Global Coastal Change

Author: Ivan Valiela

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Global Coastal Change

Global Coastal Change Ivan Valiela

© 2006 by Ivan Valiela BLACKWELL PUBLISHING 350 Main Street, Malden, MA 02148-5020, USA 9600 Garsington Road, Oxford OX4 2DQ, UK 550 Swanston Street, Carlton, Victoria 3053, Australia The right of Ivan Valiela to be identiﬁed as the Author of this Work has been asserted in accordance with the UK Copyright, Designs, and Patents Act 1988. All rights reserved. No part of this publication may be reproduced, stored in a retrieval system, or transmitted, in any form or by any means, electronic, mechanical, photocopying, recording or otherwise, except as permitted by the UK Copyright, Designs, and Patents Act 1988, without the prior permission of the publisher. First published 2006 by Blackwell Publishing Ltd 1 2006 Library of Congress Cataloging-in-Publication Data Valiela, Ivan. Global coastal change / Ivan Valiela. p. cm. Includes bibliographical references and index. ISBN-13: 978-1-4051-3685-3 (pbk. : alk. paper) ISBN-10: 1-4051-3685-5 (pbk. : alk. paper) 1. Coast changes. I. Title. GB451.2.V35 2006 333.91′714—dc22

2005015457

A catalogue record for this title is available from the British Library. Set in 11/13pt Palatino by Graphicraft Limited, Hong Kong Printed and bound in Singapore by Fabulous Printers Pte Ltd The publisher’s policy is to use permanent paper from mills that operate a sustainable forestry policy, and which has been manufactured from pulp processed using acid-free and elementary chlorine-free practices. Furthermore, the publisher ensures that the text paper and cover board used have met acceptable environmental accreditation standards. For further information on Blackwell Publishing, visit our website: www.blackwellpublishing.com

Contents Preface, vi 1.

Global context of coastal change, 1

2.

Atmospheric-driven changes, 22

3.

Sea level rise, 48

4.

Alteration of freshwater discharges, 79

5.

Alteration of sediment transport, 105

6.

Loss of coastal habitats, 124

7.

Petroleum hydrocarbons, 146

8.

Chlorinated hydrocarbons, 174

9.

Metals, 201

10.

Introduction of exotic species, 226

11.

Harvest of ﬁnﬁsh and shellﬁsh, 245

12.

Eutrophication, 283

13. Other agents of coastal change, 324 14.

Summing up, 347

Index, 357

Preface All environments on the earth’s surface have always been in ﬂux, and so they are today. The action of agents of change is evident everywhere, in the geological record, in the changing mosaic of land covers that carpets dry land. For a variety of reasons, environmental change has been particularly notable along the narrow coastal zones of the world. The changes have not escaped unnoticed: there has been concern among many about the alterations, especially as they have modiﬁed, thwarted, or prevented our uses of these environments. There are many books with titles such as The Empty Ocean, Sea of Slaughter, and so on, most of which emphasize the alarming degree of change attributable to human exploitation, uses, and alterations of virtually all coastal environments. Public and even scientiﬁc discussion and writing about too many of the examples discussed below provides ample evidence of widespread heedless unconcern with the environmental damage that follows so many human activities. On the other extreme, that bias is matched by the too-facile position of “viewing with alarm”, a position with which I largely symphathize. Yes, there will be no improvement in environmental quality unless cases are argued powerfully and rules enforced. On the other hand, it seems to me that there are issues that are more and less compelling, and we should act accordingly. There are so many causes for concern, all to some degree and at some scale important: which agent of coastal change should be considered as the higher and highest priorities? In each chapter, I have tried to provide the information I found, and thought relevant to, a thoughtful assessment of the agents of environmental change altering coastal environments at global and at local spatial scales. Advocates of some of the speciﬁc topics may ﬁnd my assessments wanting of conviction; all I can add is that the chapters hold what to me seemed assessments warranted by a careful scrutiny of the most current data and that often, the

evidence is ambiguous and incomplete. I would take it as a measure of success if the assessments I include in this book prompt skeptical or irate readers to review the evidence included, and to explore the reference materials added in the footnotes. In this book I review evidence of intensity and pervasiveness of effects, and recovery from the action of the major agents of change that are altering the diverse coastal habitats and populations of the world; a brief overview of the subjects of the chapters is provided at the end of Chapter 1. Throughout all chapters, I try to assess the degree of change forced by human and non-human inﬂuences. In reading otherwise quite good books covering the staggering recent changes in marine environments I have often felt the need to actually see the evidence underlying the alarming trends being discussed, rather than just read text “viewing with alarm”. For this reason, I endeavored, perhaps to a burdensome degree, to include relevant facts, tables, ﬁgures, and references throughout the chapters. To make the material clear, I have sorted the evidence into chapters that separately focus on each major agent of change, but, as will become evident, in the majority of cases there are joint effects of more than one agent of change, and powerful interactions. The separation of subject matter into the various chapters is a simple pedagogic device that should not be taken to mean that the impact of agents of change can readily be taken in isolation. I should add that my focus will be on the environmental effects. Although here and there I note certain relevant public health and other human effects, I do so simply to set the stage for the ecological discussion. A comparable treatment of the human effects is beyond the scope of this book. Looking back, I see that the chapters differ in length and detail. This disparity was more a reﬂection of the published literature than a

PREFACE

measure of the relative heft of the subjects. Some topics—petroleum hydrocarbons, effects of metals, impacts of overﬁshing, for example—seem to have led to greater and more detailed controversies, and attracted more logorrheic groups of practitioners writing more publications than other topics. The result is more lengthy chapters. Other chapters, for example the one on eutrophication, hold subject matter that extends across many disciplines, and hence required more space to deal with the diverse materials. I have sought to make this book accessible to a variety of stakeholders interested in environmental issues, including the interested lay public, the professional manager and decision-maker, and researchers and students dealing with coastal matters. To reach such a wide audience, I set out—ambitiously—to write three books in parallel: in each chapter I have added, as a case history, a vignette that showed, in an evident and compelling way, the essential dilemma and facts. This ﬁrst “book” was intended to clearly make the case that humans were involved in hastening change and were in turn affected by the changes. The vignettes were intended to bring the material to the street level, so to speak (in the case of sea level change in Venice, this will literally be so). There is a second, more technical “book”, which provides a more general review of the background, principles, evidence, effects, consequences, and possible remediation. A third “book” will be found in the footnotes and in the longer boxed material, where background matters, more arcane, but perhaps interesting, details are provided, technical points and terms are deﬁned and discussed, and references containing more information are given. The subject matter covered in this book has turned out to be dauntingly broad, but more problematic has been the extraordinary proliferation of publications as we turn into the 21st century. As an example, a search using just a single software search “engine”, entering the relatively specialized term “salt marsh” yielded 991 works during the last 10 years. This awesome plethora of publications has forced me to make use of a pitifully small portion of the available literature, limited by time to read titles, let alone digest all the information on sources. Every week, journals

vii

arriving at our library carried papers that forced revisions of chapters already written. This has been true for some time (recall Darwin’s plight about Wallace’s paper on the idea of evolution) but the current state of scientiﬁc publication has brought a new dimension to the overwhelming avalanche of new material. As a result, many worthy papers were unfortunately left unread and uncited. I apologize to the many authors of these unmentioned works; I have no solution for this dilemma, but it is clear that the world’s scientiﬁc community needs to search for ways to address this issue. In the blizzard of recent publications, there is also the danger of ignoring the older sources that constitute the intellectual history of the subject, and in many cases still merit knowing. To help readers, I endeavored to mention updated reviews of various ﬁelds, as they were available. The development of the internet has made available a rich variety of sources of information on the subjects covered in this book. Unfortunately, any thoughtful user will soon ﬁnd that the information available in the internet varies enormously in quality, from the entirely compelling to the completely unreliable. Moreover, the half-life of material posted in the internet seems unreliably short, with entries disappearing with no record of their existence. Internet sources are proliferating, and often hold important information inaccessible otherwise, and so I used them, but it is with some hesitation that in the chapters below I provide internet addresses as sources of material. This book was in a real way made possible by having access to the Marine Biological Laboratory– Woods Hole Oceanographic Institution Library. I am deeply indebted to Catherine Norton, Eleanor Uhlinger, Colleen Hurter, and the many other members of the Library staff, who displayed extraordinary patience with endless queries about obscure materials, arcane sources, and incomplete references. There can be few librarians more devoted to make their library work for users than those in the MBL–WHOI Library, and there are few libraries that make their impressive holdings as easily available. I have to thank Gabrielle Tomasky and Marci Cole for developing excellent graphics based

viii

PREFACE

only on my rough sketches, and Deborah Rutecki for detailed ferreting out of inconsistencies and references. Graduate students in my lab—Kevin Kroeger, Marci Cole, Jennifer Bowen, Joanna York, Ruth Carmichael, Sophia Fox, Mirta Teichberg, Sara Grady, Jennifer Culbertson, Ylva Olsen, and Nadine Lysiak—contributed many thoughts, information, and reactions during the long process of writing this book. In addition, Erin Kinney, Jayne Gardiner, Dhira Dale, and Scott Nickles, helped greatly by thoroughly reading and criticizing the manuscript during a course on the subject. I am indebted to John Farrington, Judy McDowell, Bruce Tripp, Anne Giblin, Max Holmes, Skee Houghton, and Sybil Seitzinger for

critiques of earlier drafts of various chapters, and am especially grateful to Pat Kramer, Boris Worm and Ken Tenore for extremely useful prepublication reviews of the entire manuscript. The remaining errors are mine, but all these students and colleagues helped clarify and update ideas. Jennifer Bowen and Ruth Carmichael provided data for some ﬁgures, and Joanna York translated Russian text. At Blackwell Publishing, I must thank Jane Andrew, who concertedly trained her editorial skills and imposed a measure of clarity, accuracy, and consistency on my manuscript, and Rosie Hayden who was invaluable in the process of bringing the manuscript to press and in ensuring its quality.

Chapter 1 Global context of coastal change

Allegorical engraving of the “opposition” of land and sea in the Venice Lagoon, taken from Bernardo Trevisan, Della Laguna di Venezia, Trattato, published in 1715. Reproduced here from Lasserre and Marzollo (2000).

2

CHAPTER 1

An allegorical engraving in an eighteenth century treatise on the lagoon of Venice (frontispiece) shows the “opposition” between land and sea. The somewhat artless design leaves some doubt about the relationship (are Sea and Land engaged in a struggle or an embrace?), but it does effectively convey that land and sea are linked in powerful, though perhaps ambiguous interactions, and that the vigorous action takes place at the land/sea boundary and unwittingly presciently, under the gathering clouds of atmospheric change. That the scene takes place in front of an urbanized area—Venice, in this case—merely adds the coda that the presence of humans might have some inﬂuence on the interaction between Land and Sea. There may not be a better summary of this book. The allegorical engraving captures the reality and extent of the coupling between land and coastal sea that research during succeeding centuries was to reveal. These couplings are real and sometimes altogether extraordinary. We know, for example, that decadal-scale meteorological disturbances in the Equatorial Indian and Paciﬁc Oceans prompt long-distance changes in the upper atmosphere. These alter visibility enough to change the number of visible stars in the Pleiades. This number was recorded by Inca astronomers, and even today, observations of the number of stars visible in the Pleiades is used as a criterion to decide on the planting date of potato crops in the high Andes (Orlove et al. 2000). Such remarkable long-distance connections speak of a far more complicated set of couplings between the ladies in the Venetian engraving than could possibly be conceived by the artist, but he captured the essence of the matter. This book is an account of the many ways in which human beings have altered the multiple couplings between land and sea in the narrow coastal zone where the two adjoin. In this ﬁrst chapter, I begin with a case history that subsumes virtually all the agents of change, and nearly all the couplings, that will be dealt with in later chapters. The recent history of the Black Sea includes a remarkable litany of environmental alterations, most owing to human activities. This extraordinary case history demonstrates that

what people do can have intense, widespread, multiple effects on the coastal zones of the world.

A case history: the remarkable variety of environmental changes in the Black Sea Some 12,000 to 7,000 years ago, the body of water we know as the Black Sea was a fresh- to brackishwater lake, fed by several major rivers (Fig. 1.1 top) whose ﬂows had increased as glaciation diminished. The lake was separated from the Aegean Sea, a branch of the Mediterranean, by what is now the Sea of Marmara (Fig. 1.1 bottom, Fig. 1.2 bottom). As the climate warmed further, the sea level in the Mediterranean rose above what is now the Strait of Dardanelles, and then higher than the Bosphorus Strait (Fig. 1.1 bottom, Fig. 1.2). Saltier sea water moved into the Black Sea in the deeper layers of water crossing the Bosphorus, and somewhat fresher sea water drained out to the Aegean in the upper layers. During a period of years to decades,1 the Black Sea became saltier, and the entire suite of organisms, and biogeochemical conditions within the Black Sea, changed across a few centuries. This is a compelling example of the ceaseless changes, driven by global-scale forcing, that have affected coastal and other environments throughout the history of the earth. The coasts of the Black Sea have continued to be affected by sea level rise. Differences in local geological subsidence, reduced freshwater delivery by rivers (owing to human water use, see below), and regional differences in heating of water all have conspired to create considerable variation in the local rates of recent sea level rise (Cazenave et al. 2002). In general, however, sea 1

Some researchers (Ryan et al. 1997, 2003) have argued that the incursion was a wall of water that suddenly and violently broke through the higher sill of the Bosphorus, and poured rapidly into the Black Lake, sweeping all before it. This is the “Noah’s Flood” viewpoint, harking on the biblical ﬂood idea, because the apparent date and the catastrophic nature of the event matched scriptural descriptions (Kaminski et al. 2002). Myers et al. (2003) suggest that it is more likely that the salinization of the Black Sea took place much more slowly, over at least decades. This view is supported by diverse sets of evidence in many papers appearing in issues of two journals dedicated to the matter and published in 2002 (Estuarine and Coastal Marine Science vol. 54 and Marine Geology vol. 190).

3

GLOBAL CONTEXT OF COASTAL CHANGE

BLACK SEA WATERSHED ea

North Sea

ic S

lt Ba

Danu

be Danu

be

Ad

Black Sea bul stan

ria

tic

M

ed

Se

a

Se

I

a

ite

rra

ne

an

TURKISH STRAIT SYSTEM

26°

27°

28°

29°

30° Black Sea

Figure 1.1 Top: outline (thick line) of the watershed of the Black Sea (from http://www.parliament.ge/ SOEGEO/ english/blacksea/ needs.htm). Bottom: map of the Turkish Straits area between the Black Sea and the Mediterranean, showing the Dardanelles, Sea of Marmara, and Bosphorus (from Myers et al. 2003).

41°

Aegean Sea (Mediterranean Sea)

Bosphorus Strait Sea of Marmara

Strait of Dardanelles

40°

level rise has been much higher in the Black Sea (27 ± 2.5 mm yr−1) than in the Mediterranean (7 ± 1.5 mm yr−1) (Cazenave et al. 2002). The higher rate of sea level rise makes for a greater threat for low-lying cities and wetlands on the margins of the Black Sea. There is little evidence of human inﬂuence in the Black Sea region before 4,000 years before the present (yr BP); earlier than this date, pollen deposits in sediments show that the watersheds were covered by forests (Mudie et al. 2002). After 4,000 yr BP, the pollen record in Black Sea sediments suggests that a few humans cleared land and engaged in limited agricultural practices. The pollen record agrees with archaeological

evidence of the earliest settlement of Troy and other settlements on the shores of the southern Black Sea (Bottema et al. 1995). During more recent centuries, human beings have joined global forces as major agents altering conditions in the Black Sea. Toward the close of the 20th century, on average, there were 67 people per km2 on the watershed of the Black Sea (Leppäkoski & Mihnea 1996).2 Freshwater 2

The density of people on earth during 2000 may have been 45 people km−2; by the year 2050, density is projected to reach 66 people km−2 (Cohen 2003). The density of people on the watershed of the Black Sea is therefore at what will be the global situation in 2050. Given the situation in the Black Sea, these statistics are a sobering tocsin as to future developments on wider geographic scales.

4

CHAPTER 1

Figure 1.2 Top: the Bosphorus at Istanbul (from Mee 1992). Bottom: sketch of a possible version of the changes in the postglacial sea level in the Aegean, Marmara, and Black Sea (after Myers et al. 2003; other versions given in Kaminski et al. 2002; Major et al. 2002).

ﬂow through the major rivers, particularly the Danube (Fig. 1.1 top) transports land-derived materials toward the Black Sea, but as human populations became more numerous, their activities changed the land cover on the watersheds from which the rivers received their water, sediment, and chemical loads. For example, during the 8th to 4th centuries BC, more than 62 Greek city state/colonies were founded and eventually ringed the coast of the Black Sea. Vast amounts of timber were required to support the commerce and defense of these Hellenic settlements. Homer reports in the Iliad that many Greek cities contributed some 1,093 ships to the Trojan War, so it is evident that ships were plentiful. Bondyrev (2003) speculated that 2,000–4,000 trees were required to build and outﬁt a ship, and that there might have been 100–800 oak trees per hectare, so that 25–40 ha might be felled to support one ship. Very roughly, perhaps 10,000–

12,000 ships might have been active across the Black Sea colonies, with a useful lifespan of 10 years. Bondyrev conjectured that more than 150 million ha of forests might have been cleared by the ancient Greeks between 800 and 400 BC. These numbers are of course mere guesswork, and do not consider regrowth of forests, but sufﬁce to make the point that large tracts of land were likely altered quite early in the history of human expansion onto the watershed of the Black Sea. Intensive deforestation on the watershed of the Black Sea continued into later centuries: the presence of large pine trees was one reason for the founding of the Russian Navy base in Sevastopol in the 18th century (Radchenko & Aleyev 2000). Ecological alterations to the coastal landscape have continued to this day, when about 162 million people live on the watershed of the Black Sea. As we will see in other chapters, popula-

5

GLOBAL CONTEXT OF COASTAL CHANGE

Nutrients

Prut Sulina

Ialomitra Siret Braila

Rousseé

Arges

Olt Iantra

Jiul Isker

Timok Lom

Orsova

Beograd Wien Bratislava

Krems

0

Linz

100

Passau

200 Regensburg

Altitude (m)

300

Drobeta-Turnu Severin

Tamis Morava Karas Nera

Tisza Sava

The yield of nitrogen and phosphorus out of watersheds by rivers increased as the density of human populations increased in the watersheds of the Danube and other European rivers (Fig. 1.4). A larger number of people, the expansion of urban areas, increased agricultural exploitation, and industrial development within the watersheds of the Black Sea all increased nutrient loads to the sea across the second half of the 20th century (Garnier et al. 2002). Drava

Ipoly Budapest

Sio

Vah Hron Komarno

Morava Raba

Kamp

400

Isar Inn Traun Enns Ybbs

Regen

tion and land use changes of this magnitude carry signiﬁcant consequences for the water bodies that receive the exports from the altered watersheds. Rivers such as the Danube (Fig. 1.3) ﬂow downslope towards receiving waters, in this case, the Black Sea. The rushing waters carry large loads of sediments and nutrients, even in pristine watersheds. Human activities on watersheds, however, can severely alter the rates of transport of land-derived materials to the sea, including nutrients, sediments, and water.

Delta

2,400

2,200

2,000

1,800

1,600

1,400

1,200

1,000

800

600

400

200

2,400

2,200

2,000

1,800

1,600 1,400 1,200 1,000 800 Distance from Black Sea (km)

600

400

200

Wetted section (m2)

40,000

30,000

20,000

10,000

0

Figure 1.3 Top: longitudinal proﬁle of the altitude (vertical axis) and depth of the Danube (indicated by the width of the black outline) as it courses from its origin to the Black Sea. Names next to the vertical lines show locations of cities. Bottom: extent of the cross-section of the Danube along its course. Modiﬁed from Garnier et al. 2002; redrawn from other sources.

6

CHAPTER 1

200 Fertilizer applied (tons × 104 yr−1)

2,400 1,800 1,200

0

100

200

300

400

500

Population density (individuals km−2)

Figure 1.4 Relationship of annual nitrogen (ﬁlled circles) and phosphorus (open circles) yields (per km2 of watershed) from European rivers to the density of people in the area. Data from Garnier et al. (2002).

Time courses of human activities on the watersheds of the rivers feeding the Black Sea were mirrored in the nutrient contents of the rivers and the receiving sea waters. The use of phosphate fertilizers on the watershed of the Danube, the major source of materials to the Black Sea, increased through much of the 20th century, but decreased in the 1990s as the Eastern European economies suffered a downturn (Fig. 1.5 top). The concentrations of phosphate and phosphate load delivered by the Danube followed the same time course (Fig. 1.5 middle), and there were similar changes in phosphate concentrations in the water of the northwest Black Sea (Fig. 1.5 bottom). Similar patterns were evident with nitrate. Concentrations of nitrate in coastal waters near the Danube mouth increased six-fold between the 1960s and the 1980s (Humborg et al. 1997), and tripled between 1970 and 1985 (Mee 1992).3 A greater delivery of nutrients increased concentrations in river water ﬂowing to the Black Sea, even though considerable portions of river-borne phosphate and nitrate were intercepted in the estuaries entering the Black Sea (Ragueneau et al. 2002). The retention or transformation of available nutrients within environments has been referred to as their “assimilative capacity”. In the Black Sea, as in so many coastal environments, human sources have increased nutrient loads to levels that overwhelm the assimilative capacity of the estuarine part of the rivers. The result is that substantial and increasing amounts of river-borne nutrients traverse the estuary, and go on to affect the adjoining coastal waters.

Phosphate load (tons yr−1)

0

3

150 100 50 0 1959

1969

1979

1989

1999

35 30 20 20 15 10 5 0 1959

1969

1979

1989

1999

14 12 10 8 6 4 2 0 1959

1969

1979 Year

1989

1999

600

Phosphate (µmol l−1)

N or P (kg km−2 yr−1)

3,000

Figure 1.5 Top: phosphate fertilizer use on the watershed of the Danube. Middle: discharge of phosphate by Danube water. Bottom: concentration of phosphate in the water of the northwest Black Sea. Data from Kroiss et al. (2003).

Fluvial nutrient delivery to the Black Sea decreased (Mee 1992) as a result of lower use of fertilizers during the more impecunious 1990s, and the capture of sediments behind dams. Sediments Sediments previously transported by the Danube into the Black Sea have accumulated on the bottom of reservoirs formed behind dams built to furnish hydroelectric power and to ease navigation. Iron Gate I and II dams started operation

GLOBAL CONTEXT OF COASTAL CHANGE

in the Danube in 1970 and 1984 (see Fig. 1.3 top). In the reservoirs formed behind these two dams (note the two large areas about 900–1,200 km from the Black Sea in Fig. 1.3 bottom), river ﬂow slows and sediment particles settle. Behind-dam capture of particles lowered sediment discharge by the Danube by 30–40%, a decrease that has led to intense erosion of wetlands and mudﬂats within the Danube delta, as well as in the sediment-starved shelf (Panin & Jipa 2002). Lower sediment supply also has led to erosion of shorelines and beaches (Radchenko & Aleyev 2000). The erosion has led to installation of unsightly artiﬁcial groins that not only alter sand movement, but trap tar, other pollutants, and litter. These declines in environmental quality, plus the lowered water quality, on aggregate, have deterred the tourism that has been economically important in the Black Sea, the only beach region available to millions of Eastern Europeans (Mee 1992). Crimea, historically the major tourist center in the region, lost 85% of its tourists during the last decade of the 20th century (Radchenko & Aleyev 2000).

7

1989), and by 27 and 52% by 1981 and 1985 (Mee 1992). The lower input of fresh water has made for a much shallower upper layer through the 20th century, perhaps by as much as 30 m. This shoaling of the oxygenated upper layer has in turn exposed a larger area of coastal sea ﬂoor to the anoxic lower layer, a serious development for any organism living near the bottom. Biological responses and other changes

In the Black Sea there is an oxygenated upper layer of water that is relatively fresh (22‰), and whose relatively low salinity was historically maintained by river ﬂow. This upper layer is underlain throughout the deeper areas of the Black Sea by an anoxic deep layer of salty water of Mediterranean origin.4 Use of fresh water for agricultural, industrial, and municipal purposes within the watershed of the Black Sea reduced the ﬂow of river water by about 15% by 1981 (Murray et al.

Increased nutrient supply has altered the phytoplankton of the Black Sea (Lancelot et al. 2002a, 2002b): phytoplankton densities increased, there are more blooms of fast growing species, and the species composition has changed (Table 1.1). For example, phytoplankton biomass increased 5–10-fold between the 1960s and 1970s off the Romanian coast (Leppäkoski & Mihnea 1996). Dinoﬂagellates, Noctiluca in particular, and bluegreen bacteria increased in abundance, outpacing the growth of diatoms and other groups. There were intense blooms of Mesodinium, a ciliate, off the coast of Bulgaria in the 1980s. Some of the dinoﬂagellate and blue-green blooms were toxic and led to shell- and ﬁnﬁsh kills (Fig. 1.6). The proliferation of phytoplankton biomass in the water column in turn had at least three major consequences.5 First, the increased production created a greater fall of organic particles sedimenting to the sea ﬂoor, which increased bacterial decomposition and consumed oxygen, expanding the extent of low oxygen water beyond that occurring naturally, and beyond the effect of the smaller ﬂow of fresh water. By the 1990s, 95% of the northwest Black Sea shelf, and the entire Sea of Azov were prone to episodes of low oxygen

4

5

Water

The Bosphorus sill impairs circulation below 60–200 m in the Black Sea (about 87% of the volume). Water below these depths is not ﬂushed, so there is time for bacterial action to consume oxygen faster than the oxygen is supplied by physical processes, such as diffusion, from above. Nutrient enrichment has increased the volume of anoxic water in recent decades. Nutrients (ammonium in the case of nitrogen) are recycled and released from the anoxic waters upward. Some of the ammonium that diffuses upward from the anoxic deeper water is anoxically oxidized to nitrogen gas by bacteria using a newly described reaction (annamox) (Kuypers et al. 2003), but some ammonium manages to reach the oxygenated water above, adding to the eutrophication effect.

The intensiﬁed eutrophication of the coastal waters of the Black Sea (about 30% of the Black Sea area) has led to other effects, less well-established than the ones discussed here. For example, the increased nutrient enrichment has altered the emission of important gases to the atmosphere (Amouroux et al. 2002). The enriched sediments released nitrous oxide, dimethyl sulﬁde, and methane— gases that have substantial effects on the condition of the atmosphere, and are involved in global warming trends and destruction of atmospheric ozone. A considerable part of the methane produced is oxidized within sediments and in the water column, but nevertheless some methane does manage to enter the atmosphere (Ivanov et al. 2002).

8

CHAPTER 1

Table 1.1 Densities of different groups of phytoplankton species and the number of blooms (samples with densities > 5 × 106 cells l−1) in samples taken at Constanta Station in the northwestern Black Sea, during 1960–1970 and 1980–1990. Data from Humborg et al. (1997). 1960–1970 Cell density (106 cells l−1) Diatoms Dinoflagellates Euglenophytes Prymnesiophytes

1980–1990 Number of blooms

7–21 17–51 – –

Total blooms

8 4 – – 12

Cell density (106 cells l−1)

Number of blooms

5–300 5–810 5–108 220–1,000

19 14 6 3 42

(Bakan & Büyükgüngör 2000). Species that lived on the sea ﬂoor were badly diminished: ﬁsheries for bivalves (soft-shell clams and mussels), for instance, have nearly disappeared, as have many others (Table 1.2). A single anoxic episode in the Romanian coast killed about half the ﬁsh population (Mee 1992). The shallower low oxygen layer in the water column had widespread effects on the biodiversity of organisms living on the bottom of the Black Sea (Fig. 1.7). Surveys done in the 1980s found far fewer species in the benthos, and those that survived were restricted to shallower depths compared to similar surveys done in the 1960s. Second, the increased phytoplankton density was large enough for the depth of light penetration of the Black Sea to decrease from 50–60 m in the 1960s to 10–35 m in the 1990s (Mee 1992).6 The decreased light, plus the lower oxygen that resulted from the shoaling of the upper fresher layer, severely affected the beds of algae and plants on the sea ﬂoor. By the 1990s, the area of sea ﬂoor supporting stands of Phyllophora in the northwest Black Sea had been reduced to 5% of the area of original habitat (Zaitzev & Mamaev 1997) (Fig. 1.8). Phyllophora was formerly commercially important as a source of agar, and provided a habitat to many coastal species. The Figure 1.6 Dead ﬁsh and crustaceans on a Romanian beach during a bloom of the toxic dinoﬂagellate Prorocentrum cordatum. From Leppäkoski and Mihnea (1996).

6

Bakan and Büyükgüngör (2000) reported that nutrient limitation of phytoplankton growth ceased after the 1960s; since then nutrients have been in excess of requirements, because the density of cells made for poor light penetration, so that phytoplankton growth has been limited by insufﬁcient light throughout most of the Black Sea.

9

GLOBAL CONTEXT OF COASTAL CHANGE

Table 1.2 Status of selected prominent organisms and communities in the Black Sea, as of 1996. The changes are approximations derived from comparisons of recent estimates relative to some comparable data from earlier in the 20th century. Adapted from Bakan and Büyükgüngör (2000). Type of community

Major kind of organisms

Effects relative to previous condition

Macroalgal canopies

Red macroalgae (Phyllophora) Brown macroalgae

Area of habitat reduced to 3% Less than 1% of previous area

Benthos

Mollusks Oysters Hypanis Crabs (about 14 species) Shrimps (more than 20 species)

30% of previous abundance Less than 5% of previous abundance 50% of previous abundance 30–50% of previous abundance 40% of previous abundance

Fish (e.g. 20 species of gobids, all endemic to the Black Sea) Dolphins (3 species) Monk seal

20% of previous abundance

Water column

0

20

Number of species 40 60

80

0 1960s

Depth (m)

40

1980s

80

120

160

200

Figure 1.7 Relationship of number of large benthic invertebrate species found at different depths off the coast of the Black Sea for the 1960s and 1980s. Data from Zaika (1990).

biomass of another habitat-building alga (the brown macroalga Cystoseira barbata) also diminished during the 1970s, falling from 100–8,200 to 80–290 g m−2 in rocky bottoms off Romania (Leppäkoski & Mihnea 1996). Stands of the seagrasses Zostera marina and Z. nolti covered nearly

5–10% of previous abundance Few specimens left

all the bottom of Sevastopol Bay early in the 20th century, but have largely disappeared (Shalovenkov 2000). The seagrass meadows were lost as a result of dredging of the bottom to allow port and navigation activities, and also by lowered water transparency. The fauna associated with seagrass meadows—including commercially important species such as oysters and mussels— disappeared by the end of the 20th century (Milchakova 1999). The changes in nutrient and freshwater delivery from the watershed thus thoroughly altered the bottom habitats of the coastal Black Sea. Third, the increased density of phytoplankton favored proliferation of grazers in the water column, such as copepods and protozoans.7 The biomass of herbivorous zooplankton increased more than three orders of magnitude between 1961 and 1983 (Bakan & Büyükgüngör 2000). In turn, the marked increases in zooplankton in 7

Eventually populations of the herbivorous protozoan Noctiluca became 52–88% of the zooplankton biomass. Noctiluca is not a prey readily eaten by predators, so, as Noctiluca became relatively more dominant, much less food became available for ﬁsh and other predators as the decades wore on. Perhaps the preponderance of the unpalatable Noctiluca had an indirect effect on the reduction of ﬁsheries in the Black Sea.

10

CHAPTER 1

1980s 1970s

Figure 1.8 Time course (1950s, 1960s, 1970s, and 1980s) of the reduction of sea ﬂoor area covered by a Phyllophora canopy in the northwest corner of the Black Sea. Adapted from Zaitzev and Mamaev (1997).

1960s 1950s

turn fostered increases in the numbers of ﬁsh that fed on the zooplankton (Porumb 1989). The resulting larger abundance of ﬁsh made it possible for the ﬁshing ﬂeet, particularly small-scale Turkish purse seiners, to increase harvests (Gücü 2002). Within two decades, the commercial ﬁsh catch fell by an order of magnitude (Fig. 1.9).8 As is often the case, certain species are more susceptible to human exploitation than others: the catch of sprat hardly changed across the time period shown in Fig. 1.9, while the herring and sardine catch dropped markedly. The harvest-induced collapse of the ﬁsh stocks was species- and sizespeciﬁc: larger species of pelagic ﬁshes were depleted during the 1960s and again during the 1980s. Catch of smaller pelagic species and bottom-feeding ﬁsh were lowered during the early 1990s (Gücü 2002). In 1965 there were 23 species of ﬁsh harvested commercially in the Black Sea; 8

As usual in the case of economically important resources, other explanations have been forwarded for the collapse of the Black Sea ﬁsheries (Kideys et al. 2005). Some have supposed that unspeciﬁed pollutants have interfered with the migration of anchovies and other species. Others suggested that the ﬁsh were being eaten by the porpoises that had been protected from harvest as endangered species; Mee (1992), however, reported that “hardly a dolphin is to be found”.

only ﬁve remained so by the 1990s (Zaitzev & Mamaev 1997). The magnitude and rapidity of the substantial human-generated changes in the biomass and biodiversity of the entire food web of the water column of the Black Sea are astonishing. By 1989–1990 it became evident that the removal of ﬁsh (particularly of small pelagic species) left unconsumed their erstwhile prey, which then were available to other predators (Gücü 2002; Lancelot et al. 2002a, 2002b). An apparent result of this available food supply was the proliferation of gelatinous predators, such as the native jellyﬁsh Aurelia aurita, through the 1980s. In 1989, the alien comb jelly Mnemiopsis sp.9 was introduced from the Middle Atlantic coast of North America

9

There is some question as to the taxonomic identity of this comb jelly, hence the designation as Mnemiopsis sp. (Weisse et al. 2002). Incidentally, this comb jelly was not the ﬁrst alien species in the Black Sea. In the 1940s a predatory sea snail, Rapana thomasiana, originally from Japan, invaded the Black Sea, and was particularly damaging to the oyster shellﬁshery. The snail later was subject to a ﬁshery itself, and its populations declined in the 1990s (Mee 1992). Mya arenaria, the soft-shell clam, an invader from North America, appeared in the 1960s, competed with native species, and became a valued harvested stock, but now has become depleted by the low oxygen conditions (Mee 1992).

11

GLOBAL CONTEXT OF COASTAL CHANGE

Figure 1.9 Time course of commercial ﬁsh catch in the Black Sea, 1984–1997. The species included are whiting (M. merlangus), sprat (S. sprattus), herring (C. cultriventris), and two subspecies of the European anchovy or sardine (Engraulis). From http://www.zin.ru/projects/ invasions/gaas/mnelei_e.htm.

Fish harvest (metric tons × 104)

40

Merlangius merlangus Sprattus sprattus Clupeonella cultriventris Engraulis encrasicolus maetricus E. e. ponticus

35 30 25 20 15 10 5 0 1984

in ship ballast water, and it quickly increased its range and abundance in the Black Sea.10 There were no native predators of the gelatinous predators, and apparently sufﬁcient food, so their populations expanded freely. The largest abundance of larval Mnemiopsis was in the area near the plume produced by the ﬂow of the Danube (Weisse et al. 2002), where the impact of enrichment was greatest. The data on densities of populations of gelatinous species have been said to be unreliable (Weisse et al. 2002); at best, estimates are variable, but Mnemiopsis sp. did seem to bloom irregularly and at times profusely through the 1990s (Shiganova et al. 2000). Their voracious feeding (maybe 300 copepods per day) led to an order of magnitude lowering of zooplankton abundance during the late 1980s and early 1990s (Weisse et al. 2002). Mnemiopsis also feed on eggs and larvae of anchovies and other ﬁsh, and have been thought to have helped lower

1986

1988

1990 Year

1992

1994

1996

the abundance of smaller pelagic ﬁsh species (Shiganova et al. 2000).11 Mnemiopsis sp. have become less abundant in more recent years as their plankton food supply has sharply diminished, and as a new alien comb jelly, Beroe ovata, a specialist in feeding on other comb jellies, entered the Black Sea in 1999, and presumably fed freely on Mnemiopsis sp. The booms and busts of populations in the Black Sea food web during the second half of the 20th century demonstrate several important points. Assemblages of organisms in coastal environments can be dramatically reshufﬂed by changes in nutrient supply to producers on the bottom rung of food webs, and by changes in consumers at the top of food webs. Such bottom-up and top-down forcings interact in a complicated fashion. Human beings are heavily involved in the reshufﬂings, mainly by increasing nutrient supplies and by overﬁshing stocks, as well as by altering other features of the land/sea coupling.

10

Some have concluded that the bloom of Mnemiopsis sp. was the cause of the decrease in ﬁshery catch because the comb jellies consumed larval ﬁsh. Mnemiopsis sp. certainly eat ﬁsh larvae, but model studies suggest that the overﬁshing anteceded the proliferation of comb jellies (Gücü 2002; Lancelot et al. 2002a, 2002b). In addition, ﬁsh catch data for the Mediterranean, where there were no blooms of comb jellies, show depletion patterns quite similar to those of the Black Sea (Lleonart & Maynou 2003), with a large depletion of smaller pelagic species. This coincidence suggests that ﬁshing pressures may be involved in the time course of catch in both seas.

11

The number of ﬁsh eggs sampled in Sevastopol Bay, in the Crimean Peninsula on the north coast of the Black Sea, ranged from 2,380 to 7,990 eggs per 10 min of trawling, including 16–18 species, during the 1950s. During the 1980s, these numbers were 77–792, with 6–12 species represented. During the 1990s the numbers were lower, 60–1,027 eggs, with 3–12 species present. These reductions were most likely the result of several factors, including lower oxygen, contaminants, grazing, and overﬁshing (Gordina et al. 2001).

12

CHAPTER 1

The different conclusions reached by reasonable scholars viewing the same phenomena about the goings on in the Black Sea illustrate another feature that we will have to deal with throughout this book: data are often insufﬁcient, and no matter their quantity or quality, different people may interpret the information differently. Because data and interpretation are so manifestly controversial issues, here and in the following chapters, within reasonable space limits, I make it a point to present the actual evidence (in ﬁgures, tables, and the like) supporting the assertions made. Inputs of other pollutants In addition to nutrients, the rivers entering the Black Sea carry many other contaminants, including metals, petroleum hydrocarbons, and synthetic organochlorines. In terms of metals, during the 1980s, the Danube alone discharged 4,500 tons of lead, 1,000 tons of chromium, 900 tons of copper, and up to 60 tons of mercury (Mee 1992). Signiﬁcant concentrations of mercury were found within mussels (Ryabushko et al. 2002). Anthropogenic contributions have in some cases tripled the concentrations of heavy metals found in shallow coastal Black Sea sediments (Secrieru & Secrieru 2002). Top predators such as marine mammals, however, do not carry high metal loads in their tissues: mercury in porpoises from the Black Sea was one order of magnitude lower than in the North Sea, for example ( Joiris et al. 2001). Rivers also transport petroleum hydrocarbons into the Black Sea. About 50,000 tons of petroleum compounds were discharged annually during the 1980s by the Danube into the Black Sea (Mee 1992), but only about 2,600 tons during the 1990s (Bakan & Büyükgüngör 2000). Small-scale discharges across many different sites, derived from discharges of municipal, domestic, and industrial sources, were more than an order of magnitude larger than amounts released from accidents and ballast water discharge from tankers, but high concentrations of oil have been recorded along shipping lanes (Mee 1992).

Concentrations of persistent organochlorine residues in the sediments of the Danube delta are among the highest recorded (Fillmann et al. 2002). High concentrations of chlorinated hydrocarbons and polychlorinated biphenyls (PCBs) were also found in the plume of the Danube (Maldonado & Bayona 2002). During the 1980s, concentrations of DDT in sediments of the Danube delta were about 1,000 times as large as comparable Mediterranean sediments. DDT contents in ﬁsh caught in Turkish waters neared the limit for human consumption, and were 5–10 times as large as concentrations found in the Baltic, an environment considered contaminated (Mee 1992). The recently deposited and high concentrations of chlorinated hydrocarbons, including DDT, suggest that these compounds were used recently within the watersheds of rivers that carry the compounds to the Black Sea (Fillmann et al. 2002), even though some of these organochlorines are banned in many countries. If the inputs of metals and organic compounds were not enough, the Black Sea watershed was also exposed to the largest ever accident involving the release of radioactive materials. Radionuclides from the Chernobyl accident fell widely on the watershed of the Danube, and adsorbed into soil particles. Erosion and ﬂuvial transport then conspired to convey the adsorbed radionuclides to the Black Sea. Radioactive materials were largely deposited in near-shore sediments under the Danube plume (Gulin et al. 2002). The accumulation in sediments showed a 5-year delay following peak atmospheric fallout, and the radionuclides in the surface sediments have decreased substantially since (Gulin et al. 2002). Fortunately, radioactive contamination of organisms within the Black Sea did not reach levels that might have biological consequences (Egorov et al. 2002). Threats to habitats Much as the erstwhile algal and seagrass canopies discussed earlier, coastal wetlands on the coast of the Black Sea were productive habitats. These wetlands are essential for many birds and mammals, and, as we will see in later chapters,

GLOBAL CONTEXT OF COASTAL CHANGE

play important ecological functions at the interface between the land and sea. In the Black Sea, coastal wetlands have been subject to alteration and destruction by wastewater discharge and toxic pollutants, “reclamation” of wetlands for agriculture, ﬁlling for construction, and the dumping of dredge spoils and solid waste (Bakan & Büyükgüngör 2000). The losses of these valuable habitats have not been measured, nor the consequent effects assessed, but they are surely not minor.

13

Black Sea as we turn towards the 21st century is an unfortunate harbinger of the complex suite of alterations that can be expected in coastal environments as people make more intensive use of land on watersheds, and on the water itself. The accelerating and bewilderingly diverse changes that seem the fate of environments along the coasts of the world, do, however, have some common root causes.

Underlying causes of global coastal change Signs of some recovery Within recent years, there are some signs of recovery in the coastal Black Sea. The nitrogen and phosphorus loads to the Black Sea have decreased by 25 and 50%, respectively (Lancelot et al. 2002a, 2002b). Nutrient concentrations in the Danube outﬂow decreased, perhaps by half or so, between the 1980s and 1990s (Lancelot et al. 2002a, 2002b). The biomass of Mnemiopsis sp. during 1995 was an order of magnitude lower than during 1990 (Weisse et al. 2002). It would be reassuring to report that the lowering of nutrient loads might be the result of improved management practices, but it seems more likely that the economic decline suffered by central European countries within the Black Sea watershed might be a more reasonable explanation (Garnier et al. 2002). This speculation again demonstrates the powerful ties between environmental changes and human activities, and the inﬂuence of factors completely external to ecology on environmental conditions. In the Black Sea we have a remarkable case history where eutrophication has become a primary environmental challenge, exacerbated by overﬁshing, loss of habitats, reduction of fresh water and sediment supplies, introduction of alien species, and a host of other agents of environmental change (Bakan & Büyükgüngör 2000). Perhaps the examples to be reviewed in the chapters that follow lack the remarkable compounded variety and intensity of perturbations visited upon the Black Sea shores, but we will see that similar changes are occurring in many locales, and at global scales. The situation in the

The central facts about change in the coastal zones of the world are that: i) there are increasingly more people in the world and their activities have historically been focused at the land/sea boundary, for biological, cultural, economic, and geographic reasons; and ii) these people consume resources. These two central features, as we will see throughout the chapters of this book, bring about changes in coastal environments. The changes generated by humans, moreover, have, during the 20th century, become large enough to exceed the changes pressed upon coastal environments by “natural” or non-anthropogenic forces. Increases in human populations Global population numbers and growth

The best estimates of total numbers of humans available suggest that the human population on earth has increased, particularly during the 20th century (Fig. 1.10). We can take a closer look at estimates of recent human population growth by examining the probable course of events between 1950 and 2050 (Fig. 1.11 top). Human populations across the last half of the 20th century increased dramatically, although there is a glimmer of future relief from the pace of growth. Predicted rates, based on censuses and wellestablished population models, predict a lower rate of people added per year (Fig. 1.11 middle), and a decrease in annual growth as we move into the 21st century (Fig. 1.11 bottom). In fact, there is an 80% chance that by the year 2100—two to three generations from now—the size of the

14

CHAPTER 1

12 11

2100

Number of people (billions)

10 Modern Age

9 Old 8 Stone Age 7

New Stone Age

Bronze Age

Iron Age

Middle Ages 2000

6 5

1975

4 3

1950

2 Plague

1 7,000

5,000 3,000 Years BC

1900 1800

1,000 1 1,000

human population will have stopped growing or could decrease (Lutz et al. 2001).12 Global population numbers are shown as a well-deﬁned curve in Fig. 1.11, but there is a substantial uncertainty in the estimates (Keilman 2001). The curve represents best estimates, and there is a range of probable values lower and higher at each year. For the year 2050, for instance, when the world’s population is on average expected to reach about 9 billion people,

12

The causes of the decreases are beyond the scope of this book. Sufﬁce it to say that powerful economic and social mechanisms have altered human population dynamics. Economic stringencies, increased availability of birth control measures, emergence of the issue of women’s rights, delayed ﬁrst reproduction, demand for more education, improved infant mortality, social security, and health plans, are among the underlying factors. In certain regions, diseases such HIV/AIDS will certainly have some effect, although it is the case that except for the Black Plague in the Middle Ages (Fig. 1.10), diseases (and wars) have not prevented human population growth. The decreased growth rates will bring other difﬁcult issues to the fore, including the impending social and economic crisis inherent in the increasing preponderance of older age groups in many countries (Cohen 2003). We might note, in passing, that the posited mechanisms for decreased human population growth are not ecological: we have successfully forced ourselves outside the ambit of the ecological forces that govern population numbers in all other species.

3,000 5,000 Years AD

Figure 1.10 Estimated population across human history. Adapted from Population Reference Bureau and United Nations, http://www. prb.org, http://www.un.org/ popin/wdtrends.html.

estimates by Lutz et al. (2001) are that there is a 95% probability that the human population value will lie between 6.6 and 11.4 billion people. In addition, there are some discrepancies among different methods of arriving at the estimates: Bongaarts and Bulatao (2000), for example, give that same range as 7.9–10.9 billion. Even just considering the Lutz et al. (2001) values, there is an uncertainty of about 27% in the best estimates. Regardless of the uncertainty, however, we can expect substantial increases in the number of human inhabitants of the earth before population growth tapers off. We need to note, however, two salient facts. First, we will still have to deal with in the course of our next two generations—regardless of the uncertainty of the estimates—substantial increases in human numbers. By the end of the present century there will be about 11 billion people on earth, a substantial increase from the current population of about 6 billion (Fig. 1.10). Second, during the 21st century the number of people living in urban environments will increase faster than the rural population (Fig. 1.11 top). In 1800 about 2% of humans lived in cities; this percentage increased to 12% by 1900, and to 47%

10

8 Number of watersheds

World population estimate ( ) and projection ( ) (× 109)

GLOBAL CONTEXT OF COASTAL CHANGE

8 6 Urban 4

Rural

2 0 1950

1970

1990

2010

2030

2050

4 2 0

−5 −4 −3 −2 −1 0 1 2 3 4 5 6 Area in watershed with changing land use (%)

80 16

60

Increase Decrease No change

40 12

20 0 1980

1990

2000

2010

2020

2.0

Frequency

People added per year (× 106)

Agriculture Forest Urban

6

100

% increase per year

15

8 4

1.6 1.2

0 0

0.8 0.4 0.0 1980

1990

2000 Year

2010

2020

Figure 1.11 Top: estimated and projected total world population, and breakdown into rural and urban populations, 1950–2050 (data from FAO, http://www.fao.org). Middle: number of people added annually; and bottom: annual percentage increase, 1980–2020 (data from Population Reference Bureau and United Nations, http://www.prb.org, http://www.un.org/popin/wdtrends.html).

by 2000 (Cohen 2003). The years 2005–2010 are a momentous transition in human history; by this half decade, half of us will be living in urban settings. The increased urbanization of the earth’s surface manifests itself directly by the changing land uses on watersheds (Fig. 1.12). The percentage of urban land covers on watersheds of the eastern USA has increased in recent decades, at the expense of forest and agricultural land covers (Fig. 1.12 top). In addition, the largest increases in urbanization are taking place in those water-

0.1 0.2 0.4 0.8 1.6 3.1 6.3 12.5 25 50 100 Urban area (%)

Figure 1.12 Top: frequency distributions of gains and losses of area covered by agricultural, forest, or urbanized land covers, 1982–1992 in 16 watersheds of the eastern USA. USDA data, from Van Breemen et al. (2002, ﬁg. 5). Bottom: number of watersheds with different percentages of urbanized land cover within 51 eastern US watersheds. Each bar is broken up into those watersheds where the human population within the urbanized areas were increasing, decreasing, or showed no change across the previous decade. Data from Dow and DeWalle (2000).

sheds with the greater proportion of urban land covers (Fig. 1.12 bottom). At present rates of urbanization in Europe, urban areas will double in less than a century (El Pais, Nov. 3, 2005). Urban sprawl is therefore becoming pervasive worldwide, and its spread is accelerating. The remarkable image shown in Fig. 1.13 conveys the startling degree to which the earth is already urbanized. The development of large metropolises and innumerable smaller urban areas throughout the sample area (Europe) is evident.

16

CHAPTER 1

Local population density and growth

The environmental consequences of the spread of urban sprawl cannot be overemphasized, as will become evident in the chapters that follow. The concentration of people in urban centers necessarily leads to greater demands in the consumption of energy, water, food, and other resources, as well as complicating the disposal of liquid and solid wastes, and concentrating contamination of air, water, and soils. This uncoupling of day-to-day human experience from the source of their sustenance may make it more difﬁcult to develop social and political support for measures to maintain sustainable environmental uses. 500 Number of people (× 106) ( )

180 400 150 300 120 200 90 100 60 0

0

20

40 60 80 Distance from coast (km)

100

Population density (individuals km−2) (–)

Figure 1.13 Mosaic of enhanced night images of the European region. From http://www.gsfc.nasa.gov/ topstory/2003/0815citylights.html.

The facts about global population growth are evident. Population density, however, varies at local spatial scales, particularly near coasts. In addition to the proliferation of urban centers and accompanying sprawl, Fig. 1.13 highlights a second spatial feature: people accumulate near the coast. People have preferred to settle along coastlines through human history. Whether it was because waterways were the major means of transport, or because resources were more readily available along shores, humans seemed to settle along coasts, and still do so today. As of 1990, about 23% of the human population live within the narrow strip within 100 km from coasts (Nicholls & Small 2002). The density of people in near-coast zones is 112 people km−2; this is 2.5 times the mean global population density of 44 people km−2 (Nicholls & Small 2002). We can see this pattern in the shorelines that are so clearly outlined by the night lighting in the image of Fig. 1.13. Near-shore settlement (and activities) is evident throughout the world (Fig. 1.14): there are a surprisingly large number of us that live within 5 km of the shoreline. The pattern of emphasized human use of nearcoast areas occurs not only at a global spatial scale, but at smaller scales as well. The urbanization of landscapes and the tendency for faster development nearer to coasts are evident at a

Figure 1.14 Worldwide estimates of the number of people (left vertical axis) and people per unit area (right vertical axis) living in 5 km bands away from the coast. From Nicholls and Small (2002).

GLOBAL CONTEXT OF COASTAL CHANGE

people km−2. About 40% of coastal people live in small towns or settlements at densities of more than 1,000 people km−2. The remaining half of the coastal population lives in rural or isolated areas. The quite uneven along-shore distribution means that there are going to be quite disparate effects on the coastal environments receiving inputs in different parts of the world’s coasts. Such marked spatial heterogeneities of human pressures on coastal environments are going to appear more than once in the chapters that follow.

Number of houses

600 1984 1974 1955 1938

400

200

0

200

400 600 800 Distance from shore (m)

17

1,000

Figure 1.15 Number of houses located in 100 m bands away from shore along the coast of Waquoit Bay, Cape Cod, Massachusetts, 1938–1984. The number of houses built by each of four dates are also indicated. From Valiela et al. (1992).

global scale (Fig. 1.14), as well as at smaller local spatial scales (Fig. 1.15). People built dwellings close to the coast of Waquoit Bay, a small water body on the coast of Cape Cod, Massachusetts, and, as the decades passed, those who found the nearer-shore plots occupied ﬁlled in as close to water as was then possible (Fig. 1.15). The pattern of impressive near-shore development we have been documenting is widespread. Culliton et al. (1990) compiled information on the rate of population increases in each county throughout the United States, and found the fastest rates of human population growth in coastal counties. This was true for every stretch of coast throughout the United States, and, we might suspect, for other areas of the world. We may conclude that rates of increase in human populations are greater near shores, and that these will increase in parallel but faster than world totals. The human populations accumulated against the shorelines of the world are not, however, uniformly distributed. There are pronounced local differences in human density along the coastal strips of the world (Nicholls & Small 2002). About 10% of coastal people live concentrated in urban areas at densities greater than 10,000

Use of resources Human beings of course consume resources, and more people means more pressure on the environments that produce the required resources. For the sake of brevity, here we will deal with only one aspect of one important human activity, agricultural production of food. This may seem an odd choice in a book about coastal environments. There are at least two reasons for discussing food production here. First, people grow food on land, and what happens to coastal environments is in large measure—but of course not completely—caused by what people happen to do on land, a point to be made repeatedly throughout the chapters below. Second, as in the allegorical Venetian engraving, land and sea are coupled: increasing food yields on land often leads to increased transport of materials to receiving coastal environments. Such exports, as we will see in subsequent chapters, force substantial coastal environmental alterations. As human populations increased through the 19th and 20th centuries (see Fig. 1.10), food production had to increase just to feed people, let alone to allow an improved standard of living and nutrition. The ﬁrst approach was to ﬁnd more and more arable land to bring under cultivation, a strategy that was effective up to about 1940. After the mid 20th century, the rate of population increase was faster than the feasible expansion of area of cultivable land, and the ratio of cropland to people slowly decreased (Fig. 1.16). The pressure to increase yield on a per parcel basis forced the development of improved

18

CHAPTER 1

Grain production (kg person−1) Cropland (ha people−1)

Cropland

Fertilizer

90 80

400

70 60

300

50

Grain

40

200

30 Meat

100

20 10 0

0 1890

1930

1960 Year

1980

agricultural practices. Chief among the new methods was the fertilization of crops, using manures as a nitrogen supply and phosphates mined from far-off guano islands (Anderson 1993), or later on in the century, industrially produced nitrate.13 The application of nitrate increased markedly through the last half of the 20th century (Fig. 1.16). The collapse of the USSR and its huge agricultural sector created the evident dip during the early 1990s, and the slowing of the world economy in general diminished fertilizer use toward the end of the 20th century. With the aid of fertilizers, the production of grains and meat—the core agricultural staples—have, in recent decades, more than kept pace with the demands of a growing population.14 The recent

13

Fritz Haber, a German chemist, applied a suggestion by Carl Bosch to develop the ﬁrst successful commercial reaction that led to mass production of nitrogen fertilizers and military explosives. Hydrogen and nitrogen, when appropriately heated and compressed, produce ammonium (NH4), a compound that can then be used for many purposes. During World War I, Haber directed German chemical warfare activities. Haber won the Nobel Prize in 1918 for his discovery. After the Nazi rise to power, Haber defected to the UK. The Haber–Bosch process is one of those crucible inventions that have changed the world as we know it, for good or bad. 14 There are unfortunate regions of the world where famine is still common. These regrettable conditions are more the result of political and social issues affecting distribution than a matter of magnitude of food supply. These aspects once again illustrate that what might be said at a global scale may not apply at local scales.

1997

Meat production (kg person−1) Fertilizer production (× 106 metric tons)

100

500

Figure 1.16 Time course of per capita cropland area, production of grain and meat (data simpliﬁed from Galloway 1998), and world nitrogen fertilizer production (from http://www.fao.org).

decrease in use of fertilizers might suggest that, since the world population has continued to increase, as the centuries turn, we might enter a period where per capita nutrition for the world’s human population diminishes, and pressures will build to increase yields again.15 The worldwide use of nitrogen fertilizer has no doubt been a life-saver for many people in many places, as well as improving the quality of nutrition worldwide. It turned out, however, that a substantial portion of the nitrogen used as fertilizer eventually reaches the sea, and the exported nitrogen has manifold effects on aquatic systems (cf. Chapter 12). The Haber–Bosch process yields ammonium, which is reduced to nitrate, either in the production of fertilizer or by bacteria in the environment. This nitrate is a highly mobile form of nitrogen and travels readily in moving water. About 11% of the nitrogen fertilizer used for crops travels downstream in rivers (Seitzinger & Kroeze 1998) and most of that reaches estuarine and near-shore

15

It seems inevitable that crops that are more effective at uptake and retention of nitrogen will be created using genetic engineering methods. There are sectors of society that are suspicious of these new developments, but given what we will ﬁnd out in Chapter 12 about the consequences of increased application of fertilizers (across poorer and poorer soils), it may be concluded that developing genetically engineered crops is an attractive option that might be a necessary recourse to meet increased demand for food.

GLOBAL CONTEXT OF COASTAL CHANGE

environments. Once in the coastal environments, the land-derived nitrate enters and stimulates many reactions and transformations, and prompts wholesale alterations that will be considered in Chapter 12. Sufﬁce it to say that human demands for food have increased transport of a highly reactive compound that thoroughly alters the receiving coastal environments. As it turns out, much of the nitrogen discharged to coastal environments (Nixon et al. 1996; Seitzinger & Giblin 1996) is converted to N2 gas (a relatively unavailable form of nitrogen) by bacteria in coastal environments or is buried in coastal sediments. Coastal ecosystems thus to a degree intercept landderived nitrogen, but in the process the coastal environments are thoroughly altered. Much like population growth in the coastal zones of the world, the use and impact of fertilizer nitrogen has a highly heterogeneous distribution. Fossil fuel combustion is widespread in many countries, and adds oxidized nitrogen compounds to the atmosphere. Depending on wind direction, atmospherically transported nitrogen compounds can be deposited in different coastal environments. In addition, humans produce prodigious amounts of waste water that contain nitrogen, and its inevitable local disposal adds to the spatial heterogeneity of nitrogen impacts on coastal waters. This example makes evident how human demand for resources forces changes on coastal systems. It also makes the point that the coastal systems themselves are not independent of what takes place on land.

Contents of this book If there is one reality about coastal environments, it is that they have always changed. There is plentiful evidence that there have been enormous changes along the coastlines of the world in geological and shorter time scales. The notion that there is something that could be called the “balance of nature” will certainly be far less compelling after we parse the available evidence. In this book we examine those changes that are affecting and altering the world’s coastal environments within time scales of days to centuries.

19

Human activities can interact with or accelerate rates of natural changes. In the various chapters we will investigate just how far anthropogenic actions have enhanced changes driven by “natural” agents. We will concentrate on evidence about ecological or biogeochemical changes, largely caused by human activities of one sort or another. As will become evident, in the current global condition, we cannot divorce humans from ecology. We will consider not just ecological alterations, but how these environmental changes interact with people’s use of coastal environments. A ﬁrst group of two chapters deals with ecological changes driven by changing global atmospheric and climate alterations. These chapters include features of, and effects on, climatic warming, increased ultraviolet radiation, and the accelerated rise in sea level. A second group of two chapters addresses coastal change created by human use of water on land, with chapters on the consequences of the interception of fresh water before it arrives at the shore (Chapter 4) and of the interception and increased erosion of terrestrial sediments (Chapter 5). Chapter 6 covers human destruction of the various habitats that occupy the near-shore zones. These include the losses of salt marshes, mangroves, and coral reefs, as well as other less threatened habitats. Three chapters deal with toxic substances— petroleum hydrocarbons (Chapter 7), chlorinated compounds (Chapter 8), and metals (Chapter 9)—that humans add to coastal environments. Three chapters deal with changes that humans force on coastal food webs, including nutrient additions (Chapter 12 on eutrophication), which alter ecosystems by effects that permeate up food webs and create pervasive effects, and humandriven alterations in the species composition of coastal environments (Chapter 10 on biological invasions, and Chapter 11 on overharvest of coastal stocks). Chapter 13 contains information on a miscellany of other agents of environmental change (litter, thermal, sound, and radioactive pollution, and human pathogens). The ﬁnal chapter addresses ways in which the agents of change and their effects interact, and how we might compare the various agents of change in

20

CHAPTER 1

terms of the intensity of the changes, the extensiveness of the effects, the possibility of recovery from the perturbations, and how to assess the relative priorities we might place on addressing the issues raised by the chapters in this book. This last chapter also evaluates how we might set priorities for management or restoration of the impending or already present changes.

References Amouroux, D., G. Roberts, S. Rapsomanikis, and M. O. Andreae. 2002. Biogenic gas (CH4, N2O, DMS) emission to the atmosphere from near-shore and shelf waters of the north-western Black Sea. Estuar. Coast. Shelf Sci. 54:575–587. Anderson, T. D. 1993. The U.S. guano islands of the Paciﬁc: A brief background. Ohio J. Sci. 93:50. Bakan, G., and H. Büyükgüngör. 2000. The Black Sea. Mar. Poll. Bull. 41:24–43. Bondyrev, I. V. 2003. Colonization of the Black Sea by the ancient Greeks and its ecological consequences. http:// www.transoxiana.com.ar/Eran/Articles/bondyrev.html. Bongaarts, J., and R. A. Bulatao (eds). 2000. Beyond Six Billion: Forecasting the World’s Population. National Research Council. Committee on Population, Commission on Behavioral and Social Sciences and Education. National Academy Press, Washington, DC. Bottema, S., H. Woldring, and B. Aytug. 1995. Late Quaternary vegetation of northern Turkey. Paleohistoria 17:53–142. Cazenave, A., P. Bonnefond, F. Mercier, K. Dominh, and V. Toumazou. 2002. Sea level variations in the Mediterranean and the Black Sea from satellite altimetry and tide gauges. Glob. Planet. Change 34:59–86. Cohen, J. E. 2003. Human population: The next half century. Science 302:1172–1175. Culliton, T. J., and 5 others. 1990. 50 years of population change along the nation’s coasts, 1960–2010. National Oceanographic and Atmospheric Administration. Rockville MD, 41 pp. Dow, C. L., and D. R. DeWalle. 2000. Trends in evaporation and Bowen ratio on urbanizing watersheds in the eastern United States. Water Res. Res. 36:1835–1843. Egorov, V. N., and 5 others. 2002. The Black Sea radioecological response to 90Sr and 137Cs after the Chernobyl Nuclear Plant accident. Mor. Ehkol. Zh. 1:5–15. Fillmann, G., and 6 others. 2002. Persistent organochlorine residues in sediments from the Black Sea. Mar. Pollut. Bull. 44:122–133. Galloway, J. N. 1998. The global nitrogen cycle: Changes and consequences. Environ. Pollut. 102 Sl:15–24.

Garnier, J., and 5 others. 2002. Modelling transfer and retention of nutrients in drainage network of the Danube River. Estuar. Coast. Shelf Sci. 54:285–308. Gordina, A. D., and 5 others. 2001. Long-term changes in Sevastopol Bay (the Black Sea) with particular reference to the ichthyoplankton and zooplankton. Estuar. Coast. Shelf Sci. 52:1–13. Gücü, A. C. 2002. Can overﬁshing be responsible for the successful establishment of Mnemiopsis leidyi in the Black Sea? Estuar. Coast. Shelf Sci. 54:439–451. Gulin, S. B., and 5 others. 2002. Radioactive contamination of the north-western Black Sea sediments. Estuar. Coast. Shelf Sci. 54:541–549. Humborg, C., V. Ittekkot, A. Cociasu, and B. V. Bodungen. 1997. Effect of Danube River dam on Black Sea biogeochemistry and ecosystem structure. Nature 386:385– 388. Ivanov, M. V., N. V. Pimenov, I. I. Rusanov, and A. Lein. 2002. Microbial processes of the methane cycle at the north-western shelf of the Black Sea. Estuar. Coast. Shelf Sci. 54:589–599. Joiris, C. R., and 7 others. 2001. Total and organic mercury in the Black Sea porpoise Phocoena phocoena relicta. Mar. Pollut. Bull. 42:905–911. Kaminski, M. A., and 5 others. 2002. Late glacial to Holocene benthic foraminifera in the Marmara Sea: Implications for the Black Sea-Mediterranean connections following the last deglaciation. Mar. Geol. 190:165–202. Kedeys, A. E., A. Roohi, S. Bagheri, G. Finenko, and L. Kamburska. 2005. Impact of invasive ctenophores on the ﬁsheries of the Black Sea. Oceanography 18:76–85. Keilman, N. 2001. Uncertain population forecasts. Nature 412:490–491. Kroiss, H., M. Zessner, and C. Lampert. 2003. Nutrient management in the Danube basin and its impact on the Black Sea. J. Coast. Res. 19:898–906. Kuypers, M. M. M., and 8 others. 2003. Anaerobic ammonium oxidation by annamox bacteria in the Black Sea. Nature 422:608–611. Lancelot, C., J.-M. Martin, N. Panin, and Y. Zaitzev. 2002a. The North-western Black Sea: A pilot site to understand the complex interaction between human activities and the coastal environment. Estuar. Coast. Shelf Sci. 54:279– 283. Lancelot, C., J. Staneva, D. Van Eeckhout, J.-M. Beckers, and E. Stanev. 2002b. Modelling the Danube-inﬂuenced North-western continental shelf of the Black Sea. II: Ecosystem response to changes in nutrient delivery by the Danube River after its damming in 1972. Estuar. Coast. Shelf Sci. 54:473–499. Lasserre, P., and A. Marzollo. 2000. The Venice Lagoon Ecosystem Project: Genesis, goals, and overview. Pp. 1–22 in Lasserre, P., and A. Marzollo (eds). The Venice Lagoon Ecosystem: Inputs and Interconnections Between Land and Sea. UNESCO and Parthenon Publishing Group, Paris.

GLOBAL CONTEXT OF COASTAL CHANGE

Leppäkoski, E., and P. E. Mihnea. 1996. Enclosed seas under man-induced change: A comparison between the Baltic and Black Seas. Ambio 25:380–389. Lleonart, J., and F. Maynou. 2003. Fish stock assessments in the Mediterranean: State of the art. Scientia Marina 67 (suppl):34–49. Lutz, W., W. Sanderson, and S. Scherbov. 2001. The end of world population growth. Nature 412:543–545. Major, C., W. Ryan, G. Lericolais, and I. Hajdas. 2002. Constraints on Black Sea outﬂow to the Sea of Marmara during the last glacial-interglacial transition. Mar. Geol. 190:19–34. Maldonado, C., and J. M. Bayona. 2002. Organochlorine compounds in the western Black Sea: Distribution and water column processes. Estuar. Coast. Shelf Sci. 54:527–540. Mee, L. D. 1992. The Black Sea in crisis: A need for concerted international action. Ambio 21:278–286. Milchakova, N. A. 1999. On the status of seagrass communities in the Black Sea. Aquat. Bot. 65:21–32. Mudie, P. J., A. Rochon, A. E. Aksu, and H. Gillespie. 2002. Dinoﬂagellate cysts, freshwater algae and fungal spores as salinity indicators in Late Quaternary cores from Marmara and Black Seas. Mar. Geol. 190:203–231. Murray, J. W., H. W. Jannasch, and S. Honjo. 1989. Unexpected changes in the oxic/anoxic interface in the Black Sea. Nature 338:411–413. Myers, P. G., C. Wiecki, S. B. Goldstein, and E. J. Rohling. 2003. Hydraulic calculations of postglacial connections between the Mediterranean and the Black Sea. Mar. Geol. 201:253–267. Nicholls, R. J., and C. Small. 2002. Improved estimates of coastal population and exposure to hazards released. Eos 83:301,305. Nixon, S. W., and 15 others. 1996. The fate of nitrogen and phosphorus at the land–sea margin of the North Atlantic Ocean. Pp. 141–180 in Howarth, R. W. (ed.). Nitrogen Cycling in the North Atlantic and its Watersheds. Kluwer, Dordrecht, The Netherlands. Orlove, B. S., J. C. H. Chiang, and M. A. Cane. 2000. Forecasting Andean rainfall and crop yield from the inﬂuence of El Niño on Pleiades visibility. Nature 403:68–71. Panin, N., and D. Jipa. 2002. Danube River sediment input and its interaction with the North-western Black Sea: Results of the EROS-2000 and EROS-21 projects. Estuar. Coast. Shelf Sci. 54:551–562. Porumb, F. 1989. On the development of Noctiluca scintillans under eutrophication of Romanian Black Sea coastal waters. Sci. Tot. Environ. Suppl. 1992:907–920. Radchenko, V. N., and M. Y. Aleyev. 2000. Environmental and social impacts of management approaches in Sevastopol Bay in a historic retrospective: A case study from the Black Sea. Ocean Coast. Manag. 43:793–817.

21

Ragueneau, O., and 12 others. 2002. Biochemical transformations of inorganic nutrients in mixing zones between the Danube River and the north-western Black Sea. Estuar. Coast. Shelf Sci. 54:321–336. Ryabushko, V. I., V. N. Egorov, A. F. Kozintsev, S. K. Kostova, and V. K. Shinkarenko. 2002. Mercury in the mussel Mytilus galloprovincialis Lam. from bays of the Crimean Peninsula of the Black Sea. Mor. Ehkol. Zh. 1:99–107. Ryan, W. B. F., and 9 others. 1997. An abrupt drowning of the Black Sea shelf. Mar. Geol. 138:119–126. Ryan, W. B. F., C. O. Major, G. Lericolais, and S. L. Goldstein. 2003. Catastrophic ﬂooding of the Black Sea. Ann. Rev. Earth Planet. Sci. 31:525–554. Secrieru, D., and A. Secrieru. 2002. Heavy metal enrichment of man-made origin of superﬁcial sediment on the continental shelf of the North-western Black Sea. Estuar. Coast. Shelf Sci. 54:513–526. Seitzinger, S. P., and A. E. Giblin. 1996. Estimating denitriﬁcation in North Atlantic shelf sediments. Pp. 235–260 in Howarth, R. W. (ed.). Nitrogen Cycling in the North Atlantic and its Watersheds. Kluwer, Dordrecht, The Netherlands. Seitzinger, S. P., and C. Kroeze. 1998. Global distribution of nitrous oxide production and N inputs in freshwater and coastal marine systems. Glob. Biogeochem. Cycles 12:93–113. Shalovenkov, N. 2000. Scales of ecological processes and anthropogeneous loads on the coastal ecosystems of the Black Sea. Estuar. Coast. Shelf Sci. 50:11–16. Shiganova, T. A., and 8 others. 2000. Population development of the invader ctenophore Mnemiopsis leidyi, in the Black Sea and in other seas of the Mediterranean basin. Mar. Biol. 139:431–445. Valiela, I., and 11 others. 1992. Couplings of watersheds and coastal waters: Sources and consequences of nutrient enrichment in Waquoit Bay, Massachusetts. Estuaries 15:443–457. Van Breemen, N., and 12 others. 2002. Where did all the nitrogen go? Fate of nitrogen inputs to large watersheds in the northeastern U.S.A. Biogeochemistry 57/58:267–293. Weisse, T., M.-T. Gomoiu, U. Scheffel, and F. Brodrecht. 2002. Biomass and composition of the comb jelly Mnemiopsis sp. in the north-western Black Sea during spring 1997 and summer 1995. Estuar. Coast. Shelf Sci. 54:423–437. Zaika, B. E. 1990. Change in macrobenthic populations in the Black Sea with depth (50–200 m). Doc. Acad. Sci. Ukrainian SSR. Ser. B 11:68–71. Zaitzev, Y., and V. Mamaev. 1997. Marine Biological Diversity in the Black Sea: A Study of Change and Decline. Black Sea Environmental Series, Vol. 3. United Nations Publications, New York, 208 pp.

Chapter 2 Atmospheric-driven changes

Appearance of coral reefs off Easter Island, before (March 1999) and after (March 2000) a major bleaching event. From Wellington et al. (2001), courtesy of the authors.

ATMOSPHERIC-DRIVEN CHANGES

A case history: coral bleaching1 The coral reefs of the world have been subject to substantial alterations throughout their history. Mass mortality of corals has been reported for many widely dispersed places since the 1870s, and has increased in more recent times. Some of the mechanisms that have altered tropical reefs are biological. In 1983–1984, a pathogen decimated populations of the long-spined sea urchin, a common species in Caribbean and Atlantic reefs. The sea urchin populations were virtually eliminated from their range, and where they survived abundance was only 2–7% of that before the disease epidemic. Algae previously eaten by the urchins then overgrew corals (Knowlton 2001). After 20 years or so, the small populations of remaining urchins produced sufﬁcient recruits for urchin density to return to prior numbers. The denser urchin populations consumed the macroalgal cover, and allowed regrowth of the corals (Edmunds & Carpenter 2001). Another widely reported case involved population outbreaks of the crown-of-thorns, a starﬁsh that feeds directly on corals. The outbreaks were ﬁrst recorded in the Indo-Paciﬁc in the early 1900s, and peaked in 1970. The crown-of-thorns outbreaks were regarded with alarm as they occurred across many widely dispersed reef sites, with concern for the integrity of reefs that appeared overrun by starﬁsh. Within a decade, however, the starﬁsh outbreak subsided and the reefs recovered. Weather-mediated disturbances—hydrographic disturbances, storms, rain, earthquakes, and unusual drops in sea level—have also altered reefs. For example, an unexplained inﬂux of upwelled water, 10°C colder than normal, and with low oxygen content, killed 60–90% of corals, eliminated sea cucumbers, and decimated sand dollars, snails, worms, and sponges from the reef off Mocorroy National Park, Venezuela during

1

This section is largely derived from Glynn (1993), HoeghGuldberg (1999), Pittock (1999), Goreau et al. (2000), Wellington et al. (2001), and Done (2000, at http://www.aims.gov.au/pages/ research/reffs/wcr-02.html).

23

January 1996 (Laboy-Nieves et al. 2001).2 Storms also have intermittent and sometimes severe effects. Hurricane Mitch, for instance, lowered the density of newly recruited coral heads by about 80% in Glovers Atoll in Belize (Mumby 1999); and records kept from 1967 to 1992 showed that up to 100% of the coral cover was removed by recurrent cyclones in Heron Island on the Great Barrier Reef in Australia (Hughes & Connell 1999). Various biological and weather phenomena3 have affected coral reefs, but, in general, the resulting losses of corals and associated organisms have come and gone, with reasonable recovery of the reefs. Coral cover has been restored after some time, and the perturbations tend to be limited to local spatial scales. Beginning in 1963 a new level of concern about the status of coral reefs was raised by widespread records of bleaching of corals and associated fauna (see the chapter frontispiece). Bleaching involves the expulsion of the symbiotic dinoﬂagellate algae (called zooxanthellae), or the pigments from these symbionts, from the cells within the coral polyps. These symbionts furnish most of the ﬁxed carbon that supports the coral’s metabolism. Bleaching became even more common during the 1980s and 1990s, and culminated in widespread global occurrences during 1998, the year of the most pronounced El Niño–Southern Oscillation4 (ENSO) on record. Bleaching spread 2

Corals seem more susceptible to climate-driven perturbations than other forms of life. In the Venezuelan reef example, seagrasses, seaweeds, and mangroves in nearby areas were unaffected by warmed water (Laboy-Nieves et al. 2001). 3 To these agents of change we need to add human interventions as another major impact on tropical reefs. For example, in the Philippines destructive ﬁshing methods have dominated the losses of coral reef habitat during 25 years of efforts in coral reef conservation (White & Vogt 2000), markedly outdoing perturbations from climatic change and outbreaks of populations. Recovery from anthropogenic disturbances may or may not occur. Such aspects of habitat destruction are discussed in Chapter 6. 4 The original meaning of “El Niño” referred to a warm water ﬂow along the coasts of Peru and Ecuador, often during the year’s end, hence the connection with the Nativity. El Niño events are associated with long-distance circulation patterns in the Indian and Paciﬁc Oceans, and are collectively called the El Niño–Southern Oscillation (ENSO) phenomenon. ENSO events involve the weakening of trade winds, which allows warm surface waters to move from Indonesia toward South America. The warm water current

24

CHAPTER 2

No bleaching

1–10%

10–30%

Inshore

13

20

12

30

25

Offshore

72

14

9

5

0

30–60% > 60%

30 Sea surface temperature (°C)

Table 2.1 Percentage of 654 sites within the Great Barrier Reef that showed different intensities of bleaching. Data are shown for sites facing landward (“inshore”) and seaward (“offshore”) separately. Data from Berkelmans and Oliver (1999).

29 28 27 26

prevents the upwelling of colder deeper water near shore, and impairs its delivery of nutrients to the nutrient-poor upper waters. The lack of nutrients lowers the production of phytoplankton, and affects the entire food web. The key anchovy stocks become depleted, causing economic harm to the large ﬁshing ﬂeet, and starving the sea bird fauna. ENSO also has powerful inﬂuences on weather patterns throughout South America and beyond, as we will see below. 5 The highly heterogeneous nature of bleaching damage is common. In the Seychelles, for example, the ratios of mean percentage bleached to the associated standard deviations ranged from 1 to 3 (Spencer et al. 2000). Such observations are highly heterogeneous, with frequency distributions with a few high values and many more low values.

1999

1997

1995

1993

1991

1989

1987

1985

1983

from south to north, from Madagascar to the Persian Gulf, across the Indian Ocean, through Indonesia, to the Great Barrier Reef of Australia, across the Paciﬁc, and, more easterly, from Brazil to Florida. Bleaching events have taken place as water warms, whether under the inﬂuence of ENSO or any other oceanographic conditions (Wilkinson et al. 1999; Wellington et al. 2001). Although bleaching was widespread, the degree or intensity of bleaching, however, was quite variable at regional (Goreau et al. 2000) and local (Berkelmans & Oliver 1999) scales. For example, in the Great Barrier Reef of Australia, the intensity of bleaching during the 1998 episode varied greatly from place to place; the proportion of corals bleached in different speciﬁc sites varied from 0 to 90% (Table 2.1). Such local differences in bleaching intensity no doubt derived from differences in coral species composition, duration of the exposure, and local details of topography and habitats.5 In general, it has been difﬁcult to estimate intensity and extent of bleaching,

1981

25

Year

Figure 2.1 Weekly sea surface temperature for Tahiti, 1981–1999. The arrows show the timing of reported coral bleaching events and the horizontal line shows the threshold temperature above which bleaching is thought to take place. From HoeghGuldberg (1999).

mortality of corals, further effects on reef communities, and recovery (Glynn 1993; Hoegh-Guldberg 1999; Goreau et al. 2000; Spencer et al. 2000). There has been much speculation as to the causes of the widespread bleaching, but most evidence points to atmospheric changes that resulted in elevated seawater temperatures.6 Correlational evidence includes many sets of measurements of sea surface temperatures, such as those from around Tahiti, which show that recorded episodes of bleaching coincided with water temperatures above certain thresholds; about 29°C in the case of the Tahitian reef (Fig. 2.1). Similar comparisons between bleaching episodes and seawater temperatures in several sites show that the “threshold” temperatures vary over a span of only a few degrees.7 Experi6

Precipitation is also affected by climate change. Seawater salinities in tropical areas (except where affected by estuaries) may range from 32 to 40‰ (Veron 1986). If precipitation increases, salinity may become low enough to affect corals. 7 Coral reefs have a deﬁned range (Kleypas et al. 1999b), occurring between 30°S and 30°N, a range of latitude that spans a seawater temperature range of roughly 16–29°C. Reefs have difﬁculty forming below 16–18°C, probably because of lowered light and carbonate availability in sea water at the higher latitudes.

ATMOSPHERIC-DRIVEN CHANGES

mental evidence showed that a warmer temperature was detrimental to corals (Glynn & D’Croz 1990; Jockiel & Coles 1990). Experiments with coral heads in sea water whose temperature was manipulated, showed that 5-day exposure to 1–3°C above normal (29°C) reduced the density of zooxanthellae in 50% of the heads of three species of corals (Berkelmans & Willis 1999). Corals occur in environments that have a long history of little variation in temperature. Seawater temperatures in tropical oceans, on average, have varied by less than 2°C for at least 18,000 years (Thunnell et al. 1994), and hence corals evolved under relatively benign conditions. Because of their geological history, corals are highly sensitive to current extra-normal changes in temperature. The remarkably low variation in high temperature thresholds for coral bleaching made it possible to develop models, based on water temperature predicted from remotely sensed atmospheric data, that have done well in forecasting bleaching episodes (Goreau & Hayes 1994). The models successfully anticipated the 1998 bleaching in the Great Barrier Reef, and had their prediction conﬁrmed by ﬁeld observers soon afterwards. Elevated temperatures lower photosynthetic activity in corals,8 both because the number of symbionts is lowered and because of lower photosynthetic rates of the remaining zooxanthellae.9 The model results, plus the physiological studies, showed that temperature

8

The effect of raised temperature was ﬁrst noted by Coles and Jockiel (1977) in corals exposed to heated sea water released by a power plant in Hawaii. This, as in so many other examples, shows that some of the most crucial basic ﬁndings often emerge in welldone applied studies. 9 Corals and their symbionts usually live in waters near their upper temperature range. As temperatures increase, the ﬂow of energy to the dark reaction of photosynthesis is interrupted, and the light energy collected by the light reaction is passed to oxygen rather than to the dark reaction. The energy-laden, active “toxic” oxygen then denatures the proteins that make up the photosynthetic apparatus of the zooxanthellae (Lesser 1997; Hoegh-Guldberg 1999). Conﬁrmation of these mechanisms was obtained by experimental treatment with antioxidants, which not only returned photosynthetic rates to normal in the symbionts, but also prevented bleaching (Lesser 1997). The mechanisms thus identiﬁed lead to the conclusion that light driving photosynthesis becomes the underlying basis for bleaching as temperatures rise.

25

increases were involved in the onset of the coral bleaching, which affected such widespread geographic regions. During the last half of the 20th century bleaching episodes were linked to unusual events, such as ENSO phenomena, that led to unusually high seawater temperatures. Of the six major recent mass bleaching episodes, the 1998 event was the most severe. Massive corals aged over 700 years died as a result of this event. Although this might have been circumstantial, the death of these longevous corals suggests that the 1998 conditions were extreme relative to changes seen across the last 3,000 years (R. Aronson, in van Woesik 2001). All predictions about future temperatures in the tropics suggest that there will be considerable increases (McCarthy et al. 2001). As tropical seawater temperatures rise further through the 21st century, temperatures above bleaching thresholds may become more commonplace (Fig. 2.2) (van Woesik 2001). Under these circumstances, the thermal threshold of corals might be surpassed routinely in decades to come. Annual bleachings already occur in the southern Red Sea, where seawater temperatures exceed 33°C every summer (M. Guillaume, in van Woesik 2001). It is unclear what will happen to coral reefs in the future, but some adaptation by the symbionts to higher temperatures might be expected (Rowan et al. 1997). The possibility of acclimation is evident in the continued presence of corals in the Red Sea, where warm water is common. In addition, some acclimation to hot temperatures may be suggested in ﬁndings by P. Glynn (in van Woesik 2001). Glynn compared damage from the 1982 and 1998 ENSO bleachings, and found less impact from the more recent bleaching, even though the warming was more marked. Studies of fossil corals show that corals managed to survive large-scale climatic shifts during the history of the earth. For example, coral reefs managed to survive temperatures thought to exceed the threshold some 6,170 years ago (M. Gagan, in van Woesik 2001). Even as temperatures changed, local environmental details, such as terrestrial inputs, reef hydrography and

26

CHAPTER 2

TAHITI 30

Sea surface temperature (°C)

28 26 34

PHUKET

32 30 28

32

SOUTH COAST OF JAMAICA

30 28 26 1860

1900

1950

2000

2050

Year

topography, weather, and so on, may have played as large a role as glacial/interglacial climatic changes in determining what species survived (Pandolﬁ 1999). Unfortunately, today’s reefs are exposed to pressure from adverse global changes as well as widespread local habitat destruction, a combination that has the potential for major alterations to reef assemblages. This alarm has to be tempered with the realization that although a large percentage of reefs have been degraded, there are no known extinctions of modern coral species (Pandolﬁ 1999). Moreover, coral reefs have survived—not without changes—across geological time, even as signiﬁcant shifts in climate dominated the earth. And, as discussed above, mass mortalities, owing to a variety of causes, have been almost de rigeur in coral reefs, and there is evidence that though sensitive, these assemblages of organisms have managed to recover from drastic alterations. In addition to rising temperatures (and changes in precipitation regimes), changes in the atmosphere have also increased the levels

2100

Figure 2.2 Sea surface temperatures for three tropical coastal areas, 1860–2100. The temperatures were generated using a global atmosphere/ocean/ice model (Roeckner et al. 1996), and forced using greenhouse gas concentrations from IPCC scenario IS92a (IPCC 1992). ENSO inﬂuences are included. The horizontal lines show the thermal threshold for bleaching at each site. From Hoegh-Guldberg (1999).

of ultraviolet (UV) radiation impinging on the surface of the earth.10 UV radiation is to some extent attenuated as it penetrates into the seawater column. The degree of attenuation depends on water quality; in richer coastal waters, the attenuation of UV radiation may reach 90% at depths of 4 m (Whitehead et al. 2000). In clear oceanic waters 90% absorption of UV radiation might not take place until depths of 12–20 m (Whitehead et al. 2000). In any case, whatever fraction reaches organisms might be injurious 10

Ozone in the stratosphere absorbs most of the UV radiation (both UVA, 320–400 nm, and UVB, 280–320 nm) arriving from the sun. In recent decades the concentrations of stratospheric ozone have been depleted by chemical reactions with compounds produced by industry, mainly chloroﬂuorocarbons used as propellants in spray cans and other uses. The decreases have been most pronounced over the poles (World Meteorological Organization 1995, 2003), especially the Antarctic. At mid-latitudes, the decreases in ozone concentrations are less marked, perhaps 4–5% per decade. As a result of the agreement signed in the 1987 Montreal Protocol, uses of these compounds are regulated, and it is anticipated that the loss of ozone will decrease. These circumstances suggest that the tropics, where corals occur, might be least affected by UVB radiation increases, and that future trends seem unlikely to worsen (Smith & Buddemeier 1992; Herman et al. 1996).

ATMOSPHERIC-DRIVEN CHANGES

to intracellular processes: experimental manipulation of UVB can cause bleaching in corals (Gleason & Wellington 1993), but there is scant evidence that such effects are important in the ﬁeld (Hoegh-Guldberg 1999).11 The coral bleaching case history highlights that current atmospheric changes are creating coastal change of a broad nature in a major coastal environment, and doing so on a global scale. This case history also shows that, to some degree, such changes have taken place across geological history, and that although mechanisms exist by which the effects might be to some extent remediated, recent changes seem to exceed historical precedents. Adaptation to temperature increases in corals and symbionts might provide some margin for survival, induced radiation-absorbing amino acids might prevent cellular damage from UV, and the patchy nature of the effects might allow recolonization with propagules from patches less affected by the bleaching. There is no question that warm temperatures—clearly linked to anthropogenic causes— are tied to bleaching, and that to a certain degree bleaching leads to coral mortality. Corals recover from bleaching, but recovery may take years to decades. Local phenomena, such as storms or disease—perhaps “natural”, perhaps not—also create similar damage, which might be as slowly repaired. In any case, the example of coral bleaching demonstrates that climate-driven factors can force substantial environmental change in coastal ecosystems. Below we explore major climate-driven factors—mainly warming and increases in UV radiation—and their effects, all

11

Corals and their symbionts to some degree may acclimate to the changes in UV radiation levels in several ways (Roy 2000). Organisms might respond behaviorally by moving deeper. They might respond physiologically by altering concentrations of lightabsorbing pigments such as xanthophyll, by increasing concentrations of mycosporine-like amino acids that absorb damaging UV radiation (Dunlap & Chalker 1986), and by a variety of active oxygen scavenging systems (Hoegh-Guldberg 1999). The very fact that mechanisms are present to absorb UVB makes the point that this radiation indeed constitutes a biological problem, but also that evolution has furnished tools to adjust to UVB, at least to some extent.

27

created by changes in the chemistry of the atmosphere.12

Global changes in the chemical composition of the atmosphere The balance between heating and cooling of the earth’s atmosphere is in large measure determined by the composition of gases [in particular carbon dioxide (CO2), methane (CH4), nitrous oxide (N2O), ozone (O3), chloroﬂuorocarbons, and several other gases], and by the concentrations of particulates and aerosols.13 Greenhouse gas concentrations The ability of water vapor, carbon dioxide, methane, nitrous oxide, ozone, and chloroﬂuorocarbons to absorb energy, and hence prevent radiative heat losses from the atmosphere, has suggested the use of the term “greenhouse gases” for these compounds. Water vapor has received less attention than the other greenhouse gases, whose concentrations have increased signiﬁcantly since the industrial revolution (Table 2.2).14 12

A multitude of other factors and effects, above and beyond warming and UV radiation, are tied in some fashion to atmospheric processes, as will become evident through many of the following chapters. For example, atmospheric warming is part of the forcing that leads to rising sea levels, the topic of the next chapter. Sea level rise is not covered in this chapter because there are some other important variables that also are involved in altering the level of the sea, and hence the topic was more comfortably dealt with in a separate chapter. Changes in the use of fresh water, and alterations to sediment transport, are another two factors related to climate; they too are dealt with in separate chapters. 13 Assessment of the relative contributions to temperature regulation by these various gases is difﬁcult for technical reasons, particularly the partial overlap in absorption by the different gases, the interference of clouds, and differences in absorption of radiation of different wavelengths (Kiehl & Trenberth 1997). 14 Ruddiman (2003) argued that the human inﬂuence on the atmosphere began considerably earlier, some 5,000–8,000 years ago. Forest clearing, rice growing, and other agricultural practices could have released CH4 and CO2, and could account for unexplained changes in atmospheric gas composition. These greenhouse gas emissions may have been enough to prompt sufﬁcient warming as to halt glaciations. Later, the spread of bubonic plague could have then led to such substantial abandonment of cultivated land that a cooler period resulted, known as the Little Ice Age (1300–1900 AD).

28

CHAPTER 2

Table 2.2 Concentrations of greenhouse gases in pre-industrial times and at the end of the 20th century. Data summarized from Wigley (1999) from measurements in gases trapped within ice cores from Antarctica. Concentrations (parts per billion by volume) Gases

Pre-industrial

CO2

270–280

CH4

700

1,700

N2O

270

310

Chlorofluorocarbons

0

End of 20th century 360

Substantial concentrations

(referred to as “forcing”16), that is attributable to the different gases varies. Toward the end of the 20th century, on a clear day, about 60% of the “greenhouse” effect derived from water vapor, about 25% from carbon dioxide, about 8% from ozone, and the rest from trace gases such as nitrous oxide and methane (Kiehl & Trenberth 1997). The relative forcing effects of different greenhouse gases, at the longer time scale of centuries (Table 2.3), were evaluated by models that estimate heat accumulation expected from increases in concentrations. Clearly, carbon dioxide is the major greenhouse gas contributing to higher temperatures, followed by methane; other gases made smaller contributions. Aerosols

Since 1750, concentrations of carbon dioxide have increased by about 31%, methane by 151%, and nitrous oxide by 17%; ozone has increased in the trophosphere but decreased in the stratosphere; and chloroﬂuorocarbons increased over the last half of the 20th century (Watson et al. 2001).15 Concentrations of UV-absorbing ozone in the stratosphere have diminished, most likely as a result of ozone destruction by the chloroﬂuorocarbons. The propensity to retention of heat, and hence to warming of the atmosphere or oceans

15

There is little question that the increases in greenhouse gases are from anthropogenic intervention (Houghton et al. 1996). The increases in CO2 concentrations derive from the combustion of fossil fuels (coal, oil, and gas) and biomass burning, slash-and-burn agricultural practices, and the use of dung as fuel [30–50% of the earth’s land surface has been altered by human activities (Vitousek et al. 1997), and about 20% of the earth’s forests have been converted to other land covers (McNeill 2000)]. The principal humanconnected sources of CH4 are rice paddies, livestock ﬂatulence, and land-ﬁll emissions. Plus there are natural releases from anoxic sediments. Use of fertilizers and subsequent microbial transformations result in N2O emissions from cultivated land. In addition, human perturbation of habitats alter the emission and consumption of these gases in natural environments. Increases in O3 are from combustion sources, and losses are from the chemical action of the chloroﬂuorocarbons used as propellants and released to the atmosphere. The chloroﬂuorocarbons are exclusively human-made. A number of other gases (carbon monoxide, nitrogen oxides, butane, and propane) might also react with trophospheric ozone, a powerful greenhouse gas.

The warming that forcing made possible by increased concentrations of greenhouse gases is to some extent countered by the presence of particles referred to as aerosols (Table 2.3). Aerosols have several effects on climate [discussed in some detail by Houghton et al. (1996)], but the main direct effect is to scatter part of solar radiation back to space, which tends to cool the earth’s atmosphere and surface. Aerosols may be of human or natural origin. Industrial release of sulfur dioxide and carbon particulates, burning of fuels and biomass, dust from poor agricultural practices, and desertiﬁcation are among the contributions under human inﬂuence. These anthropogenic sources often affect relatively restricted local areas, because particles remain at low altitudes, and hence do not remain in the atmosphere for very long. Natural particulates are also important. Natural aerosols form from various sources, including oxidation of dimethyl sulﬁde (DMS), a volatile product of marine producers. There is a likely connection between DMS release by phytoplankton and atmospheric cooling, but the controls and interactions are not well understood (Denman et al. 1996). Volcanoes are another, less important, 16

The “forcing” is intended to describe the relative inﬂuence of the changes in atmospheric composition on the balance between incoming solar short-wave radiation and the long-wave radiation reﬂected back into space.

ATMOSPHERIC-DRIVEN CHANGES

29

Table 2.3 Contribution to mean global atmospheric heat forcing, in watts per m2, from 1765 to 1990, by different gases and particulates. Data summarized from Wigley (1999). Forcing towards warmer temperatures Greenhouse gases CO2 CH4 N2O Halocarbons Trophospheric ozone

Forcing towards cooler temperatures

1.29 0.44 0.13 0.11 0.40

Aerosols Direct SO4 aerosols Indirect SO4 aerosols Biomass aerosols Totals

−0.30 −0.80 −0.20

Net total forcing

source of natural particulates. For example, ash released by the Mt Pinatubo eruption in 1991 in the Philippines led to a −0.5°C cooling of the global atmosphere for about a year after the eruption (Houghton et al. 1996). The ash from Pinatubo reached high altitudes and was carried around the world by winds.17 Water droplets in the clouds also alter heat budgets, but it is not clear whether cloud cover has increased or decreased, in part because there is such great spatial and temporal variation in cloud cover. Such major intermittent but widespread events can therefore alter the time course of rising temperatures. In general, the role of aerosols, and clouds in particular, in climatic forcing has been difﬁcult to assess, in part because of the high variation across geography and time. Although negative forcing by aerosols provided a signiﬁcant counter to greenhouse gases, on balance, forcing has remained positive after the Industrial Revolution (Table 2.3). These calculations suggest that we should have expected 17

−1.3

2.37

As one example, the eruption of Mt Pinatubo led to a cold temperature anomaly throughout the Middle East during the winter of 1992. The cooler temperatures allowed a deeper mixing of surface waters in the Gulf of Eilat, which brought nutrient-rich water up to the surface and promoted extensive algal blooms. These covered the reef and caused extensive coral death (Genin et al. 1995).

+1.07

atmospheric warming. Throughout the 1765–1990 period, of course, there were changes in the intensity of solar radiation, which prompts the question whether the changes caused by atmospheric gases are of the same magnitude as the changes to be expected from alterations in radiant energy from the sun. Increased ultraviolet radiation The concentration of stratospheric ozone is important not only because of greenhouse effects, but also because this gas is the key absorber of UV radiation. Concentrations of ozone, as well as the resulting penetration of UV radiation, have changed across the years. Ozone concentrations in the trophosphere increased, but stratospheric ozone decreased signiﬁcantly toward the end of the 20th century (IPCC 2001). The concern with such changes is that beneath the areas where there is greatest depletion of ozone, UV radiation may reach levels hazardous to organisms. The amount of UV radiation reaching the surface of the earth is spatially variable, being lowest at lower latitudes and largest over the polar zones, giving rise to the stratospheric “ozone holes” often mentioned in the press.

30

CHAPTER 2

Climate warming18 The temperature of the earth has changed impressively across the millennia. Researchers in many disciplines have made ingenious use of a diverse quiver of biogeochemical evidence—tree rings, ice cores, lake sediments, corals—that have been interpretable as proxies for temperature regimes, and of written records of weather and climate changes. From the aggregate evidence, it has been possible to reconstruct temperature regimes of reasonable accuracy for thousands of years before the present, particularly for the Northern Hemisphere. For the previous millennium, the reconstructed records show a relatively warm start, peaking in the Medieval Warm Period (900–1200 AD), cooling steadily toward the Little Ice Age (about 1550–1700), followed by a somewhat milder period between 1700 and 1800, and a renewed cooling from 1800 to 1900. These shifts are likely to have been prompted by changes in solar radiation rates, as well as by volcanic action contributing aerosols to the earth’s atmosphere.19 Atmospheric temperature records are to some extent available from the late 1880s (Fig. 2.3). Reconstruction of the temperature regimes 18

Much of this section is from Houghton et al. (1996), de Mora et al. (2000), Jones et al. (2001), and McCarthy et al. (2001). 19 There is an accumulating archaeological record of cultural change that shows remarkable coincidences with shifts in atmospheric conditions (deMenocal 2001). As just one example, the collapse of the Mayan civilization between 750–790 AD and the last dated Mayan stela in 909 AD have been variously attributed to overpopulation, overexploitation of natural resources, intercity warfare, and social upheaval. It also turns out that the period 800–1000 AD was one of severe drought, as made evident by geochemical data from lake sediments that are known proxies for warmth and dryness (Hodell et al. 2001). Since Mayan cities depended heavily on irrigated ﬁelds and water-dependent agricultural practices, such droughts could have threatened the intensive food production system needed to support the urban centers. Data from Peruvian ice cores also show a peak in wind-carried particles during that time, so that the climatic-driven change was geographically extensive, and, from all evidence, was likely involved in the collapse of a major civilization. Although this coincidence in time is not tantamount to causality, the Maya collapse and other examples do suggest that climatic changes have occurred repeatedly in the past, and that it may not have been unusual for human societies and their environments to be considerably affected by the changes.

across the globe, based on actual temperatures and supported by the biogeochemical proxy measurements, show that the 20th century has seen continuing changes in the condition of the world’s atmosphere. The contrast, though, is that in the latter part of the 20th century the atmosphere of the earth has warmed in an unusual fashion, considerably faster than seen at any time during the previous millennium. A time series of global mean temperatures (expressed as deviations from the mean of the entire record from 1880 to 1997) (Fig. 2.3) show that during the 20th century we experienced relatively cooler global temperatures up to the 1940s, followed by increasing warming through the remainder of the century, particularly after the 1980s. The aggregate temperature estimates derived from tree rings, ice cores, lake sediments, corals, and documentary sources leave little room to doubt that the last few decades of the millennium have been signiﬁcantly warmer than any others, at least in the Northern Hemisphere (Jones et al. 2001). The most recent 30-year period has been the warmest of the millennium. Temperatures during recent decades were the warmest since at least 1400. The 1990s were the warmest decade in the century and a half during which meteorological records have been kept; 1990 and 1995 were among the warmest, and 1998 was the warmest year on record. For 16 consecutive months from May 1997 to September 1998, each month broke the previous record of high temperatures (Karl et al. 2000). The year 2003 was the third warmest recorded in 150 years of meteorological records for the entire world, according to the World Meteorological Association (Boston Globe, Dec. 17, 2003). Although there were large local disparities in temperatures across different areas of the world (for example, in the East Coast of the USA, January 2000 was cooler than average, by somewhat less than 2°C), the winter of 2003–2004 was the warmest recorded in the Northern Hemisphere. Additional compelling corroboration of general warming comes from the substantial thinning and retreat evident in most mountain glaciers during the 20th century (Fig. 2.4). The area of the

ATMOSPHERIC-DRIVEN CHANGES

31

0.8

Mean temperature anomaly (°C)

0.6

0.4

0.2

0

−0.2

−0.4 1880

1900

1920

1940

1960

1980

2000

Year

Figure 2.3 Changes in mean global atmospheric temperatures for each year from 1880 to 2000. The data are expressed as the difference, in °C, between one year’s temperature relative to the mean for all the years for the period. Data from the National Climatic Data Center, NESDIS, National Oceanic and Atmospheric Administration.

ice cap on top of Mt Kilimanjaro, for example, diminished by 82% between 1912 and 2000; if melting continues at current rates, there will be no ice remaining on Kilimanjaro by about 2015 (Trenberth 2001). The extent of snow cover in the Northern Hemisphere has been lower than the average values for the period 1974–1994 since 1988. Added to the other lines of evidence, some of which are included in the box on p. 33, the widespread glacial retreat and lower snow cover strongly imply that global warming is in effect. The weight of the evidence available compellingly leads to the conclusion that extraordinary warming is occurring on a global basis. But climate has always changed to some degree throughout the history of the earth. What evid-

ence is there that human activities are responsible for recent warming, and does the recent warming exceed the variation that might be expected from non-human causes? The best way we have at present to examine the relative importance of human greenhouse gas contributions versus the “natural” forcers of climate change is to apply models. These models use the best available information on relevant atmospheric, land, and ocean processes that affect atmospheric heat content and temperature, to produce best estimates of variables such as departures in atmospheric temperature from some agreed-upon mean. To get a notion of how appropriate these model predictions may be, one approach is to compare model predictions

32

CHAPTER 2

20 Storbreen

Cumulative balance (m)

0

Hintereisferner

−20 Rhône Storglaciären

−40

Sarennes −60 South Cascade −80

−100 1890

1910

1930 1950 Year

1970

1990

Figure 2.4 Time course of the mean thickness of mountain glaciers (expressed relative to conditions in 1890), from 1890 to 1990. The glaciers included are among the few with such long records: Storbreen (Norway), Hintereisferner (Austria), Rhône (Switzerland), Storglaciären (Sweden), Sarennes (France), and South Cascade (United States). From Warrick et al. (1996).

to empirically reconstructed past records of atmospheric heat budgets and temperatures. The observed records are reconstructed from a number of proxies, such as gas concentrations in ice cores, isotopic ratios in trees, corals, and so on, as well as actual temperatures measured more recently. Reconstructed atmospheric temperature records are shown in Fig. 2.5 as the black lines that zig-zag up and down across the years; the model predictions of the same variables are included as the gray areas (the span of the areas show the relative uncertainty of the model predictions). If the model simulations predicted temperature anomalies since 1860 based only on “natural” forcing (from changes in solar

radiation and volcanic activity), there was a poor match between the measured record and the model predictions (Fig. 2.5 top left). Model runs that considered only changes from differences in human greenhouse gas emissions also produced unsatisfactory ﬁts to the measured temperature record (Fig. 2.5 top right). But model runs that included both the natural and anthropogenic drivers came up with predicted time courses of temperature anomalies that provided a far improved match to the measured estimates of atmosphere heat content and temperatures (Fig. 2.5 bottom).20 Similar modeling exercises have been done with regard to heat content and temperatures of the oceans (Barnett et al. 2001; Levitus et al. 2001). Warmer atmospheric temperatures have increased the heat content of sea water during the 20th century (Fig. 2.6) (Levitus et al. 2000). The warming was detectable during 1920–1940, followed by a cooling, and then a steeper rise after the 1970s, following the pattern of atmospheric temperatures (Fig. 2.5). The warming extends from near-surface waters to remarkable depths of thousands of meters.21 Although the warming has not been the same everywhere in the oceans, warming was evident in the Atlantic, Indian, and Paciﬁc Oceans, as well as in the aggregate world ocean (Fig. 2.6). The time course of warming for entire oceans is also evident in many local sites. The seawater

20

We may carp that the ﬁt to the empirical data is still far from perfect. We need to recall, however, that: i) the models try to capture the major features, but cannot include all the multiple variables affecting heat content and temperatures; and ii) the observed data themselves are only spliced reconstructions from various data sources, and even when actually measured as temperatures they have a degree of uncertainty. To me, it is remarkable that the models do as well as they do, given the nature of the information available. 21 The increased heat content of deeper water layers anteceded detectable increases in temperature of surface layers. This seemingly implausible situation, where warmer water underlies colder layers, may have occurred because in sea water, salinity as well as temperature determines density. Levitus et al. (2000) argue that relatively warm but salty water could be convected down from relatively small regions of the sea surface, and be transported and broadly distributed at depth. In any case, the heat-holding capacity of the oceans are some 30-fold that of the atmosphere, so that the seas are providing a large buffer against warming of the earth’s surface.

ATMOSPHERIC-DRIVEN CHANGES

Miscellaneous observations that suggest effects of warming conditions • • In 1984 visitors to the Portage Glacier Visitor’s Center, near Anchorage, Alaska, could see a dramatic view of the ice-blue glacier through its ﬂoor-to-ceiling windows. By 1998 tourists only saw a lake with small icebergs calved from the glacier, which is now quite a distance away and out of sight (Boston Globe, Oct. 5, 1998). • Sea ice in the Arctic has decreased in area and thickness (Vinnikov et al. 1999). • Coastal ice sheets in Greenland thinned during the 1990s (Thomas 2001). • Permafrost thawing has increased in polar latitudes and high mountains (Bradley 2001). • Ice breakup in the Tanana River of Alaska is now occurring 5.5 days earlier in the spring than in 1917, according to records of the Nenana Ice Classic, an annual guessing contest as to the date of thawing of the river (Sagarin & Micheli 2001). • The ﬁrst spring blooming of lilacs and honeysuckles, and the ﬁrst major pulse of snowmelt in mountain streams, have occurred earlier and paralleled the increases in temperature in the western USA during 50 years of records (Cayan et al. 2001). • Most plant species in Mediterranean climates now unfold 16 days earlier than 50 years ago (Peñuelas & Filella 2001). • Trembling poplar trees have shifted to bloom 26 days earlier in western Canada during the previous century (Peñuelas & Filella 2001). • The tree limit in the Scandes Mountains of Sweden has risen by 30–165 m during the 1990s, the highest record in about 4,000 years (Kullman 2001). • Plants turn to fall colors 5 days later, on average, in sites ranging from Scandinavia to Macedonia, than they did 30 years ago (Peñuelas & Filella 2001). • Lake Champlain, in New York state, now freezes over an average of 8 days later than 100 years ago; it failed to freeze over 16

•

•

•

•

•

•

•

•

•

•

•

times between 1815 and 1950, and 25 times between 1950 and 2004 (Rock 2004). Over the past 50 years snowfall dropped by 58 cm in Keene and 43 cm in Berlin, New Hampshire (Rock 2004). Woolly adelgids, an insect pest of hemlocks, has spread northward in eastern North America during the last 10 years, threatening the survival of these northern trees (Rock 2004). Aphids in the UK have advanced their life history stages by 3–6 days over the last 25 years (Peñuelas & Filella 2001). Butterﬂies in northeast Spain now appear 11 days earlier than in 1952 (Peñuelas & Filella 2001). Frogs begin their spring chorus in New York state about 10 days earlier than in 1900 (Peñuelas & Filella 2001). The collection of sap for the production of maple syrup started in March before the 1980s, in more recent years tapping sap has been done in February (Boston Globe, March 22, 2004). Overwintering bird species with southern geographic distributions have became more numerous in coastal Cape Cod, Massachusetts (Valiela & Bowen 2003). Warm-water species of zooplankton have spread northward in the North Sea during the last 30 years, replacing larger and more numerous northern zooplankton and yielding less food for larval cod (Beaugrand et al. 2003). Species of many different kinds of land animals have expanded northward across the world in recent decades (Parmesan & Yohe 2003). The length of the phytoplankton active growing season has increased in areas with warming water from 1948 to 1995 (Peñuelas & Filella 2001). During the 1990s, seawater temperatures off Woods Hole, MA, averaged 1.2°C warmer than they were between 1890 and 1970 (Nixon et al. 2004). And so on . . .

33

Temperature anomalies (°C)

34

CHAPTER 2

NATURAL 1.0

ANTHROPOGENIC 1.0

0.5

0.5

0.0

0.0

−0.5

−0.5

−1.0 1850

1900

1950

2000

−1.0 1850

1900

1950

2000

Year

Temperature anomalies (°C)

ALL FORCINGS 1.0 Model Observations

0.5

0.0 −0.5 −1.0 1850

1900

1950

2000

Year

Figure 2.5 Comparisons of modeled and measured (described as “observations”) estimates of temperature anomalies since 1860. Top left: modeled including forcing by solar and volcano inﬂuences only; top right: modeled using anthropogenic greenhouse gases only; and bottom: using both natural and anthropogenic inﬂuences. From McCarthy et al. (2001); summarized from various sources.

temperatures in a site in the central Indian Ocean, and another in the Arabian Gulf, show warmer mean temperatures as the 20th century came to a close (Fig. 2.7). Clearly, there were large seasonal and interannual ﬂuctuations, which make the trends hard to discern. The trends in the entire data sets, however, in these two sites, and elsewhere, indicate a consistent warming across recent decades. The warming of sea water evident at global and local scales corroborates ﬁndings that there has been an excess entry of heat into the earth’s atmosphere (Kiehl & Trenberth 1997), and that

most of the excess heat has accumulated in the oceans (Hansen et al. 1998; Levitus et al. 2001). Here again, the conclusion was that the observed heating could not be explained unless there was forcing from emission of anthropogenic greenhouse gases (Levitus et al. 2001). The timing of the measured warming, during the latter half of the 20th century, matches the suspected time course of emissions of greenhouse gases from human activities. Careful reviews by panels of experts (Houghton et al. 1996) concluded that “The balance of evidence suggests a discernible human

ATMOSPHERIC-DRIVEN CHANGES

ATLANTIC OCEAN 2

0

−2 1950

1

1970

1990

INDIAN OCEAN

Heat content anomaly (1022 J)

0 −1 −2 1950 4

1970

1990

PACIFIC OCEAN

2 0 −2 −4 1950 6

1970

1990

WORLD OCEAN

4 0 −4

1950

1970 Year

1990

35

inﬂuence on global climate.” Half a decade later, another panel of experts concluded that “There is new and stronger evidence that most of the warming observed over the past 50 years is attributable to human activities” (IPCC 2001; Watson et al. 2001). These conclusions stem largely from two lines of evidence. First, there is a remarkably coherent body of facts that all seem to point to higher-than expected temperatures in the atmosphere, oceans, and land across the earth. The evidence is convincing because it comes from highly diverse sources and because the time course of warming coincides with the time course of greenhouse gas emissions from human activities. Second, the model results just discussed above strongly suggest that temperatures across the world would be considerably different in the absence of contributions from human greenhouse emissions. The warming of recent decades is therefore not only demonstrable, but is a signiﬁcant departure from natural circumstances, and human activities seem the only likely explanation that could have forced the recent changes. A further review by a variety of other scientiﬁc bodies agreed with these conclusions, and manifested their agreement in a bold and uncompromising statement (box, p. 37). In spite of the weight of the evidence, and the overwhelming agreement by the scientiﬁc community, there are skeptics as to the reality of global warming, and as to the possible disruptive effects. The skeptics—largely, and ironically, residing in the USA, the nation that emits a quarter of the global carbon dioxide—include a handful of scientists and media commentators, and unfortunately many political ﬁgures with powerful inﬂuences. Their views are expressed in opinion columns headlined “More global hot Figure 2.6 (left) Time series through the last half of the 20th century, showing 5-year running means (± standard error) of heat content and mean temperature anomalies in the upper 300 m of the Atlantic, Indian, Paciﬁc, and World Oceans. The anomalies are the deviation of each entry relative to the mean for the entire data set. From Levitus et al. (2000).

36

CHAPTER 2

28.5

ABU DHABI

Temperature (°C)

28.0 27.5 27.0 26.5 26.0 1974 1976 1978 1980 1982 1984 1986 1988 1990 1992 1994 1996 1998 2000

Temperature (°C)

28.5

CHAGOS

28.0 27.5 27.0 26.5 26.0 1974 1976 1978 1980 1982 1984 1986 1988 1990 1992 1994 1996 1998 2000 Year

Figure 2.7 Time course, 1973–2000, of seawater temperatures in two sites, Abu Dhabi in the Arabian Gulf, with an annual range of about 20°C, and the Chagos archipelago in the Indian Ocean, with an annual range of about 4°C. The straight lines are the best ﬁt to the time courses; the jagged lines are the 12-month moving averages of temperatures. Adapted from Sheppard (2001).

air”, referring to the IPCC reports (IPCC 2001; McCarthy et al. 2001), or “The rain forest does not need saving”, referring to the complex of expressed opinions about the loss of habitat and climatic consequences of tropical deforestation. Commenting on the information available on climate warming, their opinion appears to be that “. . . the overwhelming balance of evidence shows no appreciable warming trend in the past 60 years: hence, it is unlikely to be signiﬁcant in the future” (Singer 2001). Such skepticism plays into the hands of those politicians and others that give higher priority to matters other than the environmental future of

the globe. The political, social, and economic measures—and implicit lifestyle changes—that have to be taken if we are to act on the recommendations to globally reduce emissions of greenhouse gases, are politically daunting, and will demand draconian changes in the use of energy. In fact, an ex-Vice President of the USA is reported to have said, probably in discouragement, “the minimum that is scientiﬁcally necessary [to address global warming] far exceeds that maximum that is politically feasible” (McKibben 2001). For example, a much-discussed report on the National Energy Policy for the United States (Cheney et al. 2001) focused on the means for

ATMOSPHERIC-DRIVEN CHANGES

Statement on the science of climate change “Despite increasing consensus on the science underpinning predictions of global climate change, doubts have been expressed recently about the need to mitigate the risks posed by global climate change. We do not consider such doubts justiﬁed. There will always be some uncertainty surrounding the prediction of changes in such a complex system as the world’s climate. Nevertheless, . . . it is at least 90% certain that temperatures will continue to rise, with average global surface temperature projected to increase by between 1.4 and 5.8°C above 1990 levels by 2100. This increase will be accompanied by rising sea levels; more intense precipitation events in some countries and increased risk of drought in others: and adverse effects on agriculture, health, and water resources . . . It is now evident that that human activities are already contributing adversely to global climate change. Business as usual is no longer a viable option. We urge . . . prompt action to reduce emissions of greenhouse gases . . . The balance of the scientiﬁc evidence demands effective steps now to avert damaging changes to earth’s climate.” A joint statement issued by the Australian Academy of Sciences, Royal Flemish Academy of Belgium for the Sciences and the Arts, Brazilian Academy of Sciences, Royal Society of Canada, Caribbean Academy of Sciences, Chinese Academy of Sciences, French Academy of Sciences, German Academy of Natural Scientists Leopoldina, Indian National Science Academy, Indonesian Academy of Sciences, Royal Irish Academy, Accademia Nazionale dei Lincei (Italy), Academy of Sciences Malaysia, Academy Council of the Royal Society of New Zealand, Royal Swedish Academy of Sciences, Turkish Academy of Sciences, and Royal Society (UK) (Science, May 18, 2001, p. 1261).

37

providing for energy demands by extracting fossil fuels and burning them. The priority in the Policy was to allow Americans to use about one-third more energy during the following two decades, essentially “business as usual”. Climate change was given brief treatment in six paragraphs, and measures such as energy conservation considered more a “personal virtue” than a basis for energy policy (McKibben 2001). There is a further political difﬁculty. Because of the inertia of the heat budgets of the atmosphere and seas (Wigley 2005), any action taken, even if drastic, will likely have detectable improvements only in the long term (decades to centuries). Promises of such distant results are unconvincing to citizens or decision-makers whose career time steps more likely extend to a few years at most (Hasselman et al. 2003). Nonetheless, it is encouraging to read the views of other political ﬁgures. The Chief Scientiﬁc Adviser to the British Government averred that “climate change is real, and the causal link to increased greenhouse emissions is now well established . . . climate change is the most severe problem we are facing today . . . there is still a good chance of mitigating the worst effects of climate change . . . [by reduction of] carbon dioxide and other greenhouse gas emissions worldwide” (King 2004). The measures necessary to cut emissions sufﬁciently will be difﬁcult to implement, but arguments have been made that measures such as carbon taxes may, at least in theory, signiﬁcantly lower the probability of human activity leading to dangerous levels of interference with climate (Mastrandrea & Schneider 2004). It is evident that there remain powerful economic and political constituencies that will work to ensure that the current energy policies and uses are maintained (Chow et al. 2003). It is also the case that, in time, fossil fuels will become too costly or scarce so that other, renewable, energy sources will become more economically competitive and hence more politically palatable. It is essential to foster energy conservation and develop renewable sources, such as wind and solar energy. It might even be necessary to take a second look at technically improved versions of nuclear energy production

38

CHAPTER 2

methods. The concern is whether the time when we make the technological transition away from fossil fuels will be soon enough to prevent the complex and pervasive changes to the earth’s environments (and human health, societies, and economies)—including, and importantly, coastal areas—that will surely ensue from continued use of present practices.

Effects of climate-related changes on coastal environments22 The responses of organisms to the climatedriven changes in temperature, UV radiation, precipitation, and weather are highly variable. Mangroves (Field 1995) and seagrasses (Short & Neckles 1999) seem unlikely to be strongly affected by current warming, for example. Other environments and the species therein may be more susceptible. Below we deal with the potential impacts of warmer water (which may create population dislocations, and alter the carbon chemistry of sea water), increased UV radiation, and altered precipitation and weather.

There is an emerging general pattern of recent poleward movement of distribution ranges of many animal species, coastal and otherwise (Barry et al. 1995; Holbrook et al. 1997; Hellberg et al. 2001; Beaugrand et al. 2003). In an exhaustive review of recent changes in the distribution ranges of different animal species, 39 species of marine invertebrates showed range expansions poleward, and only three did not; 24 species of marine zooplankton had expanded poleward, and none had not (Parmesan & Yohe 2003). A study that compared the ratio of birds with southern and northern geographic distributions that overwintered in Cape Cod, USA, found that relatively more southern species overwintered in this coastal area across the decades of the 20th century (Fig. 2.8). The ratio of southern to northern bird species varied with no trend, in response to unidentiﬁed variables, during the early part of the century, but as rising global temperatures became more evident after the 1960s, the ratio of southern to northern birds increased in concert with the temperature increases.

4.8

Dislocations in geography and time

There are many examples where warming has altered geographic distributions or seasonal timing of appearances of a large variety of species, including some marine ones. There are terrestrial examples of cases where the timing alterations impair the long-evolved synchrony among species, for instance, in the timing of arrival or emergence of pollinators, and the opening of ﬂowers they fertilize, or birds that migrate and may ﬁnd the food supply not available (Peñuelas & Filella 2001). 22

The reader may miss mention of many other aspects of the consequences of climate changes, such as, for one example, carbon sequestration in the oceans. Such issues are certainly among the larger issues in science and policy of coming decades. To do justice to these important issues, coverage of topics would have had to expand to the deeper oceans, land, and atmosphere, which would unduly lengthen this book. Denman et al. (1996) provide a useful review of some related topics.

Southern : northern ratio

Effects of seawater warming

9

4.2

55

99

9

9

99

8 55 8 8 65 5 88 8 5 6 333 667 8 8 57 47 7 8 8 3 763 4 44 767766 4 444 4 667 4 5

3.6

3 0

0.1 0.2 0.3 Temperature anomaly (°C)

0.4

Figure 2.8 Time course of the ratio of southern to northern bird species overwintering in coastal Cape Cod during much of the 20th century. The classiﬁcation into “northern” and “southern” was done on the basis of the geographic range of the winter range of the species. Numbers 3 to 9 represent data for the decades 1930s, 1940s, etc. Data from Valiela and Bowen (2003).

ATMOSPHERIC-DRIVEN CHANGES

The latitudinal expansions and contractions of species ranges are by no means novel. Such changes are known from detailed examination of the geological record and pollen data (Davis & Shaw 2001). It is difﬁcult, though, to ascertain whether we are dealing merely with geographic shifts, or with effects that threaten the abundance and persistence of species, or even the function of ecosystems. There have been expressions of concern about detrimental effects (Holbrook et al. 1997; Davis & Shaw 2001), but the evidence on absolute detrimental effects seems inconsistent. For example, certain coastal polar populations such as some, but not all (Barbraud & Weimerskirch 2001), penguin species (Smith et al. 1999), might increase in abundance and geographic range. Some temperate ﬁsh species affected by warming became exceptionally productive (Beamish 1993). These results contrast with observations in the food web of the North Sea (Beaugrand et al. 2003). Increasing seawater temperatures during the mid-1980s led to two detrimental effects. First, warmer seawater temperatures increased metabolic costs, requiring greater food consumption. Second, warmer temperatures shifted the species composition of zooplankton, and the altered assemblage of zooplankton provided a smaller mean size of prey available to young cod at a critical stage in their development. The smaller prey lowered growth of the young ﬁsh, as well as reduced recruitment. These climate-driven pressures, combined with overﬁshing (discussed in Chapter 11), are likely to be the explanation for the crisis in cod populations in the North Sea and elsewhere in recent decades.23 Similar relationships of increased temperatures to shifts in the phytoplankton, zooplankton, and young salmon have been also found in the North Atlantic (Beaugrand & Reid. 2003).

39

We therefore have plentiful evidence that widespread geographic and seasonal dislocations are taking place. We have suspicions and some circumstantial evidence that these dislocations may have signiﬁcant consequences on marine populations, but much more information will be needed to evaluate the likely changes. In addition to the biogeographic changes prompted by warming there are vital economic, political, and public health issues. For instance, diseases now limited to warmer latitudes may be spreading poleward. Cholera, dengue fever, and equine encephalitis may all be favored by warmer temperatures or other features fostered by climate change, including El Niño events (Colwell 1996). These speculations are circumstantial, but there is little doubt that we are witness to major adjustments in biological distributions worldwide, that there are more than environmental issues affected, and that these changes are climate-related. Moreover, there is more and more evidence that human activities, principally emissions of greenhouse gases, are behind the changes. We can be alarmed or pleased with the prospect of warming, and the biogeographic restructuring going on across the world, depending on what we value and where we live, and whether we consider ecological or public health issues. To qualify the environmental changes occurring in the coastal environments of the world as beneﬁcial or detrimental involves premature value judgments. What is undeniable is that there are major alterations in species ranges being forced by atmospheric warming, that we need far more information about the consequences of these changes, and that the shifts in species distributions will lead to substantial ecological and human changes. Alteration of inorganic carbon equilibria in sea water 24

23

During 1963–1983 there was a “gadoid outburst” in which codlike ﬁsh populations ﬂourished. Beaugrand et al. (2003) argued that the same array of climate-driven controls, but with reverse trends, might have been responsible, even though ﬁshing pressure was high. Thus, “bottom-up” controls might have been as important as “top-down” controls exerted by ﬁshing pressures. These issues will be deﬁned and discussed more extensively below in the chapters on overﬁshing and eutrophication.

Dissolved inorganic carbon (DIC) plays an important role in key biological processes such 24

Smith and Buddemeier (1992), Gattuso et al. (1999), and Kleypas et al. (1999a) provide more details of the phenomena involved in this section.

40

CHAPTER 2

as photosynthesis and calciﬁcation. In sea water under present-day conditions, DIC is made up of < 1% dissolved carbon dioxide (CO2), 90% bicarbonate (HCO3−), and 10% carbonate (CO32−) ions. The equilibrium between these three forms depends on the pH of the water, its temperature, and salinity, as well as in the partial pressure of CO2 in the overlying atmosphere. Concentrations of CO2 in the global atmosphere, however, have increased by about 25% since the 1700s, and might reach double preindustrial levels by 2065 (Houghton et al. 1996). Biogeochemical model studies show that over the past 300 million years, seawater pH has been strongly buffered, and has only been about 0.6 pH units lower than at present; continued increases in CO2 emissions may lead to pH lowering to about 0.7 pH units. This means that the foreseeable changes in coming centuries might exceed those experienced over 300 million years (Caldeira & Wickett 2003). Such signiﬁcant increases in the acidity of sea water can be expected to have many, largely unforeseen biological effects; at this point, at least two major effects are of concern. First, the increased inputs of dissolved CO2 from the atmosphere through the air/sea interface will likely increase supplies of CO2 in near-surface waters. Second, the increased partial pressure of CO2 in the atmosphere during the past century could lead, through its effects on DIC chemical equilibrium reactions in sea water, to a decrease in the carbonate saturation capacity of tropical sea water. These effects could be the basis for signiﬁcant changes in organisms that make differential use of CO2 and carbonate, such as plants, algae, and corals. More CO2 could increase production and biomass of plants such as seagrasses (Short & Neckles 1999) that are carbon-limited, but probably leave seaweeds that are not carbon-limited unaffected (Beardall et al. 1998). Thus, in the subtidal zone, more CO2 dissolved in sea water might lead to species shifts towards seagrasses. Intertidal wetland plants—grasses, mangroves—in general may be favored by increases in CO2 (Field 1995). For coralline algae and corals, lower carbonate supply could mark a critical environmental

change (Kleypas et al. 1999a). These organisms depend on a supply of carbonate (largely in its aragonite and calcite forms) to build support skeletons. Calculations based on well-established chemical reactions suggest that by the middle of the 21st century, the predicted increases in concentrations of atmospheric CO2 may decrease precipitation of aragonite (the main form of calcium carbonate used by corals for reef-building) by 14–39%. Such decreases would signiﬁcantly weaken coral skeletons, and expose coral reefs to damage by storms and other disturbances (Kleypas et al. 1999a). They could also decrease the reef’s ability to keep up with sea level rise. These of course are suppositions built upon uncertain predictions, but do seem a potential future threat to these coastal systems. Changes in seasonal cycles

The temperature of sea water is one of the factors that determines the timing of the seasonal patterns of abundance of organisms and processes in coastal ecosystems. The seasonal pattern of nutrient supply, growth of phytoplankton, rates of grazing, and so on govern not only but the timing of food supply, reproduction, and other fundamental features of the life of organisms that make up coastal food webs, but many biogeochemical functions as well, including microbial transformations, nutrient cycling, and decomposition of organic matter. In New England waters, the seasonal plankton cycle includes a late winter peak of phytoplankton as a norm, but on the occasional warmer year, the peak in phytoplankton growth shifts to a late summer peak.25 The mechanism underlying the shift may be the greater responsiveness of the herbivorous zooplankton than of phytoplankton to increases in seawater temperature. As sea water warms, grazers seem to increase their activity faster than the phytoplankton, and the increased grazing pressure prevents the phytoplankton from blooming, at least until later in the season, when the grazers are themselves 25

Many features mentioned in several previous papers on this topic are summarized by Oviatt (2004).

ATMOSPHERIC-DRIVEN CHANGES

eaten by other larger predators that proliferate best in warmer water or during the warmest part of the year. The removal of grazers in warmer water releases the phytoplankton from control by their grazers, which leads to a late summer peak in the algae. It turns out that, as in many other parts of the world’s oceans, New England coastal waters have been warming considerably during recent decades (Nixon et al. 2004). This warming might result in a future reconﬁguration of the controls of the seasonal plankton cycle. These future trends may include the late summer phytoplankton peak in phytoplankton production, as well as some of the other features, such as greater delivery of ungrazed algae to the sea ﬂoor. If temperature regimes do have such a preponderant role in constraining seasonal patterns of plankton, coastal ecosystems may change in substantial ways as warming continues (Oviatt 2004). Warming may bring unforeseen but widespread consequences concerning the timing of supply of food for, and increased occurrence of diseases of benthic species of commercial interest, storage of organic matter and regeneration of nutrients in the sea ﬂoor, changes in the species composition (favoring warmer-water taxa), and a shift from a dominant grazer/predator food web toward a microbial/decomposer food web, among other process changes. Further research will reveal the full potential for temperaturedriven changes in coastal food webs; at present we can only conjecture that wholesale changes are in store. Effects of increased ultraviolet radiation Increases in radiation within UV wavelengths have biological consequences (Vincent & Neale 2000). Nucleic acids (DNA and RNA) and proteins readily absorb and are sensitive to increases in UV radiation. Because of this, many biological processes are impaired after exposure. Inhibition of photosynthesis; dislocation of allocation of carbon compounds into lipids, carbohydrates, and proteins; damage and bleaching of pigments; lowered cell motility; and changes in respiration, nutrient uptake, and nitrogen ﬁxation rates are

41

among the reported direct symptoms of exposure to UV radiation. Indirect symptoms include the production of excess active oxidizing agents mentioned in the section on coral bleaching above. Symptoms of exposure to UV radiation vary with species, duration of exposure, intensity of radiation, degree of acclimation, cloud cover, particles in the water, and so on. In many examples of even the best experimental procedures results are uneven. Systematic studies show that in many cases, for example in copepods and ﬁsh larvae, only that small portion of the population that is exposed within the upper 1 m of the water column was seriously affected (Browman et al. 2000). Concern with the destruction of ozone has in recent years generated a huge literature dealing with the effects of exposure to UV radiation. Effects have been found for viruses and bacteria (Jeffrey et al. 2000), phytoplankton (Vernet 2000; Helbing et al. 2001), benthic invertebrates (Nozais et al. 1999), and zooplankton and ﬁsh (Zagarese & Williamson 2000). This work has been important in laying the groundwork, and ﬁnding ways to identify UV damage in the ﬁeld. Unfortunately, it has proven difﬁcult to carry out experimental work that provides relevant exposures, opportunity for avoidance, and protection responses to come into effect. Much of the work thus does not approximate exposures in the ﬁeld, and may overestimate possible damage (Vincent & Neale 2000). Exposure to UV radiation is not a biologically novel threat to organisms. In many cases, physiological and behavioral mechanisms have evolved across geological time. We discussed earlier the induction of mycosporine-like amino acids27 in corals, but many other coastal producers also manufacture these protective compounds as needed (Norris 1999). Motile organisms may also seek cover, move deeper in 27

These compounds have a long evolutionary history, and have become part of the resources used by not only the organisms that synthesize them, but also their predators. For example, Antarctic krill acquire UV-absorbing mycosporine-like amino acids from the phytoplankton they consume (Newman et al. 2000), presumably for their own use as protection against UV radiation.

42

CHAPTER 2

the water column, or carry out activities nocturnally (Roy 2000). While we have much evidence that, potentially, UV exposure creates biological difﬁculties, there are less convincing data showing signiﬁcant consequences in the ﬁeld. Perhaps longer and more intense exposure, and less ozone, might eventually create more difﬁcult— and measurable—circumstances and effects. Effects of changes in precipitation and weather There has been concern that the atmospheric changes being documented might generate more frequent extreme weather, including hurricanes and ENSO-related disturbances. Reviews of global climate change have found little evidence that warmer conditions would alter total precipitation, although climate models do suggest such changes will occur (Houghton et al. 1996; IPCC 2001). Instead of marked changes in total precipitation, it appears that there might be a shift toward greater heterogeneity in the geographic and temporal distribution and intensity of precipitation, and perhaps a greater frequency of stronger meteorological events (Karl & Trenberth 2003). Such changes could lead to more intense rainfall (Palmer & Räisänen 2002) and ﬂoods, for example, particularly at lower latitudes (Milly et al. 2002). There is evidence that at warmer temperatures relatively more of the precipitation falls in heavy (> 40 mm day−1) to extreme (> 100 mm day−1) amounts (Karl & Trenberth 2003). Such shifts could intermittently affect coastal environments at local scales, but consistent patterns have been difﬁcult to detect. Two recent papers argue, however, that there is a 30-year record of increasing intensity of tropical cyclones (Emanuel 2005; Webster et al. 2005). The intensity was related to warming temperatures and the increased destruction potential is likely to augment ecological damage as global temperatures rise. Temporal trends of storm or disturbance frequency seem strikingly local and uncertain (Michener et al. 1997). There may be recent increases in wave height in the North Atlantic (Guley & Hesse 1999), and there might have been a reduction of frequency of storm surges in the

northern Adriatic coast (Pirazzoli & Tomasin 1999). But there is little conclusive evidence that these local anomalies were different in degree compared to the variation that has taken place historically. In addition, there is no consistent evidence as to their coastal effects (McCarthy et al. 2001). In addition, even if there were increases in storm intensity, not all storms have similar consequences on all environments. Hurricane Mitch damaged coastal reefs on Belize (Mumby 1999). In other cases, such as Hurricane Bob on Cape Cod, the effects of the severe but short-lived disturbances dissipated quickly, and conditions in the coastal marshes, lagoons, and estuaries returned to normal within weeks to months (Valiela et al. 1996, 1998). Storm effects depend on the intensity of the disturbance, as well as on the type of habitat: environments dominated by hard, rigid biological structures are far more sensitive than environments with soft, unconsolidated underpinnings.

The global climate change dilemma for coastal environments Concerns about future climatic changes are not new (Storch & Stehr 2000): Arrhenius in the 19th century wrote about carbon dioxide and climate change; in the 1930s Kincer (1933) wrote a paper asking “Is our climate changing?”, sounding an alarm about a signiﬁcant warming (see Fig. 2.3). After the 1940s, industrial aerosols cooled the atmosphere (see Fig. 2.3), prompting alarm about a new ice age (Storch & Stehr 2000). These historical concerns turned out to be exaggerated. The 18th century geologist James Hutton28 believed that what would happen in the next half 28

In a ground-breaking but skeptically received lecture given in 1785 to the Royal Society of Edinburgh, Hutton (1788) argued that what has taken place is the key to understanding what occurred in the geological past, arguing that we might “. . . examine the construction of the prefent earth, in order to understand the natural operation of time paft; to acquire principles, by which we may conclude with regard to the future courfe of things. . . .” This radical innovation in thinking about science was a key to the uniﬁcation of scientiﬁc studies, and inspired many others, Charles Darwin in particular, later in the next century.

ATMOSPHERIC-DRIVEN CHANGES

century would on the whole be much like what happened in the last half century. It is not evident that this principle will apply where we have the accelerating changes we have seen in the carbon dioxide and temperature records for recent decades. For terrestrial ecosystems, it seems quite evident that on-going changes in temperature, UV radiation, and disruption of precipitation and weather are on their way to prompting novel and fundamental alterations. The most prominent effects will be on health, food production, and water supply for the increasing human population, as concluded by the IPCC updates (IPCC 2001; McCarthy et al. 2001), and the statements from the various academies of science (see box, p. 37). Issues arising from climate change may become most problematic for coastal urban areas (Deelstra 1995), agricultural regions (Brinkman 1995), and desertic zones. Plants, animals, and humans are going to be inevitably affected as are the water budgets, erosion, and untold other features of land ecosystems (Vitousek et al. 1997; Milly et al. 2002; Parmesan & Yohe 2003). We can point out remarkable instances of climate-driven changes in the ENSO phenomena, for example, which have pervasive effects on land environments at remarkably long distances. In Indonesia, ﬁres associated with the 1997–1998 ENSO event were associated with drought and burning of 5.2 million ha of rain forest, the largest ﬁre disaster ever observed (Siegert et al. 2001). Much farther away, over the last 30 years, high-quality harvests in the ﬁve main wine regions of Spain were also associated with the occurrence of El Niño that same year or the year before (Rodó & Comin 2000). Good wine years are most often those during which there is enough but not excessive rain, and sufﬁciently long, warm growing seasons—conditions associated with the occurrence of El Niño in the Paciﬁc Ocean. Such global-scale, but indirect and often arcane, linkages speak of the expected pervasive inﬂuence of global climate-driven events throughout the complex set of local environmental variables that in complicated ways result in good or bad vintages. The dilemma for coastal aquatic environments is the difﬁculty in clearly deﬁning the risks that

43

might follow current atmospheric-driven changes in temperature, UV radiation, and precipitation. Yet, we know that climate change is widespread, and has, and will, prompt changes that might be likely to affect species and ecosystems in the future. We have reviewed the worst-case history of coral bleaching, and the considerable geographic displacements in the North Sea and North Atlantic. These are perhaps the most eloquent examples of effects of climate-driven variables, examples supported by many other instances of dislocations. We lack, however, consistent evidence of serious deleterious impacts or lack of recovery of the affected biota. The widespread and certainly increasing nature of the climate-driven effects, however, suggest that we need, with a certain urgency, to learn far more about their possible effects on natural systems in coastal environments.

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Ocean: An ENSO event and a warning of future change? Ambio 28:188–196. World Meteorological Organization (WMO). 1995. Scientiﬁc Assessment of Ozone Depletion: 1994. Global Ozone Research and Monitoring Project Report No. 37. Geneva, Switzerland. World Meteorological Organization (WMO). 2003. Scientiﬁc Assessment of Ozone Depletion: 2002. Global Ozone

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Research and Monitoring Project Report No. 47. Geneva, Switzerland, 498 pp (http://www.unep.org/ozone/pdf/ scientiﬁc-assessment2002.pdf). Zagarese, H. E., and C. E. Williamson. 2000. Impact of solar UV radiation on zooplankton and ﬁsh. Pp. 279–320 in de Mora, S., S. Demers, and M. Vernet (eds). The Effects of UV Radiation in the Marine Environment. Cambridge University Press, Cambridge, UK, 324 pp.

Chapter 3 Sea level rise

View during the ﬂood of 1966, from the Piazzetta in front of the Ducal Palace, across the channel in front of Venice, to the island of San Giorgio Maggiore. From Hibbert (1989), © Topham Picture Library.

SEA LEVEL RISE

49

Figure 3.1 The northern Adriatic region, showing the shorelines at about 600 AD and the land ﬁlled in after 600 AD. The Rivoalto settlement became the city of Venice. Adapted from Lane (1973).

A case history: sea level rise in the North Adriatic1 Coastal and historical setting The sparsely populated lowlands along the northern coast of the Adriatic, called Venetia by the Romans, were by all accounts an unprepossessing province of little interest. The area, a nearly impassable mix of mudﬂats, wetlands, channels, and estuaries, supported scattered settlements of people comfortable on water and boats, whose living depended on ﬁshing and salt drying. One of the settlements was Rivoalto, a relatively high site with access to a navigable canal (Fig. 3.1). In Roman times the open lagoons were far more 1

The inclusion of this case history here owes much to Andreina Zitelli, Massimo Cardinaletti, and Davide Tagliapietra, who introduced me and my students to, and effectively communicated their pleasure and fascination with, the extraordinary nature of Venice and its lagoon.

extensive than today, reaching from south of Ravenna (the center of authority representing the Eastern Orthodox Empire in Byzantium) nearly to Trieste. These lagoons were protected from the wind and wave action of the Adriatic by strings of sandbars, lidi, which give their name to a particular site today, Lido, the Venetian beach resort. The open waters behind the bars were the “Seven Seas” that Pliny made part of the common idiom about sailing. The shallow lagoons, with their mudﬂats, sandbars, marshes, and canals have been ﬁlled in across the centuries by sediments brought by rivers and by human intervention. The rather ﬂat mainland shore today lies many kilometers seaward of its position before 600 AD (Fig. 3.1). Most of the ﬁlled lands are under 2 m of elevation, and much of the area is still below the mean sea level (Sestini 1992). The Lombard invasion of the remnant Roman Empire in 568 AD prompted the ﬂight of peoples from the mainland into the lagoons, where the muddy and shifting shallows made it difﬁcult

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Figure 3.2 Map of land uses on and around the Venice Lagoon area. Venice is in the center of the lagoon, connected to the mainland by a bridge (dashed line). Marghera is an industrial town on the inner shore of the lagoon. The lagoon has major inlets at Chioggia, Malamocco, and Lido, which lead to ship trafﬁc lanes (shown as arrows) to the major industrial and tourist port sites. The land use categories are: 1, valli diked for ﬁsh culture; 2, areas diked and used for agriculture; 3, tidal ﬂats (barene) ﬁlled for industrial development; 4, areas ﬁlled since 1900 for urban development; 5, barene open to tidal exchanges; 6, breakwaters designed to harden the inlet shores; 7, dredged channels for ocean-going ship trafﬁc. The “T” locates the petroleum transfer terminal. Adapted from Pirazzoli (1987).

for outsiders to venture, and the isolation offered some protection from the Germanic threat. The Lombard incursions were succeeded by new invasions of Franks, led by Charlemagne’s son Pepin, but the inaccessibility of the lagoon settlements made it possible to resist the waves of invasion. Venetia remained as an outpost of the Eastern Orthodox Empire, and the independentminded Venetians named a dux (that later became the Doge of more familiar Venetian usage). From this beginning the Rivoalto settlement became Venezia, which was to remain independent from 697 to 1797 as a city, a republic, and then an empire.2 Alterations to the coastal landscape In recent centuries there have been many changes to the mosaic of shallow-water environments of the Adriatic Sea. The importance of the lagoon 2

Excellent reviews of the history of the Venetian story and the Northern Adriatic can be found in Lane (1973), Lauritzen et al. (1986), Hibbert (1989), and Zorzi (1989).

was well known to the Venetians; they not only used the lagoon intensively, but also made some of the earliest efforts at management of coastal environments. Early in the 16th century, Cristoforo Sabbadini (1487–1560) noticed an expansion of the reed beds growing on sediments, presumably from the mouths of the Po River nearby. Concerned that the inﬂux of Po sediments might close the Chioggia inlet of the lagoon, Sabbadini appealed to the Venetian Republic, which ﬁnally agreed to sponsor a large regional project to redirect the ﬂows and sediments from the Po, as well as to restructure the inlets to the lagoon. Another change was that the area of marine environments—wetland, mudﬂat, and lagoon— “reclaimed” by people had increased markedly. These reclaimed areas were devoted to agricultural crops, mariculture, industry, urban areas, and more (Fig. 3.2). Fin- and shell-ﬁshing in the lagoons, as well as maricultural efforts, were, and are today, widespread. All these alterations meant that for most of ﬁve centuries, the lagoon of Venice, as well as

51

SEA LEVEL RISE

Increasing acqua alta floods Intermittent ﬂoods were always part of life in the lagoons of the North Adriatic. High tides have been an issue for Venice for centuries, but toward the close of the 20th century, high tides reached higher levels (Fig. 3.3). Beyond the yearto-year variability evident in the scatter of Fig. 3.3, there was a slow but clear rise (an acceleration) in the height of high tides (note the slope of the best-ﬁt curve). The increase in tidal height has been interpreted as the result of rising sea level: if this was so, we have to conclude that as the 20th century passed, sea level in the North Adriatic slowly rose, particularly after the 1930s. Another measure of the changing sea level regime is the decadal-scale increase in the frequency of events referred to as acqua alta

3

This material is from Sestini (1992), Pirazzoli (1987, 1991), and Bondesan et al. (1995).

Height of highest tide of the year (cm above MSL1897)

200

150

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50 1870

1890

1910

1930

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Figure 3.3 Highest tides recorded in Venice, 1872–1981, in each year, relative to the mean sea level (MSL) during 1897 (black circles). The curved line is the long-term trend for the entire data set. The thicker line is a 10-year running mean; the thinner line a 5-year running mean. Data from Pirazzoli (1987). 160 ≥ 1.3 m

Number of tides

the coastline of the northern Adriatic, became increasingly human-dominated environments. There remain remnants of pristine environments, but, by and large, the coastal mosaic has been strongly modiﬁed by people. This is not to say that there is no “nature”: there are still rich ﬁsheries, shellﬁsh beds, wetlands, and the rest. What we have in the Venice lagoon is a common feature of coastal environments worldwide: strong human modiﬁcation, coupled to a rich environment, full of resilient populations that in one way or another do well in perturbed environments, and are intensively harvested by humans. Thus, though disturbed through the centuries, the lagoon has remained very much alive ecologically, as well as furnishing ﬁsh and shellﬁsh, with the wetlands supporting abundant wildlife. All of these natural and human activities, as well as tourism—the major economic base for the region—have come into question in recent decades, threatened by ﬂooding.3 Add to this the uncertain continuing integrity of the unmatched historical and artistic heritage of Venice, and it is easy to fathom the concern about sea level rise in the northern Adriatic.

120

≥ 1.2 m ≥ 1.1 m

80

40

0 1871

≥ 1.0 m

1891

1911

1931

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1971

Year

Figure 3.4 Number of high tides higher than 1, 1.1, 1.2, and 1.3 m per year in Venice, for the decades between 1870 and 1980. Data from Pirazzoli (1987).

(Fig. 3.4). Note that once again, after the 1930s, the number of tides higher than 1 m increased markedly. By the late 1980s there were 160 tides per year that reached 1 m or higher. The one particularly high value in Fig. 3.3 is the widely publicized 1966 acqua alta (cf. Chapter 1 frontispiece). We can compare these high tide heights to elevation contours at the center of Venice (Fig. 3.5);

110

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Rialto Bridge

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St. Mark’s Square

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Doge’s Palace

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St. Mar

k’s Bas

in 0

50 100 m

94

Cartographie A. SEVESTRE, CNRS-Meudon.

Figure 3.5 Map of the center of Venice showing the relative elevation as contours spaced in tens of centimeters. The dashed lines indicate the gang-plank walkways (see Fig. 3.7). From Pirazzoli (1987).

SEA LEVEL RISE

Figure 3.6 Venetians carrying on daily life in spite of ﬂoods. From Pirazzoli (1987).

it is apparent that during the high tides commonly seen in the late 20th century, much of Venice would be ﬂooded, excepting a few small spots at the highest elevations (> 1.3 m). Venetians, as always, have made the best of it, carrying on their daily life (Fig. 3.6). Floods became so routine that the city established elevated trestle-and-plank pathways (Fig. 3.5) along the most trafﬁcked streets (Fig. 3.7). Press attention to the now-famous acqua alta of 1966 made the issue evident to the world: Venice was slowly but alarmingly sinking beneath the sea. Causes of sea level rise in Venice The increased ﬂoods affecting Venice and the northern Adriatic are the result of “natural” as well as human-derived causes of change. Meteorological and astronomical forces prompt short-

53

Figure 3.7 Photos of approximately the same location in Venice, showing elevated gang-planks being erected in preparation for a ﬂood (top), and during a ﬂood (bottom). The gang-plank walkways are located along the main thoroughfares in the city, indicated as dashed lines in Fig. 3.5. From Pirazzoli (1987).

term changes of sea level. Long-term changes have been caused by global sea level rise and by land submergence. The weather forces intermittent shifts in sea level, by surges of water associated with storms, winds, seiches, and low-pressure fronts. Combinations of seiches and Bora (cold, dry, northeast) and Sirocco (south and southeast) winds have intermittently raised high-tide levels as high as 1.8 m in Venice (Sestini 1992). These short-term events therefore have produced signiﬁcant ﬂoods, notwithstanding their relatively short duration. The occurrence of these events is highly uncertain, as evident in the variability in the highest tides

54

CHAPTER 3

of the year (the most meaningful for human concerns) across the centuries (see Fig. 3.3). These intermittent mechanisms have raised high tides above critical levels, and are partly responsible for the acqua alti that so interfere with life in Venice. The effects of weather have also been exacerbated by the dredging of deep channels for deep-draft shipping (shown by arrows in Fig. 3.2). The deeper channels ease the inﬂow of ﬂood waters. In addition, large areas of lagoon were diked to protect large lagoons (valli) used for ﬁsh culture, and other areas were ﬁlled for industrial development. These projects reduced the area of lagoon and wetland available to disperse ﬂood waters and dissipate the energy of tidal and storm-driven currents. Astronomical inﬂuences include tides (semidaily to intra-annual), as well as lesser-known multiyear cycles.4 These astronomical mechanisms—responsible for the up-and-down cycles evident in the running means in Fig. 3.3—all impose somewhat orderly ﬂuctuations on both short-term meteorological alterations and longerterm changes. They do not secularly change the magnitude of sea level ﬂuctuations. Human activities, however, have altered the tidal regime itself,5 by the deepening and building of protective walls of navigation channels,6 and by cutting off large sections of the lagoon from tidal exchange. The rise in relative sea level in Venice during the 20th century was about 22–24 cm (Sestini 1992). Of this increase, about 9 cm were due to the actual rise of global sea level, largely because

4

These astronomical cycles include lunar declination (16-plus years), sunspot (11 years), Saros (18 years and 11 days), and Méton (19 years) cycles (Pirazzoli 1987). 5 The amplitude of tidal excursions has increased within Venice Lagoon. The harmonic M2 and S2 components of tidal cycles were 5% lower in Venice compared to the open Adriatic in 1920; in the 1980s, these components were 4% greater in Venice (Aldighieri et al. 1985). 6 Deepening and hardening of the shorelines has taken place in the Lido, Malamocco, and Chioggia inlets (Fig. 3.2). For example, shoal depths (< 3 m) made entry to Venice via the Lido hazardous, so from 1880 to the present, breakwaters and navigation channels (now 11 m deep) were constructed through the Lido to Venice and to Marghera. These allow oil tankers and cruise ships to enter the lagoon, but also ease the path for ﬂood waters. Similar works have altered the two other inlets (Pirazzoli 1987).

the volume of the global sea has increased, driven by higher temperatures in the mixed upper layers of the seas. An additional 2 cm or so resulted from “natural” geological subsidence of the land in the region. A further rise of 12–14 cm or so was attributable to subsidence of the land surface owing to withdrawal of groundwater from the aquifer under the region. The water was used mainly for industry in Marghera, but also for irrigation and urban consumption. Withdrawal of groundwater became most intensive from the 1940s on, which ﬁts our ﬁnding that the relative rise of sea level might have accelerated after the 1930s. There was also some compaction of sediments beneath buildings, which lowered the surface even more. After the alarm raised by the 1966 acqua alta, withdrawal of groundwater was curtailed, and human-driven subsidence stopped by the late 1970s, when aqueducts bringing water from other rivers replaced aquifers as water sources. This is an unusual example in which a management decision was made and achieved—at least by stopping further sinking, which was the management goal on a regional scale. Effects of sea level rise on the North Adriatic coast7 The ﬂooding found in Venice and its lagoon was also seen elsewhere in the coast of the North Adriatic. The rise in relative sea level is associated with a variety of consequences. All the beaches of the area are unstable, with retreat along most of the coast, even though 60% of the shore has been furnished with coastal defense structures (56 km of groins and breakwaters, 55 km of seawalls, 42 km of jetties and dikes, and 36 km of other structures). Marshes, tidal ﬂats, and sand bars have been converted to lagoon open water. Sediments are being removed from the lagoon ﬂoor. This erosion is in part the result of removal of sediment by the sea, uncompensated by sediment brought in by the rivers. Until the 19th century, the rivers of the area were

7

This section is developed from material in Pirazzoli (1987), Sestini (1992), and Bondesan et al. (1995).

SEA LEVEL RISE

loaded with sediments, and built up such features as the delta of the Po. Signiﬁcant regression of shores began late in the 20th century when construction of dams upstream, plus better soil conservation practices, resulted in the settling of river-borne sediments behind the dams, and hence reduced sediment transport downstream to the estuaries and deltas. The sediment load of the Po, for example, was lower by 38% between 1965 and 1973, a fairly short interval. The construction of piers, jetties, and groins along the shoreline further promoted erosion and retreat of the shoreline. Dredging of substantial quantities of sand from the bed of the Po itself did not help the river’s ability to transport sediment down to its delta. The continued presence of wetlands in the area will depend on their ability to accumulate sufﬁcient sediment to accrete vertically. Because so much of the area has been diked, or provided with protective walls, the wetlands cannot grow inland as the sea rises. Because rivers have been dammed or impounded, sediment loads are lower, and hence there is little possible vertical accretion. Predicted sea level rises of as much as 13–34 cm near Venice and 28–49 cm in the Po delta, would almost certainly diminish the area of wetland habitats remaining in the Adriatic lagoons (Sestini 1992). The rise of sea level had, and has, serious human effects. Increased loss of land could affect the tourist trade, both in terms of visitors to Venice, and to summer tourism on the barrier beaches. The greatest threat, of course, is to the continued functioning of Venice and other population centers as cities. Each of these, if sea level rises as expected across the next centuries, will have to develop some means to remain functioning. All the suggestions made so far are technically challenging, are inordinately expensive, and have likely undesirable side effects. For Venice, “the only solution is to separate the Adriatic from the lagoon during the periods of high meteorological tides” (Bras et al. 2001; Pirazzoli 2001). One example of the several technically daunting options is the installation of mobile gates at the lagoon inlets. The idea is to ﬁrst narrow the inlets with transversally disposed

55

dikes, and then to build mobile gates8 across the gaps between the dikes. The gaps would span 200–300 m, with various depths appropriate for shipping (12 m deep in Lido, 15 m in Malamocco, 12 m in Chioggia). The gates would remain out of the way of navigation during normal times. If ﬂood water threatens to exceed 1.1 m above the mean sea level of the period 1884–1909 (about 0.87 m above present mean sea level), the gates would swivel, one end raised off the bottom, preventing further entrance of water into the lagoon (Pirazzoli 2001). A major issue of concern regarding restricting ﬂow between the Adriatic and Venice lagoon has been whether there might be serious environmental consequences. Venice has been able to release its wastes into the lagoon for centuries because of one fortunate key factor: the mean residence time of water in the lagoon is about 1 day (Gacic et al. 2001). This has meant that the lagoon has been effectively ﬂushed daily, a feature that makes it possible for the urban wastes to be removed, and keeps the water quality surrounding Venice within tolerable levels. There have been disputes as to the number and duration of closures that might take place under different sea level rise scenarios (Bras & Harleman 2001; Pirazzoli 2001). Many have worried that lowering exchanges across the inlets would seriously impair water quality, and lead to undesirable ecological changes in the lagoon. The feasibility of construction, operation, and maintenance of the massive gates, and the environmental and economic consequences of their construction and operation, therefore remain controversial (Ammerman & McClennen 2000; Bras et al. 2001; Pirazzoli 2001). The not-inconsequential estimated cost of US$2.65 billion did not

8

The project, the Experimental Electromechanical Module, nicknamed MOSE after its Italian initials, consists of 79 mobile gates, situated at the three inlets to the Venice lagoon. One end of the 20 m-long gates would be hinged securely to the sea ﬂoor. The internal hollow compartments would normally be ﬁlled with sea water, and the gates would lie horizontally on the bottom of the inlets much of the time. When necessary, compressed air would be pumped into the hollow gates, which would then pivot upwards to partially seal the inlets. Each gate would oscillate with waves and surges, a design requirement to avoid damage to the structures.

56

CHAPTER 3

deter the Italian government from giving the goahead to MOSE (“Work begins on plan to save Venice from the sea”, Boston Globe, May 15, 2003). If it works, MOSE may help Venice survive for a period of perhaps 50 years, perhaps allowing sufﬁcient time for development of other approaches to deal with the inevitable longer-term further rise in sea level. The point here, however, is that major efforts will be required now and in the future to save Venice from ﬂood waters, that there is uncertainty about the success of the different options, and that whatever is selected as an option will be remarkably costly. The case of Venice and its lagoon is yet another demonstration of the powerful linkages between people and coastal environments. The very waters that served to protect the early Venetian settlement now threaten present-day Venice, and the historic turn of events were caused by a complex mix of natural and anthropogenic forces. Restoring the problems in these cases are technically, economically, and politically daunting.

Changes in global sea level rise The issue of sea level rise has been in the public eye since the late 1960s, owing to the Venice ﬂoods, and the publicity was stimulated in the early 1980s when alarming predictions of rises of 56–345 cm by 2100 were issued (Hoffman et al. 1983). Since then, there has been intense controversy regarding both the rate of global sea level rise and the relative contribution by different mechanisms.9 The key question in these projections is what might be the relative inﬂuence of human activities, relative to forcing by nonanthropogenic causes. Just as with everything else on earth, the level of the world’s sea has changed across time scales of millions and thousands of years through geo9

Church (2001) reviews information about the methods and possible biases involved in estimates about measuring global sea level rise, and the models used to predict future changes. There are issues regarding the adequacy of tide gauge sites, processes involved, and terms to consider. In spite of these caveats, models and measures agree that there have been, and will continue to be, higher than “natural” increases in sea level rise.

logical history (Fig. 3.8), perhaps rising more than 600 m (Fig. 3.8 top), and falling almost 150 m below present levels (Fig. 3.8 middle). As recently as 30,000 to 18,000 years ago, the last glacial maximum, the surface of the sea was more than 100 m below present levels. Coastal salt marshes, for example, ﬂourished along the edges of a plateau in what is now the Northwest Atlantic, and their growth left an unmistakable clue in peat layers under what is now Georges Bank, in the Atlantic off the eastern coast of North America (Emory 1991). The recession of glacier ice allowed water to return to the sea. Quite coincidentally, the history of human civilization has largely, therefore, taken place during a span of steeply rising sea level (Fig. 3.8 middle): we have, as a species, proliferated during a brief, geologically speaking, interglacial period, and our civilizations have developed through rising sea levels. There are many records of continuing increases during more recent time scales of years and decades (Fig. 3.8 bottom, for sea levels in Stockholm and Amsterdam, but see Fig. 3.14 below). The year-to-year changes are most likely created by meteorological differences; the decadal trend, however, shows a steadier increase in sea level rise. The increases are also not uniform everywhere on the world’s shores, as we will see shortly, but the consensus is that current global sea level rise may be about 1–2 mm per year (Miller & Douglas 2004). The change in sea levels documented in Fig. 3.8 was created by a variety of processes. Sea level has varied across a wide range of time and spatial scales, and different processes have effects at different scales (Fig. 3.9). Wind-driven waves and tsunamis, for instance, change sea level during relatively short timespans and affect relatively small areas. Tidal effects alter sea level at diurnal and monthly scales, but across larger expanses. Meteorological and climatic phenomena [low-pressure fronts, storms and surges, wind setup, monsoons, El Niño–Southern Oscillation (ENSO) events, and droughts] may change the sea level at small or large spatial scales, for periods of days to months and even years. For example, depending on the intensity of the ENSO event, mean sea level in the vast areas of the Paciﬁc

57

SEA LEVEL RISE

Sea level above present level (mm)

600 400 200 0 −200 590

C

O 500

S

D

M-P

P

T

J

440 410 360 290 250 210 Million years before present

C 140

T 65 35

5

Relative sea level (m)

0

−50

−100

−150 140

120

100

80 60 40 Thousand years before present

20

0

Amsterdam

Stockholm +100 mm −100 1700

1800

1900

2000

Year

Figure 3.8 Time courses of sea level expressed at time scales of millions of years (top; inferred from global seismic data), tens of thousands of years (middle; relative sea level at Huon Peninsula, New Guinea, inferred from oxygen isotope data from raised reefs, and from submerged fossil corals), and hundreds of years (bottom; for Amsterdam and Stockholm). Compiled by Lambeck and Chappell (2001) from many sources and using different sorts of information.

58

CHAPTER 3

105 Tectonic 104

Tides

Polar ice melt

Polar ice recession Mountain glacier melt

102 Waves

Effect on sea level (mm)

103

101

100

Groundwater removal

Thermal expansion 10−1

10−2 10−3

10−1

101

103

105

Time scale of effect (years)

between Australia and North America may differ by as much as 50 mm (Wyrtki & Nakahoro 1984). Such differences may be more a tribute to our data-crunching prowess10 than differences that will affect coastal change, but, on the other hand, hurricanes and typhoons can bring extensive ﬂooding and damage to coasts. Intensity of effect is therefore not coupled to extensiveness or to duration of the alteration of sea level. Here we mostly focus on effects that occur at all (global, regional, and local) spatial scales, but mainly at time scales relevant to humans, or that can be recorded by us. The various mechanisms included in Fig. 3.9 can alter sea level by up to hundreds of meters, 10

Maps of sea level across the world’s oceans have been produced from satellite-collected data (for example, Semtner 1995). These maps report the sea level to the nearest 20 cm. The satellite data, of course, are aggregated across large spatial scales; clearly differences of 20 cm measurements will be imperceptible to a sailor in a ship intermittently and erratically bobbing up and down in 15 m waves at one point on the sea surface.

107

Figure 3.9 Sketch of the approximate range of magnitude of effects from several processes (tides, waves, thermal expansion, mountain glacier melting, polar ice melting, polar ice recession, groundwater removal, and tectonic movements) on sea level rise, and the time scale of these effects.

regardless of time scale. The mechanisms whose effects are relevant to human experience (up to centuries) involve thermal expansion of seawater volume and ice mass melting, with some allowance also for post-glacial rebound, freshwater interception, and atmospheric disturbances. Longer-scale effects may move the sea level to a greater extent, but those changes are beyond human time scales. Land itself may move during short-term events, such as earthquakes. Forces unleashed by tectonic activity (faults, grabens, earthquakes, tsunamis) may change the relative sea level in short time scales by rapidly lifting or sinking relatively small sections of land masses. Near Naples, Italy, intense volcanic and tectonic activity in late Roman times raised certain areas, and lowered others, in a complex patchwork of alterations that created a new coastal topography that severely affected human settlements. Today, tourists can see raised beaches, visit semi-submerged temples, or look at the remains of a mosaic ﬂoor from a Roman

SEA LEVEL RISE

59

potentate’s palace under 10 m of sea water.11 Many examples of such local perturbations have been described in the Aegean (Cundy et al. 2000; Stiros et al. 2000), Rhone delta (Vella & Provansal 2000), Papua New Guinea (McSaveney et al. 2000), China (Mei-e 1993), South Australia (Bourman et al. 2000), New Zealand (Goff et al. 2000), Tierra del Fuego (Isla & Bujalesky 2000), and in southern Chile (Fig. 3.10). We here are more interested in the somewhat longer-term phenomena that might also affect sea level at larger spatial scales. We will therefore focus on thermal-driven expansion of seawater volume, post-glacial rebound of land, withdrawal of groundwater from aquifers, changes in storage of water on land, and tectonic displacements.12 These agents of sea level change work across decades to millennia, and may affect relatively large areas (Fig. 3.9). Mechanisms forcing long-term sea level changes13 The speciﬁc mean relative sea level in any one coastal area is a result of atmospheric changes that alter global sea volume and geological forces 11

Such archaeological markers have been suggested as supplementary data to add to the geographically skimpy tide gauge data (Flemming 1992). In certain areas, such as the Mediterranean, there are hundreds of suitable sites with known ages. Data from such sites, if corrected for tectonic activity, match those derived from tide gauge records, and add needed detail to regional interpretation of sea level changes. 12 Several other mechanisms have been hypothesized as potentially being capable of inﬂuencing sea level. An interesting global mechnism is the melting of methane hydrates (a form of frozen methane referred to as clathrates) (Bratton 1999). The volume of clathrates in submarine sediments is substantial; warming of sea water might thaw the clathrates, which, when frozen, have a greater volume than the liquid forms (much as ice does relative to water). This lower volume hypothetically might produce a lowering of sea level of 10–146 cm. Of course, we lack knowledge of the time it would take to warm the water to sufﬁcient depths, and the response of sea level to the melting. This mechanism might be more appropriately an explanation of geological time scale phenomena, rather than the subjects of concern in this book. Locally, the compaction of sediments can be important, as we saw in the case of Venice. In the Netherlands, reclaimed areas of clays may be compressed to half their thickness after 100 years, and peats to oneninth of their original thickness (Jelgersma 1996). 13 Most of the contents of this section are summarized from information from National Research Council (1990), Warrick et al. (1996), and Milliman and Haq (1996). These sources review many issues, contributions, approaches, and data sets.

Figure 3.10 Salt marshes in Golfo Elefantes, Chile, growing in an area that suffered 2–2.5 m subsidence during an earthquake in 1837. The marsh grass Puccinellia has replaced the southern beech (Nothofagus) forest that formerly grew on the land. Dead tree trunks are still visible in the background. From Reed (1995).

that change vertical positioning of the land. Major mechanisms that change the level of sea water include the thermal expansion of seawater volume owing to climatic warming, melting, or creation of ice, and the interception of fresh water by humans. Major mechanisms altering the surface elevation of the earth’s crust include post-glacial rebound and tectonic movements. Thermal expansion of seawater volume As temperature increases, seawater volume expands. The volume of the sea, and hence sea

60

CHAPTER 3

level, therefore increase as global warming takes place. The process is complex, because there are strong irregularities in how heat and mass of water are redistributed across the oceans; for example, how much of the water column might be subject to vertical mixing and therefore be exposed to heat exchanges with the warmer atmosphere. Various estimates of the contribution to global sea level rise by thermal expansion are available, but the newer data suggest that this mechanism may account for only a fraction of the total (Miller & Douglas 2004). Changes in glacial ice volume Volumes of ice are found in mountain glaciers and polar ice caps. There are within-region patterns that show consistent time courses of sea level, but large among-region differences occur, and the uncertainties are difﬁcult to evaluate. The recession of mountain glaciers has been carefully noted since the early 1900s (see Fig. 2.4). Estimates of the contribution to sea level rise by the melting of glaciers vary; one researcher estimates that this source added about 0.35 mm yr−1 between 1890 and 1990, and 0.6 mm yr−1 between 1985 and 1993 (Dyurgerov 1994), so the melting of mountain glaciers may be accelerating. The relative contribution from changes in polar ice caps is far more uncertain. Polar ice caps hold by far the major ice volumes on earth.14 The uncertainty derives from a lack of data from these distant places, as well as the unknown extents to which the different forms of ice are thawing or forming. Distinctions have to be made among ice on land, ice on shelves, ground ice, and ﬂoating ice. Warrick et al. (1996) review the various estimates, and conclude that, as best guesses, the formation of Antarctic ice may contribute about −0.3 mm yr−1, and the melting of the Greenland ice sheet may have an equivalent sea level rise of 0.3 mm yr−1. 14

In Greenland the ice is mainly deposited on land. In the much colder Antarctic environment, there is a thick layer of ice on land, plus vast ice shelves extending into the Antarctic Ocean, as well as “ground” ice attached to the sea ﬂoor under ice shelves. The ice on land in Greenland is more exposed to warming atmospheric temperatures, and is hence losing volume. In Antarctica it is thought that ice volumes may be still expanding, but sufﬁcient data are lacking.

More recent estimates put the sea level effect of the melting of the Greenland ice mass at about 0.6 mm yr−1 (Mitrovica et al. 2001). The polar ice estimates have an uncertainty of at least 50%, and may, judging from their magnitudes and sign, roughly cancel each other out. There has been much recent reassessment of the relative magnitude of mountain and polar glacial melting. Reviews of recent data lead to the conclusion that melting of mountain glaciers may be larger than polar melting, and that the sum of glacier melt may be the major contributor to global sea level rise (Arendt et al. 2002; Miller & Douglas 2004), larger than the other posited mechanisms. Freshwater storage and withdrawal

Human beings have manipulated water on land sufﬁciently to alter the hydrological cycle at global scales. This is discussed in more detail in Chapter 4; here I note that the human need for water has led to considerable withdrawal of aquifer water, as well as increased storage in reservoirs. In addition, water use and storage are complexly linked to local and regional climatic effects, so that it is difﬁcult to separate natural from anthropogenic effects. Groundwater removed from aquifers in excess of recharge rates releases water to the land surface. This water may then evaporate or move to the sea; the net removal of water from aquifers also leads to subsidence of the land surface. This effect was evident in the case of Venice, but that is not an isolated instance. In Bangkok, for example, removal of groundwater dramatically lowered elevation of the land surface (Nutalaya et al. 1996; Sabhasri & Suwarnarat 1996). Bangkok spreads across ﬂat and low-lying (0–1.5 m above mean sea level) terrain next to the Gulf of Thailand, and is subject to ﬂoods from seasonal monsoon rains as well as sea water. Between the 1930s and 1978, ground surface elevations ranged from 20 to 85 cm. From 1978 to 1988, subsidence ranged from 20 to 160 cm. As pumping of water out of the aquifer beneath Bangkok increased, the water table dropped and mean sea level rose. Sea level increased markedly across the late 20th century, and repeated ﬂoods are commonplace there

SEA LEVEL RISE

now. There are many other examples (Jelgersma 1996): the Galveston–Houston area of the US has subsided 2.8 m since 1915, owing to groundwater withdrawal; in Taipei there was 2 m subsidence during 22 years of withdrawal; and in Japan groundwater withdrawal has caused subsidence in 59 sites. Aquifer water removal also lowered land elevations in Long Beach (California), Ravenna, various sites on the coast of China (Mei-e 1993), and several sites in Japan (Bird 1996). Relative sea level may, therefore, be directly and indirectly altered by human withdrawal of groundwater. The effects of this mechanism are demonstrably meaningful within quite speciﬁc areas, and appear at the relatively short time scale of decades. These topics are treated at more length in Chapters 4 and 5. Although impressive locally, the net contribution of the removal of groundwater to global sea level rise is modest, perhaps reaching 0.07–0.38 mm yr−1. The storage of fresh water in reservoirs has been a major goal of water husbandry worldwide. This topic is one of the most data-rich that we will encounter, and is discussed more fully in Chapter 4. Stored water is water that will either not reach the sea or will do so with longer time lags. Thus, reservoirs counter the trends to sea level rise, albeit by a small contribution, about −0.09 to −0.54 mm yr−1. Anthropogenic deforestation has multiple effects on the hydrological cycle: less water in plant biomass, less evapotranspiration, and more runoff from land to sea. The net effects are by no means well established. The loss of wetlands (Chapter 6) similarly must have complex, and as yet unknown, effects on sea level rise. “Reclamation” of wetlands around Manila in the Philippines has been suspected to be responsible for considerable sea level rise (Spencer & Woodworth 1993). Conjectures about other consequences of human activities—thawing of permafrost, storage in irrigated soil, evaporation from reservoirs, and so on—have not been evaluated. The uncertainty of all these various hydrological mechanisms is large indeed, and their aggregate contribution to sea level rise has perhaps been 0.1 mm yr−1, with a range of −0.4 to +0.75 mm yr−1. These mechanisms maybe contributed

61

about 0.5 cm to global sea level rise during the 20th century, a small contribution compared to the larger global effects of thermal mountain glacier melt and thermal expansion. Meteorologically driven changes A variety of weather-related factors may alter sea level. We have already mentioned waves and tides, only to add that these are short-term inﬂuences. There are meteorological disturbances that may be somewhat longer term, lasting years. Examples of such weather-driven features are the effects of ENSO events. Multiyear gauge records off Rabaul, Papua New Guinea show a trend to decreases in relative sea level, owing to uplifting of the land (Fig. 3.11). Note, however, the large drops and rises in sea level during 1982–1983 and 1986–1987. These were years with pronounced El Niño events. Another example is the increased frequency with which the Thames Barrier engineering works, which protect London from ﬂooding, have had to be closed (Fig. 3.12). These periodic high ﬂoods are a combination of more frequent unusual meteorological events and high tides. Changes in atmospheric pressure and winds can therefore signiﬁcantly alter sea level, not only in localities such as New Guinea and London, but across large sections of entire oceans, usually for months to a few years. Crustal deformations

All land surfaces on earth are subject to vertical shifts, prompted by post-glacial rebound, plate movement by tectonic forces, accumulation of sediments, as well as other minor mechanisms. Huge ice sheets, up to 3,000 m thick, covered the northern third of the world’s continents at the peak of the Wisconsin glaciation, 18,000 years ago15 (Fig. 3.13 top panels). There was enough water stored in the ice to lower the sea level to more than 100 m below what it is today. By 15,000 15

The causes of the glacial/interglacial cycles are not well understood. It is probably a complex interaction between astronomical changes in the angle of the earth’s orbit, and the distance between the earth and sun, modiﬁed by earth-bound phenomena, such as ocean circulation and effects of the ice sheets.

62

CHAPTER 3

Change in sea level relative to mean (mm)

30

20

10

0

−10

−20

−30 1973

1975

1977

1979

1981

1983

1985

1987

Year

Figure 3.11 Monthly mean sea level (solid line) and 12-month running mean of sea level (dashed line) at Rabual, Papua New Guinea, 1973–1987. From Wyrtki (1990).

Number of closures

30

20

10

0 1982

1986

1990

1994 Year

years ago the glaciers began to recede, and had largely disappeared by 7,000 years ago (Fig. 3.13 middle and bottom panels). In parallel, the global level of the sea rose, maybe as fast as by 12.5 mm yr−1. In the north, relative sea level rise initially might have been > 10 mm yr−1, decreasing to 1–2 mm yr−1 or less during the last 1,000 years (Bloom & Yonekura 1990).

1998

2002

Figure 3.12 Number of closures (during 2-year periods) of the Thames Barrier, a coastal engineering works designed to protect London from ﬂooding by the high waters of the River Thames. Data from King (2004).

The weight of the enormous mass of ice deformed the continental plates. The withdrawal of glaciers released the burden on the land below, and led to vertical realignments of land surfaces that continue even today. In most northern areas, this has meant an upwards thrust of land, lowering the sea level over large portions of the earth, and across 103–104-year time scales (see Fig. 3.9).

63

SEA LEVEL RISE

1800

30

10 2000 00

00

20 15000 100 500 00

18,000 yr BP

1200 600

18,000 yr BP

1600

00

20

80

0

1600

12,000 yr BP

12

00

20 0

1 80200 0 400

12,000 yr BP

0

60

Figure 3.13 Contours of the ice mass (in meters of thickness) formed over North America (Laurentian ice sheet, left panels) and Scandinavia (Scandian ice sheet, right panels), during three different periods of the geological record. Redrawn from Peltier (1990).

300

8,000 yr BP

This post-glacial rebound aspect is reasonably well understood. For example, models predicted that there would be 11 mm yr−1 of crustal uplift in the area of Hudson Bay, Canada (a site quite near the center of the Laurentian glacial mass, Fig. 3.13 left panels), as a result of glacial rebound. In fact, tide gauge measurements showed a rise of 11 mm yr−1 for the past 50 years, nicely corroborating the model predictions (Carter et al. 1994). The effect of glaciation on sea level was also far from uniform.16 The weight of the ice had 16

Details of the diverse deformations created by glacial loads on coastal topography are reviewed by Carter (1988), among other reference texts.

8,000 yr BP

created an upwards “forebulge” toward the southern limit of glaciation and beyond. The forebulge made for a sea level in the affected regions south of the glaciated areas about 10–20 m lower than it is today. South of the “hingeline” between glaciated and non-glaciated areas, that crosses New England in North America, the land, rebounding from the glacial forebulge, subsided, with much drowning of valleys, recession of coastlines, and continued erosion of the coast. In regions where relative sea level increased, such as the eastern coast of the USA, we ﬁnd characteristic environments, including long, shallow estuaries, coastal lagoons, and barrier beaches, all supporting vegetation such as seagrass beds and salt marshes. In fact, salt marshes, for instance,

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CHAPTER 3

are quite young environments: they are basically post-glacial features younger than 7,000 years old. Superimposed on the post-glacial rebound are regional or more local tectonic effects (Emery & Aubrey 1991; Flemming 1992) related to plate uplifts or subsidence, earthquakes, and vulcanism. These occur at widely different time and spatial scales. Raised marine terraces, drowned river valleys, and other features commonly seen along coastlines, owe their long-term origin to such forces. Because there is so much local detail, it has been difﬁcult to ascertain the relative contribution to sea level rise.

Magnitude of sea level rise Data on sea level are largely derived from tidal height data collected in gauging stations along the coasts of the world (Spencer & Woodworth 1993). Use of tidal level data to calculate relative global sea level rise is not, however, unambiguous (Miller & Douglas 2004). As we have just seen above, there are many mechanisms that prompt rises and falls in relative sea levels, and to make things more complicated, the different mechanisms may operate at different time and spatial scales. The time course of global sea level heights has been calculated based on available data from sites where tide records have been kept for considerable lengths of time (Spencer & Woodworth 1993). From the time courses in representative sites around the Paciﬁc Ocean we can see that in different places on the world’s coasts, sea level has varied in complex ways (Fig. 3.14). In most but not all places, sea level rose in the long term, but there are differences in magnitude and in details of the time scales (cf. Figs 3.8 bottom, 3.11, 3.14). There are annual shifts, runs of rises and falls of a few years, and across decades and centuries. The speciﬁc rises and falls are not parallel among stations; there is considerable heterogeneity among the time courses shown in Fig. 3.14. The global pattern of sea level toward the end of the 20th century is a heterogeneous patchwork

pattern, with different sections of the world’s coasts rising or falling (Lambeck & Chappell 2001). Among the many stations across the coasts of the world that maintain long-term gauge records of sea level, sea level is thought to be rising in 206 stations and falling in 129 stations. The patchwork pattern of sea level rise and fall can be found at different geographic scales. For example, differences in sea level rise are found across a relatively small part of the world such as the British coast (Fig. 3.15). The same range of variation in sea level rise has been reported at somewhat larger geographic scales, such as the coast of China [20 of 32 stations showed rising, the rest falling, sea level (Mei-e 1993)] and the coast of the United States (Nicholls & Leatherman 1996). Similar variation in sea level rise can be found at global scales (Lambeck & Chappell 2001). Tide gauge measurements show changes in sea level, but the inﬂuences of local geological processes and local short-term meteorological events are superimposed on the pattern of global sea level change that is of interest. To detect actual longer-term global sea level rise, the local, shorterterm variations, as well as the effects of crustal deformations have to be removed from tide gauge data.17 Once the unwanted effects are removed as much as possible, the residual values are averaged to represent the global mean sea level. There is also some question as to how many stations might be truly needed to obtain a suitable representation for the world’s seas. Further, there is the matter as to how long a record might be necessary. For example, a straight line ﬁtted to data obtained at stations on the eastern coast of the US with long-term tidal height records (> 50 years) shows a rise in sea level in all stations, with a range of 1.6–4.2 mm yr−1 (Nicholls & Leatherman 1996). If, instead, we use only the

17

Removal of the superimposed high-frequency effects can be done by various statistical ﬁltering techniques (Gornitz 1995). Unfortunately, lower-frequency effects are harder to remove, and the variation of effects that can be resolved given the available tidal gauge records is of similar amplitude to the sea level signals being quantiﬁed. The ﬁltering methods are continually being improved (for example, Mitrovica et al. 2001).

65

SEA LEVEL RISE

10 5 HONOLULU 0 10 5 SEATTLE 0 10

Sea level change (cm)

5 SAN FRANCISCO 0 10 5 BALBOA 0 10 5 SYDNEY 0 10 5

HOSOJIMA

0 −5 −10 −15 1890

1900

1910

1920

1930

1940

1950

1960

1970

1980

Year

Figure 3.14 Annual mean sea level for six stations around the Paciﬁc Ocean with relatively long records. The straight line on the Honolulu data shows a 1.5 mm yr−1 increase or 15 cm per century. From Wyrtki (1990).

data obtained between 1970 and 1991 (21 years), we ﬁnd that sea level fell in 17 out of 22 stations, and that the range varied from −4.7 to +1.0 mm yr−1. It is a challenge to discern whether the latter

data set is simply too short to appropriately depict the rise of sea level, or whether the rise of sea level has indeed decreased in the shorter span of decades.

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CHAPTER 3

+0.7

−3.7 +0.4

−4.7 −2.4

−7.0 0

+4.0

+2.1

+0.6 +2.0

+0.3

+0.3

−0.3 +1.7 +1.4 +1.5 +5.4

−3.2 +5.1 +1.7

−0.3

Lack of sufﬁciently detailed information to capture the differences from one locale to another, plus the multiplicity of mechanisms involved in setting sea level, and the differences in time and spatial scales of the various mechanisms, all conspire to complicate the estimation of the magnitude and changes in sea level. There is no doubt that the level of the global ocean is rising, but there is variation in the estimates of current and future trends in sea level rise. Compilations of estimated changes in global sea level range from 1–2.4 mm yr−1 (Raper et al. 1996) or 1–1.5 cm yr−1 (Warrick et al. 1996) to 1.5–2 mm yr−1 (Miller & Douglas 2004) measured across the 20th century. Such estimates add up to an increase of perhaps 18 cm during the 20th century, with a range of 10–25 cm. The magnitude of the effects of several mechanisms (Table 3.1) has been estimated using various models. As

+4.1

Figure 3.15 Recent sea level changes (in mm yr−1) around the British Isles. From Carter (1988).

already mentioned, melting of mountain (and to a less clear extent, polar) glaciers and thermal expansion of shallow oceanic volume appear to have made the greatest contribution to global sea level rise during the 20th century. Freshwater storage added a minor portion of the rise. There has been vigorous discussion as to whether the sum of these estimates adds up to match the measured relative global sea level rise, an issue that has not been settled. In general, however, the sums of contributions by the various mechanisms do not exactly match tide gauge data, but given the broad range of uncertainty associated with the various estimates, we cannot say that the differences are signiﬁcantly different. This at once suggests that we need to sharpen our estimates by future studies, but that we seem to have captured, to a reasonable degree, the major features of the controls of global sea level rise.

67

SEA LEVEL RISE

Table 3.1 Time scale and magnitude of the contribution to sea level change by four major mechanisms. Adapted from National Research Council (1990), Sundquist (1990). Time scale (years)

Potential contribution to sea level change (mm)

Seawater volume Shallow (0–500 m) Deep (500 –4,000 m)

10−1 to 102 102 to 104

100 to 103 100 to 104

Ice volume Mountain glaciers Greenland ice sheet Antarctic ice sheet

100 to 102 102 to 105 102 to 105

101 to 103 101 to 104 103 to 105

Water on land Aquifers Lakes and reservoirs

102 to 105 102 to 105

102 to 104 100 to 102

Crustal deformations Tectonic Post-glacial rebound

105 to 108 102 to 104

103 to 105 102 to 105

Changes in

Effects of sea level rise on coastal environments A large proportion of the coasts of the world has been, and will be, exposed to rising sea levels. The most vulnerable coastal countries are those with high concentrations of human populations near the sea, expanses of low-lying environments, and a limited ability to deal with the environmental change. Such countries might include Bangladesh, China, Egypt, Senegal, Nigeria, and Uruguay (Nicholls & Leatherman 1995), but by no means will the rest of the coastlines be exempt from sea level challenges. The challenges arise from a suite of diverse effects. Intensified flooding of uplands The area of land surface that will be ﬂooded as sea level rises will, of course, depend on slope. On the western coast of South and North America, for example, the effects of ﬂooding will be largely trivial. On the Nile (Milliman et al. 1989) and Ganges–Brahmaputra deltas (Alam 1996) substantial amounts of land could be submerged.

In the Republic of Maldives, the effects may be catastrophic.18 In human terms the most intractable problem will be in thickly settled areas. A sense of the complex dimensions of the problems can be garnered from the following passages, describing the consequences of subsidence and sea level rise for Bangkok, Thailand:

18

Vreugdenhil and Wind (1987) and Barnett and Adger (2003) review cases of atoll island nations. The Maldives is the most prominent, an archipelago of more than 1,200 tiny islands to the west of Sri Lanka. The islands are grouped into 19 atolls, with an aggregate area of 298 km2. The 269,000 Maldivians are densely (909 people km−2) dispersed over 200 islands, with more than 50,000 living in the capital, Male. The maximum elevation of the Maldives is 3 m above sea level. As they have throughout geological history, the corals that produced the atolls may, if undisturbed, keep up with rising sea level, but in the populated islands, human activities have lowered coral growth. Construction of dikes is impractical because of the high porosity of coral sand, which would permit ready inﬁltration of sea water. There are four more countries that are composed of low-lying atolls: Kiribati (population 78,000), the Marshall Islands (population 58,000), Tuvalu (population 9,000), and Tokelau (a dependency of New Zealand, population 2,000). All these countries are undeveloped economically, have serious water and food issues, and are vulnerable to storms and sea level rise.

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“During October to December, seasonally high tides rise up to 1.35 m above the mean sea level, which in turn raises the Chao Phraraya River water level, and then water ﬂows back into sewers and canals . . . areas . . . lower than 1.35 m [are ﬂooded] . . . Subsidence . . . [causes] settling of building foundations, underground pipelines, roads, walkways and bridges . . . cracking and bending of pipes, ﬂoor slabs, concrete walk-ways and steps, and detachment of septic tanks and sidewalks or steps from buildings . . . buried pipe lines can be lifted above ground . . . Buildings with shallow foundations can sink, roads buried. Overdraw of groundwater [results in] . . . salt water gradually . . . replac[ing] the freshwater. Groundwater in the Bangkok aquifer is now heavily polluted to the point that it is no longer potable. Septic tanks in the ﬂooded area . . . become water logged and sewage and associated diseases become potential . . . hazards. The transportation system [is] disrupted as [are] supplies [of] food and medicine” (Nutalaya et al. 1996).

And this is just a partial list jotted down as an aside in a paper on the geological aspects. Many of the major cities of the world lie, for reasons of commerce, navigation, and other historical contingencies, along the low-lying shores of shallow estuaries, deltas, and lagoons. The measures that will be needed to hold back ﬂood waters from these metropolitan areas will not only be remarkably costly, but their construction and maintenance will be a challenge to our technology and to our political and social fabric. Clearly, the most consequential effect of increased sea level rise will be on coastal cities, many of which have in fact developed where they are because of their location near the shore and within estuaries. Loss of coastal environments The vertical position relative to land and sea level is often important for organisms of many coastal environments. It follows that changes in relative sea level raise the question as to the resulting effects on coastal habitats. Coral reefs seem unlikely to be affected by direct effects of sea level rise. Vertical accretion rates calculated from various data suggest that reefs may accrete anywhere from 1 to 10 mm yr−1 (Smith & Buddemeier 1992). Such rates of vertical

20 Accretion rate (mm yr−1)

68

16 12 8 4 0 0

4

8

12

16

20

−1

Mean sea level rise (mm yr )

Figure 3.16 Plot of sediment accretion rates versus mean sea level rise for mangroves (open circles) and salt marshes (black circles). Values from salt marshes on the coast of the northern Gulf of Mexico are shown as “x”. Data from Armentano et al. (1988), Day and Templet (1989), Lynch et al. (1989), Wood et al. (1989), Kearney et al. (1994), Parkinson et al. (1994), Leonard et al. (1995).

movement are, on the whole, larger than the 1– 6 mm yr−1 that could be taken as a range for reasonable rates of sea level rise. The geological history of reefs suggests that they should be able to deal with signiﬁcant changes in sea level. Of all the coastal environments, wetlands—salt marshes and mangroves—may seem the most susceptible to an increased rise in sea level (Reed 1990). We know that the organisms within these environments are quite sensitive to the level of intermittent ﬂooding, and hence, as sea level rises, the bands of coastal vegetation must move upwards. The key feature of wetland maintenance versus sea level rise is the rate of accretion19 of the wetland surface relative to sea level rise (Fig. 3.16). Available data on these rates for mangroves and salt marshes suggest that many coastal marshes accrete at rates comparable to current sea level

19

Accretion in coastal wetlands may be mainly a result of accumulation of organic matter produced by the plants themselves, as occurs in New England salt marshes, or mainly by accumulation of mineral land-derived sediments, as occurs in the southeast of the USA, particularly in the Mississippi delta (cf. Chapter 5).

SEA LEVEL RISE

rise (points above the 1 : 1 line in Fig. 3.16); mangroves also seem to keep up with sea level rise. There are other coastal marshes where accretion has been insufﬁcient to keep up with sea level rise (points below the 1 : 1 line in Fig. 3.16). The majority of wetlands where subsidence outstrips accretion are from studies done in the Mississippi delta in Louisiana, where diking throughout most of the course of the Mississippi River has intercepted the source of sediment for the delta (more on this in Chapters 5 and 12). Where the marsh or mangroves are set against the land, as a fringe parallel to the shore, they can move inland as the sea level rises. A good example of this can be seen in the extent of mangrove vegetation within the mouth of the Fitzroy River in northwest Australia (Fig. 3.17). As sea level rose between 1949 and 1977, the mangroves colonized inland, and, at the same time, retreated from their seaward edge. In this example, sea level rise simply moved the habitat inland to where the topography permitted the mangroves to grow at the appropriate elevations. If the coastal topography had been one of steep slopes, the areas propitious for mangroves would be less extensive, and instead of the increase in mangrove area reported in Fig. 3.17, there would have been a decrease. Similar inland migrations can be found in salt marshes on subsiding shorelines. For example, within only 30 years of subsidence, the salt marsh grass Spartina patens colonized wet areas that were previously occupied by forests of bald cypress (Taxodium distichum) in coastal Louisiana (Fig. 3.18). Coastal marshes and mangroves growing as a coastal fringe seem capable of responding to sea level rise by moving inland. There are some exceptions to this conclusion. Wetlands that are sediment-starved by the interception of sediments behind dams may be unable to keep up with the pace of sea level rise (Chapter 5). Wetland stands on islands within estuaries have no land to move on to, and hence this type of coastal wetland is threatened by sea level rise (Kearney & Stevenson 1991; Wray et al. 1995; K. McGlathery, personal communication). The migration of wetlands inland may also be prevented in places where people have built hard barriers landward

69

of the wetlands (Brinson et al. 1995). Such situations may or may not be widespread, but can be locally important, worst-case examples of wetland loss. Sea level rise may or may not be the principal factor governing wetland losses. Snedaker et al. (1994) and Snedaker (1995) suggest that altered water and sediment transport,20 more than sea level rise, may be more inﬂuential in altering mangrove distributions. Other anthropogenic effects are also likely to cause large losses of most wetlands. For example, Nicholls et al. (1999) used modeling methods to estimate losses of coastal wetlands worldwide, and found that losses depended on the sea level rise and population growth scenarios used in the calculations, but concluded that sea level rise could lead to world wetland losses of 0–2% by the 2020s. Total wetland loses were likely to be higher, perhaps 13–31%. Habitat losses from other sources are discussed further in Chapter 6. Increased erosion of shores Erosion of different types of shorelines can occur regardless of increased sea level rise (Orviku et al. 2003), but the latter accelerates rates of erosion (Bird 1996; McCarthy et al. 2001). More than 70% of the world’s beaches are retreating as a result of sea level rise plus human intervention (Bird 1996). Rates of beach erosion in US beaches, for instance, averaged somewhat less than 1 m yr−1 during the 20th century, as the sea level rose (Leatherman et al. 2000, 2002). Such erosion can affect coastal marshes and mangroves, but the removal of sediment is often more characteristic of cliffs and beaches, where vegetation does not provide a buffer against the removal of unconsolidated sediment. Since more people have migrated towards these same shores (see discussion in Chapter 1), shore erosion has consequently affected more people. For as long as oceans and land have existed, the margins of shores have been subject to great

20

Chapters 4 and 5 deal with these topics; Chapter 6 discusses coastal habitat losses.

70

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Landmark 1

Landmark 2

1949

Seafront of mangroves in 1949

1977 Mangrove forest Salt flat

Tidal creeks

0

500 m

Figure 3.17 Changes in the distribution of mangrove forest, muddy salt ﬂats, and location of shoreline between 1949 and 1977 in the estuary of the Fitzroy River, Australia.

SEA LEVEL RISE

71

Figure 3.18 View of an area previously forested by cypress, a freshwater swamp tree, that was invaded by the salt marsh grass Spartina patens, after sea water elevated the salinity of the wetland water. From Reed (1990).

changes. Storms cause short-term erosion of beaches and other coastal environments, but on average there is recovery to long-term levels. Longterm erosion occurs mostly when sea level rises or sediment supply changes (Zhang et al. 2002). Where human interventions have captured sediment in upstream reservoirs, or added dikes, groins, rip-rap, and other shoreline protection devices, shoreline erosion is almost always accelerated.21 Increased salt water intrusion With rise in sea level, the boundary demarcating the interdigitation of salt and fresh waters will move inland. This may in turn have effects, including the replacement of freshwater species and lowering the quality of fresh water for human use. Environments (wetlands, estuaries, low-lying land) presently dominated by terrestrial or freshwater organisms may be invaded by salt-tolerant species (see Fig. 3.18). Salt water may seep across artiﬁcial dikes and dunes, and impair the current intensive agriculture, and affect the quality of drinking and irrigation water of low-lying areas

21

Shoreline erosion is discussed in more detail in Chapter 5.

in the Netherlands or Bangladesh. Both these countries have extremely high population densities, and most of their territory lies below 3 m above mean sea level. In the Netherlands, engineering waterworks have for centuries protected the extensive poldered areas from the North Sea, and a substantial portion of the country lies below sea level. Since the polder surfaces commonly lie below sea level, increased sea level rise will increase the head pressure, which may increase seepage of sea water into polders, waterways, and certainly into rivers and potable water supplies (Vreugdenhil & Wind 1987). Bangladesh has the highest agrarian population density in the world, 80% of its territory is lowland less than 3 m above sea level, and 50% of the land already suffers ﬂoods under the monsoon rains and high tides. Three-quarters of the population in Bangladesh depends on agriculture, which takes place in areas diked extensively for ﬂood protection and irrigation. The southwest part of the country, the Sunderbans, is home to 12.5 million people, but is still largely covered by mangrove forests, subject to ﬂooding, and the area is only partially protected by dikes. In the large delta of the Ganges–Brahmaputra, in the southcentral section of Bangladesh, river channels

72

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Impacts of sea level rise in the New York City metropolitan area22 The greater New York City region, with over 2,400 km of shoreline, will be vulnerable to accelerated sea level rise due to anticipated climate warming. The trajectories of sea level rise for the New York City area, predicted from current rates, or by several different models, all increase into this century (Fig. 3.19). A substantial area throughout the New York metropolitan area is now subject to potential ﬂooding (Fig. 3.20). As sea level rises further, there will be a number of consequences that need to be considered in planning for coastal management of the region. Storms and hurricanes seem unlikely to increase in frequency, but the level of the 100year storms will increase (i.e. the depth of ﬂoods will increase); the frequency of what is now 22

Gornitz et al. (2002).

are unstable, with intense erosion, accretion, and changes in course; the relatively fewer farmers in this part of the country often lose land, and are forced to move. The low-lying eastern part of Bangladesh hosts 1.3 million inhabitants; the agricultural land and the major city in this section are protected by a dike (Vreugdenhil & Wind 1987). Throughout this country, increased sea level will mean greater hardships for the population and greater alteration of the coastal environments. While in certain low-lying localities, seawater intrusion can therefore be critical, on a larger geographic scale this seems a minor effect of sea level rise, because in most coasts land slopes will constrain the effects to a smaller area. In many coastal areas, human withdrawal of groundwater has led to seawater intrusion into aquifers. In Japan, 38 coastal cities have reported salt intrusion into groundwater supplies (Yama-

the 100-year storm will therefore be shorter. The areas subject to ﬂoods will increase beyond current areas at risk. Coastal erosion rates will increase (to 4–10 times larger by the 2080s). Wetlands may be able to keep pace with sea level rise into the 2030s if current rates of sea level rise were to occur; or for less time if the other estimates are what actually happens. After that, a steady decline in coastal marshes will take place. These changes will prompt efforts at beach nourishment 2–7 times larger by 2050–2080 than at present, and such nourishment will have to be repeated at intervals. In the densely populated urban areas, lower ﬂoors may likely be abandoned. In less urbanized areas, people may retreat, or buildings may be relocated. Shoreline armoring will need to be applied it is decided to protect urban areas (but armoring will often accelerate erosion). Groins will be built (but will often also increase erosion further down-drift).

moto & Kobayashi 1986). Sea level rise might exacerbate this issue in the future.

Restoration potential The dependency of sea level rise on seawater expansion and the melting of glaciers means that any campaign to restore sea levels must aim ﬁrst at the control of global atmospheric temperatures, a restoration measure that seems pressing but daunting to achieve (Chapter 2). Local measures of relief might be achieved by restoring the supply of sediment for accretion on wetlands and deltas, or by halting the exploitation of groundwater under low-lying coastal areas. The restoration of sediment-deprived wetlands in estuaries is theoretically possible if sediment loads are restored to previous magnitudes. This would require demolition of dams

73

SEA LEVEL RISE

Port Jefferson, NY

80

Sea level rise (cm)

Sea level rise (cm)

Willets Point, NY

60 40 20 0

80 60 40 20 0

2000 2020 2040 2060 2080 2100

New York

Long

80 New York City

60

Montauk, NY

Connecticut

und

d So

Islan

Long lsland

Atlantic Ocean

40 New Jersey

20 0

Sea level rise (cm)

Sea level rise (cm)

The Battery, NYC

2000 2020 2040 2060 2080 2100

80 60 40 20 0

2000 2020 2040 2060 2080 2100

Sea level rise (cm)

Current trend HCGG HCGS CCGG CCGS

Sandy Hook, NJ

2000 2020 2040 2060 2080 2100

80 60 40 20 0 2000 2020 2040 2060 2080 2100

Figure 3.19 Time courses for predicted sea levels in ﬁve sites within the New York City metropolitan area through the remainder of the present century. Predicted sea levels were obtained using current rates of sea level rise, or British (HC) or Canadian (CC) models that considered the effects of predicted increases in temperature based on increased greenhouse gases (GG), or on increases in greenhouse gases balanced versus the effects of increased sulfur aerosols (GS). From Gornitz et al. (2002).

upstream, a politically and economically drastic step. To my knowledge, this has not been attempted successfully. Chapter 5 expands this discussion of altered sediment loads to wetlands and deltas. Other ways to try to restore eroded coastal sites are noted in the box opposite, which summarizes the situation for sea level rise and its consequences for the New York metropolitan area.

There is an example of success in control of a local factor—the subsidence of Venice caused by the removal of water from an aquifer. For over 400 years the lagoon of Venice remained an invaluable setting for one of the world’s most splendid cities. In the face of increasing pressure, evident even in the 16th century, the Magistrato alle Acque had inscribed in marble the following announcement:

74

CHAPTER 3

Westchester Co.

N

BRONX

NEW

JERSEY

n

ke

o ob

H

Newark

Laguardia Airport

Jersey City

QUEENS JF Kennedy International Airport

The Battery

BROOKLYN Ja m

Elizabeth

Nassau Co

Newark International Airport

MA NH AT TA N

Edgewate

aica

STATEN ISLAND h

eac

yB awa

Long Beach

k

Rockway Breezy Point Coney Island Point Beach

Roc

0

1

2

3

4

5 km

Figure 3.20 Zones at risk of a 100-year ﬂooding event during the start of the 21st century in the New York metropolitan area. These areas will inevitably increase in extent as projected sea levels rise. From Gornitz et al. (2002).

SEA LEVEL RISE

VENETORUM URBS DIVINA DISPONENTE PROVIDENTIA AQUIS FUNDATA AQUARUM AMBITU CIRCUMSEPTA AQUIS PRO MURA MUMITUR QUISQUIS IGITUR QUOQUO MODO DETRIMENTUM PUBLICIS AQUIS ANFERRE AUSUS FUERIT HOSTIS PATRIAE JUDICETUR NEC MINORI PLECTATOR PEOMA QUAM QUI SACROS MUROS PATRIAE VIOLASSET HUJUS EDICTUS JUS RATUM PERPETUVMQUE ESTO23

The edict appears to have been difﬁcult to enforce, as those that in fact “dared to damage the public waters” were most everyone in the region. Of course no one set out to damage the lagoon, but the net effect of human activities, from the 16th to the 21st centuries, as in so many cases of environmental change, was to inadvertently but persistently degrade the natural setting of Venice and the other northern Adriatic cities. Venetians quickly perceived the problems, living as they do so close to the waters, and have had some success in eliminating the withdrawal of groundwater from the aquifer under the region. This was an option because fresh water was obtainable from the Alps; such a ready alternative might not be so accessible in most other places. We should recall, though, that the effects of halting withdrawal were only to stop sinking of the land, not to restore previous elevations. It is much more difﬁcult to elevate land that has subsided; experimental injections of ground water into an aquifer under islands in the Venice lagoon have yielded only 5% restoration of previous land elevations. Efforts in pumping water into aquifers in California and Shanghai suggest that it is possible to inject clean fresh

23

Lauritzen et al. (1986) translate this as “The city of the Venetians/ with the aid of Divine Providence/was founded on water/enclosed by water/defended by water instead of walls/ Whoever in any way dares damage the public waters/shall be declared an enemy of the State/and shall not deserve less punishment/than he who broaches the sacred walls of the State/This edict is valid for evermore.”

75

water to counter subsidence, and even to some degree raise land surface elevations (Jelgersma 1996). Clearly, the global aspects driving sea level changes are outside what Venetians by themselves might hope to manage. Control of atmospheric temperatures will require global human action to minimize the rise of atmospheric temperatures —a daunting political and economic prospect that we discussed in the last chapter. In the larger scheme of things, we could at best hope to counter only that portion of sea level rise that responds in the relatively short term, and is under human inﬂuence. It seems unlikely that rising sea levels can practically be brought under control as we might wish, at least in the short term. In any case, even if all we will face is a continuation of rising sea level at current paces, it will be necessary for many urban centers to revamp their basic design and erect protective works to prevent the sorts of problems that occurred in Bangkok. Reworking of much of the infrastructure will surely be needed, as well as the construction of protective structures to divert ﬂooding sea waters. Many examples mentioned in this chapter— Venice and Bangkok, and the wetlands of the Mississippi, among others—show that the issue of sea level rise is closely linked to other agents of coastal change, including the use of fresh water by humans and the supply of sediment. These are the subjects of the next two chapters.

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Arendt, A. A., K. A. Eschelmeyer, W. D. Harrison, C. S. Lingle, and V. B. Valentine. 2002. Rapid wastage of Alaska glaciers and their contribution to rising sea level. Science 297:382–386. Armentano, T. V., R. A. Park, and C. L. Cloonan. 1988. Impacts on coastal wetlands throughout the United States. Pp. 87–149 in Titus, J. G. (ed.). Greenhouse Effect, Sea Level Rise, and Coastal Wetlands. US Environmental Protection Agency, Washington, DC, 186 pp. Barnett, J., and W. N. Adger. 2003. Climate dangers and atoll countries. Climatic Change 61:321–337. Bird, E. C. F. 1996. Coastal erosion and rising sea-level. Pp. 87–103 in Milliman, J. D., and B. U. Haq (eds). SeaLevel Rise and Coastal Subsidence. Kluwer Academic Publishers, Dordrecht, 369 pp. Bloom, A. L., and N. Yonekura. 1990. Graphic analysis of dislocated quaternary shorelines. Pp. 104–115 in SeaLevel Change. National Research Council, National Academy Press, Washington, DC, 234 pp. Bondesan, M., and 6 others. 1995. Coastal areas at risk from storm surges and sea-level rise in northeastern Italy. J. Coast. Res. 11:1354–1379. Bourman, R. P., C. V. Murray-Wallace, A. P. Belperio, and N. Harvey. 2000. Rapid coastal geomorphic change in the River Murray estuary of Australia. Mar. Geol. 170:141–168. Bras, R. L., and D. R. F. Harleman. 2001. Obsolete? No. Necessary? Yes. The gates will save Venice. Eos Trans. Am. Geophys. Union 83:217, 224. Bras, R. L., D. R. F. Harleman, A. Rinaldo, and P. Rizzoli. 2001. Rescuing Venice from a watery grave. Science 291:231–232. Bratton, J. F. 1999. Clathrate eustasy: Methane hydrate melting as a mechanism for geologically rapid sea-level fall. Geology 27:915–918. Brinson, M. M., R. R. Christian, and L. K. Blum. 1995. Multiple states in the sea-level induced transition from terrestrial forest to estuary. Estuaries 18:648–659. Carter, R. W. G. 1988. Coastal Environments: An Introduction to the Physical, Ecological and Cultural Systems of Coastlines. Academic Press, London, 617 pp. Carter, W. E., and 13 others. 1994. New gravity meter improved measurements. EOS Trans. Am. Geophys. Union 75:90–92. Church, J. A. 2001. How fast are sea levels rising? Science 294:802–803. Cundy, A. B., and 10 others. 2000. Coastal wetlands as recorders of earthquake subsidence in the Aegean: A case study of the 1894 Gulf of Atalanti earthquakes, central Greece. Mar. Geol. 170:3–26. Day, J. W. Jr., and P. H. Templet. 1989. Consequences of sea level rise: Implication from the Mississippi Delta. Coast. Manage. 17:241–257. Dyurgerov, M. 1994. Global Mass Balance Monitoring. Report to U.S.A. Department of State. Institute of Geography, Russian Academy of Sciences.

Emory, K. O. 1991. Georges Cape, Georges Island, Georges Bank. Pp. 38–39 in Backus, R. H., and D. W. Bourne (eds). Georges Bank. MIT Press, Cambridge, MA, 593 pp. Emory, K. O., and D. G. Aubrey. 1991. Sea Levels, Land Levels, and Tide Gauges. Springer-Verlag, New York, 237 pp. Flemming, N. C. 1992. Predictions of relative coastal sealevel change in the Mediterranean based on archaeological, historical and tide-gauge data. Pp. 247–281 in Jeftic, L., J. D. Milliman, and G. Sestini (eds). Climatic Change and the Mediterranean. Edward Arnold, London, 673 pp. Gacic, M., P. M. Poulain, M. Zore-Armanda, and V. Barale. 2001. Overview. Pp. 1–44 in Cushman-Roisin, B., M. Gacic, P. Poulin, and A. Artegiani (eds). Physical Oceanography of the Adriatic Sea: Past, Present and Future. Kluwer Academic Publishers, Dordrecht, the Netherlands. Goff, J. R., and 7 others. 2000. Evidence for an earthquake and tsunami about 3100 to 3400 years ago, and other catastrophic salt water inundations recorded in a coastal lagoon, New Zealand. Mar. Geol. 170:233–251. Gornitz, V. 1995. Sea-level rise: A review of recent past and near-future trends. Earth Surf. Process. Landf. 20:7– 20. Gornitz, V., S. Couch, and E. K. Hartig. 2002. Impacts of sea level rise in the New York City metropolitan area. Glob. Planet. Changes 32:61–88. Hibbert, C. 1989. Venice: The Biography of a City. W. W. Norton & Co., New York, 435 pp. Hoffman, J. S., D. Keyes, and J. G. Titus. 1983. Projecting Future Sea Level Rise: Methodology, Estimates to the Year 2100 and Research Needs. US Environmental Protection Agency, Washington, DC, 121 pp. IPCC (Intergovernmental Panel on Climate Change). 2001. Climate Change 2001: Impacts, Adaptation, and Vulnerability. Contribution of Working Group II to the Third Assessment Report of the IPCC. Cambridge University Press, Cambridge, UK, 1032 pp. Isla, F. I., and G. G. Bujalesky. 2000. Cannibalisation of Holocene gravel beach-ridge plains, northern Tierra del Fuego, Argentina. Mar. Geol. 170:105–122. Jelgersma, S. 1996. Land subsidence in coastal lowlands. Chapter 2 in Milliman, J. D. and B. U. Haq (eds). SeaLevel Rise and Coastal Subsidence. Kluwer Academic Publishers, Dordrecht, the Netherlands, 369 pp. Kearney, M. S., and J. C. Stevenson. 1991. Island land loss and marsh vertical accretion rate evidence for historical sea-level changes in Chesapeake Bay. J. Coast. Res. 7:403–415. Kearney, M. S., J. C. Stevenson, and L. G. Ward. 1994. Spatial and temporal changes in marsh vertical accretion rates at Monie Bay: Implication for sea-level rise. J. Coast. Res. 10:1010–1020. King, D. A. 2004. Climate change science: Adapt, mitigate, or ignore? Science 3003:176–177.

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Lambeck, K., and J. Chappell. 2001. Sea level change through the last glacial cycle. Science 292:679–686. Lane, F. C. 1973. Venice: A Maritime Republic. Johns Hopkins University Press, Baltimore, 505 pp. Lauritzen, P., J. Lewinski, and J. J. Norwich. 1986. Venice Preserved. Adler & Adler, Bethesda, MD, 176 pp. Leatherman, S. P., B. C. Douglas, and J. L. LaBreque. 2002. Sea level and coastal erosion require large-scale monitoring. Eos Trans. Am. Geophys. Union 84:13,16. Leatherman, S. P., K. Zhang, and B. C. Douglas. 2000. Sea level shown to drive coastal erosion. Eos Trans. Am. Geophys. Union 81:55–57. Leonard, L. A., A. C. Hine, and M. E. Luther. 1995. Surﬁcial sediment transport and deposition processes in a Juncus roemerianus marsh, West-Central Florida. J. Coast. Res. 11:322–336. Lynch, J. C., J. R. Meriwether, B. A. McKee, F. Vera-Herrera, and R. R. Twilley. 1989. Recent accretion in mangrove ecosystems based on 137Cs and 210Pb. Estuaries 12:284– 299. McCarthy, J. J., O. F. Canziani, N. A. Leary, D. J. Dokken, and K. S. White (eds). 2001. Climate Change 2001: Impacts, Adaptation, and Vulnerability. Cambridge University Press, Cambridge, UK, 1032 pp. McSaveney, M. J., and 6 others. 2000. The 17 July 1998 tsunami, Papua New Guinea: Evidence and initial interpretation. Mar. Geol. 170:81–92. Mei-e, R. 1993. Relative sea-level changes in China over the last 80 years. J. Coast. Res. 9:229–241. Miller, L., and B. C. Douglas. 2004. Mass and volume contributions to twentieth-century global sea level rise. Nature 428:406–409. Milliman, J. D., and B. U. Haq. 1996. Sea Level Rise and Coastal Subsidence: Causes, Consequences, and Strategies. Kluwer Academic Publishers, Dordrecht, the Netherlands, 369 pp. Milliman, J. M., J. M. Broadus, and F. Gable. 1989. Environmental and economic implications of rising sea level and subsiding deltas: The Nile and Bengal examples. Ambio 18:340–345. Mitrovica, J. X., M. E. Tamisiea, J. L. Davis, and G. A. Milne. 2001. Recent mass balance of polar ice sheets inferred from patterns of global sea-level change. Nature 409: 1026–1029. National Research Council (ed.). 1990. Sea-Level Change. National Academy Press, Washington, DC, 234 pp. Nicholls, R. J., and S. P. Leatherman. 1995. Potential impacts of accelerated sea-level rise for developing countries; a discussion. J. Coast. Res. Special Issue 14:303–323. Nicholls, R. J., and S. P. Leatherman. 1996. Adapting to sea-level rise: Relative sea-level trends to 2100 for the United States. Coast. Manage. 24:301–324. Nicholls, R. J., F. M. J. Hoozemans, and M. Marchand. 1999. Increasing ﬂood risk and wetland losses due to global sealevel rise: Regional and global analyses. Global Environ. Change 9:S69–S87.

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Nutalaya, P., R. N. Yong, T. Ghumankit, and S. Buapeng. 1996. Land subsidence in Bangkok during 1978–1988. Pp. 105–130 in Milliman, J. D., and B. U. Haq (eds). Sea-Level Rise and Coastal Subsidence: Causes, Consequences, and Strategies. Kluwer Academic Publishers, Dordretch, the Netherlands, 369 pp. Orviku, K., J. Jaagus, A. Kont, U. Ratas, and R. Rivis. 2003. Increasing activity of coastal processes associated with climate change in Estonia. J. Coast. Res. 19:364–375. Parkinson, R. W., R. D. DeLaune, and J. R. White. 1994. Holocene sea-level rise and the fate of mangrove forests within the wider Caribbean region. J. Coast. Res. 10:1077–1086. Peltier, W. R. 1990. Glacial isostatic adjustment and relative sea-level change. Pp. 73–87 in National Research Council (ed.). Sea-Level Change. National Academies Press, Washington, DC, 256 pp. Pirazzoli, P. A. 1987. Recent sea-level changes and related engineering problems in the Lagoon of Venice. Prog. Oceanogr. 18:323–346. Pirazzoli, P. A. 1991. Possible defenses against a sea-level rise in the Venice area. J. Coast. Res. 7:231–248. Pirazzoli, P. A. 2001. Did the Italian government approve an obsolete project to save Venice? Eos Trans. Am. Geophys. Union 83:217, 223. Raper, S. C. B., T. M. L. Wigley, and R. A. Warrick. 1996. Global sea-level rise: Past and future. Pp. 11–46 in Milliman, J. D., and B. U. Haq (eds). Sea-Level Rise and Coastal Subsidence: Causes, Consequences, and Strategies. Kluwer Academic Publishers, Dordrecht, the Netherlands, 369 pp. Reed, D. J. 1990. The impact of sea-level rise on coastal salt marshes. Prog. Phys. Geog. 14:465–481. Reed, D. J. 1995. The response of coastal salt marshes to sealevel rise: Survival or submergence? Earth Surf. Process. Landf. 20:39–48. Sabhasri, S., and K. Suwarnarat. 1996. Impact of sea level rise on ﬂood control in Bangkok and vicinity. Chapter 18 in Milliman, J. D. and B. U. Haq (eds). Sea-Level Rise and Coastal Subsidence. Kluwer Academic Publishers, Dordrecht, the Netherlands, 369 pp. Semtner, A. J. 1995. Modeling ocean circulation. Science 269:1379–1385. Sestini, G. 1992. Implications of climatic changes for the Po Delta and Venice Lagoon. Pp. 428–494 in Jeftic, L., J. D. Milliman, and G. Sestini (eds). Climatic Change and the Mediterranean. Edward Arnold, London, 673 pp. Smith, S. V., and R. W. Buddemeier. 1992. Global change and coral reef ecosystems. Ann. Rev. Ecol. Syst. 23:89– 118. Snedaker, S. C. 1995. Mangroves and climate change in the Florida and Caribbean region: Scenarios and hypotheses. Hydrobiologia 295:43–49. Snedaker, S. C., J. F. Meeder, M. S. Ross, and R. G. Ford. 1994. Discussion of Ellison, Joanna C. and Stoddart,

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David R., 1991. Mangrove ecosystem collapse during predicted sea-level rise. Holocene analogues and implications. J. Coast. Res. 7: 151–165, 10:497–498. Spencer, N. E., and P. L. Woodworth. 1993. Data Holdings of the Permanent Service for Mean Sea Level. Permanent Service for Mean Sea Level. Bidston, Birkenhead, UK, 81 pp. Stiros, S. C., and 5 others. 2000. Seismic coastal uplift in a region of subsidence: Holocene raised shorelines of Samos Island, Aegean Sea, Greece. Mar. Geol. 170:41–58. Sundquist, E. T. 1990. Long-term aspects of future atmospheric CO2 and sea level changes. Pp. 193–207 in National Research Council (ed.). Sea-Level Change. National Academies Press, Washington, DC, 256 pp. Vella, C., and M. Provansal. 2000. Relative sea-level rise and neotectonic events during the last 6,500 yr on the southern eastern Rhône delta, France. Mar. Geol. 170:27–40. Vreugdenhil, C. B., and H. G. Wind. 1987. Impact of Sea Level Rise on Society: Framework Analysis and Recommendations. A. A. Balkema, Rotterdam, the Netherlands. Warrick, R. A., C. Le Provost, M. F. Meier, J. Oerlemans, and P. L. Woodworth. 1996. Changes in sea level. Chapter 7 in Houghton, J. T., L. G. Meira Filho, B. A. Callander, N.

Harris, A. Kattenberg, and K. Maskell (eds). Climate Change 1995. Cambridge University Press, Cambridge, UK, 572 pp. Wood, M. E., J. T. Kelley, and D. F. Belknap. 1989. Patterns of sediment accumulation in the tidal marshes of Maine. Estuaries 12:237–246. Wray, R. D., S. P. Leatherman, and R. J. Nicholls. 1995. Historic and future land loss for upland and marsh islands in the Chesapeake Bay, Maryland, U.S.A. J. Coast. Res. 11:1195–1203. Wyrtki, K. 1990. Sea level rise: The facts and the future. Pac. Sci. 44:1–16. Wyrtki, K., and S. Nakahoro. 1984. Monthly Maps of Sea Level Anomalies in the Paciﬁc 1975–1981. Hawaii Institute of Geophysics Report HIG-84-3. Hawaii Institute of Geophysics, Hawaii. Yamamoto, S., and A. Kobayashi. 1986. Groundwater resources of Japan with special reference to its use and conservation. IAHS-AISH Publication 151:381–389. Zhang, K., B. Douglas, and S. Leatherman. 2002. Do storms cause long-term beach erosion along the U.S. east barrier coast? J. Geol. 110:493–502. Zorzi, A. 1989. Una Città, una Repubblica, un Impero: Venezia 697–1797. Arnoldo Mondadore Editore, Milan, 227 pp.

Chapter 4 Alteration of freshwater discharges

Derelict ships stranded in the now-exposed bottom of the Aral Sea. Photo by Roman Jashenko; from http://www.kz/gallery/aral/cha12.html.

Introduction Discharge of fresh water from land to sea across the diverse shorelines of the world has been altered locally owing to human activities on speciﬁc watersheds, as well as being affected by globalscale climatic changes. It is difﬁcult to generalize about human effects on discharges of fresh water to coastal environments—there are probably as many different responses as there are rivers. Human use of fresh water most often reduces the volume of ﬂow, which results in alterations to the coastal environments that depend on the

water, sediment, and nutrients brought from the land. In contrast, the global-scale effects might be related to changes in global atmospheric conditions, which also change freshwater delivery to coastal environments, with consequences that were unforeseen until recently. Human water use and climatic effects are not spread out evenly across the world’s surface, so that effects of freshwater interception are more severe in some places than in others. In some parts of the world, human interception of water can become a dominant agent of coastal change. To demonstrate the full impact of human ability to intercept freshwater delivery to a sea, there is no

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more compelling example than that of the disastrous history of the Aral Sea during the 20th century.

A case history: freshwater interception in the Aral Sea1 The Aral Sea is a large land-locked water body in Central Asia. The watershed includes the present-day Republics of Kazakstan, Uzbekistan, and Turkmenistan. To many of us in the Western Hemisphere, the area resounds of ancient adventure and conquest—Alexander, Marco Polo, Genghis Khan, Huns, Tartars, Tamerlane, and the Golden Horde all passed through this area. The place names still provide a sense of ancient and exotic glamour; double-humped Bactrian camels bearing oriental silks, precious spices, rich bukhara rugs through Samarkand, Tashkent, and the other cities along the trade routes: the riches of the then-known world. The reality has been otherwise. In this historically arid area of the world, people through the centuries depended on two rivers, the Amudarya2 and Syrdarya, to supply water to support the mixed agriculture that furnished their food supply. In the 1920s, the government of the Soviet Union made the decision to become self-sufﬁcient in the production of cotton, and fostered the cultivation of cotton in the fertile soils of the plains of Central Asia. As it happens, cotton is a crop that requires signiﬁcant water, so the plan included intensive irrigation efforts, which required increased interception of river water then ﬂowing into the Aral Sea to grow the new crop. From the start, it was clearly realized by all parties that the result would be a sharp diminution in the size of the Aral Sea. The decision to

proceed was based on a considered balancing of expert opinions. Some experts estimated the economic gains from cotton cultivation versus the tangible beneﬁts derived from the Aral Sea as it then existed. They concluded that the value of a cubic meter of water used for irrigation was far larger than the value furnished by that same volume of water delivered to the Aral Sea.3 Other experts believed that the reduction in the Aral would have minor or no environmental effects on the surrounding lands. A few more cautious scientists disagreed, citing estimates of the likely drastic environmental changes, made as early as 1927, but their opinion went unheeded (Glantz 1999). As is almost the rule, short-term economic estimates won over potential environmental concerns. By the mid-1950s hundreds of kilometers of unlined canals were dug from the two rivers into the surrounding dry plains to supply water for growing cotton, as well as other crops. This made it possible to signiﬁcantly increase the area of land under irrigation (Fig. 4.1). The water management decisions were also accompanied by intensive use of pesticides, as required by the cotton crop. The strategy worked, and during the 1960s to 1990s, the Soviet Union became the leading producer of cotton, ahead of the United States and China, the next largest producers in the world. By the 1970s and 1980s the world began to perceive the consequences of the push for what the Soviet government described as “white gold”. The land area devoted to agriculture and under the inﬂuence of managed water works had increased notably (Fig. 4.2). The diversion of river water for agriculture, as was expected, diminished river ﬂow into the Aral Sea (Fig. 4.3 top panel). The ﬂow of water through the Syrdarya to the Aral Sea declined during the 1960s. In 1975, no water from this river reached the sea. Water 3

1

Most of this material is taken from Micklin (1988), Létolle and Mainguet (1993), Micklin and Williams (1996), Glantz (1999), and Stone (1999), who review information on the transformations suffered by the Aral Sea. 2 The historically minded might best recognize this river as the Oxus, a waterway that became a notable historical boundary for Alexander, the Persians, and the Arabs, and passed through an area later to be overrun by the Russian tsars.

In most cases of environmental use choices, options are taken in the absence of analyses of the pros and cons, so this case was unusual. Regardless of the merit of the comparison used here, however, the case illustrates the dilemma of using economic worth as the currency with which to evaluate environmental issues. In virtually all cases, the result will be to go with the option that appears most valuable, or for which someone is willing to pay the highest price. It will seldom be the case that environments are conserved using this means of comparison.

ALTERATION OF FRESHWATER DISCHARGES

Irrigated area (ha × 106)

8 7 6 5 4 3 1920

1940

1960

1980

2000

Year

Figure 4.1 Increase in irrigated land in the Aral Sea basin, 1930–1990. Data from Zonn (1999).

from the Amudarya ceased to ﬂow to the sea during 1982. During the 1990s some ﬂow was restored as favorable weather lowered evaporation, and increased snowfall in the Pamir Mountains helped feed the rivers (Bortnik 1999). The completion of canal construction after the mid-1990s drew even more water off the Amudarya. As a result of lower river ﬂow, the water level in the Aral Sea decreased (Fig. 4.3 second panel). Decadal-scale changes in water level of the Aral were considerable through historical times; ﬂuctuations of 3–4 m were thought to have taken place several times between the 1780s and 1900 (Létolle & Mainguet 1993). These “background” changes in sea level, however, were considerably smaller than the reductions in water level seen after 1960. The effect of latter-day water withdrawal therefore exerted far more inﬂuence than the action of climatic changes. The surface area and volume of the Aral also decreased (Fig. 4.3 third and bottom panels). The reduction in area became so prominent as to be readily visible in satellite images; maps show the drastic reduction of surface area and exposure of the Aral Sea’s ﬂoor (Fig. 4.4). As volume decreased, salinity of the water in the Aral Sea increased beyond 40‰ (Fig. 4.3 bottom panel), levels that exceed the salinity of the oceans. The ecological effects that followed were dramatic. Coastal habitats (wetlands, lagoons,

81

shoals, and littoral zones) important for many species of plants and animals nearly disappeared. As just one example of the change in habitats, we can take the reed-covered wetlands. A celluloseproducing industry depended on the 700,000 ha of reeds growing on coastal wetlands to furnish its raw material. This enterprise cannot now ﬁnd enough reeds, because only 30,000 ha of wetlands remain, and this remnant only survives through managed ﬂooding of restricted areas with eversmaller water supplies. Large portions of the previous Aral Sea bottom were exposed as the 20th century ebbed. The changed balance between water and soil surface altered regional weather patterns. Summers in the region became warmer, and winters colder, with a shorter growing season, too short for cotton to grow properly. The frequency of major dust storms has increased from one every 5 years during the 1950s to about ﬁve per year now. Salts (as well as pesticides and other contaminants) from the more than 30,000 km2 of newly exposed sea bed were spread by winds. The salts carried by wind salinized surrounding soils, making millions of hectares of soil unproductive, and contaminating potable water and groundwater. The salts spread across the globe: Aral Sea salts have been identiﬁed in air over India as well as the Arctic, Paciﬁc, and Atlantic Oceans. Almost all the surface waters and sediments have become contaminated with agricultural chemicals (Zholdasova 1999). The intensive use of agricultural chemicals has left residues in the entire area. In the arid climate, soil is easily blown away, and the soil particles, with the adsorbed contaminants, are spread widely by the winds. In addition, as irrigation water drains back into the network of canals, it carries pesticide residues from the agricultural ﬁelds. The reduced habitat areas, the increased salinity, and the contamination by agricultural chemicals severely altered the fauna of the Aral Sea and nearby waters (Zholdasova 1999). The sharp changes can be seen in the dramatic change in makeup of the ﬁsh present in the Aral Sea, and in the diminished harvests. Twenty or so species of ﬁsh native to the Aral became extinct by the mid-1980s, except for a salt-tolerant species of

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Syrdarya Aral Sea

Amudarya Samarkand Bukhara

Syrdarya

Amudarya Bukhara

Samarkand

0

100

200 300 km

Figure 4.2 The Aral Sea area before (top) and after (bottom) the development of agriculture. The maps show the increase in irrigated land, construction of canals, and shift in areas of the Aral Sea as well as the creation of other water bodies. Adapted from Létolle and Mainguet (1993).

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ALTERATION OF FRESHWATER DISCHARGES

River flow (km3 yr−1)

80

40

0 1920

1940

1960

1980

2000

1940

1960

1980

2000

1940

1960

1980

2000

Water level (m)

60 50 40 30 1920

50

30 1920

Figure 4.3 Changes in river ﬂow, water level, surface area, and volume and salinity of the water in the Aral Sea, 1920–1990. Data from Gleick (1993, table F.20, compiled by P. P. Micklin); salinity data from Gleick (1992).

Volume (km3) ( )

1,200

40 30

800

20 400 0 1920

stickleback. The ﬁshing industry in the Aral employed about 60,000 people in the 1950s. The ﬁsh harvest from the Aral ﬂuctuated between 25,000 and 45,000 metric tons per year from 1930 to 1960; after 1960, the ﬁsh harvest precipitously collapsed (Fig. 4.5). Before the mid-1960s, the ﬁsheries of the Aral were a valuable part of the region’s economy. The ﬁsheries were based on about 12 species of commercial importance, mainly local species of carp and perch. After 1983 commercial ﬁshing ceased. Former ﬁshing villages were abandoned, and ﬁshing and other vessels were stranded many kilometers from

10

Salinity (‰) ( )

Surface area (km2 ×103)

70

0 1940

1960

1980

2000

Year

shore (see Chapter 4 frontispiece). As a last-ditch effort to maintain workers in the canning plants that previously had processed Aral ﬁsh, ﬁsh were imported from the Arctic, Paciﬁc, and Baltic. This became economically unsustainable and the industry stopped altogether in 1994 (Glantz 1999), creating unemployment and economic decline of the former coastal cities around the Aral. Efforts had been made since the 1920s to foster the Aral ﬁsheries, principally by the introduction of many alien species, most of which failed to establish themselves. The spectacular collapse of the ﬁshing industry in the 1960s and 1980s

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1998

1978

2010

Fish harvest (metric tons × 103 yr−1)

1960

CHAPTER 4

50 40 30 20 10 0 1930

1950

1970

1990

Year

Figure 4.5 Time course of the ﬁsh harvest from the Aral Sea, 1930–1992. Data from Létolle and Mainguet (1993), Zholdasova (1999).

Figure 4.4 Sketches of the reduced water surface area of the Aral Sea in 1960, 1978, 1998, and that predicted for 2010. Adapted from Micklin and Williams (1996).

stimulated further efforts at introductions. Many species of ﬁsh, crab, mollusk, and shrimp were also introduced accidentally along with the intended ﬁsh species. All told, 18 ﬁsh species have been introduced into the Aral, and four of these now dominate the ﬁsh fauna of the Aral Sea. The Aral ﬁshery therefore not only has collapsed, but the remaining assemblage of species is drastically different from what was present previous to the campaign to intercept fresh water. The changes in weather, plus the salinization of soils and scarcity of clean water, have, in a sad and ironic turn of events, led to decreased cotton yields. Cotton yields around the Aral during 1999 were 2 tons ha−1, less than one-third the yields achieved in Israel, a country with a similarly arid climate. Attempts to introduce salt-tolerant strains of cotton have failed to catch on because of suspicion from a nearly desperate community of growers.

The lowered level of the Aral was accompanied by an increased depth of the groundwater table around most of the region, and by contamination of the groundwater with salts. This meant that supplies of drinkable water have diminished. In some regions, however, groundwater rose as a result of the irrigation needed to grow crops in the desert (Orlovsky 1999). After years of irrigation, the amounts of water used resulted in the groundwater table becoming shallower by 1 to 5 m. Since the groundwater was salty, salts diffused upwards as a result of intense evaporation and made soils unsuitable for crops.4 The environmental changes in the Aral Sea region affected people in many ways, including increases in certain—but not all—human diseases (Fig. 4.6). The incidence of typhoid fever and viral hepatitis in a region near the Aral Sea (KyzykOrda in Fig. 4.6) increased after the Aral Sea transformations compared to incidence across the whole of Kazakhistan. The public health effects probably were associated with political indifference and lowered economic and social 4

Salinization of irrigated soils is an unfortunately common phenomenon. Ghassemi et al. (1995) estimated that the proportion of irrigated land affected by salinization might reach 21% worldwide, with greater percentages in certain countries such as Egypt (30%), Uzbekistan (60%), and Turkmenistan (80%).

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ALTERATION OF FRESHWATER DISCHARGES

100

TYPHOID FEVER

80

60

Number of cases per 100,000 people

40

Kyzyk-Orda

20

Kazakhistan

0 1970 2,800

1974

1978

1982

1986

VIRAL HEPATITIS

2,400 2,000 1,600

Kyzyk-Orda

1,200

Figure 4.6 Time course of two infectious diseases, typhoid fever and viral hepatitis, in the period 1970–1987. Lines show data for the whole of the Kazakhistan Republic and within the Kyzyl-Orda district, which is nearer the Aral Sea. Data modiﬁed from Elpiner (1999).

800 400 Kazakhistan 0 1970

conditions, as well as poorer diet and worsening air and water quality resulting from the environmental changes. There can be no accurate estimates of the economic losses that resulted from the decision to proceed with the cotton cultivation plan. What is certain is that “expert” valuation of the relative beneﬁts of that cubic meter of water was a shortsighted exercise that should serve as an object lesson. We can almost be sure that any such economic-based valuation will underestimate environmental damage, and that, in addition, the non-economic imponderables not included in the decision may turn out to have far more (some-

1974

1978 Year

1982

1986

times economic!) consequences than could be imagined. Toward the end of the 20th century the dimensions of the disaster affecting the Aral Sea became evident to local and international authorities (Bortnik 1999). The Soviet Union even printed a stamp that commemorates the issue (Fig. 4.7). Remediation efforts have been discussed or planned, but resources are limited at best, even with the assistance of international agencies. Restoration has, realistically, been limited to engineering works that will capture smaller volumes of fresh water within the limited aquatic environments that remain. No one will

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Figure 4.7 Commemorative stamp printed by the Soviet Union, showing a ship stranded on the dry bed of the Aral Sea. The lettering reads “Aral, region of ecological disaster”. The inset features a saiga antelope, endemic to the region, whose numbers have become sufﬁciently low that commercial hunting has been banned since 1992 (Krutov 1999).

insist on the full release of river water into the Aral, because over 90% of the harvest of all crops in the region still comes from irrigated lands, and the main supply of fresh water is still the two rivers. The importance of cotton to the republics bounding the Aral was evident during the Soviet era: the National Seals of four of the ﬁve cotton-growing republics of the Soviet Union bore a symbolic representation of cotton (Létolle & Mainguet 1993). In spite of the political upheaval that followed the dissolution of the Soviet Union, the fundamental reliance on cotton remains, as suggested by the continued inclusion of cotton in the changed seals of the new post-Soviet republics (Fig. 4.8). With critical conditions in the agriculturebased economies of the countries surrounding the Aral Sea5, it is implausible to expect early 5

As examples of the relatively impecunious state of the cottongrowing republics, we can examine some data from Létolle and Mainguet (1993). In the relatively more prosperous Ukraine and Russia, 1,076 and 295 TW h−1 of electricity were produced, but only 12–90 TW h−1 in the cotton republics. Industrial production was 3,848 and 3,085 rubles per person per year in the more prosperous two countries, and only 1,100–1,684 in the cotton republics. Infant mortality during 1988 was 19 and 15‰ versus 29–58‰.

action, or much success in restoration of the Aral Sea. The conditions in the Aral Sea have to some extent been repeated in the nearby Caspian Sea (Aleem 1972). There, the fresh water brought in by the Volga River has diminished from agricultural water use. The level of the Caspian has dropped by about 3 m (Bird & Koike 1986), its area has shrunk, exposing tens of kilometers of sea ﬂoor all around the shore, and salinity has increased. These examples constitute the worstcase evidence for the fearful impact of human interception of fresh water on aquatic and environments and more.

Human demand for fresh water Current trends Throughout the 20th century demand for fresh water increased faster than human populations (Fig. 4.9): the 10-fold increase in water use was larger than the ﬁve-fold increase in number of people (L’vovich & White 1990). Agricultural uses consumed by far the most fresh water (Fig. 4.9; Table 4.1). About 63% of the fresh

ALTERATION OF FRESHWATER DISCHARGES

87

Figure 4.8 Seals of Turkmenistan (top) and Uzbekistan (bottom), before (left) and after (right) the dissolution of the Soviet Union. From http://www.ur.ru/∼ap/ussr, http://www.oxuscom.com/ uz-seal.htm, and http://www. turkmenistan-online.com/ turkmen_txt.html.

water intercepted by people was devoted to agricultural practices; industrial and municipal uses were smaller, as was the amount stored in reservoirs.6 The increased pressure for agricultural use of water derived largely from the large expansion of land areas throughout the world irrigated for crop production (Fig. 4.10). The increase in irrigated land was particularly fast during the mid-20th century, and although still climbing has slowed more recently. Actually, although agricultural water use increased across the 20th century, its increase was considerably smaller than the increases for industrial and municipal purposes, judging from the ratios of water use during the year 2000 relative to use during 1900 (Table 4.1). In the near future, urban populations will swell, particularly in developing countries 6

Not all intercepted water is available for our use. About 20% of water intercepted is lost to evaporation and other unrecoverable losses during the course of human activities (Postel 1998).

and in China, and, as noted in Chapter 1, more and more of the world’s population will become urban (Postel 1999). The forecasts claim that the use of fresh water by cities and industry will climb to perhaps 27% of total freshwater supply; this increase will have to come at the expense of water for irrigation, adding to the pressures on worldwide water supply.7 Future concerns about water supply will need to increasingly consider conservation and management of water 7

Uneven distribution of wealth and use of water exacerbate the issues. For example, Postel (1999) reports that there are 550 golf courses, plus another 530 already planned, in urban areas of relatively prosperous countries of the Paciﬁc Rim (Malaysia, Thailand, Indonesia, South Korea, the Philippines, and so on). Much the same is true for desertic regions of the USA (southern California, Arizona, Nevada). These discretionary land uses require extremely high rates of irrigation to remain verdant, and do so at the same time of year that crops most require water. Such use of land usually replaces agricultural ﬁelds, and often takes place just across the border from another political unit where depauperate sharecroppers might be trying to raise a subsistence crop of corn or rice.

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CHAPTER 4

6

Water use (km3 × 103 yr−1)

Total

4 Agricultural

2 Industrial

0 1900

Municipal Residential 1920

1940

1960

1980

2000

Year

Table 4.1 Worldwide freshwater use by human activities during the year 1900, and predicted for the year 2000. Data from Shiklomanov (1993). Annual volume of water used (× 103 km3 yr−1) Use of water

During 1900

Ratio of water use in During 2000 2000 : 1900

Agriculture Industry Municipal supply Reservoirs

525 37.2 16.1 0.3

3,250 1,280 441 220

6 34 27 733

Total

579

5,190

9

20th century, reaching a peak during the 1970s (Fig. 4.11).8 Toward the end of the century, construction of large dams and reservoirs was less enthusiastic, but the cumulative total number of dams [about 40,000 worldwide (ICOLD 1988)] is impressive, and has led to a greater than 700fold increase in this water use (Table 4.1). Global freshwater impoundment is therefore quite large. The capacity of reservoirs behind dams amounts to about 9% of the world’s river ﬂow (10% in Europe, 22% in North America); the area of reservoirs, pooled across the world, adds up to a 8

use by industry and urban areas. It is no exaggeration to think that we are facing a freshwater crisis as we turn into the 21st century. Role of waterworks People have devised many ways to intercept water. In many cases, we simply draw water from rivers, streams, lakes, and groundwater. Probably the largest single method by which we manipulate and remove water from natural environments is through the use of dams and reservoirs. The construction of large dams and reservoirs accelerated during the middle decades of the

Figure 4.9 Human use of fresh water during the 20th century for the world. Time courses include total freshwater use, plus the uses for agricultural, industrial, municipal, and residential purposes. Data from Gleick (1993).

Dams, however, play important roles in people’s welfare. Dams are built largely for production of hydroelectric power; clearly, the reservoirs are also useful as sources of water for irrigation, municipal, and industrial consumption, as well as non-consumptive uses such as swimming, sailing, and ﬁshing. Although construction of large dams has slowed to some extent, the pressure for more hydroelectric power demanded that this production increase from almost 1,500 × 109 kW-h yr−1 during 1975 to more than 2,000 × 109 kW-h yr−1 in 1990 (Gleick 1993, table G.2). The overwhelming economic importance of dams is evident from the fact that hydroelectric power contributes about 24% of the world’s electricity. This is about twice the contribution from nuclear power plants, and only 2.6 times less than the energy from plants powered by fossil fuels (Gleick 1993, table G.4). In addition, the reservoirs that form upstream from dams play powerful roles in human well-being. For example, the water in Lake Nasser, the reservoir behind the Aswan High Dam on the Nile River, provided the irrigation water that is credited with avoiding massive famines in Egypt during severe droughts in the early 1970s and mid–1980s (Postel 1999). These impressive positive contributions should not be forgotten in discussions of the negative consequences of dam building.

89

ALTERATION OF FRESHWATER DISCHARGES

Irrigated area (ha × 106)

200

Figure 4.10 Increase in irrigated area for the world, 1900–1989. The percent increases across several periods are shown along the x axis. Data from Gleick (1993, table E.2).

24

100

0 1900

Dams (× 103)

18

12

6

0 1920

1940 1960 Year

1920

1940

2.1%

1960

2.4%

1.1%

1980

Year

Large dams in world Dams constructed per decade

<1900

4.3%

1.9%

1980

2000

Figure 4.11 Number of large dams, and large dams constructed per decade, in the world. Large dams are deﬁned as having a height of > 15 m from the foundation to crest. Many more smaller dams exist. Data from ICOLD (1988), cited in Rosenberg et al. (2000).

surface equivalent to that of France (L’vovich & White 1990). It is therefore not surprising that human waterworks can signiﬁcantly alter the volume of surface freshwater ﬂow that some rivers take to the sea (Fig. 4.12). Such waterworks include canals that divert river water for irrigation or other uses, and dams that allow impoundment

of river water in reservoirs. Removal of water for irrigation along the course of a river may lower ﬂows markedly, as in the case of the Syrdarya River that ﬂows into the Aral Sea (Fig. 4.12 top panel). The building of the Aswan High Dam during the early 1960s on the Nile River also sharply lowered the magnitude of discharge (Fig. 4.12 second panel). Perhaps more typical is the situation in the Columbia River, where the cumulative construction of dams across smaller tributaries during the past two centuries gradually diminished ﬂow by somewhat less than 30% (Fig. 4.12 third panel).9 Flow through the Columbia is currently regulated so that both lows and highs are nearer the mean ﬂow. Perhaps the example that best demonstrates human control of river discharge is that of the Colorado River (Fig. 4.12 fourth panel). The watershed of the Colorado River, the largest hydrological basin in southwestern USA, sup9

In discussions of reduction of ﬂows we focus on mean ﬂows. It should be noted that in all cases there are contingencies of weather that drastically alter ﬂows from one year to another. For instance, maximum ﬂows in the Columbia River reached four times mean ﬂows in the 1890s. Although the ﬂow through the Columbia was managed in the 1990s, peak ﬂows still reached more than two times mean ﬂows. In addition the maximum ﬂows themselves varied at least by two-fold from year to year. Minimum ﬂow has remained more uniform at about one-half mean annual ﬂows (Sherwood et al. 1990).

90

CHAPTER 4

SYRDARYA

1.6

0.8

0 1950

1960

1970

1980

1990

14 NILE

10 6 2

River flow (m3 × 103 s−1)

1961

1963

1965

1967

1969

1971

1973

COLUMBIA

8 6 4 2 1870

1890

1910

1930

1950

1970

1990

1.2 COLORADO

0.8 0.4 0 1900

1920

1940

1980

1960

1.2 BURNTWOOD 0.8 0.4 0 1950

1960

1970

1980

Year

plies water to more than 20 million people and to a very large area of irrigated agricultural land (Carriquiry & Sanchez 1999). Annual ﬂow in the Colorado supplies 20.3 × 109 m3 yr−1 of fresh water, and the reservoirs behind several dams

1990

Figure 4.12 Top panel: discharge of the Syrdarya into the Aral Sea, 1950–1985 (from Vörösmarty & Sahagian 2000). Second panel: discharge of the Nile River into the Mediterranean, 1960–1973 (from El Din 1977). Third panel: ﬂow of the Columbia River at The Dalles, Oregon, 1887–1985 (from Sherwood et al. 1990). Fourth panel: discharge of the Colorado River at the border between the USA and Mexico, 1905–1986 (data from the US Bureau of Reclamation, Yuma, Arizona, cited in Glenn et al. 1996). Bottom panel: discharge of the Burntwood River, 1958–1990 (from Vörösmarty & Sahagian 2000).

have a usable storage capacity about four times the volume of the annual ﬂow. Management of ﬂow through the Colorado River began in the late 1800s, and has continued —on an increasing scale—to the present day

91

ALTERATION OF FRESHWATER DISCHARGES

Flood discharge (km3 yr−1)

(Glenn et al. 1996; Carriquiry & Sanchez 1999). Water from the Colorado was diverted through canals for farming as early as 1896. During 1905– 1907 a break in a canal inadvertently diverted ﬂow of the Colorado and ﬂooded the low-lying area now known as the Salton Sea and environs. This mishap prompted increased management efforts. The ﬁrst Colorado River Compact was signed in 1922, and allotted river water to states in the upper and lower sub-basins. The control of water ﬂow in the lower basin was through regulation of Lake Mead, a very large reservoir formed behind Hoover Dam, which started operation in 1934. An international treaty signed, not entirely without political tensions, in 1944 decreed that 10% of the discharge through the Colorado would cross the border into Mexico, except during droughts. The Colorado River was further regulated after construction of the Glenn Canyon Dam in the 1950s, which allowed a total interception of 96% of the river’s ﬂow. Several other dams were completed during the 1960s, and these brought near-complete control of water ﬂow through the Colorado. Flow to the Mexican border is therefore much less than it was historically, and only occasionally, when ﬂoods ﬁll reservoirs up-river and water is released downstream (Fig. 4.12 fourth panel), is there ﬂow comparable to historical discharges. Mexico has built the Canal Central within its territory, which diverts water that crosses the border. This water is used in popula-

Figure 4.13 Peak discharge from the Nile River during ﬂood stages since 1870. Data from R. E. Quelennec, cited in Sestini (1992).

tion centers and for irrigation within Mexico. The result of all this is that, except in extraordinarily wet years, the present pressure for water use—for irrigation and, increasingly, for municipal use—does not allow any Colorado water to ﬂow to the sea (Glenn et al. 1996; Carriquiry & Sanchez 1999). To evaluate discharge data such as shown in Fig. 4.12, it seems relevant to cast the recent changes in a somewhat longer time scale. For example, the peak discharge data from the Nile River (Fig. 4.12 second panel) can be compared with the ﬂow data included in Fig. 4.13. The effects of the Aswan High Dam can be seen on the far right-hand side of Fig. 4.13, but we need to realize that throughout its history the ﬂow through the Nile has been characterized by marked changes at interannual as well as secular scales. In the 19th century, for example, ﬂow was nearly double that during the 20th century, presumably because of greater rainfall in the Sahel. Drastically lower peak ﬂows comparable to ﬂows in recent years were contemporaneous with shorter-term Sahelian droughts during 1911–1915 and 1941–1944, for instance (Fig. 4.13). Although the construction of the Aswan Dam on the Nile clearly lowered water ﬂow through the river, coastal environments such as the Nile delta have been subject to dramatic changes throughout their history; the present circumstance is therefore not a completely new condition.

120 100 80 60 40 1880

1900

1920 Year

1940

1960

92

CHAPTER 4

Table 4.2 Worldwide annual ﬂows of fresh water from and into land and oceans, and transport from land to oceans via ﬂuvial, groundwater, and glacial meltwaters. Data from Shiklomanov (1993); summarized from several sources. Range of estimates (× 106 km3 yr−1)

Midpoint of estimates (× 106 km3 yr−1)

Precipitation on land Evaporation from land

99–119 63–73

109 68 Difference: +41

Water transport from land to oceans Rivers and streams Groundwater Glacier melt

33.5–47 27–45 1.6–12 1.7–4.5

Precipitation onto oceans Evaporation from the oceans

320–458 383–505

Magnitude of human use of water It seems evident, then, that human activity is one among diverse global mechanisms that can be of sufﬁcient magnitude as to prompt signiﬁcant environmental changes. Although irrigation and dam construction may have slowed somewhat towards the end of the 20th century, we still require far more water today than we did in 1900. As we move into the 21st century is the amount of fresh water used by people large or small relative to the magnitude of other major components of the world’s hydrological cycle? To address that question, we need a look at the world’s water balances. Table 4.2 includes estimates of the amount of fresh water falling on and leaving the land and sea. To make the comparisons visually easier, we can use the rightmost column of Table 4.2, which shows the midpoint of the ranges of estimates made by various authors. Freshwater precipitation from the atmosphere onto land exceeds evaporative losses by about 41 × 106 km3 yr−1. This excess volume is then available to ﬂow away from land, mainly through rivers. The difference between precipitation and evaporation on land is remarkably similar to the estimate of fresh water transported from land to sea (40.3 × 106 km3 yr−1). We

40.3 36 6.8 3.1 383 444 Difference: −61

can add the freshwater contribution from land to the amount of direct precipitation onto the sea surface, to get a total of 423 × 106 km3 yr−1 of fresh water that enters the world’s oceans. This total input to the oceans is only 21 × 106 km3 yr−1 less than the 444 × 106 km3 yr−1 that evaporates from the oceans. Perhaps this difference is within the uncertainty of the various estimates. Alternatively, perhaps the difference comes from an underestimation of groundwater ﬂow (set at 6.8 × 106 km3 yr−1 in Table 4.2). It has been easy to underestimate groundwater ﬂuxes; recent work with radium isotope tracers suggests that there may be more groundwater ﬂow than was realized (Moore 1996, 1997). Water exchanges between atmosphere and land, and between oceans and atmosphere, are large. Although the magnitude of freshwater ﬂow from land to sea might be much smaller than the other components in the world’s hydrological cycle, it can be meaningfully estimated. This estimate provides the context to answer our question about the relative magnitude of human water use. The 5.2 × 106 km3 yr−1 thought to have been used by people worldwide during the year 2000 (see Table 4.1) is small compared to ﬂuxes across the land and sea surfaces. It is, however, about 14% of the ﬂow of all the rivers in the

93

ALTERATION OF FRESHWATER DISCHARGES

Mediterranean Sea Figure 4.14 Contours of salinity of surface waters off the Nile delta, before completion of the Aswan High Dam. After the Aswan High Dam began operation, the salinity of the entire area off the delta remained at 38–39‰, with no evident broad-scale freshening. Adapted from El Din (1977).

39

29

31

35 33

37

Burullus Lagoon Alexandria

Manzalah Lagoon

Rosetta Branch

world. That seems an impressive ﬁgure: imagine a ﬂow equivalent to more than one in 10 units of the water from all rivers and streams on earth. Consider further that this ﬂow is increasing.

Consequences of reduction in freshwater discharges In the case of the Aral Sea we saw surprisingly intensive local and widespread near-global consequences. The Aral crisis is clearly a worst-case scenario. Here we review several other examples to obtain a more general view of the likely consequences of lower freshwater ﬂow into coastal environments. Changes in the down-river transport of water inevitably also change delivery of nutrients and sediments, and it is difﬁcult to separate the effects of water, sediments, and dissolved nutrients. Lower ﬂows of fresh water through estuaries alter salinity and circulation of the receiving coastal waters, and these in turn change the conditions for the biota, and for economically important species in the coastal environments that receive the terrestrial inputs. The changes may even have public health consequences. Alteration to salinity regimes and circulation The effects of human water use on salinity and circulation of coastal waters might be applied to very different spatial scales. Below I use the examples of the Nile and Colorado deltas and Siberian rivers

Bardaweel Lagoon

Damietta Branch

to show a range from reasonably local- to possibly global-scale effects of fresh-water interception. Before the Aswan High Dam was completed, about 38% of the Nile’s total freshwater ﬂow ﬂowed through the Nile delta and reached the Mediterranean. Freshwater outﬂow from the Nile River historically lowered the salinity of sea water up to about 80 km from the Egyptian coast, and to depths of about 150 m (El Din 1977). After completion of the construction of the Aswan High Dam, as well as increased use of Nile water for irrigation, water ﬂow from the Nile decreased (see Fig. 4.12 second panel). This decrease was so large that, even during peak seasonal Nile ﬂow, Mediterranean water off the Nile delta became considerably saltier (around 39‰) than before the Aswan Dam was completed (< 30‰) (Fig. 4.14). The volumes discharged at the Aswan Dam (Vörösmarty et al. 1997) were larger than the ﬂows depicted in Fig. 4.12 (second panel). This is because more than 60% of the water discharged from the dam was lost by evaporation and seepage into aquifers, and was drawn for irrigation before the Nile reached the Mediterranean (Sestini 1992). Before the Aswan High Dam was built, waters moved across the Nile delta area in a typical estuarine circulation, with a fresher upper layer on top of the water column ﬂowing outward most of the time, underlain by a deeper saltier layer of in-ﬂowing Mediterranean water below. After the dam was completed, the bulk of the water column and ﬂow was made up of Mediterranean water, with a thin layer of “fresh water” on top.

94

CHAPTER 4

Toward the end of the 20th century, this thin layer of “fresh water” that crossed the delta of the Nile and ﬂowed into the Mediterranean consisted mainly of discharged domestic waste water, municipal sewage efﬂuent, and agricultural runoff (El Din 1977; Stanley & Warne 1993). The dearth of fresh water meant that saltier water was present further up-estuary than before (Stanley 1988).10 The lack of freshwater inﬂow into the delta of the Colorado River also led to considerable shifts in salinity and circulation regimes. Under the powerful inﬂuence of the 12 m high tides of the northern end of the Sea of Cortez, the saltwater wedge brings sea water 35 km up the Colorado estuary, and all environments are exposed to much saltier conditions than existed historically. Model studies suggest that before human intervention in the water discharge of the Colorado, the ﬂow was sufﬁcient to establish brackish conditions from the river mouth to the entire Upper Gulf of California (Carbajal et al. 1997). The interception of fresh water has drastically transformed the entire coastal zone. Not only has salinity increased across the entire coastal area, but the Colorado estuary has become an inverse estuary, with higher salinities up-estuary (owing to high rates of evaporation unmatched by river ﬂow) than toward the mouth (Lavín et al. 1998). It is evident from the Nile and Colorado examples that concerted efforts by humans can change the coastal environments in reasonably large areas of the world. As we increase the demand for fresh water, we should be concerned as to what the additive effects of more and more cases of changes, as took place in the Nile and Colorado deltas, might be.

10

The pressure to use even more water is extremely high, and awareness of the consequences just listed above does not deter governmental action to push for more water use. For example, Egypt has become a leading importer of grain, and the population of Egypt may nearly double by 2050. Concerned ofﬁcials plan to expand the 3.3 million ha of irrigated land to 4.6 million over the coming 20 years to provide for the coming numbers of people. This will require added interception of Nile water, and will certainly raise conﬂicts with other countries upstream along the Nile (Postel 1999).

Reduction of nutrient supply for phytoplankton Links between land-derived nutrient supplies and primary production are discussed in more detail in Chapter 12, which deals with eutrophication. Here I just brieﬂy mention some linkages that emerge from our examples of water interception. In many coastal areas of the world, nutrients borne by freshwater ﬂow are the major factor resulting in the relatively high production rates of phytoplankton that are characteristic of coastal waters.11 In our example of the Nile River delta, nutrient transport to the sea was radically diminished. The lack of freshwater ﬂow from the Nile, owing to pressure of the Aswan High Dam, plus other losses down-river, became evident in the makeup of water off the delta coast after the construction of the Aswan Dam (Fig. 4.14). The lack of freshwater ﬂow also implied lack of the annual resupply of nutrients, and the result was lowered phytoplankton production in the eastern Mediterranean. Phytoplankton blooms off the Nile delta reached 1–10 × 106 cells l−1 before the Aswan Dam was built, while cell densities in unaffected but nearby waters only reached 1–8 × 105 cells l−1 (Aleem 1972). A dearth of fresh water may not only lower the delivery of nutrients essential for producers in coastal waters, but may also alters ratios of the various nutrients (Ofﬁcer & Ryther 1980; Justic et al. 1995; Humborg et al. 1997). The ratio of silica to nitrogen has received particular attention. The trapping of sediment particles behind dams (a topic discussed in Chapter 5) also selectively traps silica. Silica may be in or adsorbed to particles, while nitrogen travels mainly in its oxidized inorganic form, nitrate, which travels freely dissolved in water. The result is that dams change the ratio of river-transported nutrients, favoring nitrogen at the expense of silica. This alteration promotes the growth of ﬂagellates at the expense 11

The other major source of nutrients that leads to high coastal production in certain coasts (Peru, Namibia, Paciﬁc Northwest of the US, northwestern Spain, and elsewhere) is the upwelling of nutrient-rich deeper water. This mechanism is limited to the western shores of continents, and to places with quite narrow continental shelves. Nutrients derived from land and transported by rivers are more widespread geographically.

ALTERATION OF FRESHWATER DISCHARGES

of diatoms. Though this might seem a minor change, it affects coastal planktonic food webs, might encourage toxic algal blooms, and might change the fauna that depends on the larger-sized diatoms. These topics are also discussed further in Chapter 12. Alteration of estuarine vegetation Salinity incursions resulting from reductions in freshwater ﬂow can change the vegetation of habitats present within coastal environments. For example, the delta of the Colorado River historically covered 780,000 ha, with many lagoons and salt marshes, and included two depressions below sea level, now known as the Salton Sea and Laguna Salada. Much of the area of the historical delta has been converted to agricultural land now characteristic of the Imperial Valley of California. Much of the salt marsh area has become bare mudﬂats (the topic of habitat loss through conversion to human uses is discussed in Chapter 6). The coastal lagoons that remain in the Colorado delta receive their water from irrigation runoff or ﬂood waters, or, in the case of the Salton Sea, mainly from municipal wastewater efﬂuent (Glenn et al. 1996). The remaining habitats are therefore thoroughly altered, enriched in nutrients, as well as salinized in the dry climate of the area. The vegetation of all these areas has been drastically changed. In the Columbia River, farther north in the Paciﬁc coast of North America, lower freshwater ﬂows allowed incursions of salty water up-river. The salt may have affected the extensive freshwater wetlands that provided detrital foods for the benthic invertebrates, which support the salmon runs that are so economically important to the region (Sherwood et al. 1990). Interception of river water in the watershed of the Indus River increased salinity in the Indus delta estuary to the extent of killing most of the mangroves that apparently are intolerant of saltier conditions (Snedaker 1984). On the other hand, freshwater impoundment up-river of the Kromme estuary of South Africa allowed an expansion of seagrass beds (Adams & Talbot 1992). The stabilization of freshwater ﬂows furnished by the upstream

95

dams provided salinity conditions beneﬁcial to the seagrass beds. Estuarine vegetation thus depends on freshwater supply, a resource strongly affected by humans. Whether the results of human management of water are good or bad, however, is a matter of what type of vegetation we may give the greater priority. In the Indus River, average freshwater discharge decreased from about 100 km3 yr−1 before the completion of the Kotri Barrage, to 50 km3 yr−1 after the barrage was built (Milliman et al. 1984). The halving of the water discharge was accompanied by a reduction of more than half the sediment load. These changes reduced the tall mangrove forests described in colonial records to remnant stands of mangroves along the betterﬂushed tidal channels, with largely bare sediments left elsewhere (Snedeker 1984). Although salt-tolerant, mangrove species appear not to survive where freshwater supply dwindles. These examples clearly demonstrate that human alteration of the freshwater supply has considerable consequences for the vegetation at the estuarine end of rivers. Alteration of fisheries and wildlife Reduced freshwater ﬂow moving through estuaries to coastal waters commonly impoverishes the value of coastal environments for ﬁsh and wildlife, and can alter ﬁshery yields. Following inauguration of the operation of the Aswan High Dam, there were signiﬁcant changes in the ﬁsheries off the Egyptian end of the Mediterranean (Fig. 4.15). Yields of commercial ﬁsh and shrimp from the eastern Mediterranean off the Nile River delta became much reduced after the completion of the dam (Aleem 1972). The catch of ﬁsh in the Mediterranean waters near the Nile delta before the completion of the Aswan High Dam reached up to 38,000 tons yr−1 (Sestini 1992). By the mid-1970s catches were as low as 7,000 tons yr−1. The sardine ﬁshery, which had furnished 30–50% of ﬁsh landings, was especially hard hit. The addition of newer ﬁshing technology increased catches in the late 1970s, but ﬁsh harvests declined again after 1980 (Sestini 1992), only to increase through the 1990s (Fig. 4.15 top).

CHAPTER 4

The coastal lagoons of the Nile delta yielded 60% of the ﬁsh caught in Egypt during the 1970s to 1980s, and sustained perhaps 90,000 ﬁshermen (Sestini 1992). Yields from these coastal lagoons actually increased as excess drainage water was emptied into the lagoons from irrigated ﬁelds. The kinds of ﬁsh did change, however, since the drainage water freshened the lagoons, and fostered a freshwater ﬁsh assemblage rather than an estuarine fauna. More recently, municipal wastewater efﬂuent and agricultural drainage have become the main freshwater source into the lagoons (and out to sea), and the catch of ﬁsh and shrimp has diminished because of poor water quality within the lagoons. On the other hand, the catch of ﬁsh, shrimps, and prawns off the delta have increased markedly (Fig. 4.15). Nixon (2003) calculated that the increase could be a result of the shift whereby the bulk of ﬂow through the delta is contributed by wastewater and agricultural drainage, which contain and transport large quantities of nutrients to the rather depauperate eastern Mediterranean. As Nixon (2003) concluded, these events simultaneously show the powerful inﬂuence of freshwater nutrient transport to large marine systems, as well making evident the massive impact of human activities. Reduced commercial ﬁsh harvests also followed construction of the Ghulam Mohammed Barrage (a low structure to hold back and manage water) on the Indus River in 1955 (Milliman et al. 1984; Snedaker 1984). The lower harvests were thought to be associated with lower freshwater inputs to the delta of the Indus River in the Sind coast of present-day Pakistan. Although annual ﬁsh catch in the area increased from 30 to > 200 × 103 tons, the catch per boat decreased markedly, from about 200 to only 50 tons per boat per year. Reduced water supply makes coastal wetlands less desirable for wildlife in general, because of the resulting reduction in suitable wetland habitats. As just one example, I use the instantly recognizable white stork, which breeds in Europe during the northern summer (Fig. 4.16). Once breeding is completed, the storks start a welldocumented migration to their wintering grounds.

80

FISH

60

40 Landings (metric tons × 103 yr−1)

96

Aswan Dam built

20

0 1960 8

1970

1980

1990

2000

1990

2000

SHRIMP AND PRAWNS

6

4

2

0 1960

1970

1980 Year

Figure 4.15 Landings of ﬁsh (top) and shrimps and prawns (bottom) from Egyptian Mediterranean waters, 1961–1998. Data from Nixon (2003), compiled from several sources.

The bulk of the world’s white storks funnel through two major bottlenecks, the Strait of Gibraltar and the eastern coast of the Mediterranean. It is not a coincidence that major wetland areas are located in both these regions. The delta of the Guadalquivir River in Spain, the location of the celebrated Coto Doñana area, provides food and respite to storks, and to innumerable other birds and animals. The delta of the Nile plays a similar role. We have already seen the drastic alterations to the Nile delta, but both these wetlands are threatened severely by the diversion of freshwater supply, as well as by contamination with wastewater and industrial efﬂuents. The more that people reduce these

ALTERATION OF FRESHWATER DISCHARGES

97

Breeding range Migration direction

Figure 4.16 Breeding range and directions of migration for the white stork, Ciconia ciconia.

areas of delta wetlands, the less hospitable these bottlenecks will be for migrants such as the white stork. We can well share concern for the welfare of the white stork as a species worthy of conservation. But there are larger issues involved as well. In this chapter, and others (Chapters 5, 6, and 7), we ﬁnd that what people do on the great continental expanses greatly alters conditions within coastal environments. We can look again at the map of Fig. 4.16, to notice that coastal wetlands,

that thin skin of the continents, may matter far beyond their modest expanse. Coastal wetlands such as we have found in the Nile delta turn out to be a vital link in the life of organisms found across huge stretches of continents. Here we have a telling example where human-driven changes in a local situation have the potential for repercussions on a near-global spatial scale. In Chapters 6, 7, and 12 we will run into other examples where coastal environments, in spite of their limited spatial extent, play a fundamental role in the

98

CHAPTER 4

coupling of far larger areas of continents and seas. Alterations of freshwater vegetation The high nutrient concentrations in the “fresh water” entering the Nile estuary and delta, and ﬁlling the lagoons and waterways, led to water bodies being choked by freshwater aquatic vegetation. These changes had unforeseen consequences. The proliferation of vegetation may have favored the abundance of snails that are hosts for schistosomes, and hence prompted a higher incidence of schistosomiasis. Human infection rates in some areas of the Nile valley reached 100% soon after the completion of the Aswan High Dam (Nash 1999). Higher disease incidence also followed other irrigation projects elsewhere in Africa (Nash 1999). In addition, in another unforeseen effect, the proliferation of aquatic vegetation also made the lagoons and waterways less available to birds, in general lowering the quality of these aquatic habitats for migrant and resident birds and waterfowl.12 Alteration of rates of sea level rise The volumes of fresh water retained in new reservoirs across the 20th century have been thought to be large enough to make a difference to the rate at which the sea level rises (Sahagian et al. 1994; Vörösmarty & Sahagian 2000). The estimations are fraught with substantial uncertainties, but they are worth some discussion. Correcting for presumed underestimation, the volume of 20th century freshwater impound12

The wetlands of the Nile delta make up about 25% of the wetlands on Mediterranean shorelines (Sestini 1992); this makes these areas key stopovers for birds moving from Europe to Africa and back, and an essential haven for resident waterfowl. The increasing sprawl of metropolitan Cairo, as well as the increasing population growth in Egypt, will increase demand for fresh surface water, groundwater, and land conversion for agricultural production, and will further lower the area and quality of the natural habitats of the Nile delta. This will add to the concentrations of contaminants and nutrients in the remaining waters, and affect the stocks of exploitable ﬁsh and shellﬁsh. Countermeasures to manage these trends will require heroic efforts on the part of government and people of the region.

ment might perhaps be equivalent to a sea level rise of 0.5–1 mm yr−1. Chapter 3 considered anthropogenic increases in the ﬂow of fresh water from land to sea (via glacial melt, wetland destruction, deforestation, and so on); these various ways in which people increase delivery of fresh water to the sea might add up to about 0.54 mm yr−1 (Sahagian et al. 1994). Given the uncertainties in the calculations, we must conclude that the potential slowing of sea level rise owing to impoundment of fresh water on land is more or less counterbalanced by the increased delivery of fresh water to the sea by other human activities across the last century.

Increases in freshwater discharges So far we have emphasized losses of freshwater discharge, caused mainly by the human uses of fresh water on land. There are, however, rivers where water discharge shows no secular change, and rivers whose discharge might have actually increased.13 In a very few rivers, water management plans transfer water from one river to another. One example is that of the Burntwood River of Canada (see Fig. 4.12 bottom panel). Such increases in ﬂow may be unusual, but do demonstrate the remarkable control over surfacewater discharge exerted by humans. A more widespread case of increased discharge of fresh water has been recently documented (Peterson et al. 2002). The aggregate discharge of fresh water by major Eurasian rivers into the Arctic Ocean has been highly variable year to year. The discharge did, however, increase slightly during the 20th century (Fig. 4.17). The reasons for such increases are being investigated, but a signiﬁcant correlation of discharge and global mean temperature (Peterson et al. 2002) suggests that the increased discharge might be in some way 13

Milly et al. (2003) concluded that the increased delivery of fresh water owing to climate-driven changes in water storage on continents (less snowpack, soil water, and groundwater) was signiﬁcantly smaller than the sea level rise created by changes in sea water volumes, but not negligibly small.

ALTERATION OF FRESHWATER DISCHARGES

might seem an esoteric consequence, until one thinks of the human and natural upheaval that could be foisted on Northern Europe by even a small shift in the position of the Gulf Stream.

Eurasian river discharge (km3 yr−1)

2,400 2,200 2,000 1,800

Extent of alteration of freshwater discharges

1,600 1,400 1,200 1940 1950 1960 1970 1980 1990 2000 Year

Figure 4.17 Aggregate discharge from the six largest Eurasian rivers (the Yenisey, Lena, Ob’, Severnaya Dvina, Pechora, and Kolyma) into the Arctic Ocean, 1935–1999. From Peterson et al. (2002).

associated with the warming trends discussed in Chapter 2.14 Previous sections devoted considerable space to discussions of possible consequences of decreased discharges. Increased river discharges may also have consequences for the receiving environments, and even beyond. For instance, the increased discharges of fresh water into the Arctic Ocean might be of sufﬁcient magnitude to alter the temperature and salinity conditions, which would force alterations of deep-water formation throughout the oceans and seawater circulation in the North Atlantic (Peterson et al. 2002).15 This 14

99

Others report lowered freshwater delivery through Siberian rivers (Aagaard & Carmack 1989; Barry & Serreze 1999), so some caution might be appropriate regarding inferences of global-scale effects of altered freshwater regimes. 15 Changes in salinity in the northern North Atlantic may have important climatic repercussions because colder denser surface waters formed in this region sink into the deep ocean, and are replaced by a northward ﬂow of warmer water. In part, these ﬂows set the path of the Gulf Stream, a major climate regulator. These hydrodynamic features have maintained reasonably temperate weather for most of Western Europe, at latitudes that elsewhere are associated with much colder climate. In fact, the waters of the northen North Atlantic have been slightly diluted with fresh waters in recent decades; the magnitude of current freshening suggest that it might take a further century before oceanic circulation is altered sufﬁciently to change regional climate patterns, although there are considerable uncertainties in all these speculations (Curry & Mauritzen 2005).

In earlier sections we reviewed examples that show the ready human capacity to interfere with freshwater ﬂows to coastal areas. It is difﬁcult to assess just how widespread the ensuing losses and gains of discharge might be. A reasonably broad geographic assessment of human interception of river ﬂow is a review of ﬂow alteration in rivers in the northern third of the world (Dynesius & Nilsson 1994). The area encompasses about two-thirds of North America and Eurasia. This area comprises a large part of the continental masses of the world, so it may be reasonably representative. One hundred and forty-four rivers were included in the survey. Of these 72% were unaffected. Of the 43 rivers that were altered by water management (Fig. 4.18), 26 were affected by water manipulations of less than 10%. We therefore have that 17 rivers suffered waterﬂow alterations of greater than 10%, or about 12% of the rivers on the northern third of the world. A few of these affected rivers have lost much of their ﬂow, including our examples of the Colorado, Amudarya, and Syrdarya (Fig. 4.18). If the data of Fig. 4.18 do represent the status of river ﬂow worldwide, it seems that at the start of the 21st century only a few rivers have suffered a large interception of water ﬂow. Another way to portray the status of presentday demand for fresh water, and its geographic distribution, is to use estimates of the use of fresh water relative to the renewable freshwater supply per country, for each continent (Fig. 4.19). First, taking the data set as a whole, it is evident that the majority (89 out of 134 countries represented in Fig. 4.19) of countries in the world used— during the last half of the 20th century—less than 10% of their annual renewable water supply. Somewhat fewer than half (45 out of 106) consumed more than 10%. These rough estimates

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24

16

8

Colorado Eastmain

Syrdarya

La Grande

4

Naskaupi

4

Amudarya

12 8

gained flow

Number of rivers that

lost flow

20

20 40 80 100 60 Volume flow used in irrigation or diversion (%)

Figure 4.18 Number of rivers in North America and Eurasia that suffer estimated losses (or gains) of discharge owing to interception for irrigation or diversion. Those rivers that are the most altered are indicated by a name above or below the histogram bars. The Eastmain, La Grande, and Naskaupi Rivers are in northern Canada, and are involved in major water management schemes. Data from Dynesius and Nilsson (1994).

suggest that, at global scales, our exploitation of the world’s rivers may be modest at present, but of course it will increase as demand inexorably grows. Second, water use is unevenly distributed from one continent to another. Asia and Europe include the countries most likely to withdraw larger portions of their annual supply, with a mean withdrawal of 15% (Fig. 4.19). The other continents were less intense users of water supplies. Even in Asia and Europe, however, many countries used less than 10% of their annual supply: the larger consumption averages derive from a few countries that use large proportions of their own supply, or even have to import water from elsewhere for their populations. These more extreme users are countries in desert regions, or smaller countries that are industrialized and urban. Locally, then, exploitation of freshwater sources are more

marked, and certainly have reached critical stages. At the start of the 21st century, therefore, humans appear to have severely threatened freshwater supplies in selected local situations, many of which we have seen as examples. Although these severely affected rivers may be few in number, the local effects are certainly intense and comprehensive. Globally, however, human beings seem to still be a modest feature in the hydrological cycle, at least in terms of river discharge interception and consumption of annual water supplies. Anthropogenic demand for fresh water will no doubt increase in the future. It remains to be seen at which point we begin to “Aralize” the remainder of the planet. The issues associated with freshwater discharges will certainly need to be carefully monitored since demand will but increase.

Restoration potential16 In most coastal areas, freshwater interception has not intensiﬁed enough to create serious problems as yet, but we have evidence from local examples that the effects can be severe. We can expect that pressure on the world’s freshwater supply will intensify. Supply of fresh water may indeed become one of the most contentious environmental issues of the 21st century, as populations increase and supplies fail to keep pace. What is certain is that human demand for fresh water will increase. We also know that the inevitable increases in diversion of fresh water for human uses will have serious effects on coastal environments. Most alterations created by water withdrawal seem reversible if ﬂow were to be restored. To achieve this would require less water use, but it seems extremely unlikely that there will be action leading to such restoration, for three major reasons. First, human populations are still, 16

Gleick (2003) provides an excellent review of the current status of the balance between global freshwater supply and increasing demands, and discusses possible avenues toward addressing the issues.

101

ALTERATION OF FRESHWATER DISCHARGES

24 ASIA

Bahrain

6

Israel

Cyprus

12

Mean: 15%

Afghanistan

18

0 24

EUROPE

15% Belgium

Spain

12 6

Luxemburg

18

24 NORTH AND CENTRAL AMERICA

10%

18 Barbados

Number of countries

0

12 6 0 24 AFRICA

3%

6

Egypt

12

Tunisia

Morocco

Malagasy

18

0 24 SOUTH AMERICA

18

Figure 4.19 Number of countries in Asia, Europe, North and Central America, Africa, and South America that withdraw given percentages of their annual renewable freshwater supply. Data from Gleick (1993, table H.1).

1%

12 6 0 0

10

20

30

40

50

60

70

80

90

100

Consumption of annual renewable freshwater supply (%)

as mentioned in Chapter 1, increasing. Although populations may be increasing somewhat more slowly than earlier in the 20th century, more people will simply require more fresh water, just to assure survival. To furnish sustenance for more

people, more land has been irrigated. Since the 1980s, however, the ratio of irrigated area per person has declined (Postel 1993). Since there are no obvious sources of new natural water supplies available (Postel 1999), other means of

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supporting increased agricultural yields will have to be found. Desalinization plants come to mind as a technologically possible alternative,17 but this alternative may be too costly for uses such as irrigation. Water conservation, new or water-saving techniques in irrigation and water transport, as well as the development of less water-demanding crops varieties will be needed. It seems unavoidable that traditional breeding and genetically engineered crops will have to be broadly used. Second, rivers often cross national boundaries, so that any plans for river water conservation will run up against difﬁcult inter-nation hurdles. Agreement on efforts to moderate water use by what have to be stringent restrictions on water use and conservation are likely to be impaired by the certain lack of common agreement as to what might be fair access to water and about reasonable standards for water use. The differences among nations will be especially problematic where economic status differs widely among the nations that share a watershed. There will be complex issues of different cultural values, as well as fairness in access to a limited resource to be solved. Third, economic costs would be prohibitive. The regional schemes that would be required to address freshwater interception issues are extremely costly because of the large scale of any likely remediation measure. Moreover, the requisite reductions in water use will be, in almost all cases, politically intractable because of the threat to existing sectors of economic life. We are at a reasonably early stage regarding intensity of use of fresh water, but we also have certain knowledge about the severity of impacts of water overuse. It therefore seems that we should start to face the imperative need to engage in 17

There are examples where circumstances dictate that desalinization must be applied. In the Canary Islands, for example, water supplies come from “mining” a very limited supply of deep fossil groundwater, plus a larger amount of desalinized sea water. Water there is simply more costly, and the expense is internalized as the cost of living in such a place. Mining of fossil groundwater, or pumping of groundwater out of aquifers faster than it may be recharged from precipitation, is prevalent also in much of India, China, and the western USA.

efforts at prevention of further pressures on water demand. This will be helped by fostering effective and frugal use of fresh water, and developing technical improvements that make saving water easier. In addition, we should encourage technical development of desalinization processes that more economically and effectively transform salt water into water useful for consumption and irrigation, and de-emphasize works of massive water impoundment.

References Aagaard, K., and E. C. Carmack. 1989. The role of sea ice and other freshwater in Arctic circulation. J. Geophys. Res. 94:14485–14498. Adams, J. B., and M. M. B. Talbot. 1992. The inﬂuence of river impoundment on the estuarine seagrass Zostera capensis Setchell. Bot. Mar. 35:69–75. Aleem, A. A. 1972. Effect of river outﬂow management on marine life. Mar. Biol. 15:200–208. Barry, R. G., and M. C. Serreze. 1999. Atmospheric components of the Arctic Ocean freshwater balance and their inter-annual variability. Pp. 45–65 in Lewis, E. L. (ed.) The Freshwater Budget of the Arctic Ocean. Kluwer Academic Press, Dordrecht, the Netherlands. Bird, E. C. F., and K. Koike. 1986. Man’s impact on sealevel changes: A review. J. Coast. Res. Special Issue 1:83– 88. Bortnik, V. N. 1999. Alteration of water level and salinity in the Aral Sea. Pp. 47–65 in Glantz, M. H. (ed.). Creeping Environmental Problems and Sustainable Development in the Aral Sea Basin. Cambridge University Press, Cambridge, UK, 291 pp. Carbajal, N., A. Souza, and R. Durazo. 1997. A numerical study of the ex-ROFI of the Colorado River. J. Mar. Syst. 12:17–33. Carriquiry, J. D., and A. Sanchez. 1999. Sedimentation in the Colorado River delta and the Upper Gulf of California after nearly a century of discharge loss. Mar. Geol. 158:125–145. Curry, R., and C. Mauritzen. 2005. Dilution of the northern North Atlantic in recent decades. Science 308:1772–1774. Dynesius, M., and C. Nilsson. 1994. Fragmentation and ﬂow regulation of river systems in the northern third of the world. Science 266:753–762. El Din, S. H. S. 1977. Effect of the Aswan High Dam on the Nile ﬂood and on estuarine and coastal circulation pattern along the Mediterranean Egyptian coast. Limnol. Oceanogr. 22:194–207. Elpiner, L. I. 1999. Public health in the Aral Sea coastal region and the dynamics of changes in the ecological situ-

ALTERATION OF FRESHWATER DISCHARGES

ation. Pp. 128–156 in Glantz, M. H. (ed.). Creeping Environmental Problems and Sustainable Development in the Aral Sea Basin. Cambridge University Press, Cambridge, UK, 291 pp. Ghassemi, F., A. J. Jakeman, and H. A. Nix. 1995. Salinisation of Land and Water Resources: Human Causes, Extent, Management, and Case Studies. University of South Wales Press, Sidney, 526 pp. Glantz, M. H. 1999. Sustainable development and creeping environmental problems in the Aral Sea Region. Pp. 1–25 in Glantz, M. H. (ed.). Creeping Environmental Problems and Sustainable Development in the Aral Sea Basin. Cambridge University Press, Cambridge, UK, 291 pp. Gleick, P. H. (ed.). 1993. Water in Crisis: A Guide to the World’s Fresh Water Resources. Oxford University Press, Oxford, 473 pp. Gleick, P. H. 2003. Global freshwater resources: Soft-path solutions for the 21st century. Science 302:1524–1528. Glenn, E. P., C. Lee, R. Felger, and S. Zengel. 1996. Effects of water management on the wetlands of the Colorado River delta, Mexico. Conserv. Biol. 10:1175–1186. Humborg, C., V. Ittekkot, A. Cosiasu, and B. Bodungen. 1997. Effect of Danube River dam on Black Sea biogeochemistry and ecosystem structure. Nature 386:385–388. ICOLD (International Commission on Large Dams). 1988. World Register of Dams, Updating the 1984 Full Edition. ICOLD, Paris. Justic, D., N. N. Rabalais, and R. E. Turner. 1995. Stoichiometric balance and the origin of coastal eutrophiction. Mar. Pollut. Bull. 30:41–46. Krutov, A. N. 1999. Environmental changes in the Uzbek part of the Aral Sea basin. Pp. 245–260 in Glantz, M. H. (ed.). Creeping Environmental Problems and Sustainable Development in the Aral Sea Basin. Cambridge University Press, Cambridge, UK, 291 pp. Lavín, M. F., V. M. Godínez, and L. G. Alvarez. 1998. Inverseestuarine features of the Upper Gulf of California. Estuar. Coast. Shelf Sci. 47:769–795. Létolle, R., and M. Mainguet. 1993. Aral. Springer-Verlag, Paris. L’vovich, M. I., and G. F. White. 1990. Use and transformation of terrestrial water systems. Pp. 235–252 in Turner, B. L., W. C. Clark, R. W. Kates, J. F. Richards, J. T. Matthews, and W. B. Meyer (eds). The Earth as Transformed by Human Action. Cambridge University Press, Cambridge, UK. Micklin, P. P. 1988. Dessiccation of the Aral Sea: A water management disaster in the Soviet Union. Science 241:1170–1176. Micklin, P., and W. D. Williams (eds). 1996. The Aral Sea Basin. NATO ASI Series, Environment, Vol. 12. SpringerVerlag, New York. Milliman, J. D., G. S. Quraishee, and M. N. A. Beg. 1984. Sediment discharge from the Indus River to the Ocean: Past, present and future. Pp. 65–70 in Haq, B. U., and J. D. Milliman (eds). Marine Geology and Oceanography of

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the Arabian Sea and Coastal Pakistan. Van Nostrand Reinhold Co., New York. Milly, P. C. D., A. Cazenave, and M. C. Gennero. 2003. Contribution to climate-driven change in continental water storage to recent sea-level rise. Proc. Nat. Acad Sci. 100:13158–13161. Moore, W. S. 1996. Large groundwater inputs to coastal waters revealed by Ra-226 enrichments. Nature 380:612– 614. Moore, W. S. 1997. High ﬂuxes of radium and barium from the mouth of the Ganges–Brahmaputra during low river discharge suggest a large groundwater source. Earth Planet. Sci. Lett. 150:141–150. Nash, L. 1999. Water quality and health. Pp. 25–39 in Gleick, P. H. (ed.). Water in Crisis: A Guide to the World’s Fresh Water Resources. Oxford University Press, Oxford, 473 pp. Nixon, S. W. 2003. Replacing the Nile: Are anthropogenic nutrients providing the fertility once brought to the Mediterranean by a great river? Ambio 32:30–39. Ofﬁcer, C. B., and J. H. Ryther. 1980. The possible importance of silicon in marine eutrophication. Mar. Ecol. Prog. Ser. 3:83–91. Orlovsky, N. S. 1999. Creeping environmental changes in biological communities in the Karakum Canal’s zone of impact. Pp. 225–244 in Glantz, M. H. (ed.). Creeping Environmental Problems and Sustainable Development in the Aral Sea Basin. Cambridge University Press, Cambridge, UK. Peterson, B. J., and 7 others. 2002. Increasing river discharge to the Artic Ocean. Science 298:2171–2173. Postel, S. L. 1993. Water and agriculture. Pp. 56–66 in Gleick, P. H. (ed.). Water in Crisis: A Guide to the World’s Fresh Water Resources. Oxford University Press, Oxford, 473 pp. Postel, S. L. 1998. Water for food production: Will there be enough in 2025? BioScience 48:629–637. Postel, S. 1999. Pillar of Sand: Can the Irrigation Miracle Last? W. W. Norton & Co., New York. Rosenberg, D. M., P. McCully, and C. M. Pringle. 2000. Global-scale environmental effects of hydrological alterations: Introduction. BioScience 50:746–751. Sahagian, D. L., F. W. Schwartz, and D. K. Jacobs. 1994. Direct anthropogenic contributions to sea level rise in the twentieth century. Nature 367:54–57. Sestini, G. 1992. Implications of climatic changes for the Po Delta and Venice Lagoon. Pp. 428–494 in Jeftic, L., J. D. Milliman, and G. Sestini (eds). Climatic Change and the Mediterranean. Edward Arnold, London, 673 pp. Sherwood, C. R., D. A. Jay, R. B. Harvey, P. Hamilton, and C. A. Simenstad. 1990. Historical changes in the Columbia River estuary. Prog. Oceanog. 25:299–352. Shiklomanov, I. A. 1993. World fresh water resources. Pp. 13–24 in Gleick, P. H. (ed.). Water in Crisis: A Guide to the World’s Fresh Water Resources. Oxford University Press, Oxford, 473 pp.

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Snedaker, S. C. 1984. Mangroves: A summary of knowledge with emphasis on Pakistan. Pp. 255–262 in Haq, B. U., and J. D. Milliman (eds). Marine Geology and Oceanography of the Arabian Sea and Coastal Pakistan. Van Nostrand Reinhold Co., New York. Stanley, D. J. 1988. Subsidence in the northeastern Nile delta: Rapid rates, possible causes, and consequences. Science 240:497–500. Stanley, D. J., and A. G. Warne. 1993. Nile delta: Recent geological evolution and human impact. Science 260:628– 634. Stone, R. 1999. Coming to grips with the Aral Sea’s grim legacy. Science 284:30–33. Vörösmarty, C. J., and D. Sahagian. 2000. Anthropogenic disturbance of the terrestrial water cycle. BioScience 50:753–765.

Vörösmarty, C. J., K. P. Sharma, B. M. Fekete, A. H. Copeland, J. Holden, J. Marble, and J. A. Lough. 1997. The storage and aging of continental runoff in large reservoir systems of the world. Ambio 26:210–219. Zholdasova, I. 1999. Fish populations as an ecosystem component and economic object in the Aral Sea basin. Pp. 204–224. In Glantz, M. H. (ed.). Creeping Environmental Problems and Sustainable Development in the Aral Sea Basin. Cambridge University Press, Cambridge, UK, 291 pp. Zonn, I. S. 1999. The impact of political ideology on creeping environmental changes in the Aral Sea basin. Pp. 157–190 in Glantz, M. H. (ed.). Creeping Environmental Problems and Sustainable Development in the Aral Sea Basin. Cambridge University Press, Cambridge, UK, 291 pp.

Chapter 5 Alteration of sediment transport

Left: composite mosaic of images of the Ebro delta (ortofotomapa 1:25,000 v 3 color geotiff ), courtesy of the Institut Cartogràﬁc de Catalunya, Generalitat de Catalunya. Right: aerial view of the mouth of the Ebro delta, courtesy of the Department de Comerc i Turisme, Generalitat de Catalunya.

Introduction Humans have interfered with the fate of coastal unconsolidated sediments at a global scale, through sea level changes (discussed in Chapter 3) and by altering ﬂuvial transport of sediment to the coast. Humans have interfered in the transport of sediments since the cultural shift from hunting-gathering to agricultural food production. The cultivation of land inevitably makes soil more subject to erosion, and some of the soil particles that are moved by running water eventually reach the coastlines of the world oceans. Clearance of forests and cultivation may increase the erosion of sediments by 2–500 times, and this

is followed by eight-fold increases in sediment loads in small rivers downstream, and perhaps 3.5-fold increases in larger rivers (Douglas 1990). As more and more people have tilled larger areas of land, ﬂuvial transport of sediments downstream has increased in most areas of the populated world. At some point, the accumulation of people in settlements and the pressure of agriculture also required that water be husbanded in new ways and in greater amounts. The management of fresh water eventually evolved into our present proliferation of waterworks—dams, dikes, barrages, levees, and canals. Of course, any time that running water is made to stand still, the sediment loads drop out of suspension; the net result of waterworks has therefore been that the

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sediments that we once helped accelerate on their way towards the seas, now accumulate behind dams and the like. Humans have also created important but localscale alterations in the distribution of sediments within different stretches of coastline, as a result of the construction of groins, jetties, harbors, breakwaters, and the like. The construction of such coastal engineering works aims to protect property and beaches—economic incentives that can be powerful arguments.1 In this chapter I use the case history of the Ebro delta—supplemented by information from the Mississippi, Nile, Huanghe (Yellow), Changjiang (Yangtze), Indus, and other rivers—to portray how humans have changed the land-to-sea transport of sediment, and altered the within-coast distribution of sediment through shoreline protection schemes. Then I review some effects and consequences of these alterations. The difﬁculty in attempting to unambiguously separate the effects of alterations of sediment transport from those of freshwater interception (Chapter 4), or from sea level rise (Chapter 3), will become obvious. There are many examples and aspects shared by these chapters, and there is additional overlap with material to be covered in Chapter 12 on eutrophication, because sediment particles carry incorporated and adsorbed nutrients as they move to the sea.

A case history: the Ebro watershed and delta2 The Ebro delta lies on the Mediterranean coast of Spain, and drains water from one of the largest watersheds in the Iberian Peninsula (Fig. 5.1). Sediment transport has been active in the region across geological time, and there is a long and 1

These issues, plus other climate-driven alterations, their effects, and consequences are reviewed by Boesch et al. (2000); a general review of the prospects for sandy shore ecosystems is provided by Brown and McLachlan (2002). 2 The material in this section was selected from Maldonado (1972, 1977), Nelson (1990), and Mariño (1992), and from discussions with Francisco Comin, Xavier Ferrer, Jacint Nadal, and Jordic Camp, all of whom have had much experience in the Ebro delta.

France

Barcelona Ebro delta Spain Valencia

Figure 5.1 Diagramatic view of the watershed of the Ebro River, its major tributaries, and positions of the major dams (dashes across rivers). Compiled from Palanques et al. (1990) and Mariño (1992).

complex geological history of accretion and erosion of river-borne sediment on the sea ﬂoor off the mouth of the Ebro. The watershed of the Ebro River was inhabited by people for millennia, as made evident by the artistic testament left behind in the Altamira caves and other archaeological sites. For our purposes, I concentrate on the better-documented and larger changes through historical times. At the time of the Roman Empire (Fig. 5.2 top left), the Ebro estuary spanned the reach between the Roman town of Iulia Llercavonia Dertosa (the presentday town of Tortosa) and the site now occupied by the town of Amposta. Long-shore currents eroded away sediments discharged by the Ebro, leaving little surﬁcial evidence of a delta. The population of the Iberian Peninsula increased and agricultural practices expanded during the rule of the Moorish emirate (710 to the 900s) and caliphate (after 929). In spite of the relative political chaos of the Taifa kingdoms after 1002, more and more of the forests within the catchment basin of the Ebro were cut down, and more land was brought under cultivation. In later centuries the development of the Merino breed made sheepraising proﬁtable throughout much of the Iberian

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ALTERATION OF SEDIMENT TRANSPORT

lulia Llercavonia Dertosa

Tortosa

Gola de Llevant Amposta

Gola Sud-est

4–10th century

Tortosa

15th century

Tortosa

L'Ampolla

Punta del Fangar

Illa de Mar Amposta Amposta Sant Carles de la Ràpita Punta del Galatxo

L'Ampolla Tortosa

Figure 5.2 Sketches of the extent of the Ebro delta during the 4–10th centuries, 15th century, 17th century, 18–19th centuries, 20th century, and forecast for the 21st century. Compiled from Maldonado (1972), Ferrer and Martinez-Vilalta (1986), Mariño (1992). Land cover details for the 20th century from F. Comin and X. Ferrer.

17th century

18–19th century

Land use Agriculture Urban Natural

Amposta

Erosion Accretion Sediment transport

Sant Carles de la Ràpita

Peninsula, which fostered the area of pastureland at the further expense of natural vegetation. The progressive replacement of natural vegetation with agricultural lands and, later, pastures for sheep, accelerated the erosion of soils and increased the amounts of eroded sediments brought to the rivers. The rivers in turn transported the particles seaward, and the buildup of sediments near-shore expanded the delta from the 10th to the 15th centuries (Fig. 5.2 top right). Erosion within the watershed continued, and the delta

20th century

21st century

grew from the 17th to the 19th centuries (Fig. 5.2 middle panels).3 Sediment loads carried by the Ebro increased, making the courses through the delta shallower. Traversing the delta became too difﬁcult for ship trafﬁc, so the Bourbon kings had a navigation channel constructed from Amposta to a monastic site that became in short order the 3

The time course of events and social and geographic changes in the Iberian Peninsula during the 6–15th centuries are well described in Jackson (1972) and Hillgarth (1979).

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1,000

Discharge (m3 s−1)

800

1912–1935 mean = 578

600 400

1951–1965 mean = 485

200 0 J

F

M

A

M

J J Month

A

S

O

N

D

Figure 5.3 Discharge of water through the Ebro River at Tortosa during two periods of time (1912–1935 and 1951–1965). Data from Maldonado (1977).

Table 5.1 Percentage of sediment transport by rivers that is trapped behind dams. River

Dam

US rivers Ebro Indus Nile

Many Mequinenza–Ribarroja* Tarbela Aswan High

% intercepted behind dams 8–64 73 99 98

Source Crowder 1987 Palanques et al. 1990 Milliman et al. 1984 Stanley & Warne 1993

*These two are only two of 21 large dams on the Ebro and its tributaries. On the whole, sediment discharge from the watershed to the Ebro delta is likely to be at least 95% of inputs (Palanques et al. 1990).

prosperous town of Sant Carles de la Ràpita (Fig. 5.2 middle right). Sediment continued to accrue during the early 20th century (Fig. 5.2 bottom left). Since 1970 the marked growth of the delta has stopped, and, instead, the recent evidence shows a redistribution and recession of part of the coastline of the delta (Fig. 5.2 bottom left). The recent coastal erosion noted in the Ebro delta can be linked to increased use of river water for irrigation, and the construction of dams upriver during the second half of the 20th century. Water discharge was reduced by 15% (Fig. 5.3), even though there was no reduction in rainfall. We saw in Chapter 4 that such reductions in ﬂow are not unusual in rivers in populated regions that depend on irrigated agricultural lands. Sediment load delivered by the Ebro to the delta decreased by an order of magnitude between 1965 and 1983. The study of sediment deposition behind the Ribarroja and Mequinenza dams up-river show quantitatively that sediments that would have reached the delta have indeed accumulated

behind the dams (Varela et al. 1983), as occurs in other sites (Table 5.1). Forecasts of future developments predict a continuing recession of the present delta (Fig. 5.2 bottom right).4 The human-driven nature of the Ebro delta does not stop at responsibility for the geological underpinnings. As soon as the natural wetland, lagoon, bays, and mudﬂat environments developed, as the sediments accumulated on the Ebro delta, human beings began to use the delta for their own purposes. The rich soils, being made 4

The history of the Ebro delta is mirrored in the northern Adriatic, where there is a similar example in the Po delta (Bondesan et al. 1995). This delta expanded until, by the end of the 16th century, the Venetians began to be concerned that the sediment load transported by the Po River might block inlets into their lagoon (cf. Chapter 3). To prevent this, they excavated a channel 5 km long, and diverted the Po to the southeast, away from the lagoon. The river then built the delta seaward of its new outlet. The story then follows much as in the case of the Ebro delta. The new land was taken over for cultivation, the rivers upstream eventually were dammed, river-borne sediment loads dropped, erosion of the delta and barrier beaches began, and engineering efforts to protect against the erosional work of the sea had to be deployed.

ALTERATION OF SEDIMENT TRANSPORT

up of relatively unweathered materials, have for centuries supported substantial cultivation, with some irrigation with river water and groundwater. Toward the end of the 20th century the Ebro delta grew up to 20% of the rice produced in Spain, among many other crops. The wetlands, lagoons, and bays of the delta are also quite productive, yielding 25% of the ﬁsh, and 40% of the shellﬁsh harvest in Catalunya (Mariño 1992). Maricultural practices, saltworks, and tourism also developed on the delta, adding to the economic impact of the region. With this intensive human use, only a small fraction of the Ebro delta remains as natural wetland or dune vegetation; by far most of the land is under rice or other cultivation (see Fig. 5.2 bottom left). The biological richness of the delta depends on its linkage to its watershed. The nutrients and sediments that underpin the productivity of the delta environment are brought into the delta by the river. Large quantities of nutrients are transported by river water (for example, concentrations of 179 µg phosphate l−1 and 137 µg ammonium l−1 have been measured, values that are quite high, even for estuarine waters, and can have major effects stimulating algal growth), but as we have seen, the most dramatic linkage between watershed and delta is through sediment transport. The groundwater has become too salty for irrigation, so river water is used instead; in fact the salty groundwater has to be actively pumped out of aquifers below agricultural ﬁelds to prevent salinization of the soils. As elsewhere, one human alteration demands another, often greater intervention in the environment. Even though the Ebro delta is a major producer of foodstuffs for Spain it still retains a major part of the coastal wetland areas in the Iberian Peninsula. The Ebro delta plays a fundamental role as a site for migratory stopovers and wintering grounds for much of the avifauna of Europe. For example, between 1966 and 1976, a range of 44 to 61 species of aquatic birds, with up to 72,000 individuals, could be recorded on the delta, especially in winter (Ferrer 1977). Further reduction of sediment loads from the Ebro, plus intensiﬁed exploitation, will alter the marshes, lagoons, dunes, and other habitats important to

109

wildlife, ﬁsh, and shellﬁsh, as well as the agricultural lands. As this book went to press, a plan to divert the waters of the upper reaches of the Ebro River was shelved by the new, more environmentally conscious, Spanish administration. The diversion plan was to provide water for the arid province of Murcia, where water is needed for fruit and horticultural crops for sale to Northern Europe, and to water the golf courses and yards of tourist developments. The proponents of the plan argued that Ebro River water was being wasted by merely ﬂowing out to sea, and that the economy of the dry lands would beneﬁt by use of that water. The new policy-makers—based on knowledge of the functions of freshwater export (Chapter 4), the importance of wetlands (Chapter 6), and the maintenance of sediment loads to deltas—plan instead to build desalinization plants on the Mediterranean coast to provide water for Murcia.

Human-mediated changes in sediment transport by rivers The human-mediated sequence of events in the Ebro delta is not an isolated example. What we know about changes in sediment transport through the Huanghe of China, the Nile of Egypt, the Indus of India, the Mississippi of the USA, and other rivers corroborates the fact that human impacts— made so evident in the case of the Ebro— do have the potential to locally alter coastal environments. Milliman et al. (1987) compiled information that suggests that before 200 BC the loess5 plateau drained by the Huanghe River was a forested steppe, and the sediment load was relatively low, perhaps an order of magnitude lower than during the 20th century. The river was simply called Dahe, the Great River. As the pressure for 5

Loess consists of wind-deposited ﬁne sediment particles, newly eroded from highlands, deserts, or glacial outwash, and often transported from considerable distances away. The relatively unweathered particles provide rich sources of minerals for plants, and after they accumulate and become soils, loess deposits make excellent agricultural lands. The pampas of Argentina and much of the great prairies of the Mississippi valley, other “breadbasket” areas of the world, are also loess plains.

110

CHAPTER 5

ADVANCE (1500–1900) Mediterranean Sea 19 00

RETREAT (1900–1991) Mediterranean Sea 190

0

N

0 190

1900 '41

1941 1964 1971

'64 '71

1982

1800

199

1

1800

'82 '91

1800 1700 1700

1,000 m

1600 e Nil

1600

er

Riv

1500

1500

Figure 5.4 Sketches of the extension (left; 1500–1900) and recession (right; 1900–1991) phases of the recent history of the Rosetta Promontory of the Nile River delta. The black circles are lighthouses and forts that serve to anchor the maps of the changing coastlines to immobile landmarks. From Fanos (1995).

agricultural land use grew, soil eroded from ﬁelds entered the river, and tinted the water, and the name of the river was changed to Huanghe, the Yellow River. Milliman and colleagues calculated that between 200 BC and 60 AD the Yellow River may have carried about 80% as much sediment as during the 20th century. There is a period after 60 AD that might have been unfortunate for Chinese civilization, but strikingly speaks of the powerful role of farming in setting sediment loads. This date marks the invasion of northern China by nomadic Mongols, with such social disruption that the cultivated areas of the watershed area returned to pastoral land uses. The farms became grassy and wooded steppes, and records suggest that during this time the river was clearer, perhaps carrying half as much

sediment. After 600 AD the plateau again became more heavily farmed, and loads resembled those to be seen later during the 20th century. The sediments carried by the river were such that the coastline expanded seaward by as much as 50 km across an interval of 130 years. The sediment loads were high, and varied seasonally and interannually enough to drastically alter the course of of the Yellow River estuary, with great instability, but continued expansion of the delta. The delta of the Nile River shows similar changes. Old maps, particularly the detailed surveys made by Napoleon’s French Expeditionary Force in 1800, were used by Fanos (1995) to reconstruct the shorelines of the Nile delta, including the point of land referred to as the Rosetta Promontory (Fig. 5.4). Natural and human alterations of the

111

ALTERATION OF SEDIMENT TRANSPORT

course of the Nile created the Rosetta Branch through the delta between 500 and 1000 AD, which brought an increased ﬂow of Nile water through the area, and silty sediments. The increased input of silt, and subsequent trapping within the delta, caused the Rosetta Promontory to advance into the Mediterranean Sea during the years 1500 and 1900. Changes in down-river transport of sediments are common in many areas of the world, but the underlying mechanisms may not be the same in all cases. In North America, for example, the Colorado River seems to have historically carried heavy sediment loads (its name—the “red river” in Castilian—was owed to the eroded sienna-colored sediments from its arid watershed), but it lost the bulk of the sediment load delivered down-river (Fig. 5.5). In this case, the decreased sediment

load was surely associated with the loss of most of its water ﬂow during the latter part of the 20th century after the start of operation of the Hoover Dam upstream (see Fig. 4.12 fourth panel). A quite different pattern was that of the Mississippi, where erosion from the agriculturally developing area of central North America increased through the 1800s and into the 1900s; when dikes were built after World War 2 to control the too-frequent ﬂoods, they effectively trapped the suspended sediments and lowered sediment transport downriver. Clear evidence of the signiﬁcant secularscale changes in sediment loads arriving at the mouth of the Mississippi is provided by the changes in area of the Mississippi delta across these decades (Fig. 5.6). Maps of a section of the delta (Fig. 5.7) give dramatic evidence that the changes in sediment loads arriving at the mouth of

Figure 5.5 Annual discharge of suspended sediment, 1911–1979, from the Colorado River at Yuma, Arizona. The abrupt decrease in the mid–1930s was caused by the operation of the Hoover Dam up-river. From Meade (1996), compiled from various sources.

Sediment discharge (tons × 106 yr−1)

300

200

100

0 1911

600

Area (km2)

500

1920

1930

1940 Year

1950

1960

1970

West Bay Cubits Gap Garden Island Bay Baptiste Collette Total land

400 300 200 100

Figure 5.6 Changes in the areas of sub-deltas and the total area for the Mississippi River delta, 1816–1978. Adapted from Wells (1996).

0 1800

1840

1880

1920 Year

1960

112

CHAPTER 5

Figure 5.7 Outlines of the extent of the Cubits Bay sub-delta of the Mississippi River, from maps dated 1838, 1884, 1905, 1922, 1946, and 1971. From Wells (1996).

the Mississippi drastically altered the landscape of the delta, and reformed the mosaic of associated coastal environments, including the wetlands of this coastal zone. There is a common historical pattern that emerges in areas of human settlement. This pattern is one in which as human populations increase in density, and require agricultural yields to support themselves, sediment loads carried by rivers draining the area increase. As human popula-

tions increase further, and it becomes necessary to increase agricultural production, water interception works are built, and the sediment load settles within the reservoirs, rather than being transported down-river. This then decreases the loads entering coastal waters. The rivers discussed so far are examples that demonstrate how human land uses originally increased sediment loads down-rivers, but not all rivers (and their estuaries) are similarly affected.

Annual suspended sediment flux (million tons yr−1) ( )

Figure 5.8 Time series of annual sediment transport (black circles) and freshwater discharge (gray histogram) for the Ob’ River. From Holmes et al. (2002).

40

800

30

600

20

400

10

200

0 1930

In this chapter I have not emphasized the biological effects of changed sediment loads, but such effects are common and manifold. We have seen the devastating impact of lower sediment load on the salt marshes of the Mississippi in the chapter on sea level rise, a matter taken up below in connection with Figs 5.11 and 5.12. Increased loads can be major threats to mangrove swamps, where turbid, sediment-laden water clogs the aerial roots of mangrove trees and prevents oxygenation of the mangroves. There are no doubt many rivers that do not show an anthropogenic imprint; I would guess that water and sediment transport in the Amazon and Orinoco hardly show any such alterations through the last few centuries. In spite of signiﬁcant land use changes owing to Soviet-era Five-Year Plans for development in the watershed of the Ob’ in Siberia, there were few indications of long-term changes in the discharge of fresh water or in the transport of sediments between the years 1935 and 2000 (Fig. 5.8). The record from the Ob’ is dominated by marked year-to-year variation, probably associated with weather changes, but shows none of the long-term trends of less water and changed sediment loads so often seen in other rivers of the world. We have long been aware that river water transport of terrestrial sediments is a major fea-

1940

1950

1960 1970 Year

1980

1990

Annual water discharge (km3 yr−1) ( )

113

ALTERATION OF SEDIMENT TRANSPORT

0 2000

ture of coastal and shelf environments, but river water discharge is not the sole factor inﬂuencing sediment transport to coastal environments. For example, in a careful review of information on rivers discharging into the Arctic Ocean, Holmes et al. (2002) found that discharge of water and sediment transport were related in six of the eight rivers studied, but were not related in the remaining two (examples of the two patterns are shown in Fig. 5.9). It seems, therefore, that there may or may not be a clear relationship between water and sediment loads. The Yukon and MacKenzie rivers carried, on average, only 21% of the fresh water and 73% of the sediment transported by the eight major rivers emptying into the Arctic. In contrast, the Yenisey, Lena, and Ob’ carried 65% of the fresh water, but only 17% of the sediment. Such uncoupling of sediment load from freshwater discharge might be related to differences in underlying soils and geology, precipitation regimes, slopes and relief of landscape, and human land use practices within the watersheds. For any one river, these factors might interact in complex ways to determine the rates at which terrigenous sediments move toward receiving marine environments (Milliman 1991; Meade et al. 2000). Nonetheless, a few general patterns might be evident (Milliman et al. 1987). Rivers of the warm, wet, mountainous basins of northern

114

CHAPTER 5

Annual suspended sediment concentration (mg l−1)

600 Yenisey Lena Ob’ MacKenzie Yukon

500

400

300

200

100

0 0

800

200 400 600 Annual water discharge (km3 yr −1)

South America, Southeast Asia, and Indo-Malaya, particularly those with high human population densities, were the largest sources of sediment to the seas. The rivers of Taiwan, surprisingly, discharged nearly as much sediment as the rivers of the coterminous United States. Rivers from colder, drier, ﬂatter, and less intensively used terrain made more modest contributions.

Fate of river-borne sediment loads in coastal environments On reaching the zone where fresh and sea water mix, river-borne particles tend to aggregate into larger, heavier particles. These aggregates settle largely within a reasonably narrow strip near the mouth of rivers, so that the discharge of fresh water and dissolved substances to the sea becomes uncoupled from the particulate materials carried by the river. Most of the sediment loads transported by the world’s major sedimentdischarging rivers remains within the receiving

Figure 5.9 Annual sediment concentration in river water in relation to annual freshwater discharge for examples of large rivers emptying into the Arctic Ocean. Points are for each year of data available; note the large year-toyear variation in most rivers. From Holmes et al. (2002).

estuaries and coastal areas (Fig. 5.10).6 Across the land/sea boundary for the world, perhaps less than 25–30% of the loads manage to course through beyond the immediate near-shore environments (Milliman 1991). Of course, the estimates of interception of terrigenous sediment loads for individual rivers vary hugely, but in general it appears that the bulk of river-borne sediments remains in the near-shore zone (Table 5.2). Estuarine environments are therefore in general accreting sedimentary systems that trap terrigenous particles,7 as well as particles that marine currents may bring in from deeper waters (Meade 1972). The bulk of ﬂuvially transported particles are trapped in estuaries, but the sediments that do 6

This is the current situation; during glacial periods, when the sea level was much lower, the shelf break was of course nearer shore, coastal sedimentary environments must have had a more limited area, and the discharge of ﬂuvial-transported particles to deeper water might have been much larger than today. 7 This tendency to trap particulate materials is an important feature of near-coastal environments. I will return to the issue when discussing toxic substances and nutrients in Chapters 7–9 and 13.

115

ALTERATION OF SEDIMENT TRANSPORT

Amazon

Ganges

Yellow

Head of delta

Delta

Coastline

Figure 5.10 Relative proportions of discharge of suspended sediment delivered to deltas, crossing the coastline, and moving either longshore or to the shelf, and thence to the deep sea off the shelf break, for the three major sedimentdischarging rivers of the world. From Meade (1996), compiled from many sources.

Long-shore

Long-shore

Long-shore Shelf

Shelf break Deep sea

Table 5.2 Percentage of land-derived sediment transported by rivers that is retained within the estuarine environments. River Huanghe Huanghe + Chagjiang Lena MacKenzie Yenisey

% retained within estuarine environments 64 85–95 0–90 ∼50 Nearly 100

traverse estuaries are essential in the formation and maintenance of sea ﬂoor characteristics offshore (Sternberg 1984). For example, sediments carried by the Changjang River are responsible for submarine deltas, sand bars, and other structures in the East China Sea (Wang & Aubrey 1987). Huanghe River sediments can be found throughout the South China Sea and the Yellow Sea (Milliman et al. 1987). Reworking of sediments eroded from the Mississippi delta puts a fundamental imprint on the conﬁguration of shelf sediments off the Mississippi delta (Brooks et al. 1995). Tectonic changes through the Oligocene to the present have altered sediment transport of the Indus River (Coumes & Kolla 1984). Indus

Source Wang & Aubrey 1987 Milliman et al. 1987 Rachold et al. 2000 MacDonald et al. 1998 Meade et al. 2000

sediment may be deposited in the near-shore or shelf areas or may be carried farther by long-shore currents (Wells & Coleman 1984). The different rates of sediment transport by the Indus in turn markedly alters the sedimentary fan offshore, creating erosion, meanders, slumps, and ﬁlling of canyons up to 15 km wide, as well as creating a series of sedimentary basins manifested in a series of sedimentary strata offshore. These examples suggest we should expect the transport of riverine sediments to have a history of consistent change, and that these changes have broad-scale effects on adjacent marine sedimentary environments and, inevitably, on the organisms within the surface sediments. There

116

CHAPTER 5

has been too little study of the consequences of the interception of annual sediment loads from rivers on the submarine environments that are known to depend on ﬂuvial sediments to maintain sea ﬂoor topography and sediment character. When we obtain such data, no doubt we will ﬁnd that there are signiﬁcant biological effects from such substantial changes.

to increase irrigation of ﬁelds on the Nile ﬂood plain lowered the ﬂow of water through the Nile delta. The withdrawal of water has resulted in a steady recession of the Rosetta Promontory since 1900, particularly after the Aswan High Dam started operation (see Fig. 5.4 right). The shoreline was back to where it might have been between the years 1700 and 1800. Submergence of coastal wetlands

Decreased sediment loads The remarkable power of human beings in the alteration of natural environments is certainly made evident by the previous section, where we read about increases in sediment loads across historical time. But humans go beyond that: as we have become technologically more able, we carry out land management practices that can diminish the land-derived sediment loads to a signiﬁcant degree. Dams, barrages, levees, dikes, irrigation canals, and other waterworks can effectively intercept suspended particles that would otherwise have moved into rivers and be taken to coastal environments. It helps to grasp the magnitude of sediment interception to learn that the volume of accumulated settled sediments in reservoirs on the Yellow River in China amounted to 50–87% of the water-storage capacity of the reservoirs (Vörösmarty et al. 1997). Such trapping raises difﬁcult management issues as to the future of dam constructions. Sediment trapping in reservoirs can reach remarkably high percentages of the loads transported by many rivers (see Fig. 5.5 and Table 5.1). It is not surprising, then, that the sediment inputs to coastal environments may be decreased signiﬁcantly by human intervention upstream, and the reductions are large enough to have important consequences. Erosion of deltas and barrier bars The supply of sediment by rivers sustains deltas. As human activities on land lower sediment transport, delta shores retreat and the habitats of the deltas are lost. For example, after 1900 efforts

The reduction of sediment loads can markedly diminish the area of coastal wetlands. A worst-case example of this mechanism at work is the loss of salt marshes in the Mississippi delta. It is not possible to deﬁnitively say how much of the apparent subsidence may be the result of the extraction of groundwater (Chapter 4), sea level rise (Chapter 3), or be attributable to a reduction of sediment loads. Most researchers share the notion that a dearth of sediment supply is a principal cause (Hatton et al. 1983; Phillips 1986; Salinas et al. 1986). The Mississippi River has changed course every 1,000–2,000 years, and the changes have drastically altered the geography of its delta and the wetlands it supports. These wetlands make up about 41% of all US wetlands (Turner & Gosselink 1975), and are disappearing at rates as high as 130 km2 yr−1 (Gagliano et al. 1981). Elsewhere in the USA, salt marsh accretion has managed to keep up with the pace of submergence (Letzsch & Frey 1980). In Louisiana the supply of sediment brought into the wetlands by the Mississippi has been insufﬁcient to compensate for the sediments that are eroded from the estuary, and wetland surfaces are becoming lower (Fig. 5.11). The various species of salt marsh plants survive within certain ranges of duration and depth of tidal immersion, so increases in submergence lead to changes. These changes in Louisiana mean that the bands of different species have to move landward. Erstwhile salt marsh areas become open water areas, and, on the whole, the coastal landscape becomes more and more dominated by open water instead of marshes (Figs 5.7, 5.11), and the remaining wetland habitats become dominated by the more salt-tolerant marsh plants. Of course, saltier water becomes prevalent farther

117

ALTERATION OF SEDIMENT TRANSPORT

2000 1970

1930

New Orleans

2000 1970

1970 2000 1930 N

0

10

20

30

Gulf of Mexico

km

Figure 5.11 Map of the Mississippi delta area, with lines that show, for 1930, 1970, and 2000, the position of sites with a 1 : 1 ratio of open water to wetland area. Note the marked recession of the lines landward across the decades. From Salinas et al. (1986).

Loss of wetland area (% yr−1)

1.6

1.2

0.8

0.4

0.0 1880

1900

1920 1940 Year

1960

1980

Figure 5.12 Rates of loss of wetland habitats in the coast of Louisiana through the 20th century. Data compiled by Salinas et al. (1986).

up the estuary than before, and this has several consequences, including possible salt incursions into aquifers and the loss, or at least reduction, of brackish water ﬂora and fauna (Salinas et al. 1986).

The major cause of the reduced sediment load underpinning all these changes is that the main course of the Mississippi River is nearly completely surrounded by levees, constructed since the early 1900s as ﬂood-control devices to protect towns and agricultural areas. These levees isolate the river from sources of terrigenous sediments, as well as the wetlands within the delta. The effects of the levees, compounded by other kinds of wetland loss in Louisiana—including construction of canals, dredging, removal of groundwater, and agricultural development (Deegan et al. 1983) —led to a 70% decrease in sediment transport by the Mississippi (Milliman et al. 1989). The reduction in sediment load in turn led to an increasing loss of wetland habitats throughout the 20th century 8 (Fig. 5.12) (Britsch & Dunbar 1993). The im8

The loss of these wetlands was not immaterial to the devastation brought about by hurricane Katrina during the 2005 hurricane season. Wetlands absorb energy and hold ﬂooding waters; New Orleans, Biloxi, and other urban areas therefore had lost part of the natural defenses against ﬂooding and storm surges.

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CHAPTER 5

Figure 5.13 Sea walls and rock revetments built all around the shores of the Oosterschelde estuary in the Netherlands to protect ﬁlled areas from the sea. Photo by NIOOCEMO, Yerseke, reproduced in Nienhuis and Smaal (1994).

portance of sufﬁcient riverine sediment loads is emphasized by the fact that the mouth of the Atchafalaya River, which is relatively free of diking, is the only place in this area where wetlands have expanded in recent years (Steward & Berry 1990).

Remediation of coastal erosion A variety of engineering works have been devised to prevent or modify the erosion of unconsolidated sediments along coasts. These include the hardening of shores by installing rocks or concrete structures and sea walls (Fig. 5.13), the building of groins, jetties, and breakwaters (Fig. 5.14) to protect from wave action, and many other interventions. Another type of remediation is to artiﬁcially replenish beaches and coasts. The impacts of the construction of coastal structures are usually local in extent, and subject to complex inﬂuences of local features (Bird 1985;

Karambas 2003). The net effect of such devices is, in general, to redistribute erosion: they may for a time protect a given area, but lead to loss of sediments elsewhere (Leatherman et al. 2000; Pilkey et al. 2000). There is considerable disagreement on the utility and effects of coastal works. Certain parcels of shoreline have no doubt been preserved by such works, but these works may have increased, or at least redistributed, erosion. Wiegel (2002), speaking of one type of coastal works, concluded “seawalls do not cause erosion . . . except where they act as a groin . . . seawalls can decrease erosion of beaches by preventing transport of sand [during storms]”. On the other hand, others conclude that structures such as jetties may be responsible for over 85% of the severe erosion in long stretches of shores in Florida, for instance (Finkl 1996). Suspended sediment and long-shore transport off beaches in Devon, UK, were threefold, and an order of magnitude, higher where sea walls were present (suggesting greater erosion) (Miles et al. 2001).

ALTERATION OF SEDIMENT TRANSPORT

119

Figure 5.14 Aerial view of part of the delta of the Llobregat River, near Barcelona, showing sea walls, groins, and much alteration of the wetlands of the delta, designed to direct river ﬂow, catch transported sand, and protect the sewage treatment plant, agricultural ﬁelds, and other urban features from ﬂoods and storm damage.

Coastal protective works therefore do work to save certain coastal sites, but at a cost to the shorelines at broader scales. The decision as to where to protect shores therefore depends on the relative values we place on the speciﬁc sites in question versus regional spatial scales. The erosion of shores, particularly beaches, has prompted many efforts at artiﬁcial beach nourishment. This involves dredging sand from some nearby source, and transporting the dredge material to ﬁll in beaches that have been eroded. This practice is widespread (Trembanis et al. 1999; Hanson et al. 2002). In the Mediterranean coast of Spain, for example, beach erosion, owing largely to sea level rise and various types of shoreline works, occurs widely; pressure to maintain tour-

ism has prompted beach replenishment in about 400 sites throughout the Iberian Peninsula as well as in many other areas (Fig. 5.15). Replenishment does temporarily produce a usable beach (Fig. 5.16), but it is costly,9 often has to repeated annually or more frequently,10 and has biological consequences. The dredging necessary to obtain sand for replenishment thoroughly disturbs the bottom and 9

The cost of replenishing US beaches reached about US$2.5 billion in 1996 (Trembanis et al. 1999). Beach replenishment in the Gulf of Cadiz sites in Spain required a minimum of US$3.7 million per year between 1989 and 1998 (Muñoz-Perez et al. 2001). 10 Beach replenishment is by no means permanent (Daniel 2001); 27% of replenished beaches disappear in less than a year, 62% last 2–5 years, and only 12% remain for more than 5 years.

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Toulouse

Gijón

¨ La Coruna

Bilbao

Biarritz

Santander Pamplona

Vitoria Orense

Vigo

Burgos

Logrono ¨

Valladolid

Zaragoza Barcelona

Porto Tarragona S

P

A Madrid

I

N

Tortosa Castellón de la Plana Palma

PORTUGAL

Valencia Albacete

Badajoz Lisbon

M E D I T E R R E A N Alicante Córdoba Huelva

Sevilla

Murcia Cartagena

Jaén Granada

Tenerife Teide

Almería Málaga Gibraltar

Figure 5.15

Gran Canaria

Sites throughout Spain where beach nourishment has taken place. From Hanson et al. (2002).

associated fauna of dredged sites. After dredging was carried out in an area off the Tordera River in Catalunya, Spain, the remaining sediments were considerably coarser, and were almost completely defaunated (Sardá et al. 2000). Colonization of the defaunated sediments by benthic fauna was relatively rapid, with densities returning to previous values within 2 years for most, but not all, species. It appears that some elements of fauna of these habitats, presumably habituated to storm and other disturbances, are able to recover from the dredging. There is little information on the fate of beach benthos after dredging.

It is evident from the evidence gathered in this chapter that human activities on land have increased and decreased sediment transport from land to the near-shore to a signiﬁcant degree at large regional scales. Although not all rivers are similarly affected, there are a considerable number of important rivers that have been altered during human history. In a handful of cases, the coastal environments were drastically altered, but not all the biological consequences are well documented. At local scales, humans have been singularly successful at reshufﬂing sediment distribution so as to protect valued sites, with perhaps some biological consequences.

ALTERATION OF SEDIMENT TRANSPORT

121

Figure 5.16 View of the coastline near S. Vicent de Montalt’, el Maresme, Barcelona, before (left) and after (right) beach nourishment. From Hamm et al. (2002).

References Bird, E. C. F. 1985. Coastline Changes: A Global Review. Wiley, Chichester, UK, 219 pp. Boesch, D. F., J. C. Field, and D. Scavia (eds). 2000. The Potential Consequences of Climate Variability and Change on Coastal Areas and Marine Resources: Report of the Coastal Areas and Marine Resources Sector Team, U. S. National Assessment of the Potential Consequences of Climate Variability and Change, U.S. Global Change Research Program. NOAA Coastal Ocean Program Decision Analysis Series No. 21. NOAA Coastal Ocean Program, Silver Spring, MD, 163 pp. Bondesan, M., and 6 others. 1995. Coastal areas at risk from storm surges and sea-level rise in northeastern Italy. J. Coast. Res. 11:1354–1379. Britsch, L. D., and J. B. Dunbar. 1993. Land loss rates: Louisiana coastal plain. J. Coast. Res. 9: 324–338. Brooks, G. R., J. L. Kindinger, S. Penland, S. J. Williams, and R. A. McBride. 1995. East Louisiana continental shelf sediments: A product of delta reworking. J. Coast. Res. 11:1026–1036.

Brown, A. C., and A. McLachlan. 2002. Sandy shore ecosystems and the threats facing them: Some predictions for the year 2025. Environ. Cons. 29:62–77. Coumes, F., and V. Kolla. 1984. Indus fan: Seismic structure, channel migration and sediment thickness in the upper fan. Pp. 101–112 in Haq, B. U., and J. D. Milliman (eds). Marine Geology and Oceanography of the Arabian Sea and Coastal Pakistan. Van Nostrand Reinhold Co., New York, 382 pp. Crowder, B. M. 1987. Economic costs of reservoir sedimentation: A regional approach to estimating cropland erosion damages. J. Soil Water Conserv. 42:194–197. Daniel, H. 2001. Replenishment versus retreat: The cost of maintaining Delaware’s beaches. Ocean Coast. Manag. 44:87–104. Deegan, L. A., H. M. Kennedy, and R. Costanza. 1983. Factors contributing to marsh land loss in Louisiana’s coastal zone. Pp. 915–920 in Lauenroth, W. K., G. V. Skogerboe, and M. Flug (eds). Analysis of Ecological Systems: State of the Art in Ecological Modeling. Elsevier Publishing, New York, 992 pp. Douglas, I. 1990. Sediment transfer and siltation. Pp. 215–234 in Turner, B. L., W. C. Clark, R. W. Kates, J. F.

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Richards, J. T. Matthews, and W. B. Meyer (eds). The Earth as Transformed by Human Action. Cambridge University Press, Cambridge, UK. Fanos, A. M. 1995. The impact of human activities on the erosion and accretion of the Nile delta coast. J. Coast. Res. 11:821–833. Ferrer, X. 1977. Introducció ornitològica al delta de l’Ebre. Treb. Inst. Cat. Hist. Nat. 8:227–302. Ferrer, X., and A. Martinez-Vilalta. 1986. Fluctuations of the gull and tern population in the Ebro delta, north-east Spain (1960–85). Pp. 273–294 in MEDMARAVIS, and X. Monbailliu (eds). Mediterranean Marine Avifauna: Population Studies and Conservation. NATO ASI Series Vol. G12. Springer-Verlag, Berlin, 535 pp. Finkl, C. W. 1996. What might happen to America’s shoreline if artiﬁcial replenishment is curtailed: A prognosis for southeastern Florida and other sandy regions along regressive coasts. J. Coast. Res. 12:iii–ix. Gagliano, S. M., J. Meyer-Arendt, and K. M. Wicker. 1981. Land loss in the Mississippi River deltaic plain. Trans. Gulf Coast Assoc. Geol. Soc. 31:295–300. Hamm, L., and 5 others. 2002. A summary of European experience with shore nourishment. Coast. Engineer. 47:237–264. Hanson, H., and 7 others. 2002. Beach nourishment projects, practices, and objectives—an European overview. Coast. Engineer. 47:81–111. Hatton, R. S., R. D. DeLaune, and W. H. Patrick, Jr. 1983. Sedimentation, accretion, and subsidence in marshes of Barataria Basin, Louisiana. Limnol. Oceanogr. 28:494–502. Hillgarth, J. N. 1979. Los Reinos Hispánicos 1250–1516. Vol. 1. Un Equilibrio Precario: 1250–1410. Ediciones Grijalbo S. A., Barcelona, 501 pp. Holmes, R. M., and 7 others. 2002. A circumpolar perspective on ﬂuvial sediment ﬂux to the Arctic Ocean. Glob. Biogeochem. Cycles 16:45.1–45.14. Jackson, G. 1972. The Making of Medieval Spain. Thames & Hudson, London, 216 pp. Karambas, T. V. 2003. Modeling of sea-level rise effects on cross-shore erosion. J. Mar. Environ. Eng. 7:15–24. Leatherman, S. P., K. Zheng, and B. C. Douglas. 2000. Sea level rise shown to drive coastal erosion. Eos Trans. Am. Geophys. Union 81:55–57. Letzsch, W. S., and R. W. Frey. 1980. Deposition and erosion in a Holocene salt marsh, Sapelo Island, Georgia. J. Sediment Petrol. 50:529–542. MacDonald, R. W., S. M. Solomon, R. E. Cranston, H. E. Welch, M. B. Yunke, and C. Gobeil. 1998. A sediment and organic carbon budget for the Canadian Beaufort Shelf. Mar. Geol. 144:255–273. Maldonado, A. 1972. El Delta del Ebro: Estudio Sedimentológico y Estratigráﬁco. Boletín de Estratigrafía 1 (vol. extr.): 1–476. Departamento de Estratigrafía y Geología Histórica, Universidad de Barcelona, Barcelona. Maldonado, A. 1977. Introducción geológica al delta del Ebro. Treb. Inst. Cat. Hist. Nat. 8:7–45.

Mariño, M. G. 1992. Implications of climatic change on the Ebro delta. Pp. 304–327 in Jeftic, L., J. D. Milliman, and G. Sestini (eds). Climatic Change and the Mediterranean. Edward Arnold, London, 673 pp. Meade, R. H. 1972. Sources and sinks of suspended matter in river systems. Pp. 249–262 in Swift, D. J. P., D. B. Duane, and O. H. Pilkey (eds). Shelf Sediment Transport: Processes and Patterns. Dowden Hutchinson & Ross Inc., Stroudsburg. Meade, R. H. 1996. River-sediment inputs to major deltas. Pp. 63–85 in Milliman, J. D., and B. U. Haq (eds). SeaLevel Rise and Coastal Subsidence. Kluwer Academic Publishers, Dordrecht, the Netherlands, 369 pp. Meade, R. H., N. N. Bobrovitskaya, and V. I. Babkin. 2000. Suspended-sediment and fresh-water discharges in the Ob’ and Yenisey rivers, 1960–1988. Int. J. Earth Sci. 89:578–591. Miles, J. R., P. E. Russell, and D. A. Huntley. 2001. Field measurements of sediment dynamics in front of a seawall. J. Coast. Res. 17:195–206. Milliman, J. D. 1991. Flux and fate of ﬂuvial sediment and water in coastal seas. Pp. 69–110 in Mantoura, R. F. C., J.-M. Martin, and R. Wollast (eds). Ocean Margin Processes in Global Change. John Wiley & Sons, Chichester, UK, 469 pp. Milliman, J. D., J. Broadhaus, and F. Gable. 1989. Environmental and economic implications of rising sea level and subsiding deltas, the Nile and Bangal examples. Ambio 18:340–345. Milliman, J. D., Y. S. Qin, M. E. Ren, and Y. Saito. 1987. Man’s inﬂuence on the erosion and transport of sediments by Asian rivers: The Yellow River (Huanghe) example. J. Geol. 95:751–762. Milliman, J. D., G. S. Quraishee, and M. A. A. Beg. 1984. Sediment discharge from the Insus River to the ocean. Past, present and future. Pp. 65–70 in Haq, B. U., and J. D. Milliman (eds). Marine Geology and Oceanography of the Arabian Sea and Coastal Pakistan. Van Nostrand Reinhold Co., New York, 382 pp. Muñoz-Perez, J. J., B. L. de San Roman-Blanco, J. M. Gutierrez-Mas, L. Moreno, and G. J. Cuena. 2001. Cost of beach maintenance in the Gulf of Cadiz (SW Spain). Coast. Engineer. 42:143–153. Nelson, C. H. 1990. Estimated post-Messinian sediment supply and sedimentation rates on the Ebro continental margin, Spain. Mar. Geol. 95:395–418. Nienhuis, P. H., and A. C. Smaal (eds). 1994. The Ooosterschelde Estuary (the Netherlands): A Case-Study of a Changing Ecosystem. Kluwer Academic Publishers, Dordrecht, the Netherlands, 597 pp. Palanques, A., F. Plana, and A. Maldonado. 1990. Recent inﬂuence of man on the Ebro margin sedimentation system, northwestern Mediterranean Sea. Mar. Geol. 95:247–263. Phillips, J. D. 1986. Coastal submergence and marsh fringe erosion. J. Coast. Res. 2:427–436.

ALTERATION OF SEDIMENT TRANSPORT

Pilkey, O. H., and 11 others. 2000. Sea rise level shown to drive coastal erosion; discussion and reply. Eos Trans. Am. Geophys. Union 81:436–437, 439–441. Rachold, V., and 6 others. 2000. Coastal erosion vs. riverine sediment discharge in the Arctic Shelf Seas. Int. J. Earth Sci. 89:450–460. Salinas, L. M., R. D. DeLaune, and W. H. Patrick, Jr. 1986. Changes occurring along a rapidly submerging coastal area: Louisiana, USA. J. Coast. Res. 2:269–284. Sardá, R., S. Pinedo, A. Greamre, and S. Taboada. 2000. Changes in the dynamics of shallow sandy-bottom assemblages due to sand extraction in the Catalan Western Mediterranean. ICES J. Mar. Sci. 57:1446–1453. Stanley, D. J., and A. G. Warne. 1993. Nile Delta: recent geological evolution and human impact. Science 260:628–634. Sternberg, R. W. 1984. Sedimentation processes on continental shelves. In Haq, B. U., and J. D. Milliman (eds). Marine Geology and Oceanography of Arabian Sea and Coastal Pakistan. Van Nostrand Reinhold Co., New York, 382 pp. Steward, H. L., and W. L. Berry. 1990. Louisiana’s vanishing coastal wetlands. Geotimes 35:19–21. Trembanis, A. C., O. H. Pilkey, and H. R. Valverde. 1999. Comparison of beach nourishment along the U.S. Atlantic, Great Lakes, Gulf of Mexico, and New England shorelines. Coast. Manage. 27:329–340.

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Turner, R. E., and J. G. Gosselink. 1975. A note on standing crops of Spartina alterniﬂora in Texas and Florida. Contrib. Mar. Sci. Univ. Texas 19:113–118. Varela, J., A. Gallardo, and A. López de Velasco. 1983. Retención de sólidos por los embalses de Mequinenza y Ribarroja: Efectos sobre los aportes al delta del Ebro. Pp. 203–219 in Mariño, M. (ed.). Sistema Integrado del Ebro. Madrid. Vörösmarty, C. D., and 7 others. 1997. The storage and aging of continental runoff in large reservoir systems of the world. Ambio 26:210–219. Wang, Y., and D. G. Aubrey. 1987. The characteristics of the China coastline. Coast. Shelf Res. 7:329–349. Wells, J. T. 1996. Subsidence, sea-level rise, and wetland loss in the Lower Mississippi River delta. Pp. 281–312 in Milliman, J. D., and B. U. Haq (eds). Sea-Level Rise and Coastal Subsidence. Kluwer Academic Publishers, Dordrecht, the Netherlands, 369 pp. Wells, J. T., and J. M. Coleman. 1984. Deltaic morphology and sedimentology, with special reference to the Indus River delta. Pp. 85–100 in Haq, B. U., and J. D. Milliman (eds). Marine Geology and Oceanography of the Arabian Sea and Coastal Pakistan. Van Nostrand Reinhold Co., New York, 382 pp. Wiegel, R. L. 2002. Seawalls, seacliffs, beachrock: What beach effects? Shore Beach 70:17–27.

Chapter 6 Loss of coastal habitats

Views in the Mondego estuary, Portugal. Much of the ertswhile salt marsh has been ﬁlled (top, left side), leaving only a margin of salt marsh (top, right side) populated with salt-tolerant plants, Halimione (light gray) and Spartina (darker gray). The salt marsh habitat has also been converted to roads, residential areas, ﬁsh ponds, and evaporation pans for the production of salt (bottom).

LOSS OF COASTAL HABITATS

A case history: changes in the coastal environments of New England1 Excavation for the Boylston Street subway through the Back Bay section of Boston was brought to a halt in 1913. Workers had unearthed, within what was now the urban core of Boston, the remains of what was to be eventually said to be a large ﬁsh weir (Fig. 6.1). The weir had more than 65,000 wooden stakes and extended to perhaps a hectare, spread out on an area of what was, at the time, a marshy mudﬂat. Radiocarbon and pollen dating done some years later suggested that the weir dated from about 5,630 yr BP (Rosen et al. 1993; Newby & Webb 1994), and that the site was in use for about 1,000 years. The weir was installed onto peaty sediments (“lower peat” in Fig. 6.1) characteristic of salt marshes. The elevation of the ﬁsh weir was considerably lower than present-day intertidal levels; the weir caught ﬁsh2 at a time when the level of the sea was considerably different than at the time of its unearthing. The coastal environment changed drastically as sea level rose: marine deposits accumulated rapidly in what became a shallow, silty, ﬂat bay bottom. Further sea level rise created the opportunity for another period of establishment of a newer salt marsh habitat, evident as a shallower peaty layer (“upper peat” in Fig. 6.1). The discovery of the ﬁsh weir brought sharp attention to a number of issues, including the relative rise of sea level of the New England coast, a process that has shaped its deeply indented bays, and shaped its many islands and peninsulas. The long-term change in sea level, of course, must also have altered the distribution of coastal environments, a topic dealt with in Chapter 3.

125

Considerably after the weir became ﬁnally buried in accumulating sediments, French traders explored the area. In 1524–1525, Esteban Gomez, sent by Charles V to ﬁnd riches and a better passage to the Paciﬁc than the Strait of Magellan, coasted the area from Nova Scotia to Rhode Island, describing the coastline. In 1621 the self-exiled pilgrim Miles Standish described the estuary that is Boston Harbor today. The Europeans entering the bay and estuary found it fringed by green ribbons of salt marshes, with extensive mudﬂats exposed during low tides, and with great expanses of clean sands with eelgrass meadows and oyster reefs, all down-estuary of many streams ﬂowing down the drowned glacial melt valleys (Fig. 6.2).3 More than a century later, an English promoter (Wood 1634) tried to convince settlers to move to the Shawmut Peninsula—what was to be the heart of downtown Boston. The sales pitch described the area as “. . . very pleasant, hem’d in on the South-side with the Bay of Roxberry, on the North-side with Charles-river, the Marshes on the backe-side. . . .” The peninsula was “. . . a necke and bare of wood . . . [free from] . . . the three great annoyances of Woolues, Rattle-snakes and Musketoes.” This latter claim was a slight exaggeration, since the mosquitoes are still there, and as to the wolves, well, “. . . a little fencing will secure the . . . Cattle from Woolues”. The neck joining Shawmut to the mainland was low lying, mainly salt marsh subject to ﬂooding during high tides: the pathway from Boston to the mainland (now Washington Street) was “. . . well nigh impassable in the spring, when the horses waded knee-deep in water at full tides, . . . and the marshes on either side were the favorite hunting-ground of the sportsman” (Winsor 1880).

1

Material in this section is largely from Wood (1634), Winsor (1880), Johnson et al. (1942), Johnson (1949), Whitehill (1968), Morison (1971), Russell (1976), and Nixon (1982). 2 The ﬁsh weir idea is still presumptive. Nelson (1942) thought that the spacing among the brush wattles was too loose to catch ﬁsh, and instead suggested that the brush wattles laid in between the weir stakes were there to collect oyster spat. Such shellﬁsh management practices were known to many peoples from long ago. The Romans passed the practice to current-day Italians, who use bundles of brush (fascines) to collect spat. In Japan, bamboo is used for similar purposes, as are mangrove branches in Cuba.

3

The rich environmental resources seem unlikely to have impressed Gomez, who was sent by Charles V on a mission whose reward for Gomez was to be in proportion to the value of his discoveries. Finding no gold (“no hay alla de oro” is inscribed on a map of his trip), trading opportunity, or passage, Gomez impulsively took a load of people for slaves. Charles V had explicitly forbidden this, and forced Gomez to free the surviving Amerindians on arrival. There is no record whether Gomez actually sailed into the present-day Boston Harbor, but he did describe in some detail the coast from Maine to Cape Cod.

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18'0'' Boylston St.

Highest tide (April 16, 1851)

15'9.3''

Mean high water US Navy Yard Commonwealth Pier

10'5.2'' 10'2.6''

Mean sea level Commonwealth Pier US Navy Yard

5'7.8'' 5'6.8''

Mean low water, Commonwealth Pier Boston low water datum Mean low water, US Navy Yard Boston City Base

Lowest tide (Feb. 27, 1869) Commonwealth Pier

Recent fill

9.8'' 9.3'' 8.8'' 0.00

0.00 Upper peat

−2'8.6''

−12'6.4''

−12'10.9''

−14'1.4'' Shell layer 3

−6'10.2'' Silt

−9' 1 4 4 Main oyster bed 2 4 4 3 −13' −12'9.9'' Shell layer 2 Shell layer 1 Upper layer Lower peat

Sand layer Amorphous layer

Sand

Upper layer

Amorphous layer −12'10.9'' −14'1.4''

−14'1.9''

Upper wattle Lower wattle

−15'8.9'' −16'1.5'' −16'2.4''

Blue clay

Top of yellow clay

−23'8.4''

Figure 6.1 Vertical section under Boylston Street, Boston, showing the layers of different deposits, depths, and ages of the layers, as well as the position of the ﬁsh weir (inset). The weir consisted of vertical stakes, and smaller-diameter wattles interwoven among the stakes. The site of the section is shown as an “X” in Fig. 6.4. From Johnson et al. (1942), Rosen et al. (1993), Newby and Webb (1994).

LOSS OF COASTAL HABITATS

127

Figure 6.2 View of a nearly pristine salt marsh-fringed estuary, much as it could have been seen when the European settlement of New England began. The only modern mark is the ditching, mainly done during the 1900s to 1930s with some a bit later, to allow ﬁsh to enter marshes and prey on mosquitoes and other biting ﬂies. Nixon (1982) shows that ditching of salt marshes was an old practice: Eliot (1748) reports digging “a Ditch of four foot” though a marsh to a swamp, “to turn it into a Salt Meadow, that being the best I could do with it . . . the Tide now ﬂowing, where I suppose it never reach’d before.” The intent of such land management was to create a wetland covered with salt marsh grasses that were palatable to livestock.

From the 1600s on, the settlers came, inexorably, and brought with them practices of environment use well established in the Old World, but adapted to the new circumstances. In much of the British Isles, salt marshes were grazed by sheep and cattle. Livestock in New England were pastured on salt marshes if the peat could support their weight; alternatively, salt marsh grasses were cut by the farmers, and set to dry in cedar post staddles that held the crop of marsh hay above the high tides (Fig. 6.3). Salt marshes became important because they were one of the few environments where the laborious task of clearing virgin forest was unnecessary. In many New England towns, from the 17th to the 18th centuries, prized parcels of property were often a com-

bination of an area for the homestead, a forested tract as a wood lot for fuel and lumber for construction, and a parcel of salt marsh as a pasture for livestock. Today there survives on the North Shore of Massachusetts a modest industry that harvests salt marsh hay for use as horticultural mulches (this hay does not create a weed problem because the salt marsh seeds do not sprout in normal soils). The practice of poldering was widespread on the shores of the North Sea, with large areas “recovered” from the sea (see Chapter 10 frontispiece). Along the East Anglian and Dutch coasts, levees, ﬁlling, and pumping maintained land gained from the mudﬂats, sand dunes, and salt marshes, and allowed agriculture and other human

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Figure 6.3 Salt Marsh, Haystacks by Everett Warner (1877–1963), courtesy of the Florence Griswold Museum, Old Lyme, Connecticut. This pastel shows a view of a New England salt marsh used for the harvest of salt hay, seen piled onto staddles for dry storage.

uses. It is not therefore surprising, considering the English and Dutch origins of most settlers in New England, that residents on the Shawmut Peninsula soon began to extend its boundaries at the expense of the adjoining coastal environments. Within decades, the outline of the peninsula expanded, eventually obliterating large areas of coastal habitats—mainly salt marshes (Fig. 6.4). The gradual expansion of downtown Boston to today’s conﬁguration basically took place by ﬁlling in wetlands and mudﬂats. The major human alteration to the coastal environments took place between 1856 and 1894. During that period extensive areas of wetlands were ﬁlled with materials obtained by quarrying of the Trimountain Hills (Copps, Fort, and Beacon), which characterized the Boston “skyline” then; today only a street name, Tremont Street, and a remnant elevation

known as Beacon Hill remain of the three hills. It was these three hills that furnished the 6 m of “recent” ﬁll overlying the upper peat layer created by the most recent salt marshes (see Fig. 6.1). The value of salt marshes to early settlers in New England was clear: “All along the winding Massachusetts Bay shore, wherever salt grass caught the eye, exploring stockmen were petitioning the General Court to set up new townships” (Russell 1976). It was convenient to settle near wetlands, but the irony of the pattern of settlement was that as the newly established towns grew, they did so often at the expense of the adjoining salt mashes and other coastal environments. As centuries passed, larger and larger proportions of the coastal wetland area was converted to dry land by ﬁlling and by the creation of waterways by dredging.

129

LOSS OF COASTAL HABITATS

Cambridge

East Boston Boston Inner Harbor

Boston Neck

Charles River X Back Bay

N

0

1

2 km

Figure 6.4 Map showing the probable upland and shoreline of the Boston Harbor area in 1630 (black area), and the extent of land ﬁlling leading to the 1990s shoreline (gray area). The area designated Boston Neck on the map is the Shawmut Peninsula. The ﬁshing weir site in the Back Bay area is shown as an “X” (see Fig. 6.1). Adapted by Rosen et al. (1993) from Kaye (1976).

100 Area of salt marsh (ha × 103)

The expansion of European settlements radically altered the bounding coastal land and seascapes. Early data on the losses of coastal habitats in the area have not been compiled, but since 1886 about 50% of the salt marsh area in the northeastern United States have been lost (Fig. 6.5). This New England case history provides a time course of examples of human alteration of coastal environments, as the human population increased in density, and settlements became urban areas. The process, typical of such population centers, eventually leads to wholesale redesign of the coastal environments. These demographic developments changed the natural distribution of coastal habitats to a new array, with the remnant habitats

80 60 40 20 0 1886

1906

1922

1954 1968 1976

Year

Figure 6.5 Area of salt marsh lost in the New England states of the USA. Data from Nixon (1982).

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becoming human-dominated environments. Streets, buildings, piers, restaurants, theaters, museums, parks, and other features of a dynamic city—except for the occasional remnant bits of natural habitats, and the extension of bare, welldisturbed estuarine sediments—have replaced the huge extensions of salt marshes, mudﬂats, eelgrass beds, and oyster reefs.

Losses of coastal habitats Most coastal habitats have been altered to some degree by diverse agents of change. These may be globally or locally driven, and may result from natural and anthropogenic agents of environmental change, discussed in various chapters. This chapter highlights the loss of habitats by outright destruction of coastal environments by human activity, such as dredging, ﬁlling, draining, and so on. Human beings like to live near water, as pointed out in Chapter 1. Historically, we have found coasts attractive: we like the perspective, we ﬁnd food plentiful on coasts, and we like the proﬁts from coastal commerce. Our propensity to live near the coast has meant that many settlements have developed on coastal habitats. Alterations to rocky shores, shallow sea ﬂoors, seagrass meadows, and coral reefs are dealt with in other chapters on warming and sea level rise (Chapter 3), oil spills (Chapter 7), eutrophication (Chapter 12), overharvesting (Chapter 11), and the introduction of alien species (Chapter 11). Some of these environments have suffered a degree of direct damage from disturbances during ﬁshing; that topic is discussed in Chapter 11 (Thompson et al. 2002; Verity et al. 2002). Beaches of the world, as already discussed in Chapter 5, are affected by erosion and transport associated with sea level rise, locally accelerated by human coastal works. Loss of this habitat by outright construction on beaches is less widespread, so I am not treating beaches in this section. Seagrass meadows have suffered considerable worldwide losses (Short & Wyllie-Echeverria 1996; Duarte 2002), although statistics are lacking. Limited losses of seagrass meadows have been attributed to disturbances during shellﬁsh

harvests, dredging, and oil spills. The largest losses of seagrass meadows seem connected with changes in water quality, and as such, I discuss them later as part of the chapter on eutrophication (Chapter 12). Coral reefs are also under threat (Souter & Lindén 2000), but opinion on the relative importance of the different agents of change has changed since the 1990s. Earlier reviews, based on extrapolations of local damage information, were that human pressures (destructive ﬁshing with explosives,4 mining of coral rock and sand, engineering works, increased sedimentation, and so on) “. . . pose[d] a far greater immediate threat to coral reefs than climate change . . .” (Wilkinson & Buddemeier 1994).5 More recently, evidence on coral bleaching, incidence of disease, storms, and other symptoms have led to the view that accelerating global climatic change, through warming and sea level rise, poses a far more consequential potential threat to coral reefs (Wilkinson 1999; Almada-Villela et al. 2002; Sweatman et al. 2002). Reefs worldwide are showing slow to moderate rates of recovery from bleaching episodes, and there have been efforts at reef restoration (Wilkinson 2002). Since the newer information suggests that agents of change other than outright destruction may be affecting coral reefs at 4 Use of explosives to kill or stun ﬁsh is widespread in Southeast Asia and the east coast of Africa, where overly dense human populations are in dire need for food or incomes (Souter & Lindén 2000). The blasting is a short-term strategy of desperate people, since the explosions destroy the coral structures, thoroughly degrade the habitat, and severely impair the sustainability of the ﬁshery, as well as the rest of the ecosystem. Further discussion of the potentially large effects of removal of ﬁsh on reefs is provided in Chapter 11. 5 The early reports made estimates that 10% of the world’s reefs were irreparably damaged, a further 30% would be degraded after the turn of the 21st century, and a further 30% would be similarly damaged by about 2030 or so (Wilkinson 1999). HoeghGuldberg (1999) predicted that, if then-current global climate trends were to continue, there would be complete loss of coral refs during the 21st century. Wilkinson (1999) suggests that such predictions be regarded with caution. Clearly, there are severely degraded reefs in many localities. As it turns out, there are recent ﬁndings that there are larger areas of reefs in deeper waters, so the area of this habitat may be considerably larger than was realized (R. Cooke, Boston Globe, Feb. 24, 2004). Recovery from bleaching seems to be taking place. These new developments convey a degree of uncertainty as to the worldwide status of coral reefs. Much additional information is needed in this ﬁeld.

131

LOSS OF COASTAL HABITATS

Table 6.1 Condition of estuarine marsh areas in the Chesapeake and Delaware Bays, based on 1993 satellite imagery. Data from Stevenson et al. (2002). Condition (as % of the area of wetland)

Non-degraded

Slightly to moderately degraded

Severely to completely degraded

Chesapeake Bay Upper and middle bay Lower bay

31 28

50 52

19 20

Delaware Bay North shore (New Jersey) South shore (Delaware)

38 55

43 35

19 10

Sites

a global scale, I treat reef alterations in other chapters rather than here. On the other hand, there are other coastal habitats—salt marshes and mangroves—that have been far more susceptible to outright ﬁlling, dredging, construction, and other coastal works. This chapter therefore focuses on these coastal environments, and the changes affecting them during the past century. For each habitat, it also seems worthwhile to discuss just why we might be concerned about these direct losses. As in all the chapters, it will become evident that the separation of habitat loss caused by outright destruction from other sources of loss is more difﬁcult than might have been thought. Salt marshes6 Salt marsh refers to coastal wetlands with grass and herbaceous vegetation. These habitats are widespread across the world, and dominate shallow coasts with unconsolidated sediments, to the north and south of the distribution of mangroves in the tropics, up to the Polar circles. Salt marshes throughout the world have been ﬁlled, dredged, or otherwise directly destroyed for a long time, at least since Roman times in Britain, and for a thousand years in the Netherlands (Allen 6 Much of the material in this section is taken from reviews by Valiela et al. (2004) and Adam (2002).

2000). Humans have a long historical antipathy for wetlands. This antipathy is made evident by some of the words used in connection with wetlands. Fear of miasmas emitted by marshes, presumed to cause disease, gave rise to the word “malaria” (“bad air” in Italian) to describe gaseous efﬂuvia from wetlands. The very term “reclamation”, so often used to describe the process of eliminating wetlands, betrays a sense that we are somehow restoring a previously better state, “winning” land from the sea. The historical negativity, plus the fact that marshes are situated in the very places where rivers meet the coast, and hence are loci of important commerce, communication, and settlements, has meant that marshes have suffered extensive direct losses in area. Coastal wetlands in every part of the world have undergone signiﬁcant losses. A few examples from North America convey the dimensions of the problem. In the Chesapeake and Delaware Bays, only 28–55% of the estuarine marshes remained non-degraded by 1993 (Table 6.1). Orson et al. (1998) estimated that by the year 2000 in the state of Connecticut, USA, 45% of the salt marsh area would be gone, an estimated 41% would be in the course of destruction, and 14% was likely to remain in conservation. In San Francisco Bay, 79% of the tidal marsh habitats have been converted to human-dominated habitats (Fig. 6.6). There have been additional losses of open water areas in the bay, tidal ﬂats, and adjoining coastal

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Figure 6.6 Past (left) and present (right) area of tidal marsh in San Francisco Bay. A version of this ﬁgure appeared in Goals Project (1999); data ﬁles for this version are courtesy of Robin Grossinger of the San Francisco Bay Institute.

land covers; the percentage increase in humandominated environments is startling (Table 6.2). For the entire United States, comprehensive information is fragmentary (Table 6.3). For the coterminous USA, there must have been early losses parallel to those reported for New England (see Fig. 6.5). More recent data (Table 6.3) suggest that, up to the mid-20th century, there were losses of less than 10,000 ha yr−1, which amounted to less than 1% per year. For the USA, it therefore seems that about half the salt marsh area has been eliminated or transformed to some other land cover.7 In certain places, much more than

7

Losses of freshwater wetland in the USA have affected considerably larger areas than that of coastal wetlands; losses of freshwater wetlands have been estimated at about 53% of the area before European settlement (Dahl 1990). Especially hard hit have been areas such as the Florida Everglades and the forested wetlands in the southeast of the USA.

half the area is gone. In more recent decades, as the American public became more aware of the importance of coastal wetlands, losses have diminished notably, although there is still some destruction of salt marshes. The situation in the USA is by no means extraordinary: in the Thames estuary, for example, in southern England, over 70% of the marsh has been converted to dry land (Doody 1992). Most losses of salt marsh habitats have come about by direct destruction, including ﬁlling, draining, and other means of wetland “reclamation” to create land for urban development and roads, ﬁsh and salt ponds, and agricultural lands. We saw an example of such conversions in the Boston Harbor case history. Many more are available, for instance, in the Mondego estuary in Portugal (see Chapter 6 frontispiece), and in the polders of the North Sea (see Chapter 10 frontispiece).

133

LOSS OF COASTAL HABITATS

Table 6.2 Conversion of coastal wetland habitats in San Francisco Bay, across nearly two centuries, from natural systems to human-dominated land covers. Data from Goals Project (1999). Area (acresa)

Native aquatic habitats: Open bay water Tidal flats Tidal marsh Human-dominated aquatic habitats: Lagoons Salt pondsb Other altered areas Native coastal land habitats a1 acre

Ca. 1800

Ca. 1988

273,911 50,469 189,931

254,228 29,212 40,191

84 1,594 266

3,620 34,455b 155,021

89,357

23,286

% change −7 −42 −79 4,209 2,062 58,179 −74

= 0.4 ha.

bNow being restored to marshland under federal and state support and management (L. Valiela, US Environmental

Protection Agency, personal communication; http://www.southbayrestoration.org/Project_description.html).

Table 6.3 Losses of coastal wetlands in the United States (excluding Alaska and Hawaii), and in Louisiana, 1920s to 1990s. From data compiled by Mitsch and Gosselink (2000), Brady and Flather (1994), Dahl and Johnson (1991), and Gosselink and Baumann (1980). Area lost Years Coastal USA 1922–1954 1950s–1970s 1970s–1980s 1975–1985 1982–1987 Coastal Louisiana 1958–1974 1983–1990

ha × 103

ha × 103 yr−1

% loss of wetlands

% of original area lost per year

260 146 29 24 0.4

8.1 7.3a 2.9 2.4 0.06

6.5 – 1.7 1.1 1.1

0.2 – 0.15 0.11 0.18

10.8 6.6

0.86 –

Another value for this period of losses of 19 ha × 103 per year for the period 1954–1974 seems too high and was not included in this table.

a

In many coastal wetlands, other alterations, prompted indirectly by other human-driven agents of change,8 have been added to direct destruction of these habitats. This once again illustrates

the difﬁculty of separating the effects of diverse agents of ecological change.

8 For example, wetland losses in Louisiana (Table 6.3) (discussed in Chapters 3 and 5) were forced by sea level rise and sediment interception within watersheds, rather than by outright destruction (Reed & De Luca 1997).

The mass media and scientiﬁc press have widely reported losses of tropical environments such as

Mangrove forests9

9

Much of this section is taken from Valiela et al. (2001b).

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Table 6.4 Current mangrove forest areas, percent loss, annual loss rate, and percent of original area lost per year, for the mangroves of different continents and the world. Data from Valiela et al. (2001b).

Asia Africa Australia Americas World

Current mangrove area (km2)

% loss of mangrove area

Annual rate of loss (km2 yr−1)a

% of original area lost per year

77,169 36,529 10,287 43,161

36 32 14 38

628 274 231 2,251

1.52 1.25 1.99 3.62

166,876

35

2.834

2.07

aAnnual loss rates are calculated from the mean number of years between the earliest and latest information available

for the countries within each continent. For the world, for example, the data average out to span the period 1980–1997.

the felling of rain forests and the bleaching of coral reefs. This well-merited attention has created a worldwide constituency that supports conservation and restoration efforts in both these threatened ecosystems. The remarkable degree of public awareness and support is manifest in beneﬁt rock concerts at Carnegie Hall, and in the designation of ice cream ﬂavors after rainforest products. Mangrove forests are another important tropical environment, but one that has received much less publicity. The anonymity of mangroves is in spite of concern about the magnitude of losses of such forests, which has long been voiced in the specialized literature (Saenger et al. 1983; Spalding et al. 1997). Mangrove trees grow ubiquitously as a relatively narrow fringe between the land and sea, between latitudes 25°N and 30°S. They form forests10 of salt-tolerant species, with associated complex food webs and ecosystem dynamics (Macnae 1968; Lugo & Snedaker 1974; Tomlinson 1986). The destruction of mangrove forests is occurring globally. Losses of mangrove forests across 10

I use “mangrove forests” where others have used “mangrove swamp”. “Swamps” are variously used in different places of the world to describe wetlands (Mitsch & Gosselink 2000). The term “swamp” applies more narrowly to wetlands with trees and shrubs on them, so it is appropriate to use for mangrove habitats; I arbitrarily used “mangrove forests” instead here. I do, however, prefer “mangrove swamps” to “mangal”, another term used to refer to mangrove forests. The latter sounds, to my non-anglophone ear, too much like the name of a sugar beet (mangel) grown in Europe.

Table 6.5 Uses of mangrove forest area leading to loss of habitat. The data used in this table cover 66% of the world’s mangrove forest area; data were not readily available for the remainder area. Data from Valiela et al. (2001b), compiled from many sources.

Shrimp culture Forestry uses Fish culture Diversion of fresh water Land reclamation Herbicides Agriculture Salt ponds Coastal development

World total (103 km2)

% of total

14 9.5 4.9 4.1 1.9 1 0.8 0.05 0.05

38 26 14 11 5 3 1 – –

the tropics have reached alarming proportions. On a worldwide basis, perhaps 2% of the mangrove forest area has been lost per year since 1980, and this is higher in certain areas such as the Americas (Table 6.4). These losses have amounted to an estimated loss of 35% of the world’s area of mangroves since 1980. More detailed regional studies (for example, Honculada-Primavera 1995; Blasco et al. 2001) conﬁrm the worldwide losses. These loss estimates make mangrove forests the most threatened major coastal habitat in the world.

LOSS OF COASTAL HABITATS

135

Figure 6.7 Aerial view of a mangrove forest in Borneo, showing dykes and enclosed shrimp ponds carved out of mangrove habitat. From Valiela et al. (2001b), photo by Frans Lanting, Minden Pictures.

There are many causes of mangrove destruction (Table 6.5). Global changes such as increased rise of sea level may affect mangroves (Ellison 1993; Field 1995), although accretion rates in mangrove forests may be large enough to compensate for the present-day rise in sea level (Field 1995). Human alterations, mainly conversion of mangroves to mariculture (Fig. 6.7), agriculture, and urbanization, as well as forestry uses and the effects of warfare have led to the remarkable recent losses of mangrove habitats (Table 6.5) (Saenger et al. 1983; Fortes 1988; Marshall 1994; Honculada-Primavera 1995; Twilley 1998). As in all cases of environmental damage, there are good economic and social reasons for the losses. Aquaculturists who raise shrimp and ﬁsh in ponds dug out of erstwhile mangrove forest areas (Fig. 6.7) do so because of imperative economic pressures. In warmer waters fringed with mangroves, there is often artesanal harvest of mullet, shrimp,

mangrove oysters, and mangrove cockles from Mexico to Peru (Mackenzie 2001), but these more sustainable activities fail to furnish the short-term economic incentives of intensive maricultural practices.11 The lack of alternative economic activities forces the loss of mangrove habitats. 11

Shrimp culture may be highly proﬁtable in the short term, but costly, and perhaps unsustainable, in the long term, and is a difﬁcult environmental challenge (Paez-Osuna 2001a, 2001b). Shrimp culture requires stocking the ponds with wild-caught juvenile shrimp, a practice that has decimated wild shrimp stocks in certain areas such as Ecuador. The intensive feeding and use of biocides and antibiotics required to maintain fast shrimp growth under intensive culture somehow poisons ponds after a few years (PaezOsuna et al. 2003; Visuthismajarn et al. 2005), so many ponds are abandoned. Shrimp pond efﬂuents have high concentrations of nitrogen and organic matter, which promotes eutrophication down-estuary (Naylor et al. 1998; Paez-Osuna et al. 1999; AlsonoRodriguez & Paez-Osuna 2003; Jackson et al. 2003), which has manifold impacts, including alteration of down-estuary vegetation and habitats (Ruiz et al. 2001).

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Significance of coastal wetland habitat losses It is evident that there have been major losses of salt marsh and mangrove forests from the coasts of the world, by outright destruction of the habitats. Do these considerable losses of coastal wetlands matter? To answer this question it is useful to review the roles that these environments play in the mosaic of interlinked coastal landscapes. Salt marsh functions Export of energy-rich materials important to food webs of deeper waters

Most salt marsh ecosystems export energy-rich substances (reduced nitrogen compounds, dissolved and particulate organic matter) to deeper waters (Table 6.6). These exports may amount to a considerable subsidy that supports the metabolism of the receiving ecosystems, as suggested some years ago in the “outwelling” hypothesis of E. Odum, A. de la Cruz, and J. Teal. For example, measured exports of organic matter from Georgia salt marshes were large enough to furnish the energy needed to support the high rates of metabolism in the near-shore Georgia Bight ecosystem (Hopkinson 1985). The role that salt marshes

Table 6.6 Percentage of salt marshes (n = 19) that exported materials out to deeper waters. Data from Valiela et al. (2001a), compiled from many sources.

Materials Ammonium Nitrate Dissolved organic nitrogen Particulate organic nitrogen Total nitrogen Dissolved organic carbon Particulate organic carbon Total carbon

% salt marshes that exported materials to deeper waters 64 36 100 67 100 91 59 82

play through the export of materials that support populations in deeper waters has become a leading argument for the conservation of salt marshes, at least in the West Coast of the North Atlantic. Nurseries to many species, including commercially important fisheries stocks

Many species of commercial and ecological importance use coastal wetlands as nurseries (Robertson & Duke 1987; van der Velde et al. 1992; Nagelkerken et al. 2000). An example of the importance of coastal wetlands to coastal stocks was given by Turner (1992), who reported that shrimp yields along the coast of the Gulf of Mexico were proportional to the area of coastal marsh landward of the harvest area (Fig. 6.8 top). During the 20th century the shrimp harvest fell, and the reduction in catch was correlated to the cumulative loss of marshland in the Louisiana area (Fig. 6.8 bottom). These relationships follow from the nursery role played by the marshes for juvenile shrimp, as well as from the energy-rich materials exported from these marshes to deeper waters offshore. Many commercially important species use salt marshes as foraging areas during their early life stages. These include ﬁsh such as menhaden (a species that contributes a major proportion of the North American ﬁsh harvest), blueﬁsh, and striped bass, among many other species of commercial note. A study by Werme (1981) provided evidence as to why ﬁsh from deeper waters might use wetlands as nursery areas (Table 6.7). Within the salt marsh estuary, the size of ﬁsh that resided yearround in the estuary did not differ from that of the juveniles of species whose adults live in deeper waters (“invaders” in Table 6.7). There are disadvantages to having a larger size in these systems, because larger ﬁsh are easily stranded in shallow pools at low tide, and are readily eaten by the many top predators (herons, egrets, terns, and many other birds) often common in salt marshes. Juveniles from deeper-water species had fuller guts, and were far more carnivorous than resident ﬁsh (Table 6.7). The invaders achieved these feats by their larger gapes, which

LOSS OF COASTAL HABITATS

Shrimp yield (kg yr−1)

108

based, at least in part, on higher relative abundance of larger food items, accounts for the nursery function of salt marshes for juveniles of species from deeper waters. An additional, if minor, consequence of the nursery role of wetlands is that the movement of the erstwhile juveniles to deeper waters add up to a measurable, but modest, export of larger, marsh-grown units of highquality organic matter to coastal waters (Deegan 1993).

y = 3.709x 1.082 r 2 = 0.94

107

106

105 104

105

106

107

Habitat for shellfish and finfish stocks

Area of vegetated estuary (ha)

Annual catch (mt)

1,600 1,200 800 400

C = 1353 − 6.3*(Area) R 2 = 0.85

0 0

20

137

40

60

80

100

Cumulative area (km2)

Figure 6.8 Top: harvest by the shrimp ﬁshery off Louisiana in relation to the area of adjoining coastal wetland. Bottom: metric tons of shrimp caught annually off Louisiana in relation to the cumulative loss of adjoining coastal wetland. From Valiela et al. (2001b), drawn from data compiled by Turner (1992).

arise from the allometry of vertebrates: the young have relatively larger heads than adults, and, as it turns out, invaders are all juveniles. Invader species, on average, therefore had access to larger food items than the resident species (Valiela et al. 2004). Feeding on larger prey places stringent bounds on the abundance of invaders, because, as is well known, larger prey are much less abundant than small prey. Invaders have much lower densities than those of resident species, but achieve growth rates an order of magnitude larger than resident species. Thus, although obligatorily less numerous, juvenile invaders from the adjoining deeper-water environment achieved fast growth rates in salt marsh estuaries. This faster growth,

The shallow, protected bays, inlets, and lagoons that are fringed by wetlands are rich in phytoplankton and other particles that are prime food for suspension feeders, and support reasonably dense faunas of other consumers. Wetland-fringed environments are almost inevitably areas where humans harvest a variety of stocks. For example, oysters, quahogs, scallops, soft-shell clams, blue crabs, and winter ﬂounder are among the many valuable crops taken from such environments in temperate latitudes of eastern North America. The value of crops from eastern North American marsh-fringed environments are typically an order of magnitude larger, on a per unit area basis, than those obtained from grains in terrestrial agriculture (Mackenzie 1989; Ver et al. 1999). Sites for aquaculture and other commercial crops

The food-rich waters in shallow waters fringed by wetlands are potentially useful sites for mariculture efforts (Shumway et al. 2003). For example, it is common practice for New England municipalities to issue permits for commercial growing of clams in marsh-fringed estuaries. The culture of bivalves may have other beneﬁts: for example stocks of suspension-feeding shellﬁsh grown in high densities in wetland-fringed estuaries could be a tool to improve or restore water transparency, as has been argued in attempts to restore oyster banks in Chesapeake Bay (Cloern 1982; Ulanowicz & Tuttle 1992). A concern with the cultivation of shellﬁsh at high densities is the production of fecal material by the shellﬁsh, a process that adds organic matter to sediments,

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Table 6.7 Comparisons between species of ﬁsh resident in Great Sippewissett Marsh and species of ﬁsh whose adults live in deeper water, but whose young invade salt marsh estuaries. Results of paired t-tests are non-signiﬁcant (n.s.), signiﬁcant at 0.05 (*), or highly signiﬁcant (**). From Valiela et al. (2001a); data from Werme (1981).

Mean length of fish in marsh (cm) Mean length of adults (cm) % full guts % carnivory in diet Mouth gape (mm) No. fish/100 m of shoreline % growth/month

Resident speciesa

Invader speciesb

Paired t-test

41 ± 4 106 ± 19 26 ± 4 48 ± 13 1.9 ± 0.3 57 ± 31 18 ± 4

59 ± 9 422 ± 121 59 ± 9 78 ± 16c 4 ± 0.6 0.8 ± 0.3 127 ± 27

n.s. * ** ** ** ** **

a

Resident species included Menidia menidia, Apeltes quadracus, Fundulus heteroclitus, Fundulus majalis, and Cyprinodon variegatus. b Invader species included Alosa pseudoharengus, Brevoortia tyrannus, Gasterosteus aculeatus, Tautoga onitis, Centropristes striatus, and Pseudopleuronectes americanus. c > 90% if Brevoortia tyrannus was excluded.

depletes near-bottom oxygen, and alters the composition of the fauna beneath the culture rafts, sometimes favoring the fauna by adding food and sometimes increasing mortality of the sea organisms (Tenore & Gonzalez 1976; Crawford et al. 2003; Danovaro et al. 2004; Harstein & Rowden 2004).12 One other commercial crop has been the extraction of salt in salt pans constructed on what was salt marsh (see Chapter 6 frontispiece). In Western Australia evaporation ponds occupy extensive areas of what were salt ﬂats, salt marsh, and mangroves (Adam 2002).13 This practice was more important in the past than at present, since mining of rock salt is more practical, at least in economically developed countries. Salt pans in South San Francisco Bay ﬁrst operated in 1854, were widespread, and have been gradually acquired for conservation purposes during the 1970s to the 1990s.

12

More on ﬁnﬁsh culture and its environmental effects can be found in Chapter 11. 13 Evaporation ponds are not necessarily always detrimental to wildlife. In Bonaire, a leading exporter of sea salt, the extensive evaporation ﬂats have become home to over 10,000 ﬂamingos, probably one of the largest congregations of these birds in the Americas.

Contaminant interception

To a certain extent, salt marsh sediments retain contaminants of many kinds, including heavy metals (cf. Chapter 9), chlorinated hydrocarbons (cf. Chapter 8), and petroleum hydrocarbons (cf. Chapter 7). To the degree to which any contaminant from land is buried in marsh or mangrove sediments, these wetlands are preventing more widespread contamination of coastal waters. As will be discussed in a later chapter, heavy metals seem to have few discernible effects on wetland plants or marsh fauna (cf. Chapter 9). Shoreline stabilization

The roots and rhizomes of marsh plants add coherence to sediments, as do the dense roots of mangrove trees (Savage 1972). By their very presence these macrophytes therefore consolidate otherwise loose sediments. Storms often fail to disturb marsh sediments that are covered by grasses (Valiela et al. 1996). Salt marsh canopies also, to an extent, offer a buffer that dissipates wave energy, and hence lowers erosion of vegetated sediments. Salt marshes have other important engineering features: for example, the narrower the fringe of salt marsh before a protective berm,

LOSS OF COASTAL HABITATS

the more likely that storm surges will overtop the berm (Brampton 1992). Sources of forage and hay

Marshes have long been used as grazing lands where livestock assimilate the forage as well as obtain essential salts. Grazing by livestock and making of hay may have taken place in about 4000 BC in salt marshes in the Danish Baltic and by 600 BC in Dutch salt marshes (Adam 2002). Remnants of such practices can be seen in various parts of the world where salt marshes are still used as pastures. For example, visitors to most Scottish marshes will see grazing sheep on them. In the Americas, making of hay and pasturing livestock on marshes began as early as 1650. Many colonial-era property deeds in Massachusetts included a parcel of marshland for pasture purposes. Marshland was highly desirable because little work—only some logging and uprooting— was needed to create pasture. During more recent centuries in the East Coast of North America, marsh grass was cut, and the dry matter used as marsh hay to feed livestock and as horticultural mulch. Currently, there is a small market in developed countries for salt marsh hay, valuable because it does not sprout weeds when used as garden mulch. The use of salt marshes as pasture is also a pattern in the Southern Hemisphere. Cattle in coastal areas of Buenos Aires Province, Argentina can frequently be seen feeding on salt marsh hay; ranchers often burn the salt marsh swards to promote the appearance of protein-richer young shoots (P. Martinetto, personal communication), a practice also used in North Queensland (Anning 1982).

139

wetland habitats is therefore essential for this charismatic species, but there are many other waders, shorebirds, birds of prey, herons, egrets, and so on, that, though less well known to the public, also depend on such wetlands during migration for refueling and rest. Interception of land-derived nutrients

Wetlands intercept certain material transported from land and moving toward the sea. Note, for example, in Table 6.6 that nitrate is exported by a minority of coastal marshes: marshes tend to intercept rather than export nitrate, a major form of the nitrogen that limits the growth of most coastal algae and plants in most coastal environments. Salt marshes may provide a substantial interception of land-derived nitrogen because of their position between land and estuary, as well as because of their high rates of denitriﬁcation and burial of nitrogen (Valiela 1983). An example of the potential importance of nitrate interception is evident in the relationship between salt marshes and adjacent seagrass meadows. Evidence for this relationship is that the larger the area of salt marsh (or of mangrove forest), the greater the production by seagrasses in adjoining meadows (Fig. 6.9 top), and the smaller the loss of seagrass meadows that took place as nitrogen loads increased (Fig. 6.9 bottom). Seagrass meadows are highly sensitive to increased nitrogen loads: interception of land-derived nitrogen in coastal wetlands can be interpreted as an important protective ecological “subsidy” furnished by salt marshes (and mangrove forests) to adjoining coastal environments such as seagrass meadows. Aesthetic value

Waterfowl refuges and migratory stop-over sites

A large and diverse set of migratory birds depend on having adequate wetland areas as stop-over sites during migration and as wintering sites. Most of the population of European storks passes through or winters in the threatened Coto Doñana wetland in southern Spain, plus what is left of the Nile delta wetlands. Preservation of these

In discussions about coastal wetland conservation, there is a certain hesitation to mention that many of us simply take great pleasure in wetlands and the organisms and settings they provide. One cannot—we are tempted to say ought not— put a price tag on this, but the aesthetic appreciation of wetlands by many people is a powerful force that should be harnessed. Arguments

140

CHAPTER 6

(Seagrass production/total production)/100

100

% sg production = 1.117(% wetland) + 2.668 F = 33.5***

80 60

that it seems unnecessary to develop details. Here I provide an abbreviated treatment of the functions, noting references that provide the speciﬁc evidence. The functions of mangroves include the following:

40

1 20 0 0

10

20

30

40

50

60

2

(Wetland area/total estuary area)/100

% Seagrass area lost (%)

100

% sg lost = 118.130* 10−0.019(% wetland ) F = 6.7*

80

3 4

60 40

5

20 0 0

10

20

30

40

50

60

(Wetland area/total estuary area)/100

Figure 6.9 Top: seagrass production expressed as a percent of the total production in many estuaries, plotted versus the area of fringing wetland expressed as a percent of the total estuary area. Bottom: percent of the area of seagrass habitat lost (over the last 10–30 years) plotted versus the percent of the area of the estuary made up by fringing wetland. From Valiela and Cole (2002).

highlighting the aesthetic worth of wetlands, as well as the list of natural subsidies furnished by wetlands, need to be mustered, and repeated, in reaching out to the public and politicians. Success in raising awareness of wetland losses might ensure that more people will continue to delight in wetlands. Mangrove forest functions Mangrove ecosystems provide a series of ecological functions similar to those of salt marshes, so

Interception of land-derived pollutants and suspended matter before these contaminants reach deeper water (Marshall 1994; RiveraMonroy & Twilley 1996; Tam & Wong 1999).14 Export of energy-rich materials that support near-shore food webs, including shrimp and prawns (Rodelli et al. 1984; Twilley 1988; Sasekumar et al. 1992). Prevent coastal erosion by stabilizing sediments (Marshall 1994; Tam & Wong 1999). Furnish nursery and spawning areas for commercially important coastal ﬁsh and shellﬁsh species (Rodelli et al. 1984; Sasekumar et al. 1992; Mumby et al. 2004). Provide stop-over sites for migratory species (Saenger et al. 1983).

Mangrove forests are also habitats for a diversity of species that have considerable importance, both economically and ecologically. The fauna of mangroves includes ﬁsh and shellﬁsh taxa that support subsistence ﬁshing, as well as rare endemic proboscis monkeys in Borneo, the scarlet ibis and the vulnerable straight-billed woodcreeper in Trinidad, threatened Bengal tigers in India and Bangladesh, rare Bulbophyllum and other orchids in Singapore, endangered manatees in Florida, and long lists of other key species (Saenger et al. 1983). Mangroves also furnish various extractive beneﬁts, including wood for the production of charcoal, extraction of tannins, paper pulp, ﬁrewood, lumber, as well as honey, mariculture crops, salt, and so on (Saenger et al. 1983; Spalding et al. 1997).

14

Nutrient processing down-estuary from mariculture facilities might mitigate eutrophication caused by efﬂuent from mariculture ponds built in mangrove habitats (Wolanski et al. 2000). Burial and denitriﬁcation within mangrove forests may signiﬁcantly lower export of terrigenous nutrients to deeper waters (Robertson & Phillips 1995; Rivera-Monroy & Twilley 1996; Rivera-Monroy et al. 1999).

LOSS OF COASTAL HABITATS

For all these reasons any loss of mangrove forest means a loss of important subsidies to subsistence uses, and ecological, economic, and conservation functions.

141

unlikely to share our priorities about conservation of wetlands in their natural state.

Coastal wetland restoration efforts Conclusion It should be apparent from the list of functions of coastal wetlands just reviewed that these habitats play important multifaceted functions in the coasts of the world. Coastal wetlands are of signiﬁcance by themselves, though they also play a large role in linkages that affect adjoining environments in important ways. The large worldwide loss of area of these habitats has to be considered as part of the major alteration to coastal environments in general. These losses should be of concern to people. The extent of loss of this key environment places added pressure on the maintenance and restoration of what remains of this habitat. Economic pressures have historically been more powerful than arguments for the conservation of coastal wetlands. The reason we have lost wetlands is that we have repeatedly made the economic decision that other land covers are more proﬁtable and desirable. If we believe it is worthwhile to maintain coastal wetlands as natural systems, we will have to redouble efforts to help the public and political sectors of society better reconcile the balance between economic imperatives and the less apparent beneﬁts provided by coastal wetlands. The items listed above are reasons we can muster to point out that human interests in coastal wetlands might include many important natural subsidies, rather than just cash crops or building sites. In some quarters it has become fashionable to develop valuations of ecological features, seeking equivalencies of natural services with currency. Two problems with this approach are, ﬁrst, that in many cases it simply is not realistic to make such conversions; in almost all cases, the methodology used in valuation does not withstand critical scrutiny, although space precludes an exegesis of the procedures here. Second, and far more important, is that any time a price is placed on anything, it is for sale, and the highest bidder is

It is to some extent reassuring that a rising awareness by public and political sectors has prompted wetland conservation efforts in many countries. Among these efforts are widespread attempts at the restoration of coastal wetlands. Salt marsh restoration efforts of some magnitude are under way in a number of places in the USA, including the Delaware River estuary, North Carolina, the Paciﬁc Northwest, southern New England, Louisiana, and California (Weinstein & Kreeger 2000; Zedler 2001; Craft et al. 2003). There are plans to restore extensive areas of salt ﬂats and evaporation lagoons in South San Francisco Bay to their original salt marsh status (L. Valiela, US Environmental Protection Agency, personal communication; http://www.southbayrestoraion. org/Project_Description.html). Many European sites report restoration sites in which efforts are under way to rebuild or create new areas of salt marshes (Crook et al. 2002), part of a long history of European wetland manipulation. Similarly, there are innumerable mangrove restoration projects ongoing in the Caribbean, South America, Southeast Asia, and many other places. These are particularly important as means to restore areas that were converted to shrimp ponds, and were subsequently abandoned after the maricultural efforts ceased. Quite often there is a natural regrowth of marsh and mangrove vegetation if disturbed sites are left to be recolonized (Fig. 6.10). Mangrove and salt marsh restoration efforts have had mixed results, and there is some question as to whether restored marshes and mangroves fully achieve functions equivalent to the original habitat (Ellison 2000; Zedler 2000; Warren et al. 2002). Criteria that might improve the success of salt marsh and mangrove restoration have been itemized (Kaly & Jones 1998; Ellison 2000; Zedler 2001), but at the very least, re-establishing salt marsh stands and mangrove canopies is a start to developing functioning coastal wetlands, and should be pursued.

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References

Figure 6.10 Images of the same site (Hollywood, Florida, USA) in 1989, after the site had been prepared (top), in 1991, after mangrove seedlings had naturally colonized the site (middle), and in 1996 (bottom). Planted seedlings also have similar effects. Modiﬁed from Lewis (2005).

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Chapter 7 Petroleum hydrocarbons1

Oil slicks ﬂowing around island salt marshes. From NRC (2003).

A case history: the Exxon Valdez accident The very early spring of 1989 was unusually cold in southern Alaska (Spies et al. 1996). Wintering sea ducks still remained in Prince William Sound in huge numbers. Salmon fry were just about ready to leave their natal freshwater streams and migrate to coastal waters. Paciﬁc herring were about to enter the sound. Young sea otters and

seals had been born, and masses of sea birds were arriving to start their nesting season. Almost a million and a half barrels of oil from the North Slope of Alaska, transported across the entire span of Alaska via the 1,287 km of the Trans Alaska Pipeline from Prudhoe Bay to Valdez, 1

Much of the content of this chapter was garnered from conversations over the years with John Farrington, John Teal, Judy McDowell (Capuzzo), and Bruce Tripp, all of the Woods Hole Oceanographic Institution.

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Grounding site KENAL PENINSULA

3/27 64 km 3/30 144 km

Cook Inlet

ALASKA PENINSULA

4/3 225 km

200

of

St

ra it

4/7 290 km

Sh el ik

Figure 7.1 Map of southern Alaska, with the area affected by the Exxon Valdez oil spill shown in black. The inset shows the area of the map relative to the whole of Alaska, with the dashed lines showing the dates of the spread of the spill. Adapted from http://www.unu.edu/unupress/un upbooks/uu21le/uu21le0l.htm.

Prince William Sound

4/11 402 km 4/30 450 km

Kodiak Island

GULF OF ALASKA

5/2 563 km 5/18 756 km

had been loaded onto the 300 m-long supertanker Exxon Valdez. As had uneventfully happened for more than 8,700 trips during almost 12 years, the Exxon Valdez left port, a routine that moved 2.1 million barrels per day of North Slope oil out of the ice-free port of Valdez, through the narrow Valdez Arm of Prince William Sound, to points south. During the early hours of March 24, 1989, shortly after the Exxon Valdez departed, Captain Joe Hazelwood was in his cabin, having left a less experienced hand at the helm. In an effort to avoid calved icebergs, Exxon Valdez went off its planned course, and struck Bligh Reef in Prince William Sound, Alaska (Fig. 7.1). Captain Hazelwood, roused from his quarters, rushed to the ship’s bridge and after a quick look, radioed a scratchy message ashore, reporting that the ship had “fetched up hard aground”, and was “evidently leaking some oil”. In the aftermath, some 35,000 tons of crude oil were released into the waters of the sound, the largest oil spill on the shores of the USA (Spies et al. 1996). Approximately 40% of the oil found its way to the intertidal shores of Prince William Sound. Currents and wind from a northerly gale

carried the ﬂoating oil to the southwest, and eventually about 10% of the crude oil reached the shores of the Kenai Peninsula and the lower Cook Inlet (Fig. 7.1).2 The immediate effects of the spill were widely reported in the press, and images of oiled birds (Fig. 7.2), dead otters, oil-encrusted rocks, and oil-stained workers engaged in the clean-up (Fig. 7.3) were featured prominently. The toll of this spill will never be accurately gauged, but there were impressive initial effects. It was claimed that perhaps a quarter of the shorelines surveyed were exposed to oil,3 and that hundreds of thousands of birds, thousands of sea otters, hundreds of harbor seals and bald eagles, and tens of killer whales, and uncountable smaller organisms, were killed. The spill prompted a massive response to assess the damage, as well as to restore the affected environments—reaching an estimated cost of US$3.2 billion spent by 1996. The research and restoration work took place 2

The drift of the slicks spanned almost 800 km, a quite substantial distance, equivalent to that between Boston and Washington, DC, Paris and Barcelona, or the latitudinal span of Great Britain. 3 Later estimates were that 10 or 18% of the shorelines had been oiled, according to estimates by the Exxon Corporation or to Alaska’s Department of Environmental Conservation (Wheelright 1994).

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Figure 7.2 Sea bird coated by the spilled oil from the Exxon Valdez accident. Photo from http://www. fathom.com/coarse/21701790/session5.html.

under dauntingly difﬁcult conditions, stressful weather, and complicated logistics of work in remote locations, all complicated by the pressure of time, the press, and public opinion; and all the work was shadowed by the litigation to assign responsibility and punitive ﬁnes. Intensive efforts were made to remove oil from the water and shores from mid-1989 through 1992, as demanded by the existing laws. These efforts included, principally, the application of hot or cold water under high pressure (Fig. 7.3) to wash oil away, as well as some mechanical removal and hand cleaning of oiled rocks, and nutrient fertilization of beaches to enhance microbial degradation of the oil (Mearns 1996).4 Skimming of ﬂoating oil may have removed about 8–9%, and mechanical removal of oil from the shore may have accounted for about 5–6% of the spilled crude oil (Spies et al. 1996). The majority of the spilt volume remained in the sea, where it was subject to a variety of processes as time passed. Soon after 4

Claims that the fertilization successfully removed oil (Pritchard & Costa 1991) were not supported by later evidence (Stone 1992).

Figure 7.3 High-pressure steam and hot water cleaning of the shore exposed to oil from the Exxon Valdez. Photo from http://evostc.state. ak.us/facts/photo19.html.

the spill, about 20% of the spilled oil evaporated, about 50% was stranded in beaches, and about 14% was recovered by clean-up crews. Mixing of oil into the water column, microbial activity, and photochemical oxidation further dispersed, or degraded, much of the oil, so that by the fall of 1992, only about 2% of the oil may have remained in beaches, and 13% in sediments (Spies et al. 1996). The intensive effort5 to deﬁne effects of the Exxon Valdez oil spill is reﬂected in the several hundred papers that have been published on the subject. The immediate impact and the longer-term 5 At the height of the response to the Exxon Valdez accident, more than 11,000 personnel, 1,400 vessels, and 85 aircraft were in action (NRC 2003). No other oil spill compares with the intensity of the response effort.

PETROLEUM HYDROCARBONS

response of many species have been examined6 as a follow-up on the Exxon Valdez spill. Here we can only summarily review some major features reported in just a few of the many references. Intertidal rockweed cover was highly sensitive to oil, but regrew actively and recovered7 within 5 years. Oiled parts of the coast lost a substantial part of the rockweed canopy compared to nonoiled shores in the immediate aftermath of the spill (Fig. 7.4; Table 7.1). The initial loss of algal cover, however, was even greater where highpressure washing took place. The oil and washing also removed other organisms, including grazers that fed on the rockweeds. Up to 1993–1994, the rockweeds actively regrew on rock surfaces presumably relatively free of grazers,8 so that in 6

Comparisons of the immediate responses of the various species were done either by data from oiled vs. un-oiled shores, or post-spill vs. pre-spill information. In almost all cases there were difﬁculties with these comparisons. Comparisons of data collected from arrays of oiled vs. un-oiled sites inherently include preexisting differences among the sites. These inherent differences do not invalidate comparisons, but do require greater degrees of observation replication, and care with undetected biases conveyed by the unplanned choice of sites. Pre-spill data were often incomplete or largely absent, and there were signiﬁcant multiyear changes in the environment of Alaska, all of which certainly altered conditions, confounding pre- vs. post-spill comparisons. 7 The issue of recovery was another problem in the interpretation of information about accidental effects. Much controversy was generated by different and unrealistic conceptions of what recovery might mean. The government’s view was that a resource recovered when it returned to where it might have been before the damage. This is far too general and unrealistic. The Exxon-supported view was the recovery was complete when characteristic plants and animals returned and were functioning. This view fails to incorporate the many features of function of natural systems beyond mere presence of the right taxa. In addition, the way researchers tried to assess recovery differed substantially, leading to different conclusions. Some preferred to compare abundances, reproductive rates, abnormalities of morphology, rates of growth, molecular traces of oil exposure, and many other variables, usually in the context of a comparison of oiled vs. non-oiled sites (cf. previous footnote). Others opted to compare time courses of variables such as abundance, and ascertain whether the slopes of time courses were parallel across years or not (this was an attempt to compensate for possible bias inherent in the assignation of different sites to the oiled vs. nonoiled categories). In most cases the nature of the available data determined what the testing strategy was, but in any case, the results of the different approaches may have altered the conclusions reached. 8 A “greening” of habitats is a common follow-up to an oil spill. Additional nitrogen released from dying organisms may be responsible for the quick responses a few days after a spill; longer-term responses may be related to absence of grazers, as in the Exxon Valdez example.

149

oiled and oiled and washed parcels rockweed cover became more extensive than in non-oiled areas (Fig. 7.4). By 5 years after the spill, grazer settlement may have restored grazing pressure sufﬁciently that all areas showed similar rockweed covers. There were no apparent effects of the oil on kelp forests (Table 7.1). The slight initial damage to seagrass meadows was gone by 1991 (Table 7.1). Eelgrass environments were more impacted than kelp beds, and were slower (6+ years) to recover their faunal assemblages than kelp beds (2 years). Oil persisted longer in soft sediment environments where seagrasses grew than in the hard substrates under kelp (Dean & Jewett 2001). With time, however, oil concentrations diminished even in soft sediments. For example, oiled salmon spawning habitats may have led to reduced embryonic survival before 1993, but petroleum hydrocarbon concentrations were lowered below critical thresholds by 1994 (Murphy et al. 1999). Some invertebrates were initially affected by oiling but others were not; in the weeks and months after the spill, the abundance of some invertebrates decreased, and others became more numerous (Table 7.2). A species of clam, however, was sensitive to the residual oil in sediments even 5–6 years after the spill (Table 7.1). The responses of ﬁsh were equally varied. Paciﬁc herring in Prince William Sound were widely exposed to the oil, since the stocks were coming into the breeding areas as the spill occurred. Herring in 1989 suffered increased physiological damage, larval mortality, and reductions in growth (Table 7.1). These impairments appeared to have disappeared by 1990.9 Pink and sockeye salmon, a fundamental resource to various coastal environments10 and to the human population of Alaska were a prime 9

These responses are similar to those reported elsewhere. For instance, after the 1978 Amoco Cadiz spill, growth rates of sole (but not of plaice) decreased signiﬁcantly for the year of the spill, and returned to initial conditions by 1979 (Conan 1982). 10 For example, the carcasses of Paciﬁc salmon that die after spawning in the shallow streams where they spawn supply perhaps a quarter of the nitrogen supporting the growth of streamside vegetation; better vegetative growth then indirectly supports maintenance of the high quality of the streams, as well as the salmon stock (Helﬁeld & Naiman 2001).

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CHAPTER 7

100

Non-oiled Oiled Oiled + washed

Rockweed cover (%)

80

60

40

20

0 1989

1990

1991

1992

1993 Year

1994

concern. Mortality of the early life stages, and rate of growth of juveniles, could have been impaired in those streams that were oiled (Table 7.1). These deleterious effects were not detectable by 5 years after the spill. Similar effects were reported for two trout species, with recovery by 2–3 years. Abundance of ﬁsh in the intertidal zone was reduced by half, but recovered after 2 years. Paciﬁc cod and Arctic shammy, in contrast, increased in abundance immediately after the Exxon Valdez incident. Birds received much attention, not the least prompted by public reaction to heart-rending images of oiled waterfowl (see Fig. 7.2). It was fortunate that surveys of water birds were being recorded before and after the spill. For example, there was an excellent data set for the number of nests and chicks hatched per nest of the blacklegged kittiwake in the Prince William Sound area from 1984 to 1994 (Fig. 7.5). This marine gull was the most abundant colony nester in the Prince William Sound area (Irons 1996). Ten colonies surveyed were exposed to oil after March 1989, and 14 colonies of kittiwakes were in non-oiled areas. There are no evident effects of the spill (Fig. 7.5 left panels), i.e. the number of nests and chicks per nest in oiled sites do not clearly show a different pattern pre- versus post-1989. Interpretation of the impact of oil is made difﬁcult by the fortuitous circumstance that colonies destined to be oiled also held many fewer nests through-

1995

1996

Figure 7.4 Percentage cover of the brown rockweed Fucus gardneri in sites within Prince William Sound that were non-oiled, oiled, and oiled plus washed with hot water under high pressure. Data from Driskell et al. (2001).

out the sampling period. Calculation of the ratio of non-oiled to oiled data in Fig. 7.5 partially removes this confounding bias; even after this treatment, however, there still seems to be no evident effect of the 1989 oil spill (Fig. 7.5 right panels). Somewhat lowered reproduction might have occurred in oiled colonies, based on other ancillary information (Irons 1996), but the evidence of Fig. 7.5 makes clear that the most numerous colonial species was not devastated by the oiling. The oil spill imposed a heavy initial toll of birds (Table 7.1); more than 30,000 carcasses were recovered and losses of up to 700,000 were estimated from these recoveries. The best estimate is that perhaps 250,000 birds were killed by the oil (Piatt & Ford 1996). There was, of course, great interest in assessment of the impact of the spill on the sea birds of the affected areas, and extensive surveys were undertaken. The results of the longer-term assessments of the effects of the spill on birds have been inconsistent. The numbers and kinds of birds found to be affected by the oil differed from study to study. For example, Table 7.2 lists sea birds that did not recover from the spill, as recorded in several papers. There are many mismatches in the results: only red-necked grebe, Barrow’s goldeneye, merganser, and pigeon guillemot are present in more than one list. The kinds of birds affected in the long term in the Kenai Peninsula also differ from those affected in Prince William Sound (Table 7.2).

Table 7.1 Compilation of time courses (1989–1996) of some of the effects of exposure to the Exxon Valdez oil spill (EVOS) on a variety of selected organisms. Organism

1989

1990

1991

1992

1993

1994

30% gain

No effects

1995

1996

Rockweeds (Stekoll & Deysher 2000) (Fig. 7.4)

40% loss

20% loss

Kelps

No apparent effects on subtidal macrophytes

Seagrasses (Zostera marina)

Slight negative effects

Benthic invertebrates

Bivalves, gastropods, and polychaete worms were significantly more abundant; Oiled + washed amphipods were less abundant, oiled population recovered by 1993–1995 areas recovered (Jewett et al. 1999). Marked increase in abundance of harpacticoid copepods by 1997 in 1990 (Wertheimer et al. 1996) Clam (Protothaca staminea) residual oil Seastar (Desmarestia imbricata): significantly lower lowered growth 5–6 years after spill density in 1989 (Dean et al. 1996) (Fukuyama et al. 2000) Crab (Telmessus cheiragonus): 80% loss No differences in density in 1989 (Dean et al. 1996) Seastar (Pycnopodia heliantoides): higher densities in 1989 (Dean et al. 1996) Mussels (Mytilus troussulis) exposed to oil for 3–4 years were not physiologically affected by internal oil concentrations (Thomas et al. 1999). Oil contents of mussels were below thresholds injurious to birds eating mussels (Boehm et al. 1996)

Pacific herring (Clupea pallasi)

Half the spawning streams were oiled (Murphy et al. 1999). Egg/larval mortality increased by 2, larval growth decreased by half (McGurk & Brown 1996); 25–32% of embryos were damaged by oil (Carls et al. 2002). Liver necrosis, higher PAH concentrations, and slower growth were found (Marty et al. 1999)

Pink and sockeye salmon

Embryonic mortality increased by 2, lower growth rate (Rice et al. 2001), but effect could have been an artifact of measurement (Brannon et al. 2001)

Pink salmon egg mortality in oiled streams continued for at least 4 years (Bue et al. 1998; Rice et al. 2001). Juveniles exposed to oil suffered a 15% greater mortality at sea than unexposed fish (Heintz et al. 2000)

Trout

Lower growth of Dolly Varden by 24–27%; cutthroat 36–43% lower (Hepler et al. 1996)

No differences in growth

Fish of shallow subtidal zones

Pacific cod and Arctic shammies were more abundant, with greater stomach fullness (Laur & Haldorson 1996)

Density lower in 1990, no differences in 1991

Shoot No detectable effects (Dean et al. 1998) density 24% and flowering 64% lower (Dean et al. 1998)

PAH in sediments of streams lower by 1993, spawning habitats recovered to sublethal threshold by 1994 (Murphy et al. 1999). Embryonic and liver effects not detected (Marty et al. 1999). No effect on reproduction. Record harvests of herring (Pearson et al. 1999). In 1993, 75% loss of adult population, but not related to oil (Pearson et al. 1999) No detectable effect on embryo mortality (Johnson et al. 1997; Bue et al. 1998)

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CHAPTER 7

Table 7.1 (cont’d) Organism

1989

Intertidal zone fish (Barber et al. 1995)

Abundance reduced by half

Bald eagles

Initial loss, no post-spill effects on survival

Sea birds (Fig. 7.2)

250,000 lost (Piatt & Ford 1996), many exposed to oil (Trust et al. 2000)

Most species unaffected, some others recovered, a handful of taxa still not recovered after 1998 (Table 7.2). Winter survival for harlequin ducks 78 ± 3.3% in oiled areas and 83.7 ± 2.9% in unoiled areas for 1995–1998 (Esler et al. 2000a, 2000b). Breeding by black oystercatchers slightly affected by oil in 1989, but no measurable effects by 1991 (Andres 1997, 1999) or 1998 (Murphy & Mabee 2000). Goldeneyes not affected by exposure to oil after 1997 (Esler et al. 2000b). Murre colonies not demonstrably affected by oil exposure (Boersma et al. 1995; Piatt & Anderson 1996)

Sea otters

2,800 (1,000– 4,000) lost (Burn 1994; Ballachey et al. 1994)

Post-EVOS multiyear reductions in abundance in oiled shores (Bodkin et al. 2002) not evident in other reports (Burn 1994; Garshelis & Johnson 2001). Decreased survival suspected in most heavily oiled sites, with effect dissipating through 1998 (Monson et al. 2000), and few effects in less-oiled sites (Garshelis & Johnson 2001). Otters still exposed to oil, but reproduction has increased abundances in Prince William Sound as a whole through 2000, though perhaps not in more intensively oiled areas (Dean et al. 2000; Bodkin et al. 2002).

Harbor seals

Claim that 135–302 lost soon after EVOS (Frost et al. 1994) likely to be too large, and “loss” may be by migration out of oiled area (Hoover-Miller et al. 2001)

Killer whales (Matkin et al. 1994)

6 lost

Humpback whales Other cetaceans

1990

1991

1992

1993

1994

1995

1996

No differences Estimated return to initial numbers by 1992 (Bowman et al. 1995)

6 lost

1 lost

Increased by 24 in 1990, increased by 7 in 1991 (Ziegesar et al. 1994) Beached carcasses of 26 gray whales, 2 minke whales, 1 fin whale, 5 harbor porpoises, and 3 unidentified porpoises, soon after EVOS (Harvey & Dahlheim 1994)

PAH, polycyclic aromatic hydrocarbons.

Birds in oiled areas incorporated oil in their bodies, judging from biochemical assays (Esler et al. 2002; Golet et al. 2002).11 The exposure to oil that was made evident by the internal chemistry 11

Simple measurement of concentrations of hydrocarbons and like compounds in tissues does not reliably reﬂect exposure to such compounds, because many aromatics, for instance, are rapidly metabolized or excreted by organisms. One of the more sensitive and speciﬁc biochemical assays of exposure to aromatics is the detection of the enzyme cytochrome P450 (a mixed function oxygenase) (Stegeman 1981; Woodin et al. 1997). Exposure to aromatics induces the synthesis of P450, and hence its concentration quantitatively, but indirectly, indicates exposure.

of the birds was not, however, inevitably followed by demonstrable population effects. The compilation of results in Table 7.1 hints at, but does not truly convey, the remarkably contradictory sets of results that seem to characterize the multiple efforts to assess the consequences of the Exxon Valdez oil spill.12 The patchwork results characteristic of the various studies suggest that 12

An example of the inconsistencies that plague the record is the population status of harlequin ducks. Day et al. (1997a, 1997b) reported that densities were related to the degree of oiling in 1989–1990, but the effect disappeared in 1991. Murphy et al. (1997) found abundance during 1990–1991 unchanged relative to pre-spill

153

PETROLEUM HYDROCARBONS

Non-oiled Oiled

20 Number of nests in non-oiled/oiled sites

Number of nests (× 103)

20 16 12 8 4

12 8 4 0

0 1984

1986

1988

1990

1992

1994

1984

1986

1988

1990

1992

1994

1984

1986

1988

1990 Year

1992

1994

10 Number of chicks per nest in non-oiled/oiled sites

0.5 Number of chicks per nest

16

0.4 0.3 0.2 0.1 0.0 1984

1986

1988 1990 Year

1992

1994

8 6 4 2 0

Figure 7.5 Time course (1984–1994) of number of nests of black-legged kittiwake (top left) and number of chicks per nest (bottom left) in 10 colonies that were exposed to Exxon Valdez oil in 1989, and in 14 colonies that were not oiled. Ratios of nests and chicks/nest in oiled sites were divided by nests and chicks/nest in non-oiled sites (right panels). All colonies were within Prince William Sound. Data from Irons (1996).

impacts on the bird fauna were not so pervasive and extensive across a variety of data treatments and in data collected across different sites. Several studies report information on the numbers of sea bird types that were either not affected, increased, or decreased temporarily or through to the end of the study periods (Table 7.2). On the whole, about half the bird species were harmed by the spill in 1989; by the end of 1991, 10–20% of years. Rosenberg and Patrula (1998) say that densities were lower in oiled areas 1995–1997, and unchanged in unoiled sites. Irons et al. (2000) reported lower densities in oiled areas 1990–1991. Esler et al. (2000b) found that densities were lower in oiled sites 1995–1997. Esler et al. (2000a) used a population model to predict lower densities in oiled but not unoiled sites. Lance et al. (2001) found that summer populations were unchanged, but winter densities were increased in oiled sites 1989–1998. Esler et al. (2002) concluded that the population had not yet fully recovered by 1998. Inconsistencies such as these can be readily compiled for virtually all the investigated aspects of the spill.

the bird species were still impaired; clearly the trend is one of recovery (Fig. 7.6). The results, taken on aggregate, suggest that most sea birds were not or were only slightly affected after the initial unarguably catastrophic impact. A handful of species betrayed protracted signs of exposure to oil, and that exposure had largely sublethal effects, some of which could still be detected at the end of the 1990s. Long-term effects of large oil spills may therefore linger on for many years, in a handful, but variable, list of sea birds, and in some areas within the exposed region, but such effects of oiling are rather patchy. Sea otters were clearly affected by the initial Exxon Valdez oiling (see Table 7.1). Otter densities in oiled sites were 35% lower during 1989, but increased by 14% in non-oiled areas (Burn 1994). Such differences could be the result of local mortality within oiled sites (Monson et al. 2000),

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CHAPTER 7

Table 7.2 Comparison of results from studies that assessed the status of sea birds after the Exxon Valdez oil spill. No. of sea bird taxa

Period of study

Study

That decreased and Not recovered Included affected That by end in study by oil increased of study

Studies in Prince William Sound Day et al. 1989–1991 42 (1995, 1997b)

That decreased and did not recover by end of study

Seabird taxa that had not recovered by the end of study

23

0

13

6

Horned grebe, red-necked grebe, Barrow’s goldeneye, bufflehead, mew gull, northwestern crow

Murphy et al. (1997) 1989–1991

12

7

2

0

3

Red-necked grebe, pelagic cormorant, pigeon guillemot

Irons et al. (2000)a

1989–1991

14

4

1

2

5

Cormorants, Barrow’s goldeneye, mergansers, pigeon guillemot, murres

Lance et al. (2001)a

1989–1998

17

12

0

–

3

Black-legged kittiwake (July only), scoters, mergansers, Barrow’s goldeneye (March only)

Study in Kenai Peninsula Day et al. (1997a) 1989–1991

34

22

0

6

6b

Common mergansers, glaucouswinged gull

a

Sampled in March and in July. Three of these taxa may have begun recovery by 1991, data on other species were too few to assess.

Species negatively affected (%)

b

60

Prince William Sound Kenai Peninsula

50 40 30 20 10 0 1989

1990 Year

1991

Figure 7.6 Percentage of bird species in Prince William Sound and on the Kenai Peninsula coast that were negatively affected by the Exxon Valdez oil spill, 1989–1991. Data from Irons et al. (2000).

or simply by otters moving away from oiled sites. In the longer term, sea otters actually increased in numbers in Prince William Sound, probably owing to improved conditions from large-scale atmospheric events; one effect of the spill might have been to redistribute sea otters within the region. Much effort was devoted to rehabilitate oilexposed mammals and birds: 357 oiled sea otters were collected alive and brought to rehabilitation centers. Of these, 197 were released, and 25 kept in captivity. The cost of each sea otter that was treated, survived, and was released amounted to US$80,000 (Estes 1991). This high cost, plus the high mortality of released sea otters, suggests that these well-meant humanitarian efforts may be disproportionate.

PETROLEUM HYDROCARBONS

In sum, there is no doubt that within a few days of the spill, there was a considerable loss of specimens of a variety of large species, and a massive death of invertebrates across considerable portions of the shore of Prince William Sound and elsewhere in Alaska. The losses due to oil were spatially patchy, never blanketed the entire shoreline, and were mainly limited to shallower waters (Feder & Blanchard 1998). Even in the affected areas, however, the majority of large species were not affected in crucial or measurable ways. In spite of headlines such as “Slow recovery after Exxon Valdez oil spill” (Christen 1999) and “Exxon Valdez spill still fouls beaches” (Clark 2001), most affected populations could be said to have recovered 1–6 years after the spill. Recovery was aided by the incomplete initial destruction of the populations, and by the availability of recruits from unaffected patches, as well as by a certain tolerance by the biota (see Table 7.1). There was no instance of the extinction of any species, even at a local scale. By 8 years after the spill, oil residues in Prince William Sound that penetrated deeply into sediments, or had accumulated under mussel beds (Babcock et al. 1997; Carls et al. 2001), were only moderately weathered. Oil deposited in more exposed sediments or rocks was extremely weathered (Michel & Hayes 1999), and although oil was certainly still present years later,13 concentrations decreased year by year (Jewett et al. 1999). Presumably, the lower concentrations, and the weathering of residues would be less threatening to organisms as years passed, allowing recovery of the biota to take place, with some lingering sublethal effects on a few species of large, longerlived animals of Prince William Sound. Recent reviews of the longer-term effects cite the continued presence of oil along the shorelines, elevated mortality of pink salmon eggs 4 years after the spill, slow recovery of otter populations, declines in harlequin duck numbers in 13

Short et al. (2004) found Exxon Valdez oil in 78 out of 91 beaches sampled 12 years after the spill. The volume of oil on the shores was less than 1% of the oil that was initially beached, but still enough to expose organisms to contamination.

155

oiled shores in the late 1990s, and many indirect “cascading” effects of the changes in abundance of key taxa (Peterson et al. 2003). Thus, for some taxa, and for different periods of time, the Exxon Valdez spill has had lingering effects. It seems evident that in spite of massive and costly effort, it has proved quite difﬁcult to assess the extent and duration of effects. Certain populations were directly and strongly affected, but reports did not always coincide, and there was controversy as to the interpretation of results (Wiens & Parker 1995; Wiens 1996; Paine et al. 1996; Wiens et al. 1996; Peterson 2001; Peterson et al. 2001). It could well be that the effects of the oil were simply too hard to detect in the diverse kinds of environments and with many types of organisms involved, with the sampling approaches used, in the face of strong interannual14 and seasonal changes affecting the area, and complicated spatial patchiness. This dilemma suggests that new thinking be brought to bear on future work on such accidents (Paine et al. 1996; Wiens 1996; Peterson et al. 2001). In part, the inconsistencies lie with problems inherent in post-event research. In this case, there is the additional aspect that multiple agencies, as well as the responsible entity, Exxon Corporation, were involved in the survey work. Although a fair effort was committed to ﬁnd out the truth, there was too little cooperation, and even competition, among the various participants.15 Interagency pressures, the litigation that persisted throughout the post-spill period, and the laws requiring clean-up and restoration instead of deﬁnition of

14

Climatic shifts taking place across the Paciﬁc may be behind the multiple long-term changes in the bird faunas of the Gulf of Alaska, the Bering Sea, and the coast of California. These large-scale changes have prompted marked multiyear changes in biota, large enough to confound teasing out the local changes that might have been associated with events such as the Exxon Valdez oil spill (Agler et al. 1999). 15 This becomes obvious in published compilations of the major research efforts, for example, those edited by Wells et al. (1995) and Rice et al. (1996). These books cover a similar range of topics, yet, surprisingly, the lists of authors of the articles within the books are totally different. A third compilation (Loughlin 1994) shares one author’s name with the Rice et al. (1996) volume. These researchers largely worked separately, maybe even in seeming competition with one another.

156

CHAPTER 7

effects, conspired to thwart the investment of the enormous resources16 that could have added much new knowledge and useful guidelines for spill management (Paine et al. 1996). The Exxon Valdez spill also took a toll from the humans living near the affected areas, as well as the larger world beyond. About 15 Aleut communities exploited the resources of the Prince William Sound area on a subsistence basis. Subsistence harvests in villages near oiled shores declined by 31 to 77% compared to pre-spill averages, largely because of the villagers concern about safety of the food items collected from the sea (Fall & Field 1996). Villages away from the spill showed no such declines. Aleuts in the villages that were most exposed to the oil spill also revealed a considerably greater frequency of emotional disorders and depression, and more drinking, drug use, and social conﬂicts than villagers in less exposed villages. For example, in the year after the spill, 48% of 593 Aleuts reported increased conﬂicts with outsiders, compared to only 3% in non-exposed villages (Russell et al. 1996). The Exxon Valdez oil spill was featured in the news for weeks after the accident. The attention prompted many individuals and institutions, even those far from Alaska, to action. For example, as a result of the spill, the US Congress passed the Oil Pollution Act of 1990. The public became so sensitized to the issue of oil pollution that pressure of public opinion derailed efforts by the administration of then-President George Bush to tap oil resources below the Alaska National Wildlife Refuge. Litigation brought about by many groups forced Exxon Corporation, an archetype international conglomerate, to fund a huge research effort. None of these issues are closed, even as this is written. The continuing pressures and conﬂicts are evident in the adversarial tone of the local press and internet coverage still given to the Exxon Valdez spill. Disparate opinions and results led to a circumstance where science as a whole lost credibility in the eyes of the public 16

The punitive ﬁne (above the US$3 billion cost of the clean-up) levied on Exxon Corporation amounted to 5 billion dollars, an award that was appealed. Much more was also required to bring the suit, so the costs are far larger than the ﬁne.

and the courts.17 Thus, an ecological disaster had surprising and pervasive consequences for the subsistence inhabitants of the south coast of Alaska as well as for the people in our industrial society.

Sources of petroleum to coastal waters Our industrialized society in the 20th and 21st centuries has been and is powered by the combustion of fossil fuels. Oil continues to contribute the largest portion of world energy demand, and forecasts predict the trend to continue (Fig. 7.7), in particular for transportation.18 In addition, petroleum products are heavily used in the synthesis of plastics and in construction. Use of natural gas as an energy source is foreseen to increase, but this energy source is closely tied to petroleum production. Coal remains important, but produces noxious contamination of the atmosphere. Nuclear sources are viewed skeptically through much of the world, hence the nearly ﬂat trend after the 1990s (Fig. 7.7). Renewable sources are at present primarily hydroelectric, a source that is only available in those parts of the world with sufﬁcient water supplies. Use of energy from other renewable sources—wind and solar—increased as the 20th century ended, but need far more institutional and political support, as well as technical development to become economically more attractive. Given present public and political attitudes, we are, for the foreseeable decades, inextricably bound to petroleum. We have already seen (Chapters 2 and 3) some of the substantial global effects of the release of 17

One of the reasons for the confusion and differences in views was that legal constraints prevented scientists working for government agencies from cooperating or coordinating work with scientists working for the petroleum interests ( J. Farrington, personal communication). In addition, the opposite or competing views presented by scientists on either side of the case created skepticism about the credibility of scientists and their data (Barker 1994). 18 As we start the 21st century, the industrialized world consumes about 40% of the energy made available by all sources and the developing nations use roughly 25% (NRC 2003). World energy consumption is predicted to increase by about one-third by 2020; the increases are expected to take place in the developing nations, however, and the environmental consequences will therefore fall disproportionately on parts of the world perhaps less prepared to take remedial action or manage the environmental challenges.

157

Figure 7.7 Estimated and projected world energy consumption [British thermal units (Btu) × 1013) from oil, coal, natural gas, renewable energy (solar, wind, hydroelectric), and nuclear sources, 1970–2020. Adapted from US Department of Energy (http://www/eia.doe.gov/ oiaf/ieo))/images/ﬁgure_15.jpg).

Energy consumption (Btu × 1013 yr−1)

PETROLEUM HYDROCARBONS

Oil 200 Coal Gas 100

Renewable Nuclear 1970

carbon dioxide from the combustion of fossil fuels into the atmosphere in warming the globe, and changing the chemistry of the atmosphere and oceans—alterations fraught with consequences for coastal environments. Here we deal with another major consequence of use of fossil fuel: the release of petroleum-generated compounds into coastal waters. Exploitable petroleum deposits are limited to a certain few areas of the world, so that wide-ranging transport of petroleum and its products has been necessary. Transport of petroleum, as well as routine use in a variety of ways, inevitably result in the release of petroleum compounds into coastal environments. Petroleum enters the world’s oceans via various mechanisms (Table 7.3). Some of these mechanisms are natural; most are connected with human activities. In all cases, it has been difﬁcult to obtain precise estimates of the various inputs (note the broad ranges of Table 7.3) because of the innumerable speciﬁc sources and diverse distribution of discharges around the world. The most recent re-evaluation of petroleum sources to the sea (NRC 2003) concludes that oil seeping naturally from oil-containing sediments accounts for the largest single source to the world’s oceans (Table 7.3). These seeps ﬂow chronically at relatively slow rates, but appear to have modest, if any, effects on organisms. Long-term chronic exposure may have resulted in the development of assemblages of organisms that tolerate exposure to the oil. Most oil seeps are local, limited to

1980

1990

2000

2010

2020

Year

sites such as the Gulf of Mexico, southern California, offshore Alaska, and a few other localities (NRC 2003). The sources of petroleum associated with human activities add oil to previously unexposed sites, and have far more damaging effects than natural seeps. The total amount of oil added as a result of human activities is somewhat larger than that of natural seepage (Table 7.3). Procedures involved in the extraction of oil release a modest amount of petroleum. Transport is responsible for larger amounts. Tanker accidents, certainly the most widely publicized instance of marine pollution, contribute substantially to this category. The various activities involved in the consumption of oil, apparently far less newsworthy sources of contamination, add more oil to the sea than the widely reported tanker spills (Table 7.3).19 The largest item in this category is the dis19

Certain politicians, and the petroleum industry, have made use of arguments about the natural production of hydrocarbons by plants and algae to minimize our responsibility for the release of carbon compounds. Indeed, marine algae produce perhaps 26,000 million tons of hydrocarbons per year, but the types of naturally synthesized organic materials differ substantially from those made by industrial processes and by the alteration and combustion of petroleum products. Natural hydrocarbons are less diverse than petroleum hydrocarbons; among aliphatic hydrocarbons, compounds with an odd number of carbons predominate; alkenes may be abundant; and aromatics are scarce. Natural hydrocarbons are in general not as toxic as industrial hydrocarbons. The naturally made hydrocarbons are, once released in coastal environments, also not as long-lived as many industrial hydrocarbons, except where anoxic conditions exist.

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Table 7.3 Estimates (best and ranges) of annual releases of petroleum to the world’s oceans via various inputs, 1990–1999. Data adapted from NRC (2003). Amount of released oil (tons × 103 yr−1) Best estimate

Range of estimates

% of total, based on best estimate

Natural sources Seeps

600

200 –2,000

46

Human sources Extraction Platforms Atmospheric deposition Effluent waters Transport Pipeline spills Tanker spills Rinsing cargo holds Coastal facility spills Uses Rivers and runoff Spills during operation Atmospheric deposition Jettisoned aircraft fuel

38 0.9 1.3 36 150 12 100 36 4.9 480 140 97.1 52 7.5

20 –62 0.3 –1.4 0.4 –2.6 19 –58 120 –260 6.1 –37 93 –130 18 –72 2.4 –15 130 –6,000 6.8 –5,000 96.5 –818.8 23 –200 5 –22

Oil inputs to sea

Total

1,300

charge by streams, rivers, and sewer outfalls. The oil borne by these ﬂows is the aggregate contribution from many small spills on land and in the streams, added as a result of miscellaneous human activities on densely populated areas. Our use of fossil fuels has of course increased markedly after the Industrial Revolution. It is therefore not unexpected to ﬁnd that as urban/ industrial uses of coastal areas develop, disposal of waste water into nearby receiving waters—plus leaks from transportation-related activities—leave a permanent record in coastal sediments, as can be seen in Halifax Harbor (Fig. 7.8). The salient feature of the sedimentary record in Halifax Harbor is that hydrocarbons derived from both petroleum spills (Fig. 7.8 left) and from human combustion of hydrocarbons (Fig. 7.8 right) increased markedly in layers of sediment that accumulated through the 20th century. These compounds no doubt were brought to the sediments via the various sources included in Table 7.3. Note that, in contrast to the human-related

6.1

12

37

470 –8,300

hydrocarbons, hydrocarbons produced by algae and plants do not increase. Of course, the global estimates of Table 7.3 might not apply to speciﬁc portions of the coastline, where the balance of inputs from the different mechanisms will surely vary greatly. Oil tanker accidents, for example, as already noted, are not the major single source of oil to the sea.20 Nonetheless, in places where tanker accidents have occurred, there is no doubt of the impact of such accidents. The consequent public and scientiﬁc attention to these catastrophic spills provide dramatic examples, and a rich body of information with which to introduce the topic of petroleum in the sea, as we saw in the case history of the Exxon Valdez spill.

20

The largest oil spills have not been the result of oil tanker accidents. A list of spill events ranked by volume spilled would ﬁnd the 1991 Gulf War, the Ixtoc-1 blow-out, and the Norwuz Field spill itemized before the largest tanker spills (the Castillo de Belver ﬁre and the Amoco Cadiz grounding).

159

PETROLEUM HYDROCARBONS

Figure 7.8 Vertical proﬁles of some hydrocarbons in sediments of Halifax Harbor, Nova Scotia, Canada. Left: aliphatic hydrocarbon residues derived from fossil fuels and aliphatic compounds from natural sources (largely phytoplankton production). Right: aromatic residues derived from the combustion of wood and fossil fuels, probably transported by atmospheric deposition and runoff from land. Data from Gearing et al. (1991).

Depth in sediment (cm)

Aliphatic hydrocarbons (ppm) 0.1 10 1,000 100,000 0 10

10

20

20

30

30

40

40

50

50

60

A primer on oil21 Composition of petroleum

Petroleum is a complex mixture of thousands of organic compounds that have accumulated across geological time scales as a result of incomplete decomposition of plant and algal tissues. The material extracted from the geological strata is called crude oil, which may contain thousands of speciﬁc compounds, ranging from volatile gases with low molecular weight to heavy residues that only boil at high temperatures. Crude oils may range from 50 to 98% hydrocarbons by weight. Composition of crude oils varies characteristically depending on the place of origin, a convenient feature that allows oils to be identiﬁed as to source, and allows detailed identiﬁcation of the geographic source of the oil in a spill. Petroleum consists primarily of thousands of different hydrocarbons, compounds containing carbon and hydrogen. In addition, petroleum contains compounds that incorporate small amounts of nitrogen (N), sulfur (S), oxygen (O), and metals, particularly nickel and vanadium. Petroleum hydrocarbons include various classes 21

Aromatic hydrocarbons (ppm) 100 0.1 1 10 0

An impressively comprehensive review of this ﬁeld is provided by Hunt (1996).

Natural sources Fossil fuel residues

Combustion products 60

of organic compounds; some knowledge of their identity is needed here because the different classes have quite different fates and effects once they are released into the marine environment. Aliphatic22 hydrocarbons have carbons connected in straight-chain, branched, or ring structures. Aliphatic compounds in which carbon–carbon bonds are single (“saturated”) are known as alkanes (or parafﬁns in the petroleum industry). Aliphatic compounds with carbon–carbon double bonds (unsaturated) are called alkenes (or oleﬁns). Alkenes do not occur naturally in crude oils, but are the product of industrial “cracking” of petroleum. The double bonds make alkenes somewhat more reactive than alkanes, and, hence, they do not last as long in the environment after spills. Cyclic aliphatic compounds (called naphthenes in the oil industry) have carbons linked in a ring structure and occur in substantial quantities in crude oil. Aromatic hydrocarbons are characterized by the presence of six-carbon rings among which 22

Names used by chemists and the petroleum industry differ, and are of some interest. “Aliphatics” derive from a Greek word for “fatty”. “Parafﬁn” derives from the Latin, “parum afﬁnis”, meaning of little afﬁnity, a reminder of the insolubility in water, no doubt. “Oleﬁns” were so named because certain hydrocarbon gases reacted to yield oily products in early petroleum work. “Aromatic” hydrocarbons were so named by the oil researchers because they often have pleasant odors.

160

CHAPTER 7

electrons are shared. This structure favors reactions where some other element of the compound is substituted for hydrogens attached to the carbon ring, a feature that will become important when we discuss chlorinated hydrocarbons. Aromatics can be single-ring (benzenes) or two-ring (naphthalenes), or their derivatives with side chains (alkyl benzenes or alkyl naphthalenes); polycyclic aromatic hydrocarbons (PAHs) contain three or more fused rings, and make up perhaps 0.2–7% of crude oils. PAHs often include compounds that include O, S, and N. PAHs bearing many substituted groups are major fractions of the higher molecular weight petroleum constituents.23 Such complex compounds may constitute almost 50% of a crude oil, and often are the most problematic in terms of toxicity and longevity in the environment. Aromatic compounds with larger number of rings tend to be less subject than aliphatics to degradation once in the marine environment. PAHs with more than ﬁve rings degrade most slowly. Toxicity of oil is to a large extent associated with the one- and two-ring compounds, and the more substitutions in each, the lower the toxicity. Aromatic compounds are more soluble in water than aliphatics, and therefore more likely to travel and to come in contact with organisms. PAHs might include N, S, or O in their structure; such compounds are formed by molecular transformations of residues under high temperatures and pressures deep below the earth’s surface. The transformations lead to complex structures, including naphtheno-aromatics, organo-sulfur compounds, acids, phenols, pyridines, pyrroles, and highly complex asphaltenes. Each of these has different properties, which lead to different fates and effects once in the coastal environment. In particular, substituted aromatic compounds

23

Conditions that create crude oil create many aromatic hydrocarbons in which hydrocarbon chains have been substituted. When petroleum products are combusted as in vehicular or industrial uses, the relatively high temperatures foster the formation of unsubstituted polycyclic aromatic hydrocarbons (PAHs). These arrays of PAHs confer the ability to differentiate, to an extent, residues delivered from combustion processes (usually delivered via atmospheric deposition) from residues delivered by oil spills on water.

resist degradation, and hence serve as longterm markers for the presence of oil. Processing of petroleum fractions

Once petroleum reaches a reﬁnery, various fractions are separated by distillation processes. First, hydrocarbons with 1–4 carbons (methane, ethane, propane, and butane) are distilled from petroleum heated to less than 20°C at atmospheric pressure. These gases are separated for use as natural gas. Further distillations are done at higher temperatures and in a vacuum, so as to lower boiling points and reduce thermally driven cleavage of molecules (pyrolysis or “cracking”). Distillation of petroleum at 20– 200°C yields alkanes and cycloalkanes with 6–10 carbons. Such compounds are used as petroleum ether and gasolines. At 185–345°C, alkanes and aromatics with more than 12 carbons are distilled. These compounds are used as kerosene, heating oils, and aviation fuels. Distillation at about 345–540°C separates compounds with large numbers of carbon atoms, useful as lubricating oils, waxes, and compounds to be subsequently “cracked” into smaller compounds to yield more gasoline. These materials consist mostly of long chains attached to cyclic structures. Compounds that survive these distillation treatments have mostly polycyclic structures and are used for construction and industrial purposes. Different crude oils yield different ratios of the distillation products. Oil from the Prudhoe Bay ﬁelds off Alaska, for example, produce 3% natural gas, 18% gasoline, 2% kerosene, 25% heating oils and aviation fuels, 35% waxes and lubricating oils, and 18% residual oils. The crude oils or distilled fractions may of course be shipped separately: when accidents occur, a major concern is to identify the source as well as the type of material, to establish the responsible party, and to anticipate possible effects of the spill. Effects of different fractions will be markedly different: a gasoline spill will be highly toxic immediately, but will evaporate rapidly. Spills of other forms of petroleum will have longer-term effects, as in the Exxon Valdez case. Processes affecting the different fractions, once the petroleum is in the marine environment, are summarized in Table 7.5.

161

PETROLEUM HYDROCARBONS

Table 7.4 Estimates of the fate of oil released from spills (ﬁrst four columns of data), and experimental additions of oil (last column of data). Percentage of spilled (or added) oil volume (%) Ixtoc-1 a

Amoco Cadizb

50 25 12

30 8 4.5

7 6 – –

28 – 13 20.5

Evaporated To sediments Decomposed by microbes and photochemical oxidation Stranded on shore Recovered or burned Dissolved in water columnf Unaccounted for

Exxon Valdezc 20 (13)

50 (2) 14

Gulf Ward

MERLe

40

40–50 10–20 20–30

15 10 10 25

5

aSoon after the spill. Data from Atwood (1980). bOne month after the spill (Gundlach et al. 1983). c

Soon after the spill, and, in parentheses, 2 years after the spill (Wolfe et al. 1994).

dSum of dissolved and emulsiﬁed oil. e

This was an experimental addition of No. 2 fuel oil to large tanks containing sediments and water in the Marine Ecosystems Research Laboratory, University of Rhode Island (Elmgren & Frithsen 1982). f Data from Tawﬁq and Olsen (1993).

Fate of petroleum in coastal environments Once a petroleum spill takes place, the fate of the released material seems likely to depend on the mixture of compounds involved, local conditions, and whether the oil spill is chronic or a single spill. In spite of such likely variation, there are a few regularities (Table 7.4) (Whittle et al. 1982; Gundlach et al. 1983; NRC 2003). There are a series of weathering processes that alter the chemical composition of crude oil remaining in the environment (Table 7.5). Evaporation removes the more volatile compounds (lighter aliphatics and one-ring aromatics) from the water.24 Insoluble compounds such as the alkanes form tiny droplets that are suspended in a seawater matrix to create emulsions, color-

fully referred to as “chocolate mousse” because of their frothy brown appearance. A small fraction of crude oil—usually the more soluble aromatics and substituted complex compounds—may become truly dissolved in sea water. Microbial oxidation is often a large sink for spilled oil; the microbes tend to target aliphatics that contain fewer than 14 carbons, principally straight-chain alkanes and to a lesser extent branched alkanes and naphthalenes (Gundlach et al. 1983). Microbial activity is often limited by the supply of nitrogen and phosphorus, hence the attempts to fertilize environments exposed to oil spills to stimulate oxidation of spilled oil. Photo-oxidation lowers the concentrations of hydrocarbons that remain near the surface, but this fate is not likely to be an important hydrocarbon loss in most cases.25 Microbial degradation and photo-oxidation are

24

25

For example if a spill were only of gasoline, all the material would soon evaporate. In a spill of No. 2 fuel oil, perhaps 75% might evaporate, but if heavier bunker oils were released, only 10% might evaporate. In spills of crude oil, 30–50% typically evaporate, which might be more than reported for the Exxon Valdez spill.

In the case of the Amoco Cadiz, the high energy of the sea conspired to increase the sediment load and turbidity of the water, lowering photo-oxidation; in contrast, in the case of the Ixtoc-1 spill, the clarity and calm waters made it more likely that photooxidative losses would be more signiﬁcant (Gundlach et al. 1983).

162

CHAPTER 7

Table 7.5 Relative effects of processes that weather or move petroleum released from seeps or spills once the petroleum compounds are in the marine environment. ***, High; **, moderate; *, low; –, not relevant. Adapted from NRC (2003). Movement processes Weathering processes Type of oil inputa

Duration of persistence

Evaporation

Emulsification

Dissolution

Oxidation

Transportb

Sedimentationc

Stranding, tarballsd

Natural seepage

Years

***

**

**

**

***

**

***

Spills Gasoline Light distillates Crude oils Heavy distillates

Days Days Months Years

*** ** ** *

– * ** **

** *** ** *

* * ** *

* * ** ***

* – ** ***

– – *** ***

a

Note that petroleum spilled from natural seeps is a complex mixture, and parts of the petroleum are subject to all the processes; instead, the distilled fractions tend to be subject to different sets of processes. For example, the lighter fractions tend to be subject to processes toward the left of the table, and the heavier fractions are more subject to processes on the right side of the table. b Transport by hydrodynamic forces can be horizontal, moving petroleum away from the source point by advection or dispersion, or vertical, such as might occur in langmuir circulation, vertical dispersion, or sinking with water masses. c This refers to adsorption to particles and subsequent sinking. d This refers to residues that largely remain on the sea surface, and may accumulate on the shores, or aggregate into clumps.

diminished by burial in anoxic coastal sediments; hydrocarbons buried in soft sediments may survive for quite long periods.26 Certain physical processes move oil around (Table 7.5). Spilled hydrocarbons spread out in a slick and, of course, horizontal and vertical mixing and mass transport by wind, waves, and currents disperse and redistribute oil; we saw in the case of the Exxon Valdez just how far oil could be carried away from a spill (see Fig. 7.1). Some of the oil, particularly aliphatic compounds and large aromatics, adsorbs to particles and sinks to the sea bottom or becomes stranded on shore. The general qualitative summary of Table 7.5 is based on the reviews of many studies, but few studies have assessed the relative role of the various processes that affect the fate and concentrations of spilled oil (see Table 7.4). From the few estimates available, it is apparent that a quite

substantial portion of oil evaporates, and some lesser amount reaches and remains for some time in the sediments below. A modest portion of spilled oil is degraded by microbes, largely in the sediments. The amount of oil recovered, despite strenuous and costly efforts by crews and equipment, tends to be small compared to the total spilled. Of course, there is the additional complexity that the different processes do not act completely synchronously.27 For the case of the Amoco Cadiz spill, for example, during the ﬁrst month, the fate of the crude oil was as depicted in Table 7.4. Later on, there was much redistribution of the oil stranded in the shore and in sediments, some entering the water column. Throughout the following months, the lighter fractions were degraded by photo- and microbial oxidation, and there was evaporation of the volatile smaller residues. The

26

27

Reddy et al. (2002) reported that petroleum concentrations remained at 10 cm below the surface of anoxic salt marsh sediments, 30 years after the spill of the barge Florida. Such protection from oxidation is a reason why oil spill residues may have considerably longer histories in environments with soft unconsolidated sediments.

In addition, there may be differences in the impact of petroleum that is delivered by sources that are chronic and of low input, versus one-time and high-volume inputs. There is too little known about such effects, even though Elmgren and Frithsen (1982) have suggested that low-level chronic oil inputs may suffer similar fates as the oil from larger one-time spills.

PETROLEUM HYDROCARBONS

result of this redistribution and degradation was that concentrations of oil were lowered to near pre-spill levels by 3 years after the accident, except in speciﬁc parts of the shoreline where there were soft unconsolidated sediments and weak wave action. In such protected environments, concentrations of oil were lowered to oneﬁfth or one-tenth of post-spill concentrations by 4 years after the spill (Gundlach et al. 1983). As in the case of the Exxon Valdez, the long-term legacy of oil spills is spatially patchy, with some oiled patches still oiled, and others recovered. In general, most of an oiled shoreline recovers within a few years, except where there are soft sediments, where oil may persist for a long time, with unknown effects. As oil becomes less pervasive, and concentrations become lower, the biological effects also diminish.

Effects of petroleum in coastal environments Extent of effects Studies of the aftermath of many oil spills28 do not reveal a very consistent set of effects, much as we saw in the case history of the Exxon Valdez spill. Some spills hardly affected receiving ecological systems, and some showed marked alterations. Here I review a series of spills that resulted in a range from minor to major effects. The examples that follow range from open-water, highenergy waters to enclosed, calmer estuaries: to anticipate the result, it may be that the relative impact of oil might be felt most powerfully in the latter, particularly where oil can reside in soft unconsolidated sediments for long periods of time. 28

The consequences of chronic low-level releases of petroleum hydrocarbons has been even more difﬁcult to assess. There is a large literature on the concentrations and molecular responses to this oil, but fewer studies compellingly link these to population effects. Studies of oil impacts in a Norwegian fjord and the Baltic revealed only minor or equivocal effects on organisms (Nelson 1982; Gray 1987). Experimental oil additions too had no or weak effects on plankton (Oviatt et al. 1982; Vargo et al. 1982; Stacey & Marcotte 1987).

163

In cases where the oil did not reach shore, as in the accident of the Argo Merchant,29 prevailing westerly winds moved the oil slick away from shore, and the impact of oil on the populations of coastal organisms was minimal (Center for Ocean Management Studies 1978). There were some slight sublethal effects on the tissues, biochemistry, and physiology of marine invertebrates collected in sediments below the area of the spill, but these effects were transient, with recovery a few weeks later. Within the area of the oil slick, 18% of the embryos of pollock, but none of cod, were malformed. Even though ﬁsh certainly came in contact with the oil, there was no evidence of large-scale mortality of juvenile or adult ﬁsh in the area for the 12 months following the spill. Fishermen reported that catches and earnings for the 1977 season were as good or better than during the 1976 ﬁshing season. In the case of the Ixtoc-1 blow-out (Fig. 7.9),30 the second-largest release of oil recorded, most of the oil also did not reach shore. Reductions in zooplankton abundance were hard to attribute to oil (Guzman del Proo et al. 1986). There was a minor impact on marshlands, and the oil was not toxic to subtidal amphipods and zooplankton (Getter et al. 1981). There was no discernible change in the prawn landings in the Gulf of Mexico or Cuba following the blow-out (Matthews et al. 1993). Oysters contained oil but were not otherwise affected by the spill (Castro Gessner 1981). Tar ball residues did manage to reach Texas beaches, but the effects were minor. Certain beach ﬂeas were displaced from oiled areas (Rabalais & Flint 1983), and a report records that the number of visits by tourists to the oiled beaches after the 29

The oil tanker Argo Merchant ran aground southeast of Nantucket Island, off the coast of Massachusetts, on December 15, 1976 (Center for Ocean Management Studies 1978). The ship subsequently broke up, releasing its cargo of 7,700,000 gallons of No. 6 fuel oil by December 31, 1976. 30 On June 2, 1979, Ixtoc-1, a 3.2 km deep exploratory well, blew out in Bahia de Campeche in the Gulf of Mexico. A problem with the ﬂow of drilling mud caused the explosion. Oil and gas out of the well ignited (Fig. 7.9). About 10,000–30,000 barrels of oil per day spilled out from Ixtoc-1, until it was capped on March 3, 1980. Currents carried the spilled oil towards the Texas coast, about 1,000 km away, in a patch with visible sheen roughly 100 × 100 km. Most of the Texas coast was oiled to some degree.

164

CHAPTER 7

Figure 7.9 Aerial view of Ixtoc-1 during the blow-out, 1979–1980. From http://oils.gpa.unep.org/ framework/region-8.htm.

spill remained about the same (Freeman et al. 1985), except for a 60% reduction during the event. Another example of relatively modest effects is that of the largest oil spill to marine environments, the release of 1,770,000 tons31 from war-damaged oil terminals in Kuwait during the 1991 Persian Gulf War (Tawﬁq & Olsen 1993). An additional 500 million barrels were emitted or ignited from oil wells during 1991 with oil aerosols, soot, and combustion products that could have reached the Persian Gulf via atmospheric deposition (Readman et al. 1996). Severe oil pollution from the war damage was restricted to the Saudi Arabian coast, within 400 km of the origin of the spills; beached oil was common within that area in 1991. By 1992, oil contamination in the Arabian coast was markedly lower (Price et al. 1993; Readman et al. 1996): beached oil was covered by sand, and weathering, dissolution, and degradation selectively depleted alkanes and aromatics within the sediments, so that by 1992 only the most resistant compounds remained. The residues derived from the oil spills, rather than from the air-transported combustion products. Petroleum residues (polycyclic aromatics) and their metabolites were present in sediments, in clams (Readman et al. 1996), and in ﬁsh tissues (Krahn et al. 1993). A comparison of concentrations of 31

This amount was three times as large as the next largest spill, the 1979 Ixtoc-1 well blow-out in the Gulf of Mexico (NRC 2003).

petroleum compounds in clams showed no postwar increase in alkanes, but notable increases in aromatic compounds, which, however, returned to pre-war levels within 2 years (Vazquez et al. 2000). No evidence of oil pollution damage was detected in coral reefs (Downing & Roberts 1993), seagrass meadows (Kenworthy et al. 1993), and zooplankton abundance and distribution (AlYamani et al. 1993), or changes were inconclusive (Price et al. 1993). Salt marshes were impacted, and showed little recovery as of 1997. There were immediate kills of birds and other organisms (Burger 1997). At least 30,000 sea birds were killed, shrimp stocks were reduced to 27% of pre-war levels, and the recruitment of ﬁsh was reduced (Evans et al. 1993; Matthews et al. 1993), but follow-up reports of the biological effects of this spill largely found no major effects. Breeding terns were not affected, but the abundance of four species of sea birds (out of 25 species in the area) was lowered by 22 to ≥ 50% (Evans et al. 1993). These initial casualties were not, however, large enough to affect counts of most bird species after the initial period. In spite of the large volume of oil, the data reported suggest that there were measurable immediate effects, but relatively modest longterm biological effects. The Persian Gulf is a coastal environment that has been exposed to intermittent small oil spills for many decades as

165

PETROLEUM HYDROCARBONS

32

Low-molecular-weight parafﬁns have narcotic effects and could have caused the immediate disorientation of crabs seen in the ﬁeld soon after the spill (Krebs & Burns 1977).

160 Petroleum hydrocarbon Biogenic hydrocarbon

120 Crabs per m2

a result of the development of the Middle East oil trafﬁc; perhaps this, as in the case of natural seeps, “prepared” the organisms for the major spill of 1991. Conclusions about the results of examples such as the Argo Merchant, Ixtoc-1, and Persian Gulf oil contamination need to be tempered with the certainty that to some degree inadequacy of sampling in highly heterogeneous and variable environments might prevent ﬁnding signiﬁcant effects. On the other hand, if the longer-term effects of oil were large, even cursory sampling might be enough to reveal the magnitude of the impact. In contrast to the above examples, there is a case where sampling was intensive, appropriate controls were taken, and pervasive short- and long-term effects were convincingly documented. This work was done in the aftermath of the grounding of the barge Florida in Buzzards Bay, Massachusetts (Krebs & Burns 1977; Burns & Teal 1979; Sanders et al. 1980; Teal et al. 1992; Reddy et al. 2002). On September 16, 1969, the Florida hit a shoal off West Falmouth, eventually releasing about 650,000–700,000 l of No. 2 fuel oil. The wind churned the oil into an emulsion, and drove the slick into the Wild Harbor River, with dramatic consequences. Mass mortalities of the larger organisms took place almost immediately. Intertidal marsh grasses reached by the emulsion died. Windrows of ﬁsh, crabs, snails, clams, and many other animals appeared at high tide marks. In sediments saturated by oil, almost all benthic species died. Fiddler crabs that survived acted disoriented, took on out-of-season breeding colors, molted unexpectedly, and otherwise showed physiological and behavioral malfunctions,32 which resulted in increased exposure to predators and winter mortality. Studies done over several years showed increased concentrations of oil in tissues of ﬁddler crabs and other organisms from oiled sites. The abundance level of many benthic species for a few years following the spill was clearly related to the concentration of petroleum hydrocarbons

80

40

0 0

1,000 2,000 Hydrocarbon (ppm)

3,000

Figure 7.10 Density of adult ﬁddler crabs in salt marsh sites with different concentrations of petroleum hydrocarbons from the barge Florida spill. Naturally produced (biogenic) hydrocarbons were in low concentrations. Data from Krebs and Burns (1977).

in the sediment: increased concentrations of petroleum led to lower abundances (Fig. 7.10). Sediments left bare by the oiling were quickly and heavily colonized by worms characteristic of polluted muds, as well as by other species. The recolonization took place even though there were high concentrations of two- and three-ring aromatics in the sediments 6 years after the spill. By that time, oil residues in the animals had disappeared, and salt marsh grasses had returned to the oiled areas. None of these effects were detected in a nearby non-oiled estuary. By 20–30 years after the spill, there was no visual evidence of damage, even though there was still oil present at more than 10 cm below the sediment surface.33 These speciﬁc results are corroborated by results from monitoring studies on recovery from oil spills on a variety of salt marsh sites distributed throughout the world. Depending on

33

Oil can be degraded in salt marsh sediments, particularly where the grass roots are actively supplied with oxygen, presumably fostering aerobic decomposition of the petroleum compounds (Reddy et al. 2002). Such ability to oxidize petroleum compounds has encouraged the proposal of wetlands as treatment areas for petroleum industry efﬂuents (Knight et al. 1999). In the case of the Wild Harbor study, the oil-free upper 10–14 cm of the marsh sediment proﬁle was the result of accretion of marsh peat following sea level rise during the 30-plus years after the spill, rather than evidence of oil degradation.

CHAPTER 7

the criteria used for assessment, recovery in oiled salt marshes was considered to have taken place between 8 months to more than 20 years (Hoff 1996). Studies of accidental spills, such as just reviewed, are difﬁcult to compare and synthesize because it is hard to assess the relative concentrations of exposure and the types of compounds or oil fractions. To address such issues, researchers have opted to do experimental additions of speciﬁc concentrations of different oil types to enclosed units containing water and sediments. The effects of such experimental additions in a Norwegian fjord were not dramatic (Nelson 1982; Gray 1987). Seaweeds showed few changes owing to exposure to the oil, and there were inconsistent effects on density and growth of snails. There was high mortality of mussels, but less impact on other benthic invertebrates. Similarly, there were mixed results on plankton and on benthos resulting from the addition of low dosages of No. 2 fuel oil in a study in Rhode Island (Oviatt et al. 1982; Vargo et al. 1982; Stacey & Marcotte 1987). Overall, we might safely conclude that, as in the case of the Exxon Valdez, oil spills have inconsistent impacts on the receiving environments, ranging from modest to dramatic. Probably the most important factor is oil dosage, and this varies enormously from spill to spill, and from place to place within any affected area. The initial effects (in sites where the spills do have a measurable effect) are almost always far more impressive and widespread than the longer-term effects. Recovery from oiling occurs after some years, and lingering longer-term effects, usually on a handful of species, are often at a sublethal level. Spills occurring in coastal environments protected from strong wave and current action seem the most susceptible to damage by oil, and environments underlain by soft sediments seem to hold oil for the longest periods.34 Despite major efforts, we are still, as Clark (1982) put it when summarizing an early meet34

These environments include estuaries with mangrove and salt marsh fringes. The oil spill of the barge Florida, discussed above, showed the susceptibility and the persistence of oil in these environments. Proﬁtt and Devlin (1998) and Duke and Watkinson (2002) review recent information on oil pollution in mangrove forests.

ing on the issues, having to deal with “considerable divergence of views . . . oil pollution may have serious local and temporary consequences, but they are no greater, and generally less than natural ﬂuctuations . . .”. The documented cases of catastrophic oil spills occur at local spatial scales, and show recovery after some time. But there is a worrisome aspect that has been much less well documented, and may have more global dimensions. If we return to Table 7.3 we can see that the large human inputs of oil to coastal waters are the aggregate drip-drip of daily inputs from areas with human populations, plus routine vessel operations. Maybe we have been too mesmerized by the attention to spectacular tanker accidents to notice what may be more pervasive, though largely unnoticed, threats posed by local chronic releases of oil at many sites across the world’s shores. This issue has been recognized for some time (Clark 1982; Farrington 1985), but there is remarkably little awareness of it in the public and political sectors. There are troubling circumstantial bits of information: for example, the number of dead birds found on Northern European shores is greatest in places where the highest percent of those dead birds were noticeably oiled. We should not make too much of such correlational data, but if Fig. 7.11 is representative, it Mean number of birds per km

166

6

4

2

0 0

20

40 60 Birds oiled (%)

80

100

Figure 7.11 Relationship of mean number of dead birds found per kilometer of shore and percentage of those birds that were visibly oiled. Data collected in February, 1967–1973, from the coasts of the British Isles and the north coast of Europe from Brittany to the mouth of the Baltic Sea. From Clark (1997).

PETROLEUM HYDROCARBONS

might point out that maybe there is a widespread, chronic, and largely unnoticed death tithe of marine life taken by the small but widespread release of petroleum around areas with much human activity. Each of these local inputs is a renewed initial kill; the effects of each new input will recede as the oil weathers, but there are always new little inputs, each of which can have its impact in the same or a nearby area. Across decades, a pattern of local, but widespread, repeated perturbations could have more degrading and certainly more wide-ranging effects than major tanker spills. This is certainly an area requiring closer enquiry; at present we can say very little about it. Kinds of effects There is a rather heterogenous response of organisms to oil. This bewildering variation derives from differences in exposure to different concentrations of oil, the composition and persistence of the petroleum compounds present, the differential ability of various organisms35 to accumulate and metabolize oil compounds, and the differing degree to which the different compounds interfere with many metabolic processes (Capuzzo & Kester 1987). Cellular and biochemical,36 physiological, and behavioral effects of many different kinds have been unambiguously linked to expos-

167

ure to oil. We have seen examples of organisms that show biochemical evidence of exposure, but no measurable population-level consequences. It would be desirable to be able to say clearly that the detection of increased mixed function oxygenases, for example, translates into changes at the level of the individual or population—what we would see as changes in abundance. Linking exposure, external or internal oil concentrations, and oxydase responses, to population consequences has been a challenging task (Hansen et al. 1999). One study that linked spilled oil to responses from the biochemical to the population levels is that on ﬁddler crabs in the Florida spill (Krebs & Burns 1977). Increased concentrations of naphthalenes ingested from the sediments apparently impaired muscular coordination in the ﬁddler crabs. These compounds acted as possible narcotics, slowing escape responses and making burrowing difﬁcult, both of which could have led to the greater mortality, and lowered abundance observed in the ﬁeld for some years after the spill.37 This study hence linked exposure to oil, through incorporation of oil into organisms, through to the population consequences, so that we can conclude that effects on abundance are the likely result of long-term oil exposure detected by biochemical markers.

Managing oil in coastal waters 35

A plethora of kinds of organisms have been involved in the many hundreds of papers on oil spills and their consequences. There is one notable gap: minimal attention has been devoted to microbial responses. For instance, in the two thick volumes reporting the results of the Exxon Valdez work, there is only one paper (Braddock et al. 1996) devoted to microbes. Characteristically, the one paper on microbes focuses on oil-degrading bacteria (Leahy & Colwell 1990), as do most other papers on the subject (Karrick 1977). It would be of some interest to know more about how oil alters the many biogeochemical tasks carried out by microbes, tasks that are key to maintaining the functioning of coastal ecosystems. 36 Exposure to oils has been linked to a number of abnormalities in cell and tissue chemistry, detectable by different methods. Probably the most widely used method to assess exposure to oils is to measure the appearance of certain enzymes, called mixed function oxygenases, that become active after exposure to oil or other fat-soluble compounds (Stegeman 1981). Interpretation of these indicators is uncertain, since the increase in these indicators may mean that the organism is detoxifying its oil burden, or that there is the potential for damage (Capuzzo & Kester 1987).

Even as demand for petroleum products in the last few decades has increased worldwide (see Fig. 7.7), the number of accidental spills and the volume of oil released into marine environments have decreased (Fig. 7.12). Both the number of tanker accidents and the mean volume of oil spilled per incident has decreased in recent decades, and the trend seems downward still. Large war-related or blow-out spills prominently and intermittently appear in the historical record. Such events unpredictably add large volumes of oil to local areas of sea. The more diffuse, but large, inputs from urban waste waters, carried 37

Exposure to low concentrations of pesticides have similar effects (Ward & Busch 1976).

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80 60 40 20 0 1972 1976 1980 1984 1988 1992 1996 2000 60

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40

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Figure 7.12 Yearly volume of oil spilled from accidents involving oil transport vessels, where the volume of oil was greater than 1,000 barrels (about 136 metric tons) (top; data from Environment Canada, http://www.etcentre.org) and for warrelated or blow-out accidents (bottom; data from US Ofﬁce of Technology Assessment), for the world, 1974–1997.

by streams and sewers (see Table 7.3), might parallel the downward trend of the tanker data (GESAMP 1993). If so, on aggregate, signiﬁcantly less oil may be entering the world’s seas than earlier in the 20th century. In fact, the sediment record suggests that oil inputs to control sediments have decreased in recent years in the Gulf of Mexico (Fig. 7.13) and elsewhere. The reduced world-scale loading of oil into the sea is, of course, no consolation to those people in places where spills will still inevitably occur. As this is written, 12,000 tons of crude oil spilled after the sinking of the tanker Prestige off the ﬁnand shellﬁsh-rich coast of Galicia is spreading

6,000 12,000 18,000 −1 PAHs (ng g dry wt)

24,000

Figure 7.13 Vertical proﬁles of polycyclic aromatic hydrocarbon (PAH) concentrations in sediments of three sites on the northern coast of the Gulf of Mexico. Data courtesy of Santschi et al. (2001).

along the north coast of Spain and threatening to reach French shores. The slick is followed with grave concern by the important ﬁn- and shellﬁshing and mariculture industries and the tourist trade in the region. Fortunately, so far, the Prestige event has “. . . been a tragedy but not a disaster”, with only modest wildlife losses this far, but it does point out that oil will be spilled for as long as we need oil. Many ideas have been proposed to recover spilled oil, or to minimize the release of petroleum products into the sea. In terms of dealing with spilled oil, collection, dispersants, washing, and microbial enhancement have been tried. The idea of using technological devices to collect

PETROLEUM HYDROCARBONS

spilled oil has been put into action, and, in general, booms, raking, and skimming of spilled oil seem to recover perhaps 10% of the volume spilled (GESAMP 1993). Chemical dispersants have been used to lower concentrations of oil by mixing the oil more effectively in water. Reviews of the effectiveness of dispersants range from positive (NRC 1989) to questionable (GESAMP 1993). Dispersants may make oil less conspicuous, and dilute concentrations in a spill. Their use might be useful in the open sea, where mixing with large volumes effectively dilutes hydrocarbons, but use of dispersants near-shore may lead to greater ecological change than the oil itself. Enrichment with nitrogen and phosphorus fertilizers to increase microbial degradation of petroleum has been tried, with variable success (GESAMP 1993); from the data available, enhancement of microbial degradation seems to have modest prospects for general use. Washing with high-pressure hoses has been used in many cases, but as we saw in the Exxon Valdez instance, such treatment can have greater ecological effects than oiling itself. In terms of prevention, it has been proposed to require double hulls on tankers as a prevention against leakage after groundings. An enhancement of awareness of the dangers of groundings, and deployment of safety procedures, no doubt have been instrumental in lowering the release of oil to the world’s coasts in recent decades (Fig. 7.12). Improvement of sewage treatment and deviation of waste outfalls have also likely been effective at somewhat lowering worldwide (but in many instances not local) municipal releases of oil. Given the attention to the dangers of oil, it has been surprising how little attention has been given to plans to lower dependence on fossil fuels or to increase efﬁciency in the use of the fuels we do use. It seems obvious that the extent to which we shift to other energy sources and increase efﬁciency, will lower the chances of releasing oil into coastal waters. All preventive measures require economic and political action of massive proportions, and are likely to be opposed by the petroleum industry, one of the world’s most powerful sectors.

169

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Laur, D., and L. Haldorson. 1996. Coastal habitat studies: The effect of the Exxon Valdez oil spill on shallow subtidal ﬁshes on Prince William Sound. Am. Fish. Soc. 18:659– 670. Leahy, J. G., and R. R. Colwell. 1990. Microbial degradation of hydrocarbons in the environment. Microbiol. Rev. 54:305–315. Loughlin, T. R. (ed.). 1994. Marine Mammals and the Exxon Valdez. Academic Press, London, 395 pp. McGurk, M. D., and E. D. Brown. 1996. Egg–larval mortality of Paciﬁc herring in Prince William Sound, Alaska, after the Exxon Valdez oil spill. Can. J. Fish. Aquat. Sci. 53:2343– 2354. Marty, G. D., M. S. Okihiro, E. D. Brown, D. Hanes, and D. E. Hinton. 1999. Histopathology of adult Paciﬁc herring in Prince William Sound, Alaska, after the Exxon Valdez oil spill. Can. J. Fish. Aquat. Sci. 56:419–426. Matkin, C. O., G. M. Ellis, M. E. Dahlheim, and J. Zeh. 1994. Status of killer whales in Prince William Sound, 1985– 1992. Pp. 141–162 in Loughlin, T. R. (ed.). Marine Mammals and the Exxon Valdez. Academic Press, London, 395 pp. Matthews, C. P., S. Kedidi, N. I. Fita, A. I. Al-Yahya, and K. Al-Rasheed. 1993. Preliminary assessment of the effects of the 1991 Gulf War on Saudi Arabian prawn stocks. Mar. Pollut. Bull. 27: 251–271. Mearns, A. J. 1996. Exxon Valdez shoreline treatment and operations: Implications for response, assessment, monitoring, and research. Am. Fish. Soc. Symp. 18:309–328. Michel, J., and M. O. Hayes. 1999. Weathering patterns of oil residues eight years after the Exxon Valdez oil spill. Mar. Pollut. Bull. 38:855–863. Monson, D. H., D. F. Doak, B. E. Ballechey, A. Johnson, and J. L. Bodkin. 2000. Long-term impacts of the Exxon Valdez oil spill on sea otters, assessed through age-dependent mortality patterns. Proc. Nat. Acad. Sci. 97:6562–6567. Murphy, M. L., R. A. Heintz, J. W. Short, M. L. Larsen, and S. D. Rice. 1999. Recovery of pink salmon spawning areas after the Exxon Valdez oil spill. Trans. Am. Fish. Soc. 128:909–918. Murphy, S. M., and T. J. Mabee. 2000. Status of black oystercatchers in Prince William Sound, Alaska, nine years after the Exxon Valdez oil spill. Waterbirds 23:204–213. Murphy, S. M., R. H. Day, J. A., Wiens, and K. R. Parker. 1997. Effects of the Exxon Valdez oil spill on birds: Comparisons of pre- and post-spill surveys in Prince William Sound, Alaska. Condor 99:299–313. Nelson, W. G. 1982. Experimental studies of oil pollution on the rocky intertidal community of a Norwegian fjord. J. Exp. Mar. Biol. Ecol. 65:121–138. NRC (National Research Council). 1989. Using Oil Dispersants on the Sea. National Academy Press, Washington, DC, 335 pp. NRC (National Research Council). 2003. Oil in the Sea III: Inputs, Fates, and Effects. National Academies Press, Washington, DC, 265 pp. Oviatt, C., J. Frithsen, J. Gearing, and P. Gearing. 1982. Low chronic additions of No. 2 fuel oil: Chemical behavior,

biological impact and recovery in a simulated estuarine environment. Mar. Ecol. Prog. Ser. 9:121–136. Paine, R. T., and 7 othrers. 1996. Trouble on oiled waters: Lessons from the Exxon Valdez oil spill. Annu. Rev. Ecol. Syst. 27:197–235. Pearson, W. H., R. A. Elston, R. W. Bienert, A. S. Drum, and L. D. Atrim. 1999. Why did the Prince William Sound, Alaska, Paciﬁc herring (Clupea pallasi) ﬁsheries collapse in 1993 and 1994? Review of hypotheses. Can. J. Fish. Aquat. Sci. 56:711–737. Peterson, C. H. 2001. The “Exxon Valdez” oil spill in Alaska: Acute, indirect, and chronic effects on the ecosystem. Adv. Mar. Biol. 39:3–103. Peterson, C. H., L. L. McDonald, R. H. Green, and W. P. Erickson. 2001. Sampling design begets conclusions: The statistical basis for detection of injury to and recovery of shoreline communities after the “Exxon Valdez” oil spill. Mar. Ecol. Prog. Ser. 210:255–283. Peterson, C. H., and 6 others. 2003. Long-term ecosystem response to the Exxon Valdez oil spill. Science 302:2082– 2086. Piatt, J. F., and P. Anderson. 1996. Response of common murres to the Exxon Valdez oil spill and long-term changes in the Gulf of Alaska marine ecosystem. Am. Fish. Soc. Symp. 18:720–737. Piatt, J. F., and R. G. Ford. 1996. How many seabirds were killed by the Exxon Valdez oil spill? Am. Fish. Soc. Symp. 18:712–719. Price A. R. G., T. J. Wrathall, P. A. H. Medley, and A. H. AlMoamen. 1993. Broadscale changes in coastal ecosystems of the Western Gulf following the 1991 Gulf War. Mar. Pollut. Bull. 27:143–148. Pritchard, P. J., and C. F. Costa. 1991. EPA’s Alaska oil spill bioremediation project. Environ. Sci. Technol. 25:372– 379. Profﬁtt, C. E., and D. J. Devlin. 1998. Are there cumulative effects in red mangroves from oil spills during seedling and sapling stages? Ecol. Appl. 8:121–127. Rabalais, S. C., and R. W. Flint. 1983. IXTOC-1 effects on intertidal and subtidal infauna of south Texas Gulf beaches. Contr. Mar. Sci. 26:23–35. Readman, J. W., J. Bartocci, I. Tolosa, S. W. Fowler, B. Oregoni, and M. Y. Abdulraheem. 1996. Recovery of the coastal marine environment in the Gulf following the 1991 war-related oil spills. Mar. Pollut. Bull. 32:493– 498. Reddy, C. M., and 6 others. 2002. The West Falmouth oil spill after thirty years: The persistence of petroleum hydrocarbons in marsh sediments. Environ. Sci. Technol. 36:4754–4760. Rice, S. D., R. B. Spies, D. A. Wolfe, and B. A. Wright (eds). 1996. Proceedings of the Exxon Valdez Oil Spill Symposium. American Fisheries Society Symposium No. 18. Bethesda, MD. Rice, S. D., and 7 others. 2001. Impacts to pink salmon following the Exxon Valdez oil spill: Persistence, toxicity, sensitivity, and controversy. Rev. Fish. Sci. 9:165–211.

PETROLEUM HYDROCARBONS

Rosenberg, D. H., and M. J. Petrula. 1998. Status of harlequin duck in Prince William Sound, Alaska, after the Exxon Valdez oil spill, 1995–1997. Exxon Valdez Oil Spill Restoration Project (97427). Final Report. Alaska Department of Fish and Game, Division of Wildlife Conservation, Anchorage, AK. Russell, J. C., M. A. Downs, J. S. Petterson, and L. A. Palinkas. 1996. Psychological and social impacts of the Exxon Valdez oil spill and cleanup. Am. Fish. Soc. Symp. 18:867–894. Sanders, H. L., J. F. Grassle, G. R. Hampson, L. S. Morse, S. Garner-Price, and C. C. Jones. 1980. Anatomy of an oil spill: Long-term effects from the grounding of the Barge Florida off West Falmouth, Massachusetts. J. Mar. Res. 38:265–380. Short, J. W., and 5 others. 2004. Estimate of oil persisting on the beaches of Prince William Sound 12 years after the Exxon Valdez oil spill. Environ. Sci. Technol. 38:19– 25. Spies, R. B., S. D. Rice, D. A. Wolfe, and B. A. Wright. 1996. The effects of the Exxon Valdez oil spill on the Alaskan coastal environment. Am. Fish. Soc. Symp. 18:1–16. Stacey, B. M., and B. M. Marcotte. 1987. Chronic effect of No. 2 fuel oil on population dynamics of harpacticoid copepods in experimental marine mesocosms. Mar. Ecol. Prog. Ser. 40:61–68. Stegeman, J. J. 1981. Polynuclear aromatic hydrocarbons and their metabolism in the marine environment. Pp. 1–60 in Gelboin, H. V., and P. O. P. Ts’o (eds). Polycyclic Hydrocarbons and Cancer, Vol. 3. Academic Press, New York. Stekoll, M. S., and L. Deysher. 2000. Response of the dominant alga Fucus gardneri (Silva) (Phaeophyceae) to the Exxon Valdez oil spill and clean-up. Mar. Pollut. Bull. 40:1028– 1041. Stone, R. 1992. Oil clean-up method questioned. Science 257:320–321. Tawﬁq, N. I., and D. A. Olsen. 1993. Saudi Arabia’s response to the 1991 Gulf oil spill. Mar. Pollut. Bull. 27:333–345. Teal, J. M., and 6 others. 1992. The West Falmouth oil spill after 20 years: Fate of fuel oil compounds and effects on animals. Mar. Pollut. Bull. 24:607–614. Thomas, R. E., C. Brodersen, M. G. Carls, M. Babcock, and S. D. Rice. 1999. Lack of physiological responses to hydrocarbon accumulation by Mytilus trossulus after 3–4 years chronic exposure to spilled Exxon Valdez crude oil in Prince William Sound. Comp. Biochem. Physiol. 122C: 153–163. Trust, K. A., D. Esler, B. R. Woodin, and J. J. Stegeman. 2000. Cytochrome P450 1a induction in sea ducks inhabiting

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nearshore areas of Prince William Sound, Alaska. Mar. Pollut. Bull. 40:397–403. Vargo, G. A., M. Hutchins, and G. Almquist. 1982. The effect of low, chronic levels of No. 2 fuel oil on natural phytoplankton assemblages in microcosms: 1. Species composition and seasonal succession. Mar. Environ. Res. 6:245–264. Vazquez, M. A., K. W. Allen, and Y. M. Kattan. 2000. Longterm effects of the 1991 Gulf War on the hydrocarbon levels in clams at selected areas of the Saudi Arabian coastline. Mar. Pollut. Bull. 40:440–448. Ward, D. (Valiela), and D. A. Busch. 1976. Effects of Temefos, an organophosphorous insecticide, on survival and escape behavior of the marsh ﬁddler crab Uca pugnax. Oikos 27:331–335. Wells, P. G., J. N. Butler, and J. S. Hughes (eds). 1995. Exxon Valdez Oil Spill: Fate and Effects in Alaskan Waters. American Society of Testing and Materials, Philadelphia, 955 pp. Wertheimer, A. C., N. J. Bax, A. G. Celewycz, M. G. Carls, and J. H. Landingham. 1996. Harpacticoid copepod abundance and population structure in Prince William Sound, one year after the Exxon Valdez oil spill. Am. Fish. Soc. Symp. 18:551–563. Wheelright, J. 1994. Degrees of Disaster: Prince William Sound: How Nature Reels and Rebounds. Simon & Schuster, New York, 348 pp. Whittle, K. J., R. Hardy, P. R. Mackie, and A. S. McGill. 1982. A quantitative assessment of the sources and fate of petroleum compounds in the marine environment. Phil. Trans. R. Soc. Lond. B 297:193–218. Wiens, J. A. 1996. Oil, seabirds, and science: The effects of the Exxon Valdez oil spill. BioScience 46:587–597. Wiens, J. A., and K. R. Parker. 1995. Analyzing the effects of accidental environmental impacts: Approaches and assumptions. Ecol. Appl. 5:1069–1083. Wiens, J. A., T. O. Crist, R. H. Day, S. M. Murphy, and G. D. Hayward. 1996. Effects of the Exxon Valdez oil spill on marine communities in Prince William Sound, Alaska. Ecol. Appl. 6:828–841. Wolfe, D. A., and 11 others. 1994. The fate of oil spilled from the Exxon Valdez. Environ. Sci. Technol. 28:561–568. Woodin, B. R., R. M. Smolowitz, and J. J. Stegeman. 1997. Induction of cytochrome P4501A in the intertidal ﬁsh (Anoplarchus purpurescens) by Prudhoe Bay crude oil and environmental induction in ﬁsh from Prince William Sound. Environ. Sci. Technol. 31:1198–1205. Ziegesar, O. von, E. Miller, and M. E. Dahlheim. 1994. Impacts on humpback whales in Prince William Sound. Pp. 173–192 in Loughlin, T. R. (ed.). Marine Mammals and the Exxon Valdez. Academic Press, London, 395 pp.

Chapter 8 Chlorinated hydrocarbons

Whale ship in New Bedford Harbor. This vessel set out September 26, 1871 and returned 4.5 years later with a load of sperm oil (note casks for oil in foreground) and 36,085 lb of baleen. The Helen Mar struck an ice ﬂoe in the Arctic on October 6, 1892 and was lost, with only ﬁve survivors out of 39 crew (Whitman 1994). Photo, “Bark Helen Mar at New Bedford”, taken by Stephen F. Adams, September 1871. Used by courtesy of the Trustees of the Old Dartmouth Historical Society.

CHLORINATED HYDROCARBONS

A case history: New Bedford Harbor1 The city of New Bedford, on the southeastern coast of Massachusetts, has had to deal with serious shifts in economic and environmental fortunes in its 250 years of history. The watershed emptying into the New Bedford Harbor area was mainly farms and forests, ever since the early European settlers arrived in the 1600s. It was not until after the 1850s that New Bedford, through an unforeseen juxtaposition of hydrographic, economic, and ﬁshery circumstances, gained prominence. Nantucket Island, off the coast of Massachusetts, was the world center for whaling in the late 1700s and early 1800s. As whalers increased their harvest of nearby whale stocks, it became necessary to use and outﬁt larger and larger vessels to make the many-month—even multiyear—trips economically feasible. In the mid-1800s the drafts of the whaling ships became too large to enter the shallow Nantucket Harbor, and whalers were eventually forced to ﬁnd deeper ports, the major one of which turned out to be New Bedford Harbor. “. . . New Bedford has of late been gradually monopolizing the business of whaling . . . poor old Nantucket is now much behind her . . .” (Chapter ii, Moby Dick, by H. Melville).

The establishment of New Bedford as the premier whaling center (see Chapter 8 frontispiece), along with other related marine activities and whale product processing, brought prosperity to the city, and increased urbanization, including the construction of sewer lines that discharged directly into the harbor. “Had it not been for us whalemen, [New Bedford] would have been in as howling condition as the coast of Labrador . . . the town itself is perhaps the dearest place to live in, in all New England . . . nowhere in all America will you ﬁnd more patrician-like houses;

1

Material in this section is from Nelson et al. (1996), Weaver (1983), Pesch and Garber (2001), Summerhayes et al. (1977), and the New Bedford Whaling Museum webpage.

175

parks and gardens more opulent, than in New Bedford” (Chapter vi, Moby Dick, by H. Melville).

No doubt, the larger population, increased urban activity, and the whale product industry released many contaminants, including hydrocarbons and metals into the water and sediments of the harbor via the newly built sewer lines and from burning. Whaling ﬂourished in New Bedford during the 1850s, with 447 ships, barks, and schooners. This ﬂeet brought in about 64% of the US whale tonnage, and amounted to about half of the world’s harvest. In the following decades, whaling faltered. In 1859 ready sources of oil were discovered in Pennsylvania, and oil quickly replaced whale oil for use in providing light. Historical accidents led to the destruction of many ships. The Confederates destroyed many Yankee whale ships during the US Civil War (1861–1865). In 1871, 33 whale ships (22 from New Bedford) were lost as ice closed in on them before they could sail south for the winter. In 1876, of the 20 vessels that plied the Arctic for whales, another 12 ships were lost. These ship losses were never recouped by the construction of new vessels. By the early 1900s, use of baleen for clothing and other purposes was abandoned, and whaling ﬁnally ended in New England. The economy of New Bedford, based so strongly on whale products, suffered. Efforts to revive economic enterprise in New Bedford turned to textile mills after the disappearance of whaling. Textile work led to sharp increases in the city’s population, with more waste water released into the harbor, and ﬁlling in of wetlands, as well as increased ﬂow of chemical efﬂuents from the mills and other lesser industries. As a local newspaper described the upper harbor, “. . . water [was] thick with slime and shores [were] covered with ﬁlth from the sewers”. The textile industry could not survive the economic crisis of the 1920s and 1930s, and New Bedford was forced to diversify its income base. Fishing became a mainstay. Fishing-related business boomed, with New Bedford ﬁshermen being in the lead in exploiting the rich stocks of nearby Georges Bank. Even through to the 1980s, the port of New Bedford was ﬁrst in the USA in

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terms of the value of ﬁsh landings. Fish-processing plants released large loads of offal directly into the harbor, and decay of this organic matter must have sharply lowered the oxygen content of harbor water. Among the other industries that were encouraged after the textile industry faltered was the manufacturing of electronic parts. Two companies set up business in 1939–1941 in abandoned textile mills near the harbor. These companies used large quantities of certain industrially manufactured compounds, known as polychlorinated biphenyls (PCBs), in their products (capacitors and transformers). For several decades, these manufacturing plants produced much waste water, hence PCB-containing wastes were released directly by pipes from the plants, and through the outfall of the municipal wastewater treatment plant, into the harbor. In 1974–1975, scientiﬁc reports showed PCBs in birds and sediments of New Bedford Harbor and Buzzards Bay. In 1976, a monitoring survey by the US Environmental Protection Agency (EPA) revealed high concentrations of PCBs in the harbor (Fig. 8.1), in addition to many other contaminants contributed by the long history of human use of the harbor. Concern about the high concentrations, and about the wildlife and human effects connected with other PCB contamination cases in Sweden, Hudson River, and Lake Michigan, prompted an intensive series of studies to assess the degree of contamination, deﬁne the effects, and identify possible ways to remedy the situation. The concentrations of PCBs measured in sediments of New Bedford Harbor were among the highest recorded. Concentrations were highest in the inner harbor, and decreased down-estuary (Fig. 8.1). In certain places, concentrations of PCBs within sediments reached values up to four orders of magnitude larger than standards said to lead to adverse biological effects (Pruell et al. 1990; Long et al. 1995). The impressive degree of contamination is conveyed by the calculation that up to 15% of the weight of sediment in some samples was made up by PCBs. These compounds have a long life in coastal environments; PCBs released

PCBs in sediment (µg g−1 dry wt) >50 11–50 1–10 <1 no data Electronic industry sites N

0

500 1,000 m

Hurricane Barrier

Figure 8.1 Distribution of polychlorinated biphenyl (PCB) concentrations in sediments of New Bedford Harbor. From Pesch and Garber (2001); data from Nelson et al. (1996).

many decades ago are still there as a historical record in the sediments. The consequences of PCB contamination in New Bedford Harbor were made evident by elevated concentrations within organisms, and by their cellular, physiological, and population responses (Table 8.1). Microorganisms (bacteria, dinoﬂagellates, tintinnids) seemed unfettered by the presence of PCBs (Table 8.1). High concentrations of the many other human-derived substances, such as nutrients, fecal matter, and ﬁsh offal, probably

177

CHLORINATED HYDROCARBONS

Table 8.1 Selected summary of reported effects of polychlorinated biphenyls (PCBs) on organisms in New Bedford Harbor (NBH). Type of organism

Effects

Source

Bacteria

Diversity of distinct taxonomic units increased in NBH relative to uncontaminated sediments

Sorci et al. 1999

Dinoflagellates

Much more abundant in NBH than outside the harbor

Pierce & Turner 1994a

Tintinnids

Most abundant near the sewage outfall in NBH

Pierce & Turner 1994b

Arthropods

10–110 ppb PCBs are in toxic range for amphipods and mysid shrimp Lobsters had high PCB concentrations early in life, with a lower load later

Ho et al. 1997

See Fig. 8.2 Condition index (= dry wt/shell vol.) 32% lower, and reproductive effort 52% lower, in NBH relative to reference site offshore Tissue PCB content could be lowered 33–53% and 67–84% by remediation that lowered sediment PCB content to 50 ppb and to 10 ppb, respectively

Farrington et al. 1987 McDowell Capuzzo 1996

Soft-shell clams

High PCB content of tissues Incidence of leukemia (hemocytic neoplasia) 17% of clams introduced to NBH developed leukemia, compared to 1–6% in sites outside the harbor

McDowell Capuzzo 1996 Dopp et al. 1996; Strandberg et al. 1998 Craig et al. 1993

Eels

Higher cytochrome P450 activity in NBH eels than in those from less polluted sites

Schlezinger & Stegeman 2000

Killifish

324, 163, and 2.4 µg g−1 dry wt in livers of adult killifish from inner, mid-, and outside NBH, respectively; resistance developed in native fish PCB in livers and mortality increased; growth lowered by diets containing PCBs in non-native fish See Table 8.2

Nacci et al. 1999

Scup

Increased P450 in fish caught in outer NBH

Stegeman et al. 1991

Winter flounder

Increased P450 in fish caught in outer NBH Smaller young hatch from eggs with more PCB

Stegeman et al. 1987; Elskus et al. 1989 Black et al. 1988

Tomatoes

See Table 8.3

Cullen et al. 1996

Humans

Higher PCB concentrations in blood of New Bedford residents (in 1981)

Mussels

stimulated microbe abundance in New Bedford Harbor. Concentrations of PCBs in New Bedford Harbor were well within the range that would be toxic to arthropods, but in lobsters concentrations of

Mercaldo-Allen et al. 1994

Abdelrhaman et al. 1998

Gutjahr-Gobell et al. 1999 Black et al. 1998

PCBs decreased as the lobsters aged (Table 8.1), probably a result of depuration or elimination of PCBs in the shells during molting. There was, moreover, no lack of lobsters in the harbor and surrounding areas.

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(NBH: 31,000)

31,000 Blue mussels Oysters

500 PCB conc. (ng g–1)

PCB concentration (ng g–1 dry wt)

500

400

400 300 200 100

300

0 0

400

1,200

No. people km−2

200

100

0

ME

MA

RI

CT

NY

NJ DE US state

MD

VA

Analyses of the PCB content of mussels and oysters collected in New Bedford Harbor2—and elsewhere along the eastern seaboard of the USA —showed high concentrations of PCBs, but New Bedford Harbor mussels were clearly a remarkable outlier, with concentrations perhaps two orders of magnitudes higher than other places sampled (Fig. 8.2). Concentration of PCBs in shellﬁsh was related to density of people: concentrations of bivalves collected from sites near coastal areas with larger densities of people contained larger amounts of PCBs (Fig. 8.2 inset).3 The high concentrations of PCBs in clams in New 2

800

The US and International Mussel Watch Programs made good use of bivalves as sentinels to assess the degree of contamination of coastal waters by polycyclic and chlorinated hydrocarbons, heavy metals, and radionuclides (Farrington et al. 1983). It cannot be assumed that concentrations in all organisms is a general indicator of degree of contamination or exposure, however. Concentrations of PCBs, for example, in ribbed mussels and killiﬁsh of New Bedford Harbor were correlated to PCBs in water and in sediments, but concentrations in shrimp and eels were not (Lake et al. 1995). Similarly, the responses of physiological indicators of exposure also differed from one organism type to another (Table 8.1). 3 This pattern seems to be quite general, although perhaps not as marked as in the New Bedford example. For instance, in coastal waters off Denmark, PCB concentrations in organisms and sediments in coastal sites near urban/industrial complexes were some 3.5 times those recorded in offshore sites (Gunnarsson & Sköld 1999).

NC

SC

GA

Figure 8.2 Concentrations of polychlorinated biphenyls (PCBs) in mussels and oysters collected from different states along the eastern seaboard of the USA. Inset: PCB concentrations plotted vs. rough estimates of the density of people on coastal areas within the watersheds of sites where PCBs were measured. NBH, New Bedford Harbour. Data on bivalve PCB concentrations from Farrington et al. (1982).

Bedford Harbor did not lead to lower populations, but there were suggestions that clams with larger internal PCB contents showed increased incidence of disease (Table 8.1). Killiﬁsh from more contaminated parts of the harbor contained more PCBs in their livers (Table 8.2). These higher concentrations were associated with greater mortality and somewhat lowered growth rates in the adult killiﬁsh. The progeny of these killiﬁsh may have suffered greater rates of abnormalities after hatching, but the data were scanty. There was evidence that exposure to the PCB concentrations found in New Bedford Harbor selected for PCB-resistant ﬁsh (Table 8.1). The unusually high concentrations of PCBs (and the many other contaminants), however, did not eradicate marine life in New Bedford Harbor. The reported impairments to health and reproduction of certain organisms did not translate into reductions in the local populations. Many species of ﬁsh and shellﬁsh continue to be abundant, and grow well in the harbor (Nacci et al. 1999; D. Leavitt, personal communication). Local concern about PCB contamination rose in 1981 after public disclosure that PCB concentrations in the blood of New Bedford residents were

179

CHLORINATED HYDROCARBONS

Table 8.2 Differences in polychlorinated biphenyl (PCB) content of liver and selected responses in killiﬁsh (Fundulus heteroclitus) from the inner New Bedford Harbor, and from West Island, a site outside the harbor (two sites from each area; nd = no data). Data from Black et al. (1998).

PCB in liver (µg g−1 dry wt) Mortality of females in laboratory (% of initial number) Growth of females in laboratory Mortality of progeny (%) Spinal abnormalities in progeny

Outside harbor

Inner New Bedford Harbor

0.5, 9.5 0, 0

20.8, 29.3 30, 23

– 30, nd 7, nd

Reduced nd, 51 nd, 26

higher than in other US regions (Farrington et al. 1985). This discovery led to closing down the harvest of lobsters and shellﬁsh in the harbor. The inner and middle harbor are still closed to harvest. The lack of exploitation has contributed to the high densities of lobsters and clams present, in spite of the PCB residues in the sediments. Increased awareness of the PCB problem, in particular the public health aspects,4 prompted the US Congress to pass the Toxic Substances Act. This allowed the EPA to ban the synthesis of PCBs in the US Monsanto Corporation, the only US producer of PCBs, who ceased production of PCBs in 1978. New Bedford electronics companies then replaced PCBs with other compounds in their manufacturing. In 1982 the New Bedford Harbor area most affected by PCBs was designated as a US EPA superfund hazardous waste site and remedial action planning began (Farrington et al. 1985). Treatment of industrial waste efﬂuent and municipal sewage have improved signiﬁcantly since the 1980s, with the investment of hundreds of millions of US dollars. These improvements have diminished the discharge of PCBs into New Bedford Harbor, but some PCBs still enter the

harbor from street overﬂow pipes and by atmospheric deposition of wind-blown soil particles with adsorbed PCBs (Weaver 1983; Pesch & Garber 2001). A pilot dredging study took place in 1995 within the area of the inner harbor known to be a PCB “hotspot” with concentrations greater than 4,000 ppm of PCBs in sediments.5 About 7,600 m3 of the highly contaminated sediments are still in sealed containers, awaiting a decision as to what should be done with their contents. The issue of human exposure to PCBs surfaced again during the dredging project, since PCBs might be released to the air during the bringing up and exposure of PCB-laden sediments. Studies done on the air-borne delivery of PCBs to other locally grown vegetables showed no effects, but tomatoes were another story (Table 8.3). Windborne PCBs managed to be incorporated into tomatoes grown downwind of the dredging site, reaching about an eight-fold increase in PCB content. This demonstrated that air-borne transport was possible, and might have accounted at least for some of the increases in the PCB content of blood in people living near New Bedford Harbor.

4

Health effects of PCBs came to the attention of the public after the Yusho and Yucheng incidents in Japan and Taiwan. These were cases of poisoning of thousands of people from consumption of rice-bran oil contaminated with PCBs (Chen & Hsu 1986; Takayama et al. 1991). Perhaps just as attention-getting was the ability of PCBs to be transmitted from lactating mothers to their babies through breast-feeding (Harris et al. 2001).

5

Although dredging was proposed as an effective management approach (Fraser 1993), study of post-dredging distributions of PCB concentrations did not clearly demonstrate the effectiveness of the dredging (Bergen et al. 1998). The redistribution and transformation of PCB compounds up- and down-estuary confounded interpretations of the effect of dredging on patterns of PCB abundance.

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Table 8.3 Concentrations of polychlorinated biphenyls (PCBs) measured in tomatoes grown within the New Bedford area, in sites located upwind and downwind from where dredging of harbor sediments from a PCB “hotspot” took place. Data from Cullen et al. (1996). Concentrations of PCBs (ng g−1) in tomatoesa Upwind from dredging

Downwind from dredging

Before dredging

1.4 ± 0.5 2.1 ± 0.6

3.0 ± 1.3 –

During dredging

3.8 ± 2.0

25 ± 23.5

a

Samples of tomatoes grown in places other than the New Bedford study area had PCB levels of 4.4 ± 1.2 ng g−1.

The long history of increasing use of New Bedford Harbor, culminating in the remarkable degree of PCB contamination, illustrates a pervasive pattern of incremental human pressure on coastal environments. As human populations increase in density and intensity of land use near coasts, we tax the ability of the affected environments to sustain these uses, and often push degradation to levels where not only environmental damage occurs, but even human public health may be threatened. I have used the case of PCB contamination of New Bedford Harbor as a worst-case instance, but there are many other human-generated compounds that have created parallel situations. In the rest of this chapter we look at just one set of such compounds, the chlorinated hydrocarbons.

Structure and properties of chlorinated hydrocarbons6 Chemists found long ago that the addition of chlorine (or other halogen elements) conferred

useful new properties to hydrocarbons. In the mid-20th century, synthesis of compounds such as DDT led to great advances in public health and agriculture, and the use of PCBs facilitated development of the electrical and other industries. Such chlorinated hydrocarbon compounds7 are now widespread in natural environments. In general, human-made chlorinated compounds of low molecular weight (dichlorethane, vinyl chloride, tetrachloride, trichlorethane, and trichlorethylene), although widely used worldwide as solvents or cleaning agents, are rather volatile, and hence do not remain or accumulate in coastal or aquatic environments. Another group of industrial small-molecular-weight halogenated hydrocarbons, the chloroﬂuorocarbons (CFCs) or freons, were mentioned in Chapter 2. These compounds were made to use as coolants, aerosol propellants, and foams; they have been involved in the destruction of atmospheric ozone, but are not of direct consequence in aquatic environments. In contrast, there has been considerable concern about human-made chlorinated hydrocarbons of larger molecular weight, in particular those manufactured as pesticides (DDT and related substances), and those used for industrial purposes (PCBs). For the sake of brevity, this chapter deals only with these two groups of compounds (see box, p. 181). There are many more groups of industrially produced compounds, used for industrial or agricultural purposes, that can reach and alter coastal environments. Chemicals used as deterrents to marine fouling are discussed in Chapter 9 on heavy metals. Compounds used as insecticides, molluskicides, acaricides, fungicides, and herbicides include a large variety of chemical structures. Here we discuss the DDT family as a representative case, but there are many more types of compounds used for agricultural and public health purposes. Similarly, we focus on PCBs as one prominent type of industrial chlorinated hydrocarbon, even though there are many more.

7

6

Material in this section is largely from Kimbrough (1980), Crine (1986), and Waid (1986).

Some small-molecular-weight hydrocarbons containing chlorine, bromine, iodine, and other elements are manufactured naturally by marine algae and cyanobacteria, and have been found in sponges, turtles, dugongs, and porpoises (Moore et al. 2002).

CHLORINATED HYDROCARBONS

181

A primer on DDTs and PCBs

Polychlorinated biphenyls

DDTs

DDT played a major role during the Second World War, preventing the spread of infectious arthropod-borne diseases such as typhus among troops and people affected by war damage. Recognition of the importance of DDT was demonstrated by the awarding of the 1948 Nobel Prize for medicine to Paul Muller, the Swiss chemist who discovered the insecticidal properties of DDT. After the Second World War, use of DDT8 became widespread in most countries of the world as a control of agricultural and forestry pests. To this day, DDT continues to be the major tool available for controlling malaria in underdeveloped tropical areas of the world. DDT as applied for various uses is actually a mix of DDT, DDD, and DDE (Fig. 8.3). These compounds can be thought of as two-carbon aliphatics where one of the carbons has suffered substitution by two aromatic rings. The rings themselves may carry substituted chlorines. The non-ring carbon atoms may or may not bear chlorines (Fig. 8.3). DDT is the predominant form used in commercial insecticides, but once in the environment, degradation of DDT may lead to different proportions of the three compounds, in particular to the formation of DDE.

All PCBs share a basic two-phenyl ring structure (Fig. 8.3). As many as 209 different PCB isomeric compounds can be made by substituting 1–10 chlorine atoms (to make 10 groups of possible congeners) on the carbons making up the rings, but in reality about 19 isomers are major constituents of commercially made PCBs (Cairns et al. 1986).9 The ecologically relevant features of PCBs include resistance to degradation, solubility in lipids rather than water, and propensity to adsorb to particles. These characteristics have led to their long-term persistence in natural environments, accumulation in organisms, and wide distribution. PCBs became widespread in industrial uses10 because they are among the most stable organic compounds known, are excellent conductors of electricity, and are resistant to heat. PCBs, as a group, are generally more resistant to microbial degradation than other hydrocarbons, although they are subject to some degree of degradation in both aerobic or anaerobic environments (Alder et al. 1993; Chen et al. 2001). PCBs with more substituted chlorines are more resistant to microbes. PCBs are soluble in lipids,11 relatively insoluble in water, and are present in sea water as suspended colloidal complexes (Brownawell & Farrington 1986).12 Concentrations of PCBs

8

11

DDT became widely used because it has special properties: it was markedly toxic to insects, but was much less so to mammals, it lasted for many years in the area applied, and was inexpensive. As it turned out, insects soon began to develop resistance to DDT, and evidence of its effects on ﬁsh and birds (shell-thinning in raptors, etc.) accumulated. Use of DDT has been abandoned in developed countries, superceded by many other compounds that are much more short-lived in the natural environment, and target speciﬁc pests more narrowly. 9 Manufacturers marketed mixtures of PCBs. For example, Monsanto made preparations labeled Aroclor 1242, 1248, 1254, or 1260. The ﬁrst two numbers refer to the number of carbon atoms in the biphenyl group, and the last two numbers refer to the approximate percent chlorine content of the mix of compounds. 10 PCBs have been used in a surprising variety of products: capacitors, transformers, plastics, ﬁre retardants, adhesives, carbonless copying paper, paints, newsprint, hydraulic oils, and heat transfer systems (Cairns et al. 1986).

Relative solubility in lipids is estimated by measuring the relative concentration of a given compound in an experimental situation where the compound may dissolve in water or an oil (octanol). Compounds with high octanol/water partition coefﬁcients, referred to as Kow, tend to accumulate in fatty tissues. This is somewhat confusing, since compounds with high Kow also have quite small fractions in the dissolved form. As it happens, compounds with high Kow also adsorb strongly to sediment and organic particles, and are ingested by animals. The PCBs can then move to tissue lipids after ingestion of the particles (Gray 2002). 12 The more chorines in the PCB, the lower its solubility in water, but the positions of the substitutions also matter. The amount of organic matter strongly affects solubility (Chou & Grifﬁn 1986; Sawhney 1986). The less soluble and chlorinated a PCB, the more likely it is that it will be in a colloidal complex (40% of 2 Cl PCBs are in colloidal form, and about 80% of 4–8 Cl PCBs are in colloidal form) (Burgess et al. 1996).

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CHAPTER 8

within lipids in species that are the top predators in marine food webs, such as killer whales, can be several orders of magnitude higher than concentrations found in sea water (Schulz-Bull et al. 1991; Law et al. 1995; Hayteas & Dufﬁeld 2000; Loizeau et al. 2001). PCBs may be vaporized from water and transported by air (Sawhney 1986). This subject is controversial since other properties of PCBs suggest volatility ought to be low (Atlas et al.

Cl

Cl Cl Cl

Cl

Cl Cl H

Cl

DDT

Cl

Cl

Cl

DDD 3

2

1986). PCB compounds adsorb to surfaces of particles. PCBs in which the two rings are on the same plane are most strongly adsorbed to surfaces of particles. PCBs with more irregular shapes and more chlorine atoms are weakly adsorbed, because the irregular shapes interfere with their adsorption to charged particle surfaces (Shaw & Connell 1986). Transport by air-borne and water-borne particles may be the main mode of PCB transport worldwide.

Cl

DDE

2′ 3′ 4′

4 5

6′ 5′ 6 PCB

Concentrations of chlorinated hydrocarbons in organisms Factors affecting chlorinated hydrocarbon content The concentrations of persistent chlorinated hydrocarbons in organisms vary greatly. Concentrations in organisms are affected by the degree of contamination (the supply of hydrocarbons available in the environment), biomagniﬁcation (the transfer of contaminant from food to consumer), bioconcentration (storage within consumers), depuration (the relative ability to metabolize chlorinated hydrocarbons), and the kind and age of the organisms.13

Figure 8.3 Chemical structures of molecules of the DDT family of compounds, and the basic polychlorinated biphenyl (PCB) molecule. Chlorines may appear in any of the numbered sites on the two phenyl rings of PCBs, making for a large number of possible compounds.

Degree of contamination in the environment

Although it seems reasonable to suppose that there has to be some relationship between the concentration of a pollutant of sediments or water, and the internal load of contaminant borne by an organism, this may or may not be applicable to many chlorinated hydrocarbons. For example, sculpin, amphipods, and isopods did not acquire burdens of polychlorinated naphthalenes (PCNs) in proportion to those in sediments (Lundgren et al. 2002). Other processes—perhaps metabolism of PCNs in this case, other mechanisms in others —determine the magnitude of the body burdens of these hydrocarbons. Biomagnification

13

Duration of exposure has been suggested as a possible additional complicating factor. In trout, for example, larger individuals may contain greater concentrations of PCBs, but this seems untrue in perch (Olsson et al. 2000), or for seals (Ruus et al. 1999) and beluga whales (Muir et al. 1996).

Several decades ago, evidence appeared that DDT accumulated in animals, particularly those species that fed near the top of natural food webs

183

CHLORINATED HYDROCARBONS

14

Gray (2002) also notes the difﬁculty in critical comparisons of concentrations among different organisms in a food web. The methods and units used differ widely among different published studies. Small organisms are invariably analyzed as a whole, while most often parts of larger organisms (liver, lipids, feathers, and so on) are analyzed, and accumulations in these different parts may vary widely. In addition, dry weights, wet weights, and weights normalized to lipid contents are often reported, with little attention for providing means to make these different units comparable. 15 Larger δ15N values indicate that the organism considered has fed farther up in the rungs of a food web. Stable isotopic ratios such as δ15N and δ13C have become useful to deﬁne the links in food webs, as well as discern the external sources of elements entering a food web and moving up food webs (Peterson & Fry 1987; Lajtha & Michener 1994; McClelland et al. 1997; McClelland & Valiela 1998).

elsewhere in this section, could also produce increased burdens of contaminants. Probably the most convincing information about biomagniﬁcation comes from studies that relate contaminant content in organisms to stable isotope signatures of nitrogen.15 The signatures (expressed as the δ15N in parts per thousand) provide a quantiﬁcation of the relative position in a food web of the organisms being considered. In general, these studies conclude that the concentration of contaminants such as DDTs and PCBs (Fig. 8.4), as well as other organic contaminants (Broman et al. 1992; Cabana & Rasmussen Conc. (DDE or PCB: ng g–1 lipid wt)

105

DDE F = 18.85** PCB F = 15.26**

104 103 102 101 100 7 102

PCB conc. (ng g–1 wet wt)

(Barker 1958; Woodwell et al. 1967). These accumulations were in the course of time associated with adverse effects on wildlife, particularly on coastal birds (Risebrough et al. 1971; Risebrough 1986), effects that prompted much research and public attention. The underlying mechanism behind the accumulation was thought to be transfer and storage of the contaminant from food to the consumer. This idea was called biomagniﬁcation (also called biological or food chain ampliﬁcation, or trophic magniﬁcation). Most compounds in food eaten by an animal are used metabolically to support life activities; in contrast, pollutants ingested along with the food may be less metabolizable, and hence remain and accumulate within the consumer. Biomagniﬁcation was thought to be the principal pathway for the accumulation of contaminants in invertebrates and ﬁsh (Addison 1976; Young et al. 1980; Thomann 1981), but the most convincing reports of food web biomagniﬁcation involved birds (Gray 2002). Other marine examples cited by Harding et al. (1997) variously showed that biomagniﬁcation took place, did not take place, or occurred in some species within the food web and not in others. A review of the many papers published on biomagniﬁcation since the 1970s show that biomagniﬁcation of contaminants is by no means the rule. Clear evidence of biomagniﬁcation was found by 42% of the studies, a possible increase reported in a further 11%, and no increases were found in the remaining 41% of the papers reviewed (Gray 2002).14 Moreover, it appeared that factors other than biomagniﬁcation, discussed

8

9

10

11

12

13

6 δ15N (‰)

8

14

15

16

y = 0.36e0.40 r 2 = 0.82

101

100

10–1 0

2

4

10

12

Figure 8.4 Top: relationship of concentrations of PCBs and DDEs vs. trophic position (indicated as δ15N values) for an array of prey and predator species from Barents Sea. Data from Hop et al. (2002). Bottom: relationship of PCB concentration to δ15N in ﬁsh from Canadian lakes. Data from Vander Zanden and Rasmussen (1999).

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CHAPTER 8

1,000,000

Capelin Smelt Sliversides Hake Mackerel Gaspareau Herring

100,000

PCBs (ng)

10,000

1,000

100

10

1 1

10

100 1,000 Lipid weight (mg)

1994; Jarman et al. 1996; Tittlemier et al. 2002), tend to be higher in species that are higher in the trophic steps within food webs. Bioconcentration

Bioconcentration refers to the uptake and storage of hydrocarbons in lipid-containing tissues, not necessarily owing just to the ingestion of contaminants from food (Mackay 1982; Barber et al. 1988). Chlorinated hydrocarbons are in general soluble in lipids, and therefore directly penetrate into an animal through lipid membranes lining the gills or other uptake organs;16 they are then stored in fatty tissues.17 16

The propensity of a hydrocarbon to leave the compartment where it is, and to move to another compartment, is called the fugacity of the compound (Mackay 1979; Mackay & Paterson 1982). Fugacity (estimated as the octanol : water partition coefﬁcient) between water and lipids is high for lipophilic compounds such as PCBs, and hence these congeners may move readily from sediment pore water or overlying water to lipid-containing tissues (Gray 2002). 17 The importance of lipids was noted in studies that showed that increases in chlorinated hydrocarbon contents within species at higher trophic steps were eliminated when the PCB content data were normalized to the lipid content of the species (Ten Berge & Hillebrand 1974; Hargrave et al. 1992).

10,000

100,000

Figure 8.5 Relationship between polychlorinated biphenyl (PCB) content and lipid weight in seven ﬁsh species collected in St Georges Bay, Nova Scotia. From Harding et al. (1997).

Bioconcentration has been reported in phytoplankton, zooplankton, other invertebrates, and ﬁsh (Shaw & Connell 1980; Randall et al. 1998). In ﬁsh, for example, there is a reasonably close relationship between PCB concentration and lipid content (Fig. 8.5). Similarly, there is a linkage between PCB and lipid content in invertebrates such as oysters. In oysters, PCB content is strongly affected by the loss of lipid content that occurs when eggs are released; this linkage results in a seasonal cycle of lipid and PCB content of oysters, with peaks associated with the accumulation of lipid before the release of eggs (Fig. 8.6).18 In contrast, relationships between chlorinated hydrocarbon and lipid contents did not emerge in data from a broad worldwide survey of bivalve contaminants done under the International Mussel Watch Project (Fig. 8.7): neither PCBs nor DDTs increased with greater lipid content in a variety of shellﬁsh species. Lipid content, just like biomagniﬁcation, is by no means the sole explanation 18

We note also that in Fig. 8.6 the multiyear trend in concentrations is downward, suggesting that the availability of the PCBs decreased through the period of measurement, so supply also affected concentrations in the oysters.

185

CHLORINATED HYDROCARBONS

Figure 8.6 Changes in polychlorinated biphenyl (PCB) concentrations in oysters collected at three sites on the north coast of the Gulf of Mexico, on a within-year and multiyear (1969–1976) time course. From Stout (1986), modiﬁed from Wilson and Forester (1978).

PCB 1254 (ng g−1 wet wt)

3.0 2.5 2.0 1.5 1.0 0.5 0 1969

behind changes in concentrations of either PCB or DDT; other mechanisms must also play important roles.19 Depuration

Concentrations of chlorinated hydrocarbons in coastal animals may also depend on rate of depuration. Although chlorinated hydrocarbons are resistant, some degree of depuration usually takes place after uptake of chlorinated hydrocarbons. The degree of elimination of a contaminant depends on the type of organism: for example, whales have a relatively low capacity to decompose such contaminants (Tanabe 2002). Depuration rates also depend on the speciﬁc array of compounds involved: note the large differences in rate of depuration of different PCB congener compounds in Fig. 8.8. Chlorinated hydrocarbon compounds are of course quite diverse in structure, and this diversity adds further complexity to the bioaccumulation issue (Boon et al. 1994; Paterson et al. 1998). For example, rates of depuration by shrimp, ﬁsh, and worms in the ﬁeld differed for the different PCB compounds (Fig. 8.8). The different slopes and end points of the data 19

For freshwater fauna (Oliver & Niimi 1988; Van der Oost et al. 1988; Zaranko et al. 1997), and for marine mammals (Boon et al. 1989, 1994), concentrations of halogenated hydrocarbons in organisms were also higher than concentrations that could be accounted for by lipid storage. In perch, a freshwater ﬁsh, both trophic level and lipid in the ﬁsh inﬂuenced the load of chlorinated hydrocarbons (Olsson et al. 2000).

1970

1971

1972 1973 Year

1974

1975

1976

for these organisms betray the rather different ways that different consumers deal with speciﬁc hydrocarbon compounds. The rate and effectiveness of depuration therefore depends on speciﬁc metabolic pathways available to the species, as well as on the speciﬁc molecular structure of the chlorinated hydrocarbon. Type and age of consumer

The type and age of animals involved might give some clue as to the possible retention of pollutants. After all, belonging to taxonomic groups such as invertebrates, ﬁsh, and mammals, for example, could subsume aggregate effects of metabolism, depuration, fat content, food habit, and more. Compilations of estimates of bioaccumulation (the sum of effects of bioconcentration plus biomagniﬁcation) (Fig. 8.9), however, show that differences based on taxonomic classiﬁcation are not clearly discernible. Such possible differences are masked by the large variation of values within each taxonomic group. There are a few salient taxonomic aspects, for example, birds show the most accumulation—indeed it was birds that were the prime example of accumulation that raised public opinion against persistent pesticides. Gulls, unaccountably, seem to consistently bioaccumulate more than other types of bird, but in most cases, the type of animal is not a reliable predictor of pollutant retention.

186

CHAPTER 8

10,000

PCB (ng g−1 dry wt)

1,000 +

+

+

100

+ +

10

1 1

10

100

10,000

A.s. B.r. C.c. D.d. M.c. M. S.a.

DDT (ng g−1 dry wt)

1,000

100

A.t. C.u. + Cr.c. I.a. + M.e. P.p T.m.

A.b. A.a. C.c. C.f. Cr.r C.s. M.f. M.g. M.g. M.p. Pe.p. + P.g. T.b.i. T.i.

+ + + +

10 +

+ +

+ +

+

1 0

10 Lipid content (mg g−1 dry wt)

100

Figure 8.7 Concentrations of polychlorinated biphenyls (PCBs; top) and DDTs (bottom) in the tissues of several species of bivalves (indicated merely by the initials of their scientiﬁc names) plotted vs. their lipid content. Data from Farrington and Tripp (1995).

The age of the animal also seems to affect chlorinated hydrocarbon content. PCB concentration, divided by lipid content so as to eliminate effects of differences in fatty tissues among different plankton and ﬁsh species, increased by up to two orders of magnitude between short-lived and longer-lived specimens (Harding et al. 1997). Perhaps duration of exposure to these contaminants leads to increased body burdens.

Interactions and combinations of factors None of the factors just reviewed seem dominant; in fact, all the supposed relationships either have notable exceptions, or are marked by substantial variation around the trend lines. It seems evident that body burdens of chlorinated hydrocarbons are determined in different circumstances by idiosyncratic combinations of factors for different

187

CHLORINATED HYDROCARBONS

100 CI

CI

CI CI

CI

80 CI

CI

CI

60 Fish Worm Shrimp

40 20

% of initial concentration

0 100 CI

CI CI

80

CI

CI

CI CI

CI CI

CI

60 40 20 0 100 CI CI

80

CI

CI CI

Figure 8.8 Diagrammatic representation of the time course of elimination of six different polychlorinated biphenyl (PCB) congeners in a ﬁsh, a worm, and a shrimp. Adapted from Goerke and Weber (2001).

60

CI CI

40

CI

20 0 0

species of organisms, or even different subpopulations of a given species. One telling instance of an arcane and circumstantial combination of speciﬁc factors is the case of chlorinated hydrocarbons in the tissues of killer whales in coastal waters off British Columbia and Washington. There are two resident groups, one southern, one northern: the southern whales, regardless of sex, are, on average, more exposed to the industrialized Vancouver/Seattle metropolitan areas, and carry greater PCB concentrations than the northern group of residents.20 There is also another group consisting of transient killer whales that have a far larger geographic 20

CI

CI CI

Comparison with other cetaceans is possible in Ross et al. (2000), who noted that male and female beluga whales in the Gulf of St Lawrence, Canada, contained 78.8 and 29.6 mg kg−1 lipid weight PCBs, respectively. Measurements in the Mediterranean striped dolphin yielded 282 mg kg−1 lipid weight PCBs. The values for Paciﬁc Northwest cetaceans span similar ranges.

10

20 30 0 10 Elimination time (weeks)

20

30

range. The transient whales also have large PCB burdens. It is not known where the transients acquire their PBC burden, but, for unknown reasons, transients feed exclusively on seals and porpoises, but not on ﬁsh, in contrast to the resident whales (Table 8.4). Transient whales are therefore at least one food web step higher than their resident cousins, and their body burdens of PCBs are higher than those in resident whales, possibly providing evidence of the effect of magniﬁcation associated with the additional step. There was also a wide disparity in PCB content of males and females. This disparity occurs because females during their reproductive life pass on their PCB burdens to their young, to some degree while the fetus is in the placenta (Tanabe et al. 1982), but more so through the lipids contained in their rather fat-rich milk. This is a common pattern for mammals in general, including humans (Fuller & Hobson 1986; Mes 1986). This

188

CHAPTER 8

Frequency distribution (f )

14

INVERTEBRATES (53 ± 9)

12 10 8 6 4 2 0 0

Frequency distribution (f )

14

20

40

80

160

320

640

1,280 2,560

40

80

160

320

640

1,280 2,560

FISH (34 ± 5)

12 10 8 6 4 2 0

0

14 Frequency distribution (f )

10

10

20

BIRDS (210 ± 68)

12 10 8 6 4 2 0 0

10

20

40 80 320 160 Bioaccumulation factor

640

example conveys the notion of just how relatively arcane details of life histories, geography, and food preferences, with consequent intervention of multiple factors can alter the burden of contaminants carried by a species. Another example of multiple factors at work is the experimental demonstration by Rubenstein et al. (1984) that a coastal ﬁsh, scup, can acquire PCB loads by direct uptake from contaminated

1,280 2,560

Figure 8.9 Frequency distribution of estimates of bioaccumulation of PCBs for invertebrates, ﬁsh, and birds. The bioaccumulation factor is the ratio of PCB concentration in the organisms relative to the concentration in their food. Data from Harding and Addison (1986), Peakall (1986), and Borgå et al. (2001).

sediments (Fig. 8.10)—presumably through lipid membranes in their gills and storage in tissue lipids—or from eating contaminated prey. The trophic pathway led to somewhat larger PCB burdens in tissues, and, moreover, the concentrations acquired via the two different pathways seemed additive. In a similar experiment with brittle stars, Gunnarsson and Sköld (1999) found PCB concentrations in brittle stars higher than

189

CHLORINATED HYDROCARBONS

Table 8.4 Concentrations (mean ± s.e.) of polychlorinated biphenyls (PCBs) in male and female killer whales that were part of northern or southern resident populations, or transient populations, in waters off British Columbia, Washington, and Alaska. Data from Ross et al. (2000), compiled from many sources. Residents, northern group

Residents, southern group

Transients

No. of individuals in group

212

89

219

Preferred food

Fisha

PCB concentration (µg kg−1 lipid wet wt) In males In females

37.4 ± 6.1 9.3 ± 2.8

Marine mammalsb

146.3 ± 32.7 55.4 ± 19.3

251.2 ± 54.7 58.8 ± 20.6

a

96% of diet is salmon.

b53% of diet is harbor seals, 13% Steller sea lions, 12% Dall’s porpoises, and 11% harbor porpoises.

Exposure to sediments

Figure 8.10 Concentrations of polychlorinated biphenyls (PCBs) in scup, a coastal ﬁsh, under different experimental conditions: ﬁrst, exposed only to sediments and unfed for 39 days, then fed until day 60. See text for discussion. From Rubenstein et al. 1984.

PCB conc. (µg g−1 wet wt)

1.0

Fed contaminated worms

Contaminated sediments Uncontaminated sediments No sediments

0.8 0.6

Fed contaminated worms Fed uncontaminated worms

0.4 0.2 0.0 0

10

could be accounted for by passive incorporation of PCBs into the lipids. Some additional uptake took place, probably by selective feeding by the brittle stars on contaminated particles resuspended from surface sediments.21 There is no simple explanation as to the factors that determine body burdens of chlorinated hydrocarbons. Even the commonly argued biomagniﬁcation effects are of ambiguous interpre21

Feeding

Reviews of efforts to sort out mechanisms determining concentrations of toxic compounds in organisms can be found in Connell (1990), Chapman (1997), and Sharpe and Mackay (2000). Those looking for a more biogeochemical treatment of the subject will ﬁnd Macdonald et al. (2002) of interest.

20

30 40 Time (days)

50

Fed uncontaminated worms 60

tation. Higher concentrations of contaminants in species near the top of food webs could be a result of food chain biomagniﬁcation, or could come about through other means. For instance, top predators could contain more fats, and hence accumulate more of the fat-soluble contaminants. In most cases, there are probably several mechanisms at play. The variety of biological species involved, plus differences in age, ability to assimilate, store, or depurate the many different hydrocarbon isomers, results in a bewildering diversity of responses by different species to exposure to chlorinated hydrocarbons. The most that can be safely said is

190

CHAPTER 8

that concentrations of chlorinated hydrocarbons in exposed animals are controlled to some degree by the supply of speciﬁc hydrocarbon isomers, food type, isomer-speciﬁc depuration ability, fat content, and age of the different species of animals involved. The effects of these variables usually occur in some combination, and are modiﬁed by details of the biology of the species and their geographic distribution.

Effects of chlorinated hydrocarbons22 Concerns about the effects of chlorinated compounds in coastal and other environments have been voiced for decades in the professional literature.23 But it was not until Rachel Carson’s 1963 book, Silent Spring, that the public and political sectors took notice of these issues. This one crucible publication wrested attention away from the notion of DDT as the magic treatment for pest control. Silent Spring made people aware that these chemicals were a cause of ﬁsh and bird kills, decimated beneﬁcial species, and led to build-up of resistance in the pest species. Carson emphasized the human health risks, citing evidence that suggested that DDT and like compounds may be connected to human nervous disorders, tumors, leukemia, and even deaths; that the compounds readily passed and were dispersed through atmosphere, soils, water, and wastewater treatment plants; and that contamination was widespread. At a deeper level, Carson’s book made the public aware that human beings were part of ecosystems whose parts were coupled in complex ways, and that these elements and linkages had a great deal to do with human welfare on earth. It could be argued that Silent Spring, whatever faults of exaggeration it may have carried, in actual fact 22

McDowell Capuzzo (1996) reviews the many physiological, developmental, and reproductive effects of burdens of chlorinated hydrocarbons. 23 Effects of DDT on estuarine ﬁsh and blue crabs, and on freshwater ﬁsh, were reported from the mid-1940s to the 1960s (Nimmo et al. 1987). Hazards that pesticides posed to coastal environments, including the disappearance of bald eagles in Florida, and of osprey and peregrine falcons in North America, sounded alarms about the contamination of coastal environments by chlorinated hydrocarbons in the 1950s to the 1960s (Risebrough 1989).

brought ecology to the public as a meaningful and important aspect of science. An explosion of studies on the effects of agricultural and industrial compounds followed in the wake of Silent Spring. This plethora of publications continues at even faster rates today. The published number and range of effects of the DDT family, and of PCBs, are legion.24 Certainly the most publicized effect of the DDT family of chlorinated hydrocarbons was the thinning of bird eggshells after the 1940s, when DDT became broadly used (Risebrough 1989). This seemingly minor direct effect had serious indirect consequences because the parents inadvertently crushed the affected eggs during incubation. Actually, DDE was speciﬁcally responsible for the thinning of eggshells. Shell-thinning affected double-crested cormorants in California, and was responsible for the near disappearance of brown pelicans in the West Coast of the USA, of osprey from a large portion of the world, and of peregrine falcons from North America and Europe. PCBs have been claimed to reduce growth and photosynthesis in algae and plants (Mahanty 1986). In invertebrates, ﬁsh, and birds, PCBs create sublethal conditions involving lower reproduction, malformations,25 altered liver, thyroid, and circulatory functions, suppression of immunity,26 increased nutritional deﬁciencies, and more (Harding & Addison 1986; Peakall 1986). PCBs, in contrast to DDT and DDE, do not affect shell thickness or have behavioral effects in birds (Peakall 1986). PCBs are not sufﬁciently toxic to pose hazards to bird populations (Peakall 1986), but there may be impairment of various reproductive processes in mammals (Fuller & Hobson 24

Earlier work is summarized in Kimbrough (1980), Waid (1986), and Crine (1986). 25 Tumors, lesions, reproductive impairments, and ﬁn abnormalities are induced by PCBs within ﬁsh in the vicinity of discharges of waste water near-shore. Surprisingly, in spite of these symptoms, there is little evidence that there are consequent effects on the population dynamics, or abundance of the affected species (Risebrough 1989). 26 High concentrations of chlorinated hydrocarbons in tissues might lower avian immune responses: the incidence of parasitic nematodes in the digestive systems of glaucous gulls was associated with higher liver concentrations of DDT and PCBs (Sagerup et al. 2000).

191

CHLORINATED HYDROCARBONS

27

About one-quarter of births in Californian sea lions were premature, a condition that was connected to exceptionally high DDT and PCB concentrations in the body fat of the mothers (DeLong et al. 1973). Reijnders (1986) reported increases in mortality of juveniles, and lowered pup production in harbor seals in the Wadden Sea, apparently associated with higher internal concentrations of PCBs. No evidence was provided as to just how these demographic alterations affected the population at large. 28 In this example, there were also increased mercury concentrations relative to those in the control site. Although it would have been preferable, for the purposes of this chapter, to have only elevated PCB concentrations, it mattered little, as neither the PCBs nor the mercury contamination seemed to have created a clear set of biological responses.

1:1

1,000 B 100 Values in control site

1986).27 There was a lowered reproductive effort of mussels in New Bedford Harbor (see Table 8.1), but there have been no demonstrable population-level effects that could be associated with lower reproduction in the population within New Bedford Harbor. The lack of local population-level effects may be a measure of the usual extraordinarily high mortality of larval shellﬁsh, the large input of larvae from surrounding areas, and the lack of direct mortality effects of PCBs on adult mussels, rather than evidence of no local effects. The relative weak population effects of PCBs, plus features of reproduction and recruitment in shellﬁsh, thus make it quite difﬁcult to ﬁnd measurable effects of PCBs on the abundance of settled mussels and other shellﬁsh, even in this highly contaminated site. Reasonably comprehensive comparisons of microbial, plant, and animal abundance and activities in salt marsh sites exposed to high and low levels of PCB contamination (Fig. 8.11), showed no evident effects, in spite of the substantial difference in PCB concentrations present in the sediments of the two sites.28 In these examples and other instances, convincing demonstrations of population-level effects of PCBs on microbes, plants, and invertebrates in the ﬁeld are rare. We thus have the odd situation that we know that exposure to PCBs could foster a plethora of sublethal effects (McDowell Capuzzo 1996), but it is hard to ﬁnd documentation that these effects result in consequent substantial population-level impacts (Risebrough 1989). One paper on the effects on striped bass characteristically concluded “. . . we found no relationships between . . . exposure [to PCBs] and any measure of striped

L S

10 1 0.1 0.01 0.01

F

G

C

P

H PCB

E

1 10 100 0.1 Values in contaminated site

1,000

Figure 8.11 Comparison of the differences in contaminants [polychlorinated biphenyls (PCBs) and mercury (H)] and biological responses by cordgrass [stem density (S, in no. per 0.0625 m2), canopy height (C, in cm), glutathione content (G, in pmol mg−1 wet wt), and rate of photosynthesis (P, in µmol CO2 m−2 s−1)], fungi [biomass (F, in mg ergosterol per g dead shoots)], and grass shrimp [length (L, in mm), brood size (B, in number of eggs), and mass of individual eggs (E, in mg)], between a highly contaminated and a non-contaminated salt marsh environment in Georgia, USA. Data from Wall et al. (2001).

bass abundance or reproduction” (Barnthouse et al. 2003).

Distribution and time courses of chlorinated hydrocarbons PCBs and other chlorinated compounds have been found widely distributed throughout the entire world (Atlas et al. 1986). About 1,000,000 tons were produced before 1976, and perhaps 100,000 tons have managed to enter natural environments (Axelman & Broman 2001).29 PCBs were manufactured in a few places within the Northern Hemisphere, and the latitudinal distribution of PCB concentration in marine surface waters still 29

Some 610,000 tons of PCBs were produced up to 1976 in the USA, 67,000 tons in the UK, about 60,000 tons in Japan, 100,000 tons in the former Soviet Union, and 8,000 tons in China; a few other countries (e.g. Austria, Czechoslovakia) produced smaller amounts (Brinkman & de Kok 1980; Axelman & Broman 2001).

192

CHAPTER 8

PCB concentration (ng l–1)

1.0 0.8 0.6 0.4 0.2 0 60° 50° 40° 30° 20° 10° N

0°

10° 20° 30° 40° 50° 60° 70° S Latitude

reﬂects that geographic origin (Fig. 8.12). Concentrations of PCBs in the surface sea water of the temperate latitudes range about one order of magnitude higher than in sea water at similar latitudes in the Southern Hemisphere. Nonetheless, there has been broad north–south, air–water, shallow–deep water transports and exchanges, and PCBs have managed to spread to every part of the earth.30 PCBs have been found in astoundingly diverse places around the world. Air in the mid-Paciﬁc and Antarctica, deep sea water (Harvey et al. 1973), Antarctic sediments (Montone et al. 2001), Paciﬁc coral reefs (Miao et al. 2000), Australian estuaries (Shaw & Connell 1980), marine mammals (Ross et al. 2000), and migratory oceanic albatrosses (Guruge et al. 2001) all bear some PCB burden. Such observations are among the many widespread measurements that betray surprisingly rapid dispersal of PCBs.31 Atmospheric and 30

Stout (1986) raises a word of caution before making too many comparisons of measurements of PCB concentrations. Methods to measure these compounds have made great strides across the last half of the 20th century. In most cases, different numbers and types of congeners are reported, so that total PCB values might refer to quite different chemical assemblages. Concentrations are variously reported on the basis of wet weights, dry weights, and per lipid weights. Such changes make comparisons difﬁcult, and reduce the number of data sets that can deﬁne the time courses and global distribution of these compounds. 31 Water at great depths in the oceans is renewed only at time scales of thousands of years. It was somewhat of a surprise to ﬁnd that PCBs (compounds synthesized only in the last half of the 20th century)

Figure 8.12 Latitudinal distribution of concentrations of polychlorinated biphenyls (PCBs) in ocean surface waters in the western Paciﬁc and eastern Indian Ocean, 1975–1982. From Tanabe and Tatsukawa (1986).

river transport of PCB-bearing particles are the most likely pathways behind the rapid spread of these compounds from land toward coastal environments. As a result, PCB concentrations in water, sediments, and organisms are higher in the near-shore than offshore. Coastal environments have received more than their share of PCBs: perhaps up to 6% of the world’s industrial production of these compounds may be found in sediments between the coast and the shelf slopes, with larger shares of the more chlorinated congeners (Jönsson et al. 2003).32 With the sharp worldwide reductions in use of DDT and PCBs that have taken place during recent decades, residues in sediments and concentrations in animals have diminished in many coastal areas. Populations of raptors, certainly the most sensitive indicators of the presence of the DDT family of chlorinated hydrocarbons, have recovered. could be measured in deep sea water and sediments (Harvey et al. 1973; Harvey & Steinhauer 1976). Such deep-sea concentrations could not reach the deep sea by transport as dissolved materials. PCBs manage to reach the greatest ocean depths by traveling adsorbed onto particles—mostly zooplankton fecal pellets—that fall through the marine water column to the sea ﬂoor in a matter of weeks (Elder & Fowler 1977). 32 Tanabe and Tatsukawa (1986) estimate that the open sea may hold about 20% of the world’s PCB production. They claim that roughly 12% of the PCBs used in the Seto/inland area of Japan could be found in terrestrial and coastal environments. If this number is representative for other terrestrial and coastal areas, then perhaps somewhat more than 30% of the world PCB production has been released to natural environments.

CHLORINATED HYDROCARBONS

Ospreys, for example, have returned to previous abundance in Chesapeake Bay (Ambrose 2001) and elsewhere in the USA. In a very real sense, there is substantial recovery from the environmental alterations caused by these chlorinated hydrocarbons. Many studies have shown recent decreases in the concentration of PCBs in water, sediments, and organisms in coastal environments.33 These reductions no doubt derive from the ceasing of further production of PCBs, reduced disposal of these substances, and the degradation and burial of some fraction of the residues in coastal systems. One example of the evidence for recent decadalscale decreases comes from concentrations of PCBs in sediment cores taken from the northern coast of the Gulf of Mexico (Santschi et al. 2001). Much like the recent decreases in polycyclic aromatic hydrocarbons (see Fig. 7.13), concentrations of PCBs have become lower in sediments of many coastal areas. One interpretation of such vertical proﬁles is that the delivery of PCBs to these sediments peaked in the 1960s and 1970s, and either decreased markedly since or remained variably similar since then.34 Similarly, chlorinated hydrocarbon concentrations (PCBs as well as DDTs) in coastal organisms have decreased after the 1970s. The concentration of PCBs in seaweeds in the Baltic Sea decreased by an order of magnitude between 1983 and 1989 (Fig. 8.13 top). Concentrations of PCBs and DDTs in Baltic Sea herring decreased between 1972 and

33

Concentrations of PCBs in oceanic air and water have not shown decreases that parallel the decreases in coastal areas during the 1980s and 1990s. This may be a result of the dispersion of PCBs from the vicinity of their points of release, and redistribution into the more remote areas of the seas. Compare data in tables 1 and 2 in Tanabe and Tatsukawa (1986) with those in tables 3 and 4 in Axelman and Broman (2001), for example. 34 Clearly, other processes are also involved in establishing the proﬁles. There is certainly some degree of microbial degradation, which is likely more intense nearer the surface where there is more organic material, and where oxygen might be somewhat more available. In addition, there is a slow release of DDTs and PCBs from sediments to the overlying water (DiToro & Horzempa 1982; Gunnarsson and Sköld 1999; Zeng et al. 1999) by desorption and resuspension of particles bearing adsorbed hydrocarbons. Such processes certainly modify the proﬁles in sediments, but do not altogether erase the chronicle of contamination history represented in proﬁles of chlorinated hydrocarbons.

193

1988 (Fig. 8.13 bottom); PCBs and DDTs in other ﬁsh in the Baltic were low enough in 1998–1999 to lie below public health standards set by the World Health Organization (Roots 2001). Reductions in chlorinated hydrocarbons were also found in mussels off the coast of California (Fig. 8.14), paralleling decreased concentrations of chlorinated compounds in wastewater efﬂuents. There were also clear distance effects: mussels near the Los Angeles sewage outfall contained the largest concentrations; mussels growing farther away, where there were lower releases, contained lower concentrations. Many other examples exist that show decreases since the peak concentrations in PCB and DDT that occurred in the early 1970s. The compilation of information in Table 8.5 shows that lowering of PCB concentrations by about 4–10% per year was widespread in coastal organisms through the 1980s. Apparently the decreases have slowed, or reached asymptotes following the 1980s. The National Status and Trends data collected by the US National Oceanic and Atmospheric Agency show considerably lower concentrations of PCBs than during the 1970s, but no clear further decrease in mussels is evident during the last decade of the 20th century. This long latency may apply to many species: concentrations of PCBs and DDTS in Antarctic minke whales, for instance, have remained at about the same level between 1984 and 1999 (Tanabe 2002). Even though there is decisive evidence that inputs of chlorinated hydrocarbons have decreased recently, we are still faced with an almost global low level of these compounds in most environments. A low but long-term residuum of chlorinated hydrocarbons is present in organisms in most places in the world. The persistence of the degradation products guarantee that there will be residues for a long time in certain reservoirs such as soils and aquatic sediments, and these reservoirs will slowly release DDT and PCB residues to the general environment. It is not certain that these low-concentration residues will not have long-term biological consequences. Many researchers are concerned, for instance, that the low levels of PCBs, DDTs, and related compounds might have consequential effects,

194

CHAPTER 8

Enteromorpha intestinalis Cladophora glomerata Fucus vesiculosus Other seaweed spp. PCB and DDT concentration (µg kg–1 muscle wet wt)

PCB concentration (ng g–1 dry wt)

10,000

1,000

100

10

1 1983 1984 1985 1986 1987 1988 1989 1990 Year

30 PCB DDT

24 18 12 6 0 1970

1974

1978 1982 Year

1986

1990

Figure 8.13 Left: concentrations of polychlorinated biphenyls (PCBs) in different species of seaweeds collected on the northern Baltic Sea, 1983 –1989 (data from Talvari et al. 1990, 1992, in Roots 1996). Right: concentrations of PCBs and DDTs in herring collected from the northern Baltic Sea, 1972–1988 (data from Roots 1996, original data from A. Bignert).

2.5

17 Sewer outfall (Royal Palms)

16

14

2.0

13 12 11 1.5

10 9 8

Coastal City 1.0 (Corona del Mar)

7 6

Coastal City (La Jolla) 0.5

5 4 3

Pristine Island Site (Anacapa Island) 0

2 1 0

1971 1972 1973 1974 1975 1976 1977 1978 1979 1980 Year

DDE concentration (µg g−1 dry wt) ( )

PCB concentration (µg g−1 dry wt) ( )

15

Figure 8.14 Concentrations of polychlorinated biphenyls (PCBs) and DDT in mussels at the Los Angeles sewer outfall, and of PCBs in mussels at three other coastal southern California sites, 1971–1980. From Stout (1986), compiled from various sources.

195

CHLORINATED HYDROCARBONS

Table 8.5 Approximate rates of change in concentrations of polychlorinated biphenyls (PCBs) in various organisms, relative to the peak concentration reported in the time course. % change in concentration

Organism

Period

California mussel

1971–1981

−93

Oyster

1969–1976

−80

Common tern

1971–1981 1973–78

−47 −50

Menhaden

1972–1977 1972–1977 1972–1977

Dover sole

1972–1980

Striped bass

1978–1979 1979–1980 1980–1981

Mussels

1971–1980 1971–1979

−55.6 −96.4

Brown pelican eggs

1969–1982 1974–1980

−80 −50

Mussels

1976–1977 vs. 1991–1997

Black-legged kittiwakesa

% change per year

Site

Source

Los Angeles outfall

Young et al. 1988

Escambia Bay, FL

Wilson & Forester 1978

−4.2 −8.3

Buzzards Bay, MA Monomoy Island, MA

Nisbet & Reynolds 1984 Nisbet & Reynolds 1984

−60 −58 −43

−10 −9.7 −7.1

Middle Atlantic coast Chesapeake Bay Gulf of Mexico

Stout et al. 1981 Stout et al. 1981 Stout et al. 1981

−95

−11.9

−9.3 −10

Young et al. 1988

−60 −20 −30

Hudson River Hudson River Hudson River

Stout 1986 Stout 1986 Stout 1986

−5.6 −10.7

Corona del Mar, CA La Jolla, CA

Stout 1986 Stout 1986

−5.7 −7.1

California South Carolina

Stout 1986 Stout 1986

−54.7

−3.4

Gulf of Maine

Farrington et al. 1982; Chase et al. 2001

1975–1998

−90

−3.9

Canadian Arctic coast

Braune et al. 2001

Northern fulmarsa

1975–1998

−63

−2.7

Canadian Arctic coast

Braune et al. 2001

Thick-billed murresa

1975–1998

−57

−2.3

Canadian Arctic coast

Braune et al. 2001

a

Similar reductions also found for DDTs in these sea birds.

and, in the long term, could be disruptive to complex physiological functions (Lye et al. 1999; Tanabe 2002). Such endocrine disruptions could lower resistance to diseases, alter sexual hormone levels, break down steroids, and create other physiological effects in many species. There is a need for ﬁeld studies to determine if these suppositions have, or may become, a problem in natural environments.

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PCB accumulation in the sea bass food web from the Seine estuary (eastern English Channel). Mar. Pollut. Bull. 43:242–255. Long, E. R., D. D. MacDonald, S. L. Smith, and F. D. Calder. 1995. Incidence of adverse biological effects within ranges of chemical concentrations in marine and estuarine sediments. Environ. Manage. 19:81–97. Lundgren, K., M. Tyskind, R. Ishaq, D. Broman, and B. van Bavel. 2002. Polychlorinated naphthalene levels, distribution, and biomagniﬁcation in a benthic food chain in the Baltic Sea. Environ. Sci. Technol. 36:5006– 5013. Lye, C. M., C. L. J. Frid, M. E. Gill, D. W. Cooper, and D. M. Jones. 1999. Estrogenic alkylphenols in ﬁsh tissues, sediments, and waters from the U. K. Tyne and Tees estuaries. Environ. Sci. Technol. 33:1009–1014. McClelland, J. W., and I. Valiela. 1998. Linking nitrogen in estuarine producers to land-derived sources. Limnol. Oceanogr. 43:577–585. McClelland, J. W., I. Valiela, and R. H. Michener. 1997. Nitrogen stable isotope signatures in estuarine food webs: A record of increasing urbanization in coastal watersheds. Limnol. Oceanogr. 42:930–937. Macdonald, R., D. Mackay, and B. Hickie. 2002. A new approach suggests that phenomena, such as bioconcentration, biomagniﬁcation, and bioaccumulation, result from two fundamental processes. Environ. Sci. Technol. 36:457A–462A. McDowell Capuzzo, J. 1996. Biological effects of contaminants on shellﬁsh populations in coastal habitats: A case history of New Bedford, Massachusetts. Pp. 457–466 in Sherman, K., N. J. Jaworski, and T. J. Smayda (eds). The Northeast Shelf Ecosystem: Assessment, Sustainability, and Management. Blackwell Science, Oxford, 564 pp. Mackay, D. 1979. Finding fugacity feasible. Environ. Sci. Technol. 13:1218–1223. Mackay, D. 1982. Correlation of bioconcentration factors. Environ. Sci. Technol. 16:274–278. Mackay, D., and S. Paterson. 1982. Fugacity revisited. Environ. Sci. Technol. 16:654A–660A. Mahanty, H. K. 1986. Polychlorinated biphenyls: Accumulation and effects on plants. Pp. 1–8 in Waid, J. S. (ed.). PCBs and the Environment, Vol. II. CRC Press, Boca Raton, FL, 191 pp. Mercaldo-Allen, R., C. A. Kuropat, R. A. Craig, and G. Sennefelder. 1994. PCB and metal concentrations in American lobsters from the Acushnet River estuary and Long Island Sound. Bull. Environ. Contam. Toxicol. 53:820–827. Mes, J. 1986. PCBs in human populations. Pp. 39–61 in Waid, J. S. (ed.). PCBs and the Environment, Vol. III. CRC Press, Boca Raton, FL, 272 pp. Miao, X.-S., C. Swenson, L. A. Woodward, and Q. X. Li. 2000. Distribution of polychlorinated biphenyls in marine species from French Frigate Shoals, North Paciﬁc Ocean. Sci. Tot. Environ. 257:17–28. Montone, R. C., S. Taniguchi, and R. R. Weber. 2001. Polychlorinated biphenyls in marine sediments of Admiralty

Bay, King George Island, Antarctica. Mar. Pollut. Bull. 42:611–614. Moore, M. R., W. Vetter, C. Gaus, G. R. Shaw, and J. F. Müller. 2002. Trace organic compounds in the marine environment. Mar. Pollut. Bull. 45:62–68. Muir, D. C. G., K. Koczanski, B. Rosenberg, and P. Beland. 1996. Persistent organochlorines in beluga whales (Delphinapterus leucas) from the St. Lawrence River Estuary. 2. Temporal trends, 1982–1994. Environ. Pollut. 93:235– 245. Nacci, D., and 8 others. 1999. Adaptations of wild populations of the estuarine ﬁsh Fundulus heteroclitus to persistent environmental contaminants. Mar. Biol. 134:9– 17. Nelson, W. G., and 8 others. 1996. New Bedford Harbor Long-term Monitoring Assessment Report: Baseline Sampling. US Environmental Protection Agency report EPA/600/R–96/097. National Health and Environmental Effects Research Laboratory, Atlantic Ecology Division, Narragansett, RI. Nimmo, D. R., D. L. Coppage, Q. H. Pickering, and D. H. Hansen. 1987. Assessing the toxicity of pesticides to aquatic organisms. Pp. 49–69 in Marco, G. J., R. M. Hollingworth, and W. Durham (eds). Silent Spring Revisited. American Chemical Society, Washington, DC, 214 pp. Nisbet, I. C. T., and L. M. Reynolds 1984. Organochlorine residues in common terns and associated estuarine organisms, Massachusetts, USA, 1971–81. Mar. Environ. Res. 11:33–66. Oliver, B. G., and A. J. Niimi. 1988. Trophodynamic analysis of polychlorinated biphenyl congeners and other chlorinated hydrocarbons in the Lake Ontario ecosystem. Environ. Sci. Technol. 22:388–397. Olsson, A., K. Valters, and S. Burreau. 2000. Concentrations of organochlorine substances in relation to ﬁsh size and trophic position: A study on perch (Perca ﬂuviatilis L.). Environ. Sci. Technol. 34:4878–4886. Paterson, M. J., D. C. C. Muir, B. Rosenberg, E. J. Fee, C. Anema, and W. Franzin. 1998. Does lake size affect concentration of atmospherically derived polychlorinated biphenyls in water sediment, zooplankton, and ﬁsh? Can. J. Fish. Aquat. Sci. 55:544–553. Peakall, D. B. 1986. Accumulation and effects on birds. Pp. 31–47 in Waid, J. S. (ed.). PCBs and the Environment, Vol II. CRC Press, Boca Raton, FL, 191 pp. Pesch, C. E., and J. Garber. 2001. Historical analysis, a valuable tool in community-based environmental protection. Mar. Pollut. Bull. 42:339–349. Peterson, B. J., and B. Fry. 1987. Stable isotopes in ecosystem studies. Ann. Rev. Ecol. Syst. 18:293–320. Pierce, R. W., and J. T. Turner. 1994a. Plankton studies in Buzzards Bay, Massachusetts, U.S.A., 3. Dinoﬂagellates, 1987 to 1988. Mar. Ecol. Prog. Ser. 112:225– 234. Pierce, R. W., and J. T. Turner. 1994b. Plankton studies in Buzzards Bay, Massachusetts, U.S.A., 4. Tintinnids, 1987 to 1988. Mar. Ecol. Prog. Ser. 112:235–240.

CHLORINATED HYDROCARBONS

Pruell, R. J., and 6 others. 1990. Geochemical study of sediment contamination in New Bedford Harbor, Massachusetts. Mar. Environ. Res. 29:77–101. Randall, R. C., D. R. Young, H. Lee, and S. F. Echols, 1998. Lipid methodology and pollutant normalization relationships for neutral nonpolar organic pollutants. Environ. Toxicol. Chem. 17:788–791. Reijnders, P. J. H. 1986. Reproductive failure in common seals feeding on ﬁsh from polluted coastal waters. Nature 324:456–457. Risebrough, R. W. 1986. Pesticides and bird populations. Curr. Ornithol. 3:397–427. Risebrough, R. W. 1989. Chemical changes in the marine environment: Origins and concepts. Pp. 11–33 in Albaigés, J. (ed.). Marine Pollution. Hemisphere Publishing Corp., New York, 365 pp. Risebrough, R. W., F. C. Sibley, and M. N. Kirven. 1971. Reproductive failure of the brown pelican on Ancapa Island in 1969. Am. Birds 25:8–9. Roots, O. 1996. Toxic Chlororganic Compounds in the Ecosystem of the Baltic Sea. Eesti Vabariigi Keskkonnaministeerium Info-ja Tehnokeskus, Tallinn. Roots, O. 2001. Halogenated environmental contaminants in ﬁsh from Estonian coastal waters. Chemosphere 43:623–632. Ross, P. S., G. M. Ellis, M. G. Ikonomou, L. G. BarrettLennards, and R. F. Addison. 2000. High PCB concentrations in free-ranging Paciﬁc killer whales, Orcinus orca: Effects of age, sex and dietary preference. Mar. Pollut. Bull. 40:504–515. Rubenstein, N., W. T. Gilliam, and N. R. Gregory. 1984. Dietary accumulation of PCBs from a contaminated sediment source by a demersal ﬁsh (Leiostomus xanthurus). Aquat. Toxicol. 5:331–342. Ruus, A., K. I. Ugland, O. Espeland, and J. U. Skaare. 1999. Organochlorine contaminants in a local marine food chain from Jarfjord, Northern Norway. Mar. Environ. Res. 48:131–146. Sagerup, K., E. O. Henriksen, A. Skorping, J. U. Skaare, and G. W. Gabrielsen. 2000. Intensity of parasitic nematodes increases with organochlorine levels in the glaucous gull. J. Appl. Ecol. 37:532–539. Santschi, P. H., B. J. Presley, T. L. Wade, B. Garcia Romero, and M. Baskaran. 2001. Historical contamination of PAHs, PCBs, DDTs, and heavy metals in Mississippi River delta, Galveston Bay, and Tampa Bay sediment cores. Mar. Environ. Res. 52:51–79. Sawhney, B. L. 1986. Chemistry and properties of PCBs in relation to environmental effects. Pp. 47–64 in Waid, J. S. (ed.). PCBs and the Environment, Vol. I. CRC Press, Boca Raton, FL, 228 pp. Schlezinger, J. J., and J. J. Stegeman. 2000. Induction of cytochrome P450 1A in the American eel by model halogenated and non-halogenated aryl hydrocarbon receptor agonists. Aquat. Toxicol. 50:375–386. Schulz-Bull, D. E., G. Petrick, and J. C. Duiker. 1991. Polychlorinated biphenyls in North Sea water. Mar. Chem. 36:364–384.

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Sharpe, S., and D. Mackay. 2000. A framework for evaluating bioaccumulation in food webs. Environ. Sci. Technol. 34:2373–2379. Shaw, G. R., and D. W. Connell. 1980. Polychlorinated biphenyls in the Brisbane River estuary, Australia. Mar. Pollut. Bull. 11:356–358. Shaw, G. R., and D. W. Connell. 1986. Factors controlling bioaccumulation of PCBs. Pp. 121–134 in Waid, J. S. (ed.). PCBs and the Environment, Vol. I. CRC Press, Boca Raton, FL, 228 pp. Sorci, J. J., J. D. Paulauskis, and T. E. Ford. 1999. 16S rRNA restriction fragment length polymorphism analysis of bacterial diversity as a biomarker of ecological health in polluted sediments from New Bedford Harbor, Massachusetts, USA. Mar. Pollut. Bull. 38:663–675. Stegeman, J. J., R. M. Smolowitz, and M. E. Hahn. 1991. Immunohistochemical localization of environmentally induced cytochrome P450IA1 in multiple organs of the marine teleost Stenotomus chrysops. Toxicol. Appl. Pharmacol. 110:486–504. Stegeman, J. J., F. Y. Teng, and E. A. Snowberger. 1987. Induced cytochrome P450 in winter ﬂounder (Pseudopleuronectes americanus) from coastal Massachusetts evaluated by catalytic assay and monoclonal antibody probes. Can. J. Fish. Aquat. Sci. 44:1270–1277. Stout, V. F. 1986. What is happening to PCBs? Elements of environmental management as illustrated by an analysis of PCB trends in terrestrial and aquatic habitats. Pp. 163–206 in Waid, J. S. (ed.). PCBs and the Environment, Vol. I. CRC Press, Boca Raton, FL, 228 pp. Stout, V. F., C. R. Houle, and F. L. Beezhold. 1981. A survey of chlorinated hydrocarbon residues in menhaden ﬁshery products. Mar. Fish. Rev. 43:1–13. Strandberg, J. D., J. Rosenﬁeld, I. K. Berzins, and C. L. Reinisch. 1998. Speciﬁc localization of polychlorinated biphenyls in clams (Mya arenaria) from environmentally impacted sites. Aquatic Toxic. 41:343–354. Summerhayes, C. P., J. P. Ellis, P. Stoffer, S. R. Briggs, and M. G. Fitzgerald. 1977. Fine-Grain Sediment and Industrial Waste Distribution and Dispersal in New Bedford Harbor and Western Buzzards Bay, Massachusetts. WHOI Technical Report WHOI-76-115. Woods Hole Oceanographic Institution (WHOI), Woods Hole, MA. Takayama, K., H. Miyata, M. Mimura, S. Ohta, and T. Kashimoto. 1991. Evaluation of biological effects of polychlorinated compounds found in contaminated oil responsible for the disease “Yusho”. Chemosphere 22: 537–546. Talvari, A., H. Heinlaid, H. Kukk, and H. Jankovski. 1992. Accumulation of chloroorganic compounds and hydrocarbons in Baltic seaweeds. Proc. Est. Acad. Sci. Ecol. 2:50–55. Tanabe, S. 2002. Contamination and toxic effects of persistent endocrine disrupters in marine mammals and birds. Mar. Pollut. Bull. 45:69–77. Tanabe, S., and R. Tatsukawa. 1986. Distribution, behavior, and load of PCBs in the oceans. Pp. 143–162 in Waid, J. S.

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(ed.). PCBs and the Environment, Vol. I. CRC Press, Boca Raton, FL, 228 pp. Tanabe, S., M. Kawano, and R. Tatsukawa. 1982. Chlorinated hydrocarbons in the Antarctic, western Paciﬁc and eastern Indian Oceans. Trans. Tokyo Univ. Fish. 5:97. Ten Berge, W. F., and M. Hillebrand. 1974. Organochlorine compounds in several marine organisms from the North Sea and the Dutch Wadden Sea. Neth. J. Sea Res. 8:361– 368. Thomann, R. V. 1981. Equilibrium fate of micronutrients in diverse aquatic food webs. Can. J. Fish. Aquat. Sci. 38:280. Tittlemier, S. A., A. T. Fisk, K. A. Hobson, and R. J. Norstrom. 2002. Examination of the bioaccumulation of halogenated dimethyl bipyrroles in an Arctic marine food web using stable nitrogen isotope analysis. Environ. Pollut. 116:85– 93. Van der Oost, R., R. H. Heida, and A. Opperhuizen. 1988. Polychlorinated biphenyl congeners in sediments, plankton, molluscs, crustaceans, and eels in a freshwater lake. Implication of using reference chemicals and indicator organisms in bioaccumulation studies. Arch. Environ. Contam. Toxicol. 17:721–729. Vander Zanden, M. J., and J. B. Rasmussen. 1999. Primary consumer δ13C and δ15N and the trophic position of aquatic consumers. Ecology 80:1394–1404. Waid, J. S. (ed.) 1986. PCBs and the Environment, Vol. III. CRC Press, Boca Raton, FL, 272 pp. Wall, V. D., J. J. Alberts, D. J. Moore, S. Y. Newell, M. Pattanayek, and S. C. Pennings. 2001. The effect of mer-

cury and PCBs on organisms from lower trophic levels of a Georgia salt marsh. Arch. Environ. Contam. Toxicol. 40:10–17. Weaver, G. 1983. PCB contamination in and around New Bedford, Mass. Environ. Sci. Technol. 18:22A–27A. Whitman, N. 1994. A Window Back: Photography in a Whaling Port. New Bedford Historical Society, New Bedford, MA, 175 pp. Wilson, A. J., and J. Forester. 1978. Persistence of Aroclor 1254 in a contaminated estuary. Bull. Environ. Contam. Toxicol. 19:637–640. Woodwell, G. M., C. F. Wurster, and P. A. Isaacson. 1967. DDT residues in an East coast estuary: A case of biological concentration of a persistent insecticide. Science 156:821–824. Young, D. R., and 7 others. 1980. Trophic structure and pollutant concentrations in marine ecosystems of Southern California. Cal. COFI Rep. 21:197. Young, D. R., R. W. Gossert, and T. C. Heesen. 1988. Persistence of chlorinated contamination in a California marine ecosystem. Pp. 33–41 in Wolfe, D. A., and T. P. O’Connor (eds). Oceanic Processes in Marine Pollution, Vol. 5. R. E. Krieger Publishing Co., Malabar, FL, 273 pp. Zaranko, D. T., R. W. Grifﬁths, and N. K. Kaushik. 1997. Biomagniﬁcation of polychlorinated biphenyls through a riverine food web. Environ. Toxic. Chem. 16:1463–1471. Zeng, E. Y., C. C. Yu, and K. Trang. 1999. In situ measurements of chlorinated hydrocarbons in the water column off the Palos Verdes Peninsula, California. Environ. Sci. Technol. 33:392–398.

Chapter 9 Metals

The ﬁrst Chisso Company manufacturing plant in Minamata. Photograph taken early in the 1910s. From http://www.fsinet.or.jp/∼soshisha/10tisiki/10_4_e.htm.

A case history: mercury contamination in Minamata Bay1 During the early 20th century, the population of Kyushu, one of the string of islands that makes 1

Material in this section is from Smith and Smith (1975), Nishigaki and Harada (1975), Jun Ui, Minamata disease (http://www.unu. edu/unuppress/unupbooks/uu35ie/uu35ieOc.htm#chapter%20% 20%04%20minamata%20disease), D. Allchin, The poisoning of Minamata (http://www1.umn.edu/ships/ethics/minamata.htm), TED Case Studies, Minamata disaster (http://www.american. edu/ted/minamata.htm), Harada (1995), and Tomiyasu et al. (2000).

up Japan, lived largely in small villages dependent on paddy rice farming and ﬁshing for subsistence, and salt production from shallow evaporation ponds for a meager external income. Shortly after the turn of the century, salt production became nationalized in Japan, and villages such as Minamata lost their only source of outside income. At about the same time Jun Noguchi, a young and ambitious engineer, was looking for a site to develop a carbide production plant that could make use of the hydroelectric power plant he had helped construct in the mountain hinterland above Minamata. The villagers saw an opportunity for

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<3

6 9 12 >15 Chisso factory Minamata City

Minamata River

income and jobs, and offered Noguchi the use of the salt ﬂats to build the plant, at no cost, as well as access to water and the port. Noguchi, as head of the enterprise that was to become the Nippon Chisso Company, duly built the plant in 1908 (see Chapter 9 frontispiece), and soon added production of ammonium sulfate to their list of products, for use as fertilizer (“chisso”, incidentally, means nitrogen). The company expanded greatly through the 1920s and into the 1940s, and became a major industrial power in the country, with powerful connections in the political and military sectors. The town of Minamata grew around the prosperity of the plant, and became economically and socially dependent on the company.

Figure 9.1 Map of the Yatushiro Sea area off the island of Kyushu, including the Minamata region. Black triangles show the sites where cats with Minamata disease symptoms were reported, white triangles show where there were ﬁsh kills. Contours show concentrations of mercury in surface sediments; the concentration intervals are shown as numbers along the shore, from < 3 to > 15 ppm dry weight. Adapted from Harada (1995) and Tomiyasu et al. (2000).

The Chisso enterprise was characterized by high technology and samurai toughness. Only the top graduates of the most prestigious universities were hired as scientiﬁc staff, and only very top school students were considered even for low-paying jobs such as assistants. Noguchi said workers should be treated “as cows and horses”, and indeed Chisso workers were expected to work in hazardous conditions for trivial pay. The company held little regard for the townspeople or their environment. During the entire history of development and production of the massive Chisso complex, all waste products were dumped into Minamata Bay with no treatment, and ﬂoating dead ﬁsh became obvious near Minamata (Fig. 9.1). The Minamata ﬁshermen protested

METALS

many times to the powerful company, and managed to receive some ﬁnancial compensation in 1926. That they managed to get additional compensation again in 1946 speaks of the evident damage to the ﬁshery. Chisso agreed to the second compensation only after demanding that no further compensation was to be sought by the ﬁshermen. The plant in Minamata was destroyed during the Second World War, but it rose from its ashes phoenix-like, and restarted production of fertilizers and plastics a few months after the war ended. By the 1950s and 1960s the company had regained its leading industrial position in Japan, and thoroughly dominated the town of Minamata (60% of Minamata taxes came from Chisso-related activities, and 60% of the jobs in Minamata were in the Chisso plant). During this period, Chisso perfected syntheses of acetaldehyde, as a principal compound essential for other syntheses. This compound was produced using mercury compounds as catalysts for reactions. Waste products were still released with no treatment into the bay, and more dead ﬁsh appeared in the water, and the ﬁshery was again sharply reduced. The ﬁshing community appealed to the company for a third time, and managed to be given modest compensation. About this time Minamata residents began to notice that, increasingly, neighborhood cats would engage in a frenzied dance and collapse, at times falling, or seeming to jump, into the bay (Fig. 9.1). The villagers were troubled about whether these suicide-prone “dancing” cats were an omen of bad things to come. Even more troubling was the undiagnosed illness that was affecting villagers from the area around Minamata Bay. First, four people in May 1956 were taken to the hospital with slurred speech, numbness, spasms, loss of consciousness, dementia, and coma leading to death. More patients with these symptoms and outcome appeared soon after. These symptoms had been seen sporadically by doctors before in the area. The only common factor among the victims was that their diet was mainly ﬁsh caught in Minamata Bay. Diagnosis of the disease was difﬁcult. The disease was not contagious, and was likely related

203

to some poison, and a ﬁsh diet. The contaminants in the sediments (and ﬁsh) of Minamata Bay were so diverse that identiﬁcation of the speciﬁc toxic material was a challenge. Moreover, Chisso was aggressively uncooperative in helping the search for a cause. After 2 years plus of work, medical teams concluded that mercury was the heavy metal responsible for the disease, which by now had acquired the name “Minamata disease”. The use of mercury by Chisso was kept as a trade secret by the company, but it became known that there were large concentrations of this metal in the waste sludge and sediments of Minamata Bay, in the ﬁsh, and in the victims of the disease (Table 9.1). Chisso denied responsibility throughout this period. The crisis peaked when similar symptoms began to appear in villagers in another area of the coast where the company had relocated its waste drainage outfalls, and people simply avoided all ﬁsh from the region. By 1959 the fear of the disease and loss of livelihood pushed the ﬁshermen to demand compensation from the company.2 After governmental intervention, Chisso agreed to pay the equivalent of US$27,800 (at 1959 exchange rates), if the causes of the disease were not further discussed. Money was to be paid to families of dead victims, in the sum of US$830 per victim, and US$278 for each living victim, as long as the money was understood to be an expression of condolence to the families, not an admission of guilt. The agreement that was signed included clauses that forbade further demands for compensation, even if it were proven that company wastes were the cause of the disease. About 100 patients were affected, and about 20 died. Several efforts to investigate the cause of Minamata disease foundered in the early 1960s, in part because of pressure from the industrial sector. A research group at Kumamoto University isolated methyl mercury from waste sludge, and suggested this was the toxin. Another group, supported by Chisso, exposed cats to waste sludge 2

The demands were simply for compensation for damages, rather than for a more fundamental reform to prevent the release of the toxic mercury. This speaks for a more ready acceptance of given conditions in Japan than might have been the case elsewhere.

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Tsuginori Hamamoto described the plight of his father, Sohachi, a ﬁsherman. Virtually overnight, Sohachi lost his ability to keep his balance, or stay aﬂoat once he had fallen off the boat. He could not put on his sandals, walk properly, or understand what others were saying to him. Once hardy and strongly self-willed, his condition quickly degenerated, and he was hospitalized on the fourth day. There, even tied to his bed with bandages, he “craze-danced,” and said words that were not words; he salivated; he convulsed . . . “Mother would look at Dad,” Tsuginori recalled, “and would just stand there—tears dropping from her eyes— looking dazed. Then we realized that the same symptoms were developing in Mother”. The father died within 7 weeks, the mother 9 years later. From D. Allchin, The poisoning of Minamata, (http://www1.umn.edu./ships/ethics/ minamata.htm)

and came to the same conclusion, but the company suppressed the results.3 Monitoring of mercury in the hair of villagers of the region also showed high mercury contents, but these data did not reach the public. Minamata disease temporarily disappeared from public attention. After 1960, however, more cases emerged on a broader geographic range. Children of victims were born with severe retardation and physical malformations. By 1963 research at the Public Health Service conclusively tied Minamata disease to poisoning by methyl mercury released by Chisso. In 1964, Minamata disease broke out along the Agano River in Niigata Prefecture, creating more pressure on the industry. In 1968, the Japanese government concluded “Minamata . . . is a disease of the central nervous system caused 3

Out of deep loyalty to the company, Dr Hajime Hosokawa, the head of the research group supported by Chisso, did not divulge his ﬁndings that the disease was caused by Chisso wastes until his deathbed (D. Allchin, The poisoning of Minamata, http://www1. umn.edu/ships/ethics/minamata.htm).

Table 9.1 Concentrations of mercury (µg Hg g−1 dry wt) in tissues of sea food from Minamata Bay, and in cats and humans from the area, and concentrations in cats and humans not exposed to mercury in their food. Data from D. Allchin, The poisoning of Minamata (http://www1.umn.edu/ ships/ethics/minamata.htm). Exposed to methyl mercury Foods Oyster Gray mullet Short-necked clam China fish Crab Consumers Cats Humans

Not exposed to methyl mercury

5.6 10.6 20.0 24.1 35.7 8–145.5 0.1–144.0

0.9–3.66 < 3.0

by methyl mercury . . . produced as a by-product . . . of manufactur[e of] acetaldehyde at Chisso Co. Ltd. . . .” Not until 1970 did a district court rule that Chisso compensate the original victims and support their families, and newer cases made ﬁnancial agreements with the company. As of March 2000, Chisso owed about 257 billion yen to patients and to the city of Minamata for the restoration work. In 1971 the bay was closed to all ﬁshing; ﬁshermen such as Jinichi Hamatsuki, a survivor of Minamata disease, recalled scavenging for sweet potatoes to survive after the closing of ﬁshing. The close link of mercury disease and acetaldehyde was made clear by epidemiological work. Some of the most telling information was evidence obtained, years after, by analyses of methyl mercury content in umbilical cords. People in many parts of Southeast Asia, from the Ainu in Hokkaido to the Okinawans, and certain parts of Indonesia and Malaysia, preserve umbilical cords of newborns, letting them dry in special wooden boxes (umbilical cord infusions are used as treatment for serious illnesses). Analyses of methyl mercury in a large set of umbilical cords of babies born from the 1930s through the 1970s showed a time course reasonably coincident with that of

Methyl mercury in umbilical cords (ppm dry wt) ( )

Figure 9.2 Time course of production of acetaldehyde in the Nisso plant, and concentration of methyl mercury in umbilical cords. Adapted from Harada (1995), original from S. Nishigaki and M. Harada.

5

4

4,000

3

3,000

2

2,000

1

1,000

1935

1940

the production of acetaldehyde by the Chisso Company (Fig. 9.2). In particular, as production declined in the 1960s, methyl mercury concentrations decreased steeply. Chisso stopped production of acetaldehyde in 1968, when a substitute method for synthesis was developed, and the release of mercury to the coastal waters stopped. In total, it was estimated that Chisso discharged 70–150 tons of mercury into Minamata Bay during its manufacture of acetaldehyde. Chisso still produces fertilizers, chemicals, and other plastic products in its Minamata plant. Some of the contaminated sediments in the bay were dredged, and the 58 ha area was ﬁlled in and was developed as a recreational facility (Fig. 9.3), at the cost of 48.5 billion yen over 14 years. A minimum of 12,617 people were recognized as patients affected by Minamata disease, of which at least 1,408 have died. More than US$6 billion has been paid in compensation, under the condition that the victims not press for further compensation in the future. Mercury poisoning and its effects went beyond illness and death. Victims of the poisoning were held in ridicule and segregated, called “freaks” even by their neighbors. Minimatans were tarred by the association with the very disease that threatened them. The stigma went beyond

1945

1950

1955 Year

1960

1965

1970

Average production of acetaldehyde (tons month–1) (–)

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Minimata: natives of Minimata were ridiculed elsewhere in Japan. Koichi Hara, for example, moved to Osaka as a youngster, but found there constant bullying by schoolmates that mimicked the jerky movements of Minimata disease sufferers; he eventually returned to Minimata, unable to face the discrimination. After the halting of acetaldehyde production, concentrations of mercury in the sediments of Minimata Bay fell. By 1994, concentrations of mercury in ﬁsh of the bay were below the national standards (0.4 ppm total mercury, 0.4 ppm methyl mercury), and the bay was open for ﬁshing in 1997, after 26 years. More recent information on the concentrations of mercury across the sediments beneath Minamata Bay show higher concentrations relatively near the original source, the Chisso plant site (contours, Fig. 9.1), and lower but measurable concentrations farther away. Near the Chisso site, the proﬁles of concentrations of mercury reveal the time course of events. Sediments lower than 15–20 cm show little trace of contamination; either they were deposited pre-Chisso or the mercury has not been transported downward. Above 10 cm, the peak concentrations evidently manifest the effect of contamination, but the proﬁles from sites near-shore show lower con-

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Figure 9.3 Aerial view of the Minimata City area. In the center of photo is the area ﬁlled to isolate the mercury-laden sediments, now dedicated to public uses, and named, with no intended irony, “Eco Park Minamata”. From http://www.fsinnet.or.jp/ ∼soshisha/10tisiki/10_6_e.htm.

0

Mercury concentration (mg kg–1) 1 2 3

4

0 12 km 10

3 km 0 km

Depth (cm)

centrations near the upper layers. These lower concentrations might be evidence of the lower mercury discharges since 1968, or that mercury in near-shore sediments may be dispersed into the outer coastal waters (see Fig. 9.1). This latter possibility may be the reason for an increase in mercury concentrations evident in the uppermost layers of sediments even in areas farthest away from the original source (Fig. 9.4). In any case, a half century later, mercury is still detectable in reasonably high concentrations in coastal sediments near the discharge area. On the other hand, the mercury content of copepods from Minamata Bay decreased by 94% between the early 1980s to 1990, from 8.2 to 0.5 ppm, although the concentrations are still somewhat higher than those of pristine Paciﬁc waters (Hirota et al. 1993). In strong contrast to the effects on humans, there is little evidence with which to gauge effects of the mercury discharges to Minamata Bay on marine life or the coastal environment in general. There are only anecdotal descriptions of ﬂoating ﬁsh and “dancing” cats, and some early studies of ﬁsh toxicity by mercury (Matida et al. 1971). Even during the period of peak mercury discharge there were many shellﬁsh and ﬁnﬁsh in Minamata Bay, perhaps toxic to people and other vertebrates, but alive and numerous. Minamata provides a worst-case instance of metal contamination, with powerful public health

20

30

40

50

Figure 9.4 Vertical proﬁles of mercury concentrations in sediments in sites located 0, 3, and 12 km away from the Minamata shoreline. Data from Tomiyasu et al. (2000).

effects. Minamata also illustrates how, as a result of being tied to speciﬁc sources, the metal contamination was local and generally transient. In addition, the Minamata case, in spite of the obvious human consequences, provided scant evidence of broader environmental impacts.

207

METALS

Table 9.2 Major inputs of selected metals to the North Sea, as percentages per source, and (bottom line) as total metric tons of metals per year. Data originally from R. L. Norton, summarized from Salomons and Förstner (1984). Sources Atmosphere Riversa Direct discharge Disposal of sewage sludge Disposal of other wastes Total (metric tons yr−1)

Copper

Zinc

58 33 5 3 1 5,930

Lead

35 45 18 2 0.4 30,900

Chromium

Cadmium

58 31 7 2 1

25 63 7 3 1

63 34 2 1 –

7,740

2,880

840

Nickel 45 44 8 2 0.3 3,650

Mercury 14 68 14 7 – 44

a

A substantial part of metals delivered through rivers is buried in riverine sediments before reaching the sea, and some major rivers are dredged, with the spoil taken to land. The percentage contribution by rivers reported in this table is therefore an overestimate, and the percentage from atmospheric and other sources is larger than shown.

Sources of metals to coastal waters We use metals for a plethora of purposes, and in doing so, release metals into natural environments.4 Tailings from extraction processes, combustion of fuels, and disposal of waste all contribute metals that then enter coastal environments via streams and rivers, waste water or disposal of solids, or via atmospheric deposition. In most coastal regions, atmospheric and river sources are the two major avenues transporting metals to coastal waters (Table 9.2). In general, atmospheric transport is the agent that distributes metals at global geographic scales. For instance, atmospheric transport is responsible for the appearance of anthropogenic lead and other metals, presumably from industrial sources in

South America or farther distances, in water off the coast of the Antarctic Peninsula (SañudoWilhelmy et al. 2002). Lead is released to the atmosphere by industrial sources, combustion, and mining in the mainland of North America. This lead, transported via the atmosphere, is the major source of lead into coastal and oceanic water and organisms of the West Atlantic, including the Sargasso Sea (Marcantonio et al. 2002). In general, as would be expected from the multiplicity of human activities, urban coastal areas are exposed to far larger metal loading than non-urban areas (Lindegarth & Underwood 2002). In Australia, for example, sediments near urban sites held about 1,100, 800, and 300 mg metal g−1 sediment for, respectively, zinc, lead, and copper. In more pristine areas, metals in near-shore sediments held roughly an order of magnitude lower concentrations (100, 50, and 25 of the same units for the same three metals).

4

The relative balance between natural vs. anthropogenic sources of metals is still in some dispute. For example, for mercury, natural sources (volcanoes, geysers, weathering, degassing of rocks, evaporation from the sea, and river and glacial runoff) to the atmosphere may range from 25 to 182 × 103 tons of mercury per year. Human sources (industry, mining, burning fuels) may range from 7.6 to 38 × 103 tons of mercury per year (Kaiser & Tölg 1980). Globally, then, human sources may be of lesser magnitude than natural sources, and hence should not result in detectable increases across historical time. This runs counter to ﬁndings in cores of Greenland ice, where secular increases were evident. It may be that unaccounted-for increases in mercury degassing from soils exposed by land use by people add a considerably larger mercury input to the atmosphere (Kaiser & Tölg 1980).

Retention of heavy metals in coastal ecosystems Once within estuaries or similar coastal systems, metals may be bound to or complexed with other matter, buried in estuarine sediments, or transported out to deeper waters. Studies of different metals in different estuaries variously conclude that large or small portions of metals entering

208

CHAPTER 9

estuaries are retained within the estuaries, so that generalizations are difﬁcult. In a local-scale study in New Haven Harbor, USA, salt marshes fringing the estuary and harbor removed about 20–30% of the metal ﬂux from land before the metals reached the harbor (Rozan & Benoit 2001). Metal retention in estuaries is also of a signiﬁcant magnitude. Measurements in the St Lawrence estuary of Canada suggested that metal retention in estuaries may range from 54% for zinc, 66% for copper, and 69% for cadmium and manganese, to 92% for nickel and 99% for iron (Yeats &

A primer on certain properties of metals

Brewers 1982). Retention rates are tied to different propensities to exist as a free ion, adsorbed, bound, or complexed, as noted in the box. In general, metal concentrations in water and sediments decrease sharply as distance from speciﬁc sources increases, as we saw in the case of sedimentary mercury in Minamata Bay (see Fig. 9.1). Such gradients are also common for other metals, such as cadmium (Fig. 9.5), and conﬁrm that human-derived contamination of coastal environments by heavy metals tends to be localized.

Elements such as chromium, manganese, iron, copper, zinc, silver, cadmium, tin, gold, mercury, and lead differ greatly in biological functions. Some of these elements are essential, in trace amounts, for biological functions: zinc is part of many enzymes, including carbonic hydrase; iron is needed in hemoglobin, the molecule involved in oxygen uptake in vertebrates and many invertebrates; copper is at the center of hemocyanin molecules in mollusks and arthropods; vanadium is part of a respiratory pigment in tunicates; cobalt is esssential in vitamin B12; and cadmium plays a role in carbonic anhydrase in diatoms (Cullen et al. 1999; Rainbow 2002). Some metals, such as mercury, lead, copper, tin, and cadmium may be toxic to organisms in higher concentrations. These are singled out in this section because they include the most environmentally troublesome of the metals, and also span a wide range of environmentally relevant properties. Metals may occur in natural environments in various forms, with various degrees of

biological impact. Perhaps the principal distinction that can be made is that certain metals can bind to organic (alkyl) groups. These alkyl groups can be, for example, methyls (CH−3 ) or butyls (four-carbon groups). For mercury, for instance, the principal alkyl of concern is methyl, and its binding to metals is a process referred to as methylation.5 Methylation can occur in sediments, where it is microbially mediated, particularly in anaerobic sediments, or within larger organisms. Methylated metals are lipophilic, that is, they readily penetrate fat-containing biological membranes, and hence are likely to enter organisms and be stored in fatty tissues. Mercury occurs in the sea mainly in Hg0 and Hg2+ states6 (Craig & Moreton 1985; Morel et al. 1998). Free mercury occurs in very low concentrations in marine environments7 because it volatilizes, combines with sulﬁdes, or is complexed with organic matter (Rolfhus & Fitzgerald 2001). The most common mercury compounds in sea water are mercury-halogen species, and often these are tightly bound in organic complexes. Mercury accumulates in sediments by adsorption to particles and by

5

6

Mercury has 14 additional electrons in its fourth outer orbital, which may support the formation of highly stable carbon– mercury bonds. The fewer electrons in the outer orbital of cadmium and zinc may lead to less stable metal–carbon bonds. As a result, organo-metal compounds of cadmium and zinc can be rapidly degraded once they are released to natural waters, and are less prone to become serious pollutants (Förstner 1980).

The designation of mercury as “Hg” derives from its Latin name, “hydrargyrum”, meaning “liquid silver”, a term related to “argentum vivum”, referring to live or quick silver, a description of the liquid state of the elemental metal. 7 Concentrations reached in Minamata sediments (>2,000 µg Hg g−1 sediment) have not been matched elsewhere. Some Wisconsin lake and river sediments reached more than 600 µg

METALS

209

forming sulﬁdes.8 Mercury sulﬁdes may be involved in transformations and the release of mercury. Microbial activity (primarily by sulfate-reducing bacteria) creates methyl mercury in sediments. Lead in sea water occurs mainly in the Pb2+ oxidation state. Concentrations of lead as a free ion in sea water are usually quite low, with greater amounts of chloride, carbonate, hydroxide, and other compounds (Chow 1978). Exchangeable (available) lead is normally only a tiny portion of the total lead found in natural environments. Most lead compounds are adsorbed to clay particles, strongly complexed with organic matter, or covalently bound to a variety of compounds; for example, lead phosphates are signiﬁcant and permanent sinks in soils and sediments. In anaerobic sediments, insoluble lead sulﬁdes are important. Methylation of lead is rare; alkyl lead, derived from anti-knock agents in gasoline, readily enters and is toxic to organisms, but tends to be degraded in air and natural waters near the sources of contamination. Copper occurs as the Cu+ and Cu2+ oxidation states, and precipitates out of solution in sulﬁdes, carbonates, oxides, and other compounds, and is often complexed by organic matter. Cu2+ is very strongly adsorbed in clay–metal– organic complexes, and hence is less mobile than other metals, with the possible exception of lead. Coastal sediments are usually a sink for copper, but where sulﬁdes are taken up in iron compounds, the cuprous copper may not be precipitated (Davies-Colley et al. 1985). Cadmium exists mainly in the Cd2+ oxidation state. Cd2+ is only weakly adsorbed, and

so about two-thirds of cadmium occurs as the free ion, with lesser proportions of CdCO3, Cd(OH)2, CdCl2, and CdSO4. The exact proportions may depend on the total amount of cadmium, the amount of organic matter present, and the redox state. In sea water and sediments, cadmium may be found in complexes (chelated by humic materials) that are somewhat degradable and can to some degree release cadmium in aerobic conditions, less so in anaerobic conditions. Methylation of cadmium is not common. As in the case of copper, where the sulﬁdes are taken up in iron complexes, cadmium may not remain within sediments (Davis-Colley et al. 1985) Tin occurs in sea water mainly in the Sn2+ state, and most commonly as a hydroxide. In sediments, alkylated tin and tin oxide can be found. Organo-tin compounds, particularly tributyl and dibutyl tin may be produced by microbes, but in minor amounts. Tributyl tin manufactured industrially is a far more prominent source for coastal waters, and can be highly toxic. Tributyl tin has been widely used in disinfectants and antifouling compounds. In water it degrades at reasonably fast rates to dibutyl tin and then to monobutyl tin, which is persistent; in sediments, tributyl tin does not seem to be readily degraded (Adelman et al. 1990). This too-brief review of the features of metals demonstrates the signiﬁcant differences in likely biogical effects and biogeochemical behavior of important metals. These differences highlight that each has idiosyncratic properties that make certain metals more likely to have environmental impacts than others.

Hg g−1 sediment just next to an industrial source, but more generally, aquatic sediments range around 0.016 µg Hg g−1 sediment (Kaiser & Tölg 1980). 8 Mercury in the form of cinnabar (HgS) ore was known to the ancient Chinese, who used it to prolong life and as a red ink. Roman writers describe how cinnabar was used in an amal-gamation process to recover gold from cloth garments. Twelfth century Egyptians used amalgamation for large-scale extraction of metals, and the method was used in the 16th

century to process silver in the Americas, with the use and distillation of mercury as a gas as part of the process. The life expectancy of the miners involved in amalgamation was 6 months. Paracelsus in the 16th century described a therapy for treating mercury miners, and in spite of the known danger of mercury vapors, mercury was still used as a remedy for syphilis as late as the 19th century. The toxic nature of mercury had to await the tragedy of Minamata to become fully recognized (Kaiser & Tölg 1980).

210

CHAPTER 9

Cadmium (µg l–1)

8

SEA WATER

6 4 2 0

Cadmium (mg kg–1)

0

20

40

60

80

ROCKWEED 200

100 Metal 0 0

Cadmium (mg kg–1)

Table 9.3 Long-term retention of selected metals in the top 8 cm of salt marsh sediments, as a percentage of metals added experimentally as chronic treatment every 2 weeks during the growing season (1970–1981) to two series of experimental plots. The two series of plots were treated with somewhat different metal doses, and the sediments in the two series also differed in redox condition, so they were not true replications; the means calculated in this table should be considered merely as a convenient rough indication of retention, with considerable uncertainty. Data from Giblin et al. (1983) and Breteler et al. (1981).

20

40

60

80

ANIMAL

600 Snail 400 Limpet

200

Copper Iron Lead Cadmium Manganese Zinc Chromium Mercurya

First series of plots

Second series of plots

Mean % retention

49 24 60 15 27 28 45 100

70 70 39 39 30 40 59 50

59.5 47 49.5 27 28.5 34 52 75

aBased on only 2 years of data, 1977–1978.

0 0

20 40 60 Distance away from major source (km)

80

Figure 9.5 Cadmium concentration in sea water, a rockweed (Fucus), and two animals, a suspensionfeeding limpet (Patella) and a predator snail (Thais, now called Nucella), from the Severn Estuary and Bristol Channel. The data were collected from sites at different distances away from the principal cadmium source, the Bristol metropolitan area, UK. Data from Butterworth et al. (1972).

Most metals are to some degree retained and buried within coastal sediments, in particular those under mangroves and salt marshes,9 where the ﬁne organic sediments offer enormous surface areas for adsorption and organic chelators for complexation, as well as sulﬁdes and other precipitants. Many papers suggest that such sedimentary coastal environments are therefore liable to sequester contaminants—including metals— 9

Mechanisms, retention rates, and contrasts among different metals are discussed by Tam and Wong (1993, 1994, 1999), and have already been mentioned as one of the important functions of these habitats in Chapter 6.

and hence provide us with a water quality subsidy. An extension of this idea is that, under certain speciﬁcations, these environments might be engineered as adjuncts to wastewater treatment plants. This idea was examined in Great Sippewissett salt marsh, on Cape Cod, Massachusetts, where metals such as found in sewage sludge were added experimentally to salt marsh plots during many years. There was a high initial retention of heavy metals; in the longer term, after several more years of biogeochemical reworking of the metals, part of the buried metals were recycled and made newly available for transport out of the marsh sediments. This biogeochemical reworking lowered the retention rates of metals applied to the sediments (Table 9.3). The slowing of retention can be seen in the long-term time course shown in Fig. 9.6, which shows that increases in metal concentrations in sediments failed to keep pace with the potential accumulation that could have occurred had all the metal applied been

211

Lead (ppm) Lead (ppm)

Figure 9.6 Assessment of experimental, chronic addition of lead (Pb) to salt marsh plots in Great Sippewissett salt marsh, Massachusetts. Top panel: time courses of the cumulative amount of lead experimentally added to salt marsh plots, and concentrations in the top 6 cm of sediments of leadtreated and control treatments. Second panel: lead concentrations in salt marsh cordgrass (Spartina alterniﬂora) in lead-treated and control plots. Third panel: lead concentrations in ribbed mussels (Geukensia demissa) in lead-treated and control treatments. Bottom panel: lead concentrations in ﬁddler crabs (Uca pugnax) in lead-treated and control treatments. Data from Giblin et al. (1983).

Lead (ppm)

Lead (g m–2)

METALS

4

Salt marsh Cumulative amount of Pb added Pb in treated sediment Pb in control sediment

2

10

S. alterniflora Treated Control

5

4 2

40 20 1974

G. demissa Treated Control

U. pugnax Treated Control 1975

retained. Of course, different metals nonetheless showed different retention rates, even in the long term (Table 9.3). Cadmium and zinc, as we could have predicted from our review of binding ability, were retained far less than copper and lead, for example. On the whole, the long-term Great Sippewissett data suggest that although these wetland sediments are by no means perfect sinks for pollutants, some substantial portion of metals arriving onto such sediments remain buried in place, and that there are large differences in retention from one metal to another.

Uptake and bioaccumulation of metals Much as in the case of hydrocarbons, the metal content found in organisms depends on supply of the contaminant, uptake, depuration, and accumulation.

1976

1977

1978 Year

1979

1980

1981

The supply of metals available in the environment can be reﬂected in the metal content in certain organisms.10 Ambient metal concentrations result in recognizably consistent concentrations in phytoplankton and macroalgae (Costa et al. 2000; Vasconcelos & Leal 2001; Slaveykova et al. 2003).11 External metal supply inﬂuences the metal concentrations found in plants: lead concentrations in the above-ground shoots of salt marsh cordgrass decreased (Fig. 9.6), likely as the result of sharply lowered atmospheric delivery of lead created by the ban on leaded gasoline enacted in the 1960s in the USA. The lower lead concentrations also suggest that the source of the lead for the shoots was atmospheric deposition rather than uptake from the sediment. 10

Lee et al. (2000) and Shulkin et al. (2002) review different relationships of metals in sediments and uptake by benthic organisms. 11 Metal accumulation in seaweeds also depends on ambient nutrient supply (Lee & Wang 2001): fronds of sea lettuce accumulated more cadmium, chromium, and zinc if they had more nitrate available.

Copper concentration (µg g–1)

212

CHAPTER 9

300

200

100

R 2 = 0.92 y = 0.048x + 33.393

0 0

1,000 2,000 3,000 4,000 5,000 6,000 Annual copper load (kg yr –1)

Figure 9.7 Mean annual concentration of copper in the bivalve Macoma balthica vs. the annual copper load to South San Francisco Bay, data collected from 1977 to 1998. Adapted from Hornberger et al. (2000).

Ambient metal supply may also affect metal content in animals. The copper content of a deposit-feeding bivalve was closely associated with copper loads to the South San Francisco Bay (Fig. 9.7): greater metal contents appear in bivalves living contemporaneously with larger copper loads entering the water body. Bivalves, mussels in particular, have been used as sentinels for the contamination of coastal waters in monitoring efforts such as the International Mussel Watch, and similar plans (Phillips & Rainbow 1993). In other instances, just as we saw earlier with hydrocarbons, the link between ambient metals and content in organisms is less evident. Metal contents in mangrove tissues in Guanabara Bay, near Rio de Janeiro, Brazil, were unrelated to, and mostly lower, than concentrations within sediments where the mangroves grew (Machado et al. 2002). Otero et al. (2000) found no relationship of concentration of the various forms of nickel, chromium, zinc, or copper in sediments and the concentrations of these metals in a marine worm. Bryan and Gibbs (1987) reported no evident relationship between metal concentrations in two species of polychaetes and sedimentary concentrations of copper, iron, zinc, or manganese; metal concentrations in a third species showed a positive relationship to sediment concentrations of copper, zinc, cobalt, silver, iron, and lead, but not to

manganese. The bioaccumulation of 18 trace elements in green and hawksbill turtles differed considerably among these two species of marine turtles, and among the different elements in each species (Anan et al. 2001). The uptake and bioaccumulation of different metals and even the forms of each metal differ widely among different organisms (for example, recall that methyl mercury accumulates more effectively in organisms than do other forms of mercury) (Wang 2002). In the Great Sippewissett salt marsh work, plants and animals accumulated cadmium, a non-essential and toxic element, in plots experimentally enriched in metals; in contrast, plants, but not animals, showed increased zinc concentrations, a metal that is a necessary trace element (Giblin et al. 1983). Lead, a nonessential toxic, did not accumulate in either plants or animals (Fig. 9.6). Concentrations of metals in different organisms vary owing to differences in the biology of the species, even without anthropogenic contributions. A zinc content that is low for oysters would be high for mussels, and zinc concentrations that are high for a shrimp would be below what might ever be found in a barnacle (Rainbow 2002). Vanadium is quite high in sea squirts, but low in most other marine organisms. Even within related organisms there are large ranges in metal contents. It is evident that different types of organisms have idiosyncratic propensities to differentially take up and retain different metals.12 As in the case of hydrocarbons, there has been concern about whether metal contaminants may become concentrated in organisms in the upper steps of coastal food webs. Some data, for example the rockweed, limpet, and snail data of Fig. 9.5, tempt one to see such trophic biomagniﬁcation. A few studies provide more convincing evidence. 12

In addition, different parts of organisms may hold differing metal concentrations (livers tend to have higher levels). There are also certain species in which larger specimens show proportionately larger metal contents than smaller specimens, owing to time of exposure and physiological state (Abreu et al. 2000; Cubbada et al. 2001), as we found in the case of chlorinated hydrocarbons (Chapter 8). The size/concentration relation is not general, because, as in the example of polychaete worms, there was no relationship between metal concentrations in tissues and size of the worms (Bryan & Gibbs 1987).

METALS

213

in metals, species, and circumstances to be able to generalize too far.

A study of relative bioaccumulation of methyl mercury relative to total mercury found that the ratio of methylated to total mercury was 10% in the water column, 15% in phytoplankton algae, 30% in zooplankton that grazed the algae, and reached 95% in ﬁsh (Watras & Bloom 1992). Another study also found that methyl mercury— but not other mercury compounds—bioaccumulated from water to phytoplankton to zooplankton to ﬁsh (Cabana & Rasmussen 1994).13 Other studies, however, show inconsistent evidence of biomagniﬁcation of metals. In an Arctic marine food web, biomagniﬁcation of mercury was found among species of marine animals, but not animals of different size within species, and not in bivalves, ringed seals, or polar bears (Atwell et al. 1998). Copper concentrations in deposit-feeding bivalves rose and fell across decades in South San Francisco Bay (Fig. 9.8). The salient point here, though, is that initially the bivalves contained almost four times as much copper as the sediments, but by the end of the data record, ended up containing less copper than the sediments: bioaccumulation was hardly a constant. Concentrations of cadmium, chromium, and zinc were lower in mussels and clams of Jinzhou Bay, China, than in the particles they fed upon (Fan et al. 2002). Transfer and magniﬁcation of methylated metals may therefore occur in certain trophic links, but it can be hardly thought of as unvarying, and many other comparisons have not conﬁrmed the ubiquity of metal magniﬁcation. Some reviewers of information on metal bioaccumulation concluded, in fact, that “there is no convincing evidence that lead is [biomagniﬁed] across food chains . . .” (Wong et al. 1978), and that “trace metal concentrations are not as a rule biomagniﬁed along food chains . . .” (Rainbow 2002). There is simply too much case-speciﬁc variation

Trace metals bind not just to sediment and organic particles in the external environment, but also bind readily to protein sulfur and nitrogen inside cells. Such binding makes the metals less able to alter life activities of the organisms. Thus, there are two more or less well-deﬁned pools of metals within organisms, one in which the metal is bound in a detoxiﬁed form, and another that is sufﬁciently metabolically active as to cause biological effects. This second pool is the issue of concern here, and its magnitude might not be related to total metal content, and varies markedly from one species to another. Deﬁning the metabolically active pool of metals is complicated by a species-speciﬁc need for metals for essential metabolic functions. Concentrations in the active pool may surpass the essential titer, and may reach levels that lead to toxicity. These details have made it difﬁcult to generalize about linking concentrations in organisms to harmful biological effects. Copper, for example, might be necessary in trace amounts, but becomes toxic at higher concentrations; lead, in contrast, is toxic to all groups of organisms at some level, and even at the generally low concentrations in coastal environments14 may have sublethal effects (Wong et al. 1978). The list of harmful biological effects of exposure to high concentrations of metals is far too long to include here.15 Sufﬁce it to say that many metals have deleterious effects that alter many aspects of biological function of a remarkable diversity of organisms. These injurious effects include inhibition of growth, increased respira-

13

14

This paper also showed that the concentration of methyl mercury was linearly related to δ15N, which, as we saw in Chapter 8, is a measure of position on a food web. In this case, the interpretation that methyl mercury was magniﬁed along the food web is not completely convincing. The heavier isotopic signature could also be derived from different nitrogen sources (waste water, for instance) that may enter the receiving waters, and furnish heavier δ15N values to the water bodies in the comparison.

Biological effects of metals

There are certain exceptional, and characteristically local, circumstances where lead is clearly a problem. One instance is the curious link to lead pellets from hunter’s missed shots and poisoning of waterfowl that inadvertently ingest the pellets as they forage in the shallow water bodies (Sileo et al. 2001; Tavecchia et al. 2001). 15 Some of the harmful effects are discussed in Nriagu (1978), Hutzinger (1980), GESAMP (1985), George (1990), Sorensen (1991), Dallinger and Rainbow (1992), and Rainbow (2002).

214

CHAPTER 9

Copper (kg yr –1)

6,000

ANNUAL COPPER LOAD

4,000

2,000

0

Copper (µg g–1 sediment)

100

COPPER IN SEDIMENT

80 60 40 20 0

Copper (µg g–1 bivalve)

300

COPPER IN BIVALVES

200

100

0 1977 1979 1981 1983 1985 1987 1989 1991 1993 1995 1997 Year

tion, lowered feeding and digestion, inﬂamed tissues, abnormal development, immunodeﬁciency, and much else among animals, and many effects such as loss of dinoﬂagellate bioluminescence (Heimann et al. 2002) among other kinds of organisms. Nevertheless, in spite of concerted investigations, no clear relationship has been deﬁned that can relate concentrations of metals in organisms speciﬁcally to harmful effects. There

Figure 9.8 Time course (1977–1998) of annual copper loads (top), and copper concentrations in sediments (middle) and in bivalves (Macoma balthica) (bottom) in South San Francisco Bay. Data from Hornberger et al. (2000).

are also many examples of investigations showing that effects of metals on coastal organisms may be modest and sublethal (Sze & Lee 2000; Méndez 2002), insigniﬁcant, or ambiguous (for instance Selck et al. 1998; Harris & Santos 2000). There are many suggestions that there are indirect or synergistic effects of metals. For example, Sunda (1987) concluded that increased concentrations of non-essential metals in coastal

METALS

215

waters may reach levels where they may be taken up by phytoplankton instead of essential trace metals. This would interfere with many key metabolic pathways since essential trace metals are involved in so many metabolic activities. There are also suggestions that the impact of metals may be lowered by changes in exposed organisms. Many organisms, from bacteria (Nakamura et al. 1986) to animals (Weis et al. 1981; Karande et al. 1993, Mouneyrac et al. 2003) may develop some degree of tolerance or even resistance16 to metals. It is surprisingly difﬁcult, moreover, to ﬁnd compelling data demonstrating signiﬁcant population-level effects of heavy metal contamination of coastal waters. There were anecdotal observations that ﬁsh were found ﬂoating in Minamata Bay, but in spite of the large literature on that classic example, few actual data on the population status appear to have been recorded. In the majority of cases where biological effects have been found, the effects of heavy metals are evident primarily at extraordinarily high concentrations of available metals. The classic example of high levels of mercury contamination is that of Minamata Bay; in the local vicinity of the Nisso plant, methyl mercury became extraordinarily high in concentration, far higher than seen elsewhere. After the reduction of local sources, the contamination and effects disappeared. The same seemed true for organotin contamination, probably the best documented example of population effects on marine organisms due to contamination with a metal (see box on next page). In both of these prominent cases of metal contamination, it has in general been difﬁcult to demonstrate widespread and continuing deleterious impacts of metals at the population level in the ﬁeld, even at local spatial scales. We earlier noted the lack of effects of polychlorinated biphenyls (PCBs) and of mercury on salt marsh biota (see Fig. 8.11). In the Great Sippewisset salt marsh work, in spite of the

presence of signiﬁcant increases in metals in sediments, and increased metals in plants and animals, we could ﬁnd no detrimental population-level effects that were attributable explicitly to metals. Similarly, experimental increases in zinc, copper, and lead performed on Australian sandﬂats led to insigniﬁcant or at most weak and inconsistent responses by the assemblages of benthic animals (Lindegarth & Underwood 2002). These Australian experiments were done in estuarine areas whose watersheds were urbanized and in similar non-urbanized areas. The assemblages of benthic fauna of the two sites differed markedly in metal content17 and in faunal composition and abundance, but the increased concentrations of metals did not cause the assemblages of fauna in pristine sites to become similar to those in urbanized areas. The initial differences in fauna between the two areas were therefore more than likely caused by agents of ecological change other than metals (see box on next page). The extensive work on the effects of organotins, much like the case of mercury in Minamata Bay, corroborates the notions that anthropogenic metal contamination of coastal environments tends to have local, transient, and limited ecological impacts on selected subsets of organisms. This, as in the case of PCBs, raises the perplexing situation where we have clear physiological impacts, but it has been hard to discern how these impacts translate to the population level.

16

17

The resistance mechanism is believed to be the enhanced synthesis of metal-binding proteins (metallothioneins). Metallothioneins are rich in cysteine, whose thiol groups readily bind to the metals. The bound metals would be less toxic than the free ions (Dixon & Sprague 1981; Mouneyrac et al. 2003).

Sources and time course of metal contamination Human beings have signiﬁcantly altered metal cycling through the world’s environments (Lantzy & Mackenzie 1979; Mackenzie et al. 1979; Pacyna et al. 2003). Mining of metals far exceeds the natural weathering rates, and atmospheric deposition of most metals onto the earth’s surface may be 3–350 times higher than estimated from natural The urban sites showed concentrations of zinc, lead, and copper (1,100, 800, and 300 µg metal g−1 sediment, respectively) about one order of magnitude higher than those in the pristine areas (100, 50, and 25 µg metal g−1 sediment, respectively) (Lindegarth & Underwood 2002).

216

CHAPTER 9

The curious case of organotins and imposex A determined search for effective ways to prevent fouling of ship hulls and marine structures resulted in the identiﬁcation and marketing of tributyl tin (TBT), a type of compound that may be one of the most poisonous compounds deliberately introduced into marine environments (Goldberg 1986).18 The ﬁrst notice that TBT might have unwanted effects came from extensive oyster farming operations19 that were located next to piers and boat berths in Arcachon, France. From 1975 to 1982 oysters were stunted, shells that were grew were malformed, there was virtually no reproduction or recruitment,20 and new oysters transplanted to the area suffered a mortality of 50% within a month. Follow-up investigations showed that concentrations of TBT as low as parts per trillion in water led to deformations and multiple anomalies in bivalves and other mollusks (Alzieu et al. 1986). In particular, female snails developed penises after exposure to TBT, a phenomenon that came to be known as imposex.21 The frequency of imposex was attributable to exposure to organotins in the environment. Field evidence showed that dogwhelks transplanted to organotin-contaminated sites 18

The negative effects, which are discussed below, need to be weighed against the positive effects of effective lowering of fouling. For example, vessel hulls with no fouling organisms burn less fossil fuels, and hence release less greenhouse gases and lower the chance of introducing alien species to new coastal habitats (Evans et al. 2000a). 19 This maricultural operation was massive, producing 10,000– 15,000 metric tons of oysters (the giant Paciﬁc oyster Crassostrea gigas) (Alzieu 2000). The maricultural facilities surrounded marinas and seasonal moorings that accommodated about 7,800 vessels and boats. The lost revenue from 1977 to 1983 amounted to about US$147 million. 20 Later research (Labare et al. 1997) suggested that larvae were not affected by dissolved TBT, but rather that settlement and metamorphosis were impaired by the higher concentrations of TBT that bioaccumulated in microbial ﬁlms on substrates where oyster larvae attempted to settle. 21 Imposex has been described for 118 species of marine snails (Evans et al. 2000b). Imposex is quantiﬁed in diverse ways: by reference to the percentage of a population that shows the females with enlarged penises; or by the relative penis size index (RPSI), a measure of size of the penis in female snails rela-

developed imposex (Fig. 9.9 left). These transplanted snails also acquired higher concentrations of tributyl and dibutyl tin (Fig. 9.9 right). The degree of imposex in snails was correlated to tin concentration in the substrate (Fig. 9.10), but the speciﬁc shape of the relationship differed between one snail species and another. Because of the afﬁnity of metals, including tin, to be bound to sediments, the geographic extent of the effects of tin contamination was quite local. The impact of tin contamination in small ﬁshing ports in the Isle of Cumbrae (Fig. 9.11 top) or in larger ﬁshing and commercial harbors in Iceland (Fig. 9.11 bottom) were restricted to the vicinity of the sources of the organotins. French authorities established butyl tin control policies in 1982 only 3 months after the evidence was made public, an extraordinarily fast response by the political sector to a scientiﬁc demonstration. France prohibited the use of antifouling paints containing more than 3% organotins on boats smaller than 25 m long, within areas with extensive mariculture. Similar regulations were enacted in other developed countries in subsequent years (Evans et al. 2000b). TBT regulations led to lowered concentrations of organotins22 in water, sediments (Table 9.4), and animals. TBT concentration in tive to that of male snails. Other described effects of organotins include decreased oxidative degradation and nitriﬁcation in bacteria (Dahllöf et al. 1999), lower growth and altered chemistry in phytoplankton (Beaumont & Newman 1986; Sidharthan et al. 2002), toxicity to oyster larvae (Labare et al. 1997), impaired osmoregulation and lower survival in shrimp (Lignot et al. 1998), toxicity to crabs and lobsters (Laughlin & French 1980), liver accumulations in cetaceans (Le et al. 1999), and, indirectly, to have even been accused of leading to algal blooms because of the impairment of bivalve suspension feeding (Ruiz 1999). These are but a few of the many papers reporting effects of butyl tins on marine organisms. 22 The decreased use of organotin antifouling paints produced one undesirable effect: use of the more traditional copper-based paints increased, and there was a detectable increase in copper in shellﬁsh after 1982 (Alzieu 2000). In fact, there is little evidence that the many compounds proposed as antifoulants are less injurious than organotins, and each has been shown to have potential biological effects (Evans et al. 2000a). It very well may be that the continued use of organotins, but in lower dosages and under more careful regulatory scrutiny, may be a preferable option.

217

METALS

70

% imposex

60 Transplanted to harbor

50 40 30 20 10

In clean site

0 M

A

M

J

J

A S Month

O

N

Iceland (Table 9.5). Ship activity, particularly of large vessels, largely unaffected by regulations, still seems related to the effects of organotins. The more recent data show lowered effects, apparently tied to the stricter controls of organotin antifouling use. It is still difﬁcult, however, to ﬁrmly establish population-level effects of the organotins. For example, surveys done in 26 sites throughout the Portuguese coast in the late 1990s showed no relationship between the percentage imposex, or RPSI, and density of dogwhelks. In fact, the highest densities of these snails occurred in the places with 80–100% imposex (Santos et al. 2000). This is counter to expectations if organotins were signiﬁcantly detrimental to Portuguese dogwhelk populations. In the Crouch estuary in England, where concentrations of TBT decreased signiﬁcantly in all seven sampling sites (Table 9.4), the density of populations of benthic invertebrates did not change in four sites, decreased in two, and increased (the expected result!) in only one site (Rees et al. 2001). Such results suggest that TBT concentrations likely played at best a modest part in determining densities of these estuarine populations.

Tin concentration (µg g–1 dry wt)

Arcachon oysters decreased 5–10-fold between 1982 and 1985. Lower concentrations of TBT in sediments and in periwinkles were accompanied by marked increases in recruitment of new periwinkle cohorts (Matthiessen et al. 1995). Organotins are still released from larger vessels, so the effects have not disappeared. Long-term surveys of sterile female dogwhelks and relative penis size index (RPSI) in dogwhelks revealed that smaller, but still signiﬁcant, portions of the female population remained sterile into the 1990s (Fig. 9.12 left), no doubt because these snails are relatively long-lived. The RPSI data showed greater trends toward recovery (Fig. 9.12 right): these surveys show consistent reductions of imposex in affected populations, relative to the same variables recorded in sites outside the inlet frequented by large oil tankers. There is no question that the initial data from France clearly showed marked population-level effects of intense organotin contamination. Levels of organotin contamination have fallen worldwide after the 1980s, but the effects have persisted in many places throughout much of the world’s shorelines, such as Malta and

D

J

F

1.2 TBT Transplanted to harbor

0.8

DBT

0.4 In clean site 0 M

A

M

J

J

A S Month

O

N

D

J

F

Figure 9.9 Left: percentage of dogwhelks (Nucella lapillus) with imposex in a clean site in southwest England and in dogwhelks transplanted from the clean site to a harbor used by many commercial and pleasure vessels. Right: concentrations of tributyl tin (TBT) and dibutyl tin (DBT) in the tissues of dogwhelks in the clean site, and in those transplanted to the harbor site. The black and open symbols refer to two sets of snails. Data from Brian et al. (1986).

218

CHAPTER 9

100 % imposex

% imposex

60

N. lapillus L. littorea

40

20

80 60 40 20

0

0 0

0.2 0.4 0.6 TBT concentration (µg g–1 dry tissue)

0.8

1

10 100 1,000 DBT concentration (ng)

10,000

14

500

Isle of Cumbrae

% imposex ( )

12

400

10 8

300

6

200

4 100

2 0 100 100

80

60

40

20

0

20

40

60

80

0 100

Organotin concentration (ng g–1) ( )

Figure 9.10 Percentage of populations of dogwhelks (Nucella lapillus, open circles) and periwinkles (Littorina littorea, black circles) showing the relationship between levels of imposex and tributyl tin (TBT) and dibutyl tin (DBT) concentrations in their tissues. Data from Brian et al. (1986) and Birchenough et al. (2002).

Iceland

% imposex

80 60 40 20 0 0

10

20

30

Distance from source (m)

ﬂuxes. Anthropogenic metal sources also dominate coastal environments (Fitzgerald et al. 2000). Corals (Guzman & Garcia 2002) and ice cores in Greenland (Boutron et al. 1998) in regions remote from urban areas show records with signiﬁcant degrees of mercury deposition. These depositional

Figure 9.11 Organotin concentrations and levels of imposex at different distances away from sources in a small harbor in the Isle of Cumbrae (top) and a harbor in Iceland (bottom). From Evans (1999).

records can be attributed mainly to emissions from distant industrial and other anthropogenic activities. Since, as noted earlier, there is retention and burial of metals in estuarine sediments we might expect that, if indeed humans have become

219

METALS

Table 9.4 Time course of the range of concentration of tributyl tin (TBT) in the sediments of two European sites. Range in concentrations of TBT (ng l−1) 1986

1987

1988

1989

1990

1991

Arcachon Bay, France

< 2–89

< 2–50

–

–

–

–

< 0.2–7

Alzieu 2002

Crouch estuary, UK

10–44

7–32

8–22

2–12

1–8

1–7

1–4

Rees et al. 2001

70

70

60

60

50

50

RPSI (%)

FS (%)

Site

40 30 20

1992

Source

Inside inlet Outside inlet

40 30 20

10

10

0

0 1987 1988 1989 1990 1991 1992 1993 1994 1995 Year

1987 1988 1989 1990 1991 1992 1993 1994 1995 Year

Figure 9.12 History of organotin effects. Time courses (1987–1995) of the percentage of females that were sterile (FS) and the relative penis size index (RPSI = volume of female penis/volume of male penis × 100; volume estimates as cube of length). Data from within and outside Sullom Voe, an inlet in the Shetland Islands, UK, where there is a large terminal for oil tankers serving the North Sea oil ﬁelds. Data from Harding et al. (1997).

Table 9.5 Relative penis size index (RPSI) for snails (Hexaplex trunculus in Malta, Nucella lapillus in Iceland) found in three types of harbors in Malta and Iceland. Large ports, shipping

Small harbors, high boat activity

Malta Late 1990s No. of sites

93–107 3

63–80 4

Iceland 1992–1993 1998 No. of sites

0–63 0–9 11

0.6–89 0–25 20

Small harbors, modest boat activity

0–15 14 – –

Sources

Axiak et al. 2000

Svarvarsson 2000 Svarvarsson 2000

220

0.1 2000

CHAPTER 9

Metal / Al (mg metal g–1 Al) 0.5 0.2 0.4 0.6 200

0.3

400

600

1980 1960 1940 Sediment age (yr)

1920 1900

Cobalt

Copper

Iron

1880 0.2 2000

0.5

0.8

0

2

4

1980 1960 1940 1920 1900

Lead

Zinc

1880

increasingly responsible for metal emissions and deposition across the centuries, there ought to be a sedimentary record of these changes in metal inputs to estuaries. The sedimentary record23 in many coastal areas24 around the world clearly reﬂects changes across historical time (Fig. 9.13).25 Most metals were found in relatively 23

Similar time courses have been found in Greenland and Antarctic ice cores. 24 Including Narragansett Bay, Rhode Island (Goldberg et al. 1977), coral reefs (Shen et al. 1987), Halifax Harbor (Gearing et al. 1991), Minamata Bay (Fig. 9.1), Chesapeake Bay (Fig. 9.13; Owens & Cornwell 1995), Turkish Black Sea coast (Tuncer et al. 2001), Mississippi River delta, Galveston Bay, Tampa Bay (Santschi et al. 2001), Puget Sound (Crecelius & Bloom 1988), California coast (Stull et al. 1988), Tokyo Bay (Matsumoto 1988), and Jin Zhou Bay (Guoxian et al. 1988) among many others. 25 Concentrations of metals recorded in vertical proﬁles in sediments may also decrease if there are increased accumulations of other materials that also respond to human activities. Increased organic content because of eutrophication, and greater sedimentation rates owing to increased import of sediments, may “dilute” the metal concentrations. Concentrations of aluminum—the metal by far greater in natural concentrations—are less affected by human activity, and can be used to normalize the concentrations of other metals, hence minimizing the possible biases introduced by changes in sedimentation rates or eutrophication. Such normalization was done for Fig. 9.13 (Owens & Cornwell 1995).

Figure 9.13 Proﬁles of different metals (normalized in relation to aluminum, Al) in two stations in Chesapeake Bay with intermediate salinity and a depth of 13 m. The year of sediment deposition is plotted as the y axis instead of depth in the sediment proﬁle. Adapted from data in Owens and Cornwell (1995).

low concentrations up to about the time of the Industrial Revolution. Certain metals (copper, nickel, lead, zinc) became broadly employed in early industries, and their concentrations increased accordingly in waters receiving inputs from land. Other metals (e.g. cobalt) increased more slowly. Marked increases in concentrations appear in estuarine sediments deposited from the 1950s to 1970s. The concentrations of many metals decreased consistently after the 1970s (Fig. 9.13). The relatively unchanged iron proﬁle in Fig. 9.13 merely shows the effect of data normalization to aluminum, the other major natural metal in sediments. There are also many examples of local-scale reductions of metal loads to coastal environments. It seems worthwhile to review a few cases. Between 1968 and 1988 primary treated sewage from about half a million people in the greater Vancouver area was discharged into an intertidal ditch in Sturgeon Bank, in the Fraser River estuary (Arvai et al. 2002). Since 1988, the sewage has been discharged subtidally 5 km beyond the estuary, at a site at a depth of 250 cm in the Strait

METALS

of Georgia. One result of this reduction of waste contamination with metals was that the mercury content of surface sediments of Sturgeon Bank fell from about 0.9 µg g−1 before to about 0.24 µg g−1 after the diversion of sewage (Arvai et al. 2002). Metals were released to the South San Francisco Bay in sewage efﬂuent from a wastewater treatment plant. Concentrations of copper (and of other metals, not shown in Figs 9.7, 9.8) in sediments and in clams on mudﬂats in South San Francisco Bay increased in concert in the 1970s, and decreased after 1980 (see Fig. 9.8). The decreases took place after secondary treatment facilities were added to the wastewater treatment plant. The microbial digestion plus separate disposal of the digested sludge involved in secondary treatment removed most of the metals from the efﬂuent released to the bay, and the metal content of both the sediments and clams mirrored the decreased releases. In fact, through the years, there was a close relationship between the amount of copper released from the plant and the copper content of clams (see Fig. 9.7): as delivered copper decreased in recent decades, there was a parallel decrease in copper in the clams. Waste water from the Boston Metropolitan area was discharged for decades through a nearshore outfall in the harbor, and sediments in the harbor became of prominent note when they were made an election-year issue as the most polluted harbor in the USA. Since 2000, after considerable expense, waste efﬂuent is treated to a higher degree, and the efﬂuent is released via diffusers at the end of a remarkable 8 m-diameter tunnel carved through solid rock to an offshore site 15.2 km out on Massachussetts Bay. In contrast to the earlier condition of sediments within Boston Harbor, Bothner et al. (2002) found little evidence of metal contamination in Massachusetts Bay sediments near the site of the diffusers through which the efﬂuent is currently discharged. Metal content within the near-shore sediments in Boston Harbor is decreasing, and pollution in general is decreasing enough that beaches are now open for bathing and other uses. Local sources other than wastewater discharge can be also force their imprint on the time course of metal contamination. The vertical proﬁles of

221

concentrations of lead, zinc, and tin in sediment cores taken from Central Park Lake, in New York City, suggested that added metals were delivered to the site from the incineration of solid waste, before the advent of leaded gasoline. This local source of metals has decreased signiﬁcantly since incineration was replaced with other methods of garbage disposal (Chillrud et al. 1999). The recent decreases in metal loads evident in the sedimentary records are a result of lower global emissions and deposition from the atmosphere (Nriagu 1996) as well as lower local discharges (Sañudo-Wilhelmy et al. 2004). During the last decades of the past century, the implementation of clean air legislation in many industrial nations of the Northern Hemisphere substantially lowered the emissions of several metals. Regulations about discharges through smokestacks, and, in particular, the widespread bans on leaded gasoline, have been particularly effective at reducing emissions to the world’s atmosphere. Metals in waste water have also been reduced. These changes, largely in the industrialized nations, have led to a global-scale reduction in atmospheric metal loads to the surface of the earth. We saw one example of the effects of decreases in the time course of lead in salt marsh plants and sediments (see Fig. 9.6). As populous developing nations evolve their economies, there is great uncertainty whether present decreased releases of metals will continue. The cases of historical changes in the local delivery of metals reviewed in the above paragraphs suggest, ﬁrst, that concentrations of metals in sediments within coastal environments reﬂect changes in inputs. Second, the local contributions in the majority of cases are far more substantive than the global contributions.26 From these observations, we might conclude that improved local management and disposal of wastewater efﬂuents and solid wastes can make for signiﬁcant improvements in the contamination of receiving environments, and provide a reasonable expectation of success for continued efforts to further lower or at least prevent new 26

Obviously, in coastal areas far from dense human populations, global atmospheric sources can be more signiﬁcant. In these cases, however, the degree of contamination might be inconsequential.

222

CHAPTER 9

metal inputs. Obviously, public will, political action, substantial ﬁnancial resources, and international cooperation will be necessary to bring the improvements about, and to maintain the efforts.

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Machado, W., E. V. Silva-Filho, R. R. Oliveira, and L. D. Lacerda. 2002. Trace metal retention in mangrove ecosystems in Guanabara Bay, SE Brazil. Mar. Pollut. Bull. 44:1277–1280. Mackenzie, F. T., R. J. Lantzy, and V. Paterson. 1979. Global trace metal cycles and predictions. Math. Geol. 11:99–135. Marcantonio, F., A. Zimmerman, Y. Xu, and E. Canuel. 2002. A Pb isotope record of mid-Atlantic US atmospheric Pb emissions in Chesapeake Bay sediments. Mar. Chem. 77:123–132. Matida, Y., H. Kumada, S. Kimura, Y. Saiga, T. Nose, M. Yokote, and H. Kawatsu. 1971. Toxicity of mercury compounds to aquatic organisms and accumulation of the compounds by the organisms. Bull. Freshwater Fish. Res. Lab. 21:197–227. Matsumoto, E. 1988. Residence times of trace metals and nutrients in Tokyo Bay water. Oceanic Processes Mar. Pollut. 5: 211–218. Matthiessen, P., R. Waldcock, J. E. Thain, M. E. Waite, and S. Scrope-Howe. 1995. Changes in periwinkle (Littorina littorea) populations following the ban on TBT-based antifoulings on small boats in the United Kingdom. Ecotoxicol. Environ. Saf. 30:180–194. Méndez, N. 2002. Preliminary observations on the effect of cadmium on larval development of Capitella sp. B from Barcelona. Bull. Mar. Sci. 70:899–908. Morel, F. M. M., A. M. L. Kraepiel, and M. Amyot. 1998. The chemical cycle and bioaccumulation of mercury. Annu. Rev. Ecol. Syst. 29:543–566. Mouneyrac, C., and 7 others. 2003. Trace metal detoxiﬁcation and tolerance of the estuarine worm Hediste diversicolor chronically exposed in their environment. Mar. Biol. 143:731–744. Nakamura, K., T. Fujisaki, and H. Tamashiro. 1986. Characteristics of Hg-resistant bacteria isolated from Minamata Bay. Environ. Res. 40:58–67. Nishigaki, S., and M. Harada. 1975. Methylmercury and selenium in umbilical cords of inhabitants of the Minamata area. Nature 258:324–325. Nriagu, J. O. 1978. Biogeochemistry of Lead in the Environment. Elsevier/North-Holland Biomedical Press, New York. Nriagu, J. 1996. A history of global metal pollution. Science 272:223–224. Otero, X. L., J. M. Sanchez, and F. Macías. 2000. Bioaccumulation of heavy metals in thionic ﬂuvisols by a marine polychaete: The role of metal sulﬁdes. J. Environ. Qual. 29:1133–1141. Owens, M., and J. C. Cornwell. Sedimentary evidence for decreased heavy-metal inputs to the Chesapeake Bay. Ambio 24:24–27. Pacyna, J. M., E. G. Pacyna, F. Steenhuisen, and S. Wilson. 2003. Mapping 1995 global anthropogenic emissions of mercury. Atmosph. Environ. 37:S109–S117. Phillips, D. J. H., and P. S. Rainbow. 1993. Biomonitoring of Trace Aquatic Contaminants. Elsevier Applied Science, London.

Rainbow, P. S. 2002. Trace metal concentrations in aquatic invertebrates: Why and so what? Environ. Pollut. 120:497–507. Rees, H. L., R. Waldock, P. Matthiessen, and M. A. Pendle. 2001. Improvements in the epifauna of the Crouch estuary (United Kingdom) following a decline in TBT concentrations. Mar. Pollut. Bull. 42:137–144. Rolfhus, K. R., and W. F. Fitzgerald. 2001. The evasion and spatial/temporal distribution of mercury species in Long Island Sound, CT-NY. Geochim. Cosmochim. Acta 65:407–418. Rozan, T. F., and G. Benoit. 2001. Mass balance of heavy metals in New Haven Harbor, Connecticut: Predominance of non-point sources. Limnol. Oceanogr. 46:2032–2049. Ruiz, J. M. 1999. Bivalves, tributyltin and green tides: Ecosystem-level impact? Mar. Ecol. 20:1–9. Salomons, W., and U. Förstner. 1984. Metals in the Hydrocycle. Springer-Verlag, Berlin, 349 pp. Santos, M. M., N. Vieira, and A. M. Santos. 2000. Imposex in the dogwhelk Nucella lapillus (L.) along the Portuguese coast. Mar. Pollut. Bull. 40:643–646. Sañudo-Wilhelmy, S. A., K. A. Olsen, J. M. Scelfo, T. D. Foster, and A. R. Flegal. 2002. Trace metal distributions off the Antarctic Peninsula in the Wedell Sea. Mar. Chem. 77:157–170. Sañudo-Wilhelmy, S. A., A. Tovar-Sanchez, N. S. Fisher, and A. R. Flegal. 2004. Examining dissolved toxic metals in U.S. estuaries. Environ. Sci. Technol. 38:34A–38A. Selck, H., V. E. Forbes, and T. L. Forbes. 1998. Toxicity and toxicokinetics of cadmium in Capitella sp. I: Relative importance of water and sediment as routes of cadmium uptake. Mar. Ecol. Prog. Ser. 164:167–178. Shen, G. T., E. A. Boyle, and D. W. Lea. 1987. Cadmium in corals: Chronicles of historical upwelling and industrial fallout. Nature 328:794–796. Shulkin, V. M., V. Y. Kavun, A. V. Tkalin, and B. J. Presley. 2002. The effect of metal concentration in bottom sediments on the accumulation of metals by the mytilids Crenomytilus grayanus and Modiolus kurilensis. Russ. J. Mar. Sci. 28:43–51. Sidharthan, M., K. S. Young, L. H. Woul, P. K. Soon, and H. W. Shin. 2002. TBT toxicity on the marine microalga Nannochloropsis oculata. Mar. Pollut. Bull. 45:177–180. Sileo, L., and 8 others. 2001. Lead poisoning of waterfowl by contaminated sediment in the Coeur d’Alene River. Arch. Environ. Contam. Toxicol. 41:364–368. Slaveykova, V., K. J. Wilkinson, A. Ceresa, and E. Pretsch. 2003. Role of fulvic acid on lead bioaccumulation by Chlorella kesslerii. Environ. Sci. Technol. 37:1114–1121. Smith, W. E., and A. M. Smith. 1975. Minamata. Holt, Rinehart & Winston, New York, 192 pp. Sorensen, E. M. B. 1991. Metal Poisoning in Fish. CRC Press, Boca Raton, FL, 374 pp. Stull, J., R. Baird, and T. Heesen, T. 1988. Relationship between declining discharges of municipal wastewater contaminants and marine sediment core proﬁles. Oceanic Processes Mar. Pollut. 5:23–32.

METALS

Sunda, W. G. 1987. Neritic-oceanic trends in trace metal toxicity to phytoplankton communities. Pp. 19–29 in Capuzzo, J. M., and D. R. Kester (eds). Oceanic Processes in Marine Pollution, Vol. 1. Urban Wastes in Coastal Marine Environments. R. E. Krieger Publishers Co., Malabar, FL, 264 pp. Svavarsson, J. 2000. Imposex in the dogwhelk (Nucella lapillus) due to TBT contamination: Improvement at high latitudes. Mar. Pollut. Bull. 40:893–897. Sze, P. W. C., and S. Y. Lee. 2000. Effects of chronic exposure on the green mussel Perna viridis. Mar. Biol. 137:379–392. Tam, N. F. Y., and Y. S. Wong. 1993. Retention of nutrients and heavy metals in mangrove sediment receiving wastewater of different strengths. Environ. Sci. Technol. 14:719–729. Tam, N. F. Y., and Y. S. Wong. 1994. Nutrient and heavy metal retention in mangrove sediment receiving wastewater. Wat. Sci. Tech. 29:193–200. Tam, N. F. Y., and Y. S. Wong. 1999. Mangrove soils in removing pollutants from municipal wastewater of different salinities. J. Environ. Qual. 28:556–564. Tavecchia, G., R. Pradel, J.-D. Lebreton, A. R. Johnson, and J.-Y. Mondain-Moval. 2001. The effect of lead exposure on survival of adult mallards in the Camargue, Southern France. J. Appl. Ecol. 38:1197–1207.

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Tomiyasu, T., and 6 others. 2000. Mercury contamination in the Yatsushiro Sea, South-western Japan: Spatial variations of mercury in sediment. Sci. Tot. Environ. 257:121–132. Tuncer, G., G. Tuncel, and T. I. Balkas. 2001. Evolution of metal pollution in the Golden Horn (Turkey) sediments between 1912 and 1987. Mar. Pollut. Bull. 42:350–360. Vasconcelos, M. T. S. D., and M. F. C. Leal. 2001. Seasonal variability in the kinetics of Cu, Pb, Cd, and Hg accumulation by macroalgae. Mar. Chem. 74:65–85. Wang, W.-X. 2002. Interactions of trace metals and different marine food chains. Mar. Ecol. Prog. Ser. 243:295–309. Watras, C. J., and N. S. Bloom. 1992. Mercury and methylmercury in individual zooplankton: Implications and bioaccumulation. Limnol. Oceanogr. 37:1313–1318. Weis, J. S., P. Weis, M. Heber, and S. Vaidya. 1981. Methylmercury tolerance of killiﬁsh (Fundulus heteroclitus) embryos from a polluted vs. a non-polluted environment. Mar. Biol. 65:283–287. Wong, P. T. S., B. A. Silverberg, Y. K. Chau, and P.V. Hodson. 1978. Lead and the aquatic biota. Pp. 279–342 in Nriagu, J. O. (ed.). The Biogeochemistry of Lead in the Environment. Part B. Elsevier/North Holland Biomedical Press, Amsterdam, 397 pp. Yeats, P. A., and J. M. Brewers. 1982. Discharge of metals from the St. Lawrence River. Can. J. Earth Sci. 19:982–992.

Chapter 10 Introduction of exotic species

Use of an alien species in the management of estuarine areas. Top: view of a mudﬂat and marsh area in the southern Netherlands where the British cordgrass, Spartina townsendii, was purposely introduced in 1924. Middle: the same area in 1938, after extensive colonization and sediment trapping by S. townsendii. Bottom: the same area (photo taken from the extreme right hand corner of the views in the top and middle panels), after the erstwhile marsh mudﬂat area was ploughed up in preparation for agricultural use in 1950. Photos by Professor J. W. Oliver, reproduced in Ranwell (1967).

INTRODUCTION OF EXOTIC SPECIES

Two case histories Invasions into San Francisco Bay The history of the fauna and ﬂora of San Francisco Bay is an unending tale of species replacement. The bivalve species are but one of the elements that have been changed by successive invasions and introductions (Nichols et al. 1986; Herbold et al. 1992). All but two of the common bivalves in the bay were introduced (Nichols & Pamatmat 1988); in all, 234 exotic species have been recorded within the San Francisco estuary (Carlton & Geller 1993). San Francisco Bay must have always supported large populations of bivalves: pre-Columbian middens of California Indian tribes included abundant shells of native oysters (Ostrea lurida), a species whose ﬂavor was not favored by new settlers, who introduced eastern oysters (Crassostrea virginica) for commercial harvest in the late 1800s. The eastern oyster did not reproduce in the bay, so repeated imports of seed oysters were needed, which brought a series of hitchhiker species into the bay. In the 1930s the Paciﬁc oyster (Crassostrea gigas) was introduced from Japan to boost the sagging harvests of the eastern oyster. Bent-nose clams (Macoma nasuta) were initially exploited, but the abundance of these native clams diminished in parallel to the introduction in 1874 of the eastern soft-shell clam (Mya arenaria), a species that proliferated and yielded huge harvests. Overharvest, habitat loss, increasing pollution, and probably competition from Japanese little-neck clams (newly introduced with Paciﬁc oysters during the 1930s) led to the disappearance of the soft-shell harvest by 1949. In 1946 the Asiatic clam entered the bay, and since then it has become the dominant benthic mollusk in the less saline reaches of the bay. As trade across the Paciﬁc rim increased during the 20th century, new opportunities for alien introductions rose. The Asian clam (Potamocorbula amurensis) was ﬁrst noticed in the San Francisco Bay in 1986, and quickly spread widely throughout the bay, probably aided by the neareradication of benthic invertebrates by a powerful

227

storm during February 1986 (Nichols et al. 1990). The Asian clam spread to all sediment types, and even reached into the intertidal range. The succession of invasions has continued. By 1992, San Francisco Bay included 19 exotic taxa among its 30 species of mollusks (Carlton 1992). The story of invasion goes on, however: more recently, there were reports that another exotic, the green crab, was successfully expanding its range within the San Francisco Bay (Cohen et al. 1995). The green crab is a voracious feeder on bivalves, and hence could potentially control populations of the Asian clam as the crab extends its range across San Francisco Bay (Grosholz & Ruiz 1995).1 The reduction of abundance of Asian clams by the green crab has had, in turn, further ecological consequences, since predation by the green crabs may lower food supplies for benthicfeeding shorebirds that winter along the Paciﬁc shore (Grosholz & Ruiz 1999). Many species of ﬁsh have been purposely introduced to the San Francisco estuary to improve ﬁsheries; of these, 30 have successfully colonized (compared to 27 native species) (Cohen & Carlton 1998) and another, the yellow perch, introduced in 1881, became extinct in the 1950s (Herbold & Moyle 1989). The most prominent ﬁsh in the bay is the striped bass, introduced from the Atlantic coast in 1871 to improve ﬁshing harvests. The American shad, also from the Atlantic coast, was purposely introduced before the 1880s. Both of these species became abundant enough in the 1900s to become important commercial species. The success of these two exotics might be in part tied to the tolerance their eggs have for suspended sediments in the water of the smaller streams where they breed. Mining and erosion added sediment loads to streams throughout California as the area became populated. The native species such as salmon require clear water to breed successfully in the streams, and hence were readily replaced by the turbidity-tolerant aliens. 1

This is analogous to the destruction of exotic zebra mussels in Lake Michigan by a sponge that smothers the mussels (Early & Glonek 1999). Even though, initially, established exotics might have escaped their own predators or diseases, they may eventually become susceptible to other predators or diseases.

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Two additional exotics were probably introduced in discharged ballast water from ships coming from Asia. Several other exotic ﬁsh have become established in the estuary, and several native species have become rare, and two have disappeared during the same period (Herbold & Moyle 1989). Bivalves, macrophytes, and ﬁsh are far from the only invaders of the San Francisco estuary. For instance, when eastern oysters were shipped to San Francisco Bay to bolster harvests, they brought with them an array of other invertebrates attached to their shells (Nichols et al. 1986). For the sake of brevity, I only note that dominant zooplankton (Herbold & Moyle 1989; Modlin and Orsi 1997), snails (Race 1982), crabs (Grosholz & Ruiz 1995), and jellyﬁsh (Greenberg et al. 1996) in the bay are all imported species. There are few corners of the food web of San Francisco Bay that have not been affected by invading species during the last two centuries. The list of species found there has been thoroughly altered by human intervention and the establishment of aliens. In spite of the invasions, the remarkable loss of habitats mentioned in Chapter 6, and many other perturbations, there is much biological activity in the bay and reasonably rich stocks of living populations. The residents might, much like the humans on the surrounding land, be mainly immigrants, but there are a lot of them there. Spread of smooth cordgrass The smooth cordgrass, Spartina alterniﬂora, is the principal plant of the salt marshes of the eastern coast of North America. During the 20th century, this grass has spread far beyond its native geographic range and invaded or was introduced on purpose, into Europe, the Paciﬁc coast of North America, and elsewhere. This example of an invasive species illustrates the proclivity for range expansion, and that invasions may involve considerable genetic and evolutionary change. The estuaries of the Paciﬁc coast of North America are geologically young. They are mostly river valleys drowned by the rising sea level that accompanied the melting of world glaciers since

10,000 years ago (Atwater et al. 1979). The salt marshes that developed in the shallows of these estuaries are themselves only a few hundred years old (Macdonald & Barbour 1974). The mudﬂats of these western estuaries were devoid of cordgrasses, except along the upper tidal reaches, where a native species, the rough cordgrass, S. foliosa, grew. S. foliosa is quite similar to the eastern S. alterniﬂora, except that the latter is far more tolerant of submergence (Daehler & Strong 1996) and can grow considerably taller. S. alterniﬂora was purposely introduced into several western estuaries in salt marsh restoration and land reclamation efforts2 during the mid-1970s, and has spread since to other sites. The spread of the new species was particularly strong in San Francisco Bay (Callaway & Josselyn 1992), where the two species hybridized, with hybrids showing more vigorous growth than the rough cordgrass, and a new ecotype of smooth cordgrass that more aggressively colonized mudﬂats (Daehler et al. 1999). The post-invasion growth of the invader species, and of the hybrid ecotypes, thus tends to convert bare mudﬂats to vegetated salt marsh. In addition, other species of cordgrass have been introduced to Paciﬁc estuaries. S. anglica invaded Puget Sound some time ago, and in 1977 transplants were brought to San Francisco Bay for marsh restoration purposes; S. densiﬂora, a South American species, was transplanted from Humboldt Bay, California to San Francisco Bay, and has spread; S. patens, an eastern species, was introduced into estuaries in Oregon and Washington (Daehler & Strong 1996). Seeds of S. alterniﬂora carried in ship ballast water entered Southampton water, on the southern coast of England, in the early 19th century (Thompson 1991). Once established, S. alterniﬂora hybridized with the native S. maritima, described

2

The use of marsh vegetation to foster sediment trapping and stabilize sediments, as well as accumulate sediments, is not new. In fact, there is a town in California called Reclamation, so named because land in its vicinity was gained by using rough cordgrass in the 19th century (Merrill 1902). The frontispiece to this chapter shows a similar effort at land reclamation using the S. townsendii hybrid in the Netherlands.

229

INTRODUCTION OF EXOTIC SPECIES

57.5 48

48 40° 23.5° 0°

Figure 10.1 World distribution of Spartina townsendii, as of 1965. The black circles show sites where plantings became established, and open circles show sites where plantings have failed. Modiﬁed from Ranwell (1967).

23.5° 35

40°

46

90°

0°

90°

as S. townsendii, which eventually led to an allopolyploid new species, S. anglica.3 As in the case of the Paciﬁc estuaries, English mudﬂats were largely devoid of vegetation, with the native species occupying only the upper reaches of the tidal range. After the appearance of S. anglica, this new species expanded into the lower elevations and covered erstwhile bare ﬂats. The potential for stabilizing sediments, and restoring coastal wetlands, of this species made it attractive in plantings in many other sites (see Chapter 10 frontispiece). As a result, English cordgrasses were widely introduced and became established in many parts of the world (Fig. 10.1). There are further changes that may prevent further expansion and perhaps even reduce dominance by exotics such as cordgrasses. For

instance, smooth cordgrass introduced into Willapa Bay, Washington (Dahler & Strong 1997) appears to have lost resistance to grazers after a century of herbivore-free growth in Willapa. Seed production has been lowered by infections of ergot fungus in most British populations of S. anglica (Gray et al. 1990).4 Perhaps such potential controls foreshadow a longer-term readjustment of dominance in these systems. Both in western North America, and in the UK, there have been expressions of concern regarding the many changes that the Spartina expansion may bring. These include alteration of water ﬂow, excessive sediment trapping, displacement of native plant species, alteration of invertebrate faunas, particularly shellﬁsh of commercial interest, and reduction of feeding areas for

3

4

The ﬁrst hybrid between the diploid S. alterniﬂora and S. maritima was recorded in 1872, and was sterile because the differences between the parental chromosomes prevented chromosome pairings during meiosis (Marchant 1968). This sterile form produced a fertile hybrid by a doubling of chromosome number, referred to as allopolyploidy, where chromosome pairings do occur, but the cell carries twice as many chromosomes (Gray et al. 1990; Raybould et al. 1991). The polyploid species has genetic complements from both parents, and thus the potential for greater vigor. The spread of the hybrids was slow initially, but increased in pace after 1890, presumably when the fertile hybrid arose and produced seeds that could be more readily dispersed (Thompson 1991).

There has been repeated mention of Spartina “die-back” in many older stands within British sites (Goodman et al. 1959). This die-back is more likely to be related to increased submergence (Goodman & Williams 1961) than to biological control by diseases or grazers. I note that it is the older stands that suffer most from die-back. My interpretation is that S. anglica initially invaded as much of the estuaries as the depth of submergence allowed, but as sea level rose, parts of the stands were submerged longer than the cordgrass tolerated, and hence these parts died back. Incidentally, salt marsh die-back has recently been reported throughout the eastern coast of the USA (Georgia Coastal Research Council 2004), but there is little information as to the mechanisms involved.

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shorebirds and waterfowl. This is not the place to assess the validity of all these claims of harmful effects of cordgrass expansion, but the preponderance of the evidence is not completely compelling.5 Nonetheless, there is widespread concern about the invasion of cordgrass in Paciﬁc states (Daehler & Strong 1996), with a recommendation to “. . . survey the vulnerable sites frequently and eliminate introduced Spartina sp. propagules before they spread” (Daehler & Strong 1996). It is not unusual in invaded Paciﬁc marshes, for instance, to ﬁnd managers, ﬁeld staff, and volunteers involved in Spartina control by bulldozing or plowing stands, hand removing shoots, spraying substantial doses of herbicides, covering stands with black plastic, and otherwise engaged in planticide. These efforts, though worksome, are likely in vain, since most invasions are irreversible (Ewel et al. 1999). Those of us on the Atlantic coast of North America witness the western reactions to cordgrass with a certain incomprehension. On the eastern coast of North America, S. alterniﬂora has been the unquestioned avatar of conservation of coastal wetlands, based on studies done beginning in the 1950s and continuing to today, as noted in Chapter 6. Nearly every law that protects wetlands in the eastern coastal states of the USA and Maritime Provinces of Canada departs from the premise that this grass is the singlemost important component of coastal wetlands and that its protection must be guaranteed. The litany of reasons why coastal wetlands are important that were listed Chapter 6, were based

on the performance of ecological services by cordgrass. Eastern estuaries are replete with cordgrass, and yet these salt marsh-dominated estuaries are renowned for their high yields of shellﬁsh, for harboring huge numbers of resident and migrant shorebirds and waterfowl, and for contributing to ﬁsheries by serving as nurseries for other species, among other beneﬁts. Thus, there is a remarkable discrepancy as to how cordgrass is viewed on either side of the continent: an essential item to be strictly protected on the East Coast, a destructive invader in need of eradication on the West Coast.6 Ecosystems of eastern and western estuaries are not that different; the same ecological and biogeochemical processes necessarily operate in both. The two contrasting views regarding cordgrass seem incompatible. It seems timely to revisit the issues with a critical view, and carry out the necessary studies to evaluate these disparate views. The multiplicity of site invasions by cordgrass, accidental or purposive, attests to the plasticity of certain invasive species. Cordgrass managed to be present in so many places in part through its genetic, as well as ecological and physiological, ﬂexibility. It, however, is quite a specialist, surviving in sites where few other plants can establish a foothold. This is a case of special contingencies that prompted adaptations to a drastic environment (variation from fresh to salty water, anoxic sediments, and so on). These adaptations then in a counterintuitive fashion poise the species as invaders of “empty” environments.

5

We can take the issue of effects on birds as an example. First, there are some bird species that can certainly use the vegetated marsh habitat, for instance, rails. Second, the areas of estuary that have been converted to vegetated habitats may not be large; in English estuaries Goss-Custard and Moser (1988) show that at most 10% of the area was converted. Third, the linkage of reduction of shorebird abundance to cordgrass increase is at best weak and correlational. Moreover, the data suggest that implausibly tiny changes in percentage of the area of mud converted to marsh (of the order of 1%, see ﬁg. 5 in Goss-Custard & Moser 1988) account for virtually the whole loss of shorebird abundance. This seems ecologically implausible. Fourth, other work suggests that bottom faunas of sediments colonized by S. alterniﬂora are richer and more productive of animals that might become food items for predators (Da Cunha Lana & Guiss 1991).

6

A parallel divergence of views arises in regard to another invasive wetland plant, the common reed, Phragmites australis. Concern has been shown about its recent spread in the eastern USA (Chambers et al. 1999), because of its deleterious effects on wetlands and their faunas (but see Meyerson et al. 2000). On the other side of the Atlantic, authors bemoan the loss of stands of Phragmites because of the loss of wildlife habitat (Fogli et al. 2002). Able and Hagan (2000) suggest that the effects of the replacement of cordgrass by reeds might be modest for the fauna on the marsh surface. A 2003 issue of the journal Estuaries, dedicated to the common reed problem, was subtitled “A sheep in wolf’s clothing?” A review of many publications about reeds as invaders concludes that the effects of Phragmites invasions tend to have minor impacts (Weis & Weis 2003).

INTRODUCTION OF EXOTIC SPECIES

200

• Spread of early humans out of Africa into Europe and Asia. • Epidemics of Black Plague in Europe, 1348– 1841. • Introduction of Old World diseases in the Americas after 1492. • Conversion of prairies to agricultural crops in the breadbaskets of the world after the 1800s. • “Spanish” ﬂu epidemic in North America, 1919.

Number of invasions

Common features of biological invasions

All of these are but a few of the innumerable examples of biotic invasions by alien species;7 all of them were prompted by some human mediation, demographic, political, or economic; and all of them had impressive consequences, both for the natural environments being invaded, and for the humans involved. During their short history on earth, humans have characteristically altered their environment, in large measure by moving species around the globe. This is what we have done, and what we continue to do. It is the consequence of growing numbers of people, and of human activities. Some of the introductions were purposive, to further food yields, for instance, but others were accidental. All environments are continuously subject to the arrival of propagules of alien species (Fernando 1991). Human beings, however, have aided the arrivals unknowingly or purposely by introducing species with desirable properties— as crops, for the management of water quality, or as biological controls of undesired organisms. The issue of biological invasions of all environments has been viewed with alarm in recent years, and many authors have been concerned with the environmental and economic impact of introduced exotic species (Carlton 1996c; Williamson 1999; Mack et al. 2000; Pimentel et al. 2000). Moreover, rates of invasion appear to be increasing: alien species on the shores of North

Figure 10.2 Number of invasions of North American coasts by species of macroinvertebrates and macroalgae, with data compiled in 30-year intervals. Data from Ruiz et al. (2000a).

7

Species that evolved elsewhere and ﬁnd themselves in a new geographic range are referred as aliens, exotics, non-indigenous, introduced species, or invaders. I use the terms interchangeably in this chapter.

231

100

0 1790 1820 1850 1880 1910 1930 1970 1999 Year

America, for example, have increased since the 1980s (Fig. 10.2). Coastal environments are commonly subject to changes in taxonomic composition owing to invasions of species from elsewhere. The invasion of the Black Sea by a species of comb jelly from the Northwest Atlantic during the 1980s, already noted in Chapter 1, is but one impressive example of a common pattern in coastal environments. Although the instance of San Francisco Bay may be a worst-case example, frequent occurrences of exotic species are not unusual in coastal environments elsewhere. For example, the range of the European green crab has expanded to Monterey Bay, California, and to Coos Bay in Oregon (Grosholz & Ruiz 1995). In Willapa Bay, Washington, there were 12 exotic mollusk species; in Puget Sound, Washington, there were 11; and in Boundary Bay, British Columbia, there were 13 exotic species of mollusks (Carlton 1992). An additional and illustrative instance is the Asiatic clam population that was introduced into the Potomac estuary in Chesapeake Bay (Phelps 1994). Abundance of the clam increased from the mid-1970s to the mid-1980s (Fig. 10.3 top). Densities reached 1,000–8,000 individuals m−2. This species made up to 95% of invertebrate biomass in certain parts of the San Francisco estuary (Thompson & Schemel 1991). Where

232

CHAPTER 10

Clam biomass (kg m–2) ( )

CLAMS AND SAV 3 1

2 1 0 1976

0 1980

1984

1988

1992

1988

1992

SAV area (km2 × 0.1) ( )

2

4

40 Number of birds

BIRD ABUNDANCE 30 20

Black crowned night heron Mallard duck Great blue heron

10 0 1976

1980

1984

Number of birds

800 600

BIRD ABUNDANCE Ring-necked duck Bufflehead Canvasback

400 200 0 1976

1980

1984 Year

1988

Asiatic clams increased, abundance of other bivalves diminished. After a 2 year lag, the area of bottom covered by submerged aquatic vegetation (Fig. 10.3 top), and the abundance of aquatic birds also increased (Fig. 10.3 middle and bottom). Fish abundance also increased in parallel with the increased area of submerged vegetation habitat (Phelps 1994). The key mechanisms seemed to be that the Asiatic clam population effectively increased water clarity by ﬁltering particles out of the water column, and hence allowed sufﬁcient light to reach the bottom. This fostered vegetation, and generated favorable habitats for aquatic birds. Similar examples of an

1992

Figure 10.3 Top: time course of the invasion of brackish environments of Chesapeake Bay by the Asiatic clam, and the following changes in the area of submerged aquatic vegetation (SAV). Middle and bottom: changes in the abundance (number of individuals in Christmas counts done in December) of different species of aquatic birds. Data from Phelps (1994).

invasive species improving water quality by lowering water turbidity have been reported in the case of the zebra mussel (Leach 1993; Phelps 1994), and the Asian clam in San Francisco Bay (Table 10.1) and in Chesapeake Bay (Newell 1988).8 8

The lowering of water turbidity by dense populations of bivalves is well established (Riemann et al. 1988; Dame et al. 1991; Leach 1993), and is even proposed as a management option for improving water conditions (Gottlieb & Schweighofer 1996). Where poor water quality and disease have eliminated native eastern oysters, stocking of the Japanese oyster may re-establish populations of suspension feeders, and is suggested as a means to clear the water column.

INTRODUCTION OF EXOTIC SPECIES

Table 10.1 Reduction of phytoplankton biomass and annual primary production during the years before and after 1986, when the Asian clam became well established in the San Francisco Bay. Data from Alpine and Cloern (1992).

Phytoplankton biomass (mg chlorophyll m−3) Phytoplankton production (g C m−2 yr−1)

Pre-1986

Post-1986

30–40

3

110

20

In coastal environments of North America, Ruiz et al. (2000a) reported that 298 nonindigenous species of invertebrates and algae have become established in at least one coast area. An additional 76 species have managed to colonize more than one coast, so that there are a total of 374 species of recognized, established cases of invasions by invertebrates and algae. Although these are signiﬁcant numbers, alien species are relatively few compared to the number of species present in most environments.9 Ruiz et al. (2000a) record more invader species in coastal sites than in freshwater sites, but are coastal environments really more invaded than other environments, as a percentage of the species present? The percentage of alien species in ﬁsh faunas of freshwater lakes may be large: 44% in Italy, 35% in Portugal, 33% in Spain, 12% in Greece (Elvira 1995), and 8% in the USA (Courtenay et al. 1984). Alien species make up considerable proportions (10–30%) of terrestrial ﬂoras (Mills et al. 1993). Probably the most invaded is New Zealand, with 47% of its ﬂora made up of exotic species (Heywood 1989). Elsewhere, alien plants make up smaller but 9

Most studies of biodiversity underestimate species numbers. In long-term studies of a Massachusetts salt marsh—an environment that is usually thought to be taxonomically relatively simple—we have recorded well over 400 species of invertebrates on the vegetated sediments, more than 100 species on the vegetation itself, about 50 species of macrofauna and more than 100 meiofauna species on bare sediments, 27 species of vascular plants, and many species of algae, ﬁsh, and birds.

233

signiﬁcant portions: 11% in California (Mooney et al. 1986), a perhaps too-large estimate of 10.8% in the USA in general, and 22% in Canada seem to be exotic (Vitousek et al. 1996). Eight percent of insect species in Florida are exotic (Frank & McCoy 1995b). Such percentages show that assemblages of taxa in most environments suffer from a considerable, and continuous, inﬂux of alien species. What we see when we do a survey at one time is really a snapshot of “nature”, a sample of what is really a series of shifting assemblages, continually altered to some degree as time passes by a complement of itinerant taxa. Only a select few invading taxa establish a sufﬁcient foothold to become important restructurers of the assemblage of species present. The new arrangements emerge, only to be further changed as a few more new immigrants add to the mix. It seems evident that the widely held idea of a “balance” of nature might need to be modiﬁed. Below I review mechanisms favoring such frequent invasions, and go on to evaluate limitations in the interpretation of invasion data, assess the magnitude of invasions in coastal environments, and discuss patterns and effects of invasions.

Mechanisms of invasions The case histories reviewed above provide evidence of the major transport mechanisms supporting the pace of invasions into coastal environments; these can be subsumed into a few categories (Table 10.2). Of course, there are many stragglers and accidental transports of species around the world, caused by near-random acts of fate, such as storms, winds, and misdirections of one kind or another. Every year, for example, in my own area of Massachusetts, birders expect to see several species of birds that are normally not in this area: such accidental sightings are a regular feature of any natural environment. Natural assemblages normally are exposed to a low, but continual, number of new species, and only a few of these stragglers become established. Here, though, I focus on anthropogenic

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Table 10.2 Percentages of alien species (ﬂora and fauna) introduced by various mechanisms into British (Eno et al. 1997), Mediterranean (Boudouresque 1999), and North American (Ruiz et al. 2000a) coasts. No. of species Mechanism of entry Shipping Ballast Fouling Fisheries Stocking, bait Mariculture Biocontrol Multiple mechanisms Accidental releases Research Canals Unknown

Britain

Mediterranean

18 25

2 13

51

8 31

1 8

16

12

mechanisms that hasten rates of transport of alien species. Shipping Many recorded coastal invaders have been transported by ships, either attached to their hulls or in ballast water. In general, ballast water seems the major pathway of invasion (Table 10.2). Ballast water is taken up by ships to maintain appropriate buoyancy and balance. Some ships have dedicated ballast tanks, some load ballast water into empty cargo holds. The ballast water may be carried for long distances, and is emptied—most often near the shore—before the holds can be reﬁlled with new cargo. This collection and release of sea water containing live propagules has promoted high rates of organism exchanges. Live larvae and plankton are taken in to ships with ballast water, and even though there is considerable mortality during the trip, many kinds of organisms survive long periods in ballast water. For example, diatoms, protozoans, and copepods may last up to 10 months; dinoﬂagellates 5 months; and ﬂagellates, ciliates, nematodes, larvae of worms, bivalves, and barnacles, daphnids, and mites 1–2 months (Cohen 1998). Ballast importation is of course not the only route of invasion. In many ports of the

Decreased 67 7

North America

6 26 4 0.3 0.6

world, annual volumes of ballast water have diminished in the last 30 years (Cohen 1998), even as the rate of invasions might have increased (see Fig. 10.2). Some sites, such as Chesapeake Bay, receive more ballast water than San Francisco Bay, yet host far fewer alien species (Ruiz et al. 1998). Fisheries Many alien species enter coastal environments as hitchhikers in shipments of stocks of bivalves, particularly oysters. Some alien ﬁsh and shellﬁsh have been introduced as imported shellﬁsh for mariculture, stocking of natural environments, or for use as bait. Biocontrol Plants have been imported to help in marsh restoration or creation, and for stabilizing eroding shores. Certain animal species have been imported as a means of controlling undesired alien species. Accidental releases Hobbyists, museums, and aquacultural establishments have more than once accidentally released

235

INTRODUCTION OF EXOTIC SPECIES

exotic species, and a few of these escapees may become established. Wellcome (1988) concluded that accidental releases of ﬁsh species may contribute about 10% of established alien species in fresh waters. There is no better example of this mechanism than the establishment and spread of a green alga, Caulerpa taxifolia, in coasts of the Mediterranean Sea. C. taxifolia is a green alga from warm tropical coastal waters. In its normal geographic range, it is usually found growing in clumps of modest dimensions. It was displayed in the aquaria of the Wilhemina Zoo in Stuttgart, Germany, from which live fronds were sent to other institutions, including the Monaco Oceanographic Museum. Apparently, during cleaning of exhibit tanks before 1984, some bits of C. taxifolia were released accidentally onto the Mediterranean shore by the Monaco Museum. Sometime during the life of these captive C. taxifolia, some genetic change had made it possible for the alga to survive colder temperatures (Ballesteros 1989), as well as to grow far more vigorously (Rodriguez-Prieto 1999). The escaped strain only reproduces vegetatively (Meinesz & Boudouresque 1996), possesses strong grazer deterrents (Boudouresque et al. 1996), and can obtain nutrients from sediments—an unusual feature for algae (Chisholm et al. 1996). From a patch of only 1 m2 discovered in front of the Monaco Oceanographic Museum in 1984, the range of C. taxifolia has spread to thousands of hectares, spanning sites across a substantial part of the northern Mediterranean coast (Fig. 10.4). The spread of C. taxifolia has prompted concern with the potential changes it might bring about (Boudouresque 1999; Rodriguez-Prieto 1999), including alterations to other algae and seagrasses, benthic invertebrates, and ﬁsh, simpliﬁcation of assemblages of species, and lowered attractiveness to ecotourism. Compelling data are scarce. In addition, there are other congeners; C. racemosa, a complex of subspecies, may have been an import from elsewhere, perhaps from the Atlantic, perhaps from the Red Sea (Verlaque et al. 2000), and was ﬁrst recorded in 1924. It has since appeared all over the Mediterranean. This invasive species has therefore spread widely, but has not been recorded as

1984 1990 1991

1992

1992 1992 1993

1995 1995

1992 1993

Figure 10.4 Spread of Caulerpa taxifolia in the northern Mediterranean coast after its accidental release from Monaco Oceanographic Museum in 1984. Black circles show records of presence, with the year it was recorded. Data compiled by Dini (1999).

having deleterious long-term effects. It seems difﬁcult to say whether C. taxifolia will be more consequential, or will merely follow the history of its congener. Canal construction The joining of two water bodies by a new canal has provided the opportunity for organism exchanges. In particular, the Mediterranean has been subject to Lessepsian10 invasions from species on the Red Sea side of the Suez canal (see Table 10.2). In the Mediterranean there are considerable numbers of Lessepsian aliens, particularly nearer the Suez Canal (Fig. 10.5). Canals did not appear prominently in the North American data of Table 10.2 because I omitted the Great Lakes; had these been included, we would have found several canal-introduced species (e.g. sea lamprey, zebra mussel) that have had a prominent history of aggressive and consequential invasions. 10

So named (Por 1990) after Ferdinand Marie, Vicompte de Lesseps, a French diplomat and engineer (b. 1805, d. 1894) that conceived, obtained the concession and funds, and organized the construction of the Suez Canal, 1859–1869.

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2

1 28 0

0 4

1 10

0 1

7 2

6 7

2 6 5

2 6

2 3

10

7 2

3 4

1 3

15 0

0 18

Figure 10.5 Numbers of Lessepsian (i.e. trans-Suez Canal species, indicated in bold) and non-Lessepsian (in italic) species of macrophytes in different areas of the Mediterranean Sea. Data compiled by Boudouresque (1999).

Multiple mechanisms Many known aliens may have had multiple mechanisms of invasion, and often have invaded in more than one place. Shipping and mariculture are the major contributors to multiple invasions. The zebra mussel in particular was transported by multiple means, from its Caspian Sea original range, ﬁrst across land, then the Black Sea, the Mediterranean, and the Atlantic, probably via shipping, and then into the Great Lakes via canals.

Limitations of the interpretation of invasion data It is difﬁcult to evaluate present circumstances regarding invasions of coastal environments because species replacements have always occurred

in all environments, available information is biased toward large evident species, and we are not always sure which species have exotic origins. Species composition inevitably changes through time in all environments (Sprugel 1991). Even a cursory examination of the fossil record of any geological deposit will show great secular shifts in the species present. Vermeij (2001), for instance, compiled data from mollusk assemblages found in various formations, and found that 28–82% of the species became extinct from the early to middle Pliocene. Historical data on the disappearances and expansions of species are also common. The American chestnut, historically one of the most dominant tree species in eastern US forests, and the passenger pigeon, certainly an overwhelmingly abundant bird of these forests, disappeared in past decades because of an alien blight and from overhunting.

237

INTRODUCTION OF EXOTIC SPECIES

100

80

60

Helminths

Sponges

Protozoans

Bryozoans

Chordates

Cnidaria

Insects

Algae

Annelids

20

Mollusca

40 Crustaceans

Number of species

Red maple has expanded its range across these same environments, probably because forest ﬁres are less frequent (Abrams 1998). Thus, disappearances and expansions of species in natural environments seem the rule rather than the exception. We should not expect any different circumstances in the coastal zones. The questions that emerge are whether coastal environments more subject to invasion than others, and whether humans have driven invasion rates beyond the “natural background” levels of invasion. In many cases it is not always clear whether a species is an exotic or not. Carlton (1996a) coined the term “cryptogenic” (unclear origin) to refer to the many species with unknown geographic origin. One hundred and twenty-ﬁve species in San Francisco may be cryptogenic, and in Chesapeake Bay the number of aliens may be only one-third of the number of species with no known origin. These large proportions impair efforts to identify common features of successful invaders as well as in the assessment of the severity of the alien problem. Assessments of the magnitudes of exotic invasions are further hampered by a bias in the detection of invaders. Studies of alien invertebrate and vertebrate species in coastal environments are strongly skewed toward large, evident species (Ruiz et al. 2000a). To the extent that data are available, crustaceans and mollusks seem to be the most prominent taxa of invaders and smaller, less obvious taxa appear less often as invaders (Fig. 10.6). There is very little known about the transport of microbial organisms. Bacteria, including cholera organisms, and virus-like particles are abundant in ballast water (Ruiz et al. 2000b). Toxic dinoﬂagellates may also be transported by ballast water (Hallegraeff 1998). The identiﬁcation of microbial species is more difﬁcult than in the case of larger organisms, and microbial taxa are also widely distributed across the world, both circumstances that make detection of true microbial invaders even more difﬁcult than with larger species. These difﬁculties in interpretation of the record of invasion make it difﬁcult to assess the degree to which we might be concerned with the issue. The difﬁculties, however, do not obscure the fact

0 Alien species

Figure 10.6 Number of alien species of different groups of macroinvertebrates and macroalgae established along the coasts of North America. Data compiled by Ruiz et al. (2000a).

that in a few selected places, and with a few strongly invasive species, invasions have had a signiﬁcant impact on the species composition of coastal communities and waters. It may also be difﬁcult to separate the effects of invasion from those of other agents of coastal change, chieﬂy eutrophication (Fernando 1991; Phelps 1994) and global change (Simberloff 2000). As discussed in Chapter 1, the invasion of the Black Sea by a comb jelly Mnemiopsis leidyi from the Atlantic coast of the Americas, is associated with a loss to ﬁsheries of US$250 million, but it is not certain how much of the responsibility for the loss may be due to the invasion, to overﬁshing, or to the eutrophication that took place synchronously (Harbison & Volovik 1994). We may therefore ﬁnd it difﬁcult to blame exotic species for the entirety of the alterations to recipient communities.

Patterns of invasion in coastal environments The dominance of coastal environments by speciﬁc aliens seems characteristically transient.

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The abundance of the Asiatic clam in the Potomac was ephemeral: by 1994 the situation resembled pre-invasion conditions (see Fig. 10.3 top). We have already referred to several cases where invader populations may be threatened by the emergence of new predators or competitors. Predation by green crabs may eventually lower the dominance of Asian clams in saline areas of San Francisco Bay. Sponges may attack zebra mussels, and grazers may feed on cordgrass. Native ﬁsh seem to have returned to Gatún Lake, Panama, where a predaceous exotic had previously eliminated many of the indigenous ﬁsh species (Wellcome 1988). The long-term pattern of exotic invasions in coastal environments therefore seems to be one of repeated species ﬂux. Invasion, expansion, and subsequent decline, perhaps by replacement, perhaps by predation or disease by other organisms, is followed by another invasion cycle, and each new invasion has consequences that propagate throughout coastal food webs. There are active, complex food webs throughout the replacements, but they just happen to be carried out by a continually changing cast of species. Texts (for example, Shigesada & Kawasaki 1999) tend to describe invasions by successful invaders, rather than to explain underlying mechanisms. We lack theory that might predict which taxa are likely to be vicariant, and become established, or where invaders might become established. We do not understand the wide disparity in invasiveness and invasibility among successful aliens; why, for example, are some cordgrass species so much more invasive than their closely related congeners? Why is the zebra mussel a successful but innocuous invader in Europe, but has become a major agent of change in North American fresh waters (Williamson 1999)? There is a near-random and quite low chance that any species will be transported beyond its range. By far the majority of those species that by chance manage to become new arrivals fail to establish themselves, and only a minor portion of established aliens become agents of substantial ecological change in their new environments (Williamson 1999; Mack et al. 2000). Of the about

25,000 exotic plant species in cultivation in the State of Florida, about 925 have established themselves in the wild (Frank & McCoy 1995a). In Hawaii, 908 species out of the 4,275 cultivated species (not including accidental introductions) have become established in the wild (Eldredge & Miller 1998). Of the 4,500 alien species that have become established in the USA, only 15% have become a problem (OTA 1993). Only about a third of the taxa imported as biological control agents of pests become established, and about a third of these actually have an effect on the target species (Sheppard 1992; Williamson 1999). These random and low probability features make it difﬁcult to discern patterns of proclivity to invasiveness. Similarly, we cannot predict what environments are more likely to be invaded. It seems reasonable to think that species-poor, more disturbed environments might be more likely to be invaded (Baskin 1998; Davis et al. 2000). Pristine, species-rich environments do not seem, however, disproportionately devoid of invaders (Baltz 1991; Carlton 1996a; Case 1996; Goodwin et al. 1999; Williamson 1999). Biogeographic theory suggests that islands may be less resistant to invasions than continents, but this is also not borne out, at least in birds (Sol 2000). Even if we knew much more, Dayton et al. (1998) and Carlton (1996c) doubt that we can expect any reasonable degree of prediction. It may be impossible to deﬁne meaningful benchmarks as to what the “natural pristine” state of an environment might have been, because there is so much change driven by so many different agents of change altering coastal environments (Dayton et al. 1998). Fishermen removed sea otters, black sea bass, yellowtail, white sea bass, and abalones, and affected many other species found in kelp forests. The kelp themselves are sensitive to large-scale El Niño–Southern Oscillations and longer-term shifts, including global warming trends, and hence change markedly across years and decades. It has therefore been difﬁcult to ﬁnd underlying rules that set invasiveness and invasibility. Invasions have been described as “ecological roulette” (Carlton & Geller 1993), and

INTRODUCTION OF EXOTIC SPECIES

Williamson (1999) suggests that invasions may be as unpredictable as earthquakes (Geller et al. 1997), for nearly the same reasons: multiplicity of controls and low probabilities of occurrence. There are many species that happen to be transported beyond their geographic range, there are many and diverse agents of transport, and there is a quite low probability that any one species makes it as an established invader in the new site. In addition, any of the vicariant species may arrive at any of a range of sites, each with differing conditions. The near random nature of vicariance, the multiplicity of conditions that may prevent colonization, and the enormous variation in conditions from one site to another conspire to prevent generalities from emerging, at least at the current state of knowledge.

Effects of exotic species: balancing positive and negative We can use the review by Mack et al. (2000) or the case histories discussed earlier to compile a list of the detrimental effects of invasions. To this list we may add the introduction of parasites of native species of commercial or ecological importance (Fernando 1991),11 and transoceanic transport of toxic microorganisms that might be responsible for the increased incidence of brown and red tides in coastal waters (Morton 1997). The case histories reviewed above show evidence that, in selected estuaries, a few species of alien origin create major rearrangements of species’ composition in the invaded communities. To evaluate the priority to give the alien invasion issue, it would be useful to have more information on the extensiveness of invasion; that is, how widespread the occurrence of exotics with major effects might be. We have too little information of the extent to which the many other coastal environments share in the pervas-

11

Carlton (1996b) mentions viruses of Central American shrimp that were transported during 1994–1995 to shrimp culture facilities in Texas, where they infected local shrimp.

239

ive ecological consequences experienced by San Francisco Bay, for example. Much has been made of economic costs attributable to invasions by alien species. Costs associated with fouling damage caused by the Asian clam in San Francisco Bay have been estimated at US$1 billion (OTA 1993). Damage to wood docks by shipworms, another species introduced into San Francisco Bay, have been thought to reach about US$200,000 per year (Cohen & Carlton 1995). The dramatic case histories, and the cost estimates that have been compiled, certainly cast the issue of alien species negatively. There also is an unspoken ecological nativism in the notion that maintenance of the local extant species is somehow intrinsically to be preferred to the introduction of alien species. This seems a value judgement that may be difﬁcult to argue (why are Asian clams somehow less worthy than the soft-shell clams? why is smooth cordgrass less desirable than rough cordgrass?). The detrimental effects of invaders are indeed important alterations, but the negative aspects need to be balanced versus positive aspects, since many introduced species have furnished tangible beneﬁts. The introduction of new species of ﬁsh has signiﬁcantly increased food yields in Sri Lanka, Mexico, Brazil, and elsewhere, in many cases without perceived damage to indigenous species, except where piscivore species were used (Fernando 1991). In San Francisco Bay, Chesapeake Bay, and elsewhere, exotic species (e.g. shad, striped bass, Japanese oyster) have improved harvests and have been an economic boon, as well as enhancing water quality, as already mentioned. We need to note also that whatever detrimental effects aliens might prompt, these negatives pale in comparison to the beneﬁts as well as the alterations brought about by massive anthropogenic introductions of exotic species carried out by agriculture, aquaculture, and forestry. To obtain some perspective on the enormous magnitude of human interests in exotic species’ introductions, consider that introduced species yield more than 98% of the harvest of food in the USA (US Bureau of the Census 1998). More than 70%

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of the world’s food comes from only nine crop species: wheat, maize, rice, potato, barley, cassava, soybean, sugar cane, and oats (PrescottAllen & Prescott-Allen 1990). Just three genera, Eucalyptus, Pinus, and Tectona, make up 85% of industrial forestry plantations (Evans 1992). These agricultural and forestry crops are all grown far beyond their native range; they are true humantransported exotics, and have clearly replaced native plants in areas devoted to harvests. Such crops are fundamentally important economically; for example, more than 95% of New Zealand’s export earnings derive from alien species (New Zealand Department of Statistics 1996). The spread of food organisms by people is truly a biogeographic transformation of global scale. In fact, it is evident that human history is a story of widespread species transport. The ecological consequences of this transport have been transcendental,12 not only to ensure human food supplies, but also to sustain human expansion to far away places and, as a result, in the alteration of the world’s landscapes. Coastal species’ invasions are a small mirror of the larger pattern, and need to be thought of in the larger context. In a very real way, the negative effects of exotics may be the cost that we, and more so-called natural environments, pay as a result of human expansion, commerce, transportation, and the exploitation and management of natural resources.

Mitigation or prevention of invasions If we are unwilling to allow continued invasions and spreading of alien taxa, what management options might be available? Is it likely that the frequency of new invasions may be lowered, and 12

Consider what food production of the world might be without American-derived corn; what nutrition might have been in Ireland and other European countries without the potato, a native Andean. Perhaps in a less weighty vein, we could imagine what the cuisine of the Italian Peninsula might be without the tomato and beans brought from the Americas, or the Indian and Southeast Asian cuisine might have been without the Capsicum peppers brought from Mexico. Not to mention other items such as Mesoamerican cacao and vanilla, Arabian coffee, Southeast Asian tea, and so on.

that the expansion of established aliens might be curtailed? Prevention is difﬁcult, not the least because the mechanisms of alien introduction are so varied. A number of proposals to lower introduction via ballast water have been made (these are reviewed by Cohen 1998). The alternatives include: 1

2

3

The exchange of ballast water while the ship is in the open sea, with the expectation that organisms from the open water would be less likely to survive coastal conditions. Flowthrough replacement of ballast water, rather than emptying and reﬁlling, would be the safer, but costlier, alternative for this option. Onboard treatment to rid ballast water of its burden of live organisms by ultraviolet radiation, heat, ultrasound, microwaves, and so on, plus various biocides. These are costly, efﬁcacy is uneven, and treatments may result in noxious products being released into coastal waters. Ballast water may be emptied on shore and treated to kill organisms. This is costly and leaves residues to deal with.

At sea, ﬂow-through exchange seems at present to be the most compelling alternative. The treatment of hulls has been used to remove settled hitchhikers for as long as there have been ships. Most treatments are imperfect, and have to be done repeatedly. The use of tributyl tin on hulls has helped reduce hitchhikers (Minchin & Sheehan 1998), but these compounds have other consequences (as reviewed in Chapter 9). Extirpation of established coastal invaders has proven quite difﬁcult. Perhaps when that ﬁrst 1 m2 of Caulerpa taxifolia was found near the Monaco Museum, or when that ﬁrst bit of Spartina alterniﬂora entered western US estuaries, there was some hope of extirpation. After these species spread, the chances of eradication diminish. The intensive efforts at removal of smooth cordgrass from western estuaries have not been successful. In a few cases extirpation of aliens from terrestrial environments has succeeded (Mack et al. 2000), but these are fewer than the failures, and required long-term and costly campaigns.

INTRODUCTION OF EXOTIC SPECIES

Eradication in aquatic systems may more difﬁcult than in terrestrial environments, because of problems in access and active water-borne transport.

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Dayton, P. K., M. J. Tegner, P. B. Edwards, and K. L. Riser. 1998. Sliding baselines, ghosts, and reduced expectations in kelp forest communities. Ecol. Appl. 8:309–322. Dini, F. 1999. Insights into the ecotoxicological risk related to the rapid spreading of the pantropical green seaweed, Caulerpa taxifolia, along the Mediterranean coasts. Pp. 101– 114 in Proceedings of a Workshop on Invasive Caulerpa Species in the Mediterranean. Mediterranean Action Plan Reports Series No. 125. United Nations Environment Programme, Athens, 317 pp. Early, T. A., and T. Glonek. 1999. Zebra mussel destruction by a Lake Michigan sponge: Populations, in vivo 31P nuclear magnetic resonance, and phospholipid proﬁling. Environ. Sci. Technol. 33:1957–1962. Eldredge, L. G., and S. E. Miller. 1998. Numbers of Hawaiian species: Supplement 3, with notes on fossil species. Bishop Museum Occas. Pap. 55:3–15. Elvira, B. 1995. Native and exotic freshwater dishes in Spanish river basins. Freshw. Biol. 33:103–108. Eno, N. C., R. A. Clark, and W. G. Sanderson. 1997. Nonnative Species in British Waters: A Review and Directory. Joint Nature Conservation Committee, Petersborough, UK, 152 pp. Evans, J. 1992. Plantation Forestry in the Tropics, 2nd edn. Clarendon Press. Oxford. Ewel, J. J., and 20 others. 1999. Deliberate introductions of species: Research needs. BioScience 49:619–630. Fernando, C. H. 1991. Impacts of ﬁsh introductions in tropical Asia and America. Can. J. Fish. Aquat. Sci. 48 (Suppl. 1):24–32. Fogli, S., R. Marchesini, and R. Gerdol. 2002. Reed (Phragmites australis) decline in a brackish wetland in Italy. Mar. Environ. Res. 53:465–479. Frank, J. H., and E. D. McCoy. 1995a. Invasive adventive insects and other organisms in Florida. Florida Entom. 78:1–15. Frank, J. H., and E. D. McCoy. 1995b. Precinctive insect species in Florida. Florida Entom. 78:21–35. Geller, R. J., D. D. Jackson, Y. Y. Kagan, and F. Mulargia. 1997. Earthquakes cannot be predicted. Science 275:1616– 1617. Georgia Coastal Research Council. 2004. Proceedings of a Marsh Dieback Workshop. Savannah, GA, Feb. 3–4, 2004. University of Georgia, Athens, GA, 56 pp. Goodman, P. J., and W. T. Williams. 1961. Investigations into “die-back” in Spartina townsendii agg. III. Physiological correlates of die-back. J. Ecol. 49:391–398. Goodman, P. J., E. M. Braybrooks, and J. M. Lambert. 1959. Investigations into “die-back” in Spartina townsendii agg. I. The present status of Spartina townsendii in Britain. J. Ecol. 47:651–677. Goodwin, B. J., A. J. McAllister, and L. Fahrig. 1999. Predicting invasiveness of plant species based on biological information. Conserv. Biol. 13:422–426. Goss-Custard, J. D., and M. E. Moser. 1988. Rates of change in the numbers of dunlin, Calidris alpina, wintering in

British estuaries in relation to the spread of Spartina anglica. J. Appl. Ecol. 25:95–109. Gottlieb, S. J., and M. E. Schweighofer. 1996. Oysters and the Chesapeake Bay ecosystem: A case for exotic species introduction to improve environmental quality? Estuaries 19:639–650. Gray, A. J., P. E. M. Benham, and A. R. Raybould. 1990. Spartina anglica—the evolutionary and ecological background. Pp. 5–10 in A. J. Gray, and P. E. M. Benham (eds). Spartina anglica—A Research Review. HMSO, London. Greenberg, N., R. L. Garthwaite, and D. C. Potts. 1996. Allozyme and morphological evidence for a newly introduced species of Aurelia in San Francisco Bay, California. Mar. Biol. 125:401–410. Grosholz, E. D., and G. M. Ruiz. 1995. Spread and potential impact of the recently introduced European green crab, Carcinus maenas, in Central California. Mar. Biol. 122:239– 247. Grosholz, E. D., and G. M. Ruiz. 1999. Ecological and evolutionary consequences of invasions: The impacts of the European green crab on multiple trophic levels in central California. Pp. 408–409 in Marine Bioinvasions. Proceedings of the First National Conference. Hallegraeff, G. M. 1998. Transport of toxic dinoﬂagellates via ship’s ballast water: Bio-economic risk assessment and efﬁcacy of possible ballast water management strategies. Mar. Ecol. Prog. Ser. 168:297–309. Harbison, G. R., and S. P. Volovik. 1994. The ctenophore, Mnemiopsis leidyi, in the Black Sea: A holoplanktonic organism transported in the ballast water of ships. Pp. 25–36 in Cottingham, D. (ed.). Nonindigenous Estuarine and Marine Species (NEMO). NOAA, US Department of Commerce, US Government Printing Ofﬁce, Washington, DC, 125 pp. Herbold, B., and P. B. Moyle. 1989. The Ecology of the Sacramento–San Joaquin Delta: A Community Proﬁle. Biological Report No. 85(7.22). US Fish and Wildlife Service, US Department of the Interior. Herbold, B., A. D. Jassby, and P. B. Boyle. 1992. Status and Trends Report on Aquatic Resources in the San Francisco Estuary. San Francisco Estuary Project, Public Report, 257 pp. Heywood, V. H. 1989. Patterns, extents and modes of invasions by terrestrial plants. Pp. 31–60 in J. A. Drake, H. A. Mooney, F. DiCastri, R. H. Groves, F. J. Kruger, M. Rejmánek, and M. Williamson (eds). Biological Invasions: A Global Perspective. Wiley & Sons, Chichester, UK. Leach, J. H. 1993. Impacts of the zebra mussel (Dreissena polymorpha) on water quality and ﬁsh spawning reefs in western Lake Erie. Pp. 381–397 in T. F. Nalepa, and D. W. Schloesser (eds). Zebra Mussels: Biology, Impacts, and Control. CRC Press, Boca Raton, FL. Macdonald, K. B., and M. G. Barbour. 1974. Beach and salt marsh vegetation of the North American Paciﬁc coast. Pp. 175–233 in R. J. Reimold, and W. H. Queen (eds). Ecology of Halophytes. Academic Press, New York.

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Mack, R. C., D. Simberloff, W. M. Lonsdale, H. Evans, M. Clout, and F. Bazzaz. 2000. Biotic invasions: Causes, epidemiology, global consequences, and control. Ecol. Appl. 10:689–710. Marchant, C. J. 1968. Evolution of Spartina (Gramineae), II. Chromosomes, basic relationships and the problems of S. towsendii agg. Bot. J. Linn. Soc. 63:321–326. Meinesz, A., and C. F. Boudouresque. 1996. Sur l’origine de Caulerpa taxifolia en Méditerranée. Comp. Rend. Acad. Sci. Life Sci. Fr. 319:603–613. Merrill, E. D. 1902. The North American species of Spartina. Bull. US Bur. Plant Ind. 9:5–15. Meyerson, L. A., K. Saltonstall, L. Windham, E. Kiviat, and S. Findlay. 2000. A comparison of Phragmites australis in freshwater and brackish marsh environments in North America. Wetlands Ecol. Manage. 8:89–103. Mills, E. L., J. H. Leach, J. T. Carlton, and C. L. Secor. 1993. Exotic species in the Great Lakes: A history of biotic crises and anthropogenic introductions. J. Great Lakes Res. 19:1–54. Minchin, D., and J. Sheehan. 1998. The signiﬁcance of ballast water in the introduction of exotic marine organisms to Cork Harbour, Ireland. ICES Coop. Res. Rep. 224:12– 23. Modlin, R. F., and J. J. Orsi. 1997. Acanthomysis bowmani, a new species, and A. aspera Ii, mysidae newly reported from the Sacramento–San Joaquin estuary. Proc. Biol. Soc. Washington 110:439–446. Mooney, H. A., S. P. Hamberg, and J. A. Drake. 1986. The invasions of plants and animals into California. Pp. 250–272 in H. A. Mooney and J. A. Drake (eds). Ecology of Biological Invasions of North America and Hawaii. Springer-Verlag, New York. Morton, B. 1997. The aquatic nuisance problem: A global perspective and review. Pp. 1–54 in F. M. D’Itri (ed.). Zebra Mussels and Aquatic Nuisance Species. Ann Arbor Press, Chelsea, MI. New Zealand Department of Statistics. 1996. New Zealand Ofﬁcial Yearbook, 99th edn. Department of Statistics, Wellington. Newell, R. I. E. 1988. Ecological changes in Chesapeake Bay: Are they the result of over-harvesting the American oyster, Crassotrea virginica? Pp. 536–546 in M. P. Lynch, and E. C. Krome (eds). Understanding the Estuary: Advances in Chesapeake Bay Research. CRC Publication No. 129. Chesapeake Bay Consortium, Solomons, MD. Nichols, F. H., and M. M. Pamatmat. 1988. The Ecology of the Soft-bottom Benthos of San Francisco Bay: A Community Proﬁle. Biological Report No. 85(7.19). US Fish and Wildlife Service, US Department of the Interior. Nichols, F. H., J. E. Cloern, S. H. Luoma, and D. H. Peterson. 1986. The modiﬁcation of an estuary. Science 231:567– 573. Nichols, F. H., J. K. Thompson, and L. E. Schemel. 1990. Remarkable invasion of San Francisco Bay (California, U. S.) by the Asian clam, Potamocorbula amurensis. II. Displace-

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ment of a former community. Mar. Ecol. Prog. Ser. 66:95–101. OTA (Ofﬁce of Technology Assessment). 1993. Harmful Non-indigenous Species in the United States. US Congress, Washington, DC. Phelps, H. L. 1994. The Asiatic clam (Corbicula ﬂuminea) invasion and system-level ecological change in the Potomac River estuary near Washington, D.C. Estuaries 17:614– 621. Pimentel, D., L. Lach, R. Zuñiga, and D. Morrison. 2000. Environmental and economic costs of nonindigenous species in the United States. BioScience 50:53–65. Por, F. D. 1990. Lessepsian migrations. An appraisal and new data. Bull. Inst. Océanogr. Monaco 7 (spec. vol.):1–10. Prescott-Allen, R., and E. Prescott-Allen. 1990. How many plants feed the world? Conserv. Biol.4:365–374. Race, M. S. 1982. Competitive displacement and predation between introduced and native mud snails. Oecologia 54:337–347. Ranwell, D. S. 1967. World resources of Spartina townsendii (senso lato) and economic use of Spartina marshland. J. Appl. Ecol. 4:239–256. Raybould, A. F., A. J. Gray, M. J. Lawrence, and D. F. Marshall. 1991. The evolution of Spartina anglica C. E. Hubbard (Gramineae): Origin and genetic variability. Biol. J. Linn. Soc. 43:111–126. Riemann, B., T. K. Nielsen, S. J. Horsted, P. K. Bjornsen, and J. Pock-Steen. 1988. Regulation of phytoplankton biomass in estuarine enclosures. Mar. Ecol. Prog. Ser. 48:205– 215. Rodriguez-Prieto, C. 1999. Ecological and possible economical consequences of the spread of Caulerpa taxifolia in the Mediterranean. Pp. 85–99 in Proceedings of a Workshop on Invasive Caulerpa Species in the Mediterranean. Mediterranean Action Plan Technical Reports Series No. 125. United Nations Environment Programme, Athens, 317 pp Ruiz, G. M., P. W. Fofonoff, J. T. Carlton, M. J. Wonham, and A. H. Hines. 2000. Invasion of coastal marine communities in North America: Apparent patterns, processes, biases. Annu. Rev. Ecol. Syst. 31:481–531. Ruiz, G. M., and 6 others. 2000a. Global spread of microrganisms by ships. Nature 408:49–50. Ruiz, G. M., L. D., Smith, A. H. Hines, J. T. Carlton, and D. W. Coats. 1998. Ballast water and non-indigenous species in U.S. coastal waters. ICES Coop. Res. Rep. 224:5–11. Sheppard, A. W. 1992. Predicting biological weed control. Trends Ecol. Evol. 7:290–291. Shigesada, N., and K. Kawasaki. 1997. Biological Invasions: Theory and Practice. Oxford University Press, Oxford, 205 pp. Simberloff, D. 2000. Global climate change and introduced species in the United States forests. Sci. Tot. Environ. 262:253–261. Sol, D. 2000. Are islands more susceptible to be invaded than continents? Birds say no. Ecography 23:687–692.

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Sprugel, D. G. 1991. Disturbance, equilibrium and environmental variability: What is natural vegetation in a changing environment? Biol. Conserv. 58:1–18. Thompson, J. D. 1991. The biology of an invasive plant. BioScience 41:393–401. Thompson, J. K., and L. E. Schemel. 1991. An Asian bivalve, Potamocorbula amurensis, invades San Francisco Bay with remarkable speed and success. J. Shellf. Res. 10:259. US Bureau of the Census. 1998. Statistical Abstract of the United States 1996, 200th edn. US Government Printing Ofﬁc, Washington, DC. Verlaque, M., C. F. Boudouresque, A. Meinesz, and V. Gravez. 2000. The Caulerpa racemosa complex (Caulerpales, Ulvophyceae) in the Mediterranean Sea. Bot. Mar. 43:49–68.

Vermeij, G. J. 2001. Community assembly in the sea: Geologic history of the living shore biota. Pp. 39–60 in Bertness, M. D., S. D. Gaines, and M. E. Hay (eds). Marine Community Ecology. Sinauer Associates, Sunderland, MA, 550 pp. Vitousek, P. M., C. M. D’Antonio, L. L. Loope, and R. Westbrooks. 1996. Biological invasions as global environmental change. Am. Sci. 84:468–478. Weis, J. S., and P. Weis. 2003. Is the invasion of the common reed, Phragmites australis, into tidal marshes of the eastern U.S. an ecological disaster? Mar. Pollut. Bull. 46:816–820. Wellcome, R. L. 1988. International introductions of inland aquatic species. FAO Fish. Tech. Pap. 294:1–318. Williamson, M. 1999. Invasions. Ecography 22:5–12.

Chapter 11 Harvest of finfish and shellfish

Cod. Photo from http://www.dfo-mpo.gc.ca/media/infocus/2003/20030424_e.htm.

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A case study of fishermen and fish: the shallow banks off North America1 Towards the end of the 9th century, coastal ﬁshermen in Western Europe found fewer and fewer places for proﬁtable harvest, and sought better ﬁshing grounds farther aﬁeld than the coastal shallows they had been exploiting for some time. The Bretons, Germans, Flemish, and English went northward, towards Iceland. Perhaps the most venturesome were the Basque ﬁshermen who exploited waters off northwest Iberia, the Irish Sea, the Canary Islands, and, eventually, the richest ﬁshery grounds off North America. The Basque are a remarkable and distinctive people (Casado Soto et al. 1995). Their language— Euskera—antedates and seems unrelated to most Indo-European languages, except perhaps Finnish and Hungarian. Basque anatomical details and blood types suggest close linkages to prehistoric Cro-Magnon humans. Throughout their history, the tough, self-contained Basques managed to resist multiple invasions, retained their distinctiveness, and made do with a minimum of resources; all in all, a people well preadapted for the saga they came to play in the demanding and dangerous ﬁshery off the North American coast. The basque homeland, Euskal Herria, is a tiny mountainous area, draped on either side of the Pyrenees, and stretching along the coast of the Bay of Biscay. The poverty of the soil forced the Basque to use marine resources from the beginning. Images that seem to depict sea bream were discovered in a Paleolithic cave on the Spanish side of the Pyrenees. Middens with ﬁsh bones and shellﬁsh remains have also been found. By the 600s, Basques were the major providers of whale oil, whalebone, whale teeth, and whale meat to Europe. After the 9th century the Basque developed ways to use salt2 to preserve the meat, and 1

Much of the material in this section is adapted from Collins (1986), Mowatt (1986, chapter 11), Kurlansky (1997, 1999), Lear (1998), and from a remarkably comprehensive compendium of articles on all aspects of Georges Bank (Backus & Bourne 1991). 2 The practice of salting meat was of course much older, but the Basque made it their own in a large way. The Basque had the advantage of living in Iberia, near areas that were sufﬁciently hot

they learned to build more seaworthy ships by adopting the Viking technique of overlapping planks. With these ships they ranged farther and farther aﬁeld in pursuit of whales. The whaling became so established that in a 1237 charter, Fernando III of Castile decreed a royal right to the tongue (then considered the choicest) of any whale caught, as did other kings (Collins 1986). The Basque proclivity for maritime endeavors was tailormade to take advantage of economic opportunity provided by religious practices. The medieval Church prohibited the consumption of red-blooded meat on Fridays and the many holy days of the year. Such meat was considered “hot”, and too closely associated with sex, which was also banned on such days. Meat from animals or parts of animals that lived or were submerged in water, such as ﬁsh, whales, and the tail of beavers, were thought to be “cold”, and hence acceptable for consumption on the plethora of holy days. The Basque (and, to some extent, ﬁshermen from other regions) took advantage of the enormous commercial potential provided by the demand by an entire continent of observant folk for “cold” protein. During the 15th century, as they ventured north to the Faroes, Iceland, and beyond, Basque whalers encountered cod, which they also brought back to Europe, preserved in salt, and dried, much as they had processed whale meat (Fig. 11.1). Salt cod did not rot, was inexpensive, and tasted better than the dried cod processed by older Viking methods. Basques were so successful in selling their salt cod all over Europe that even today dried salt cod is a staple in many countries far away from the geographic range of cod. The Basque therefore had the ships, technology, economic incentive, character, and steadfastness required for long trips at sea. There is clear evidence that the Norse reached Newfoundland,3 and dry so that salt could be evaporated from sea water. The more northern nations, with wetter climates, could not avail themselves of the salt supply from the south as readily, which furnished a ready-made advantage to the Basque ﬁshermen, an advantage that lasted for centuries until underground salt mining was developed. 3 Jones (1986) and Fitzhugh and Ward (2000) are excellent sources on the subject of Viking settlements in the American continent.

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Figure 11.1 A seasonal factory set up on shore to process and preserve ﬁsh caught in the shallow banks of eastern North America before transport to Europe. The steps involved the unloading of ﬁsh from the ship, cleaning, salting, and setting out the split carcasses in drying racks. Detail from a map by Herman Moll in his Atlas Royal, published in 1712.

but little about Basque travels. Fishermen seldom are interested in revealing favorite ﬁshing places. The Basque have demonstrated an ability to keep their own counsel through their history, and managed to keep the location of their ﬁshing grounds secret for centuries. Several lines of evidence, some quite hypothetical, suggest that the Basque exploited the enormously productive ﬁsh and whale stocks of the shallows off North America, and did so quite early, perhaps in pre-Columbian times. There are reports of historical rumors about the Basque ﬁnding land across the sea, and there are records—apparently dated in pre-1492 times— of taxes paid on ﬁsh caught there. There are many early place names of Basque origin in the region (Ille-aux-Basques, Placencia, Port-aux-Basques, and so on), as well as some words used by the

native Beothuk inhabitants of Newfoundland that could be of pre-Columbian Basque origin (Montero 1998). When Giovanni Caboto (John Cabot) came on his 1497 voyage, searching for a northern route to the Spice Islands, he found hundreds of Basque ﬁshing vessels already in the area, so that a Basque presence appeared well established. There are also documented archaelogical remains of substantial Basque whaling stations in Placentia Bay, Newfoundland, and Red Bay, Labrador, dating from at least the early 1500s. The most compelling, albeit deductive, argument is the sheer harvest of cod that the Basque imported to European markets. It seems quite implausible to think that ﬁshing around the Faroes or Iceland, where the Basque were seldom reported to be ﬁshing, could have yielded the harvest taken. The huge cod stocks that could be

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• From a letter written by Raimondo di Soncini, to the Duke of Milan, December 18, 1497, reporting on a conversation with Giovanni Caboto, who had just returned to England from Newfoundland: “. . . and they afﬁrm that the sea is covered with ﬁsh which are caught not merely with nets but with baskets, a stone being attached to make the baskets sink with the water . . .” (from Judah 1933). • Crewmen of the Grace of Bristol, 1594: “. . . in two hours space we tooke with our hooks 3 or 4 hundred great Cod . . .” • Charles Leigh, reconnoitering the waters around the Magdalen Islands, “. . . there is as great an abundance of cods as in any place to be found. In a little more than an houre we caught with hookes 250 of them.” • From the journal of Gabriel Archer, on the ship Concord, having left Falmouth, England on March 26, 1604, captained by Batholomew Gosnold, writing May 14, 1604, on reaching “. . . a mighty headland . . . near this cape we came to anchor at 15 fathoms, where we took great store of codﬁsh, for which we . . . called it Cape Cod” (from Jensen 1972). • Baron Lahontan, a traveler: “You can scarce imagine what quantities of codﬁsh were catched by our Seamen in the space of a quarter of an hour . . . the Hook was no sooner at the bottom than a Fish was catched . . .”

found off the coasts of northern North America became widely known and soon exploited by Breton, Norman, and Portuguese ﬁshermen following Cabot’s voyage (Pope 1997; Lear 1998). The evident, plentiful cod supply may lead one to suspect that these waters might have been the unpublicized Basque ﬁshing grounds. Samuel Eliot Morison, the dean of maritime historians, barely mentioned Basque ﬁshermen in his review of the voyages of Europeans to the

western reaches of the Atlantic (Morison 1971). Morison disbelieved the notion that natives adopted Basque words, for example the widely quoted idea that the Beothuks of Newfoundland adopted the Basque name for cod (bacailhaba), forsaking their native bobboosoret. Morison did allow, however, that the Basque and whalers had extended their activities to Newfoundland for centuries, and their presence was routine by the mid-1500s (Morison 1971). Morison’s skepticism may be merited in view of the sparse historical record. The evidence available does suggest that Basque and Breton whalers and ﬁshermen were well established in the area for a very long time, and routinely exploited the rich shallows of the Western Atlantic by the mid-1500s, and most probably considerably earlier. Beyond what might be speculated, the earliest documented sale of salt cod in Europe dates from 1517 in Bourdeaux (Casado Soto et al. 1995), only 25 short years after 1492. By the turn of the 16th century, many others joined the ﬁshery in the shallows of the Northwest Atlantic, and as many as 650 vessels were catching cod, using hooks and handlines. The ﬁsh caught then were up to 2 m in length, and weighed as much as 90 kg. By 1620, the ﬂeet catching cod in these waters counted more than 1,000 vessels, most making two trips per year. By the end of the 16th century, the average ship loaded about 125,000 cod, and the ﬁsh were 7– 9 kg in weight (by contrast, cod caught at the end of the 20th century weighed on average about 2–3 kg). “Scarcely a harbor where there are not several ﬁshing vessels . . . taking every day 15,000 [to] 30,000 ﬁsh . . .” (Nicolas Denys, English ﬁsherman).

In the course of three centuries, the Basque developed an industrial-scale procedure to exploit the rich cod supply (Pope 1997). They brought salt to the banks, caught huge numbers of ﬁsh using small three-man row boats, and dried cod on the rocky headlands available on the continent and islands (see Fig. 11.1). Their ﬁsh salting and drying methods became the rule through to the 19th century (Fig. 11.2), and are used even today in the Canadian Maritime Provinces.

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Figure 11.2 Codﬁsh being gutted and split (bottom of picture), and set out on drying racks on the docks. The ﬁsh were brought into the harbor at Gloucester, Massachusetts, by the schooners seen in the top right. Photograph in the Gordon W. Thomas Collection of the Cape Ann Historical Association. Photo taken April 19, 1912 by a photographer named Parsons. Reproduced from Garland (1995).

The chain of shallows off the North American coast—from Flemish Cap in the northeast extreme, to the Grand Banks of Newfoundland, the Gulf of St Lawrence, the Gulf of Maine, and to Georges Bank in the southwest— offered an enormous ﬁshing potential. The potential harvests were so large that they justiﬁed hazardous, long, open-sea trips in tiny, yet seaworthy vessels, to catch ﬁsh to sell at substantial proﬁt in Europe. The ﬁsh of these shallows have been a natural resource that

marked the social, economic, and political development of the surrounding areas through the centuries. The Maritime Provinces of Canada and the New England states of the USA all depended in powerful ways on ﬁsh catches from the shallows (Fig. 11.3). The markets stretched out to many other parts of the world: during the 1700s lower-grade salt cod was sent to the West Indies as cheap protein to feed slaves needed to grow sugar. The Canadian and US ﬁshing industry

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“There are three kinds of human beings: those that are alive, those that are dead, and those that go to sea” (Anacarsis, 600 AD).

Figure 11.3 The fundamental importance of ﬁshing, particularly of cod, to the economy of the coastal areas of northeast North America is vouchsafed, for example, by this postage stamp from Newfoundland. Note the mass of harvested cod, and the telling notation “Newfoundland currency”. From http:// ﬁlaman.unikiel.de/manual/ﬁshbaseﬁsh_sta-mps.htm.

persisted—with uncertain ups and downs—for centuries, in spite of the daunting hazard of the endeavor. Many New England and Maritime Province ports (Fig. 11.4) developed long traditions of ﬁshing the banks, despite the hazards (Fig. 11.5). Fishermen from Gloucester, a major Massachusetts ﬁshing center, continued to ﬁsh, taking as a given the frightening losses of vessels and human lives as an inevitable part of ﬁshing (Fig. 11.6). Whether Basque, Breton, Canadian, or American, ﬁshermen had to be impressively willing to take risks. Whether engaged in long trips in factory ships, or in near-shore ﬁshing in one-person dorys (Fig. 11.7), ﬁshermen are a different sort of human, persisting in the face of daunting conditions.

To get closer to the subject of this chapter, as early as 1683 there was evidence that it might be possible to overﬁsh stocks in English waters. In the frontierless New World, natural resources seemed nearly inﬁnite, and overexploitation seemed inconceivable. In 1653 in Massachusetts, for example, encouragement was given to English businesses to invest in the ﬁshing industry, and the General Court of Massachusetts (the equivalent of a local parliament) exempted vessels and ﬁshing equipment from taxation (Lear 1998). By 1740, there were 5,000–6,000 ﬁshermen in Massachusetts alone, plus about 400 fully decked ﬁshing vessels and about the same number of undecked smaller boats. Even more people were employed ashore for processing and marketing ﬁsh caught by Massachusetts vessels. There was little doubt that the region’s ﬁshing industry was going to make great progress, with growing demand and sales to Europe and the West Indies. Sailing ﬁshing vessels were numerous in the ports (Fig. 11.4). A Canadian ofﬁcial averred: “. . . as to . . . cod, mackerel, herring . . . protection would be useless to them—I say it is impossible, not merely to exhaust them, but even to lessen their numbers by the means now used for their capture . . . although enormous quantities of ﬁsh have been caught, there are no indications of exhaustion” (L. Z. Joncas, in a report to the Canadian Ministry of Agriculture, 1885, cited in Kurlansky 1997). But even in the midst of plenty, there were a few warnings. In the 1720s, a certain M. Charlevoix, after duly remarking that “the number of cod seems to equal that of the grains of sand”, went on to say “It might not, however, be amiss to discontinue this ﬁshery from time to time.” The operative element in the quote from the Canadian ofﬁcial above is the “means now used”. By the 1880s steam-powered trawlers were working British waters, and in 1892, otter trawls (Fig. 11.8) were developed. These advances made it possible to ﬁsh far more effectively, and increase catches substantially. In fact, the catches were

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Figure 11.4 Fishing vessels in Gloucester Harbor, Massachusetts, 1919. Photograph in the Gordon W. Thomas Collection of the Cape Ann Historical Association. Reproduced from Garland (1995).

Figure 11.5 Weather in the Northwest Atlantic was and is a challenge for vessels. Reproduced from Garland (1995).

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1,200

120

80

40

Number of people lost

Number of vessels lost

160

1,000 800 600 400 200 0 1820 1830 1840 1850 1860 1870 1880 1890 Year

0 1820 1830 1840 1850 1860 1870 1880 1890 Year

Figure 11.6 Numbers of ﬁshing vessels (left) and crew from the port of Gloucester, Massachusetts, that were lost at sea (right) during the 1800s. Information from records available in http://www.downto sea.com/list.htm.

large enough to occasionally drive ﬁsh prices to ruinously low levels, and, more worrisomely, to deplete local ﬁsh stocks in the North Sea by the turn of the century. Fishermen had to go farther away to catch ﬁsh, and buy larger ships. Although the ﬁrst otter trawl was offered to Cape Cod ﬁshermen in 1893, the Georges Bank ﬁshery remained stubbornly based on sailing schooners for 30 more years. The pressure to catch more cod and similar ﬁsh increased in the early decades of the 20th century by the development of ﬁsh-freezing facilities, which opened large inland markets, and the design of dieselfueled ships, which made ﬁshing more effective. Eventually, these technological advances produced what became known as the factory ship. These vessels operated at sea, and processed, froze, and stored ﬁsh caught by several to many trawlers during long periods of time.4 The large size of these ships also allowed ﬁshing in weather that hampered the older, smaller ﬁshing vessels. 4

Figure 11.7 Dory ﬁsherman coming ashore with cleaned catch, Gloucester, Massachusetts. Reproduced from Garland (1995).

It is difﬁcult to convey the scale of these industrial-scale ﬁshing efforts. Factory ships as large as 4,000 ton capacity were used, with, as Kurlansky (1997) points out, the ability to pull trawls large enough to “swallow jumbo jets”, hauling its trawl every 4 hours, 24 hours a day. A pair of such ships could ﬁsh by pulling a huge trawl held between the two ships, and taking turns processing and operating the trawl, so that ﬁshing never stopped. Spanish companies, mostly based on Euskal Herria, were on the forefront of these developments.

253

HARVEST OF FINFISH AND SHELLFISH

In the latter part of the 20th century, high technology means of ﬁnding ﬁsh, including sonar or spotter aircraft, replaced centuries-old methods such as following feeding ﬂocks of sea birds, or lowering clumps of sticky grease to the bottom to see if the grease picked up red sediment favored by cod. These advances made ﬁshermen even more effective predators. With factory ships operating at high efﬁciency across the world, most of the world’s ﬁsh harvests increased; in the Northwest Atlantic, our area of immediate concern, the cod harvest more than tripled by the mid-1960s (Fig. 11.9). Factory ships from many nations—Spain, Japan, Poland, the then Soviet Union, East and West Germany, and more—plied the rich shallow banks off the North American coast, and took away enormous quantities of ﬁsh. During the 1970s, ﬁsh catches began to evidently diminish (Fig. 11.9), and much effort went into international meetings, regulations of net mesh to conserve juvenile ﬁsh, and even quotas. The causes of the precipitous collapse of such ﬁsheries have been debated,5 but it is inescapable 5 Increases in natural mortality, predation by seals, and failure of recruitment have been invoked as possible causes of the decline in ﬁsh stocks in Georges Bank and similar environments. Critical studies of these possible alternatives conclude that ﬁshing-related mortality appears to be the major cause (Myers et al. 1996a, 1996b, 1997).

Warp

Trawl door

Bridle Net

Danleno Ground rope

2,000 Catch (× 103 tons)

Figure 11.8 View of a modern otter trawl as it might look during a trawling run. The two “doors” keep the trawl open laterally. A series of heavy rollers hold the bottom line near the bottom, while a series of ﬂoats attached to the top line hold the upper part of the trawl open vertically. This trawl design was a marked improvement over the “beam” trawl that was held open along the bottom by a wooden beam. From Churchill (1989).

1,600 1,200 800 400 0 1895

1915

1935 1955 Year

1975

1995

Figure 11.9 Catches of cod in the Northwest Atlantic, 1985–1993. From Lear (1998).

that ﬁshing pressure had increased above levels made sustainable by reproductive and recruitment rates of the ﬁsh. The rates of mortality due to ﬁshing increased or have remained at a high level for many ﬁsh stocks in Georges Bank since the 1970s (Fig. 11.10 top). At the same time the population of ﬁsh that could spawn became smaller and smaller (Fig. 11.10 middle). The inexorable result from these circumstances was that the numbers of ﬁsh “recruiting”—that is entering reproductive age—into the Georges Bank populations diminished to rather low levels (Fig. 11.10 bottom). The intense ﬁshing effort from ﬂeets coming from far away to the North American banks

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Fishing mortality rate (F )

2.5 Yellowtail Cod Haddock

2.0 1.5 1.0 0.5 0.0 1976

1978

1980

1982

1984

1986

1988

1990

1992

1994

Spawning biomass (× 106 kg)

100 80 60 40 20 0 1973 1975 1977 1979 1981 1983 1985 1987 1989 1991 1993 1995

Recruitment (× 106 individuals)

100 80 60 40 20 0 1973 1975 1977 1979 1981 1983 1985 1987 1989 1991 1993 1995 Year

prompted the USA and Canada to take action to protect what they saw as their patrimony. The USA passed the Magnuson Fishery Conservation and Management Act, which took

Figure 11.10 Mortality due to ﬁshing (top), biomass of spawning ﬁsh (middle), and recruitment, expressed as numbers of ﬁsh entering their third year of life (bottom), for yellowtail ﬂounder, haddock, and cod, three major species of ﬁsh of commercial interest in Georges Bank. From Fogarty and Murawsky (1998).

effect in 1977. This law established a 200-mile protected band around the shores of the USA where ﬁshing was to be limited to US ﬁshermen. Canada passed a similar law, creating a compet-

HARVEST OF FINFISH AND SHELLFISH

160 140

Biomass index

120

Figure 11.11 Relative abundance (“biomass index”, a measure of weighted mean abundance, in kg tow−1) of gadids, ﬂatﬁsh, other groundﬁsh, skates, and dogﬁsh sharks, in Georges Bank 1963–1996. From Fogarty and Murawsky (1998).

255

Dogfish Skates Other Flatfish Gadids

100 80 60 40 20 0 1963 1966 1969 1972 1975 1978 1981 1984 1987 1990 1993 1996 Year

ing claim on Georges Bank. The dispute was partially resolved in the International Court in The Hague, where the northeast section of Georges Bank was granted to Canada, and the rest came under US control. As a result of the protection provided by the Magnuson Act, US ﬁshermen began a major expansion. The numbers of New England trawlers increased from 570 in 1976 to 921 in 1981. Technological ability to ﬁnd ﬁsh also increased, and resulted in an increase in catch for a few years (note the small peak on the right-hand side of Fig. 11.9). In spite of warnings from scientists and resource managers, the industry (through a Fishery Council also established by the Act) actively fought restrictions in ﬁshing. As a result of the increased ﬁshing pressure from the new American and Canadian vessels, most populations of commercially important bottom ﬁsh decreased markedly in a few short years, so that by the mid-1990s few ﬁsh could be caught even within the 200-mile limit. An unforeseen consequence of the depletion of commercially sought groundﬁsh (gadoids and ﬂatﬁsh that forage near the sea ﬂoor) during the period of intense US ﬁshing was that other species gained access to the prey left unconsumed by the

disappearance of cod, haddock, ﬂounder, etc. Sometime in the 1970s, surveys of Georges Bank revealed a surge in the populations of dogﬁsh sharks and skates (Fig. 11.11).6 These species were harvested by the large ﬁshing ﬂeets before the Magnuson Act, but were initially not desired by US trawlers. In the late 1990s some harvest of these sharks and rays took place,7 and the abundance of dogﬁsh and skates diminished too (Fig. 11.11). 6

Similarly, the loss of the intensely harvested mackerel and herring on Georges Bank led to a proliferation of sand lance (Fogarty et al. 1991). These changes in ﬁsh species have led to major readjustments in the species that make up the food webs of Georges Bank and similar ecosystems (Fogarty & Murawsky 1998). 7 In 2000 in Massachusetts, dogﬁsh shark made up 47% of the landings in the port of Chatham, and up to 91% in Plymouth and Scituate; dogﬁsh represented at the time a ﬁshery worth about US$10 million in New England. Dogﬁsh were sold to Asian and European markets as ﬁsh ﬁllets, shark ﬁns, oils, and other products. The dogﬁsh shark catch brought into port by ﬁshermen was limited by regulations, but there were huge losses of dogﬁsh as bycatch during ﬁshing for other species. “All the codﬁshing boats are getting loaded with dogﬁsh . . . [but] you can’t keep more than 7,000 pounds . . . gill nets sometimes catch 15,000–20,000 [pounds], and we’re not even looking at them,” said Chatham ﬁsherman Ken Tolley (Cape Cod Times, Aug. 8, 2000). Markets for dogﬁsh in New England and Great Britain (where it replaced cod in ﬁsh and chips) developed, and great pressure was placed on the stocks. The US government placed strict limits on dogﬁsh catches, and, as stocks dwindled, made plans to close the ﬁshery (Boston Globe, July 12, 2003).

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In the early 1980s, as the crisis in ﬁsh stocks in Georges Bank was becoming ever more apparent, another development added further conﬂict. After the 1973 oil shortage crisis thoroughly affected the US economy, pressure developed to ﬁnd fossil fuel sources within the country, avoiding the morass of dependency on foreign oil. Oil exploration suggested the possible presence of petroleum deposits under Georges Bank, among several other offshore sites. Coming after several widely reported spills (the Ixtoc blow-out, the Exxon Valdez and Florida spills, etc., reviewed in Chapter 7), the possibility of catastrophic oil spills, as well as slow degradation from chronic occurrence of small spills, became a potent environmental and political issue. Permits for exploration were issued, but environmental and ﬁshing groups brought suits to prevent exploitation. After several exploration wells found few proven reserves, and much litigation and political conﬂict, court injunctions and reduced ﬁnancial potential have left this issue aside. Although the oil issue did not become a reality, the continuing decreases in commercially desirable ﬁsh in Georges Bank and other shallows off North America had severe social and economic consequences. New England ﬁshermen spent nearly 65% more time at sea in 1988 than they did in 1978, but caught 25% fewer ﬁsh. Fishermen could not keep up payments on their trawlers, and many ﬁshermen had to leave the profession across the New England coast; 20,000 people lost jobs in Canada in 2001 as the ﬁshery collapsed, in addition to the 40,000 who had lost jobs since 1992 (http://archive.greenpeace.org/∼comms/ cbio/cancod.html). The evident scarcity of groundﬁsh led to the enacting of moratoria in the early 1990s. In 1999, the National Marine Fisheries Service (NMFS) corroborated the continued decline in Georges Bank ﬁsh stocks, particularly cod, and supported strictly regulated ﬁshing in the US sector. In 2002, spawning cod remained in sharp decline, in spite of more than a decade of sharply reduced harvests, and Canada announced that the cod ﬁshery in much of their waters was to be closed to take of groundﬁsh.

“We used to land 80,000 or 100,000 pounds of ﬁsh on each seven- or eight-day trip. Haddock would be knee-deep on the deck. But by 1988, it took 10 to 12 days just to land 30,000 pounds” said Joe Brancaleone, last of a respected family of outstanding Gloucester ﬁshermen, speaking of ﬁshing on Georges Bank. Joe is still involved in ﬁshing as Chair of the Fishery Management Council, an oversight panel that has attempted to limit ﬁsh catch. Brancaleone added, “it’s tough when you see guys that you’ve known all your life yelling and hollering that the Council is trying to put them out of business. But they don’t realize that there are so few ﬁsh that they are going out of business anyway.” Joe now manages a Burger King for a living. From a letter by Deborah Cramer, in http:// www.motherjones.com/mother_ jones/ JA94/cramer.html.

“This is the end for the few of us left. It’s over. This way of life is ﬁnished” (Kelvin Letto, ﬁsherman of L’Anse-au-Clair, Labrador). “The government complains that it wasn’t an easy decision. I’ll tell you what’s not easy—having to board up your house at age 50, having everything you own reduced to nothing, and having to walk away and leave it all behind as your community is reduced to ruins” (Earle McCurdy, head of Newfoundland’s Fish, Food, and Allied Workers Union). Both quoted in Boston Globe, April 25, 2003

In general, the reproductive potential in most ﬁsh is high enough that population recoveries became evident after ﬁshing pressures on Georges Bank were lowered (Murawski et al. 2000). In the

257

Scotian Shelf, recovery has been slower than might have been thought (Fu et al. 2001), even though ﬁshing has been restricted by limits on catch and by the creation of closed areas. Recovery was also slow in the Southern Gulf of St Lawrence (Swain & Sinclair 2000). Slow (decades rather than years) recovery may be related to other competing predators having established populations in the area, predation by seals, to some continuing, albeit reduced, ﬁshing, and because of unpredictable low recruitment, perhaps associated with large-scale global ocean changes. The conﬂict about ﬁshing in the region continues. As of 2003, ﬁshermen claim that there are plentiful groundﬁsh, and that they are readily ﬁnding them using new techniques. Moreover, there are so many ﬁsh to be caught that even with the 60 day per year ﬁshing limit in much of the area because of the regulatory limits, they have to discard a large portion of catches. Fishermen use newer methods, and target ﬁsh accumulated in speciﬁc sites. Fishermen may therefore indeed ﬁnd ﬁsh in large numbers, but probably do so in very localized areas. The NMFS surveys, in contrast, were designed to assess stocks over the entire Georges Bank, purposely using methods identical to those used in the past, so that time trends can be detected. Based on these surveys, federal scientists insist that the ﬁshery is severely depleted, and continue to press for stringent limits.8 Differences in goals and perceptions, as happened through the history of these conﬂicts, thus continue to impair the management and use of these resources.

Effects of increased fishing pressure The history of the ﬁsheries of the Northwest Atlantic may be somewhat more dramatic than elsewhere in the world, but the trends are not unlike what has happened in these other places. 8

The stringent controls on catch being considered in mid-2003 may lead to a loss of up to 3,000 jobs during their ﬁrst year, with a loss of revenues up to US$88 million. The New England ﬁshing industry has 1,000 ﬁshing boats, and grosses US$1.1 billion per year (Cape Cod Times, 11 July 2003).

World fish catch (million tons)

HARVEST OF FINFISH AND SHELLFISH

100 China

80 60 40

World excluding China

20 0 1950

1960

1970 1980 Year

1990

2000

Figure 11.12 Estimated capture by ﬁsheries for the world, 1950–2000. Data for China is set out separately because its reliability has been questioned (FAO 2000, 2002; Pauly et al. 2002). Data include both marine and inland ﬁsh harvests; for comparison, for the year 2000, inland ﬁsheries contributed only 8.3% of the world ﬁsh catch. The bulk of the harvest reported in this ﬁgure is derived from marine, mostly coastal, waters. From FAO (2002).

Much has been written about various direct and indirect consequences stemming from the intensiﬁcation of ﬁshing in most parts of the coastal seas. These consequences include lowered abundance of targeted stocks, smaller mean size of remaining individuals, loss of non-target species as bycatch, shifts in the composition of species in food webs of ﬁshed environments, and habitat disturbance.9 Lowered populations of commercially targeted stocks The world catch of commercially exploited ﬁsh species has increased during the last half of the 20th century (Fig. 11.12). There are certain ups and downs in the record, associated with the collapses of speciﬁc stocks, the development of new methods, and so on, but, in general, world ﬁsh catch increased steadily, and by 1990 was 9

Blaber et al. (2000) review these and other additional effects associated with various types of ﬁshing. The impact of human harvest of other species, such as otters and sea urchins, also have serious consequences for the communities where they were removed (Tegner & Dayton 2000). Review of the numerous studies in many sites throughout the world revealed that such removals restructured assemblages of species, affecting the abundance of many species, often of commercial as well as ecological importance.

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16

Diadromous Flatfishes

14

Cods, etc. Redfish, etc.

Catch (× 105 tons)

12

Jacks, etc. 10

Herrings, etc. Tunas, etc.

8

Mackerels, etc. 6

Sharks, etc. Other fishes

4

Crustaceans Mollusks

2

Other invertebrates 1950

1960

1970 Year

1980

1990

near 100 million tons per year. During the 1990s world catch declined slowly, by about 0.7 million tons per year since the late 1980s (Pauly et al. 2002). To some extent, it is deceptive to merely look at total ﬁshery catch for the world, because total catch subsumes the contributions by many different species, not all of which are in decline, as we have already seen. Total catch is also greatly inﬂuenced by the highly variable capture of certain species, such as anchoveta off Peru. It is more informative to consider catches from more speciﬁc regions. Biomass of reef ﬁsh in unﬁshed areas of Hawaii were more than 2.5 times larger than in intensively ﬁshed areas (Friedlander & DeMartini 2002). A time series of the catch of various marine organisms from the Northwest Atlantic off Canada during 1950–1997 (Fig. 11.13), reveals that in the mid-20th century the ﬁshery centered on cod and related species. The ﬁshing for herring-like ﬁsh increased after the mid-1960s. Cod regained dominance in the catch, through the 1980s, only to collapse—as we have noted above—by the mid-1990s. Flatﬁshes also became much scarcer during the latter period. Through all these ﬂuctuations, catches of other organisms varied little, or even increased, such as in the case of “other invertebrates”. These estimates show the many changes in the catch of different species throughout the period, and the

Figure 11.13 Time course of Canadian commercial catch of different kinds of ﬁsh and invertebrates, 1950–1997, in the Northwest Atlantic. From FAO (2002).

recent collapse that affected a few prominent species far more than others. The large decreases in recent catch are related to these important affected species, in this regional example as well as in the world catch. A thorough review of regional catch (Myers & Worm 2003) showed clearly that there has been rapid worldwide depletion of large predatory ﬁsh populations. Current abundance of populations of these predators seems to be about a tenth of the magnitude of population numbers existing before the development of industrial ﬁshing practices during the last half of the 20th century. Different types of organisms have different capacities to sustain ﬁshing pressure. Hutchins (2000) examined available time series data for 90 exploited stocks (including 38 species and 11 families of ﬁsh) that had declined in abundance for at least 15 years at some time in the history of exploitation. Of these 90 stocks, 8% recovered to earlier abundances within 5 years, 51% recovered to some extent, and 41% continued to decline in abundance. Species with faster reproductive rates and lower positions in the food web are more prone to recover from ﬁshing-related losses. Clupeids are therefore most likely to recover, with gadids, tunas, and ﬂatﬁsh less likely to recover. Other ﬁsh groups, for instance elasmobranchs (sharks, skates, and rays), show little ability to

259

HARVEST OF FINFISH AND SHELLFISH

retain abundance in the face of exploitation. Reviews of the status of elasmobranch populations subject to ﬁshing or caught as bycatch in different areas revealed signiﬁcant problems (Stevens et al. 2000).10 There was only one area with sustainable ﬁshing, one area where there was concern about declines in the ﬁshery, 16 areas with notable declines in stocks, seven areas with stocks close to local extinction, and three with stocks recovering after signiﬁcant declines brought about by overﬁshing.11 The human demand for marine products through the last half of the 20th century was high enough worldwide to lead to widespread reports of collapsing stocks. Press and media reports have made the facts about the overﬁshing of commercial stocks evident to an alarmed public. It is generally recognized that industrialized ﬁshing during the latter half of the 20th century, carried out mainly by trawlers, along with purse seining and longline ﬁshing,12 decimated many commercial stocks of the coastal world ocean.13 “. . . factory trawlers are destroying U.S. ﬁsheries and marine ecosystems” (http://archive.greenpeace.org). 10

Elasmobranchs mature only after several years of life, and bear a few live young at a time. These demographic features make for rather slow population growth, and hence any agent that removes a considerable part of the mature female population is a threat to the maintenance of the population. The same can be said about most large whales; the story of the decimation of large whale stocks by whalers in the Southern Ocean in particular, and the world ocean in general, is too well known to merit repeating here. Reviews of the facts about the effects of industrial-scale whaling can be found in Valiela (1995, chapter 9), Clapham et al. (1999), and Gerber et al. (2000). The Atlantic gray whale is the only large whale to have become extinct in recent history (Gerber et al. 2000). This whale was harvested until the 17th century, but there is little evidence to assess whether ﬁshing pressure or oceanographic changes brought about its disappearance. 11 Similar results were reported by Baum et al. (2003) for sharks in the Northwest Atlantic. Corroboration of the susceptibility of elasmobranch ﬁshes to harvest is found in the only cases where ﬁshingrelated near-extinction of ﬁsh species is suspected: these involve the Irish (Brander 1981) and the barndoor skates. Half a century ago, barndoor skates were reported in 10% of the tows taken off southern Newfoundland, but none has been found since (Casey & Myers 1998). 12 Detailed explanations of these methods, as well as the ships and gear, are available in FAO/FIIT Fishing Vessels Fact Sheets (http:// www.oceanatlas.com/world_ﬁsheries_and aquaculture/html). 13 The industrial-scale ﬁshing may compete in certain shallow waters with local small-scale artisanal ﬁshing. Large shrimpers, for example, may harvest (and even discard as bycatch) species sought by small boats in tropical waters by sustenance ﬁshers (Pauly et al. 2002).

Table 11.1 Compilation of the relative status of commercially exploited stocks, expressed as ﬁshing pressure vs. the sustainable limits, and grouped into different categories, for the world’s major commercial ﬁshed stocks, 1999. Data from FAO (2000). Status relative to sustainable fishing pressure Depleted Overfished Fully fished Moderately fished Underfished Recovering

% of stocks 9 18 47 21 4 1

In 2000, the Food and Agriculture Organization (FAO) undertook a review of the status of exploited stocks across the world (Table 11.1). These assessments are to some degree value judgements based on uncertain estimations of the available stocks versus speculative estimates of the catch that could be sustained by the populations. The assessments of Table 11.1 may be viewed by a pessimist as showing that by 1999 74% of commercial stocks were fully- or overexploited, and we have only 26% of stocks left to further increase harvests. An optimist might ﬁnd that only 27% of the stocks are in decline, and that 73% of the stocks may still provide sustained yields in years to come. In any case, the worldwide ﬁsh harvest seems near a point where we should not expect much greater yields of wild-caught marine protein.14 Some parts of the world ocean (e.g. the Indian Ocean, west-central Paciﬁc), however, still have relatively undeveloped ﬁsheries, with harvests still climbing (FAO 2002). Areas where harvests are still increasing are those that require the 14

As already mentioned, Myers and Worm (2003) reported a loss of up to 90% of the biomass of larger predatory ﬁsh from shelf and oceanic areas across many parts of the seas of the world between the middle and end of the 20th century. These estimates convey a more alarming message than the FAO results (Table 11.1). The different conclusions need reconciliation; perhaps the data for these two studies refer to different kinds and size of ﬁsh, and come from different areas subject to different ﬁshing pressures.

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development of higher technology and expensive investments to exploit, are distant from centers of population, and are of oceanic rather than coastal location. The considerable reduction of biomass of large ﬁsh that has taken place in so many places around the world’s coasts has involved the removal of major predators near the top of the coastal and marine food webs. Some time ago, I mentioned that most of our knowledge about the functioning of marine food webs came from systems where the top predators had been largely decimated (Valiela 1984, p. 268). There are many more recent reports that aver just how widespread and ecologically important such removals of top predators may be. Such removal of predators amounts to a wholesale potential release from so-called “topdown” controls of prey abundance and food web composition (Hughes 1994; Jackson et al. 2001; Myers & Worm 2003).15 If the claimed losses of larger top predators is true,16 the consequent release from top-down control must have led to either a marked bloom of benthic animals in the sea ﬂoor, or to increases in alternate predators. Some evidence for such effects is discussed below. We will return to this intriguing ecological dilemma in Chapter 14. Fishery managers have estimated sustainable yields from various stocks for many decades, although there are continuing discussions as to how to best assess what can be safely taken as a ﬁsh crop. This is a complex issue, not least because the target ﬁsh themselves eat or compete with other target ﬁsh. Blueﬁsh, a choice ﬁsh sought along the shallow waters of the eastern USA, for

example, is a voracious feeder that eats 4–12% of its body weight per day in prey ﬁsh (Buckel et al. 1999a). Blueﬁsh consume four times as much squid and eight times as much butterﬁsh than the harvest by commercial ﬁsheries, and one-sixth the commercial menhaden harvest (one of the largest US ﬁsheries) (Buckel et al. 1999b). To manage the blueﬁsh ﬁshery (as well as the squid and butterﬁsh catch), it might be useful to consider the relative role of the various species.17 Thus, arguments have been made for multispecies management plans, but these demand much additional information, which is usually not readily available. In any case, it is evident that sustainable ﬁshery management needs to be based on a multispecies basis, and much study effort is going into these practices. In addition to the individual ﬁsh brought to market, there are large portions of the netted ﬁsh that are too small to be lawfully marketed. These are discarded and are returned to the sea. Unfortunately, these discards have a low survival rate (Alverson et al. 1994). In the Northeast Paciﬁc, for instance, mortality owing to discards may range from trivial to more than 80% of total mortality, depending on the species of ﬁsh involved (Table 11.2). Mortality of discards too small to be legal or economically desirable hence adds to the toll suffered by the targeted populations; this additional source of mortality may not be considered in many estimates of stock losses, which would be, therefore, underestimates of total mortality. The mounting recent evidence shows that ﬁshing, if intense enough, does markedly reduce abundance of targeted species.

15

More detailed discussion of the idea of “top-down” vs. “bottomup” controls of coastal food webs is available in Verity and Smetacek (1996), Cury et al. (2000), and Reid et al. (2000), where a lot of empirical information and case histories are reviewed. 16 These estimates do not consider the larger abundance likely present before Westerners began ﬁshing in distant ocean waters. We only have hints of these pristine abundances through historical, circumstantial, and anecdotal evidence, as we saw in the case history of the shallows off eastern North America, where the historical abundance of ﬁsh and other stocks was hugely larger than that found at the beginning of industrialized ﬁshing. Valiela (1995, chapter 9), Jackson et al. (2001), and others have pointed out that most available estimates of stocks of ﬁsh and other top predators provide a much diminished view of the original abundance of marine stocks in many parts of the world.

17

There are many examples of similar situations where multispecies interactions may be essential to understand the dynamics of stocks. For example, populations of sand eels in the North Sea (Furness 2002) increased after ﬁshing lowered the abundance of its predators, including mackerel and groundﬁsh. This increase may not have been a result of release from predation pressure as sand eel abundance seems controlled by “bottom-up” processes rather than by “top-down” predation processes. Even as sand eel abundance increased, industrial harvest of sand eels (the largest ﬁshery in the North Sea) also increased from trivial to 106 tons yr−1, but is still modest in comparison to the consumption of sand eels by other ﬁsh, seals, and sea birds. Recovery of groundﬁsh stocks could be impaired by the present consumption patterns, or could result in competition with other wildlife.

261

HARVEST OF FINFISH AND SHELLFISH

larger individuals, as we saw in the Georges Bank cod ﬁshery. The mean weight of groundﬁsh caught in the Scotia Shelf off Canada decreased by about half as ﬁshing intensiﬁed between 1970 and 1998 (Zwanenburg 2000). The mean weight of skipjack tuna fell by a third between 1980 and 1998 in the equatorial Atlantic off the bulge of Africa, west of the Gulf of Guinea (Ménard et al. 2000). The ﬁshery for spiny dogﬁsh shark in the Georges Bank area increased ﬁve-fold between 1987 and 1993; this ﬁshery selectively took ﬁsh larger than 80 cm in length. As a result, although the abundance of the species did not change very much during this period, the ﬁsh caught became smaller (Fig. 11.14). Decreases in size of the larger top predators in food webs have consequences, because larger predators prefer and manage to eat larger prey (Valiela 1995). Where—through the action of ﬁshermen—the predators that remain in the environment are smaller, the food items they eat are perforce smaller. In marine systems, eating smaller prey means consuming prey that are at lower trophic steps, closer to the base of the food web consisting of producers such as algae. For example, during the end of the 20th century in the Northeast Atlantic, the ratio of ﬁsh-eating ﬁsh to plankton-eating ﬁsh in the ﬁshery catch decreased by half. The world ﬁshery is hence “ﬁshing down the food webs”, catching ﬁsh that are increasingly

Table 11.2 Annual estimates of discard mortality relative to ﬁshing mortality for Northeast Paciﬁc ﬁsheries. Adapted from Alverson et al. (1994), original information from US National Research Council and National Marine Fisheries Service.

Species

Discard mortality as % of total fishing mortality

Pollock Pacific cod Atka mackerel Rockfish Yellowfin tuna Sablefish Rock sole Flounder Pacific ocean perch Halibut

9.4 6.8 15.1 50.0 26.1 1.9 55.6 83.3 14.3 24.0

Reduced size of individual fish in stocks In general, ﬁshing targets the larger individuals, in a classic predator/prey strategy: the most valuable prey is the largest that the raptorial equipment (claws, teeth, or seines and nets) can handle. Hence, ﬁshermen selectively remove the largest species ﬁrst, and even within a population of ﬁsh, stocks that are ﬁshed tend to lose the

COMMERCIAL FISHERY 100

TRAWL SURVEY 100

Mean length (cm)

Females

Figure 11.14 Mean length of male and female spiny dogﬁsh shark in a commercial ﬁshery (left) and trawl (right) survey. Adapted from Rago et al. (1998).

Females 90

90 Males

80

80 Males

70 1980

1984

1988 Year

1992

70 1980

1984

1988 Year

1992

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smaller, and feed at lower steps within marine food webs.18 If the world ﬁshery is indeed catching ﬁsh lower down into food webs, we might reap a larger proportion of the energy ﬁxed by producers: the fewer the steps, the greater the harvestable portion of energy.19 On the other hand, by removing large top predator ﬁsh the ﬁshing ﬂeet might be unleashing new alterations of marine food webs. Thus, ﬁshing down food webs is a potentially important issue. The extent of ﬁshing down food webs, however, has been questioned. The calculated trophic shift for the world catch is rather small, and given the rough food habit and taxonomic identity of the catch data, the result may not be statistically or ecologically meaningful (Caddy et al. 1998; Caddy & Garibaldi 2000). Not all geographic areas show consistent ﬁshing down (Pauly et al. 2001; Badalamenti et al. 2002; Jennings et al. 2002). The lower trophic level of catch may also arise from other processes, for example, there may be a greater biomass of herbivores under increasingly eutrophic conditions or changes in global conditions (Caddy & Garibaldi 2000). This topic will proﬁt from more research on the exact trophic position of individual species making up the catch. Bycatch All ﬁshing methods to some degree harvest nontarget species. This unintended harvest has been referred to as “bycatch”, which is returned to the sea after the desired catch is removed. We have already noted the importance of discards from ﬁsheries; data for discards of too-small individuals of the target species and bycatch of species other than the target species are usually combined. 18

Pauly et al. (1998) concluded that ﬁshing pressure across the last half of the 20th century created a situation where the mean ﬁsh caught was smaller, and the world catch was at a signiﬁcantly lower mean trophic step. The estimated trophic level of the catch decreased from 3.3 to 3.1, where producers were considered to be at level 1. 19 This might occur because every time food is transferred across a feeding step (a trophic level), perhaps 90% of the assimilated food is dissipated by the respiration needed to support the organisms eating the prey. It is no accident, of course, that the largest ﬁsheries in the world, by tonnage, catch anchovies, sardines, and herring, all of which feed on plankton, reasonably low in their food webs.

Table 11.3 Comparisons of weights of landed catch and weight of bycatch plus discards in the world ﬁshery, for different groups of commercial species. Data compiled from records from the early 1980s to early 1990s. Summarized from Alverson et al. (1994).

Groups of species Shrimps, prawns Redfishes, basses, congers Herrings, sardines, anchovies Crabs Jacks, mullets, sauries Cods, haddocks, hakes Miscellaneous fish Flounders, halibuts, soles Tunas, bonitos, billfishes Squids, octopuses Lobsters Mackerels, snooks, cutlassfishes Salmons, trouts, smelt Shads Eels Total

(Bycatch + discards)/ landed weight × 100 520 63 12 249 28 20 10 75 18 9 55 3 5 10 84 35

Reviews of the available information (Table 11.3) show that perhaps 35% of the catch collected by ﬁshermen from the sea are unwanted.20 This impressive number becomes an additional loss to living populations, since survival of the released unwanted catch is thought to be low (Alverson et al. 1994). Long-lining ﬁshing practices have major effects on marine turtles, for example. Reviews of worldwide data show that more than 200,000 loggerhead and 50,000 leatherback turtles were caught as pelagic long-line bycatch (Lewison et al. 2004). Given the 80–95% decline in population numbers of these turtles during the last 20 years, losses as are occurring as bycatch are unsustainable. 20

In addition, some organisms that are caught by trawls manage to escape before the nets are brought on board. There is some additional mortality owing to this temporary capture. The majority of animals that escape the net survive, but there are taxon-speciﬁc differences; 9–10% of worms, 0–50% of shrimp and lobsters, 0–3% of snails and clams, 0–62% of seastars and related animals, and 6–97% of ﬁsh died in experiments done in actual trawl tows (Kaiser & Spencer 1995).

HARVEST OF FINFISH AND SHELLFISH

Other bycatch-prone harvests involve the ﬁsheries for shrimp, prawns, and crabs. These ﬁsheries kill several-fold more animal weight than they keep (Table 11.3) (Caillouet et al. 1996). Survival of vertebrates caught as bycatch may also be low. For instance, only ﬁve out of 1,263 mammals that were entangled were released alive from a set gillnet ﬁshery for large ﬁsh off California (Julian & Beeson 1998). Sea birds entangled in nets all die (Lewison & Crowder 2003). Other reports suggest 0% survival rates of mollusks caught as bycatch, and a higher or variable survival for crustaceans (50–100%), echinoderms (88%), and ﬁsh (0–97%) (Hill & Wassenberg 1990). Survival of bycatch is variable, but can be low in many instances. Bycatch discarded back into the sea is then available for consumption by scavengers and detrital feeders, including sea birds (gulls, gannets, petrels, skuas, shearwaters, and so on) and any other predators or scavengers. The amounts of food furnished as bycatch are signiﬁcant. In the North Sea, bycatch weight amounts to about 22% of total ﬁsh landings (Garthe et al. 1996)— and could potentially supply 1.6 times the energy required by the entire sea bird population of the North Sea. In the western Mediterranean, bycatch released from trawlers could support a gull population four-fold larger than was present (Martínez-Abraín et al. 2002). Discards and offal amounted to up to 70% of the diet of adult great skuas when their preferred prey, sand eels, were abundant, and 82% when sand eels were scarce (Furness & Hislop 1981). Breeding success was greater for Audouin’s and yellow-legged gulls during years of high trawling activity near the Ebro delta (Oro et al. 1995, 1996). Feeding on discards may be less important for other birds, such as gannets and kittiwakes, but there is no doubt that bycatch discards have considerable inﬂuence on sea bird ecology and distribution. Altering the nature of well-established ﬁsheries to reduce bycatch— obviously still a desirable goal—will no doubt change the abundance and distributions of the many sea bird populations that have come to depend on the dumped discards and offal (Garthe et al. 1996). Feeding by sea birds on bycatch has received the most attention, but other species also feed on

263

discarded material. On average, perhaps only 76% of “roundﬁsh”, 24% of ﬂatﬁsh, 12% of sharks and rays, and 9% of invertebrate weight discarded is likely to be eaten by scavenging birds (Garthe et al. 1996). The rest sinks and may be eaten by other animals in the water column (Hill & Wassenberg 1990; Castro et al. 1999) and sea ﬂoor (Bozzano & Sardà 2002). The amount of discard biomass reaching the sea ﬂoor is sufﬁcient to alter the benthic fauna (Groenewold & Fonds 2000). Certain ﬁsh actively follow recently trawled paths along the sea ﬂoor, apparently feeding on the bottom fauna stirred up by the disturbance of the trawl (Kaiser & Spencer 1994). This is very similar to the familiar sight of clusters of gulls, terns, cattle egrets, and other birds eagerly following plows as the soil is turned over in agricultural areas. There is another sort of bycatch, a delayed impact involving the entanglement of marine species in lost ﬁshing gear. This subject is taken up in more detail in Chapter 13, but here it sufﬁces to say that entanglement may be a serious conservation issue (Lewison et al. 2004) Altered composition of species Increased ﬁshing pressure, by removal of the top predators, through the action of bycatch, or owing to habitat alteration, leads to powerful shifts in the species composition of the exploited food webs (Jackson et al. 2001; Worm & Myers 2003; Lotze & Milewski 2004; Lotze 2005). The removal of top predators may release pressure on the array of prey, perhaps allowing a proliferation of the prey, or of other predators that have the opportunity to use the more abundant prey (Valiela 1995). We saw, in the earlier discussion of the indirect effects of ﬁshing in Georges Bank, that the removal of cod and other species targeted by the ﬁshery was followed by an increased abundance of dogﬁsh sharks and other ﬁsh (see Fig. 11.11) (Murawski & Idoine 1992; Fogarty & Murawski 1998; Jennings & Kaiser 1998; Rago et al. 1998). Although earlier results were equivocal, there is accumulating evidence for the restructuring of coastal food webs after the reduction of top predator abundance. There were no consistent species shifts in other heavily ﬁshed areas such as the Scotian Shelf (Duplisea et al. 1997) or the

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North Sea (Pope et al. 2000). Initial work on the effects of ﬁshing intensity in Fiji (Jennings & Polunin 1997) showed no detectable rearrangement of species, even at high ﬁshing pressures, but more recent reports suggest that declines in predatory ﬁsh were associated with increases in abundance of other taxa, such as the coral-eating crown-of-thorn starﬁsh (Dulvy et al. 2004). The species shifts brought about after the reduction of top predators are further documented in work by Friedlander and DeMartini (2002) on ﬁshing impacts on Hawaiian reefs, and analyses by Pandolﬁ et al. (2003) on reefs in general. It has been said that overﬁshing in cases such as Georges Bank has led to a degradation of the system as a result of the replacement of the gadoid-dominated food web by a dogﬁshdominated successor. This seems a value judgement, much like the issue that we encountered in the matter of alien versus native species (Chapter 10). Ecologically, the Georges Bank ecosystem is still actively functioning, albeit with a somewhat different, but still abundant, array of living things. It seems presumptuous to conclude whether the new array is better or worse ecologically. In a world with ever-changing species abundance and composition, and no agreed-upon currency with which to make ecological evaluations, it seems difﬁcult to decide whether some array constitutes a more degraded state than another array. It is evident that for the ﬁshing community, the remaining food web in Georges Bank is less economically valuable; in Georges Bank, where cod were replaced by dogﬁsh and other species, the economic beneﬁt to the ﬁshery might be altered by the relative price of the ﬁsh harvests, and hence is “degraded”, but on an ecological basis such a description seems less apt. In any event, accumulating evidence points to widespread shifts in the composition of the species in coastal food webs where harvests have removed top predator species. Disturbance of habitats In certain coastal areas—the South China Sea (Pitcher et al. 2000), the North Sea (Rijnsdorp & van Leeuwen 1996), coastal waters of the Northwest Atlantic (Churchill 1989), and many others

—the intensity of trawling21 has been remarkable across relatively long periods of time. In Tolo Harbor, Hong Kong, each square meter is trawled three times a day, and catches have declined sharply during the last decade (Pitcher et al. 2000). Records of trawling by US ﬂag ships during 1985— soon after the heyday of ﬁsh harvests (see Fig. 11.9) —show that in certain sectors of the Middle Atlantic Bight the area swept by trawl gear was up to 413% of the total area of sea ﬂoor (Fig. 11.15). Earlier, in 1981, trawlers of non-US ﬂag ships scoured 218–297% of the sea ﬂoor shallower than 200 m; in 1985, the foreign trawling effort still covered 32–81% of the areas shallower than 200 m (Churchill 1989). As early as the 14th century there have been warnings about the damage to ﬁshing grounds from ﬁshing gear (de Groot 1984). Dragging dredges and trawls over the sea ﬂoor removes fauna and damages the surface of the bottom (Fig. 11.16). Trawling, for example, leaves long trails of disturbed bottom where the doors and gear are dragged (Collie et al. 2000a; Pranovi et al. 2000).22 The effects of various methods of trawling and dredging are quite variable (Table 11.4); in part, the considerable spatial variability that characterizes benthic faunas impairs the ability to accurately measure the effects of disturbances. A compilation of many studies showed that the percentage change in abundance of benthic fauna after initial trawling and dredging ranged between +3 and −59% relative to initial abundances (Collie et al. 2000b). 21

In this section we focus on disturbances created by trawling or dredging from the surface because these are the most widespread practices, and have been most frequently studied (Collie et al. 2000b). Other, less widely used, forms of ﬁshing can create severe, but more local, habitat damage (Pet-Soede et al. 1999; Kaiser et al. 2002). Use of dynamite to stun or kill ﬁsh on coral reefs has been described for many tropical areas. Dredging for oysters in shallow waters can degrade oyster reefs (Lenihan & Peterson 1998). In the Philippines muroami ﬁshing uses stones, chains, or poles to break up corals and drive ﬁsh unto nets. Other subsistence ﬁshermen use cyanide to poison ﬁsh in an area. These practices leave devastated habitats behind, damage that takes many years to restore. 22 Trawling also potentially leads to resuspension of sediments, disturbances to the natural topography, reduction of bioturbation by native organisms, and possibly to increased mineralization of organic matter in the surface sediments (Kaiser et al. 2002). These effects occur but have been scarcely quantiﬁed. Further, the disturbance may impair recruitment of the target species, thus exacerbating impacts of industrial ﬁshing (National Research Council 2002).

265

HARVEST OF FINFISH AND SHELLFISH

76° 42°

75°

74°

73°

69°

70°

71°

72°

158 140 145 95

41° NEW YORK

Long

Island

50

75

89 321 413 271

19

10 27

67

35

48

0

0

0

1

1

22

10

0

0

0

0

0

46 131 93

8

2

6

19

53 107

7

m

40°

86

15

28

34

19 25

0

33

19

29

0

0

19 12

76

3

0

5

44

0

1 3 32 9

13

0

1

1

0

0

0

0

law

0

14

2

52

0

0

0m

20

19 0

2,000 m

are

39°

31 m

60

2

De

20

0

y

Ba

38°

36°

00

69

4,0

Figure 11.15 Distribution of trawling effort by US vessels during 1985. The numbers inside each box within the map indicate the ratio between the cumulative area trawled by ﬁshermen relative to the area of the box. From Churchill (1989).

m

37°

Faunas in areas subject to frequent natural disturbances are less affected by trawling disturbances (Kaiser et al. 2000; Kenchington et al. 2001; Tanner 2003). Soft-bodied, larger, slow-growing species are more susceptible to disturbances (Collie et al. 2000b). The level of disturbance is closely tied to frequency of the ﬁshing. A trawl disturbance experiment in the Great Barrier Reef showed that each trawl caught 5–20% of available benthic biomass, with 70–90% removed after 13 trawl passes (Poiner et al. 1998). Sites in the North Sea trawled fewer than 2.3 times per year showed little damage to benthic fauna, but those dredged more than 6.5 times per year held fewer and smaller individuals, and certain urchins and bivalves were absent (Jennings et al. 2001). Intense trawling in an area leads to the preponderance of small,

fast-growing species, and lower overall abundance of living bottom organisms (Kaiser et al. 2002). Recovery from the effects of ﬁshing disturbance is rather variable. Recovery may be rapid in physically unstable habitats, where the fauna are accustomed to disturbances (Collie et al. 2000b). In general the data from studies are quite variable so it is difﬁcult to pinpoint recovery (Collie et al. 2000b), but it seems that in most benthic assemblages, the effects of trawling disturbances are hard to detect some time between 10 and 500 days after the disturbance. For certain species, recovery may be slower: in the case of certain Australian sponges recovery may take 15 years in trawled areas (Sainsbury 1987). Where the frequency of disturbance is very high, recovery may not occur.

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Figure 11.16 Photographs of two 50 × 75 cm patches of sea ﬂoor in a gravel habitat, northern Georges Bank, November 1994. Colonial worms and hydroids are common in an undisturbed parcel of the sea ﬂoor (top); far fewer organisms are visible in a less disturbed area (bottom). Note, however, that the differences in the two sites were presumably caused by differences in disturbance due to trawling and by original differences in the faunas of the two sites. From Collie et al. (2000a), photos by Dann Blackwood, US Geological Survey.

Remedies for overfishing Depletion of commercially important ﬁsh populations poses a number of issues. The ﬁshing pressure exerted on marine environments at the start of the 21st century appears to be poised at near the limit of harvestable yield. Government agencies and distinguished panels (Lubchenco et al. 1991; National Research Council 1999) conclude that overﬁshing is a widespread reality, and insist that the goals of management of ﬁsheries aim toward sustainable ﬁshing efforts. Such sus-

tainable ﬁshing plans would not only prevent collapse of stocks, but also maintain “ecosystem health” and “ecosystem integrity”. Moreover, the management plans guiding the ﬁshing effort ought to consider the range of “ecosystem services and beneﬁts” provided by the exploited environments, not just the beneﬁts to the ﬁshing sector. While admirable in their intent, there is little agreement on how these valuable concepts might be implemented, since they are not operationally deﬁnable. Even less is agreed upon regarding just what the relevant services and beneﬁts might be.

267

HARVEST OF FINFISH AND SHELLFISH

Table 11.4 Effects of trawling on the species of the sea ﬂoor, in different sites and ﬁsheries, reported in a few studies. Comparisons with a larger set of studies available in Collie et al. (2000b). Site

Effect on benthos

Type of fishery

Source

Irish Sea

In one site, 49% of invertebrate groups had lower abundance; in another site, 0%

Sole and plaice

Kaiser & Spencer 1996

Irish Sea

None offshore, slight in shallow site

Norway lobster

Ball et al. 2000

Grand Banks

Biomass of large-size benthos 24% lower on average, but variable responses; some species increased

Experimental otter trawling

Prena et al. 1999

North Sea

20–65% mortality for bivalves, 5–40% mortality for other invertebrates; 29% of taxa showed significant decreases

Flatfish

Magda et al. 2000

Off NW Australia

Density of sponges, gorgonians, and soft corals decreased; mortality < 10% for most of area fished

Scalefish

Moran & Stephenson 2000

N Adriatic

Variable and slight effects, recovery after 1 week

Scallops, flatfish

Pranovi et al. 2000

NE Atlantic

Damage to cold-water coral reefs and associated benthos, not quantified

Various fish

Hall-Spencer et al. 2002

Barents Sea

Damage could not be detected; more shellfish, medusae, and snails

Experimental single trawling Kutti 2002

Bering Sea

Fewer sedentary benthos, mixed response in motile benthos

Long-term mixed fishery

McConnaughey et al. 2000

Adriatic Sea

No change in motile benthos, decreases in sessile organisms

Scallops

Hall-Spencer et al. 1999

While setting out desirable goals is admirable, addressing the issue of overﬁshing23—as in many other environmental questions—needs an operational, practical basis (Larkin 1996; Valiela et al. 2000). Although the grander ideals might not be deﬁnable, we can make a more prosaic list of things that could feasibly be done that might reduce the impacts of overﬁshing. Regulation of fishing Reduction in ﬁshing pressure has led to recovery of exploited stocks. Some of the evidence for such 23 Murawski (2000) pointed out that it is even difﬁcult to deﬁne just what overﬁshing is. Each target species and method of harvest offers speciﬁc details applied to speciﬁc parts of the sea. This does not even start to consider the interspecies and ecosystem couplings of each stock type, let alone the other ecosystem services and beneﬁts.

recovery after decreased ﬁshing pressure was mentioned in earlier sections of this chapter, but there is no better example of recovery than the case of the striped bass in the northeast coast of North America (Richards & Rago 1999). Adult striped bass live to perhaps 30 years, and return to natal areas in brackish to freshwater upper reaches of estuaries to spawn. The juveniles live in estuarine waters for several years, and move to feed and overwinter in deeper coastal waters as adults. Most of the striped bass along the coast of North America from North Carolina to Maine, spawn in Chesapeake Bay or the Hudson estuary. Commercial catch of bass declined during the 19th century, reaching a low in the 1930s—probably an effect of increased construction of dams and water pollution in the estuaries. Unexplained strong recruitment

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Striped bass off the northeast coast of North America Striped bass have played a prominent economic role in the northeast USA since colonial times (Richards & Rago 1999); the ﬁrst public school house was built in 1670 partially with funds from sales of striped bass. Harvests were intermittently low and high, but as early as the mid-1700s there were some expressions of concern with the “very great numbers having been imprudently, or rather wantonly taken in one season.” The ﬁshery expanded nevertheless, and by the 1980s, striped bass contributed about US$200 million to the economy of the northeast US coastal region. On November 3, 1981, Tony Stetzco caught a world record striped bass off Nauset Beach, on the outer coast of Cape Cod, Massachusetts. At 57 in and 73 lb, it was the largest striped bass ever caught surf casting from a beach. But it was an entirely deceptive clue; the ﬁshery was to soon suffer a steep decline. “It hit me one day, when I had 40 giant bass stacked behind me, that it wasn’t right”, Stetzco said (Cape Cod Times, June 5, 1997).

invigorated the ﬁshery after 1934, with a peak about 1970 (Fig. 11.17 top). The take of bass declined measurably in the following decades (Fig. 11.17 middle), enough to create concern for this valuable resource; the crash prompted research as to the causes of the lowered catch. Evidence accumulated that ﬁshing pressure was large enough to impair recruitment; contamination of spawning areas might have had some impact, but no clear evidence of the action of other factors appeared. From the results of the studies, strict state and federal regulations were put in place during the early 1980s to protect maturing females, and eventually an Act of Congress promoted a near-complete moratorium on striped bass capture. In addition, moni-

toring and restocking of striped bass were made part of the management plan.24 The result of strict regulation of the harvest of striped bass was that during the late 1980s there were clear signs of recovery in the striped bass population. Gradually, the protected young bass grew, and survived into maturity (past the fourth year of life or so) and became a signiﬁcant part of the now more numerous population. Much of the catch during these decades was recreational, and was released back because the ﬁsh were below the regulatory limits (Fig. 11.17 bottom). After a good recruitment year in 1989, the “juvenile index” (Fig. 11.17 top) increased on average. Less restrictive ﬁshing limits were allowed in subsequent years as the stocks recovered. “In my over 30 years in business, I have never seen such a great striper year—everyone is catching bass” (Karen Hill, owner of the Sports Port on Main Street, Hyannis, in the Cape Cod Times, June 5, 1997).

It is certainly the case that external—probably large-scale and global—factors created good and poor years of recruitment in the striped bass ﬁshery. It was, however, the protective measures put in place, and strictly enforced, that made it possible for the recruitment to improve. This example is one of outstanding success in ﬁshery management, and demonstrates that under favorable circumstances, with appropriate research basis, timely action, and strict enforcement, overﬁshed populations can recover and provide sustained harvests. Not all overﬁshed populations of ﬁsh show ready recovery after the easing of ﬁshing pressures. California sardines (Velarde et al. 2004), crabs in the Gulf of Alaska (Orensanz et al. 1998), and several groundﬁsh in Georges Bank (Fogarty & Murawski 1998) showed slow signs of recovery after ﬁshing pressure was lowered. 24 Later studies of the results of management of striped bass demonstrated that lowered ﬁshing pressure was far more effective than stocking (Richards & Rago 1999). Although stocking may be useful locally, it does not solve the root cause of overexploitation, and should not be used as an excuse to avoid having to regulate ﬁshing take.

269

60 40 20 0 8 6 4 2 0 12

Recreational catch (× 103 metric tons)

Figure 11.17 Time course (1954–1996) of abundance and catch of striped bass in the Maryland side of Chesapeake Bay. Top: index of juvenile abundance (index estimated by Maryland Department of Natural Resources from the mean number of ﬁrst-year bass seined from nursery areas). Middle: commercial landings of striped bass (Boreman & Austin 1985). Bottom: catch by recreational ﬁshermen along the Atlantic coast of the USA (North Carolina to Maine), 1981–1995 (data from US National Oceanic and Atmospheric Administration and US National Marine Fisheries Service). Adapted from Richards and Ragu (1999).

Commercial landings (× 103 metric tons)

Juvenile abundance index

HARVEST OF FINFISH AND SHELLFISH

Total Kept

8

4

0 1954 1958 1962 1966 1970 1974 1978 1982 1986 1990 1994 Year

The responses of marine mammals to protection from harvest are also at best variable (Gerber et al. 2000). Populations of North Paciﬁc gray whale are recovering, after protection from harvest, with a remarkable rate of population growth of 2.5% per year. The Paciﬁc gray whale, now protected, has been removed from the endangered and threatened species lists and continues to increase to more than 20,000 individuals. A population of Arctic bowhead whales that is hunted by Alaskan subsistence hunters is recovering, even as another four populations of nonharvested bowheads are not. The North Atlantic right whale, although rigorously protected, is not showing signs of recovery, perhaps owing to numbers low enough (300–350 individuals) to impair social interactions, and to accidental mor-

tality from ship collisions and entanglements in ﬁshing gear. It therefore remains to be seen just how general the striped bass example of recovery from ﬁshing pressure might be. Probably the key is strict enforcement of ﬁshing restrictions, reduction of incidental mortality, and habitat protection. The recovery of certain stocks may also be delayed by ecological reasons, such as the establishment of competitors or predators of juveniles during the period when the original stocks became rare. Alternatively, it may be that changing external factors might constrain recovery. In any case, lowering ﬁshing effort not only stops further crises of exploited populations, but may enable recovery, which may be a matter of a few years or longer, depending on circumstances.

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Bycatch-reducing devices (BRDs) A number of countries are moving toward forcing ﬁsheries to lower bycatch damage, mainly to ﬁsh species and to endangered species, by regulations that BRDs be added to ﬁshing gear. BRDs are modiﬁcations of trawling gear that allow some of the captured species to escape while retaining the target species. Different types of BRDs have been devised to reduce the bycatch of unwanted ﬁsh, invertebrates, turtles (Crowder et al. 1994), sea snakes (Milton 2001), birds (Darby & Dawson 2000; Tasker et al. 2000), and marine mammals (Caswell et al. 1998; Julian & Beeson 1998). The questions regarding this management option are whether they work at all, and whether they can be quantitatively useful. The relative importance of bycatch mortality depends on the magnitude of this source of mortality relative to other mortality sources and to the status of the population concerned (Caswell et al. 1998). Several examples demonstrate that bycatch mortalities are large enough to have significant demographic consequences. Populations of Atlantic croaker suffer bycatch mortality during shrimp ﬁshing (Diamond et al. 2000). The mortality rates were three times as large in the Gulf of Mexico compared to the Atlantic, and the stocks in these two areas have declined accordingly. Reductions of bycatch mortality by 5% in the Atlantic, and 35% in the Gulf of Mexico, could reverse croaker declines. Loggerhead turtles become entangled in nets, and about 70–80% of loggerhead deaths are thought to be related to ﬁsheries. Model predictions suggest that stringent use of turtle excluder devices could allow recovery after several decades (Crowder et al. 1994; Spotila et al. 2000). Mortality of the black-footed albatross in the central North Paciﬁc may reach 5,000–10,000 birds per year (Lewison & Crowder 2003). These mortalities, if they continue, may effectively lower albatross populations over the next three generations (about 60 years). Bycatch mortality of dusky dolphins off Patagonia is large enough to constitute a threat to the abundance of these mammals (Dans et al. 2003). Some items appear in bycatch rarely, but if the species are in fact rare, as for example the 13 sea snake species found off Australia, the loss to

bycatch may still have serious consequences. Different species of sea snakes tolerate bycatch to radically different levels; bycatch, particularly in the prawn ﬁshery off Australia, may be lowered by the use of turtle excluder devices recently ordered by law (Milton 2001). Possibly, the loss of 73 yellow-eyed penguins in gillnets in New Zealand waters between 1979 and 1997 might represent a threat, given that there are only 1,400– 1,700 breeding pairs in the entire world (Darby & Dawson 2000). Marine reserves The establishment of areas where ﬁshing is prevented is an appealing idea, and many papers (e.g. Davis 1995) and international and other agencies have espoused their creation.25 Marine reserves could protect spawning populations and habitats, as well as export larvae and adults to ﬁshed areas, but there are some difﬁculties with the concept. First, of the many reserves [over 1,300 worldwide (Kelleher et al. 1995)] that exist on paper, few [only 9% (Kelleher et al. 1995)] actually operate so as to effectively control ﬁshing, particularly those in third world countries (McClanahan 1999). Control of ﬁshing in reserves requires resources and institutional will seldom available in many places. The creation and enforcement of reserves more often than not are governed by economical and political realities rather than by environmental criteria.26 Second, it has been difﬁcult to assess the effects of reserves, in part because of insufﬁciently critical data. For example, the abundance of most commercially sought species (Fig. 11.18, black circles) recorded within areas of Kenyan reefs subject to ﬁshing by artisanal means were lower than those recorded in areas protected within a reserve. Species of ﬁsh that were not commercially sought also showed the same pattern (Fig. 11.18, open 25 Useful reviews of the beneﬁts and drawbacks, and much else about marine reserves, are compiled in a series of papers appearing in Bull. Mar. Sci. 66 (2000) including Chiappone and Sullivan Sealey (2000). Additional discussion is provided by Roberts and Polunin (1991, 1994), Wantiez et al. (1977), and Tuya et al. (2000). 26 Roberts (2000) is of the opinion that, in fact, it matters little where reserves are established, and that reserves lead to biological beneﬁts regardless of where they are located. Others think that more ecological information might improve site selection as well as chances of measurable effects (Crowder et al. 2000).

HARVEST OF FINFISH AND SHELLFISH

Not fished

100

Commercial spp. Non-commercial spp.

1:1

10

1

0.1 0.1

1 10 Exposed to fishing

100

Figure 11.18 Comparison of abundance (mean number of individuals per 2,500 m2) of ﬁsh species in a Kenyan reef area with no ﬁshing at all, and an area where artisanal ﬁshing was allowed. Species of commercial interest are shown as well as species not of commercial interest. Data from Watson and Ormond (1994).

circles). These results—because of lack of appropriate control observations—could be interpreted as a lack of effect of ﬁshing, or perhaps that the non-ﬁshed area was originally more suitable for most ﬁsh. Such results therefore resist unambiguous interpretation. A few other studies furnish more compelling information. Fish in areas off Cape Horn that were protected from harvesting were more abundant and larger, on average, than in areas where ﬁshing was permitted (Buxton & Smale 1989). Certain areas of reef in the Philippines were protected from ﬁshing since the 1970s (Russ & Alcala 1989). In one such area, ofﬁcial protection measures broke down in 1984, and artisanal ﬁshing with nets and dynamite blasting took place, while in a similar area, protection was maintained. Comparison of the kinds and numbers of ﬁsh in these two areas provided evidence of the effects of ﬁshing pressure. The numbers of species did not change signiﬁcantly—as is the case with virtually all ﬁshing, extinction is rarely an issue. The abundance of six of the species decreased signiﬁcantly, four did not change, and two increased signiﬁcantly in the ﬁshed reserve; in the nonﬁshed reserve, one species signiﬁcantly decreased in abundance, and all others were unchanged

271

(Russ & Alcala 1989). Fishing hence created change, not necessarily all leading to a consistent decreased abundance of ﬁsh species. From these examples, it appears that reserves tend to foster ﬁsh abundance, and allow larger ﬁsh classes to survive, but results are quite variable. Part of what makes it difﬁcult to assess the effects of reserves on ﬁsh abundance is that effects may not appear for some time. In addition, the efﬁcacy of a reserve may depend on the intensity of ﬁshing effort exerted on the stocks.27 Much theory has been published on this aspect, but empirical studies of the effects of ﬁshing intensity on ﬁsh population abundance are few. Both these objections are addressed in a study by Jennings and Polunin (1997), who made use of the long-term existence of qoliqoli in Fiji, welldeﬁned areas of coral reefs whose artisanal ﬁshing rights were traditionally restricted to people living in speciﬁc villages. As it turned out, these qolicoli were exposed to a 60-fold range of ﬁshing intensities, ranging from 4 to 243 people per kilometer of reef front depending year-round on the protein obtainable from the qolicoli. One hundred and forty-four ﬁsh species—belonging to six different taxonomic families of ﬁshes—were found in the surveys. Many were target species harvested by the villagers, others were not target species. The impacts of ﬁshing, even at the highest intensities, were modest. There were no detectable effects of ﬁshing on abundance for four of the families of ﬁshes, and clear decreases in two of the families. Abundances of the two affected families decreased by 75 and 83%, and most of the decreases occurred within the lower range of ﬁshing pressure. These two families were evidently particularly sensitive to even low ﬁshing pressures; groupers, one of the affected families, are slow growing and long lived, and their populations thus are highly susceptible to harvests. These studies showed that, as we saw before, longterm protection afforded by reserves relative to 27 Success of a reserve may also be affected by the proportion of the stock or habitat encompassed by the restricted areas, and by the relative “porosity” of the reserve boundaries. “Porosity” in this context refers to the extent of movement in and out of the restricted areas by species to be protected or species (predators, parasites, diseases, etc.) that affect the species to be protected (Murawski et al. 2000).

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Table 11.5 Comparison of the abundance and biomass of Atlantic sea scallop in the Georges Bank area; data calculated from dredge surveys done inside and outside areas closed to harvest, during 1994 (before establishment of the closings) and 1998 (after the closings). Data adapted from Murawski et al. (2000).

Year Abundance (mean individuals tow−1)

Biomass (kg meat

tow−1

)

Outside closed areas

% difference, inside/outside

59 588

35 196

+68 +200

+890

+454

Inside closed areas

1994 1998 % difference, 1998/1994 1994 1998 % difference, 1998/1994

ﬁshed areas was variable, and was more dependent on taxonomic identity of the ﬁsh species than on ﬁshing intensity. More recent reports of the effects of ﬁshing in Fijian reefs showed more consistent impacts of local ﬁshing (Dulvy et al. 2004), so that one might suspect that protection from ﬁshing will have a beneﬁcial effect. Broad reviews of results of studies from 31 temperate and tropical locations where marine reserves were established show great variation in reponses by various species and among sites (Micheli et al. 2004). On the whole, signiﬁcant, but small, beneﬁts accrued from the establishment of reserves, but there was so much variability that it seemed difﬁcult to generalize. A similar study of 80 reserves showed higher average abundance, variety, and size of organisms within reserves compared to control areas (Halpern & Warner 2002). Surprisingly, the increases were similar regardless of age of the reserves, perhaps suggesting that the reserves were placed, reasonably, where the local populations were larger and richer to begin with; this makes sense, but may make assessment ambiguous. We saw earlier that the obvious collapse of commercial ﬁsh stocks in Georges Bank led to closures of signiﬁcant areas to ﬁshing, beginning in 1994 and continuing through more recent years. Even though the industrialized harvest pressure reached overwhelming levels, it still has been hard to assess effects of the closings on ﬁsh, in part because of the long-distance movements of the

0.6 8 +1,310

0.3 0.5

+52 +806

+57

different life stages. More conclusive information on the effects of the post-1994 ban on ﬁshing in selected areas of Georges Bank—which therefore became a kind of reserve—is available from dredge surveys of less mobile sea scallop populations (Table 11.5). There were relative increases in abundance and biomass in 1998 relative to 1994 in all sites, probably simply a result of large-scale, externally driven improvements in recruitment or conditions. More relevant to our topic here is that the increase seen within closed areas was far larger than that recorded outside the areas closed to ﬁshing: scallops did respond to a halt in ﬁshing. Changes such as seen in sea scallops in Georges Bank show that closing areas to ﬁshing provided the opportunity for establishment—or recovery— of substantially larger populations within a few years after the closing. Marine protected areas do work, therefore, at least in certain places, and with certain populations. Difﬁculties of measurement and assessment impede ascertaining the degree to which we can generalize the beneﬁts of reserves, but they certainly do no ecological harm, and potentially are environmentally beneﬁcial. Mariculture Hunting-gathering sufﬁced early in human history to supply the nutritional needs of scattered bands of people. As humans became more numerous, it became more and more difﬁcult to support growing numbers by simply harvesting existing wildlife

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100

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MARINE

Capture Aquaculture

Yield (million tons)

80

60

40

20

0 1994

1995

1996

1997

1998

1999

Yield (million tons)

40

2000

2001

FRESHWATER

20

0 1994

1995

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1997 1998 Year

1999

and plants,28 and people turned to a variety of agricultural practices, which on aggregate provided much larger yields of food per effort and per unit of land. In contrast, ﬁshing—the catching of aquatic wildlife—has continued into our days, and reached industrial-scale levels. It is no wonder that, in spite of the enormous richness of coastal seas, ﬁsheries for most target ﬁsh are seldom sustainable (Pauly et al. 2002), and many, as we have seen, are in decline.

2000

Many observers argued that that we ought to shift away from our ancestral hunter-gatherer patterns and shift to the agricultural model, and propose that aquacultural methods be developed intensively. The recent decline of marine stocks created markets for maricultural products,29 and production by mariculture has increased, albeit slowly in recent decades; the use of cultured products still lags far behind the capture ﬁshery harvest (Fig. 11.19 top). Freshwater maricul29

28

This shift has created much ecological change. There are students of early human communities who argue that hunting pressure by increasingly numerous groups of humans might be at least in part responsible for the extinction of many species of large mammals, such as mammoths and mastodons, and it is well known that miscellaneous agricultural and animal husbandry practices have changed landscapes of very large areas, such as the Mediterranean countries, the prairie-based breadbaskets of the world, and more recently, rainforest areas of the tropics.

2001

Figure 11.19 Yield of worldwide marine (top) and freshwater (bottom) capture ﬁsheries and mariculture, 1994–2001. Data from FAO (2002).

There were still somewhat more than 25 million people involved in ﬁshing as a livelihood during the 1990s (FAO 2002), but this industry is in steep decline. The number of ﬁshing vessels remained about the same, somewhere between 1 and 1.2 million (FAO 2002), but the number of vessels lost or scrapped was considerably larger than the number of vessels built. In contrast, through this same decade, the number of people employed in aquaculture doubled, up to about 7 million worldwide, with particularly large increases in China (FAO 2002). Maricultural yields of ﬁnﬁsh, bivalves, seaweeds, and shrimp and prawns for the world was about 14 million tons, up by almost an order of magnitude since 1970 (FAO 2002).

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ture, particularly in China, has increased more markedly, and its harvest is more than twice that of inland ﬁsheries (Fig. 11.19 bottom). Note, however, that the sum of marine plus freshwater mariculture still was, as of 2001, less than half the harvest by the marine capture ﬁshery. For now, we are faced with increasing maricultural efforts, and a clearly stagnant capture ﬁshery. Increased mariculture,30 however, is a mixed blessing (Naylor et al. 2000). Cultured shrimp and prawns, bivalves, seaweeds, salmon, and other crops provide much needed sources of employment and capital to a needy sector of local populations, as well as furnishing much-desired crops that are readily sold. Mariculture offers the additional beneﬁt that it may in part replace exploitation of wild stocks, although in practice this potential seems unfulﬁlled. On the other hand, as mariculture has developed, a number of problems have arisen, including destruction of coastal habitats, alteration of waters and sediments by waste efﬂuent disposal, depletion of stocks of wild larvae and juveniles, and introduction of alien species and diseases.31 Destruction of coastal habitats

We have already discussed the widespread loss of mangrove habitat that results from the construction of maricultural ponds for the culture of shrimp and prawns in Chapter 6. 30

Mariculture is not a single practice; there are as many methods as crops. Fish are cultured in cages suspended near-shore, grown in ponds constructed for the purpose, or in tanks (and fed food made up of processed smaller ﬁshes or grains). Shellﬁsh are usually grown as hatchery or wild-caught juveniles placed on shallow sediments, cages, or hanging from ropes or other structures, all where the shellﬁsh can feed on the phytoplankton grown naturally in rich waters. Shrimp and prawn, the largest maricultural crop, are grown from wild-caught larvae in shallow ponds expressly made for culture, often by digging out coastal wetlands. Reviews of techniques, environmental consequences, and other aspects are available in Pillay (1992) and Midlen and Redding (1998). 31 To these environmental issues we might add that there is ample evidence that contaminants such as mercury, polychlorinated biphenyls (PCBs), and chlorinated hydrocarbons are appearing in cultured crops such as salmon (Easton et al. 2002; Knowles et al. 2003; Berntssen et al. 2004). The contaminant concentrations derive from the high food web position of such ﬁsh, and from the accumulation of mercury in the foods fed to the farmed salmon and other crops. The issues of bioaccumulation were discussed in previous chapters.

Alteration of water and sediment

Maricultural practices perforce release materials that alter the quality of water downstream (Gowen & Bradbury 1987; Ervik et al. 1997). Nitrogencontaining wastes are routinely released from maricultural facilities,32 sharply altering virtually all aspects of the nitrogen (and other nutrient) cycles (Tenore et al. 1982; Hargreaves 1998). The shrimp pond mariculture that has so diminished the area of mangrove habitats (Chapter 6) also releases substantial quantities of nutrients (and organic matter). These efﬂuents lower water quality in the discharge areas and create algal blooms and low oxygen (Wolanski et al. 2000; Paez-Osuna 2001; Alonso-Rodriguez & Paez-Osuna 2003; Paez-Osuna et al. 2003).33 Maricultural practices can also release large amounts of unconsumed food particles, and fecal materials from the crop (Henderson et al. 1997). A good example of this is mussel culture, a widespread practice in many countries. Deep fjords, called rias locally, along the coast of Galicia in northern Spain, are particularly well suited for the cultivation of mussels, with rich supplies of nutrients brought by upwellings into the rias from colder deeper waters (Blanton et al. 1987). The nutrients support dense phytoplankton blooms that in turn furnish an abundant food supply for mussels. In the rias mussels are grown attached to ropes hanging below rafts that are densely spread out across the rias. Mussel culture is an economic mainstay for the region, with the crop sold widely in other European countries (Pérez Camacho et al. 1991). Mussel culture in the rias has had many ecological effects (Tenore et al. 1982; Kaiser et al. 32

Of the nitrogen added as feed to culture ponds for a variety of ﬁsh crops, only 11–29% may be recovered as ﬁsh product; 60–89% is released as nitrogen in efﬂuent from culture ponds (Hargreaves 1998). The released nitrogen, mainly as ammonium and organic nitrogen, then affects the estuary downstream. The efﬁciency with which ﬁsh ponds make use of added nitrogen is therefore low; this suggests that feeding regimes could be lowered with perhaps little loss of yield. 33 This is not the case for mariculture of seaweeds; these crops in fact improve water quality by taking up nutrients from the water in which they are grown. Similarly, to some extent shellﬁsh culture might be thought of as improving water quality, because the shellﬁsh feed by removing particles from suspension, and hence lower turbidity of the water.

HARVEST OF FINFISH AND SHELLFISH

1998). The mussels clear algae from the water, increasing transparency. Mussels do not assimilate all the particles they ingest, and fecal or unassimilated particles are expelled as feces or pseudofeces, and the rain of both kinds of particles (referred to as biodeposition) markedly alters the organic content of the sediments below the rafts. Where currents are strong, organic particulates are widely transported, and biodeposits are beneﬁcial, in that they may provide food for certain inshore ﬁshery stocks. Where currents are too weak to disperse the biodeposits, the accumulation of organic matter below the culture rafts lowers oxygen concentrations and signiﬁcantly alters the fauna below the rafts. Similar effects occur elsewhere and with other kinds of mariculture (Smaal 1991; Kaiser et al. 1998; Naylor et al. 2000). Depletion of wild stocks

Although maricultural products in certain instances may be intended to replace harvests that would have been sought in the wild, the reverse can be true, because of the need to feed the crop, and the need to obtain juvenile stock before culture. The increasing demand for small pelagic ﬁshes to make meal for the cultivation of salmon and shrimp has actually increased pressure on natural ﬁsh stocks. Cultivated salmon and shrimp may consume 2–5 times as much protein as the crop provides. Naylor et al. (2000) reported that total world ﬁshery landings may be 96 million tons per year. Of this catch, 65 Mt are eaten by people, and 30 Mt (plus 2 Mt of scraps) go to produce ﬁshmeal. Of the latter, about 10 Mt is used as feed in aquaculture, and that proportion is increasing. Therefore, rather than replacing the wild-caught harvest, mariculture consumes a signiﬁcant part of the world ﬁsh catch. Certain crops, such as shrimp culture, require stocking of the maricultural ponds with wildcaught larvae or juveniles (Pannier 1979; Twilley et al. 1996). In Ecuador, it was estimated that during 1983–1984 more than 90,000 artisanal ﬁshermen were involved in catching 16.5 billion juvenile shrimp annually from the near-shore waters. At least this many juveniles were needed to stock the 120,000 ha of shrimp ponds. The catch

275

included juveniles of different shrimp species, young ﬁsh, and other species, but only two species of shrimp, Penaeus vannamei and P. stylirostris, survive in the ponds. The ﬁshermen were paid in proportion to the two survivor species in their catch; the rest became bycatch lost to the coastal ecosystem. Since in the near-shore waters only about half the catch includes the Penaeus shrimp larvae, at least 33 billion shrimp were likely caught to satisfy the maricultural efforts. For some time, shrimp farming in Ecuador was billed as “an aquaculture success story” (Aiken 1990). Farmed shrimp during 1988 yielded more than ﬁve times as much shrimp as the ﬁshing effort; 20% of the economy derived from shrimp (87% from farmed shrimp). Shrimp was more economically important than bananas and cacao combined, and twice as important as coffee (Aiken 1990). The Ecuadorian shrimp effort depended on capture of wild juvenile shrimp. By the late 1980s, climatic patterns and overharvest were such that the supply of available larvae diminished, and many ponds could no longer be stocked. After the early 1990s, shrimp farming as an industry was severely diminished. The harvest of ﬁsh for ﬁshmeal production (usually smaller ﬁsh with fast reproductive rates) may also lead to lowering abundance of the larger top predators. In the North Sea, harvest of capelin, sand eel, and pout for ﬁshmeal production have been implicated in the reduction of stocks such as cod (Hislop 1996). Introduction of alien species and diseases

In many cases, mariculturalists have preferred to rear speciﬁc species, often not of local origin, as more promising or desirable crops, following a pattern common in agriculture. As noted in Chapter 10, one of the major vehicles for the introduction of exotic species along the coasts of the world has been the exchange of maricultural products. And, again much as in agriculture, the introduction of alien crops is closely followed by diseases of the crops, which can then spread to local species.34 34

Naylor et al. (2000) reviewed and provided references for cases of disease transmission linked to aquaculture.

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CHAPTER 11

Effects of fishing compared to other effects

35

It has been argued that there is some sort of multidecadal alternation between anchoveta and sardines, governed by large-scale cold vs. warm water regimes. It is hard to see much regularity in the record of Fig. 11.21, or in the reconstruction of Fig. 11.20. In addition, there are also records of sea bird abundance (Fig. 11.21) that seem to follow entirely different decadal and secular patterns; in many coastal environments, sea birds consume 5–29% of the ﬁsh production in the environment (Hunt et al. 1996), a rather substantial portion, which makes birds potent competitors with other ﬁsh and the ﬁsheries. To some degree, oceanographic conditions, sardines, anchovetas, and sea birds must be coupled as they change across decades. It is not evident just which of the numerous possible couplings are more or less important.

SARDINE

12 Biomass (× 106 metric tons)

This chapter has focused on environmental changes that are likely to be linked to ﬁshingrelated mechanisms. In all such cases, however, the changes forced by human harvests have taken place within a context of variation, at times quite substantial, in abundance, composition, and so on, driven by a multiplicity of inﬂuences not related to ﬁsheries. Such externally driven changes can be dramatic. For example, a detailed reconstruction of the abundance of sardine and anchovy abundances off the coast of California over the last 1,700 years shows a continued long-term pattern of drastic shifts in abundance (and possibly of climate-driven movement of the ﬁsh in or out of the area) (Fig. 11.20). The reconstruction was made on the basis of identiﬁable ﬁsh scales accumulated in stratiﬁed layers of sediment below the water column inﬂuenced by the California current. The two species did not seem to follow related patterns of abundance, and the prominent changes often spanned one or two orders of magnitude within relatively short periods of time. There are few better examples that demonstrate that substantial change may be the only constant in coastal populations. It is possible to think that changing global conditions—such as mean air temperatures and carbon dioxide concentrations (Fig. 11.21 upper panels)—may be somehow connected to the local shifts in abundance in sardines and anchovies/ anchoveta (Figs. 11.21 bottom).35 These external driving variables may or may not be linked to human effects; the long-term reconstruction of

16 8 4

0 200 400 600 800 1000 1200 1400 1600 1800 2000 4 3

ANCHOVY

2 1 0 200 400 600 800 1000 1200 1400 1600 1800 2000 Year

Figure 11.20 Reconstruction of the 1,700-year record of abundance of Paciﬁc sardines and northern anchovies off California. The data on abundance of ﬁsh scales in dated sediment layers were converted to biomass using equations derived for the purpose. Adapted from Baumgartner et al. (1992).

Fig. 11.20 suggests that drastic alterations in ﬁsh abundance took place long before people became a factor. It would be surprising, however, if ﬁshing pressure was not also involved in recent decades, since these ﬁsh species make up the world’s largest single ﬁshery in terms of yield (FAO 2002). The regulation of population dynamics of commercially interesting stocks by a combination of forces (external and global, local and ﬁshery) is likely to be the rule rather than the exception. Abundance of cod off Norway seems subject to oceanographic plus ﬁshery inﬂuences (Fromentin et al. 2000). In near-shores with large human densities, oceanographic/ﬁshery forcings are likely to be also intertwined with further inﬂuences of eutrophication (Rijnsdorp & van Leeuwen 1996; Thurow 1997; Caddy & Garibaldi 2000), a powerful agent of environmental change we discuss in the next chapter. These complex sets of interactions are further complicated by demographic factors. In the North Sea, nutrient-enriched waters are near-shore, where the youngest age classes of plaice are largely found; eutrophication hence affects the

277

HARVEST OF FINFISH AND SHELLFISH

GATA (°C)

2

GLOBAL AIR TEMPERATURE

1 0 −1 −2 1910

CO2 (ml)

2

1920

1930

1940

1950

1960

1970

1980

1990

2000

1940

1950

1960

1970

1980

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2000

CO2 CONCENTRATION

1 0 −1

Sea birds (× 106) ( ) Anchoveta (× 500,000 metric tons) ( )

30

1920

1930

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FAUNA

Anchoveta

Sardine 300

Sea birds 20

200

10 100

0 1910

1920

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1950 Year

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Figure 11.21 Anomalies (deviations from mean values) of global air temperatures (GATA; top) and carbon dioxide concentration (middle) (both with long-term increases removed) during the 20th century. Bottom: sea bird abundance, and anchoveta and sardine landings from Peru. Data from several sources, compiled by Chavez et al. (2003).

abundance of the early life stages of plaice. Fishing is most intense at some distance from the coast, and affects the larger size classes of plaice. The net balance of relative effects has not been worked out, but it seems that, in addition, large-scale regional climatic or oceanographic inﬂuences can in certain cases be as important as forcing by

nutrients and ﬁshing and can create sporadic stronger-than-average age classes (Rijnsdorp & van Leeuwen 1996). These complicated and nearly unpredictable interactions are probably widespread across regions and species, and may go a long way to explain why it has been so difﬁcult to understand changes in marine populations,

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and fully sort out the role of ﬁshing pressure from that of “background” variation. Worm and Myers (2003), to the contrary, argue that in fact “. . . there is at present no evidence that worldwide declines [in marine ﬁsh abundance] are linked in any major way to climate change. Large declines in marine ﬁsh communities have always coincided precisely with the onset of industrialized ﬁshing”. They suggest that overﬁshing may make these populations more susceptible to climatic change, but that ﬁshing pressure is the driving variable. They aver that we cannot use global climatic changes as an excuse to exonerate the ﬁshing industry. Whatever turns out to be the case, whether climate change has been involved in the recent alterations of stocks or not, what is certain is that overﬁshing has had major impacts on coastal marine environments, and that a ﬁrst step in remediation of what is now clearly a crisis situation is to develop strategies to conservatively manage harvests.

References Aiken, D. 1990. Shrimp farming in Ecuador: An aquaculture success story. World Aquacult. 21:7–16. Alonso-Rodriguez, R., and F. Paez-Osuna. 2003. Nutrients, phytoplankton and harmful algal blooms in shrimp ponds: A review with special reference to the situation in the Gulf of California. Aquaculture 219:317–336. Alverson, D. L., M. H. Freeberg, S. A. Murawski, and J. G. Pope. 1994. A Global Assessment of Fisheries Bycatch and Discards. FAO Fisheries Techical Paper No. 339. Food and Agriculture Organization of the United Nations, Rome, 233 pp. Backus, R. H., and D. W. Bourne (eds) 1991. Georges Bank. MIT Press, Cambridge, MA, 593 pp. Badalamenti, F., G. D’Anna, J. K. Pinnegar, and N. V. C. Polunin. 2002. Size-related trophodynamic changes in three target ﬁsh species recovering from intensive trawling. Mar. Biol. 141:561–570. Ball, B. J., G. Fox, and B. W. Munday. 2000. Long- and shortterm consequences of a Nephrops trawl ﬁshery on the benthos and environment of the Irish Sea. ICES J. Mar. Sci. 57:1315–1320. Baum, J. K., and 5 others. 2003. Collapses and conservation of shark populations in the Northwest Atlantic. Science 299:389–392. Baumgartner, T. R., A. Soutar, and V. Ferreira-Batrina. 1992. Reconstruction of the Paciﬁc sardine and northern

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Tanner, J. E. 2003. The inﬂuence of prawn trawling on sessile benthic assemblages in Gulf St. Vincent, South Australia. Can. J. Fish. Aquat. Sci. 60:517–526. Tasker, M. L., and 5 others. 2000. The impact of ﬁshing on marine birds. ICES J. Mar. Sci. 57:531–547. Tegner, M. J., and P. K. Dayton. 2000. Ecosystem effects of ﬁshing in kelp forest communities. ICES J. Mar. Sci. 57:579–589. Tenore, K. R., J. Corral, and N. Gonzalez. 1982. Food chain patterns associated with intense mussel culture in the Ria de Arosa (NW Spain). Atlantica Rio Grande 5:118. Thurow, F. 1997. Estimation of total ﬁsh biomass in the Baltic Sea during the 20th century. ICES J. Mar. Sci. 54:444–461. Tuya, F. C., M. L. Soboil, and J. Kido. 2000. An assessment of the effectiveness of marine protected areas in the San Juan Islands, Washington, USA. ICES J. Mar. Sci. 57:1218– 1226. Twilley, R., S. Snedaker, A. Yañez, and E. Medina. 1996. Biodiversity and ecosystem processes in tropical estuaries: Perspectives of mangrove systems. Pp. 327–370 in Mooney, H. (ed.). Functional Roles of Biodiversity: A Global Perspective. J. Wiley, New York. Valiela, I. 1984. Marine Ecological Processes. SpringerVerlag, New York, 546 pp. Valiela, I. 1995. Marine Ecological Processes, 2nd edn. Springer-Verlag, New York, 686 pp. Valiela, I., and 5 others. 2000. Operationalizing sustainability: Management and risk assessment of landderived nitrogen loads to estuaries. Ecol. Appl. 10: 1006–1023. Velarde, E., E. Ezcurra, M. A. Cisneros-Mata, and M. F. Lavin. 2004. Sea bird ecology, El Niño anomalies, and prediction of sardine ﬁsheries in the Gulf of California. Ecol. Appl. 14:607–615. Verity, P. G., and V. Smetacek 1996. Organism life cycles, predation, and the structure of marine pelagic ecosystems. Mar. Ecol. Prog. Ser. 130:277–293. Wantiez, L., P. Thollot, and M. Kulbicki. 1997. Effects of marine reserves on coral reef ﬁsh communities from ﬁve islands in New Caledonia. Coral Reefs 16:215–224. Watson, M., and R. F. G. Ormond. 1994. Effect of an artisanal ﬁshery on the ﬁsh and urchin populations of a Kenyan coral reef. Mar. Ecol. Prog. Ser. 109:115–129. Wolanski, E., and 6 others. 2000. Modelling and visualizing the fate of shrimp pond efﬂuent in a mangrove-fringed tidal creek. Estuar. Coast. Shelf Sci. 50:85–97. Worm, B., and R. A. Myers. 2003. Meta-analysis of cod– shrimp interactions reveals top-down control in oceanic food webs. Ecology 84:162–173. Worm, B., and R. A. Myers. 2004. Managing ﬁsheries in a changing climate. Nature 429:15. Zwanenburg, K. C. T. 2000. The effects of ﬁshing on demersal ﬁsh communities of the Scotian shelf. ICES J. Mar. Sci. 57:503–509.

Chapter 12 Eutrophication

“Green” tide blooms in the shores of Waquoit Bay, Massachusetts. These nutrient-fed canopies of unattached drift macroalgae normally lie on the bottom, but, as in this occasion in summer 2003, sudden increases in sunlight increase the production of oxygen bubbles within the canopy, which alters the buoyancy of fronds, and the algae ﬂoat off and become stranded on shore.

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A case history: changing watersheds and estuaries of Cape Cod Cape Cod is a small peninsula on the extreme east of the state of Massachusetts. Its landscape and coastal environments are marked by centuries of increasing human pressure. In pre-Columbian times, Cape Cod was populated by Wampanoags (“the people of the east” or “of the dawn”) of the Algonquin complex of Amerindian nations. The Wampanoags witnessed the arrival of religious exiles from England and Holland in the early 1630s, and eventually became the victims of European diseases and land-taking. The Wampanoags lived in small dispersed communities, cultivated corn, squash, and beans, made intensive use of ﬁsh and shellﬁsh from the numerous ponds and estuaries in Cape Cod, and hunted a bit in the forests of the area. The early European colonists had minor interest in forests,1 and by the mid-1800s the previously forested glacial terrain had been largely cleared by settlers for ﬁrewood, and for ship and house construction. The landscape “. . . was for the most part, bare or with only a little scrubby wood left on the hills.”2 Soils were too infertile to support much agriculture so, instead, sheep grazed on pastures across the rolling hills, and swamps were converted to cranberry bogs. Cape Codders depended in large measure on the sea for food and meager incomes. Saltworks were active, drying and selling the salt that remained after the evaporation of sea water. Shellﬁsh, alewife, eels, and other ﬁsh were harvested from the many ponds and shallow estuaries in

1

Although they did harvest sassafras, a tree whose powdered roots prompted perspiration and was used by 17th century doctors as a remedy for many diseases, including plague, pox, and the “French disease” (Vuilleumier 1970). Sassafras was one of the few goods exported back to Europe in ships that came laden with manufactured goods. 2 In his book, Cape Cod, Henry David Thoreau described travel through the sand pathways of Cape Cod during 1849–1855. He found a mostly treeless peninsula jutting daringly into a daunting sea, full of difﬁcult challenges and dramatic contrasts, and was impressed by the natural settings. “A man might stand there and put all America behind him” says the last sentence of Cape Cod [a reprint of the original 1864 edition is available (Thoreau 1984) ].

the area. The many racks of cod ﬁllets hung out to dry were evidence of the links to the sea in settlements such as Provincetown. Other marine resources were opportunistically used. Thoreau noted how the locals used small boats to drive pods of pilot whales onto beaches, and how selfstranded pilot whales were eagerly butchered. Because the “shores are more fertile than the dry land”, Cape Codders were forced to use the sea, which often involved hardship and tragedy. Thoreau tells us “that it would not do to speak of shipwrecks there, for almost every family has lost some of its members at sea.” Nevertheless, “the surviving inhabitants went a-ﬁshing again . . . as usual”. Salvaging goods and timbers from the too-frequent shipwrecks was also a not-inconsequential activity; in fact, the hardy, weatherbeaten Cape Codders in Thoreau’s time often dated events relative to some memorable shipwreck. Sheep could not support the increasingly numerous Cape Cod communities, and pastures eventually became fallow. By the 1930s, growth of pitch pine and scrub oak had reforested the landscape of Cape Cod. Some small land parcels were still worked by local farmers who grew mainly potato and strawberry crops, as could be seen in the area surrounding Waquoit Bay, one of the many estuaries found on Cape Cod (Fig. 12.1 top). Cape Cod farms, though artisanal and small in area, provided the bulk of the strawberry crop in New England, and somehow set records for its harvest. By mid-century, however, the local agricultural industry declined, unable to compete with national and international industrialscale competition, and farmed parcels largely disappeared. Then, everything changed. Population increases in the Boston–New York megalopolis, improved access through new roads, and economic upsurges through the 1960s and 1970s led to a boom of construction in Cape Cod. Many houses were built for residents as well as the much larger population of visitors that were attracted to the area during the summer (Fig. 12.1 bottom). The ﬁshing and farming economies of communities that characterized the Cape Cod area for centuries became transformed into communities whose dominant sectors were tourism and construction.

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Figure 12.1 Aerial photos of the Waquoit Bay area. Top: photo taken during 1938, showing mainly forested land cover, many parcels devoted to small-scale agriculture, and small rural communities surrounding Waquoit Bay. Bottom: photo taken in 1993, showing the intense development of urbanized areas, minimal number of agricultural parcels, and reduced area covered by forests.

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4,200 3,900 Land use (ha)

Natural vegetation 3,600

1,500 1,200 Human uses 900 600 1920

1940

1960

1980

Land use (ha)

800

2000 Turf

600

Impervious surfaces

400 200 Agriculture 0 1920

1940

1960 Year

1980

2000

Figure 12.2 Time course of land uses on the watershed of Waquoit Bay, 1938–1990. Top: human land use and natural vegetation. Bottom: changes in cover by turf, impervious surfaces, and agriculture. From Bowen and Valiela (2001).

The residential urban sprawl that took over Cape Cod during the latter half of the 20th century is evident in the time course of land uses on the watershed of Waquoit Bay (Fig. 12.2). Between 1938 and 1990, the land area under natural vegetation decreased from 84 to 68%, owing to conversion to land covers associated with urbanized areas. Loss of forest area in Cape Cod has continued. The Cape Cod that so charmed Thoreau is harder to ﬁnd. The history of land on Cape Cod since the 17th century is—much like that of coastal areas anywhere in the world—a succession away from forests and toward intensely used land covers. In

late 20th century Cape Cod, the changes in land use are dominated by urban sprawl. This has led to pervasive many-fold environmental, economic, social, and aesthetic consequences for the people, land, and adjoining estuaries. One of the major consequences of urban sprawl in Cape Cod is the increase in export of nitrogen from watersheds to estuaries.3 Nitrogen from different sources entered the watershed of Waquoit Bay toward the end of the 20th century. The largest source of nitrogen to the watershed of Waquoit Bay was atmospheric deposition (Table 12.1), but that source did not increase (Fig. 12.3 top). Inputs of waste water and fertilizer nitrogen, although smaller, did increase during the last century (Fig. 12.3 top). More people discharged more waste water (by septic systems in the Waquoit watershed), and applied more fertilizer to larger areas of golf courses and lawns. The watersheds of Cape Cod estuaries retained a surprising portion of the nitrogen they received. Most (81%) of the nitrogen entering the watershed of Waquoit Bay was intercepted before reaching the bay itself (Table 12.1). The interception of external nitrogen within the watershed of Waquoit has therefore “subsidized” water quality in the bay by lowering nitrogen loads.4 In spite of the remarkable degree of nitrogen retention within the watershed, nitrogen load reaching Waquoit Bay nonetheless increased more than two-fold during the last half of the 20th century (Fig. 12.3 bottom). These increases largely

3

Cape Cod is basically a pile of sand, underlain by unconsolidated materials that were left behind by glaciers (Strahler 1966; Oldale 1992). The soils and sediments are coarse-grained enough that there is little surface runoff, and virtually all the precipitation percolates into the sediment and is transported by groundwater ﬂow toward the receiving ponds and estuaries. The few small streams on Cape Cod are springs of groundwater rather than rivers whose ﬂow is supported by runoff. The through-put of nitrogen from watershed to estuary therefore is almost exclusively via groundwater ﬂow. 4 The interception of nitrogen within the watershed also changed the relative inﬂuence of the different sources of nitrogen. Losses of atmospheric nitrogen within the watershed were larger than losses of wastewater nitrogen (Table 12.1); these differences in interception shifted the relative contribution of nitrogen by waste water, fertilizers, and atmospheric deposition to the bay itself. Waste water added 53% of the nitrogen input to Waquoit Bay, while fertilizers and atmospheric deposition became smaller contributors (Table 12.1).

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Table 12.1 Relative contribution by the major sources of nitrogen (N) to Waquoit Bay, within-watershed nitrogen losses, and nitrogen exports to the bay for each of the three major sources. Data from Valiela et al. (1997). Source of N Atmospheric deposition Fertilizer use Wastewater disposal Total

N inputs (kg × 103 N yr−1)

200

% of N input lost within the watershed

% of N load exported to the bay

56 16 28

90 67 78

30 17 53

100

81

100

TO WATERSHED Total

160 120

Atmospheric 80 Waste water

40

Fertilizer 0 1930 1940 1950 1960 1970 1980 1990 2000 28

N inputs (kg × 103 N yr−1)

% of N load to the watershed

24

TO ESTUARY Total

20 16 12 8 4

Waste water Atmospheric Fertilizer

0 1930 1940 1950 1960 1970 1980 1990 2000 Year

Figure 12.3 Calculated changes in land-derived nitrogen (N) loads by waste water, atmospheric deposition, and fertilizers to the watershed of Waquoit Bay (top), and to the bay itself (bottom), 1938–1990. From Bowen and Valiela (2001).

mirrored the increases in wastewater inputs (Fig. 12.3 bottom). The increased nitrogen loads to Waquoit Bay toward the end of the 20th century were therefore a result of increases in waste water

and fertilizer nitrogen, and of reduction of area covered by natural vegetation.5 To assess the possible effects of the increasing urbanization of watersheds, we compared presentday conditions within certain smaller subestuaries within the Waquoit Bay estuarine system whose watersheds differed in land use. Sage Lot Pond is a subestuary with an entirely forested watershed, and represents a relatively pristine condition; Quashnet River has a mixed-use watershed that serves to represent an intermediate step in the forest-to-urban transition; Childs River has a relatively urbanized watershed. The different degrees of urbanization of the watersheds of these subestuaries is made evident by the time courses of their densities of building across the second half of the 20th century (Fig. 12.4). The differences in land use on the different subwatersheds created different nitrogen loads to the subestuaries, and different concentrations of nitrate6 in fresh water about to ﬂow into the 5

The relative effectiveness of Cape Cod forests and other natural vegetation in intercepting atmospheric nitrogen derives in part from the fact that these are actively growing forests, still recovering from the deforestation prevalent in the 19th century. The forests are therefore storing a substantial part of the atmospherically derived nitrogen deposited by precipitation (Valiela et al. 1997). It is of concern that the atmospheric loads to Waquoit Bay and its watershed are now poised at rates that are just below those that saturate the ability of forests to retain or intercept atmospheric nitrogen. Further increases in atmospheric nitrogen loading may lead to large increases in through-put of atmospheric nitrogen to estuaries and bays (Bowen & Valiela 2001). 6 The fresh water also carried other nutrients, including ammonium, organic nitrogen, phosphate, silica, and so on. None of these others were transported in proportion to the degree of urbanization on the watershed, as was nitrate.

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5

80

CHILDS RIVER

Childs River 60 NO3 (µM)

Houses per ha

4 3 2

40

20 1 0

Quashnet River Sage Lot Pond

0 0

different estuaries. The greater the degree of urbanization, the larger the concentration of nitrate in groundwater discharging into the upper reaches of the estuaries (Fig. 12.5). The concentrations of nitrate within the estuaries diminished—converted to nitrogen gas, buried, or simply diluted—as water moved down-estuary (Fig. 12.5).7 In general, concentrations of nitrate were greater in those subestuaries whose watersheds were most affected by urban sprawl, and the increases in nitrogen loads led to manifold environmental changes. The changes resulted because nitrogen is the element primarily responsible for controlling the growth of algae and plants in the estuaries of Waquoit Bay.8 Single-celled algae and large algal seaweeds increased9 in response to larger nitrogen supply 7

The down-estuary losses were greater than would be predicted by simple dilution. Note that the points measured in the estuaries lie below the simple mixing line (dashed line): this implies that processes beyond dilution were active. 8 Details about nutrient limitation in Waquoit Bay can be found in Tomasky et al. (1999), Peckol et al. (1994), and Hauxwell et al. (1998). 9 Figure 12.6 oversimpliﬁes by showing responses of total phytoplankton and macroalgae. Different kinds of phytoplankton and macroalgae responded differently to increased nitrogen supply in Waquoit Bay. In the case of plankton, species of an intermediate size (nanoplankton) responded to enrichment, but not other species

20

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12

16

20

24

28

32

16

20

24

28

32

12 16 20 Salinity (‰)

24

28

32

QUASHNET RIVER

0 0

NO3 (µM)

Figure 12.4 Time course of building density, 1938–1984, in the three subwatersheds of Waquoit Bay used in the space-for-time comparisons. From Valiela et al. (1992).

NO3 (µM)

1940 1950 1960 1970 1980 1990 Year

20

4

8

12

SAGE LOT POND

0 0

4

8

Figure 12.5 Concentrations of nitrate (mean ± SE) in fresh groundwater about to enter (gray circles; gray lines show the standard error of these values), and in the water of, the three different subestuaries of the Waquoit Bay estuarine system (black circles). The nitrate (NO3) data are plotted against the salinity of the water to depict the pattern of nitrate concentrations down each estuary (salinity is a proxy for distance in this case). In addition, the dashed lines show where the nitrate concentrations would have been if the down-estuary pattern were merely the result of dilution of the fresh water with nitratepoorer sea water. The points showing nitrate contents of the estuary water lie below the dashed lines: this indicates that substantial losses (by conversion to nitrogen gas, or burial in sediments) took place as the land-derived nitrate moved downstream. Adapted from Valiela et al. (2000a). of smaller or larger sizes (G. Tomasky, unpublished data). The macroalgae were dominated by a green and a red species, both of which responded to enrichment, but several other species showed no response. The effects of eutrophication in Waquoit Bay were therefore quite species-speciﬁc.

3.0

2.2 6 1.4

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2 2

4

6

8

10

12

14

Biomass (g dry wt m−2) ( )

240

2.4 MACROALGAE

200

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120

1.2

80

0.8 0.4

40 0

2

4

6

8

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Biomass (g dry wt m−2) ( )

50

Figure 12.6 Biomass and production (mean ± SE) of phytoplankton, macroalgae, and eelgrass, plotted against mean ambient concentrations of dissolved inorganic nitrogen (DIN) (mainly nitrate plus some ammonium) in the three subestuaries of the Waquoit Bay estuarine system exposed to different rates of land-derived nitrogen loads. From Valiela et al. (2000a).

0.6 EELGRASS

40 0.4 30 0.2

20 10

0

0 −0.2 0

(Fig. 12.6 top and middle). The large crop of nitrogen-fed macroalgae became a dominant feature of Waquoit Bay. Macroalgal canopies in places reached a depth of 75 cm, and sometimes drifted on shore, prompting news headlines (“Algae clogs [sic] Cape beaches”, Cape Cod Times, June 28, 2003; “Massive algae [sic] bloom mars Waquoit Bay; nitrogen blamed”, Falmouth Enterprise, July 1, 2003) (see Chapter 12 frontispiece).

2

4 6 8 10 Mean annual DIN in estuary (µM)

12

Production (kg biomass m−2 yr−1) ( )

0

14

Production (kg biomass m−2 yr−1) ( )

Biomass (mg chlorophyll a m−3) ( )

PHYTOPLANKTON 10

Production (kg C m−2 yr−1) ( )

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The growth of macroalgae had some beneﬁcial effects, including sequestering nutrients, and furnishing more and better food particles. Substantial amounts of nitrogen were stored in macroalgal biomass; this helped maintain water quality, because the nitrogen held within the fronds would otherwise be available to foster phytoplankton blooms, which would make the water green and turbid. Another beneﬁt was that the supply of

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Shell growth (mm wk−1)

3.0 Oysters 2.4 Soft-shell clams

1.8

Quahog clams

1.2 0.6

Mussels 0.0 10

100 N load (kg N ha−1 yr −1)

1,000

Figure 12.7 Growth rates of different shellﬁsh species in relation to the nitrogen (N) loading rates to Cape Cod estuaries. From Carmichael et al. (submitted).

particles for suspension-feeders such as bivalves improved in more eutrophic estuaries, and increased the growth of shellﬁsh (quahog clams, soft-shell clams, ribbed mussels) (Fig. 12.7). Similarly, the growth of horseshoe crabs in Cape Cod estuaries increased as nitrogen loads increased (Carmichael et al. 2004), up to a limit: horseshoe crabs avoided feeding within estuaries subject to the highest loads. The macroalgal canopy, however, also had what could be considered detrimental effects. Macroalgae (Fig. 12.6 middle) overgrew and virtually eliminated eelgrass meadows (Fig. 12.6 bottom, Fig. 12.8). Eelgrass meadows are an essential habitat for scallops, a commercially important species in Cape Cod estuaries.10 The conversion of eelgrass meadows to macroalgal canopies was accompanied by a two-order of magnitude fall in the catch of scallops from Waquoit Bay (Valiela et al. 1992). There was also greater oxygen consumption. Macroalgae11 consumed enough oxygen during most nights that at dawn each day there was a low oxygen 10

Many other species also decreased as a result of the shift from seagrass meadows to macroalgal canopies. For example, ﬁsh species that depend on eelgrass habitats decrease in abundance in nitrogen-loaded estuaries (Hughes et al. 2002). Many herbivores became less abundant in eutrophic estuaries (Hauxwell et al. 1998). 11 Of course, microbes also consumed oxygen during their decomposition of the abundant detrital organic matter that accumulated on the bottom, and the fauna also respired. In Waquoit, macroalgae and microbes were the principal consumers of oxygen.

(hypoxic) condition. During the day, oxygen released as a product of photosynthesis by the macroalgal canopy restored the high oxygen conditions. If there were at least 3 days of cloudy weather, the water column oxygen was not restored sufﬁciently, and near-bottom water became anoxic. Such episodes of hypoxia and anoxia led to kills of ﬁsh, shellﬁsh, and other species, which became more frequent as landderived nitrogen loads increased (D’Avanzo & Kremer 1994). The increased rate of macroalgal growth, and lower abundance of bottom-dwelling animals in estuaries (including grazers of macroalgae), reshufﬂed key ecological control processes (Hauxwell et al. 1998). In subestuaries of the Waquoit system where land-derived nutrient loads were low, grazing rates roughly matched rates at which macroalgae grew, so that points fell along the 1 : 1 line (Fig. 12.9). In these conditions, macroalgal canopies did not proliferate. Where nitrogen loads became larger, however, the balance shifted: rates of macroalgal growth increased, but grazing rates became smaller. The result was a macroalgal bloom. In effect, the inﬂuence of increased nitrogen supply at some point overwhelmed grazing controls on macroalgal biomass. Comparisons from estuaries subject to different loads provided compelling, but inferential, evidence that increased nitrogen loads were the agents of change that led to the restructuring of function and composition of species of Waquoit food webs. Use of stable isotopic ratios made it possible to unambiguously demonstrate that it was the nitrogen from land that forced the notable environmental changes seen in Waquoit Bay. The different mosaic of land uses on the different sub-watersheds of Waquoit Bay imposed different stable nitrogen ratios on the nitrogen transported from land to the subestuaries. Isotopic ratios of nitrate emerging from each subwatershed into the receiving subestuaries increased where there were more people (Fig. 12.10 top left) or more wastewater disposal (Fig. 12.10 top right) on or from the watersheds. The degree of urbanization therefore left an imprint on nitrogen entering the estuaries, and this

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Figure 12.8 Sketch maps of the distribution of eelgrass meadows in Waquoit Bay, 1951–1992. From various sources cited in Valiela et al. (2000b).

Grazing rate (g m−2 day−1)

3

1:1 14 kg N ha−1 yr−1

1951

Childs River Quashnet River Sage Lot Pond

2 350 kg N ha−1 yr −1

601 kg N ha−1 yr −1

1

0 0

2

4 6 Growth rate (g m−2 day−1)

8

10

Figure 12.9 Estimates of grazing rates by amphipods and isopods feeding on macroalgae, plotted in relation to the growth rates of macroalgae in the three subestuaries of Waquoit Bay subject to different nitrogen (N) loads (numbers above the envelopes). The dashed line shows the line indicating where grazing and macroalgal growth rates would be equal. Adapted from Hauxwell et al. (1998).

imprint was detectable in the water in the estuaries, and in the estuarine organisms. For example, the higher the isotopic ratio in groundwater nitrate entering the estuaries, the higher the isotopic signature of the producers found within speciﬁc estuaries (Fig. 12.10 middle). Higher in the food web, isotopic ratios of consumers also paralleled signatures of food items within each estuary (Fig. 12.10 bottom). The isotopic ratios therefore unambiguously demonstrated the close coupling between land

1971

1978

1987

1992

and estuary, mediated through the export of nitrogen: different combinations of land uses on speciﬁc watersheds created recognizably different isotopic ratios in the nitrogen exported to Waquoit Bay estuaries and their food webs. The isotopic data made evident that it was the very atoms of nitrogen that left the watershed that coursed through the food web of the estuaries. The isotopic ratios conﬁrmed that changes in watershed land use powerfully affect the receiving estuaries, and that the mechanisms involve nitrogen transport from land to sea. The path of land-derived nitrogen inputs can therefore be traced, step by step, as the nitrogen moves from watershed to and into the food webs of receiving estuaries. Increased urbanization on watersheds increases the amount of nitrogen, the element that happens to often limit producer growth in coastal ecosystems, delivered to the estuaries. Nitrogen derived from urban sprawl on watersheds enters the estuarine food webs, and forces a series of substantial environmental alterations and species-speciﬁc population changes in the receiving estuaries. The recent drastic changes in land use on Cape Cod therefore have had powerful repercussions in the way the receiving estuaries function, and on the structure of estuarine food webs. What centuries of artisanal harvest of shell and ﬁnﬁsh, deforestation, sheep grazing, and strawberry and potato farming could not do, was achieved in a few decades by urban sprawl, and its increased release of nitrogen. Recognition of the pressures that are being placed on the estuarine systems by urban sprawl has prompted concern from people living on Cape

δ15N (‰) of NO3 in groundwater

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CHAPTER 12

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12

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8

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4

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0

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−4

0 4 8 12 Population density (individuals ha−1) on watershed Plants Algae Seston

16 δ15N (‰) in producers

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12 8 4 0

δ15N (‰) in bivalve tissue

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8

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Mussels Quahogs Scallops Soft-shell clams

12 9 6 3 0

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4 6 δ15N (‰) in seston

Cod and their decision-makers. There is worry about preservation of natural resources such as shell- and ﬁnﬁsheries, and about the potential damage to the tourist industry that now supports the economy of Cape Cod. Visitors come to Cape Cod to enjoy the natural setting that so impressed Thoreau, and these environments are now being severely altered by urban sprawl. We can still share Thoreau’s endless view of the Atlantic, only now we might have to peer around houses, vehicles, and parking areas. Options to lower nitrogen loads are daunting, since the population of Cape Cod will certainly increase in the future, and any measure will be economically and politically difﬁcult. Improvements in waste and fertilizer management, preser-

Figure 12.10 Top panels: stable isotopic signature of nitrate nitrogen (N) in relation to density of people on the watershed (left) or wastewater load from the watershed (right) (from Cole et al. 2005). Middle: relationship between the stable isotopic signature of nitrate nitrogen and that of producers found in the estuaries (from Carmichael et al. submitted; Cole et al. 2005). Bottom: relationship of stable isotopic signatures in shellﬁsh tissue to algal and other organic particles suspended in the water (labelled as “seston”) (adapted from Carmichael et al. submitted).

vation of forests, and conservation of nitrogenretaining wetlands, all will have to be among the approaches needed (Bowen et al. 2004). Even if Cape Codders ﬁnd the political and economic will to put such options in place, recovery will take some time. Travel time of the nitrogen entrained in the groundwater below the ground may take anywhere from 1 to 60+ years, so there will necessarily be long-term lags even under the most draconian remediation measures.

Definition and causes of eutrophication Eutrophication is an amorphous term that means many things to many people, and has been applied

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Hessen (1999) and the National Research Council (2001) review much material relevant to land-derived nutrient loads to aquatic systems.

1,600 1,400 1,200 1,000 800 600 400 200 0 0

20 40 60 80 100 120 140 160 180 200 Population density (individuals km−2)

Figure 12.11 Relationship between population density on the watersheds and export of nitrogen (N) in rivers to estuaries, for temperate regions of the North Atlantic Ocean. From Howarth et al. (2002). 8

12 8 f

N export (kg × 103 N km−2 yr −1)

to describe a complex of symptoms. The term derives from “pertaining or aiding nutrition”. Eutrophication is an increase in supply or production of organic matter in an environment (Nixon 1995), which forces a large array of ecological and biogeochemical consequences (Cloern 2001). The term eutrophication was ﬁrst used to refer to waters exposed to increased inputs of organic matter, such as might occur from sewage outfalls or disposal of urban, agricultural, or industrial sludges (National Research Council 2000; Gray et al. 2002). In the Baltic Sea, annual human discharge of organic matter during the 1980s was more than 50-fold larger than the internal photosynthetic production of organic matter (Elmgren 1989). The organic enrichment lowered oxygen and few organisms survived. Disposal of urban sewage sludge still takes place in certain coastal sites. In general, however, external organic loads to estuaries are smaller than load of organic matter produced within the estuaries. In the Neuse River of North Carolina, for example, external loads of organic matter from land amounted to only 22%, on average, of internal loads (Paerl et al. 1998). On an annual basis, the nitrogen-forced internal production within the Neuse contributed more to consumption of oxygen than the external organic load. Except in urban areas where large ﬂows of sewage are discharged, eutrophication is the result of increases in supply of limiting nutrients. This then increases the production of organic matter, which in turn leads to lower oxygen concentration. Thus, the most common cause of eutrophication of coastal environments is the addition of nutrients, largely derived from land.12 The actual amount of nutrients exported by fresh water and by air masses to coastal waters is, not surprisingly, related to the number of people on the watershed (Fig. 12.11). This is in spite of the notable interception of nutrients within watersheds (Fig. 12.12), already noted in the Waquoit case history. On

Total N export from landscape (kg km−2 yr−1)

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Figure 12.12 Comparisons of the nitrogen (N) inputs to some eastern North American watersheds, and exports from these same watersheds. The dashed line shows where the points would lie if inputs equaled exports. Inset shows the frequency (f ) distribution of the retention values within the watersheds. Data from Howarth et al. 1996; Boyer et al. 2002.

average, about 75% of the entering nitrogen was retained within the watersheds of eastern North America (Fig. 12.12 inset), and more than 75% of the nitrogen delivered to terrestrial environments

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around Europe was retained and not exported to aquatic systems (Hessen 1999). Watersheds therefore subsidize water quality in receiving water bodies by the retention of large portions of nitrogen inputs. Nutrient retention within watersheds, however, can be altered by human inﬂuence. We can increase external inputs into watersheds, so as to saturate natural processes of uptake (Aber et al. 1989), or we can change land use toward land covers that retain less of the nitrogen, as we saw in the Waquoit case history.13 Where human activities become dominant, within-watershed interception diminishes, and greater ﬂuxes of nitrogen leave the watersheds. Such land-derived nitrogen exports result in signiﬁcant and increasing eutrophication of receiving coastal waters worldwide. The remainder of this chapter focuses on sources of nutrient enrichment, how widespread enrichment is worldwide, and its ecological and biogeochemical consequences. Then we turn to recovery from eutrophication, and mention options for the restoration of coastal environments.

Sources of nutrients to coastal ecosystems It could be argued that virtually all nitrogen entering coastal environments is of atmospheric origin: N2 gas is ﬁxed industrially from the atmosphere to make fertilizer, the fertilizer supports growth of corn fed to livestock that are then eaten by people. The nitrogen in human wastes— originally atmospheric nitrogen—is then released to coastal waters through a sewage outfall. That may be true if we consider the fate of nitrogen on a global scale. To understand and manage eutrophication it seems more practical to consider a more local watershed/estuary ecosystem scale. 13

Other alternatives have been suggested, including physical removal of algae, and establishing beds of bivalves that will consume particles from the water column. Many such suggestions are interesting but on examination have limited application, owing to uncertain technical success, or because, as in the case of the bivalves, it would require literally covering the bottom with shellﬁsh to achieve measureable improvements. We have already seen that such maricultures create their own environmental problems. Many of the available management options are assessed in Bowen and Valiela (2004).

This smaller scale involves deﬁning the more limited geographic boundaries of catchment area, and the area of receiving waters that are coupled to the watershed. For these more local units of land- and seascape, we can consider that the major sources of nitrogen are atmospheric deposition, fertilizer use, and wastewater disposal.14 Atmospheric deposition Nitrogen delivered to coastal watershed and estuary surfaces mainly derives from industrial, agricultural, and vehicular sources up-wind. “Airshed” areas are usually much larger than watershed areas. Nitrogen released by combustion processes in the industrial midwest of the USA may, for example, be transported hundreds and perhaps thousands of kilometers away to the eastern coastal states. Ammonium volatilized from livestock feed lots, and oxidized nitrogen released by vehicular exhaust in urban centers, are other major sources of atmospheric nitrogen compounds. The history of atmospheric deposition varies geographically, because deposition depends on the degree of industrialization and urbanization of the airshed. In some regions, atmospheric deposition may be increasing, but elsewhere the time trend is less clear. In the eastern USA, for instance, nitrate deposition increased about fourfold between 1925 and the 1970s, while ammonium deposition decreased about three-fold. Deposition of both forms of nitrogen has been relatively uniform since the passage of clean air legislation in the USA in 1973, at the decadal scale, although it remains variable on an annual basis (Bowen & Valiela 2001; Paerl et al. 2002).

14

In studies of nutrient inputs one often sees input terms for livestock as sources. It is hard to deal with this term because it very much depends on how the boundaries of the watershed are deﬁned. Livestock manure cannot be considered as an input of new nitrogen unless we are dealing with feedlot livestock whose rations come from outside the watershed, for instance as meal from ﬁsh caught offshore. In general, pasture-fed livestock raised in the watershed are merely an intermediate step in the path of atmospheric nitrogen to the receiving coastal water, but pasture-fed livestock raised outside the watershed but consumed as meat in the watershed should be counted as a separate external contribution to the wastewater stream.

EUTROPHICATION

Evidence for the limiting role of nitrogen in coastal systems Rates of growth of many coastal producers are controlled largely by the availability of nitrogen, according to results of small-scale enrichment experiments, inadvertent quasiexperimental enrichment at whole-estuary scales, and studies of nutrient concentrations and ratios (Howarth 1988; Nixon 1992; Elmgren 2001). For the sake of brevity, therefore, the emphasis here is on inputs of nitrogen to coastal systems.15 Small-scale enrichment experiments

Experimental additions of nutrients to containers with sea water and coastal producers have repeatedly shown that the crop or production of phytoplankton and macroalgae increases in response to nitrogen additions, and seldom in response to additions of phosphate (Howarth 1988; Valiela 1995). Analyses of results from 17 such experimental enrichments done in temperate latitude waters showed that in the majority of marine waters there was an enhanced response by phytoplankton when nitrogen was added (Fig. 12.13 top left). The response to phosphate additions was far less notable (Fig. 12.13 top right). Summaries such as this indicate that the growth of these phytoplankton was limited by the supply of nitrogen, rather than of phosphorus. Nitrogen limitation was found even in waters that were more subject to nitrogen contamination. In pristine waters it seems factors other than nitrogen limited growth, perhaps availability of phosphorus or of trace metals such as iron. Nitrogen limitation of phytoplankton growth was found in all coastal waters (Fig. 12.13 bottom). Enrichment at whole-estuary scales

Increased production took place where nitrogen loads to entire estuaries increased. Nixon 15

Other nutrients that might be of interest include phosphorus and silica. Phosphorus may be the limiting nutrient for the growth of producers in coastal waters underlain by carbonate sediments, such as many coral reefs, and in certain other coastal environments (Lapointe & Hanisak 1997; Lapointe & Barile 2001) and in brackish waters where nitrogen-ﬁxing blue-green bacteria may proliferate, such as the Baltic Sea (Elmgren 2001).

295

(1995) reviewed several such cases; in the case of Waquoit Bay, groundwater ﬂow to different estuaries provided a range of nitrate loads, but similar phosphate loads, and the response to increased nitrate was signiﬁcant. Another similar example of such whole-estuary nitrate enrichment is that of Laholm Bay, Sweden (Fig. 12.14). During the second half of the 20th century there were gradual increases in the land-derived loads of nitrogen, but not of phosphorus. Producers in Laholm Bay responded to increased nitrogen supply starting in the mid-1960s. Eventually, considerable blooms of ﬁlamentous macroalgae appeared, and clogged waterways and beaches, and exceptional plankton blooms became evident in the water. Studies of concentrations and ratios

Coastal producers almost invariably ﬁnd less available nitrogen than phosphate. Producers require a ratio of roughly 16 atoms of nitrogen to 1 atom of phosphorus for growth to take place. Concentrations of dissolved inorganic nitrogen (but not of phosphate) in sea water are markedly lower than in fresh waters (Valiela 1995); rivers, for example, transport terrestrial nutrients to estuaries in ratios of 50–500 nitrogen to phosphate (Hessen 1999), but there are major transformations and losses of available nitrogen in transit. In coastal waters the ratio of dissolved N : P ranges widely (Fig. 12.15), but N : P values were usually lower than 16 : 1. N : P ratios only approximately indicate the nutrient that might limit producer growth, but do make the point that in marine waters there is a relative dearth of nitrogen. The limiting role of nitrogen is not always the case for the growth of every producer, in every site, as shown by the few values larger than 16 (Fig. 12.15), and perhaps not at very long (geological) time scales, but holds as a reliable generality for many coastal waters.16 Silica may be important in determining whether diatoms are important components of phytoplankton (see references cited in Turner 2002). 16 More extensive discussion of the issues regarding nitrogen and phosphorus limitation of growth of coastal producers is available in Valiela (1995, chapters 2 and 14), Howarth (1988), and National Research Council (2000).

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NITROGEN

PHOSPHORUS

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The relative contribution of atmospheric nitrogen delivery to the nitrogen budget of different coastal ecosystems depends on the area of the watershed and water body, and on the precipitation regime. In large water bodies surrounded by relatively dry areas, atmospheric contributions might range up to 60% of total nitrogen loads (Paerl et al. 2002). Atmospheric deliveries to the

0 1990

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Nitrogen (tons yr−1) ( )

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Figure 12.13 Results of analyses of 17 different nutrient-enrichment experiments, done in a range of waters that can be classiﬁed as ranging from very polluted to pristine (top panels), and as to different marine environments (bottom panels). The response of the phytoplankton is expressed as ∆r, an index of response, normalizing the ratio of ﬁnal to initial phytoplankton populations. The index was necessary because the experiments were done with different concentrations of nutrients and phytoplankton, and incubations lasted for different spans of time. Adapted from Downing et al. (1999).

Figure 12.14 Time course of loads of nitrogen and phosphorus to Laholm Bay on the Baltic Sea, with notes on the producer responses along the way. From Rosenberg et al. (1990).

watershed and to the water surfaces of Waquoit Bay—an environment in a wet temperate region —ranged from 31 to 79% (Valiela & Bowen 2002). We know that much atmospheric nitrogen is intercepted within watersheds, particularly by forests, but atmospheric nitrogen inputs can exceed the nitrogen-intercepting capacity of

EUTROPHICATION

N limited

16

previously pristine environment now beginning to show incipient eutrophication. Nitrate transport by the Mississippi River to the Gulf of Mexico increased 3–4-fold across the last half of the 20th century (Fig. 12.16 top), primarily driven by excess fertilizer applied in the North American midwest. There are many other examples, such as increases in fertilizer-derived nutrients from rice ﬁelds converted from wetlands in the Mondego estuary of Portugal (Martins et al. 2001).

P limited

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297

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35

Figure 12.15 Frequency distribution of the nitrogen to phosphorus ratio (N : P) in waters of different estuary enviroments. The dashed line shows the position of the theoretical Redﬁeld 16 : 1 ratio between nitrogen and phosphorus. Data from Boynton et al. (1982).

watersheds (Aber et al. 1989; Fenn et al. 1998). The excess nitrogen may then ﬂow through the watersheds, and further enrich receiving waters. Such saturation is a problem of concern, because as we saw earlier, atmospheric inputs are often the largest term in watershed nitrogen budgets, and could seriously affect receiving ecosystems. Preservation of growing forests, plus action to curtail atmospheric nitrogen deposition within airsheds far distant from the sea, would be highly desirable priorities for the management of coastal waters.

Wastewater disposal The nitrogen enrichment of receiving waters is an inexorable correlate to increased number of people on a watershed. For example, in the eastern USA—not by any means the most crowded region of the earth—there has been a shift away from agricultural land use, toward urban land covers (with some reforestation) (see Fig. 1.12 top).17 The percentage of watersheds that are covered by urban areas differs, but many areas have surprisingly large proportions of urban cover (see Fig. 1.12 bottom). Population density increased in most places with urban sprawl. We might also recall from Chapter 1 that humans tend to accumulate near shores. Taken together, these statistics suggest that coastal watersheds will be increasingly urban, and hence waste water will increasingly reach coastal waters.18

Time trends in nutrient enrichment

Fertilizer use Increasing human populations across the last century have required larger agricultural harvests on increasing areas of cultivated land, which in turn have stimulated considerable increases in the worldwide production and use of nitrogen fertilizers (see Fig. 1.16). One result of increased use of fertilizers is that rivers and groundwaters carry substantially larger loads of fertilizerderived nutrients toward the coasts of the world. Increases in the area cultivated and fertilizer use have taken place, for instance, on the watersheds draining the northeast coast of Australia toward the Great Barrier Reef (Bell & Elmetri 1995), a

The aggregate impacts of atmospheric, fertilizer, and wastewater trends is that the nutrient content of fresh waters, and the transport of nitrogen toward estuaries and other coastal environments, 17

This trend is consistent with the predicted shift toward greater proportions of the global population becoming urban during the current century (see Fig. 1.11). 18 The pattern of replacement of forest and agricultural land uses with urbanization seems widespread. Terrestrial ecologists have studied the effects of such replacements, and concluded that the resulting nitrogen export depended in part on the mix of land covers. For example, watersheds with fast rates of urbanization and far more forest cover than agricultural land uses were most likely to show larger increases in nitrogen exports to receiving waters (Wickham et al. 2002).

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1.8

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have increased across the 20th century, and are likely to continue to do so. In different parts of the world’s coastlines, the hugely variable pattern of human population density, land covers, and other factors create a complicated and variable set of circumstances, with the result that different local coastal environments have been receiving quite different nitrogen loads. Nutrients transported to the sea by large rivers (e.g. Amazon, Orinoco), by force of their huge discharges, are a main conduit of land-derived materials to coastal seas and beyond (Bonilla et al. 1993; Demaster & Pope 1996; Howarth et al. 1996; Smith & Demaster 1996). The watersheds of these rivers, in spite of current concern about loss of tropical forests, probably have changed little across recent decades, and only as they

1980

Figure 12.16 Top: annual ﬂux of nitrogen (N) and discharge from the Mississippi region to the Gulf of Mexico, 1955–1996 (data from Goolsby et al. 2001). Bottom: nitrate concentrations in the water of several British rivers during the latter half of the 20th century (from Nedwell et al. 1999).

were affected by climatic changes. As land is increasingly cleared to support human activities through the present century, and more nutrients run off the land (Lewis 1986; Williams & Melack 1997; Meyer et al. 1998), these major South American rivers will certainly carry signiﬁcantly more nutrients to coastal environments. Nutrient concentrations in European rivers (Radach 1992) and groundwater increased earlier in the mid-1900s. Stricter controls on the use of fertilizers and disposal of waste water through many parts of Europe stopped the increases in concentrations during the 1980s to 1990s (Nienhuis 1992; de Jonge & van Raaphorst 1995; Iversen et al. 1998; Juhna & Klavins 2001). Examples of the increases (about four-fold in half a decade) and more recent leveling off can be seen in the nitrate

EUTROPHICATION

A primer on human waste disposal Human wastes are discharged with and into water. The mechanism might be as simple as an outhouse or a cesspool dug into the soil, or a septic system that includes a holding tank and a diffuser, to sewage treatment plants in urban areas. In rural areas wastes are deposited in cesspools that are merely holes in the ground, or septic systems that accumulate wastes in tanks where a degree of processing occurs (Novotny & Olem 1994; Valiela et al. 1997). Septic systems are widespread; in the USA, for example, 29% of the population uses septic systems for household waste disposal. Most of the nitrogen from septic systems reaches groundwater and is transported to receiving waters (Valiela et al. 1997). Sewage treatment plants process the wastes from urbanized areas. Sewage treatment plants are most frequently of the primary treatment type, where the efﬂuent may be screened of larger objects, and some disinfection might be applied. Far fewer are treatment plants that involve some degree of secondary treatment.

content of rivers in the UK during the 20th century (Fig. 12.16 bottom). Notice, however, that few rivers have shown decreases in nitrate content. In fact, even at the close of the 20th century, nitrate concentrations carried seaward were rather high (perhaps 700–2,800 µM nitrate) compared to the 1–10 µM usually found in temperate coastal waters. Similarly, concentrations of nitrate and phosphate in Chinese rivers increased 2–4-fold during the last half of the 20th century. It is not surprising, then, to ﬁnd that in Jiaozhou Bay, for instance, which receives inputs from many rivers draining agricultural and urban areas, concentrations of ammonium, nitrate, and phosphate increased about four-fold between the 1960s and 1990s (Shen 2001). In North America there is a patchwork of local differences in time courses, with some fresh waters bearing lower, and some, such as the

299

Such plants include holding tanks where microbial action attacks part of the waste matter, perhaps some of the solids falling to the bottom of the holding tanks are removed as sludge, and the efﬂuent is aerated and chlorinated to lower the concentration of human disease-producing organisms. A handful of urban areas of the world boast tertiary treatment plants, in which nitrate concentrations are lowered in specially designed tanks that foster denitriﬁcation, and other treatments remove phosphate from the efﬂuent. Only tertiary treatment plants address the issue of nitrogen interception, the limiting nutrient that fosters growth of organic matter that leads to eutrophication of coastal waters. The nitrogen in human waste water is therefore largely released into the aquatic environment, and to a large extent, we rely on natural biogeochemical transformations to furnish the interception that has so far held back a huge portion of the nitrogen we humans process. This interception has been referred to as the “assimilative capacity”, a feature that varies greatly from one environment to another, and is quite difﬁcult to deﬁne.

Mississippi River, carrying larger nutrient loads toward receiving waters. The aggregate forcing by wastewater disposal, use of fertilizers, and atmospheric deposition, across a variety of watershed land covers, however, may smooth over local heterogeneities to yield a notable long-term pattern of increased nitrogen loads to coastal waters (Goolsby et al. 2001).19 The degree of nutrient enrichment in speciﬁc coastal environments, as we might surmise, will 19

Land-derived nutrients may be carried to deeper waters in the large discharges of major rivers, but many coastal environments—salt marshes, mangroves, lagoons, bays, and so on—intercept substantial portions of the land-derived nutrients (Chapter 6; Galloway et al. 1996), and hence strongly modify the export to deeper waters. For example, the massive through-put from the Amazon watershed dominates nutrient cycles to the northeast of the estuary and of deeper Atlantic waters. The nutrient cycles of coastal marine environments to the southeast of the Amazon estuary are more inﬂuenced by nutrient transport and transformations within the mangroves that fringe much of the coast of Brazil (Dittmar & Lara 2001).

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depend of the intensity of human use20 of the speciﬁc watersheds21 feeding into the coastal environments. Because of these inﬂuences, land-derived nutrient loads are geographically quite variable, but at a global scale, and on a secular time scale, we should be prepared to see signiﬁcant increases in the delivery of land-derived nitrogen (as well as other nutrients) to the receiving coastal waters. The examples of eutrophication we have been discussing have been largely cast at the local watershed/estuary geographic scale. Because of the widespread and increasing eutrophication of speciﬁc waters worldwide, the aggregate pattern of these local impacts coalesce into a global-scale phenomenon. This contrasts with the pattern of other agents of change we discussed earlier, such as warming, sea level rise, and UV effects. These agents of change are driven by global atmospheric changes, and impact local environments. In both contrasting patterns there are effects at local and global scales, but the difference lies in the direction of the forcing that couples large- and smallscale effects. These features are important for the basic understanding of the changes taking place, as well as for planning to remediate effects.

Effects of increased nutrient supply Coastal environments, as we have just seen, are variously affected by human-mediated increases in eutrophication.22 The effects of eutrophication can range from small shifts in production and species composition, as may take place in waters exposed to incipient enrichment, to wholesale elimination of most organisms, as occurs in basins so eutrophic as to be devoid of oxygen. This section reviews a variety of effects detected in different 20

In many cases, increased nutrient loads to estuaries may, in part, be proximately aided by greater ﬂow of fresh water rather than directly result from human activities (McComb & Humphries 1992; de Jonge 1995; McComb & Lukatelich 1995; Goolsby et al. 2001). 21 In places where atmospheric deposition is a major input, the nutrients derive from a (usually) much larger “airshed”. The area of inﬂuence of atmospheric deposition as a source of nutrients depends on air currents, and the loads carried are largely owing to industrial activity and combustion of fossil fuels. 22 Reviews are provided by Nedwell et al. (1999), National Research Council (2000), and Hessen (1999).

coastal organisms owing to increases in the supply of nutrients, and the consequent effects that might cascade up various links in coastal food webs.23 Producers Phytoplankton

The response of phytoplankton to increased nutrients is a topic that has received much research attention, and a variety of approaches have been applied. One approach has been to plot nutrient concentrations versus phytoplankton biomass or production from data collected across many coastal sites, at different times. Such comparative plots show considerable scatter, but, in general, larger nitrogen supply is signiﬁcantly associated with larger phytoplankton abundance or activity. This conclusion may be reached from data collected on the changes in nitrogen supply as sewage treatment was introduced in a Swedish estuary (Fig. 12.17 top)24 or from data comparing diverse natural and experimental systems (Fig. 12.17 bottom). On aggregate, the available evidence suggests that, in time scales of up to decades, inputs of nitrogen into marine waters can be conﬁdently expected to lead to more phytoplankton biomass or phytoplankton production. There is signiﬁcant “bottom-up” forcing of phytoplankton by nitrogen supply in coastal food webs.

23

Such a propagation of effects prompted by changing nutrient supplies, exerted up into different links in food webs has been referred to as “bottom-up” control. There is another vigorously espoused view, referred to as “top-down” control (already mentioned in Chapter 11), that argues for the primacy of controlling effects of predators and grazers on organisms at lower steps in the food web. An example might be the removal of grazers or predators by humans or disease, which could release pressure on producers or grazers (Hughes 1994; Jackson et al. 2003). Lively discussions have resulted between partisans of such “top-down” controls and advocates of “bottom-up” explanations (Valiela et al. 2004). As in all such disputes, both sides are right: nutrients and grazers seem to play some role. Although in this chapter we focus on nutrientdriven effects, in the examples reviewed below we run into cases where multiple controls seem to play a role. 24 Other examples of increased concentration of phytoplankton related to greater nitrogen loads can be found in Boynton et al. (1996).

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6 5 4 3 2 1 0 200

Primary production (g C m−2 yr−1)

In many coastal environments, the effects of nutrient supply on producers can be modiﬁed or mediated to some degree by other factors, such as hydrodynamics, light supply, grazing, and changes in climate. If the time that water resides within an estuary, for example, is short, organisms might have insufﬁcient time in which to grow in response to the nutrient or food supply within that particular estuary (Pace et al. 1992). Another effect of fast ﬂushing through an estuary might be that nutrients (or food) are swept to deeper water before estuarine organisms can take them up. Fast ﬂushing can lower within-estuary nutrient retention (Howarth et al. 1996), and hence lead to greater nutrient export to deeper waters. Insufﬁcient light may limit production. For example, at certain times of year phytoplankton production in San Francisco Bay may be limited by low light, owing to turbid conditions from sediment (Cloern 1999). Grazers can constrain phytoplankton abundance. In Chesapeake Bay, for example, oysters reefs have been drastically reduced since the late 19th century by overﬁshing. The oyster stocks present before 1870 would have been able to harvest more than 77% of the phytoplankton produced in 1982 (Newell 1988). These grazers doubtless restrict the abundance of their food particles, and hence affect the turbidity of areas such as Chesapeake Bay. Similarly, Hughes (1994) concluded that the blooms of macroalgae covering corals are the result of overﬁshing of herbivorous ﬁsh. Climatic shifts may also alter the rate and mix of nutrients delivered to coastal systems (Hessen 1999). For example, shifts in precipitation or temperature regime can alter nitrogen delivery by rivers to coastal waters, through the effect of complex, and largely poorly understood, mechanisms (altering evapotranspiration rate of vegetation, runoff volumes and velocity, and so on).

Chlorophyll a (µg l−1)

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Figure 12.17 Top: measurements of phytoplankton chlorophyll and total nitrogen (N) load entering Himmerfjarden estuary, Sweden, taken across a period of years as sewage treatment was improved and N concentrations decreased. Bottom: rate of phytoplankton production in relation to inputs of dissolved inorganic nitrogen in natural and experimental marine ecosystems. Data for both panels from several sources, adapted from National Research Council (2000). Harmful algal blooms

Certain kinds of phytoplankton are harmful or noxious because they produce compounds toxic to humans, shellﬁsh, ﬁsh, and marine mammals,25 or because their blooms lead to unsightly scums or foams, depletion of oxygen as the blooms decay, and alteration of habitats by shading. The 25

Reviews of the many types of harmful algal blooms, and their characteristics, are available in Richardson (1997), Zingone and Enevoldsen (2000), and Anderson et al. (2002).

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algae involved are taxonomically diverse.26 Their appearances have been popularly referred to as red, green, and brown tides, and on aggregate are called harmful algal blooms (HABs). Effects of truly toxic HABs on people include paralytic shellﬁsh, diarrheic, amnesic, and neurotoxic shellﬁsh poisoning, and ciguatera ﬁsh poisoning, among other illnesses. About 2,000 cases of human poisoning are reported each year, and several hundred fatal cases occur each year worldwide from the consumption of shellﬁsh contaminated with toxic phytoplankton (Hallegraeff 1993; Zingone & Enevoldsen 2000). Toxic blooms are threats to many marine populations, including some involved in shell- and ﬁnﬁshing as well as to mariculture (Shumway 1992; Bricelj et al. 2001). There are many instances of such damage. One example is a bloom that caused the death of 900 tons of cod, salmon, and trout in the North Sea off Denmark. Another instance is the starvation and recruitment failure of bay scallops on the East Coast of the USA, attributed to blooms of brown tides (Zingone & Enevoldsen 2000).27 The mortality of cultivated quahog clams in a site with brown tide bloom was 67%, compared to less than 5% in a nonbloom site (Greenﬁeld & Lonsdale 2002). In 1987, 14 humpback whales died in Cape Cod Bay after eating mackerel that contained saxitoxin produced by dinoﬂagellates (Geraci et al. 1989). “During the last two months the inhabitants in Rhode Island witnessed the following remarkable phenomenon. The water of a considerable portion of the Bay became thick and red, emitting an odor 26

Out of a total of 3,365–4,024 species of phytoplankton, only 184– 267 form blooms that are in some fashion noxious, and only 60–78 species of phytoplankton (45–57 of which are dinoﬂagellates) are truly toxic (Sournia 1995). Zingone and Enevoldsen (2000) agree, allowing that out of the approximately 4,000 marine planktonic microalgal species, there might be about 80 toxic species, and perhaps about 200 noxious species. They add that although identiﬁcation of these taxa is challenging, new taxa are being added yearly to these lists. 27 Brown tides (Bricelj & Lonsdale 1997) are caused by certain very small phytoplankton (chrysophytes) that do not take up nitrate, and have appeared in a few places (Narragansett Bay, off Long Island, NY, Laguna Madre in Texas, and the Benguela Current off South Africa). Brown tides are long-lasting, and achieve biomasses that shade and eliminate seagrass beds (Dennison et al. 1989), and reduce shellﬁsh growth and abundance (Gallagher et al. 1989).

almost intolerable to those living near by. The situation became alarming when, on the 9th and 10th of September, thousands of dead ﬁsh, crabs and shrimps were found strewn along the shores or even piled up in windrows. . . .” Mead (1898).

The occurrence of HABs may have become more common in recent decades; known toxic dinoﬂagellate species28 and the number of coastal countries reporting HABs have increased markedly. It is not evident, however, whether this emergence of blooms is a reﬂection of new interest and detection methods or constitutes an actual surge of blooms (Wyatt 1995). There is paleogeological evidence of the presence of different types of dinoﬂagellate cysts in many sites, and many species may simply have remained undetected until now. It does seem likely, however, that the proliferation of human activities in coastal waters may be somehow involved. Increased transport of water across large distances in tanker ballast water, and maricultural practices, for example, may have dispersed toxic species to new areas (Hallegraef 1993). There is no conclusive evidence about the speciﬁc mechanisms that may foster dinoﬂagellate blooms (Richardson 1997; Anderson et al. 2002); there have been arguments for nutritive effects, as well as for oceanographic effects. There is circumstantial evidence of a link between nutrients and red tides. Red tides increased by more than an order of magnitude between 1976 and 1986 in Chinese near-coastal waters, nearly in parallel with the ﬁve-fold increase in human population and their increased waste release (Lam & Ho 1989). The increased delivery of nutrients derived from fertilizer use on land may also be associated with the incidence of red tides (Fig. 12.18). Red tides were more frequent at lower N : P levels (Fig. 12.18 inset), suggesting a connection with a greater availability of nitrogen.29

28

Burkholder (1998) reported that 55 species were recognized, compared to about 20 only a decade earlier. 29 Other nutrient-related control mechanisms have also been suggested. These include altered ratios of concentrations of some nutrients relative to others; increases in certain organic compounds that might be taken up by the dinoﬂagellates, or might bind metals toxic to dinoﬂagellates; and limitation by low concentrations of essential trace metals (Boyer & Brand 2001).

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40 No. red tides

50

Number of red tides

40

Redfield 16 : 1 ratio

30 20 10 0 10 12 14 16 18 20 22 N:P

30 20 10 0 0

4 8 12 16 Fertilizer use (tons × 106 yr –1)

20

Figure 12.18 Relationship of the number of red tides per year in Chinese coastal waters (1970–1992) and annual use of fertilizer on adjoining land. Inset: number of red tides per year plotted against the average ratio of nitrogen to phosphorus (N : P) in water, in the same coastal sites. Data from sources cited in Anderson et al. (2002).

Surges in dinoﬂagellate abundance might be started and maintained by oceanographic features that may move seed populations into regions favorable for growth (Anderson et al. 2002). There might also be important global-scale oceanographic effects: ﬁsh kills by red tide blooms took place during the intense 1997–1998 El Niño in Hong Kong (Zingone & Enevoldsen 2000). Dinoﬂagellate blooms might (Usup & Azanza 2001) or might not (Azanza & Taylor 2001) be related to El Niño–Southern Oscillation (ENSO) anomalies; the mechanisms underlying such relationships are unknown. The mechanisms that control brown tides are also poorly known. Changes in abundance of certain grazers (top-down effects) may release brown tide organisms to bloom (Buskey et al. 1997, 2003). Others have argued that brown tides are inhibited by inorganic nitrogen, stimulated by organic nitrogen (LaRoche et al. 1997), and unresponsive to urea supply (Gobler & SañudoWilhelmy 2001), and that iron supply might or

303

might not be involved in brown tide blooms (Gobler & Sañudo-Wilhelmy 2001). The mechanisms that govern the appearance of HABs, judging from what is known about red and brown tides, might be quite diverse and species-speciﬁc. Moreover, there are likely to be multiple factors at play, including bottom-up as well as top-down inﬂuences (Gobler et al. 2002). The extreme species-centered diversity of HABs is evident in the case of curious organisms, known as species of Pﬁesteria, that occur in shallow muds of poorly ﬂushed and highly eutrophic salt marsh estuaries in the Middle Atlantic states of the USA (Burkholder et al. 2001a, 2001b). These organisms proliferate where there are increased anthropogenic nutrients; they ingest microalgae that increase rapidly in these conditions. Pﬁesteria are not algae: they cannot take up dissolved nutrients by themselves, nor carry out photosynthesis, but after they incorporate the organelles of their algal food they can use them to take up dissolved nutrients and ﬁx carbon. This extraordinary nutritional strategy certainly merits the sesquipedalian term “kleptochloroplastidy”. Pﬁesteria attracted public attention by their ability to rapidly swarm out of the muds—press reports were about “phantom dinoﬂagellates”—to aggressively feed on the wounds and carcasses of ﬁsh, and by producing toxic compounds that volatilized and made people ill. The Pﬁesteria phantom attacks on ﬁsh have been conﬁrmed, but there is some doubt about the production of toxins that poisoned ﬁsh or created human ailments (Berry et al. 2002; Vogelbein et al. 2002). Macroalgae

There are many reports of increased growth of macroalgae exposed to increased nutrient supplies. For the most part, nitrogen seems to be the limiting nutrient, but there are many exceptions. Photosynthesis and growth of macroalgae from southern Florida increased when ammonium was added, but not when phosphate was provided (Lapointe & Hanisak 1997). Nitrogen supply limited the growth of eight out of nine species of macroalgae growing in nutrient-depauperate

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coral reefs30 in Kanehoe Bay, Hawaii (Larned 1998). A survey of published reports on enrichment experiments found that out of 36 species tested, inorganic nitrogen supply increased growth in 22 species, and inorganic phosphate did so in 17 species (Larned 1998). In the majority of reported cases, stimulation of macroalgal growth seems associated with wastewater nitrogen sources. Stable isotopic signatures —as we already saw in the Waquoit case—of macroalgae that are blooming or forming dense canopies appear to be heavy, suggesting the inﬂuence of waste water (Lapointe & Barile 2001). In more pristine environments, the macroalgae, when found, are not in dense canopies, and bear lighter isotopic signatures, indicating that nitrogen ﬁxation or atmospheric deposition furnished the nitrogen (France et al. 1998). Different species of macroalgae respond differently to nutrient enrichment. We already saw one example in the case of the macroalgae of Waquoit Bay, where two species increased, and several others did not. The response can be more complex: growth rates of Sargassum baccularia, a common brown alga of Australian coral reefs, almost doubled within concentrations of 3–5 µM ammonium, but higher concentrations reduced growth (Schaffelke & Klumpp 1998). This response was not mirrored in other species of macroalgae. In turn, macroalgae can alter the nutrient environment in which they grow. We saw in the Waquoit example that nitrogen uptake into macroalgae was sufﬁcient to lower ambient 30

Dissolved inorganic nitrogen and phosphate are in low concentrations in waters overlying coral reefs. The mean ± SD of nitrate was 0.25 ± 0.28 µmol−1, and 0.13 ± 0.08 mmol−1 for phosphate, in data from 1,000 coral reefs worldwide (Kleypas et al. 1999). Macroalgae living in these circumstances must use a variety of strategies to obtain a sufﬁcient nutrient supply. These strategies include using morphological adaptations to take up nutrients from water ﬂowing over fronds (Fujita & Goldman 1985); storing nutrients for long periods of time in fronds (Mann 1973); creating, through their own architecture, microenvironments in which higher concentrations of nutrients, perhaps diffusing out of sediments, for example, may accumulate; and use of nitrogen-ﬁxing symbionts (Larned 1998). These mechanisms must also be useful even in waters with richer nutrient supplies. Because of the diversity of mechanisms available to procure sufﬁcient nutrients, it should not be surprising to ﬁnd it difﬁcult to generalize about which nutrient limits growth, or if nutrients are involved as the major control of macroalgal growth.

nutrient concentrations in the water column. This seems a general feature; for example macroalgae intercept a substantial part of ammonium regenerated from the sediments below the canopies (Bierzychudek et al. 1993; McGlathery et al. 1997), hence lowering the nutrient supply for other producers. So far in this section, I have been concerned with the inﬂuence on producer growth by processes cascading up food webs, but there are, of course, top-down effects, which may or may not be clear-cut. For instance, macroalgal biomass in Glover’s Reef, Belize, was more strongly inﬂuenced by herbivory than nutrient supply, but the opposite was true for macroalgal cover; and different kinds of algae showed distinctly different responses to nutrients and grazers (McClanahan et al. 2003). Such species- and site-speciﬁc differences are a pervasive feature of the biology of macroalgae. A fruitful line of future study might be to discern the circumstances that affect the balance between the two types of controls in different environments, and allow one or the other mechanism to assume a greater role in different coastal systems. Seagrasses

Much as in the Waquoit Bay example, macroalgal canopies have replaced seagrass meadows in many nutrient-enriched sites across the world’s coasts (Raffaeli et al. 1998). A review of worldwide data on seagrass meadows showed that the relative contribution by seagrasses to total production in coastal environments dropped precipitously where nitrogen loads increased (Fig. 12.19 top), and seagrass cover decreased as nitrogen loads got higher (Fig. 12.19 bottom). The overall conclusion is that seagrass meadows disappear in estuaries receiving nitrogen loads above 100– 200 kg N ha−1 yr−1. The literature on direct responses of seagrasses to increased nutrients includes remarkably variable results. In nutrient-depauperate environments such as the Great Barrier Reef, growth rates of seagrasses may be nitrogen-limited (Udy et al. 1999). Insufﬁcient nitrogen loads from the mainland may be responsible for the limited extent of

EUTROPHICATION

Most evidence suggests that increased nitrogen loads indirectly impair biomass, growth, and production of seagrasses, as was made clear in the Waquoit Bay case history. Loss of seagrass stands (Fig. 12.19) occur mainly through the shading that results from the increased biomass of other producers (macroalgae, phytoplankton, epiphytes) in the same habitats (Duarte 1995; Hauxwell et al. 2003).

(Seegrass production/ total production)/100

100 % sp = 145.653 Nload F = 15.5***

80

–0.550

60 40 20 0 0

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Seegrass cover lost (%)

100 80 60 40 % sg loss = 0.693 Nload + 14.211 F = 14.6*

20 0

0

200 400 600 N load (kg N ha–1 yr –1)

800

Figure 12.19 Top: seagrass production as a percentage of total primary production in many estuaries, plotted vs. land-derived nitrogen (N) load to each estuary. Bottom: percent of area of seagrass meadow lost (over 10–30 years, depending on site) in several estuaries, plotted vs. the corresponding land-derived nitrogen load. * and ***, statistically signiﬁcant at the 0.05 and 0.001 levels of probability, respectively. From Valiela and Cole (2002).

seagrass meadows in the southern Great Barrier Reef; local increases in seagrass beds near sites frequented by tourists may be a result of larger nitrogen loads in these sites.31 Other reports tell us that increased nitrogen concentrations in water may have a direct detrimental effect on seagrass growth (Burkholder et al. 1994). 31

305

Stable nitrogen isotope signatures of corals in the Great Barrier Reef suggest that corals—as do the seagrasses—incorporate waste water-derived nitrogen in inshore sites, and that nitrogen derived by natural nitrogen ﬁxation is relatively more important to corals in areas farther away from the coast (Sammarco et al. 1999).

Corals

Corals survived well and became more colored after experimental nitrogen enrichment (McClanahan et al. 2003). The coloration was a measure of increased density of zooxanthellar symbionts within the tissues. On the positive side, improved nitrogen supply enabled corals to counter bleaching (Chapter 2). On the other hand, a greater density of zooxanthellae may increase competition with the coral polyps for whatever carbon supply may be present within the coral tissues. This may diminish calciﬁcation by the coral polyps (Ferrier-Pagès et al. 2000). If calciﬁcation is reduced, the growth of coral heads may be slower, and corals may be overgrown by macroalgae (River & Edmunds 2001). These are interesting possibilities, but there is still much to be learned, because—as in the case of macroalgae—in corals there are species-based differences of considerable magnitude (FerrierPagès et al. 2000; McClanahan et al. 2003). Corals are indirectly affected by the overgrowth by macroalgae that is often a result of nutrient enrichment. Nitrogen-stimulated blooms of the green alga Codium isthmocladum grew to form canopies up to 2 m deep, and smothered many species in reef habitats off Florida, including corals (Lapointe & Barile 2001). Such smothering is increasingly common along the tropical shorelines of the world, and many of the bloom species are aliens newly arriving at reefs (Carlton & Scanlon 1985; Lapointe & Barile 2001). Microbes Increased supply of nutrients and organic matter alter microbial abundance and activity, changes

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that sharply reset the cycling of biogeochemically important elements, including carbon and nitrogen. Such alterations have signiﬁcant consequences within coastal environments, as well as at global scales. Alterations to the carbon cycle

Large increases in nitrogen delivery, in almost all cases from land, fosters, as we have seen above, the production of organic matter that is then available to microbes. In most aquatic systems, the particles produced within the water column or near the bottom fall to the estuary ﬂoor and are available for microbial degradation. Microbes consume oxygen32 as they degrade organic matter, and the consumption is often faster than the rate at which oxygen can be renewed by hydrodynamic processes.33 These circumstances create hypoxic or anoxic zones. In such low oxygen zones, microbes continue to metabolize organic matter, but are forced to use electron acceptors other than oxygen, including nitrate, iron, manganese, sulfate, and hydrogen, as conditions become more reducing. Microbes that use hydrogen produce methane, an active greenhouse gas, but most of the methane is oxidized within coastal sediments, and there is some question as to the magnitude of the contribution of methane from coastal sediments to the atmosphere. The major well-known microbial carbon-related consequence of eutrophication is, therefore, the increase in bacterial use of oxygen, and the creation of low oxygen conditions. 32

Of course, microbes are not the only group of organisms that consume oxygen. In the Waquoit example we saw that the respiration by macroalgae was sufﬁcient to create hypoxia; the fauna consumed oxygen as well. In most coastal environments, however, we can suppose microbes are responsible for most of the metabolism that consumes oxygen. 33 Renewal of oxygen content of water in an environment may be brought about by diffusion from the atmosphere (a slow process), or by turbulent or advective mixing (faster processes) with oxygencontaining water. Unless the hydrodynamic setting provides circumstances in which water is somehow contained, hypoxia will likely fail to develop. In Long Island Sound, for instance, development of late summer hypoxia depends on transport and containment of water masses (Torgersen et al. 1997). The effects of hydrodynamics is powerful enough to suggest to Falkowski et al. (1980) that anoxia occurring in the New York Bight could have happened even in the absence of inputs of sewage from New York City.

Anoxic zones occur naturally in many marine environments (Levin 2002), such as fjords, the Black Sea, upwellings, shallow lagoons, enclosed bays, and certain sea mounts. In these environments low oxygen conditions are a result of nutrient-rich overlying waters plus impaired ﬂow. Such low oxygen sites existed before humans made much of a difference. Human-derived nutrient inputs, however, are responsible for most cases where low oxygen has recently become a problem in coastal waters. Worsening low oxygen conditions have been reported in the Baltic (Elmgren 1989, 2001), Chesapeake Bay (Ofﬁcer et al. 1984; Boesch et al. 2001), Long Island Sound (Parker & O’Reilly 1991), and coastal Denmark (Iversen et al. 1998), among many other sites. In many Japanese coastal areas, such as Mikawa Bay, water column oxygen has decreased alarmingly through the last half of the 20th century. The place that perhaps has come to be a classic example of expanding hypoxia is that of the Northern Gulf of Mexico, where increased nitrogen discharges from the Mississippi River watershed (Fig. 12.20 top three panels) have fostered hypoxic near-bottom conditions during the last decades of the 20th century. Hypoxia, although increasing, is a highly variable condition (note, for example, the lower area in 2000; Fig. 12.20 bottom panel). Alterations to the nitrogen cycle

Increased supply of nitrogen compounds alter rates of many transformations of this element within coastal environments by altering activity of bacteria involved in the nitrogen cycle. In mangrove sediments, for example, the addition of nitrogen-containing waste water increased the density of bacteria populations in general, and, in particular, of nitriﬁers and denitriﬁers (Tam 1998). Some of the anthropogenic dissolved nitrogen enters the coastal environments as ammonium;34 34

More ammonium also lowers rates at which nitrogen gas is converted internally to organic nitrogen by certain microbes; such ﬁxation of N2 is energy demanding, and is unnecessary if reduced nitrogen is otherwise available for uptake and synthesis of protein (Van Raalte et al. 1974; Carpenter et al. 1978). Thus, the increased external ammonium in coastal waters may be, to a modest extent, compensated by a lowering of ﬁxation. This is another, if small, part of the “assimilative” capacity of aquatic environments for nitrogen loads.

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30.0°

Atchafalaya R.

29.5°

2.0 2.0 2.0

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Latitude

Mississippi R. 2.0

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July 7–18, 1986 −93.5°

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Atchafalaya R. Mississippi R.

29.5° 2.0

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July 23–27, 1990 −93.5°

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July 23–28, 1996 −93.5°

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Longitude

Area of bottom water hypoxia (km × 103)

25 20 15 10 5 0 1985

1987

1989

1991

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Year

Figure 12.20 Expanse of near-bottom water in the northern Gulf of Mexico with less than 2 mg O2 l−1 (gray area) during mid-summer, for three different years (top 3 panels). Bottom: area of hypoxic bottom water in the same sites, 1985–2000. From Rabalais and Turner (2001).

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nitrifying bacteria use energy contained in ammonium (a reduced, and hence energy-rich, form of nitrogen), releasing nitrate as an endproduct. Increased nitrate concentrations—new nitrate entering the coastal system, plus nitrate produced by nitriﬁers—is a common symptom of eutrophication. Another group of microbes (denitriﬁers) use the nitrate as an acceptor for the excess electrons that result from the metabolism of organic matter, and convert the nitrogen into nitrogen gas. Denitriﬁers can be thought of as countering the effects of eutrophication, because they convert a biologically active form, nitrate, into nitrogen gas, a form of nitrogen largely unavailable to most organisms, and hence relatively inactive. Rates of denitriﬁcation in aquatic environments increase as nitrogen loads increase (Seitzinger 1988; Valiela 1995), but are unable to keep up with the rise in nitrogen loads that humans force on aquatic systems. Denitriﬁers may process as much as half the external load, where loads are relatively low, but that portion decreases as external loads increase (Valiela 1995, chapter 14). Denitriﬁers are certainly part of what has been referred to as the assimilative capacity of coastal systems for nitrogen enrichment, but that valuable capacity has limits readily exceeded by human inputs.35,36 35

Recently it was discovered that there is an additional mechanism through which inorganic nitrogen can be converted to N2 (Daalsgard et al. 2003; Kuypers et al. 2003). Certain bacteria can oxidize ammonium with nitrate and nitrite, to yield N2. This has been called the anammox reaction, a pathway that occurs in ammoniumrich, anoxic, aquatic environments. Estimates done so far suggest that 10–35% of the marine production of N2 might be accomplished through the anammox pathway. Anammox occurs in conditions similar to those under intense eutrophication, so it may contribute an as yet undetermined assimilative capacity for inorganic nitrogen in coastal waters. 36 Transformations of nitrogen, carbon, and oxygen are tightly coupled in coastal environments, and they often interact in complex ways. For example, low oxygen may exacerbate the effects of external nitrogen loading to estuaries (Kemp et al. 1990). Low oxygen in the bottom water and sediments of Chesapeake Bay impaired the activity of nitrifying bacteria, which in turn meant that denitriﬁers lacked a supply of nitrate. The impairment of nitriﬁers therefore diminished the conversion of nitrate to the biologically more inert N2 gas. Thus, the ammonium that was regenerated from sediments remained in the water column, adding to the dissolved inorganic nitrogen content contributed by external sources. These interactions added to the level of eutrophication down the bay.

Eutrophication worldwide may be linked to global atmospheric changes. Nitriﬁcation and denitriﬁcation release nitrous oxide (N2O) into water and the atmosphere. The release of N2O from nitriﬁcation increases in low oxygen conditions, and from denitriﬁcation in high nitrate concentrations. This is of interest, because as mentioned in Chapter 2, N2O is an active greenhouse gas and is involved in ozone depletion. There is discussion in the literature about whether coastal environments are sources or sinks of N2O.37 Emissions of N2O from rivers and estuaries may amount to up 24–30% of total global anthropogenic emissions38 (Seitzinger & Kroeze 1998) or 60% of the global marine emissions (Bange et al. 1996). These are substantial contributions, even though estuaries make up a tiny area of the world’s surface. Moreover, most of the N2O released from rivers and estuaries is of anthropogenic origin, and eutrophic estuaries have higher rates of release of N2O to the atmosphere (Robinson et al. 1998; Seitzinger & Kroeze 1998). To the extent that coastal eutrophication is increasing worldwide, there will be increased release of N2O to the global atmosphere. Consumers The effects of increased nitrogen loads on species of consumers are as diverse as the consumers themselves. Eutrophication may not affect,39 or may favor, or impair, the growth and abundance of consumers in a variety of ways. These effects include direct changes in the quantity and quality of food supply, and indirect effects such 37

Discussion of the various views can be found in Seitzinger and Kroeze (1998), Robinson et al. (1998), Corredor et al. (1999), de Wilde et al. (2000), Marty et al. (2001), Dong et al. (2002), and LaMontagne et al. (2003). 38 Sources summarized by Seitzinger and Kroeze (1998) suggest that of the 4–5 Tg N yr−1 total global emissions estimated from measurements, 1.2 are from rivers and estuaries, 1 from fertilizerinduced soil emissions, 0.5 from biomass burning, 1.3 from industrial sources, and 0.4 from cattle feed lots. 39 In certain coastal waters there are no discernible effects of increased nutrient loading on consumers. The catch per unit effort of estuarine ﬁsh in a eutrophic Australian estuary with evident macroalgal and phytoplankton blooms, was indistinguishable from catch in a similar but much less eutrophic estuary (Steckis et al. 1995). Such a lack of measurable impact is by far the exception rather than the rule.

EUTROPHICATION

as reduction of suitable habitat through loss of key species and lowered oxygen. Improved food supply

Eutrophication yields more and better quality food particles for certain consumers, particularly for shellﬁsh, as we saw in the case of shellﬁsh in the Waquoit estuaries, where the growth of consumers increased in parallel with nitrogen loading rates. The particles available for suspension and other feeders in eutrophic estuaries tend to be more plentiful, and seem of greater nutritive quality, because of higher nitrogen content, and perhaps different species composition.40 The stimulatory bottom-up effect of eutrophication, via improved food supply, can thus cascade up food webs, penetrating across more than one feeding step. For example, increased nitrogen supply fosters larger phytoplankton production. Such increases in primary production may be associated with an increased ﬁshery catch across many coastal sites (Fig. 12.21). The association is only a correlation, but there is evidence of similar effects on quasiexperimental

Fish landings (kg ha−1 yr−1)

160

bases, for example in Waquoit Bay, and in actual experiments in large tanks (Nixon & Buckley 2002). Nutrient enrichment in the Baltic during recent decades provides another example of pervasive changes up food webs: more land-derived nutrients led to a 43% increase in production of organic matter, and a 34% increase in production by zooplankton and benthos (Elmgren 1989). The greater supply of prey may have led to a 122% increase in ﬁsh production, which in turn supported a 115% increase in food for predators that harvested ﬁsh. The nutrient enrichment thus forced major changes that cascaded upwards through the Baltic food web.41 Analyses of other data suggested that there was only a weak coupling between primary production and ﬁsh yield (Micheli 1999), but more recently Ware and Thompson (2005) found a rather strong association between production by phytoplankton and ﬁsh yield in the Northeast Paciﬁc. The degree of upward coupling between producers and consumers is a topic that certainly merits further study, but I expect that we will ﬁnd out that, indeed, there are bottom-up effects via nutrient supply that signiﬁcantly “cascade” into upper links of marine food webs. Changes to habitat

140 120 100 80 60 40 20 0 0

300 400 100 200 Primary production (g C m−2 yr −1)

500

Figure 12.21 Relationship of data on rates of primary phytoplankton production (forced by nitrogen inputs) relative to ﬁsh landings from many marine ecosystems. From many sources, cited in Nixon and Buckley (2002). 40

309

In certain waters, nutrient enrichment leads to blooms of phytoplankton species with noxious characteristics. In these cases, particles available to suspension-feeders may be of undesirable quality as food for consumers.

Indirect changes in habitats driven by eutrophication are of consequence for many consumers. We saw one example of this in the elimination of eelgrass meadows in the enriched waters of Waquoit Bay estuaries (see Fig. 12.8). Species that depend on seagrasses, such as scallops and certain ﬁsh, nearly disappeared after the seagrasses were replaced by macroalgae. In an estuary in Nova Scotia, the collapse of eelgrass meadows led to a decline in abundance of geese and ducks that fed on the eelgrass (Seymour et al. 2002). 41

The changes were to an extent modiﬁed in the upper trophic steps by the harvest of marine mammals by people. Because of the harvest of seals in the Baltic during the 20th century, the portion of ﬁsh eaten by these marine mammals decreased by 300%. Concomitantly, the take by ﬁshermen increased by 75% (Elmgren 1989). This interaction between eutrophication and ﬁshing once again shows that human effects on coastal environments are seldom separable into neat categories, and that bottom-up and topdown effects interact in most situations.

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16

Lowered oxygen

Combined effects of eutrophication

From the examples just reviewed, we might expect that, as nutrient enrichment increases, eutrophication will have a sequence of effects, from positive to negative, as eutrophication passes from incipient to severe. Bottom-up effects of nutrients that alter favorable habitats and lower oxygen may impair the potential control of food species by consumers. We saw in the Waquoit example that grazing could match macroalgal growth rates when nutrient supply was low. As nitrogen loads increased, that control potential was lost, and algae proliferated. Similar shifts occur in Danish estuaries (Geertz-Hansen et al. 1993). The comparisons of watersheds in the Waquoit Bay system are probably the closest depiction available of such a transition. There is an additional, more geographically diverse, data set that provides a

Chlorophyll (µg m−3)

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1,000

6 Fish catch (metric tons km−2)

Eutrophication indirectly affects consumers by lowering oxygen. In the hypoxic zone off the Mississippi delta, there were fewer zooplankton in layers of water containing less than 1–2 mg O2 l−1 (Marcus 2001; Qureshi & Rabalais 2001), and few if any benthic animals on the sea ﬂoor under those areas (Diaz & Rosenberg 2001). Where there was no oxygen in the water above, only mats of cottony sulfur bacteria were evident on the sediment surfaces (Rabalais et al. 2001). In Long Island Sound, of 18 species of commercial interest that were sampled by trawls, 15 were dramatically more frequent at sites where the bottom water oxygen concentrations were greater than 3 mg l−1 than at sites with an oxygen content below 2 mg l−1 (Howell & Simpson 1994). The establishment of macroalgal canopies often leads to anoxic sediments below the canopies (Raffaeli et al. 1998). In the Waquoit case history, this effect led to fewer benthic invertebrates (Hauxwell et al. 2003); in reefs off Florida, the low oxygen associated with dense macroalgal canopies was associated with the elimination of benthic species, including corals, sponges, and other species, and declines in reef ﬁsh (Lapointe & Barile 2001).

4

2

0 1

10 100 Index of human activity (I/A)

1,000

Figure 12.22 Estimates of mean phytoplankton mass (top) and ﬁsh catch (bottom) from 14 coastal semi-enclosed seas around Europe, plotted vs. an index of human activity on the watersheds (I/A is the ratio of number of inhabitants on the watersheds divided by the area of the marine system receiving inputs from the watersheds). Triangles denote seas with reported eutrophic conditions and lowered oxygen. Data adapted from de Leiva Moreno et al. (2000); Breitburg (2002).

glimpse at the possible sequence of the effects of increasing eutrophication on consumers. Data on ﬁsh catch, and density of people on watersheds contributing materials to the seas where the ﬁsh were caught, can be compiled from various sources, for various European coastal seas (Fig. 12.22). We saw earlier that more people on land increased nutrient loading to estuaries, and that this in turn stimulated primary production in the estuaries. We do not have direct assessments of nutrient loads for these waters, but we can calculate a proxy index for the relat-

EUTROPHICATION

ive nitrogen load entering the water bodies. Human density on the watershed divided by the area of the receiving water body (to allow for the large differences in area of the receiving water bodies involved) could be a suitable proxy. The data on phytoplankton and ﬁsh harvest were variable, but patterns may be discerned (Fig. 12.22). The amount of phytoplankton in the different coastal seas seemed to increase as the index of loading increased, throughout the range of values of the index (Fig. 12.22 top). In contrast, there were higher ﬁsh harvests as the ratio of people on the watershed to water surface increased initially. Then, as the index increased further, larger harvests become rarer (Fig. 12.22 bottom). These data, though certainly speculative, suggest that an increased human impact fostered phytoplankton production throughout. Assuming that ﬁsh harvest is a reasonable measure of abundance, we may infer an initial increase in consumers as greater human inﬂuence was exerted on the aquatic systems (presumably as a consequence of better food quality and quantity). At some point, however, some deleterious effects of human activity may impair ﬁsh harvests; perhaps higher values of the index were associated with more extensive hypoxia (de Leiva Moreno et al. 2000; Breitburg 2002). The Waquoit example, plus the other evidence discussed here, suggest that the effects of eutrophication depend on the degree of external nutrient enrichment entering the local environment, and how the direct and indirect effects affect the array of consumers present. The effects of eutrophication range from stimulatory to impairing, and seem quite species-speciﬁc. Unfortunately, nitrogen loads hardly ever stop at incipient to lower levels of eutrophication, where effects seem largely favorable. As human beings increase the intensive use of coastal watersheds, coastal waters are increasingly exposed to larger nitrogen loads that lead to unfavorable effects.

Extensiveness and scale of eutrophication It has been difﬁcult to assess the degree to which eutrophication affects coastal environments,

311

because many of these environments are naturally quite nutrient-rich, because a given level of enrichment may produce different effects in different environments, and because the necessary information may not be available for many coastal environments. In addition, as noted above, the effects of eutrophication range from minor to major, depending on the inputs and the receiving system. Any experience of traveling to coastal environments around the world, however, provides evidence that there are few coastal water bodies that seem unaltered in some fashion by eutrophication. One attempt to more quantitatively assess the extent of eutrophication was based on an expert review of information on a limited number of coastal environments (Bricker et al. 1999). Based on information collected on 139 sites within the USA, the experts assessed the condition of the estuaries, ranking the sites from pristine to highly altered by eutrophication. Of the 139 estuaries, 44 showed “high”, and 36 showed “moderate”, symptoms of eutrophication. Thus, 58% of the sites examined evidenced moderate to severe eutrophication. In addition, there were grounds to predict that, with no remediation of the nutrient inputs, by 2020, two-thirds of the sites would be sufﬁciently eutrophic as to impair human uses. Even within the limitations of available information, it was evident that eutrophication was widespread, and increasing, into the 21st century. Another effort to assess worldwide trends in eutrophication focused on the time courses of oxygen content in coastal waters. Diaz (2001) reviewed evidence on reported coastal hypoxia in sites around the world and found that hypoxic conditions were worsening in 75% of the coastal waters cited.42 It seems safe to conclude that most coastal waters are exposed to some degree of eutrophication, and that in most of these cases conditions are worsening. From the examples reviewed earlier in this chapter, it is evident that many groundwaters, streams, and rivers that deliver fresh water to 42

Low oxygen conditions were improving in 3% of the sites, were more or less stable in 26% of the cases, were increasing gradually in 52% of the sites, and were rapidly increasing in 23% of the cases (Diaz 2001).

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coastal environments contain increasing concentrations of nutrients.43 For the most part, the resulting eutrophication—via algal blooms, hypoxia, ﬁsh kills, and so on—impacts local sites. The area and number of sites so affected are increasing, and few parts of the world’s coastlines seem immune to eutrophication. Thus, although nutrient enrichment takes place at a local scale, the proliferation of affected sites around the world’s coasts is converting the local into a global phenomenum, but on a piecemeal basis. Certain other aspects of eutrophication, for instance the release of nitrous oxide, are likely to be having global-scale atmospheric effects.

Deﬁning standard values for water quality is an essential ﬁrst step before any corrective efforts are planned. Environmental managers have discussed many possible standards or indicators that might be used to quantify degree of eutrophication, including nutrient concentrations, producer biomass or production, and oxygen content of water, among others. Obviously, it is important to know what the concentrations of limiting nutrients are in the estuary in question. To deﬁne standards for nutrients is a challenge, however. What is a very low concentration for one producer species may be high for another species found in that same environment. What may be a high concentration for a coral reef (1–5 µM nitrate, for instance), may be low for a temperate latitude estuary. We could ﬁnd high nutrient concentrations in estuaries with turbid water, where producers have insufﬁcient light to grow and take up nutrients.

We could also ﬁnd low nutrients in estuaries where the producers do have enough light and hence take up nutrients. In addition, the rate of water turnover within the water body might be slow or fast, and water columns may mix or not, so that nutrient content may be inﬂuenced by ﬂushing and mixing as well as by land-derived nutrient inputs. Further, nutrients in water may exist in organic as well as inorganic forms, and can be in organic particles or organisms, all of which may be to some degree interconvertible in complex ways. The interpretation of nutrient concentrations therefore requires considerable ancillary information about the speciﬁc water body in question. Since eutrophication might be deﬁned in terms of production of organic matter, a second approach to set standards might be to classify estuaries as oligotrophic, mesotrophic, eutrophic, or hypereutrophic if their organic carbon supply were < 100, 100–300, 300–500, or > 500 g carbon m−2 yr−2, respectively (Nixon 1995). These categories may be too broad to be useful for management purposes, and the production rates might be affected by the same turbidity and ﬂushing factors as nutrient concentrations. In some estuaries production is carried out by macroalgae, seagrasses, and coral, as well as by phytoplankton, so that assessment of eutrophication level is a more daunting task.44 Similarly, standards for oxygen content have been proposed. Many studies have used a threshold of 2 mg oxygen l−1 to classify a water body as hypoxic. The US Environmental Protection Agency, for instance, recently recommended that marine waters from Cape Cod to Cape Hatteras be considered hypoxic if mean dissolved oxygen were lower than 4.8 mg O2 l−1, with a minumum for individual values of 2.3 mg oxygen l−1. Exposure to concentrations below 4 mg

43

44

Managing eutrophication Defining standards for management

There are exceptions, of course. Nutrient content in the Thames (Mann 1972), certain Baltic rivers (Iversen et al. 1998), and elsewhere have decreased during recent decades, owing to concerted management efforts, and there have been accompanying improvements in water quality. Nonetheless, in the majority of sites we may expect increasing symptoms of eutrophication in coastal waters, and, more than this, that conditions have worsened across recent decades.

There is another feature that adds complexity to the interpretation of nutrient data. Phytoplankton drift (largely) passively with water masses, and hence the relevant nutrient concentration is that of the water mass holding the cells. In contrast, macrophytes attached to the bottom have water parcels move past them, and may acquire nutrients from different water masses. Flow rates, as well as concentrations of nutrients are relevant to the regimes that set nutrient supply for macrophytes.

EUTROPHICATION

oxygen l−1 for some prolonged period will have some biological effects, ranging from lethal to sublethal, such as altered behavior, growth rates, and reproduction (Gray et al. 2002). Oxygen concentrations, moreover, are among the most labile of all measurements of water quality. Oxygen concentrations change with the passage of a cloud, hour of the day, season, and from one place to another, so that estimation of oxygen regimes in a water body, and assessment of duration of exposure, are not straightforward.45 It is therefore hard to select what variables can best furnish standard benchmarks to assess the condition of a water body. Measurements of nutrients, production, and oxygen, all may provide useful information, but much ancillary and quite site-speciﬁc information is also needed to interpret the local situation. Comparisons to set uniform standards therefore appears to be an uncertain approach. In part as a response to the dilemma of in situ standards, environmental management agencies have worked hard to establish more comprehensive ways to assess inputs and susceptibility of aquatic environments. Probably the most developed approach is the assessment of total maximum daily loads (TMDLs) of nutrients, organic matter, sediments, and other inputs. It is difﬁcult, however, to convincingly relate loads entering estuaries directly to ambient concentrations. For example, the concentrations of dissolved inorganic nitrogen in the New York, Delaware, Chesapeake, and San Francisco Bays are more than twice the concentrations found in Boston Harbor, even though their nitrogen loading rates are considerably lower (Kelly 1997; Tucker et al. 1999). Water exchanges with deeper waters are more active in Boston Harbor, and the ﬂushing lowers nutrient concentrations (Giblin 45

Another difﬁculty with oxygen as a standard is that tolerance to a given hypoxic episode differs among different organisms (Gray et al. 2002). Fish are less tolerant of a given low oxygen concentration and duration of exposure than crustaceans and echinoderms, and the latter are less tolerant than annelids. Mollusks are quite tolerant, in part because they can shut down their shells and engage in anaerobic metabolism. On the other hand, ﬁsh, though physiologically more sensitive, can move away from oxygen-poor waters, while shellﬁsh are far less mobile. This raises the issue of which organisms should be used as sentinels to assess conditions.

313

et al. 1997). Thus, any assessment of loads and standards needs to include detailed descriptions of water exchanges, as well as other complicating factors. To capture the action of complex factors, TMDLs are cast at the level of a speciﬁc coupled watershed/estuary system, and require the identiﬁcation of the pollutant of interest, its sources, its assimilation within the system, the effects on diverse critical system components, knowledge about how inputs alter water quality, and much more (National Research Council 2000). Erection of such TMDLs therefore requires much information about what the inputs are, what their effects may be, and the consequences for receiving environments. It is almost the rule that there is insufﬁcient empirical information for these requirements. Almost invariably, TMDLs are therefore based on speculative extrapolations, assumptions, and “best estimates”. Although TMDLs are conceptually better than simple standards, and are highly desirable features from a juridical and management point of view, setting convincing TMDLs is a daunting task, often requiring leaps of faith, as well as a full complement of the standards that we were concerned with earlier. The resulting TMDLs may provide a degree of comfort in that they set conservative targets to manage toward. Their application and interpretation, however, should be tempered in view of the uncertainty of the criteria used. Restoration of water quality As already mentioned, the major external inputs of nitrogen to a watershed/estuary system are atmospheric deposition to the watershed and to the open water; release of fertilizer used on the watershed surface; and disposal of waste water. As we have seen, nitrogen from these three sources has different fates within the watershed/estuary system. Quantiﬁcation of the different pathways is an essential ﬁrst step to remediation efforts. If, for example, atmospheric deposition is the major input, remediation will require large-scale regional reduction of emissions up-wind. If waste water is the major nitrogen source, action to control local, within-watershed disposal of waste water

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CHAPTER 12

Table 12.2 Estimates of the different inputs and outputs of nitrogen (N) compounds in Narragansett Bay. Adapted from Nixon et al. (1995). Dissolved inorganic N Inputs Atmospheric deposition to bay Rivers and streams to bay a Urban runoff b Sewage treatment plant effluent

24 290 17 142

Total inputs

473 – – – –

Outputs Export to deeper watersc Denitrification Shellfish harvest Burial in sediments

Dissolved organic N

5.6 96 18 30

Particulate N

Totals

– 12 2 10

30 400 37 183

150

24

650

– – – –

– – – –

85–170 3–4 88–173 44–97

aThe N carried by streams and rivers includes N brought into the watershed as atmospheric deposition, fertilizer

leachates, and wastewater disposal in areas upstream from the estuary. b This N is contributed by atmospheric deposition on urban impervious surfaces. cThis term is a result of transport of fresh water through the bay. It can be thought of as a net exchange, because there are relatively large N exchanges brought in and out of the bay during tidal seawater movement, which entrain N introduced by the fresh water.

through improved septic systems or sewage treatment plants would be the better alternate option. Use of watershed-level analyses, however difﬁcult, seems the best approach to restore water quality affected by eutrophication. To devise effective actions designed to lower eutrophication of estuaries and other coastal waters, we therefore need a whole-system level accounting of the major inputs contributing nitrogen. One such example is the nitrogen budget for Narragansett Bay (Table 12.2). The entries are approximate, but point out that most of the inputs were inorganic nitrogen, and that sewage treatment plant efﬂuent plus river ﬂow carrying nitrogen from upstream (mostly originating from waste water and fertilizer nitrogen) were the major sources of external nitrogen. In many places dominated by urban development, as we saw in the instance of Waquoit Bay (see Table 12.1), waste water was the major vehicle contributing nutrients to receiving coastal waters. In watersheds where agriculture dominates land use, fertilizers contribute the major portion of the added nitrogen; but in any case, whether by fertilizers

or by waste water, land-derived nitrogen is most often the major source to coastal waters, for instance to the Baltic Sea (Struck et al. 2000). Budgets such as that of Table 12.2 are useful in many ways, not the least being that they identify the relative magnitude of various sources. Once we have identiﬁed the major sources, we can more conﬁdently address and target those sources that promise greater return in environmental quality for the effort. The box opposite includes some examples—mainly with easily identiﬁed point sources of waste water—that show the level of success at recovery of receiving water quality that may be achieved by management actions.46 The examples listed in the box show that recovery of coastal environments eutrophied as a result of point-source wastewater inputs is possible. The recovery to be expected ranges from modest in certain respects to quite good in 46

There are certain coastal areas where natural processes (estuarine circulation, upwelling, and so on) are involved in the delivery of the bulk of the nutrients (Mackas & Harrison 1997); in such places reduction of sewage efﬂuent and fertilizer use might not yield detectable results.

EUTROPHICATION

Examples of major efforts at management of eutrophication New York City

The several sewage treatment plants discharging wastewater efﬂuent from the New York metropolitan area into the lower Hudson River were upgraded to secondary treatment between 1971 and 1991 when the upgrade of the largest plant was completed. During the 1990s, the percent saturation of oxygen in the lower Hudson was more than double the values recorded during 1923–1970 (Brosnan & O’Shea 1996). Evidence of sewage contamination, mainly from the New York City region is still found in sediments throughout the Long Island Sound region. Certain sewage-related contaminants, such as mercury, have decreased, but other sources from elsewhere in the contributing watersheds are still adding contaminants (Brink et al. 2000; Varekamp et al. 2000). Fraser River estuary

Primary-treated storm water, domestic, and industrial waste water were discharged between 1962 and 1988 from a sewage treatment plant directly into an intertidal ditch in the Fraser River estuary (Arvai et al. 2002). The plant served about 500,000 people in the Vancouver, BC, area. By 1974, an area of several hectares devoid of benthic organisms had developed on the mud- and sandﬂats near the discharge outfall. In response to concerns about the loss of habitat, sewage was redirected, and has been discharged since 1988 via a subtidal outfall at 100 m in depth in the Strait of Georgia. The intertidal outfall is used only for emergencies, and, on average, discharges less than 0.01% relative to the discharge through the subtidal outfall. Measurements in the zone near the intertidal outfall done before and after the 1988 diversion of the sewage showed clear improvements. The maximum mass of algae was lowered by 4–5-fold, and oxygen concentrations nearly

315

doubled. The sediment became far less organic and ooze-like, and supported diatoms rather than blue-green bacteria. The abundance of benthic invertebrates such as amphipods increased from none to hundreds to thousands per square meter. The mercury content of the sediment dropped to a fourth of the pre-1988 values. Kaneohe Bay

Treated urban sewage was discharged into Kaneohe Bay since the mid-20th century, and damage to coral reefs within the bay became apparent by the 1970s (Laws & Allen 1996). Human populations and land clearing have increased in the watershed, and owing to public pressure, the ﬂow of waste water was permanently diverted to deeper water in 1977–1978. Measurements taken during 1989– 1992 revealed that nutrient concentrations had become considerably lower (Fig. 12.23). Phytoplankton abundance was variable across the bay, but on average declined by only 35– 40%.47 Widespread recovery of the corals had taken place by 1983 (Maragos et al. 1985). Dictyosphaeria cavernosa, a macroalga that had proliferated during the period of sewage release, diminished in cover after the diversion, particularly in the southern areas most affected by the outfall. In these areas, macroalgal growth rates were low enough so that grazer pressure could control biomass48 of D. cavernosa (Stimson et al. 2001). In the interim, Kaneohe Bay was colonized by at least two 47

This seems a modest decline, given the drop in inorganic nutrients. Note, however, that even during the time when urban waste water entered the bay, ambient concentrations of nutrients (Fig. 12.23) were quite low compared to those in most temperate estuaries. In addition, there has been an unexplained increase in particulate carbon and nitrogen in the bay (Laws & Allen 1996), which might be supporting producers through the regeneration of nutrients. There remains much to be learned about Kaneohe Bay and reefs in general. 48 This is the same phenomenon we encountered in the case of macroalgal/grazer controls in the Waquoit Bay case history (see Fig. 12.9).

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species of alien macroalgae, which, as it turned out, were preferred by the grazers to the native D. cavernosa. The availability of more palatable algae may have relieved grazer pressure on the D. cavernosa population in certain other areas of the bay, and resulted in patchy but noticeable rich crops of the latter macroalga during recent years (Stimson et al. 2001). This example illustrates the complicated interactions of bottomup and top-down controls that may very well be a common feature in coastal environments. Tampa Bay

Increasing population density, urbanization, and industrial development led to nitrogen enrichment of the waters of Tampa Bay during the mid-20th century. Phytoplankton chlorophyll proliferated through the 1970s, and by 1982 the area of seagrass meadows had decreased by 72% compared to earlier in the century. Efforts to lower nitrogen loads, largely through interception of nitrogen in wastewater efﬂuent, began in 1980. Chlorophyll concentrations decreased and by 1985 reached goals set by the managers. Recovery of seagrass meadows has lagged, with an estimate of perhaps 25 years for recovery (Howarth et al. 2002, using data from several sources). Boston Harbor49

Inadequately treated sewage efﬂuent was discharged during many decades from the Deer Island and Nut Island treatment plants into Boston Harbor through two outfalls. The sludge collected at these primary treatment plants was also discharged into Boston Harbor after digestion and disinfection. Untreated road runoff and storm waters were also discharged through other pipes called “combined sewer overﬂows”. The contamination of Boston Harbor was a perennial concern for Bostonians, but little was 49

Information for this section is from http://www.mwra.state. ma.us/harbor/html/soh2002_42.htm, and Bothner et al. (1998).

done about it. Then, during the 1988 presidential campaign, the Republican candidate used Boston Harbor, the most polluted harbor in the country by some measures, as the background in a television commercial chastising the Democratic candidate, who was advocating environmental issues, and happened to be the Governor of Massachussetts. After this, politicians mobilized by court orders planned a complete and ambitious redesign of the wastewater disposal system. The Nut Island plant was closed, and the ﬂow was redirected by a tunnel dug beneath the Harbor to Deer Island (completed 1998–2000), where a completely redesigned treatment plant was constructed. This plant had many technical advances, including complete secondary treatment and denitrifying digestors to lower dissolved nitrogen in efﬂuent. By September 2000 the ﬁnal step was completed, when a tunnel 5 m wide and 14 km long, laboriously bored through hard rock beneath the sea, was opened. The efﬂuent from the entire Boston metropolitan area, treated at the new Deer Island plant, and now containing much lowered nutrient concentrations, began to be discharged into the deeper waters of Massachusetts Bay, beyond Boston Harbor, through a system of diffusers. Sewage sludge from the digestors is now converted and sold as fertilizer. By 2008, ﬂow through combined sewer overﬂows into Boston Harbor should be lowered by 88%, and treated. These developments required enormous political and economic commitment, and overcame strong opposition and institutional inertia. Such efforts are not easily repeated in many places, and are not completed as yet. To assess the effects of the massive redirection of waste water on water quality, there is an extensive monitoring effort in place in Boston Harbor and in Massachusetts Bay. In general, water quality has improved signiﬁcantly. Ammonium concentrations in the water column of Boston Harbor decreased by more than 80% after the opening of the new plant at Deer Island, even before the new tun-

EUTROPHICATION

nel was in use. Mean phytoplankton chlorophyll in the water column of Boston Harbor dropped by 46% after the opening of the new disposal system. Boston Harbor is reasonably well mixed and ﬂushed, and oxygen concentrations were acceptable (4.9–8.1 mg O2 l−1) but unchanged by the new disposal plan. The waters of the outer harbor now meet health standards for Enterococcus bacteria for swimming much of the time (combined sewer overﬂows, however, still add pathogens after storms, and there are still health advisories against swimming in the inner harbor and mouths of rivers). Recovery is also evident in the sediments of Boston Harbor. After the diversion of efﬂuent away from the harbor, by 1998 stable isotopic signatures in surface sediment samples had moved from values similar to those of wastewater sludge toward values resembling those of unpolluted marine sediments (Tucker et al. 1999). Sediments have become more oxygenated, concentrations of heavy metals are lower, and bottom-dwelling animals are about ﬁve-fold more abundant. Animal activities have increased microbial metabolism, which has lowered the organic matter content accumulated on sediments across earlier decades (Tucker et al. 1999). Flounder have

others. A pervasive problem that remains is that addressing major point sources (sewage treatment plants, for instance) is the only politically, economically, and technically feasible measure. It is much harder to address, and bring action to bear, on the many additional non-point nutrient inputs that are associated with urban sprawl or agricultural land uses. To address these other sources we are going to need to apply what have been termed BMPs (best management practices) to control fertilizer sources, for example.50 Such

317

50% lower incidence of liver disease, mussels have 50% lower polycyclic aromatic hydrocarbon (PAH) concentrations. Fish and shellﬁsh caught in the harbor easily meet health guidelines for consumption, including lobsters, whose meats are acceptable, but their hepatopancreas contain polychlorinated biphenyl (PCB) concentrations above health guidelines. In Massachusetts Bay it has been difﬁcult to detect the effects of the release of the treated efﬂuents via the diffusers. Nutrient concentrations are not manifestly increased. Chlorophyll concentrations vary greatly over the diffusers, and pre- and post-differences are not evident. No impact on the bottom-dwelling animals has been recorded as yet. The Boston Harbor example illustrates several important points. First, though extremely costly and difﬁcult, it is possible to take measures to lower the effects of eutrophication. Second, in general, there are manifold sources of nutrients, making it far more difﬁcult to manage inputs. Third, improved conditions are clearly perceivable after the diversion of the waste water, but improvement of water quality and the environment takes a long time, and some parts of the affected environment and species improve at different paces than others.

practices vary in effectiveness and acceptance across the coastal lands of the world. Lowering atmospheric sources of nutrients will require large-scale emission reduction in the industrial and urban areas of the world, as well vehicular sources—a major political challenge,51 in particular in developing nations, but still a major issue in industrial economies too. Control of such nonpoint nutrient sources is a far more challenging problem than merely redirecting a sewage outfall pipe or improving the treatment of efﬂuent. It is clear, however, that without public will,

50

The National Research Council (2000) reviews some of these BMPs; for example conversion to agricultural practices that avoid tilling soil after harvest lowered nitrogen loads from agricultural ﬁelds by more than 50%; controlling gully erosion also lowered nitrogen export by half.

51

Requirements for lowering contributions of nutrients to coastal waters from septic systems and diffuse urban discharges are reviewed by the National Research Council (2000).

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0.7

NUTRIENTS

Concentration, 1989–1992 (µM)

0.6

NO3 NH4 PO4

0.5 0.4 0.3 0.2 0.1 0 0

0.2

Concentration, 1989–1992 (mg m–3)

1.4

0.8

0.4 0.6 Concentration, 1978–1979 (µM) CHLOROPHYLL

1.2

Chlorophyll a

1.0 0.8 0.6 0.4 0.2 0 0

0.2

0.4

0.6 0.8 1.0 Concentration, 1978 –1979 (mg m–3)

1.2

1.4

Figure 12.23 Comparison of nutrient (top) and phytoplankton (bottom) concentrations before (1978–1979) and after (1989–1992) the diversion of wastewater efﬂuent away from Kaneohe Bay, Hawaii. Dashed lines denote 1 : 1. Data from Laws and Allen (1996).

and political effort, we will see increasing eutrophication of the remaining coastal waters of the world.

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Sournia, A. 1995. Red-tide and toxic marine phytoplankton of the world ocean: An enquiry into biodiversity. Pp. 103–112 in Lassus, P., G. Arzul, E. Gerard, P. Gentien, and C. Marcaillou (eds). Harmful Algal Blooms. Lavoisier/ Intercept, Paris. Steckis, R. A., I. C. Potter, and R. C. J. Lenanton. 1995. The commercial ﬁsheries in three southwestern Australian estuaries exposed to different degrees of eutrophication. Pp. 189–204 in McComb, A. J. (ed.). Eutrophic Shallow Estuaries and Lagoons. CRC Press, Boca Raton, FL, 240 pp. Stimson, J., S. T. Larned, and E. Conklin. 2001. Effects of herbivory, nutrient levels, and introduced algae on the distribution and abundance of the invasive macroalga Dictyospaeria cavernosa in Kanehoe Bay, Hawaii. Coral Reefs 19:343–357. Strahler, A. N. 1966. A Geologist’s View of Cape Cod. Doubleday, Orleans, MA, 115 pp. Struck, U., K.-C. Emeis, M. Voss, C. Christiansen, and H. Kunzendorf. 2000. Records of southern and central Baltic eutrophication in δ13C and δ15N of sedimentary organic matter. Mar. Geol. 164:157–171. Tam, N. F. Y. 1998. Effects of wastewater discharge on microbial populations and enzyme activities in mangrove soils. Environ. Pollut. 102:233–242. Thoreau, H. D. 1984. Cape Cod. Parnassus Imprints, Orleans, MA, 319 pp (a reprint of the 1864 original). Tomasky, G., and 5 others. 1999. Nutrient limitation of phytoplankton growth in Waquoit Bay, Massachusetts, USA: A nutrient enrichment study. Aquat. Ecol. 33:147–155. Torgersen, T., E. DeAngelo, and J. O’Donell. 1997. Calculations of horizontal mixing rates using 222Rn and the controls on hypoxia in western Long Island Sound, 1991. Estuaries 20:328–345. Tucker, J., N. Sheats, A. E. Giblin, C. S. Hopkinson, and J. P. Montoya. 1999. Using stable isotopes to trace sewagederived material through Boston Harbor and Massachusetts Bay. Mar. Environ. Res. 48:353–375. Turner, R. E. 2002. Element ratios and aquatic food webs. Estuaries 25:694–703. Udy, J. W., W. C. Dennison, W. J. L. Long, and L. J. McKenzie. 1999. Responses of seagrass to nutrients in the Great Barrier Reef, Australia. Mar. Ecol. Prog. Ser. 185: 257–271. Usup, G., and R. V. Azanza. 1998. Physiology and bloom dynamics of of the tropical dinoﬂagellate Pyrodinium bahamense. Pp. 81–94 in Anderson, D. M., A. D. Cembella, and G. M. Hallegraeff (eds). Physiological Ecology of Harmful Algal Blooms, Vol. G41. Springer-Verlag, Berlin. Valiela, I. 1995. Marine Ecological Processes, 2nd edn. Springer-Verlag New York, 628 pp. Valiela, I., and J. L. Bowen. 2002. Nitrogen sources to watersheds and estuaries: Role of land cover mosaics and losses within watersheds. Environ. Pollut. 118:239–248.

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Valiela, I., and M. L. Cole. 2002. Comparative evidence that salt marshes and mangroves may protect seagrass meadows from land-derived nitrogen loads. Ecosystems 5:92–102. Valiela, I., and 11 others. 1992. Couplings of watersheds and coastal waters: Sources and consequences of nutrient enrichment in Waquoit Bay, Massachusetts. Estuaries 15:443–457. Valiela, I., and 7 others. 1997. Nitrogen loading from coastal watersheds to receiving estuaries: New method and application. Ecol. Appl. 7:358–380. Valiela, I., and 5 others. 2000a. Operationalizing sustainability: Management and risk assessment of land-derived nitrogen loads to estuaries. Ecol. Appl. 10:1006–1023. Valiela, I., and 5 others. 2000b. Role of salt marshes as part of coastal landscapes. Pp. 23–38 in Weinstein, M. P., and D. A. Kreeger (eds). Concepts and Controversies in Tidal Marsh Ecology. Klewer Academic Publishers, Dordrecht, the Netherlands. Valiela, I., and 7 others. 2001. Following up on a margalevian concept: Interactions and exchanges among adjacent parcels of coastal landscapes. Sci. Mar. 65 (Suppl.):215–229. Valiela, I., D. Rutecki, and S. Fox. 2004. Salt marshes: Biological controls of food webs in a diminishing environment. J. Exp. Mar. Biol. Ecol. 300:131–159. Van Raalte, C., I. Valiela, E. J. Carpenter, and J. M. Teal. 1974. Nitrogen ﬁxation: Presence in salt marshes and inhibition by the additions of combined nitrogen. Estuar. Coast. Mar. Sci. 2:301–305. Varekamp, J. C., M. R. B. T. Brink, E. L. Mecray, and B. Kreulen. 2000. Mercury in Long Island Sound sediments. J. Coast. Res. 16:613–626. Vogelbein, W. K., and 6 others. 2002. Pﬁesteria shumwayae kills ﬁsh by micropredation not exotoxin secretion. Nature 418:967–970. Vuilleumeir, M. 1970. Indians on Olde Cape Cod. W. S. Sullwold Publications, Taunton, MA, 96 pp. Ware, D. M., and R. E. Thompson. 2005. Bottom-up ecosystem trophic dynamics determine ﬁsh production in the Northeast Paciﬁc. Science 308:1280–1284. Wickham, J. D., and 5 others. 2002. Geographic targeting of increases in nutrient export due to future urbanization. Ecol. Appl. 12:93–106. Williams, M. R., and J. M. Melack. 1997. Solute export from forested and partially deforested catchments in the central Amazon. Biogeochemistry 38:67–102. Wyatt, T. 1995. Global spreading, time series, models and monitoring. Pp. 755–764 in Lassus, P., G. Arzul, E. Erard, P. Gentien, and C. Marcaillou (eds). Harmful Algal Blooms, Vol. 19. Lavoisier/Intercept, Paris. Zingone, A., and H. O. Enevoldsen. 2000. The diversity of harmful algal blooms: A challenge for science and management. Ocean Coast. Manage. 43:725–748.

Chapter 13 Other agents of coastal change

A whale entangled in ﬁshing gear.

Humans have altered coastal environments in a number of additional and diverse ways, including dumping of debris, release of heated water, sound production, release of radioactive materials, and spread of pathogens. All of these activities have received their share of press attention, so there is public awareness of the topics. This chapter reviews published facts as to their impact on coastal organisms and environments.

Dumping of debris Occurrence of debris There is no part of the ocean that does not show some bit of human-derived debris. The density of discarded garbage and other items can reach prodigious numbers, and debris of many kinds

325

OTHER AGENTS OF COASTAL CHANGE

Table 13.1 Abundance of debris or litter on or in a number of marine environments; sites are selected from many recent reports. Data are for items on sea ﬂoor, unless otherwise indicated. Location

Abundance (items km−2)

Dominant type of debris or litter

Source

Sargasso Sea (on surface)

50–12,000

Plastic pellets

Carpenter & Smith 1972

Off Hokkaido

0–9,894,044

Plastic, styrofoam pellets

Ogi & Fukumoto 2000

North Pacific gyre (on surface)

334,270

Plastic fragments, plastic bags, and filaments

Moore et al. 2001

Off Kodiak Island

4.5–25

Derelict fishing gear on bottom

Hess et al. 1999

Northwest Hawaii

27–62

Derelict fishing gear on bottom

Donohue et al. 2001

Plastic, styrofoam fragments

Moore et al. 2002

California Current (on surface) Baltic Sea

126

Plastic fragments, plastic bags, plastic and glass bottles, metal and glass fragments, fishing gear

Galgani et al. 2000

North Sea

156

Plastic fragments, plastic bags, plastic and glass bottles, metal and glass fragments, fishing gear

Galgani et al. 2000

Celtic Sea

528

Plastic fragments, plastic bags, plastic and glass bottles, metal and glass fragments, fishing gear

Galgani et al. 2000

Bay of Biscay

142

Plastic fragments, plastic bags, plastic and glass bottles, metal and glass fragments, fishing gear

Galgani et al. 2000

Northwest Mediterranean

1,935

Plastic fragments, plastic bags, plastic and glass bottles, metal and glass fragments, fishing gear

Galgani et al. 2000

Adriatic Sea

378

Plastic fragments, plastic bags, plastic and glass bottles, metal and glass fragments, fishing gear

Galgani et al. 2000

Rio de La Plata estuary

15–370

Plastic fragments, plastic bags, cans

Acha et al. 2003

Plastic debris

Cunningham & Wilson 2003

Plastic, glass, metal fragments

Nagelkerken et al. 2001

Plastic, wood, glass fragments

Debrot et al. 1999

Plastic and styrofoam fragments

Ribic 1998

Sydney Harbor (on beaches) Curacao (off beaches)

103,752,000 items m 100 items m

−2

−2 −1

Curacao (on beaches)

19–253 items m of beachfront

Long Beach Island, NJ (on beaches)

0.5–6.5 items m−1 of beachfront

can be found on the sea surface, the sea ﬂoor, and on shorelines (Table 13.1). The type of garbage depends in part on the human activity nearby. Litter on New Jersey beaches, near the New York metropolitan area, included land-derived items (syringes, condoms, cans, six pack rings, balloons, straws, tampon applicators, bottles, plastic bags,

and cotton swabs), and marine-derived items (nets, lobster pots, ﬁshing lines, ﬂoats, buoys, and cruise ship logo items) (Ribic 1998). On Sydney Harbor beaches, land-derived debris made up about 28% of the items, and marine-derived materials only 4% (Cunningham & Wilson 2003). On coasts farther away from urban areas, such as the islands

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CHAPTER 13

Total debris per 500 m

3,250

2,250

1,250

250 1991

1992

1993

1994

1995

Year

northwest of Hawaii, marine ﬁshing gear made up most of the debris (trawl netting, gillnets, seines, ﬁshing lines, and other fragments) (Donohue et al. 2001). Plastic pellets, however, can be found in large densities even in remote areas such as Tonga, Rarotonga, and Fiji (Gregory 1999) or the Antarctic islands (Walker et al. 1997), so that remoteness does not guarantee freedom from our litter. It is evident that humans seem quite careless about what they do with their solid waste matter and ﬁshing gear. The speciﬁc sources of the large amounts of debris found in coastal environments are of course quite varied, and difﬁcult to quantify (Derraik 2002). Merchant ships may dump 639,000 plastic containers per day across the world. Smaller boats may be responsible for more than half the rubbish dumped into waters off the USA. The world’s ﬁshing ﬂeet may have dumped 135,400 tons of ﬁshing gear and 23,600 tons of other materials into the sea during 1975. Much litter— mainly packaging—is brought to coastal waters via rivers and outfalls that drain urban areas. Large amounts of litter are left behind by beach users. In Sydney Harbor, for example, about 63% of litter on beaches came via storm water that carried materials dumped upstream, about 20% was dropped by beachgoers, and ﬁshermen dumped about 7% and shipping about 5% (Cunningham & Wilson 2003). Debris sources are therefore widely distributed among different people and activities, and sources differ from place to place.

1996

Figure 13.1 Debris counted across the period 1991–1996 on Island Beach State Park, NJ. From Ribic (1998).

By far, manufactured plastic debris dominates the litter loads to the marine environments,1 followed by smaller amounts of glass and metal items. A review of litter types across many parts of the world’s coasts showed that plastics made up 32 to > 90% of the debris, averaging 68% (Derraik 2002). Of course, human garbage includes much other material, such as paper, food wastes, wood, and so on. These are not often seen in litter surveys because they are degraded, largely by biological action, and hence have relatively short lifespans in aquatic environments. Although time course data are scarce, abundance of litter in coastal environments may have increased during the 1990s (Fig. 13.1). Ogi and Fukumoto (2000) aver that density of marine plastic debris increased, worldwide, from several thousands of items per square kilometer during the 1970s to about 500,000 items km−2 in the 1990s, and suggest that during the 1990s there was an order of magnitude increase per decade, although they do not give the sources for these assertions. It is certain that there are continuing sources of litter: Panamanian beaches cleared of litter

1

Plastics are organic polymers synthesized industrially from petroleum, and their versatility has been exploited in innumerable products; their applications multiplied enormously during the last half of the 20th century. Plastics are cheap, resist decay, and have relatively light density. These characteristics favor their ready disposal, survival in natural environments once dumped, and dispersal by ﬂoating (Derraik 2002).

OTHER AGENTS OF COASTAL CHANGE

regained about half their original density of litter after 3 months (Garrity & Levings 1993). In any case, it is evident that there are numerous human-derived litter objects strewn about coastal environments. Given our proclivity to use plastic containers and wrapping, and our apparent inability to securely dispose of, or recycle, the materials, there will be increased visible litter in coastal environments in the future. The abrasive action of water motion, and other factors, act to break down the large visible litter into microscopic fragments that appear to be widely dispersed (Thompson et al. 2004). Microplastics were found in the sediments of 40 out of 47 beaches sampled throughout Great Britain. Plankton samples collected during the 1980s and 1990s held more than twice the microplastic fragments than during the previous two decades. Microplastic fragments are widespread, and may adsorb and release potentially damaging chemicals. Amphipods, lugworms, and barnacles all ingested microplastics offered to them as food (Thompson et al. 2004); the consequences of such ingestion are unknown. Distribution of litter The spatial distribution of marine litter is heterogenous in the extreme. Certainly, vicinity of humans—urban areas, shipping lanes, areas of intense ﬁshing—is a factor (Table 13.1), but there are also important effects owing to hydrodynamics and wind (Galgani et al. 2000). A representative example of the inﬂuences of the various factors is the distribution of litter from the large cities of Buenos Aires, La Plata, and Montevideo, on the bottom and shores of the estuary of the Rio de la Plata (Fig. 13.2 bottom). As river water moves down-estuary, it meets sea water, forming a marked front with a steep salinity gradient (Fig. 13.2 top and middle). As is known for other kinds of particles (Largier 1993; Sanford et al. 2001), litter carried by fresh water tends to accumulate on the estuary ﬂoor under the front (black circles in Fig. 13.2 bottom). Similar distributions were recorded for debris along the shorelines of the estuary (white squares in Fig. 13.2 bottom). Similarly, the currents that encircle

327

subtropical oceanic gyres envelop large areas of the sea where high densities of litter particles accumulate (see Table 13.1). For such reasons, beaches in islands far off from human population centers may be as full of litter as those near urban areas (Benton 1995). The examples demonstrate the effectiveness with which marine hydrodynamic processes control the distribution of litter throughout the seas (see box on p. 329). Effects of litter Litter and debris are readily apparent along the shores of the world. There is an obvious issue of aesthetics marring the enjoyment of coastal settings, but is there evidence of ecological effects of the garbage that is now so abundant on the coastlines of the world? There has been concern about ingestion of litter materials and entanglements of marine organisms. Ingestion of litter

Fish, turtles, birds, and marine mammals frequently ingest debris. Plastic spherules were found in eight of 14 ﬁsh species examined off the New England coast, and 21–30% of specimens of ﬁsh off New England and in the Bristol Channel contained plastic pellets (Derraik 2002). About 267 species have been found with litter in their guts, including 33 species of marine ﬁsh, and 86, 36, and 23%2 of all marine turtles, bird, and mammal species, respectively (Laist 1997). Between 20 and 80% of turtles examined across many coastal waters over the world held some sort of litter in their gut; a survey in the western Mediterranean found plastics, tar, paper, styrofoam, wood, seabird feathers, hooks and lines, net fragments, and other items (Tomás et al. 2002). A study of dead turtles stranded on the south coast of Brazil found that 61% of individual green turtles held debris, including plastic bags, ropes, cloth, hard plastic, styrofoam, oil, paper, and other items

2

These percentages are calculated relative to many species of marine ﬁsh, seven species of turtles, 312 species of sea birds, and 115 species of marine mammals (Laist 1997).

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CHAPTER 13

34° URUGUAY

Colonia

Punta del Este

Montevideo

15

La Plata

35°

Punta −10 −5 Piedras

30

Samboronbón Bay

36°

20

25

Latitude S

Buenos Aires

ARGENTINA 59°

Depth (m)

0

58°

0

57° Longitude W 15

10

5

20

55°

56°

25

30

10

20 0

20

40

60 80 Distance (km)

100

120

34° URUGUAY

Colonia

Latitude S

Buenos Aires

Montevideo

Punta del Este

La Plata

35°

Punta Piedras

Shore debris −2

(g 10 m ) Samboronbón Bay

36°

< 500 501–1,000 1,001–1,500

ARGENTINA 59°

58°

57° Longitude W

56°

Estuary debris (items km−2) 0–10 11–50 51–100 101–150 >150

55°

Figure 13.2 Distribution of debris and salinity in the Rio de la Plata estuary. Top: contours of near-bottom salinity. Middle: vertical view of the distribution of salinities measured along the straight line shown in the top panel. Bottom: distribution of debris on the estuary ﬂoor (black circles) and on the shores (open squares). From Acha et al. (2003).

329

OTHER AGENTS OF COASTAL CHANGE

The case of vicariant sneakers A notable example of the remarkable distances that litter may travel under the inﬂuence of currents and wind is the fate of the 80,000 athletic shoes lost overboard on May 27, 1990 in the North Paciﬁc Ocean (Fig. 13.3) (Ebbesmeyer & Ingraham 1992). A severe storm caused 21 containers to be lost from the container ship Hansa Carrier en route from Korea to the USA. Many shoes apparently were released from the containers, and ﬂoated on the sea surface, subject to wind and currents. Six months to a year later, thousands of shoes, hardly the worse for their travels, were washing ashore in North America, from

southern Oregon to the Queen Charlotte Islands in Canada. Since pairs of shoes were not tied together, matching shoes did not arrive together; beachcombers held shoeswap meets in Oregon to ﬁnd matches to their ﬁnds. Tracks of the likely paths of the shoes from their release point (Fig. 13.3), estimated by ocean surface circulation models that considered wind and currents across the Paciﬁc, gave new evidence about the variability of ocean ﬂows, as well as empirically conﬁrming much knowledge about circulation in the Paciﬁc. The appearance of athletic shoes so far from the point of the accidental dumping also demonstrated the wide dispersal of human litter across the world’s seas.

Recovery locations (1990–1991) March 26 (250) May 18 Jan/Feb (200) (100) Nov/Dec (200) Nov/Dec (200)

Alaska Current

ka

n

re St

am ia rn t lifo en Ca urr C

42°N

April 4 (150) May 9–10 (200)

as Al

t rren

u ic C

t arc

Sub

126°W

t as dc n i t H rif d Shoe spill (May 1990)

35°N

165°W

152°

63°N 139°W

126°W

56°N

113°W

49°N

Figure 13.3 Map of the North Paciﬁc, indicating the site where the May 27, 1990, shoe spill occurred, the modeled likely drift path (in grey), and some locations on the North American shore where shoes from the containers lost from the Hansa Carrier arrived over the following year. The numbers in brackets are the number of shoes found at each date. From Ebbesmeyer and Ingraham (1992).

330

CHAPTER 13

(Bugoni et al. 2001). Plastic debris is commonly found in sea birds (Spear et al. 1995). For example, plastics were present in 95% of the regurgitated gut contents of 19 great shearwaters, and 73–93% of blue petrels at Gough Island in the South Atlantic (Ryan 1987). At least 26 species of cetaceans have ingested plastic debris (Derraik 2002). These results speak to the widespread abundance of litter, as well as to the lack of discrimination in feeding by these animals.3 It has been difﬁcult to conclusively demonstrate effects of litter ingestion. Possible deleterious effects of ingestion of litter that have been suspected include lowered foraging effectiveness, nutrition, digestion efﬁciency, and survival (Azzarello & Van-Vleet 1987; Derraik 2002). Plastic particles may also leach harmful organic compounds such as polychlorinated biphenyls (PCBs) and DDE once in the gut (Mato et al. 2001). There is only anecdotal, circumstantial, or weak evidence of detrimental effects of the ingestion of litter. Guts in dead specimens of a West Indian manatee, a Blainville’s beaked whale, or a young pigmy sperm whale were ﬁlled with plastic (Derraik 2002). There were weak correlations of lower sea bird fat condition and higher levels of plastics in the gut for some species (Azzarello & Van-Vliet 1987). Other reports have concluded that there were no lethal effects of the debris, or blockage of the digestive tract, in spite of the variety and volumes of litter materials (Tomás et al. 2002). The items seemed to pass through the turtle digestive tracts. There was no demonstrable effect of amount of plastic in the gut on bird size in these ﬁeldcollected data (Ryan 1987). Experimental feeding of different amounts of plastic pellets to whitechinned petrels showed no gut obstruction or damage to the gut lining (Ryan & Jackson 1987). Ingested plastic did not seem to impair digestive efﬁciency. Other possible sublethal foraging and nutritional effects appear even more difﬁcult to demonstrate (Tomás et al. 2002). 3

Sea birds apparently simply test any particle that ﬂoats on the sea surface as possible food; 80% of ﬂoating plastic debris that washed ashore on a Dutch coast showed distinct peckmarks made by birds at sea (Cadée 2002).

Entanglements Marine organisms can become entangled in plastic packaging, synthetic ropes or lines, and, most commonly, ﬁshing gear (Derraik 2002). Entanglements have been reported for at least 135 marine species, including 34 species of ﬁsh, and 86, 16, and 86% of the sea turtle, bird, and mammal species, but in almost all instances data on the extent and effects of entanglement are insufﬁcient (Laist 1997). Perhaps 13–29% of the mortality incidences of gannets in the German Bight have been attributed to entanglements (Laist 1997). There were records of seven beached humpback whales, eight dead whales in the water, and nine entangled whales (two of which died) off Hawaii during 1972–1996 (Mazzuca et al. 1998). Off the Maritime Provinces and New England states there were 12 reports of dead whales that showed evidence of entanglement, and 29 entangled live whales in 2002 (Whittingham et al. 2003). Entanglement therefore does not necessarily entail mortality in whales, and cause and effect links are hard to ascertain. “Ghost-ﬁshing” and entangling by discarded and lost ﬁshing gear (fragments of various types of nets, long-lines, lobster and ﬁsh traps, and many other items) is a feature that has been difﬁcult to evaluate, but remains a potential and increasing threat to marine species. Reports of entanglements are rare enough to make detection of secular trends difﬁcult. It may be that mortality from this source is increasing, but evidence is sparse at best. For example, in Paciﬁc waters off Colombia, six humpbacks were found dead or entangled from 1986 to 1995, and 18 during 1996 to 2000 (Alzueta et al. 2001). For the 24 reported cases of dead whales, 10 were of natural or unknown causes, three showed signs of an encounter with a ship, one was the victim of a hunting attempt, and 10 were entangled; so that 42% of the deaths in this area may have been linked to entanglement. There was no information, however, as to the size of the humpback population in the area, so it is hard to evaluate what the impact of the reported mortality might be. Laist (1997) concluded a review of existing information by saying “. . . deﬁnitive proof of

OTHER AGENTS OF COASTAL CHANGE

population-level effects is lacking even for those species thought to be most affected by entanglement . . . for seabirds . . . toothed whales, and ﬁsh . . . entanglement appears unlikely to cause effects at a population level.” He did add that we must have an underestimated view of the effects of entanglement, since most victims are likely to pass unnoticed, and that for certain species such as endangered right whales or Hawaiian monk seals, whose population numbers are low, entanglement must be a threat.4 Transport of alien species

Marine organisms (bacteria, algae, barnacles, hydroids, bryozoans, tunicates) may settle and encrust on debris, particularly plastics (Derraik 2002). Most debris lacks colonizers (Powlik 1995; Donohue et al. 2001), but some items have them. As the litter is transported across large distances, it may offer its passengers the chance to expand their geographic range beyond the usual limits. The plentiful distribution of litter may therefore potentially add another means for invasions of new habitats by alien species. The geographic range of two species of bryozoans and a few barnacles are thought to have expanded via this means. Remedies It is apparent that human litter of one kind or another is widely spread across all the coasts of the world. Although at present we cannot compellingly assay the ecological effects of litter, we are near the level of contamination where the litter may begin to generate the effects that are suspected. Moreover, at the moment, in any coast of the world, we can clearly see amounts of litter that offend most people’s notion of what the natural world ought to look like.

4

Right whales seem particularly susceptible to entanglement (Laist 1997). Photographic records showed that 57% of right whales had rope scars around their mouths. If we accept the estimate of 350 right whales in the North Atlantic, about 20% of the population had entanglement-related scars on the peduncle alone. About 12% of the dead right whales reported between 1970 and 1989 were circumstantially attributed to entanglement with ﬁshing gear.

331

There are ways in which the proliferation of human debris in the sea can be diminished. Political action that provides incentives to reduce the use of packaging, strongly encourages recycling of materials, and fosters the development and use of degradable materials will lower the disposal of debris. International and national legislation discouraging at-sea disposal from vessels would help if enforced. In a few cases where the populations are very small, such as certain whales, efforts at rescue of entangled specimens seem worthwhile. Above all, efforts to educate the public as to the need to better dispose of solid waste— a too-little recognized but valuable resource— need to be put into action.

Thermal pollution The need for energy and fresh water for use in coastal zone settlements has in a few certain places prompted the construction of power plants and desalinization facilities. These industrial-scale enterprises use sea water as a coolant or source of water, and have effects on coastal organisms and ecosystems through the mortality of organisms entrained in the sea water taken into the plants, as well as by effects of discharge of heated water into the sea.5 In addition to added heat, discharged water has a changed chemistry and microbiology, for instance owing to the chlorination that is used to reduce settlement inside pipes within the cooling system (Karaas 1992; Saravanane et al. 1998). Entrainment losses There are quite variable effects on the organisms that are entrained in the water taken into power and desalinization plants. Losses of different groups of phytoplankton during passage through 5

Water discharged from power-generating plants is roughly 10°C warmer than receiving waters. One plant in Thailand released water that was 6.5–11°C warmer than water taken into the plant; discharge from a Chesapeake Bay nuclear power plant was 6.7°C warmer than bay water (Abbe 1987). A coal-ﬁred power station in Australia discharged heated water 7.5–10.5°C above ambient temperatures (Ainslie 1977).

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a Virginia power plant were 25–80% during summer; the losses were related more to dense populations of grazers inside discharge pipes than to increased water temperatures (Jordan et al. 1983). This was also found in Swedish power plants (Karaas 1992). For phytoplankton, the production and biomass of ﬂagellates, but not dinoﬂagellates or diatoms, were lowered during passage through a nuclear power plant on Chesapeake Bay; 65– 100% of the zooplankton, however, survived passage through the plant (Sellner et al. 1984). In contrast, about 80 kg of zooplankton were killed daily in the operation of the nuclear power plant in the Gulf of Bothnia, including roughly 42,000 ﬁsh larvae, 30,000 ﬁsh eggs, and 180,000 immature shrimp (Keskitalo & Ilus 1987). Passage through a Polish power station killed 37–80% of the summer zooplankton (Tunowski 1988). The mortality of juvenile herring-like ﬁsh entrained in water taken up in Canadian and US power plants ranged between 46 and 82% (Stokesbury & Dadswell 1989). Mortality of coastal organisms is therefore extremely variable, but can affect large numbers of organisms.6 Certainly, the number of organisms killed during entrainment is large, but we need to consider whether the losses are meaningful at the level of the populations involved. The basic issues are how much of the ambient water is forced to course through the plants, and what might be the impact of within-plant losses on the populations at large. Modeling efforts suggest that volumes of cooling water and numbers of organisms are such that effects on populations are in general minor (Cakiroglu & Yurteri 1998). For example, Saila et al. (1997) examined multiyear records of data on entrainment and mortality of three species of ﬁsh larvae, extrapolating the loss of larvae to estimate the potential reduction of the stocks that might be expected from the entrainment losses. They concluded that the impact of entrainment was unlikely to be ecologically signiﬁcant for the three species. 6

The number of entrained organisms can be effectively lowered by means of screens and sonic devices placed strategically at the entry ports to the plants (Rulifson & Dadswell 1987). For example, in a Maryland power plant, more than 80% of larger ﬁsh larvae were excluded by screens (Weisberg et al. 1987).

Effects of discharges of heated water In general, only limited areas or volumes of ambient sea water are affected by discharged water. The impact of heated water discharged by a South African nuclear power plant was restricted to less than 500 m from the release point (Jury & Bain 1989). In the case of the Chesapeake Bay nuclear power plant, the area enclosing water 2°C warmer than the ambient water was only 0.34 km2 (Abbe 1987). Effects of heated water discharged from a Gulf of Bothnia nuclear power plant on producers on the sea ﬂoor extended to about 1.8 km from the point of discharge (Keskitalo & Ilus 1987). A larger area, about 36 km2, was affected by heated water discharge from a power plant in Rhode Island (Mustard et al. 1999), but only during fall, when water was 0.8°C warmer than water farther from the point of discharge from the power plant. Heating effects therefore do not extend much beyond the vicinity of the discharge. Effects of the warmer water within the limited spaces where heating was detectable were both negative and positive. Effects that could be considered deleterious include instances such as in the Gulf of Thailand, where water temperatures within 20 m of the discharge pipe from a power plant were high enough to kill organisms within 400 m of the discharge pipe. Heated water from a nuclear power plant in the Gulf of Bothnia changed the species composition of producers (Ilus & Keskitalo 1987; Keskitalo & Ilus 1987; Snoeijs & Prentice 1989). Discharges from the Chesapeake Bay nuclear power plant (Abbe 1987) scoured sediments and changed the assemblage of benthic organisms. On the other hand, there are reports that suggest more benign effects. Increased temperature was found to have no effect on populations of copepods in the water receiving the discharges (Olson 1987). The heated discharge slightly increased the seasonal phytoplankton biomass, and prompted a less than 2-week earlier start to the spring bloom (Keskitalo 1987). Throughout, changes in nutrient supply driven by other factors had more powerful inﬂuences on plankton than the heated water. Fish and shellﬁsh grew

OTHER AGENTS OF COASTAL CHANGE

faster near the point of discharge of heated water, and the area near the discharge became a haven for wintering birds (Abbe 1987). Settlement of coral was stimulated near power station outfalls (Coles 1985). Incidentally, releases of radionuclides from nuclear power plants seem insufﬁcient to pose a threat to organisms in the discharge area (Abbe 1987). On the whole, given the current numbers of power plants and volumes of cooling water used, there seems to be at most modest and quite local effects of entrainment within power plant cooling water, or of heated water discharges on the populations of organisms within the waters receiving the discharges. If individuals entrained were to belong to rare or threatened populations, the effects of the mortality might become more serious.

Sound pollution Certain species of marine organisms—including ﬁsh and mammals—are highly attuned to the use and perception of sounds across a very wide range of frequencies (Enger 1968; Au 2000). Such species use sounds to maintain social systems, breed, hunt, escape predators, and more. Sound has been used for communication and perception by such species because it propagates far more readily in water than any other form of energy, and can be used at depth, in turbid water, or in the dark. Comparisons of noise perception and distance of detection are confused by differences in loudness in the noise sources, attenuation at the receptor, temperature of the water, and interspeciﬁc differences. In general, however, larger species emit and hear larger ranges of frequencies, and make greater use of lower frequency sounds (Au 2000; Ketten 2000). High frequency sounds do not propagate very far in water, and hence these frequencies are more frequently involved in the location of prey, to avoid obstacles, nearﬁeld social communication, and so on. Sound at lower frequencies propagates for far larger distances. Some large whale sounds may travel hundreds of kilometers.

333

The acoustic world in which ﬁsh and marine mammals live has been thoroughly altered by the cacophony of human-made sound now pervading the oceans in general and areas near humans in particular. In Kagoshima Port, Japan, a variety of noises from ferries, cruising jetfoils, high-speed ﬁshing boats, and pleasure boats, among other sources, created a broad mix of underwater sounds across a large range of wavelengths (Fujieda et al. 1998). Elsewhere in the sea, sounds are generated by ship engines, use of sonar, explosions during oil exploration, activities during oil extraction, and even maricultural practices. During the past half-century, human contribution to such underwater noise has increased, particularly in the low frequencies (Croll et al. 2001). In the North Paciﬁc, for example, ships were heard 55% of the time in long-term records of ambient low frequency sound (Curtis et al. 1999). Whales were heard 43% of the time, predominantly during migration of blue and ﬁn whales. Sound levels were more pronounced in coastal stations. Undersea sounds and fish Fish make, and are responsive to, noise across a wide range of frequencies. Only a few certain herring-like ﬁsh species can hear low frequencies: they may have evolved this ability as a device to elude foraging echo-locating dolphins or whales (Mann et al. 1998).7 Fish are also exposed to human noise-making. Noise from small boat engines lowered hearing in fathead minnows after 2 h exposures at different levels of loudness (Scholik & Yan 2002). Experimental exposure of ﬁsh to loud, repetitive, low frequency sounds associated with marine oil exploration led to extensive damage to the hearing organs of ﬁsh (McCauley et al. 2003). 7

Marine mammals such as killer whales use different repertoires of acoustic foraging, depending on their prey. In the northwest coast of the USA and Canada, there are resident pods of killer whales that feed preferentially on ﬁsh, and transient pods that preferentially fed on marine mammals. Barrett-Lennard et al. (1996) and Mann et al. (1998) proposed that the ﬁsh-eaters made more ready use of sonar echolocation because this detection mode was not as easily detected by their ﬁsh prey. Predators of mammals made less frequent, and intermittent, use of sonar echolocation, presumably because seals were better at detecting these frequencies.

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The repellent effects of certain noise frequencies have been proposed as useful tools to prevent bycatch of marine mammals in trawls, scare off seals feeding on salmon runs, manage the direction of movements of salmon smolts, and other purposes. The success of these strategies is modest, as the presumed target organisms often ignore or acclimate to the sounds (Knudsen et al. 1994; Akamatsu et al. 1996; Popper & Carlson 1998). Undersea noise and marine mammals Concerns with the issue of noise pollution have focused on low frequency sounds of human origin, and their effects on the larger baleen whales, whose biology and behavior are more closely attuned to lower frequencies (Ketten 2000). The literature on these subjects reports widely disparate results, from effects that merely distracted whales for a short time, to far more consequential effects. Studies off Hawaii concluded that noise from small boats was unlikely to affect hearing of humpback whales (Au & Green 2000). Blue and ﬁn whales continued to forage and feed off California even when an experimental, loud, low frequency source was transmitting (Croll et al. 2001). Although there were no other evident disturbances, vocalizations may have been altered in response to the experimental low frequency noise (Croll et al. 2001); low frequency sonar in fact lengthened humpback songs off California, perhaps as a compensation for low frequency interference with their communication (Miller et al. 2000). Belugas and narwhals 25–30 km away detected high frequency noise from vessels involved in ice-breaking and transport to mining centers in the coast of the Northwest Territories of Canada (Cosens & Dueck 1993). Bowhead whales responded in observable fashion 14% of the time, and beluga whales 38% of the time, to noise from aircraft overﬂights in the Beaufort Sea; the more frequent responses by belugas were in spite of their lower ability to hear low frequency noise (Patenaude et al. 2002). The responses were minor, involving slight alterations in behavior for a short time, with no evident distress. Killer

whales were to a degree displaced from areas where low frequency acoustic harassment devices were installed to prevent seals from the vicinity of salmon pens (Morton & Symonds 2002). On the other hand, there are instances where there were more serious effects presumed to be associated with low frequency noises. Military maneuvers off the Canary Islands were followed by mass strandings of Cuvier’s beaked whales across large distances along the coasts of the Canary Islands (Simmonds & Lopez-Jurado 1991). The beaked whales suffered damage to their inner ears and brains. Trials of extremely loud, low frequency sonar devices by NATO ships coincided with strandings of Cuvier’s beaked whales across 38 km of Greek island coastline (Frantzis 1998); 12 were found stranded alive, and an additional individual was found dead later. The simultaneous stranding of many individuals was unusual, as this rare species seldom strands. Naval maneuvers—presumably using powerful low and mid-frequencysonar— coincided with mass stranding of 17 beaked whales, plus dolphins, and the subsequent death of six beaked whales on the Bahamas (Balcomb & Claridge 2001). A multinational naval exercise near the Canary Islands once again led to deaths and strandings of 17 beaked whales during September 2002, events that repeated what followed after naval exercises in 1985, 1988, 1989, and 1991 (New Bedford Standard Times, Oct. 1, 2002). Causes of the death of the beaked whales are still to be demonstrated (New Bedford Standard Times, Oct. 9, 2003). As many as 200 melon-headed whales, an open ocean species, were found in shallow water of Hanalei Bay in the island of Kauai, Hawaii in early July 2004 (The Washington Post, July 11, 2004). This episode coincided with a six-ship US Navy sonar exercise some distance offshore; in addition, mass strandings of the most noise-sensitive whales off Japan have occurred near an American naval base, but were unknown in comparable areas elsewhere. Although mere coincidence cannot be ruled out, all these instances of disturbance to whales circumstantially but compellingly suggest that loud, low and mid-frequency noise drastically disturbs whales. It has been argued that the sound sensitivity of whales that tend to feed on

OTHER AGENTS OF COASTAL CHANGE

squid deep in the sea makes them especially susceptible to disturbance by low frequency noise, even if the noise is at low levels when it is received by the whales (Bowles et al. 1994). The acoustic world of marine fauna may be further altered by two kinds of recent technical developments. Sound has become a major way to study the oceans, because sea water is transparent to sound, and even slight differences in seawater temperature can affect the speed of sound. Thus, sound transmission is a delicate tool with which to investigate hydrodynamics, even across large distances, and has promise as a research approach deployed in experiments in what has been called acoustic thermometry of ocean climate (National Research Council 2000). Since some marine species use this range of frequencies to communicate and sense their environment, the research emissions could disrupt their life activity. In addition, military uses of very loud, low frequency sonar to detect and track newly developed “quiet” submarines has been in development. The charismatic nature of whales, plus the publicity from the Canary Islands, Bahamas, and Hawaii strandings, have generated public and political concerns. Legal suits held up experiments in thermography (Au 2000),8 naval operations testing the advanced sonar (Cape Cod Times, Nov. 2, 2002), and even experiments to assess the relative sensitivity of whales to new sonar technology (Cape Cod Times, Jan. 9, 2003). The sonar experiments were eventually approved under restricted conditions, such as maintaining a certain distance from the coast and so on. Other kinds of sonar-based research have also been the subject of litigation, including seaﬂoor mapping (Malakoff 2002).

8

Au et al. (1997) carried out a variety of tests to see whether the emissions from the proposed experiments would affect dolphins. If the source of sound was at 850 m, as planned, dolphins directly above the sound source would not be able to hear the sound until they dove to 400 m. For dolphins farther than 0.5 km away from the source, the loudness was lower than the sounds produced by the dolphins themselves.

335

There is no doubt that sound pollution saturates the world’s oceans. Most evidence suggests that for ﬁsh and most marine mammals, the effects of human-generated sound range from minimal to slight and temporary behavioral disturbances. For a few species, principally the generally rare beaked whales, there is circumstantial but convincing evidence that certain ranges of sound, particularly those involved in powerful sonar signals, can be lethal.

Radioactive pollution9 Many radioactive elements occur naturally in extremely low concentrations across all environments, but human activities have increased the abundance of radioactive isotopes of many elements, including 90Sr, 137Cs, 238Pu, 240Pu, and 241 Am. These and other isotopes, such as 3H, 14C, 99Tc, and 129I, have become important as tracers of marine biogeochemical processes, ocean currents, and assessors of environmental conditions. In fact, our insertion of radioactive materials into the atmosphere and oceans has been an extraordinary boon for biogeochemists, climatologists, and oceanographers, who have used the radioactive elements as tracers of movements and transformation processess in the atmosphere, land, and sea. Of course, the potential hazards of the released radionuclides have raised more public concern. Sources of radioactive materials to coastal waters Industrial or military uses of radioactive materials have been involved in most of the instances where radioactive materials were released to aquatic environments. Most of the radioactive materials have come as global-scale atmospheric fallout from nuclear bomb tests during the 1950s and 1960s. Accidents involving nuclear materials, such as the Chernobyl accident, have proven to be a major input into the Baltic and Black Seas.

9

Much of the information in this section is from the excellent review by Livingston and Povinec (2000).

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Releases during operations of nuclear power plants and other nuclear-based activities, such as nuclear reprocessing plants, have been sources of radioactive materials in the Irish and North Seas.

little if any evidence of radioactive contributions from dumping, and the ecological impact of these dumping practices has been reported as negligible (Livingston & Povinec 2000).

Global atmospheric fallout

Nuclear accidents

Most of the anthropogenic radioactivity on earth derived from 423 nuclear weapons tests carried out between 1945 and 1980. The USA, UK, and USSR were responsible for 90% of these tests. About three-quarters of the radioactivity fell on the Northern Hemisphere, and fallout was maximal about latitudes 30°–60°. The levels of radioactivity near the sites of the explosions— including Novaya Zemlya, Marshall Islands, Christmas Isand, French Polynesia, Lop Nop, Johnston Atoll, and the Nevada test site— were higher in some but not all cases.

In April 1986, there was a major accident in the operation of a nuclear power plant at Chernobyl, in the Ukraine. This accident released the largest amount of nuclear materials so far, and the materials were widely distributed. 137Cs was the most abundant and widespread isotope released from Chernobyl. Clouds carrying radioactive material from Chenobyl traveled north, and raised the 137 Cs content of the Baltic by a factor of 10, and its concentrations remain much higher than in any other European sea as we turn into the 21st century. The Black Sea was also affected (Chapter 1); concentrations of 137Cs increased by a factor of two. In the western part of the Black Sea, increased radioactivity arrived by river transport rather than by atmospheric deposition. A few other accidents involved planes carrying nuclear weapons. One such accident took place in Palomares, Spain, releasing isotopes; small amounts of 240Pu and 241Am from this accident reached the Mediterranean. Nuclear-powered submarines have had accidents and have released small amounts of radioactive materials into the sea. In general, concentrations of radioactive isotopes, taking 137Cs as representative of anthropogenic materials, have been highest in the Baltic, Irish, and Black Seas, followed by the North Atlantic and the North Sea (Fig. 13.4). These differences were the result of the combined sources from fallout and reprocessing plants.

Releases from nuclear plants

These include discharges to coastal waters associated with the operation of nuclear power and nuclear fuel reprocessing plants. Nuclear power plants have been involved in minor releases of radioactive materials, with quite local and low levels of contamination. Nuclear fuel reprocessing plants chemically treat spent nuclear fuel, and generate large quantities of radioactive wastes. Some of these reprocessing plants have discharged small amounts of radioactive materials, which have been detectable at considerable distances.10 Dumping of radioactive wastes

Low-level wastes in containers, and reactor assemblies with or without spent nuclear fuel, have been discharged at sea. These practices were banned in 1972, but some dumping appears to have continued until 1993. Some local (a few meters) leakage of radioactive materials have been found in some places, but in general there is 10

Only a few nuclear reprocessing plants have been involved in signiﬁcant release of radioactive materials to the sea. These are the Sellaﬁeld plant on the Irish Sea, the La Hague plant on the northwest coast of France, and two plants on the West Coast of India (Livingston & Povinec 2000).

Fate of radioactive materials in coastal environments Because of differences in the mix of sources, and rather different chemical properties of different elements released into given coastal sites, the movement and effects of radioactive isotopes are quite heterogeneous. Different isotopes have quite different behaviors in aquatic environments. Some, such as 3H, 14C, 99Tc, 129I, and 137Cs, are soluble in

337

OTHER AGENTS OF COASTAL CHANGE

sea water, and remain in water masses, which has made them effective tracers of water movements. In contrast, the plutonium isotopes, 241Am, and other isotopes have low solubility in sea water and instead adsorb to particles, and in coastal environments quickly are removed from the water to the sea ﬂoor.11 For example, 55Fe discharged from a UK nuclear power plant appeared in sediments in the immediate vicinity of the discharge pipe.12 The concentrations of 55Fe diminished by more than six-fold within 300 m away from the discharge site (Warwick et al. 2001). On the other hand, in the case of other isotopes, long-distance transport is common. Brown rockweeds Fucus vesiculosus and F. serratus along the Swedish coast took up 99Tc released as part of the operation of a UK nuclear reprocessing plant (Lindahl et al. 2003). The 99Tc concentrations in rockweeds varied in proportion to release rates from the plant, and recorded a transport time of 4–5 years between the release point in the Irish Sea and the Swedish coast.

11

In oceanic waters the particle density is lower, and most radioactive materials remain in the water column (Livingston & Povinec 2000). 12 Under aerobic conditions iron appears as Fe3+, and forms insoluble ferric hydroxide compounds; Fe3+ therefore should precipitate quite near the point of discharge, and accumulate within sediments below. Therefore the 55Fe released from power plants in such circumstances should also appear in nearby sediments.

Antarctic

S Atlantic

S Pacific

C Atlantic

C Pacific

Iceland

Indian

NW Atlantic

N Pacific

Greenland

Norwegian

Arctic

Barents

North

Black

Irish

Baltic

0.1

NE Atlantic

1

Mediterranean

10

137

Figure 13.4 Mean 137Cs concentrations in surface waters of the world’s oceans and seas, adjusted to the year 2000 to account for the half-life of the isotope. From Livingston and Povinec (2000).

Cs concentration (Bq m–3)

100

Seas

In most coastal sites, the radioactive isotopes derive from a mix of sources. The Ob estuary received global atmospheric fallout from nuclear bomb tests, releases from nuclear power and reprocessing plants, and nuclear accidents (Cochran et al. 2000). Sediment proﬁles in Dutch salt marshes showed peak concentrations of 137Cs from atmospheric deposition from nuclear bomb tests and a mix of 134Cs and 137Cs released by the Chernobyl nuclear power plant accident. There are also less prominent but more chronic isotope contributions from the nuclear reprocessing plants in Sellaﬁeld, UK, and La Hague, France (Dyer et al. 2002). The distance traveled by speciﬁc radioactive isotopes is obviously limited by the relative duration of the half-life of the isotope. For example, 55 Fe has a half-life of 2.7 years. The 55Fe discharged from a decommissioned nuclear power plant in Winfrith, UK, decreased markedly in concentrations within sediments and organisms within 2–5 years after the discharges were reduced (Warwick et al. 2001).13 Such decreases have been reported in other places; recent measurements in harp, ringed, and bearded seals from the European Arctic showed reduced concentrations of 137Cs, which matched decreased 137Cs levels in waters of the Barents Sea and other Arctic areas (Carroll et al. 2002). 13

Curiously, the concentrations in seaweeds did not decrease (Warwick et al. 2001). Such basic differences need more study.

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Concerns about radioactivity have not only been about the initial discharges, which have fortunately been limited, but also about the accumulation of the radionuclides in food webs. A classic example is the deposition of radioactive fallout from bomb tests on Arctic lichens, uptake of the isotopes by reindeer, and ultimately accumulation in Lapps and Aleuts (Warwick et al. 2001). Many coastal organisms take up isotopes. Seaweeds have been used often as an indicator of radionuclide concentrations in sea water, as we saw in the case of the Swedish rockweeds cited earlier. Such seaweeds may concentrate radionuclide concentrations more than 10,000fold relative to seawater concentrations (Warwick et al. 2001). Zooplankton collected near the site of the nuclear accident in Palomares, Spain, contained ﬁve-fold larger concentrations of radioactive isotopes of 240Pu and 241Am than found in plankton collected farther away (Sanchez-Cabeza et al. 2003). Zooplankton effectively concentrated the radioactive elements released in the accident, but did not reach concentrations above established guidelines. Zooplankton samples collected from the North Paciﬁc, away from most local sources, showed evidence of contributions of radionuclides from fallout (Hong et al. 2002). Scallops and mussels from the French Atlantic coast accumulated 210Po, a naturally produced isotope (Bustamante et al. 2002). Seals concentrated 137Cs by factors of 34 to 130 (Carroll et al. 2002).14 The accumulation of isotopes in species eaten by people has of course been a public health concern. Aleuts of northern Canada and Alaska consume large quantities of marine mammal meat. More than 90% of over 200 samples of Arctic marine mammal tissue provided by native hunters had detectable concentrations of 137Cs (Cooper et al. 2000). Nevertheless, the concentrations were low enough that even strictly marine mammal meat diets did not expose the native hunters to

doses that would elicit public health concern (Cooper et al. 2000). The radioactive isotopes thus were superb, low-level tracers of long-distance sources, but did not reach levels that led to deleterious effects. Often there is a relatively rapid turnover of elements within the tissues of many coastal organisms. The isotopic versions are therefore similarly taken up and released back into the environment relatively quickly. 134Cs and 241Am, for example, have “biological half-lives” of 29 and 41 days, respectively, in mussels (Guengoer et al. 2001). Bivalves therefore depurate their isotope content relatively quickly. Even though there is much information on accumulation and turnover in radionuclides in organisms, there is no evidence that extant radioisotopic contents of coastal waters have had any impact on the ﬂora and fauna. Similarly, there is little evidence of deleterious effects of released radionuclides on people. Data on the concentrations of representative anthropogenic (137Cs) and natural (210Po) radioactive isotopes in water from different seas can be used to assess likely isotopic content in ﬁsh caught. The latter data then allow an estimation of the likely exposure that humans eating ﬁsh and shellﬁsh would receive (Livingston & Povinec 2000). These approximate calculations suggest that on a global scale the contribution by natural 210Po in ﬁsh and shellﬁsh is 100–1,000 times larger than the radioactivity provided by anthropogenic 137Cs. Even in Western Europe, where higher radioactive concentrations are found, the anthropogenic sources furnish half the radiation provided by natural sources. In all cases, both natural and anthropogenic sources furnish radioactive exposures below recognized medical thresholds. Livingston and Povinec (2000) concluded their review by saying “. . . the world’s oceans and seas are only slightly contaminated by anthropogenic radionuclides with negligible radiological impact on the world population.”

14

Studies of radionuclide concentration ratios yield highly variable results. For example, concentration factors obtained in laboratory studies on macroalgae were 2.5, for isopods 33, for ﬁsh 2, for brown shrimp 16, and for polychaetes 11 (Topcouglu 2001). Another report concluded that concentration factors in phytoplankton ranged from 200 to 4,000 (Heldal et al. 2001).

Pathogens Microbes that cause disease in other organisms are doubtless powerful forces affecting the work-

OTHER AGENTS OF COASTAL CHANGE

ings of natural communities, as well as human uses of coastal environments. Viral attacks, for example, are as important as grazers in causing mortality of phytoplankton (Suttle et al. 1990). Diseases create die-offs of sea urchins and release kelp and other macroalgae from the inﬂuence of grazing by the urchins (Pearse & Hines 1979). Toxic dinoﬂagellates have led to kills of salt marsh ﬁsh (Burkholder & Glasgow 1997). Microbial diseases are a chief impediment for successful maricultural practices (Lee et al. 1996). It is evident that diseases therefore likely play a diverse and powerful role in coastal environments, but there is still far too little information to gauge their impact. Here I will focus on a far narrower aspect, human diseases in coastal waters, in particular relating to the harvest and consumption of shellﬁsh. The focus on human effects is an anomaly for this book. I treat the subject because in many settings (for example, National Research Council 1994) it has been considered along with the other topics dealt with in previous chapters. Since closures protect stocks from harvest, and preclude human use of a resource, they therefore indirectly alter the coastal environment. Extent of resource closures The area of near-coastal environments subject to shellﬁsh harvest is substantial. In the contiguous USA, more than 85,500 km2 (33,000 square miles) of coastal habitats were, as of 1995, considered to be shellﬁsh producers, yielding a dockside value of roughly US$200 million for about 35 million kg (77 million pounds) of oysters (50% of harvest), clams (40%), mussels (9%), and scallops (< 1%) (NOAA 1998).15 Of these 85,500 km2, 69% were approved for harvest, 19% were conditionally approved, and 13% were closed to harvest.16 The 15

An additional US$137 million was landed by the deeper-water ﬁshery, and a large but undetermined amount of coastal shellﬁsh was taken by recreational near-shore harvest (NOAA 1998). Incidentally, the oysters taken from natural populations are only 28% of the national harvest: the remainder are grown in maricultural facilities. Mariculture oyster production has been increasing during the end of the 20th century and more recently (Chapter 11). 16 Shellﬁshing areas are regulated on the basis of certain indicator standards. A shellﬁshing area is closed if there are more than 14 fecal coliform colonies per 100 ml in the overlying water.

339

latter number is considerably lower than the 48% recorded in 1985, and is the lowest in the history of the surveys by the National Oceanic and Space Adminstration (NOAA), the relevant branch of the US government. The diminishing extent of shellﬁsh bed closures is a result of improved wastewater and runoff management. Connection of shellfish and human diseases Shellﬁsh feed on particles suspended in the overlying water. To obtain a sufﬁcient ration, shellﬁsh ﬁlter remarkably large volumes of water, sorting out and ingesting detrital and live particles from bacteria to larger zooplankton. This feeding mode has made shellﬁsh excellent collectors of microbes, among which may be disease-producing species. Eating shellﬁsh has always therefore been somewhat of a gamble for people, because of the possible contraction of disease or ingestion of toxins. In New England, for example, there was a long-held tradition that one ought not to eat shellﬁsh in months without an “r” in their name. Presumably, the May–July set of months might have been most likely to be those in which shellﬁsh contain toxic harmful blooms or human-derived pathogens. In fact, the frequency of such toxic mussels may be behind the historical near-absence of mussels in the cuisine of people in the eastern North American coast, in contrast to their popularity in Europe. Eating contaminated shellﬁsh was early connected to a number of human diseases, including typhoid fever, hepatitis, diarrhea, cholera, and other intestinal maladies (Rippey 1994; Murphree & Tamplin 1995). In the 1920s a serious outbreak of typhoid fever (1,500 cases and 150 deaths) was traced to sewage-polluted oysters. This triggered efforts by the US Public Health Service to monitor and investigate ways to detect wastewater contamination in shellﬁsh. Disease organisms are quite diverse and difﬁcult to routinely screen, and often are rather rare in water samples, so that, at least early in the 20th century, other means of detection of the possible presence of human pathogens were sought. Human intestines are host to microorganisms characteristic of the guts of warm-blooded animals. These fecal coliforms,

CHAPTER 13

clostridia, enterococci, and streptococci are therefore abundant in human waste waters, as are the viral phages that attack certain coliform bacteria. These organisms have for many decades been used as indicators of the possible presence of the more rare and difﬁcult to diagnose pathogenic organisms, such as botulism, Salmonella, and staphyloccal bacteria (Grimes 1991), hepatitis and polio viruses, and protozoans such as Giardia and amoeba. In the USA and elsewhere, the abundance of enteric (deriving from human intestines) indicator organisms above some threshold concentration in water, sediments, or shellﬁsh and ﬁnﬁsh has been the criterion applied to ban the use of those resources, either by closing a shellﬁsh harvest area, or through condemning the catch. Sources of human-derived pathogens Many reports provide the suspected reasons for closures of shellﬁsh harvest areas. These include urban runoff (40% of cases), unidentiﬁed upstream sources (39%), wildlife (38%), septic systems (32%), wastewater treatment plants (24%), agricultural runoff (17%), boating facilities (17%), boats (13%), industry (9%), sewer overﬂows pipes (7%), direct discharges (4%), and livestock feed lots (3%) (NOAA 1998). From the above survey list, and many other speciﬁc studies, it is evident that sources of indicator organisms are mixed (Baudart et al. 2000; Lipp et al. 2001a); in most cases it is hard to sort out the speciﬁc sources, and there are surely additional inputs.17 As with so many other examples discussed in previous chapters, we might not be sure about what speciﬁc land covers are responsible for adding possible pathogens to estuaries. But we can be readily convinced that as the density of people (and their activities) on a watershed increases, the larger the degree of contamination—in this case, the concentration of fecal coliform bacteria—in 17

Studies of source contributions reveal that resuspension of sediment bacteria may be an additional, seldom recognized, source (Struck 1988; Valiela et al. 1991). Because of the rather high concentrations of fecal coliform bacteria so often found in anoxic sediments, resuspension of even a few centimeters of sediment can furnish much of the fecal coliform load seen in the water column after disturbances such as storms.

400 Fecal coliform (cfu 100 ml–1)

340

Salinity (‰) <20 20–29 >30

300

200

100

0 0

2 4 Population density (individuals ha–1)

6

Figure 13.5 Geometric means of fecal coliform counts (in colony forming units (cfu) per 100 ml) in reaches with different ranges of salinity, for ﬁve estuaries in the coast of North Carolina, plotted vs. the density of people on the contributing watersheds. Dashed line shows the threshold for shellﬁsh bed closure (14 cpu 100 ml−1). Data from Mallin et al. (2000).

the receiving coastal waters (Fig. 13.5). In addition, it is clearly the fresh water originating on land transports the fecal coliforms to receiving estuarine waters, as can be deduced from the much larger fecal coliform counts in fresher reaches of tidal creeks of North Carolina (Fig. 13.5). Concentrations of enteric indicators become lower in the transition from land to sea. Concentrations of fecal coliform bacteria are diluted by mixing with cleaner sea water. These bacteria also suffer high mortality owing to intolerance to salt, are killed by exposure to light (Tudor et al. 1990; Bravo & de Vicente 1992; Burkhardt et al. 2000), and are grazed by many microscopic consumers (George et al. 2001). In addition, bacteria adsorb to particles (George et al. 2001), and may then accumulate in the bottom as particles sediment to the estuary ﬂoor. In the organic, nutrientrich, anoxic sediments characteristic of shallow coastal environments, fecal coliform and related bacteria may not only survive, but also divide so that populations may therefore grow (Gerba & McCleod 1976; Struck 1988). Accumulation, survival, and division are therefore responsible for the considerable larger (by one to two orders of magnitude) densities of benthic fecal coliform

341

Sediment fecal coliforms per 100 ml

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107 106 105 Shiaris et al. 1987 Labelle et al. 1980 Gerba & McLeod 1976 Struck 1988 Van Donsel & Geldreich 1971

104 103 102 10 1 10–1 10–1

1

10 102 103 104 105 Water fecal coliforms per 100 ml

106

Figure 13.6 Fecal coliform concentrations in sediments in relation to associated measurements of fecal coliforms in overlying water in ﬁve studies (references given in Valiela et al. 1991). From Valiela et al. (1991).

counts relative to counts done on the overlying water column (Fig. 13.6) (Erkenbrecher 1981; Struck 1988; Martinez-Manzanares et al. 1992). Monitoring the status of shellfish harvest areas The use of enteric indicators from the 1920s on has in fact been quite a success, and has reduced the incidence of shellﬁsh-borne bacterial disease for people (Richards 1987). The application of indicators such as fecal coliform counts are thoroughly entrenched in the legislation governing management of coastal resources (there are also standards for swimming and boating, as well as for harvest). The results of studies to assess the usefulness of enteric indicators are mixed. Some reviews of the evidence found that enteric indicators were a useful basis for following the deterioration or recovery of coastal ecosystems (Sawyer 1988). Other studies suggest a weak or zero correlation of the abundance of enteric indicators in the overlying water column or in sediments to the occurrence of pathogens in shellﬁsh (Rodrick et al. 1988; Martinez-Manzanares 1992). Others found no relation of water column fecal coliform density to fecal coliforms in shellﬁsh, but a significant link between the far more numerous sediment fecal coliforms and shellﬁsh fecal coliforms (Valiela et al. 1991). In general, measurements in sediments provided more convincing evidence

of wastewater contamination than measurements in water (Watkins & Burkhardt 1996). Moreover, outbreaks of viral diseases such as Norwalk fever, hepatitis A, and viral gastroenteritis, still occur (Richards 1987; Conaty et al. 2000). Although enteric indicators in the water have done yeoman service through most of the past century, it seems evident that there might be a number of continuing problems with their use. First, different enteric indicators are inconsistent with each other (Fig. 13.7), so that different assessments may be reached depending on which one is preferred. Second, birds and mammals other than humans also contribute fecal coliforms indistinguishable from human sources (Watkins & Burkhardt 1996). Third, survival of enteric bacteria in the environment may differ from survival of the pathogens they are supposed to indicate (Grimes et al. 1986; Rippey & Watkins 1992; Burkhardt et al. 2000). For example, a pathogenic virus may be present in the absence of enteric indicator bacteria (Robertson & Tobin 1983; Nicholson et al. 1989; Watkins & Burkhardt 1996; Conaty et al. 2000). Fourth, these indicators may accumulate or even reproduce within sediments (Struck 1988), so that they are far from a passive tracer of possible pathogens. These drawbacks of enteric indicators have been long recognized, but their continued use has been argued on the basis that management has to take place, and the enteric indicators are the best

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400 300 200 100

Fecal coliform bacteria (cfu 100 ml–1)

0 0

200 300 100 Enterococci (cfu 100 ml–1)

400

400 300 200

References

100 0 0

40 60 80 20 C. perfringens (cfu 100 ml–1)

100

0

120 160 40 80 Coliphage (pfu 100 ml–1)

200

400 300 200 100 0

Figure 13.7 Comparisons among different enteric indicators (fecal coliforms, enterococci, Clostridium perfringens, and coliphages) using data from two estuaries (Charlotte Harbor and Sarasota Bay) in Florida, US. cfu, colony forming units; pfu, plaque forming units. Data from Lipp et al. (2001a, 2001b).

alternative available. Actually, new developments in biogeochemical analyses and molecular biology make possible a more direct assessment of the presence of human waste water18 and of pathogens. In terms of potential for human disease, there have been proposals for new methods to detect enteric microbial indicators (for example, Rhodes & Kator 1991; Fiksdal et al. 1994), but 18

it seems best to try to detect the pathogens themselves. Rather than merely follow the questioned indicators, recent advances in molecular methods make it feasible to amplify even weak signals from pathogens that may be in low but nevertheless dangerous abundance, and directly assess their presence (Pancorbo & Barnhart 1992; Hernandez et al. 1995; Jiang et al. 2001). Such new assays seem the more practical way to solve the issues that have hampered speciﬁc detection of human pathogens in coastal environments and shellﬁsh.

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Topcuoglu, S. 2001. Bioaccumulation of cesium-137 by biota in different aquatic environments. Chemosphere 44:691– 695. Tudor, J. J., M. P. McCann, and I. A. Acrich. 1990. A new model for the penetration of prey cells by bdellovibrios. J. Bacteriol. 172:2421–2426. Tunowski, J. 1988. Zooplankton losses during the passing through the cooling system of a power station. Ekil. Pol. 36:231–243. Valiela, I., M. Alber, and M. LaMontagne. 1991. Fecal coliform loadings and stocks in Buttermilk Bay, Massachusetts, USA, and management implications. Environ. Manage. 15:659–674. Walker, T. R., K. Reid, J. P. Y. Arnould, and J. P. Croxall. 1997. Marine debris surveys at Bird Island, South Georgia 1990–1995. Mar. Pollut. Bull. 34:61–65. Warwick, P. E., A. B. Cundy, I. E. Croudace, M. E. D. Bains, and A. A. Dale. 2001. The uptake of iron-55 by marine sediment, macroalgae, and biota following discharge from a nuclear power station. Environ. Sci. Technol. 35:2171–2177. Watkins, W. D., and W. Burkhardt, III. 1996. New microbiological approaches for assessing and indexing contamination loading in estuaries and marine waters. Pp. 241–263 in Vernberg, F. J., Vernberg, W. B. and Siewicki, T. (eds). Sustainable Development in the Southeastern Coastal Zone. Belle Baruch Library of Marine Science No. 20. University of South Carolina Press, Columbia, SC. Weisberg, S. B., W. H. Burton, F. Jacobs, and E. A. Ross. 1987. Reductions in ichthyoplankton entrainment with ﬁne-mesh, wedge-wire screens. N. Am. J. Fish. Manage. 7:386–393. Whittingham, A., D. Hartley, J. Kenney, T. Cole, and E. Pomfret. 2003. Large Whale Entanglement Report 2002 (updated 2003). National Marine Fisheries Service, Protected Resources Division, 1 Blackburn Drive, Gloucester, MA 01930. Unpublished report.

Chapter 14 Summing up

Artist’s conception of an asteroid impact on earth. Such an impact created the large changes in fauna that deﬁned the Cretaceous–Tertiary epochs, and included the extinction of dinosaurs and many other species. Image from NASA.

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Linkages among different agents of environmental change If any one feature emerges from the material included in the previous chapters, it is the multifaceted complexity of the subject matter. The various agents of change not only prompt different effects on different components of coastal environments, but they also interact with one another in a complicated fashion. Neat separation of subject matter in separate chapters has been hard to maintain. Throughout, the agents of change interact with one another, repeatedly threatening to burst out of their own separate chapters. For example, climate change involving altered temperatures is intimately involved in shifting freshwater and sediment transport, and hence to transport of toxic compounds through coastal environments and into food webs. In the chapter on habitat loss I principally made use of the examples of outright destruction of marshes and mangroves, but discussions of loss of many other coastal habitats appeared repeatedly in other chapters, for instance in the linkage between freshwater and ﬂood management, sea level rise, and maintenance of coastal wetlands. Sea level rise, changes in sediment loads, sea ﬂoor disturbance during trawling, destructive ﬁshing in coral reefs, all can lead to loss of various coastal habitats. Eutrophication creates anoxia, which also leads to habitat losses, and, incidentally, is also often connected with toxic substances, harmful algal blooms, and alien species. Perhaps the most widespread interaction between agents of change taking place on the world’s coastal environments is that between eutrophication and overﬁshing. We can truly say that human beings are conducting a global-scale experiment in bottom-up and top-down control of coastal ecosystems. This is a theme that has relevance to applied issues as we have highlighted in earlier chapters, and has also, appropriately, been a focus of much basic research attention (Worm et al. 2002; Valiela et al. 2004). Human meddling with these two major agents of environmental change will go a long way to determine what we see as the biological make-up of coastal environments in the future. What we can say now

is that owing to human nutrient enrichment and overﬁshing, there have been major changes in the way coastal environments look now compared to what they looked like a century ago: enrichment broadly has made for more algal growth, overﬁshing has greatly diminished the number of large grazers and carnivores, and certain habitats have decreased in extent.1 These human-mediated changes have had consequential repercussions throughout the affected environments. Many other examples of such linkages appear through the various chapters, for instance, warmer waters and increased UV radiation bleach corals, which also suffer from overﬁshing and eutrophication. Increased climatic warming may favor invasions of exotic species from more tropical areas. Toxic substances may impair ﬁsheries, and so will destruction of habitats necessary for the stocks. As we saw in the Nile, the interception of sediments and nutrients affect ﬁsheries. And so on; the connections among the various agents of coastal environmental change are endless.

Defining effects Another aspect that prevents easy summing up of the material covered is that the impacts of given agents of change are not of equivalent intensity, nor are the effects evenly spread across all types of organisms, or over space. In fact, we can assume that both intensity and extensiveness of effects varies enormously in every instance. Habitat loss locally removes all organisms within many mangrove forests, but sonar affects deep-diving beaked whales near speciﬁc places where naval maneuvers have been held. Eutrophication substantially alters most coastal waters of the world, but the Exxon Valdez spill provided a clear example of wildly varied impacts from one site to another. Warming patchily bleachs many corals across the tropical coasts of the world. These complexities have made it hard to clearly identify the nature of each agent of change, the 1

Boris Worm pointed out to me the striking collection of “before and after” images of a variety of marine environments displayed at http://www.shiftingbaselines.org/lenticulars/index.html.

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scale and impact of its effects, and the potential for management or remediation. In spite of the complexities, the cavalcade of agents of change and their heterogeneous effects, described in the chapters of this book, have demonstrated in a compelling fashion that as human populations have increased, and concentrated along coasts, our activities increasingly have thoroughly, and widely, altered or destroyed coastal organisms and environments. Having made the effort to sort out the admittedly interlinked, complicated effects of the major agents of change impinging on the world’s coasts, it seems useful to try to summarize and compare the relative inﬂuence of the various agents of coastal change.

Comparisons of impacts of the agents of change Before engaging in comparisons among the different agents of coastal change, it should be crystal clear that all the agents of change discussed in this book have important effects on some organism, in some place, at some level. The alterations the various agents of change have brought about have justiﬁably given rise to concern; that is why they merited mention in the separate chapters. Any summary comparison perforce glosses over many details, forcing classiﬁcations that oversimplify what are intrinsically interlinked, complicated topics; but to carry out a comparison, some constraints have to be assumed. Here I focus on environmental aspects, leaving out human health, economics, and political issues. The emphasis of the comparison will be on the global dimension of the changes. At the local level, I have documented, throughout chapters of this book, intense effects of all the agents of change on some organism, place, or environment; if we were to restrict comparisons to local spatial scales we would not further our comparative understanding of the relative impacts of the agents of coastal change. The intent of the exercise that is summarized in Table 14.1 is therefore to compare the impacts of the different agents of change at the larger global level.

Faced with similar complex and diverse issues, government agencies, regulatory bodies, and other institutions often call on panels of recognized experts to provide their perspectives on issues. Such panels then use their joint experience to try to reach consensus on the issues at hand. Examples of panel reviews on topics such as ours are available in GESAMP (1990) and National Research Council (1994). Here we will assume the role of such experts, since we have reviewed the facts throughout the previous chapters. The task is therefore to set up criteria that enable teasing apart key aspects of environmental change that are part of a complicated gossamer of links, interactions, and different levels of effects. The synthesis exercise assesses intensity and extensiveness of effects, potential for recovery, and future prognosis, for each agent of change. The intensity of the effects can be assessed at different levels of biological organization (population, ecosystem, or biome2), and the extensiveness can be gauged considering the breadth of the taxonomic or geographic range of the effects (Table 14.1). Recovery from impacts of the agents of change differed, and so did susceptibility to management action. In addition, it seems reasonable to consider the long-term prognosis, that is, whether expert opinion considers that the problem is improving or likely to worsen in coming decades. If we can agree that these evaluation criteria do capture the variety of environmental effects, we might then go on to compare the impacts of the various agents of change. To do this, we need a protocol that, in an informed fashion, compares the relative effects of the various agents of change. A possible protocol would be to assign values of 1 through 5 to each evaluation category: “1” meaning least, “5” meaning most pronounced. The number entries in the cells of Table 14.1 were derived from consideration of the material in the various chapters, which are here condensed in the following over-brief summaries. •

2

Warming bleaches corals, shifts distributions of species, and may alter inorganic carbon equilibria. These impacts are mainly at the Referring to the entire geographic range of a habitat.

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Table 14.1 Relative assessments (1 to 5; 5 being the worst case) of the intensity and extensiveness, potential for recovery, and forecasted prognosis about effects of the various agents of coastal environmental change at global spatial scales, based on demonstrated evidence. Intensity of effects at different levels

Agent of change Temperature increase UV increase Sea level rise Alteration of freshwater transport* Alteration of sediment transport Loss of habitats Petroleum hydrocarbons Chlorinated hydrocarbons Metals Exotic species Harvest of fin- and shellfish Eutrophication Litter Thermal pollution Sound pollution Radioactive pollution Pathogens

Extensiveness of effects

Potential Overall for Prognosis assessment Population Ecosystems Biome Taxonomic Geographic recovery for future value 4 2 2

2 1 2

2 2 2

2 2 2

4 3 4

3 1 5

4 2 4

21 13 21

3

4

1 (3)

3

1 (3)

3 (4)

3

18 (21)

3 5 5 5 3 5 5 5 3 3 2 1 1

4 5 4 2 2 3 4 5 1 1 1 1 2

1 2 2 1 1 1 3 3 1 1 1 1 1

2 5 2 2 1 2 2 5 1 2 1 1 1

2 4 2 2 2 4 5 4 1 1 1 1 1

5 3 2 2 2 2 3 3 1 1 1 1 1

3 5 2 1 2 4 4 5 5 1 3 1 2

20 29 19 15 13 21 26 29 13 10 10 7 9

*Normal entries assume a reduction in delivery of fresh water to the seas; entries in parentheses assume an increase.

population level; there is too little information as to effects at the ecosystem or biome level. Bleaching affects a narrow group of species (corals), but geographic shifts associated with warming affect many taxa; both bleaching and shifts are worldwide. Recovery will require massive economic shifts and political will, and warming will become worse before action is likely. • Increased UV radiation may have injurious consequences for many species, but evidence on intensity, extensiveness, and so on, is scanty. Recovery will follow the prohibition of aerosols, which is already taking place. • Sea level rise ﬂoods uplands, leads to habitat loss, and increases the erosion of shores. Biological (but not social and economic!) impacts seem relatively mild, probably affect few taxa, but are widespread on the world’s coasts.

The prognosis is for continued rise at a relatively fast pace, so issues will be exacerbated. • Alteration of freshwater transport changes salinity and circulation, lowers nutrients in estuaries (and perhaps adjoining open ocean waters), and alters estuarine vegetation, ﬁsheries, and wildlife. The extent and direction of the alteration may depend on whether there is less or more fresh water entering the seas,3 a topic that awaits further work, and separation of global climatic inﬂuences from human water demands. Most evidence in hand suggests mostly local, human-derived impacts, and marked reductions of ﬂow in certain estuaries. Recovery in these sites may 3

The two possibilities were covered in Table 14.1 by assuming a reduction, indicated by the normal entry, or an increased delivery of fresh water, indicated by the rating in parentheses.

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•

•

•

•

•

•

be possible if human water use is better managed, but the continuing human demand for fresh water suggests a worsening prognosis. Alteration of sediment transport conspires with changes in fresh water and sea level to erode shores and submerge wetlands. These changes affect many species and whole ecosystems, but are mainly local in area. Recovery of the affected systems is unlikely, but management practices on land could be applied to address the issues, so future sediment loads may be better managed. Loss of coastal habitats removes populations and ecosystems, but is usually spatially heterogeneous, so that not all the biomes have been lost. Loss affects all species present in the habitat, and is widespread. Coastal habitats can be restored, but prognosis has to be guarded, since more and more people are accumulating near the coast, and this is the root cause of habitat loss. Petroleum hydrocarbons may have intense local effects, which are often temporary, on populations and ecosystems. Contamination is widespread, recovery occurs, and the longterm prospects seem to be improving owing to increased awareness and precautionary measures. Chlorinated hydrocarbons are deleterious to certain, not all, organisms. Because of active research and public awareness, the use and subsequent coastal release of these chemicals into coastal environments are decreasing. Metals have largely non-lethal impacts on a selected few types of organisms, except under unusually high concentrations. Metal contamination has limited geographic extent, and the impacts are often quite local, limited to near the sources of the metals. Exotic species have recently been widely reported and are an issue that has attracted widespread attention. Exotics may have impacts partly displacing native populations. Invasions affect nearly every type of coastal environment, but blooms of alien species tend to come and go without wholesale eradication of living organisms. Preventions seems far more practical than eradication, and some

•

•

•

• •

• •

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remedies to reduce invasions are available and could be put in place. Harvest of ﬁnﬁsh and shellﬁsh leads to intense and widespread effects in coastal environments. Substantial removal of many stocks, across most coastal waters, is taking place, and there are considerable habitat disturbances associated with ﬁsheries. Means of management are available, and there is evidence that stocks can recover once harvest pressure is lessened. Eutrophication creates intense ecological changes in innumerable coastal sites worldwide. Populations are severely impacted. This agent of change is the one that may have the best-documented alterations to ecosystemlevel processes and function. Causes are well established and there are remedies for restoration, but they are costly. Litter impacts selected species to some degree, but is widespread over the seas. The occurrence of litter in the coasts will only get worse before it gets better. Thermal pollution affects populations taken into thermal plants, but the effects are quite local. Sound pollution is a disturbance that borders on a threat to whales, particularly beaked whales. Its effects are narrowly connected to sites where naval operations take place. Radioactive pollution is a minor agent of change in coastal environments. Pathogens that impact humans do not generally affect natural systems, except indirectly by the closing of shellﬁsh beds.

We all come to exercises such as that of Table 14.1 with a given background and viewpoint. When we are called on to make a judgement, we bring our background and perspectives with us. In any panel whose task is to review complex subjects such as ours there are many individual experts, and many disagreements as to details among them. The assessment values assigned to each cell of Table 14.1 came from consensus of quite diverse opinions from graduate students in a course based on the chapters included in this book. It might be of interest for the reader to create a blank version of Table 14.1, review the

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material in the chapters, and then decide on values to be assigned to each classiﬁcatory variable, for each agent of change, to compare with the values decided on here. A scan down the columns of Table 14.1 reveals that intensity of effects of many agents of change can be high at the population level, less so at the ecosystem level, and relatively rare at the level of entire biomes. The extensiveness of effects on diverse kinds of organisms was particularly marked for eutrophication and loss of habitats. Geographic extensiveness of effects was notable in the case of sea level rise, loss of habitats, exotic species, overﬁshing, and eutrophication. Recovery and prognosis were highly variable across the agents of change. In Table 14.1 the values allotted to each column were also added up to obtain an overall assessment, given in the right-hand column. It might be argued that giving equal weight to each assessment criterion is faulty: for a global assessment, for example, the extensiveness of an agent of change might be held more important than other features, and hence should be given more weight. The problem with any weighting system is that it introduces new uncertainties by making added assumptions, so as a practical default, we merely gave each criterion equal weight; the reader might prefer to do otherwise. The exercise summarized in Table 14.1 is not to be taken as an exact description of the hierarchy of global effects of agents of coastal change. It is, instead, an informed qualitative way to compare complicated change-producing variables and to synthesize much diverse information. The speciﬁc numerical values are rough guidelines that should not be overemphasized. Perhaps a reasonable use of the results of Table 14.1 is to identify tiers of agents of change, all of which have signiﬁcant impact in speciﬁc local geographic scales, but whose global-scale consequences differ. Eutrophication, loss of habitats, and harvest of ﬁsh and shellﬁsh have intense effects, and create environmental change across most of the coasts of the world. These three agents of change are substantively and extensively affecting the world’s coastal environments. As such, they merit immediate attention by global and local agencies

charged with the responsibility of management of the world’s coastal environments. Atmospheric warming, sea level rise, alteration of sediment and freshwater4 transport, and invasions by exotic species, constitute a second tier of important agents of change. Petroleum and chlorinated hydrocarbons, metals, UV radiation, and litter fall into a third tier. These agents of change appeared to be, in some variable way, less intense, or more local in range than those in the ﬁrst tier, and recovery might be more likely. Such agents of environmental change may be addressed by agencies that operate at global or at local scales. Thermal, sound, and radioactive pollution, and human pathogens, though certainly important for some organisms, in some places, and for some times, have limited ranges of intensity and extensiveness. Recovery is possible, plus prognosis is less dire than in the case of agents of environmental change that fell in the ﬁrst tier. In this last tier, action at local scales might be most appropriate, because the impacts are largely focused on speciﬁc species, places, and resources. Other groups of experts (GESAMP 1990; National Research Council 1994) have faced similar challenges, having to assess the relative importance of many causes of environmental change. These groups have compiled expert opinion on the impacts of the diverse agents of change affecting coastal environments. They divided the subject matter in a somewhat different fashion than done in this book, and although their decision protocols were not described it is of interest to compare their conclusions to our assessments from Table 14.1 (Table 14.2). Both assessments assigned high priority to eutrophication and habitat loss. The entries in Table 14.1 also gave overﬁshing high priority, and the NRC panel almost agreed. Table 14.1 also placed higher priority on global atmospheric factors than the NRC report. These differences may be due to the advantage we had of access to more recent data on overﬁshing and climate4

If indeed there is more fresh water entering waters such as the Arctic, this agent of change will have substantial global implications; assessment of this alternative awaits further research results.

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Table 14.2 Results of a review of the relative effects of various agents of change on coastal environments (National Research Council 1994) (left column), using the criteria mentioned in the header. The terms describing agents of environmental change used in the NRC report differ somewhat from the list used here; the approximate equivalent terms used here are also given. For comparison, the right column shows the results of the protocol used in Table 14.1. The order of the agents of change is intended to convey an approximate weighing of the issues. Major coastal environmental issues, chosen because of their broad geographic scope, importance to use of resources, reversibility of effects, and anthropogenic influences (approximate equivalences to terms used in the NRC report, compared to the terms used here, are shown in parentheses) Eutrophication Habitat modification (habitat loss) Hydrologic and hydrodynamic disruption (alteration of freshwater and sediment transport) Exploitation of resources (harvest) Toxic effects (petroleum and chlorinated hydrocarbons, metals) Introduction of exotic species Global climate change and variability (warming, UV radiation, sea level rise) Shoreline erosion (sea level rise) Pathogens and toxins affecting human health (pathogens)

related factors. The newer information probably made the issues more imminent. The assessment in Table 14.1 agreed with the NRC panel that toxic materials had important, but intermediate, global impacts. We concluded that human pathogens, and other topics ignored by the NRC panel as probably less notable, were indeed of lesser impact on a global scale. The assessments entered in Tables 14.1 and 14.2 are best efforts at the summary of large bodies of information and experience—attempts to objectively get to as quantitative a comparison as feasible. The assessments remain to an extent subjective, and not devoid of uncertainty. It would be unwise to use the values to set absolute priorities among each and every one of the various agents of change, everywhere in the world. The summary of assessment criteria shown in Table 14.1 was, instead, intended to serve two purposes: i) to review and compare the extent of the

Relative impacts, at a global scale, of the effects of the following agents of change, from Table 14.1 Eutrophication Habitat loss Harvest of fish and shellfish Warming Sea level rise Exotic species Alteration of sediment and water transport Petroleum hydrocarbons Chlorinated hydrocarbons Metals Litter UV radiation Sound pollution Thermal pollution Pathogens Radioactive pollution

various effects of major agents of coastal alteration, and ii) to identify those agents of change that seemed to have the greatest and broadest impact, with demonstrable effects in the widest range of coastal environments of the world. We might add an additional caveat here, in that Table 14.1 considers the impact of speciﬁc agents of change in isolation. Earlier we made the point that the agents of change we deﬁned often interacted with each other. The exercise of Table 14.1 cannot readily consider such interactions explicitly. Had we included such interactions, it would be likely that global atmospheric changes would need to be given far more attention, because atmospheric changes interact with —and frequently exacerbate—the effects of most other agents of environmental change discussed throughout this book, and given enough climatic alteration, would play increasingly major roles altering coastal environments.

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Sustainable uses? Through the preceding chapters it is obvious that humans have used coastal environments extensively. The effects of the uses range from minor to drastic. In the case of salt marshes, for instance, we saw examples of a range of impacts. The examples spanned from minor harvest of grass for livestock feed in New England in the 1700– 1800s, through various other impacts, including metals, polychlorinated biphenyls (PCBs), oil, eutrophication, and so on, to outright loss of the entire habitat in San Francisco Bay and elsewhere. It seems fashionable today to refer to sustainable uses of natural environments. The problem is just where to draw lines as to what might be a sustainable level of use and alteration and what might not be. The temptation might be to select key components, and set standards of sustainability based on them. For salt marshes in the East Coast of the USA, salt marsh cord grass would be the logical choice. Even there, however, it is a challenge to decide what standards to demand. With a plant whose height can vary between 10 cm and 3 m in salt marshes just a few kilometers apart, it is not a simple issue. And there are those experts who will aver that just to grow cord grass is insufﬁcient to provide the services expected of a functioning salt marsh. Similar arguments can be raised about every attempt at deﬁning sustainable uses. Sustainability therefore is an attractive concept that is difﬁcult to deﬁne in an operational fashion. Lacking an operational base, sustainability seems an impractical goal for environmental policy. It might pragmatically be better to simply aim to diminish the impact of human uses of coastal environments as far as possible, knowing full well that the economic, social, and population pressures will work in the opposite direction everywhere.

Prevention or restoration? In the earlier chapters we saw many instances of attempts to restore damage to coastal environ-

ments. Mangroves were replanted, oil spills were fertilized to encourage oil-degrading bacteria, ﬁsh stocks were protected from harvest, and natural reserves were set up to protect certain populations and environments. Such remediation measures work to some different degree in each case. They are admirable in all cases, and ought to be supported, and research needs to be done to improve their success. What is less widespread is the will and means to prevent, a priori, as much of the damage as possible. This is akin to the case of fossil fuel use: do we simply go on with “business as usual”, seeking more fossil fuels (with all the consequences that entails), or do we spend some political will and capital to encourage the countries of the world to conserve and be more fuel-efﬁcient? In the case of coastal environments, we cannot expect to continue our current expansion of perturbation without signiﬁcant consequences. We can, with sufﬁcient capital and knowledge, restore some of the damage, but our efﬁcacy at restoration is not high, and restoration work is costly. It seems prudent (and practical) to devote far more effort to prevention (as in the case of double hulls for oil tankers!) than to pin our hopes on the uncertain and expensive option of restoring coastal damage after it is done.

Separating natural from anthropogenic changes The chapters of this book have documented innumerable changes in coastal environments, most of them by human-driven agencies. There is no question that we have brought about a plethora of alterations, particularly during the 20th century. I selected the frontispiece to this chapter to remind us, however, that environmental changes have been a repeated and characteristic feature of the history of the earth. As already mentioned in the introduction, the fossil record is a lengthy tale of alterations, some of a catastrophic scale, to the biology of the earth. There has never been a “balance of nature” in the sense of long periods uninterrupted by alterations to the biological

SUMMING UP

make-up of natural environments. Coastal environments, situated as they are at the interface of land and sea, have been particularly subject to global-scale alterations in sea level, sediment and water shifts, and so on. The focus of this book has been agents of change owing to human activities, but the effects of these anthropogenic factors are often hard to sort out from those derived from “natural” agents of change. Sea level owing to groundwater removal combines with temperature-driven seawater expansion, and both may be masked, or emphasized, by tectonic crustal shifts. Greenhouse gas forcing of atmospheric and ocean temperature takes place in the context of solardriven climatic change. Changes in freshwater transport owing to human water use are closely linked to sediment and nutrient supplies to coastal environments, and hence to eutrophication. Certain powerful political ﬁgures have in recent years taken advantage of the uncertainty associated with natural external changes—for example, in regard to global atmospheric changes, or about hydrocarbon contamination of the atmosphere—to ﬁnd reasons to deny human responsibility for global environmental change, or at least to defer politically daunting action to reduce or manage consumption. Unfortunately, what may be politically convenient positions today might allow environmentally dangerous conditions in the longer term. Human-mediated alterations featured in the chapters of this book are indeed but the latest in the long series of small and large changes that have, since the beginning, altered the nature of the coastal zones of the world. The salient and awesome fact, however, is that the coastal changes we have prompted have begun to match and exceed those of the external sidereal agents of change—asteroid impacts, crustal plate movements, wobbles in the earth’s orbit, changed solar energy delivery, and so on. While we have no recourse but to adjust to externally driven

355

alterations, it is possible to conceive of modiﬁcations in our behavior that would lessen pressure on coastal—and other—environments. Some of the alternatives have been mentioned toward the ends of each chapter; in each case, conservation and management of consumption should be a prominent ﬁrst step. All the mitigating actions would require reassessment of our relative priorities about economics, standard of living, property rights, how we value natural environments, political and economic differences in different parts of the world, and our perception of sufﬁcient evidence and risk. This is a daunting cultural and political challenge. No doubt, the needed changes in how we make use of the earth’s resources, and in particular those in the coastal environments of the world, will not take place by themselves. We should hope that we recognize the need for action and move to address the impacts of global coastal change. We should not wait until the fate of the Black and Aral Seas become commoner, seaweeds hide beaches everywhere, the Gulf Stream no longer warms Northern Europe, more cities suffer Venice’s fate, mangroves and marshes have largely become shrimp farms or suburbia, and overﬁshing frees us from the dilemma of deciding whether to consume mercury and PCB-laden tuna.

References GESAMP (Group of Experts Assessing Marine Pollution). 1990. State of the Marine Environment. Report Study No. 39. United Nations Environment Programme, Nairobi, Kenya. National Research Council (NRC). 1994. Priorities for Coastal Science. National Academy Press, Washington, DC. Valiela, I., D. Rutecki, and S. Fox. 2004. Salt marshes: Biological controls of food webs in a diminishing environment. J. Exp. Mar. Biol. Ecol. 3000:131–160. Worm, B., H. K. Lotze, H. Hillebrand, and U. Sommer. 2002. Consumer versus resource control of species diversity and ecosystem functioning. Nature 417:848–851.

Index Page numbers in italic refer to ﬁgures, those in bold refer to tables, and numbers followed by n refer to footnotes. acetaldehyde 203, 204, 205 acoustic thermometry 335 acqua alta 48, 51–3, 53 Adriatic Sea debris 325 North ﬂood events 48, 51–3, 53 history and development 49–50 sea level rise 48, 49 –57 Aegean Sea 2, 3 aerosols effects on climate 28–9 sources 28 agar 8 agents of change comparisons of impacts 349–53 deﬁnitions of effects 348–9 interactions between 348 natural and anthropogenic 354–5 agriculture 17–19 Cape Cod 284 effect on sediment load 105, 112 introduction of exotic species 239 – 40 irrigation and its effects 80–6, 87, 89 need for fresh water 87 airshed areas 294, 300n Alaska, Exxon Valdez accident 146 – 56, 161 albatross, bycatch mortality 270 Aleut people effects of Exxon Valdez oil spill 156 radioactive isotope ingestion 338 algae see phytoplankton; seaweed alien species see exotic species alkanes 159, 161 alkenes 159 alkyl benzenes 160 alkyl naphthalenes 160 Amazon River, sediment load 113, 115 American chestnut 236

American shad, San Francisco Bay 227 ammonium atmospheric deposition 294 Black Sea 7n and eutrophication 306, 308 export by salt marshes 136 as source of nitrogen 294 Amoco Cadiz accident 149n, 158n, 161, 161n, 162–3 Amsterdam, changes in sea level 57 Amudarya River, interception for irrigation 80–6, 99, 100 anammox reaction 7n, 308n anchovies/anchoveta, abundance 276, 277 anoxic conditions 290, 306, 310 Antarctic ice sheet formation and sea level 60 antifouling paint 215, 216–17, 217, 218, 219, 219, 240 aquaculture see mariculture aquifers groundwater withdrawal from effect on sea level 60–1 and salt water intrusion 72 Venice 54, 73–5 water pumped back into 75 aragonite 40 Aral Sea effects of freshwater interception for agriculture 79, 80–6, 87 ﬁshing 81–4 salinity 81, 83 surface area 81, 82, 83 water level 81, 83 Arctic Ocean, freshwater discharge to 98–9 Arctic shammy, response to oil spills 150, 151 Argo Merchant oil spill 163 assimilative capacity 6n, 299 asteroid impact with earth 347 astronomical cycles 54 Aswan High Dam 88n, 89, 91, 116 effect on ﬁshing 95 effect on Mediterranean salinity 93–4, 93

effect on Nile nutrient transport 94–5 Atchafalaya River wetlands 118 athletic shoes, loss at sea 329, 329 Atlantic Ocean North, salinity changes 99 Northwest, ﬁshing 246–57, 258 atmosphere, global changes in chemical composition 27–9 atmospheric temperature comparison of modeled and measured estimates 31–2, 34 long-term changes 30, 31 records 30 and sea temperature 32, 34, 35, 36 see also climate change Aurelia aurita, Black Sea 10 bald eagle, effects of chlorinated hydrocarbons 190n ballast water, introduction of exotic species in 234, 240 Baltic Sea chlorinated hydrocarbon concentrations in herring 193, 194 debris 325 eutrophication 293 radioactive contamination 335, 336 Bangkok ﬂooding 67–8 subsidence due to groundwater extraction 60–1 Bangladesh, effects of sea level rise 71–2 Barrow’s goldeneye, effects of Exxon Valdez oil spill 150, 152, 154 Basque people, ﬁshing 246–9 Bay of Biscay, debris 325 beaches loss of 130 nourishment/replenishment 119–21 beluga whale chlorinated hydrocarbon accumulation 187n effects of noise 334 benthic invertebrates, effects of oil spills 149, 151, 165

358

benzenes 160 Beroe ovata, Black Sea 11 bicarbonate ions in sea water 40 biocontrol by/of exotic species 234 biodeposits from mariculture 275 biomagniﬁcation chlorinated hydrocarbons 182–4, 189 metals 212–13 birds and bycatch 263 chlorinated hydrocarbon bioaccumulation and effects 188, 190, 195 effects of oil spills 148, 150, 152–3, 153, 154, 164 importance of mangrove forests 140 importance of salt marshes 139 ingestion of debris 330 overwintering trends 38 see also individual species bivalves as markers of contamination 178n San Francisco Bay 227 see also speciﬁc types and species black-legged kittiwake chlorinated hydrocarbon accumulation 195 effects of Exxon Valdez oil spill 150, 154, 153 black oystercatcher, effects of Exxon Valdez oil spill 152 Black Sea algae 8–9, 10 ammonium 7n biological responses in 7–12 case history 2–13 eutrophication 7n, 13 ﬁshing 10, 11, 11n habitat threats 12–13 human inﬂuences 3–5 jellyﬁsh 10–11, 13, 231, 237 “Noah’s Flood” incursion 2n nutrients 5–6, 13 phytoplankton 7–9 pollen deposits in sediments 3 pollutant input 12 radioactive contamination 12, 335, 336 recovery 13 salinization 2 seagrass 9 sediments 6–7 tourism 7 water 7

INDEX

watershed 3 zooplankton 9–10, 11 blue whale, effects of noise 334 blueﬁsh 260 Bora 53 Bosphorus Strait 2, 3 Boston changes in coastal environment 125, 126, 127–9 eutrophication management 316–17 metal pollution in harbor 221 Boundary Bay, exotic species 231 bowhead whale effects of noise 334 population recovery 269 BRDs 270 breakwaters 118 British Isles, sea level changes 64, 66 brittle stars, PCB contamination 188–9 brown pelican, chlorinated hydrocarbon accumulation and effects 190, 195 brown tides 301–3 bubonic plague 27n bufﬂehead, effects of Exxon Valdez oil spill 154 Burntwood River, discharge 90, 98 butane 28n, 160 butterﬁsh 260 bycatch 262–3, 275 reducing devices 270 Caboto, Giovanni ( John Cabot) 247 cadmium concentration in sea water 208, 210 properties 208, 209 retention in estuaries 207–8 retention in salt marsh sediments 210, 211 sources 207 uptake and bioaccumulation 212 Canada, ﬁsheries 249, 250, 254–5, 256, 258 canal construction and introduction of exotic species 235, 236 Cape Cod agriculture 284 eutrophication 284–92 overwintering bird species 38 seagrass meadows 289, 290, 291 urbanization 284 –92 see also Waquoit Bay, Cape Cod capelin 275

carbon effects of eutrophication on carbon cycle 306, 307 effects of rise in sea temperature on dissolved inorganic carbon 39–40 export by salt marshes 136 sequestration by oceans 38n carbon dioxide 157 atmospheric concentration 27, 28, 40 and climate forcing 28, 29 dissolved in sea water 40 sources 28n carbon monoxide 28n carbonate ions dissolved in sea water 40 Carson, Rachel, Silent Spring 190 Caspian Sea 86 Castillo de Belver 158n cats, effects of mercury 203, 206 Caulerpa C. racemosa 235 C. taxifolia 235, 240 Celtic Sea, debris 325 Central Park Lake, New York, metal concentrations 221 CFCs see chloroﬂuorocarbons Changjang River, sediment load 115 Chernobyl 12, 335, 336 Chesapeake Bay clams 231, 232 cryptogenic species 237 degradation 131 oysters 301 Chile, salt marshes 59 Chioggia 50, 54n chlorinated hydrocarbons bioconcentration 184–5, 186 biomagniﬁcation 182–4 concentrations in organisms 182–90 Danube and Black Sea 12 degree of environmental contamination 182 depuration 185, 187 distribution and time course 191–5, 195, 195 effects 190–1, 350, 351 New Bedford Harbor case history 174, 175–80, 191 structure and properties 180–2 chloroﬂuorocarbons 180 atmospheric concentration 27, 28 and climate forcing 28, 29 sources 28n

INDEX

cholera 39 chromium Black Sea 12 retention in salt marsh sediments 210 sources 207 Ciconia ciconia, effects of reduced freshwater discharge 96–7 cinnabar 209n clams association with human disease 339 – 42 Cape Cod 290 Chesapeake Bay 231, 232 copper accumulation 212, 212 effect on water quality 232, 233 effects of harmful algal blooms 302 effects of oil spill 151 as exotic species 227, 231–2, 238 PCB contamination 177, 178 San Francisco Bay 227 value of harvest 339 clathrates 59n clean air legislation 221 climate change 30–8, 42–3 effects on coastal environments 38 – 42 historical concerns about 42–3 statement on science of 37 climate forcing effects of greenhouse gases 28, 29 role of aerosols 28–9 clouds, role in climate forcing 29 Clupea pallasi, response to oil spills 149, 151 coastal engineering works effects on sediment load 106 erosion remediation 118–21 coastal erosion Ebro delta 107, 108 prevention by mangrove forests 140 remediation 118–21 and sea level rise 71 and sediment load 107, 108, 116, 118 –20 coastal habitats chlorinated hydrocarbon contamination 192–3 effects of mariculture 274 limiting role of nitrogen 295 loss 68 –71, 124– 45, 350, 351 New England 125–30 nutrient sources 294–7

patterns of exotic species invasion 237–9 see also speciﬁc habitats cobalt properties 208 sources and time course of contamination 220, 220 cod 245 effects of harmful algal blooms 302 ﬁshing 247–53, 253, 254, 258, 284 population crisis in North Sea 39 response to oil spills 150, 151 salt 246 Colorado River dams 91, 111 estuarine vegetation 95 ﬂow management 89–91, 99, 100 sediment load 111 Columbia River estuarine vegetation 95 ﬂow regulation 89, 90 salinity 94 common reed 230n common tern, PCB concentrations 195 Connecticut, salt marshes 131 copper in antifouling paints 216n biological effects 213 Black Sea 12 properties 208, 209 retention in estuaries 208 retention in salt marsh sediments 210 sources 207, 207, 220, 220 time course of contamination 220, 220 uptake and bioaccumulation 212, 212, 213, 214 coral reefs bleaching 22, 23–7, 349, 350 and dissolved inorganic carbon 40 Easter Island 22 effects of eutrophication 305 effects of sea level rise 68 nutrients in surrounding water 304n range 24n recovery from eutrophication 315–16 seaweed 23, 304n threats to 130–1 cordgrass see Spartina

359

cormorant effects of chlorinated hydrocarbons 190 effects of Exxon Valdez oil spill 154 Coto Doñana 96 cotton production, Soviet Union 80–1, 84, 86, 87 Crimea, tourism 7 croaker, bycatch mortality 270 crown-of-thorns starﬁsh, effect on coral reefs 23 crude oil 159, 160 see also petroleum hydrocarbons crustal deformations and sea level 61–4 cryptogenic species 237 Curacao, debris 325 Cuvier’s beaked whale, effects of noise 334 Cystoseira barbata, Black Sea 9 cytochrome P450 as marker of exposure to aromatic hydrocarbons 152n dams 88–91 Colorado River 91, 111 Danube River 5, 6–7 Ebro River 108 effect on sediment load 108, 116 Danube River 3, 4 altitude and depth 5 cross-section along course 5 eutrophication 6, 7n, 13 hydroelectric dams 5, 6–7 nutrient load 6, 13 pollutants 12 sediment transport 6–7 DDD 181, 182 DDE 33, 181, 182, 190 DDT 180 applications 181 bioconcentration 184–5, 186 biomagniﬁcation 182–4 Black Sea 12 concentrations in organisms 182–90 Danube River 12 distribution and time course 191–5, 195 effects 190–1 resistance to 181n structure 181, 182 debris abundance 325 distribution 327 effects 327–30, 350, 355

360

debris (cont’d ) entanglements with 324, 330 –1 ingestion 327–30 management and prevention 331 occurrence 324–7 sources 325, 326 transport of exotic species on 331 deforestation Black Sea watershed 4 effects on hydrological cycle 61 Delaware Bay, degradation 131 deltas Ebro 105, 106 –9 erosion 105, 106 –9, 116 Ganges-Bramaputra 67, 71–2 Mississippi 69, 111–12, 116–17 Nile 95 – 6, 96, 98n, 110 –11, 116 Po 55, 108n dengue fever 39 denitriﬁcation 308 desalinization 102, 109 desalinization plants, entrainment losses 331–2 Desmarestia imbricata, effects of oil spill 151 DIC, effects of rise in sea temperature 39 – 40 dichlorethane 180 Dictosphaeria cavernosa 315–16 dimethyl sulﬁde 7n, 28 disease associated with shellﬁsh 339 – 42 dispersants used on oil spills 169 dissolved inorganic carbon, effects of rise in sea temperature 39 – 40 dogﬁsh shark 255, 255, 261 dogwhelk, imposex 216–17, 217, 218, 219, 219 dolphins bycatch mortality 270 chlorinated hydrocarbon accumulation 187n Dover sole, chlorinated hydrocarbon bioaccumulation 195 dredging for beach nourishment/ replenishment 119, 119–21 damage to habitats 264–5 PCB-contaminated sediment 179 Easter Island, coral reefs 22 Eastmain River 100 Ebro watershed and delta, sediment transport 105, 106– 9

INDEX

eelgrass meadows see seagrass meadows eels, PCB contamination 177 El Niño–Southern Oscillation (ENSO) 23–4n, 43 and coral bleaching 23–4 and harmful algal blooms 303 and sea level 56, 61 elasmobranch ﬁsheries 258–9 energy consumption 156, 157 renewable 156 sources 156 ENSO see El Niño–Southern Oscillation enteric indicators 340, 340–2 equine encephalitis 39 estuaries classiﬁcation 312 effects of reduced freshwater discharge on vegetation 95 fast ﬂushing 301 inverse 94 metal retention 207–8 nutrient enrichment 295, 296 river-borne sediment in 114–15 water quality standards 312 ethane 160 eutrophication 283–323 Baltic Sea 293 Black Sea 7n, 13 Cape Cod 284–92 Danube River 6, 7n, 13 deﬁnition and causes 292–4 due to agriculture 18–19 effects 300–11, 350, 355 extensiveness and scale 311–12 and food supply 308–9 management 312–18 nutrient sources 294–7 time trends in nutrient enrichment 297–300 and urbanization 284–92, 293 exotic species 350, 351 accidental release 234–5 agricultural introduction 239–40 balancing positive and negative effects 239–40 common features of invasions 231–3 introduction 226–44 Lessepsian invasions 235, 236 limitations on interpretation of invasion data 236 –7 mariculture 275 mechanisms of invasion 233–6

mitigation or prevention of invasions 240–1 pattern of invasion in coastal environments 237–9 San Francisco Bay 227–8 Spartina alterniﬂora 228–30 transport on debris 331 expert panels 349 extreme weather events, increase in 42 Exxon Valdez accident 146–56, 161 factory ships 252–3, 259 famine 18n fecal coliform bacteria 339–40, 340–1 fertilizers 18–19, 286–7, 297, 298 ﬁddler crabs, response to oil spills 165, 167 Fiji, qoliqoli 271–2 ﬁn whale effects of Exxon Valdez oil spill 152 effects of noise 334 ﬁsh biomass index 255 chlorinated hydrocarbon accumulation 177, 178, 179, 188, 189, 195 decreased size of individuals 261–2 effects of chlorinated hydrocarbons 190 effects of mercury 206 populations of commercially targeted stocks 257–60, 261 response to oil spills 149–50, 151–2, 163, 164, 165 San Francisco Bay 227–8 and undersea sounds 333–4 see also individual species ﬁshing 245–82, 350, 351 and altered species composition 263–4 Aral Sea 81–4 Basque people 246–9 Black Sea 10, 11, 11n bycatch 262–3, 275 bycatch reducing devices 270 Canada 249, 250, 254–5, 256, 258 cod 247–53, 253, 254, 258, 284 discards 260, 261, 262–3 effect on coral reefs 23n effects of eutrophication 310–11 effects of increases in take 257–65 effects of reduced freshwater discharge 95

INDEX

Georges Bank 252, 253n, 255 –6, 256 –7, 261, 272 habitat disturbance due to 264–5 importance of mangrove forests 140 importance of salt marsh habitats 136 –7, 138 industrial decline 273n interaction with other agents for change 276–8 introduction of exotic species 234 Massachusetts 249, 250, 251 Mediterranean 11n, 95–6 Minimata Bay 202–3, 205 New Bedford 175–6 New England 125, 126, 249 –50, 251, 255, 256 Northwest Atlantic 246–57, 258 recovery 256–7 regulation and control 253, 254–5, 256 –7, 257n, 267–9 remedies for overﬁshing 266–75 sustainable 260 technological advances 250–3 use of explosives 130n, 264n ﬁshing gear damage to habitats caused by 264 – 5, 266, 267 as debris 326 entanglement with 263, 330 ﬁshmeal production 275 Fitzroy River, mangrove swamps 69, 70 ﬂooding due to sea level rise 67–8 North Adriatic 48, 51–3, 53 Florida, oil spill 162n, 165–6, 167 Florida Everglades 132n food chain ampliﬁcation see biomagniﬁcation food production 17–19 food webs alterations in species composition caused by ﬁshing 263–4 bottom-up vs. top-down control 260, 300n decrease in size of top predators 261–2 effects of eutrophication 308–9 “ﬁshing down” 261–2 removal of top predators 260 forests, interception and retention of nitrogen 287n Fraser River estuary, eutrophication management 315

fresh water balances 92 consumption of renewable supply 99–100, 101 discharges alterations in 79–104, 350–1 consequences of reduction 93–8 Ebro delta 108 effects of reduction on vegetation 98 extent of alteration 99–100, 101 increases in 98–9 restoration potential 100–2 and sediment load 113, 114 human demand for 86–93 interception, Aral Sea 80–6 magnitude of human use 92–3 nutrient transport 109 role of waterworks 88–91 storage, effect on sea level 60–1, 61 withdrawal from aquifers effect on sea level 60–1 and salt water intrusion 72 Venice 54, 73–5 Fucus F. gardneri, effects of oil spills 150 F. serratus, radioactive contamination 337 F. vesiculosus effects of oil spills 149, 151 factors affecting 238 radioactive contamination 337 Fundulus heteroclitus, PCB contamination 177, 178, 179 gadoid outburst 39n Galveston–Houston, subsidence due to groundwater extraction 61 Ganges–Bramaputra delta, effects of sea level rise 67, 71–2 Ganges River, sediment load 115 gannet, entanglement with debris 330 garbage see debris gasoline 160, 221 genetically modiﬁed crops 18n, 102 Georges Bank ﬁshing 252, 253n, 255–6, 256–7, 261 ﬁshing ban 272 food web 264 habitat disturbance 266 oil exploration 256 Georgia Bight 136 ghost-ﬁshing 330 Ghulam Mohammed Barrage 96

361

glacial forebulge 63 glacial rebound 61–3 glaciation and sea level 63–4 glaciers ice volume and sea level 60 thinning and retreat 30–1, 32, 60 glaucous-winged gull effects of chlorinated hydrocarbons 190n effects of Exxon Valdez oil spill 154 Glenn Canyon Dam 91 global warming see climate change Gomez, Esteban 125 gray whale effects of Exxon Valdez oil spill 152 extinction 259n population recovery 269 Great Barrier Reef, bleaching 24, 25 green crab predation by 238 range 231 San Francisco Bay 227 greenhouse gases and climate change 31, 32, 34, 35 and climate forcing 28, 29 concentrations 27–8 Greenland, ice sheet melting and sea level 60 groins 118, 118 groundwater see fresh water Guadalquivir River 96 Gulf of Mexico, hypoxic conditions 306, 307 Gulf Stream 99 Gulf War (1991), oil spills associated 158n, 161, 164–5 Haber–Bosch process 18 HABs 301–3 haddock 254 Halifax Harbor, Nova Scotia, petroleum hydrocarbon proﬁles in sediments 158 Halimione 124 Hansa Carrier, loss of shoes from 329, 329 harbor porpoise, effects of Exxon Valdez oil spill 152 harbor seal effects of chlorinated hydrocarbons 191n effects of Exxon Valdez oil spill 152 harlequin duck, effects of Exxon Valdez oil spill 152, 152–3n, 155 harmful algal blooms 301–3

362

Hawaii, debris on beaches 325–6 heavy metals see metals Helen Mar 174 hepatitis, Aral Sea region 84–5 herring 255n Baltic Sea 193, 194 response to oil spills 149, 151 Hexaplex trunculus 219 Hokkaido, debris 325 Homer, Iliad 4 Hoover Dam 91, 111 horned grebe, effects of Exxon Valdez oil spill 154 horseshoe crab, Cape Cod 290 Huanghe River, sediment load 109 –10, 115, 116 humpback whale effects of Exxon Valdez oil spill 152 effects of harmful algal blooms 302 effects of noise 334 entanglement with debris 330 Hurricane Bob 42 Hurricane Mitch 23, 42 hydroelectric power 5, 6–7, 88n hypoxic conditions 290, 306, 307, 310 deﬁnition 312–13 imposex 216–17, 217, 218, 219 Indus River estuarine vegetation 95 ﬁshing 96 sediment load 115 introduced species see exotic species invertebrates effects of chlorinated hydrocarbons 190 PCB bioaccumulation 188 Irish Sea, radioactive contamination 336 iron properties 208 retention in estuaries 208 retention in salt marsh sediments 210 sources and time course of contamination 220, 220 irrigation for agriculture 80–6, 87, 89 cotton production in Aral Sea region 80–6 for leisure activities 87n and soil salinization 81, 84, 109 Ixtoc-1 blow-out 158n, 161, 161n, 163 – 4

INDEX

Jamaica, sea surface temperature 26 Japan, subsidence due to groundwater extraction 61 jellyﬁsh, Black Sea 10–11, 13, 231, 237 jetties 118, 118 Kaneohe Bay, eutrophication management 315–16, 318 kelp see Fucus, F. vesiculosus kerosene 160 killer whale acoustic foraging 333n chlorinated hydrocarbon accumulation 187, 189 effects of Exxon Valdez oil spill 152 effects of noise 334 killiﬁsh, PCB contamination 177, 178, 179 Kiribati 67n kleptochoroplastidy 303 Kodiak Island, debris 325 Kotri Barrage 95 Kromme estuary 95 Laguna Salada 95 Laholm Bay, nutrient enrichment 295, 296 Lake Mead 91 Lake Michigan, mussels 227n Lake Nasser 88n land reclamation by Spartina 226, 228 Laurentian ice sheet 63 lead biological effects 213–14 Black Sea 12 from waste incineration 221 properties 209 retention in salt marsh sediments 210, 211 sources 207, 207, 220, 220 time course of contamination 220, 220 uptake and bioaccumulation 211, 211 Lena River, sediment load 115 Lessepsian invasions, exotic species 235, 236 levees, Mississippi River 117 Lido 50, 54n litter see debris Little Ice Age 27n, 30 Littorina littorea, imposex 216–17, 218 livestock, role in nitrogen cycle 294n lobster, PCB contamination 177 loess 109n

Louisiana, wetlands 133 lunar declination 54n MacKenzie River, sediment load 115 mackerel 255n macroalgae see seaweed Magnuson Fishery Conservation and Management Act 254 Malamocco 50, 54n malaria 181 Maldives, effects of sea level rise 67n mangal see mangrove forests manganese retention in estuaries 208 in salt marsh sediments 210 mangrove forests effects of reduced freshwater discharge 95 effects of sea level rise 68–9 Fitzroy River, Australia 69, 70 functions 140–1 loss of 133–5 metal retention 210 restoration 141, 142 terminology 134n uses 134 Manila, wetland reclamation 61 mariculture 272–4 disadvantages 274–5 effects on coastal habitats 274 effects on water and sediment 274–5 exotic species 275 mangrove forests 140n oysters 216n, 339 salt marshes 137–8 and wild stock depletion 275 marine reserves 270–2 Marshall Islands 67n Massachusetts, ﬁshing 249, 250, 251, 252 Mayan civilization, collapse 30n Medieval Warm Period 30 Mediterranean debris 325 ﬁshing 11n, 95–6 radioactive contamination 336 salinity 93–4, 93 melon-headed whale, effects of noise 334 menhaden 260 chlorinated hydrocarbon bioaccumulation 195 mercury 191n biomagniﬁcation 213 Black Sea 12

INDEX

methylation 208 Minimata Bay contamination 201– 6, 206, 215 properties 208–9 retention in salt marsh sediments 210 sources 207n, 207 toxicity 209n uptake and bioaccumulation 213 merganser, effects of Exxon Valdez oil spill 150, 154 Mesodinium, Black Sea 7 metallothioneins 215n metals balance between natural and anthropogenic sources 207n biological effects 213–15, 350, 351 biomagniﬁcation 212–13 Danube and Black Sea 12 mercury contamination at Minimata Bay 201–6, 206, 215 methylation 208 properties 208–9 resistance to 215 retention in coastal ecosystems 207– 8, 210–11, 210, 211 sources in coastal waters 207, 207, 215 –22 time course of contamination 215 –22 uptake and bioaccumulation 211–13, 214 methane 160 atmospheric concentration 27, 28 and climate forcing 28, 29 emission from Black Sea 7n and eutrophication 306 sources 28n Méton cycle 54n mew gull, effects of Exxon Valdez oil spill 154 microbes denitriﬁers 308 effects of eutrophication 305–8 effects of PCBs 176, 177 pathogenic 338–42, 350, 351 response to spilled oil 167n microplastics as debris 327 Minimata Bay, mercury contamination 201–6, 206, 215 Minimata disease 203–4, 205 minke whale chlorinated hydrocarbon concentrations 193 effects of Exxon Valdez oil spill 152

Mississippi River delta 69, 111–12 loss of salt marsh 116–17 levees 117 nitrate transport 297, 298 sediment load 111–12, 115 mixed function oxygenases 167n Mnemiopsis, Black Sea 10–11, 13, 231, 237 Mondego estuary 124, 132 Montreal Protocol 26n MOSE Project 55–6 Mount Pinatubo 29 mudﬂat reclamation 226, 228n, 229 Muller, Paul 181 murre chlorinated hydrocarbon accumulation 195 effects of Exxon Valdez oil spill 152, 154 mussels association with human disease 339–42 Cape Cod 290 chlorinated hydrocarbon contamination 177, 178, 191, 193, 195, 195 copper accumulation 212 effect on water quality 232 effects of oil spills 151 as exotic species 238 Lake Michigan 227n mariculture 274–5 value of harvest 339 Mya arenaria 10n mycosporine-like amino acids 27n, 41 Mytilus see mussels Nantucket Island 175 naphthalenes 160 naphthenes 159 Narragansett Bay, nitrogen budget 314 narwhal, effects of noise 334 Naskaupi River 100 natural gas 160 Netherlands, effects of sea level rise 71 Neuse River, nitrogen/nitrates 293 New Bedford Harbor 174, 175–80, 191 New England changes in coastal environments 125–30

363

ﬁshing 125, 126, 249–50, 251, 255, 256 settlers 125, 127 New Haven Harbor, metal retention 208 New Jersey, debris on beaches 325 New York city effects of sea level rise 72, 73, 74 eutrophication management 315 Newfoundland, importance of cod ﬁshing 250 nickel retention in estuaries 208 sources 207, 220 time course of contamination 220 Nile River damming 89 delta 95–6, 96, 98n, 110–11, 116 discharge 90, 91 effects of sea level rise 67 ﬁshing 95–6 and Mediterranean salinity 93–4, 93 nutrient transport 94–5 sediment load 110–11 nitrogen cycle, effects of eutrophication 306, 308 nitrogen/nitrates in agriculture 18–19 atmospheric deposition 294, 296–7 budgets 314 Cape Cod 286–92 concentration in sea water 295 Danube and Black Sea 6, 13 enrichment at whole-estuary scales 295, 296 effects 300–11 experiments 295, 296 time trends 297–300 evidence for limiting role in coastal systems 295 interception and export by salt marshes 136, 139 interception and retention in forests 287n isotopic ratios 290–1, 292 land-derived 291 and mariculture 274 Mississippi River 297, 298 modiﬁcation of effects 301 Neuse River 293 outputs 314 and population density in European river watersheds 5, 6 sources/inputs 287, 294–7, 298, 313–14 standards for 312

364

nitrogen oxides 28n nitrous oxide 308 atmospheric concentration 27, 28 and climate forcing 28, 29 emission from Black Sea 7n sources 28n Noctiluca, Black Sea 7, 9n non-indigenous species see exotic species North Sea debris 325 effects of rise in sea temperature 39 overﬁshing 252 Northern fulmar, chlorinated hydrocarbon accumulation 195 northwestern crow, effects of Exxon Valdez oil spill 154 Norwuz oil ﬁeld spill 158n Nothofagus 59 Nova Scotia, petroleum hydrocarbon proﬁles in sediments 158 Nucella lapillus, imposex 216–17, 217, 218, 219, 219 nuclear fuel reprocessing plants 336, 337 nuclear power plants discharge of heated water 332–3 entrainment losses 331–2 radioactive contamination from 336 nutrients Black Sea 5–6 effects of reduced freshwater discharges 94–5 enrichment effects 300–11 experiments 295, 296 time trends 297–300 and eutrophication 292–4 export from mangrove forests 140 export from salt marshes 136 freshwater transport 109 land-derived, interception by wetlands 139 and mariculture 274 modiﬁcation of effects 301 sources 94n, 294–7, 298 standards for 312–13 Ob’ River, sediment load 113 Oil Pollution Act 1990 156 oil spills 157, 158 Amoco Cadiz 149n, 158n, 161, 161n, 162–3 Argo Merchant 163

INDEX

deﬁnition of recovery from 149n effects 163–7 experimental 166 Exxon Valdez 146–56, 161 fate of released oil 161–3 Florida 162n, 165–6, 167 Gulf War (1991) 158n, 161, 164–5 Ixtoc-1 blow-out 158n, 161, 161n, 163–4 management 168–9 Prestige 168 prevention 169 oleﬁns 159 organotin pollution 215, 216–17, 217, 218, 219, 219, 240 see also tin Orinoco River, sediment load 113 osprey, effects of chlorinated hydrocarbons 190, 193 otter trawls 250, 253 outwelling hypothesis 136 overﬁshing 250–2, 253–7 deﬁnition 267n effects 257–65 remedies for 266–75 oxidation of petroleum hydrocarbons 161 Oxus River, interception for irrigation 80–6, 99, 100 oxygen and eutrophication 306, 307, 310 standards for 312–13 oysters association with human disease 339–42 Cape Cod 290 Chesapeake Bay 301 dredging for 264n effects of TBT 216 mariculture 216n, 339 PCB contamination 178, 195 San Francisco Bay 227, 228 spat collection 125n value of harvest 339 and water quality 232n ozone atmospheric concentration 27, 28 and climate forcing 28, 29 depletion 26n, 28n “holes” 29 sources 28n and ultraviolet radiation 29 Paciﬁc herring, response to oil spills 149, 151

Paciﬁc Ocean debris 325 sea level 64, 65 PAHs 160 Papua New Guinea, sea level 61 parafﬁns 159 passenger pigeon 236 pathogens 338–42, 350, 351 PCBs see polychlorinated biphenyls PCNs 182 penguins bycatch mortality 270 distribution 39 peregrine falcon, effects of chlorinated hydrocarbons 190 periwinkle, imposex 216–17, 218 Persian Gulf, oil spills associated with ﬁrst Gulf War 158n, 161, 164–5 pesticides 180 petroleum hydrocarbons aliphatic 159 aromatic 159–60 Black Sea 12 in coastal environments effects 163–7, 350, 351 fate 161–3 management 167–9 sources 156–8 composition 159–60 Danube 12 Exxon Valdez accident 146–56, 161 movement within marine environment 162, 162–3 natural seeps 157 processing 160 proﬁles in sediments 158 toxicity 160 transport 157 weathering within marine environment 162, 161 Pﬁesteria 303 pH of sea water 40 phosphorus/phosphates concentration in sea water 295 Danube and Black Sea 6, 13 enrichment at whole-estuary scales 295, 296 experiments 295, 296 time trends 299 limiting role 295n and population density in European river watersheds 5, 6 photo-oxidation of petroleum hydrocarbons 161 Phragmites australis 230n

INDEX

Phuket, sea surface temperature 26 Phyllophora, Black Sea 8, 10 phytoplankton Black Sea 7–9 effects of eutrophication 288, 289, 300 –3, 310–11 effects of reduced freshwater discharges 94–5 effects of sea temperature rise on seasonal cycles 40–1 factors affecting abundance 301 passage through power plants 331–2 in salt marshes 137 pigeon guillemot, effects of Exxon Valdez oil spill 150, 154 plastics as debris 326 ingestion by animals 327–30 Pleiades 2 Po River 50 delta 55, 108n sediment load 55 polar ice cap melting and sea level 60 poldering 127–8 pollutants Black Sea 12 interception by mangrove forests 140 polychlorinated biphenyls (PCBs) accumulation in killer whales 187, 189 applications 181 bioaccumulation estimates 185, 188 bioconcentration 184–5, 186 biomagniﬁcation 183–4 concentrations in organisms 182– 90 depuration 185, 187 distribution and time course 191–5, 195, 195 effects 190–1 fugacity 184n human exposure to 177, 178, 179 ingestion 33 lipid solubility and storage 181n, 184 –5 New Bedford Harbor contamination 176–80, 191 properties 181–2 structure 181, 182 polychlorinated naphthalenes 182 polycyclic aromatic hydrocarbons 160

population and agriculture 17 coastal density and growth 16–17 decrease 14n density 3n global numbers and growth 13–16 urban/rural split 14–15 pout 275 precipitation changes, effect on coastal environments 42 Prestige oil spill 168 prevention of pollution and degradation 140, 169, 240–1, 331, 354 Prince William Sound, Exxon Valdez accident 146–56, 161 propane 28n, 160 Prorocentrum cordatum 8 Protothaca staminea, effects of oil spill 151 Puccinellia 59 Puget Sound, exotic species 231 Pycnopodia heliantoides, effects of oil spill 151 radioactive contamination 335–8, 350, 351 Black Sea 12, 335, 336 sources 335–6 Rapana thomasiana 10n red maple 237 red-necked grebe, effects of Exxon Valdez oil spill 150, 154 red tides 301–3 reserves 270–2 reservoirs 88–91 sediment trapping 116 resource consumption and use 17–19 restoration of habitats 72, 100–2, 141, 142, 228, 313–18, 354 right whale entanglement with debris 330, 331n protection 269 Rio de la Plata estuary, debris 325, 327, 328 Rivoalto 49, 50 rockweed see seaweed Rosetta Promontory 110–11, 116 Sabbadini, Cristoforo 50 St Lawrence estuary, metal retention 208 salinization of irrigated soils 81, 84, 109

365

salmon effects of harmful algal blooms 302 effects of oil spills 149–50, 151, 155 salt extraction 124, 138 use in food preservation 246 salt marshes aesthetic value 139–40 biodiversity 233n Chile 59 Connecticut 131 contaminant interception in 138 die back 229n effects of oil spills 165–6 effects of sea level rise 68–9 export of energy-rich substances 136 functions 136–40 hay 127, 128, 139 importance to coastal ﬁshery stocks 136–7, 138 interception of land-derived nutrients 139 invader species 136–7, 138 livestock grazing on 127, 139 loss of 131–3 mariculture 137–8 metal retention 210–11, 210, 211 as migratory stop-over sites 139 Mississippi delta 116–17 New England 127–9 PCB contamination 191 phytoplankton 137 Portugal 124 as post-glacial features 63–4 restoration 141, 228 role in shoreline stabilization 138–9 salt extraction 124, 138 San Francisco Bay 131, 132, 133 and seagrass meadows 139, 140 Thames estuary 132 United States 132, 133 as waterfowl refuges 139 Salton Sea 91, 95 San Francisco Bay copper concentrations 212, 214, 221 cryptogenic species 237 invasion by exotic species 227–8, 239 salt marshes 131, 132, 133 Spartina 228 sand eels 260n, 275 sand lance 255n sardines, abundance 276, 277 Sargasso Sea, debris 325

366

Sargassum baccularia 304 Saros cycle 54n sassafras 284n scallops association with human disease 339 – 42 effects of harmful algal blooms 302 Georges Bank 272 value of harvest 339 Scandian ice sheet 63 schistosomiasis 98 scoter, effects of Exxon Valdez oil spill 154 scup, PCB contamination 177, 188, 189 sea level and crustal deformations 61–4 and El Niño–Southern Oscillation 56, 61 and glacial ice volume 60 and glaciation 63–4 long-term changes in 57 mechanisms of change 59–64, 67 meteorologically driven change in 61 processes affecting 56 restoration potential 72–5 rise effect on coastal environments 67–72, 350 global 56–64 magnitude 64–6 New England 124, 125 North Adriatic 48, 49 –57 and reduced freshwater discharge 98 and tectonic activity 59, 59, 64 sea lion, effects of chlorinated hydrocarbons 191n Sea of Cortez 94 Sea of Marmara 2, 3 sea otter, effects of Exxon Valdez oil spill 152, 153– 5 sea snakes, bycatch mortality 270 sea temperature and atmospheric temperature 32, 34, 35, 36 and coral bleaching 24–5, 26 effects of a rise 38–41, 349–50 sea urchins, coral reefs 23 sea walls 118, 118 sea water acidity 40 intrusion into freshwater environments 71–2 nutrient concentration 295

INDEX

salinity and coral reefs 24n effects of reduction in freshwater discharges 93–4 thermal expansion of volume 59–60 seagrass meadows Black Sea 9 Cape Cod 289, 290, 291 effects of eutrophication 289, 290, 291, 304–5, 309 effects of oil spills 149, 151 loss of 130 and salt marshes 139, 140 seasonal cycles, effects of rise in sea temperature 40–1 seastars, effects of oil spills 151 seaweed Black Sea 8–9, 10 chlorinated hydrocarbon concentrations 193, 194 coral reefs 23, 304n effects of eutrophication 283, 288–90, 303–4 effects of oil spills 149, 150, 151 factors affecting 238 as indicator of radioactive contamination 338 mariculture 274n radioactive contamination 338 recovery from eutrophication 315–16 sediment Amazon River 113, 115 Black Sea 6–7 Changjang River 115 and coastal erosion 107, 108, 116, 118–20 Colorado River 111 compaction 59n Danube River 6–7 Ebro watershed and delta 105, 106–9 effects of agriculture 105, 112 effects of coastal engineering works 106 effects of mariculture 274–5 effects of waterworks 105–6 Ganges River 115 general load patterns 113–14 Huanghe River 109–10, 115, 116 human-mediated changes in transport 109–14 Indus River 115 interception and decreased load 108, 116 –17

Lena River 115 MacKenzie River 115 Mississippi River 111–12, 115 Nile River 110–11 Ob’ River 113 Orinoco River 113 PCB concentration 176 petroleum hydrocarbon proﬁles 158 Po/Venice Lagoon 54–5 retention of metals 210–11 river-borne, in coastal environments 114–16 transport alteration 105–23, 350, 351 and water discharge 113, 114 and wetlands 108, 109, 116–17 Yenisey River 115 seiches 53 sewage disposal and eutrophication 293 and metal pollution 220–1, 221 as source of nitrogen 286–7, 297 treatment methods 299, 315 shellﬁsh association with human disease 339–42 commercial growing 137–8 area covered 339 closure to harvest 339 see also speciﬁc types shipping, introduction of exotic species 234, 240 shipworms 239 shrimp culture and habitat destruction 274 importance of salt marshes 136, 137 invasion by viruses 239n pond stocking 275 ponds created from mangrove forests 135 silica 94, 295n Sirocco 53 skate 255, 255, 259n snails, imposex 216–17, 217, 218, 219, 219 sneakers, loss at sea 329, 329 snow, reduced cover 31 solar radiation and atmospheric temperature 29, 30 sonar, military devices 334, 335 sound pollution 333–5, 350, 351 Soviet Union/former Soviet Union, cotton production 80–1, 84, 86, 87

INDEX

Spain, beach nourishment/ replenishment 120–1 Spartina 124 die back 229n divergent views on 230 S. alterniﬂora 229n lead accumulation 211, 211 spread of 228–30 S. anglica 228, 229 S. densiﬂora 228 S. foliosa 228 S. maritima 228 –9 S. patens 69, 228 S. townsendii 229 introduction in Netherlands 226 world distribution 229 San Francisco Bay 228 species distribution ranges, alterations in due to seawater warming 38–9 Standish, Miles 125 Stockholm, changes in sea level 57 Strait of Dardanelles 2, 3 striped bass chlorinated hydrocarbon accumulation 195 recovery of stocks 267–8 San Francisco Bay 227 Sunderbans 71 sunspots 54n sustainability 260, 354 Sydney Harbour, debris on beaches 325, 326 Syrdarya River discharge 90 interception for irrigation 80–6, 99, 100 Tahiti, sea surface temperature 24, 26 Taipei, subsidence due to groundwater extraction 61 Tampa Bay, eutrophication management 316 Taxodium distichum 69 TBT 209, 216–17 tectonic activity and sea level 59, 59, 64 Telmessus cheiragonus, effects of oil spill 151 tetrachloride 180 Thames Barrier, frequency of closure 61 Thames estuary, salt marshes 132 thermal pollution 331–3, 350, 351

Thoreau, Henry David, Cape Cod 284n tidal height, North Adriatic 51 tide gauges and sea level measurement 64 tin biological effects 215, 216–17, 217, 218, 219, 219 concentrations in sediments 219 properties 209 tributyl tin 209, 215, 216–17, 217, 218, 219, 219, 240 from waste incineration 221 TMDLs 313 Tokelau 67n tomatoes, PCB contamination 177, 179, 180 total maximum daily loads 313 tourism, Black Sea 7 Toxic Substances Act (USA) 179 Trevisan, Bernardo, Della Laguna di Venezia 1 tributyl tin 209, 215, 216–17, 217, 218, 219, 219, 240 trichlorethane 180 trichlorethylene 180 Trojan War 4 trophic magniﬁcation see biomagniﬁcation trout chlorinated hydrocarbon accumulation 195 effects of harmful algal blooms 302 response to oil spills 150, 151 tuna, size of individual ﬁsh caught 261 turtles as bycatch 262–3, 270 ingestion of debris 327 Tuvalu 67n typhoid in Aral Sea region 84–5 ultraviolet (UV) radiation and coral bleaching 26–7 effects of exposure 41–2 increases in 29, 41–2, 350 United States, loss of wetlands 132, 133 urbanization 14–16 and eutrophication 284–92, 293 and need for fresh water 87 and sewage disposal 297 UV radiation see ultraviolet radiation vanadium properties 208 uptake and bioaccumulation 212

367

Venezuela, coral reefs 23 Venice acqua alta/ﬂood events 48, 51–3, 53 defence measures 55–6, 73–5 elevation contours 52 history and development 49–50 and sea level rise 48, 49–57 withdrawal of groundwater from aquifers 54, 54, 73–5 Venice Lagoon development 50–1 effect of ﬂood defence construction 55–6 usage 50 vinyl chloride 180 volcanoes and atmospheric temperature 30 as source of aerosols 28–9 Volga River 86 Wampanoag people 284 Waquoit Bay, Cape Cod 17, 284 eutrophication 286–92 nitrogen sources 296 phytoplankton 288, 289 seaweed 283, 288–90, 303–4 time course of land use 286 urbanization 285 wastewater disposal see sewage disposal water quality deﬁning standards 312–13 effects of exotic species 232, 233 restoration 313–18 water vapor as greenhouse gas 27, 28 watersheds, urbanization 15 waterworks and freshwater use 88–91 and sediment load 105–6 weather, sea level change related to 61 wetlands accretion 68–9 antipathy to 131 Aral Sea 81 Atchafalaya River 117 Black Sea 12–13 Ebro delta 108, 109 effects of reduced freshwater discharge 95, 96–8 effects of sea level rise 68–9 loss of 61, 69, 124–45 signiﬁcance 136–41 Nile delta 96, 96–7, 98n restoration 72–3, 141, 142

368

wetlandsm (cont’d ) and sediment load 108, 109, 116–17 United States 132, 133 see also mangrove forests; salt marshes whales acoustic foraging 333n chlorinated hydrocarbon accumulation 187, 189, 193 effects of Exxon Valdez oil spill 152 effects of harmful algal blooms 302 effects of noise 334 entanglement with debris 324, 330, 331n extinction 259n

INDEX

population recovery 269 protection 269 strandings 334 whaling 174, 175, 246, 247, 259n, 284 white stork, effects of reduced fresh water discharge 96–7 Willapa Bay, exotic species 231 winter ﬂounder, PCB contamination 177 xanthophyll 27n Yellow River, sediment load 109–10, 115, 116 yellowtail ﬂounder 254 Yenisey River, sediment load 115

zinc from waste incineration 221 properties 208 retention in estuaries 208 retention in salt marsh sediments 211, 210 sources 207, 207, 220, 220 time course of contamination 220, 220 uptake and bioaccumulation 212 zooplankton Black Sea 9–10, 11 effects of rise in sea temperature 39 passage through power plants 332 radioactive contamination 338 zooxanthellae, coral reefs 23, 25, 305 Zostera see seagrass meadows