Invasion Ecology
Invasion Ecology Julie L. Lockwood Martha F. Hoopes Michael P. Marchetti
© 2007 Julie Lockwood, Ma...
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Invasion Ecology
Invasion Ecology Julie L. Lockwood Martha F. Hoopes Michael P. Marchetti
© 2007 Julie Lockwood, Martha Hoopes, and Michael Marchetti 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 Julie Lockwood, Martha Hoopes, and Michael Marchetti to be identified as the Authors 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 2007 by Blackwell Publishing Ltd 1
2007
Library of Congress Cataloging-in-Publication Data Lockwood, Julie L. Invasion ecology / Julie L. Lockwood, Martha F. Hoopes, Michael P. Marchetti. p. cm. Includes bibliographical references and index. ISBN-13: 978-1-4051-1418-9 (pbk. : alk. paper) ISBN-10: 1-4051-1418-5 (pbk. : alk. paper) 1. Biological invasions. I. Hoopes, Martha F. II. Marchetti, Michael P. III. Title. QH353.L63 2007 577’.18–dc22 2006009954 A catalogue record for this title is available from the British Library. Set in 10.5/12.4 Minion by Newgen Imaging Systems (P) Ltd., Chennai, India Printed and bound in Singapore by COS 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
An Introduction to Invasion Ecology
1
2
Transport Vectors and Pathways
18
3
Trends in Numbers of Invaders
38
4
Propagules
59
5
Disturbance
76
6
Establishment Success: The Influence of Biotic Interactions
107
7
Modeling the Geographical Spread of Invasive Species
132
8
Ecological Processes and the Spread of Non-Native Species
158
9
Ecological Impacts of Invasive Species
184
10
Impact Synthesis
206
11
Evolution of Invaders
223
12
Prediction, Risk Assessment, and Management of Species Invasions
241
Literature Cited
269
Index
299
Preface
It has been our experience that placing the reams of observations, examples, and research within the broad context of invasion ecology gets more difficult as the field grows each year. Not only does relevance change with new insights, but the field itself requires an intense understanding of multiple disciplines. Invasion ecology by its nature is integrative, requiring its practioners to understand (at the least) economics, evolution, population genetics, biogeography, and ecology. Indeed, those studying biological invasions must not only understand these contributing fields, they must combine them in order to find ways to control the flow and impact of non-native species. This is a tall order. We discovered early in writing this book that we could not cover all aspects and examples of invasions within a single text. Not only are different disciplines integrated into the study of invasions, but different ecologists have distinct motivations behind their research. We ultimately decided to produce an overview of the invasion process from the point of view of an applied ecologist. Thus, we hope to highlight the salient research findings associated with each invasion stage (e.g., transport, establishment, spread, impact), while explicitly considering how these findings support societies’ efforts to prevent the influx of non-native species and control those that cause ecological harm. We would also like to humbly apologize to the many excellent invasion ecologists (including some of our friends) whose work we did not include. This textbook was designed to be used in senior-level courses for undergraduates and in graduate courses. We purposefully take a very broad view of biological invasions with the hope that students and instructors will find the book useful as a primer to the growing field of invasion ecology. We have tried to make the chapters accessible and interesting to read while conveying the history and complexity of current debates, and we offer readings that we think complement each chapter. We have assumed students have at least completed a basic ecology course and can read calculus equations without fainting (or we suggest they assume a seated position while reading Chapters 7 and 8). The boxes associated with each chapter are integral to the material there, thus we strongly recommend that students and instructors do not skip over them. Given our goal of producing a comprehensive textbook, readers will not find appreciable amounts of new material here, instead we hope they will find some insightful organization and interesting juxtapositions. The final version of this book represents three years worth of steady work on the part of the authors. Honestly, we have thoroughly enjoyed most of it, and the parts that were not so much fun were at least made tolerable by our “working” sessions. These sessions
Preface
vii
invariably lasted two days, involved lots of cookies and good wine or bad beer (only occasionally together), and were filled with much joking and laughter. For us, this book was well worth creating for the friendships its production forged and the insight we as authors gained while researching the topics; however, we certainly hope those that actually buy the book will find it worth their money. If we achieve any success with this text, it will have much to do with the help we have received from our families, partners, colleagues, and friends. In particular we would like to thank those colleagues who provided “friendly” peer review for each chapter. These folks are Kama Almasi, Tim Blackburn, Cini Brown, Phillip Cassey, Jeff Corbin, Curt Daehler, Susan Harrison, David Holway, Ruth Hufbauer, Theo Light, Kelly Lyons, Peter Moyle, Kristina Schierenbeck, Katriona Shea, Dov Sax, Dan Simberloff, and Betsy Von Holle. Our gratitude also goes out to those authors who shared original figures and photos with us including Towns Peterson, Dave Rizzo, and Marcel Holyoak. We owe our respective academic departments our thanks as they not only put up with our distraction but also provided the space we needed to work, laugh, eat cookies, and drink beer, and the reams of paper we needed to construct the final version of the book. We would also like to thank our students who had to put up with our busy schedules and distracted brains. We owe a special thanks to our original editor at Blackwell, Sarah Shannon who initially contacted Julie to propose the book, to our current editors there, Hannah Berry, Rosie Hayden, and Ward Cooper, to John Normansell for typesetting, and to the Blackwell production staff. Finally, we dedicate this book to Tanya, Eve, and Tabby and thank them for their patience, kindness and love. Without you three to support us we would never have made it to the finish line. Julie Lockwood – New Jersey Martha Hoopes – Massachusetts Michael Marchetti – California
1 An Introduction to Invasion Ecology Overview Public and academic recognition of the problems associated with biological invasions has grown exponentially over the past decade. The reasons for this growth are three-fold. First, the negative effects of some non-native species have grown too large to ignore. Thus, increasing numbers of scientists are studying and managing non-native species in an effort to minimize the effects of biological invaders on native species and human economies. Second, the number of species being moved out of their native ranges and into novel locations is itself growing. Therefore, not only are the problems caused by non-native species becoming blatantly obvious, but also the overall number of problems appears to be growing. Third, with so many invasive species, it is very hard to do ecological field research without encountering invaders and eventually including them in investigations even if those investigations are for basic research. Invaders offer some new insights, and it is very difficult for curious scientists to pass up the opportunity to explore these new avenues. This textbook has the somewhat lofty goal of providing an overview of the study of biological invasions. We cover everything from the “whys and hows” of the transportation of non-native species to the post-invasion evolutionary changes that some invaders undergo. In between these topics lies the bulk of subjects addressed by most currently published research on non-native species. We provide an organizational framework for understanding this literature and relate the major findings to societies’ efforts to stem the flow of non-natives and reduce the impacts of those non-natives that cause damage. Thus we are taking a decidedly applied approach to the topic of invasion ecology much in the same way that Charles Elton did in his classic work The Ecology of Invasions by Animals and Plants (Box 1.1). In this introductory chapter we more precisely define what we mean by a non-native species, settle on a general terminology for use throughout the book, and detail what we mean by the “invasion process.” As you will see, the decisions we had to make in regard to these issues are key to how you, the reader, should use the book to your best advantage.
2
Chapter 1
What are invaders and why do we care about them? One of the principal ways in which speciation occurs is through geographic isolation (i.e., vicariance speciation, Mayr 1963). Physical features such as oceans, mountains, ice sheets, and river valleys represent boundaries to the movement of individuals between populations of the same species. Over time these separated populations diverge via drift and selection, with each population eventually forming a unique species. Climatic and geological events sometimes remove these barriers and allow individuals representing these species to disperse over relatively long distances. During these events long-separated species intermingle, with some eventually expanding their ranges. For example, during the Miocene when the isthmus of Panama emerged from the sea to link North and South America, North American mammals tended to move south whereas birds and plants of the South American rainforests tended to move north (i.e., The Great American Interchange, Marshall et al. 1982). Such relatively rapid expansion of species groups is unusual enough to deserve special recognition, and paleontologists and ecologists have given these species a whole host of names (e.g., immigrants, waifs, colonizers; see Table 1.1). As long as humans have had the ability to disperse across continents, people have also helped to breach the geographic boundaries to species’ movement. Domesticated animals and plants have regularly been transported along with other goods as human settlers have moved into novel territory (Crosby 1986). Almost certainly representatives of non-domesticated species have hitched a ride within these people’s clothes, growing on their boats or wagons, and living within or on their domesticated animals. Like the groups of mammals, birds, and plants that expanded their geographical ranges in the Great American Interchange, these species encountered locations that were previously out of their reach, and some of them successfully colonized these novel environments. This time, however, these successful colonizers achieved this new distribution with the help of humans. In the same way that ecologists distinguish between human-mediated and natural extinction of species, they also distinguish between natural and humanmediated rapid range expansion. This book concentrates exclusively on those species that found their way out of their native range and into a novel non-native location via human actions. There is ongoing discussion within the scientific community as to whether range expansions aided by humans are substantively different from range expansions that follow shifts in the Earth’s paleoclimate (e.g., Vermeij 2005). We address this debate in Chapter 2. Suffice it to say here that, compared to natural range expansion, humans have massively increased the rate at which species invade new areas, and they have substantially changed the geographic patterns of invasion. From a purely practical point of view, however, whether there are differences in human-assisted versus natural invasions is a moot point. This is not to say that there is little to be learnt from examining invasion patterns in the paleontological record. It is never wise to ignore the lessons of history. However, there is ample evidence that non-native species can cause serious ecological and economic problems (Mooney et al. 2005). Invasive species eat, compete, and hybridize with native species often to the detriment of the natives. Invasion can result in the loss of native species and the loss of ecosystem services such as water
An Introduction to Invasion Ecology
3
Table 1.1 Terminology commonly used for non-native species in the English language. Native Non-native Transported Established Spread Impact Invasive
Adventive Alien Casual Colonizing Cryptogenic Escaped Established Exotic Foreign Immigrant Imported Introduced Invasive Naturalized Non-indigenous Noxious Nuisance Pest Ruderal Tramp Transformer Transient Translocated Transplanted Transported Waif Weed
✹
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✹ ✹ ✹ ✹ ✹ ✹ ✹ ✹ ✹ ✹ ✹ ✹ ✹ ✹ ✹ ✹ ✹ ✹ ✹ ✹ ✹ ✹ ✹ ✹ ✹ ✹ ✹
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Notes: The list was compiled from the authors’ readings of the invasion ecology literature. Other scientists may categorize the terms differently. This inconsistency of usage creates problems for synthesis and management. The columns indicate types of species and invasion stages to which the words apply. See Figure 1.2.
filtration, soil stabilization, and “pest” control. More directly, most agricultural pests are non-natives and many human diseases can be classified as non-native (Pimentel 1997). Invasive species clog waterways, impede navigation, destroy homes, and kill livestock and fisheries (Mooney et al. 2005). Whether or not we think modern invasions are historically unique, they demand our attention. The motivation for this book comes as much, if not more, from this practical source as it does from the more esoteric interest in invasion as an ecological fact of life. This approach follows directly from the seminal work of Charles Elton (Box 1.1), who was one of the first to consider biological invaders as key drivers of ecosystem change, and he takes a clear stance on their value (i.e., get rid of them). The other founding work on invasions, an edited volume by Baker and Stebbins (1965), takes a more neutral stance.
4
Chapter 1
The authors of this book see biological invaders as ‘probes’ into the inner workings of nature. The origin of non-native species is known, their arrival in a new location is often well-documented, and it is possible to directly collect information on their activities while they are adjusting (or not) to their new community. Each of these attributes allows the study of invaders to become a powerful tool in our ecological and evolutionary arsenal (Sax et al. 2005). Invasion ecology has swung back and forth through the years between these two views (Davis 2005). We recognize the value of invaders as unique sources of ecological and evolutionary knowledge in this text, but we definitely lean towards the Eltonian viewpoint. Indeed the one thing that separates this text from a basic overview of ecology is its focus on how non-native species interact with human society, and how we can use ecological knowledge to thwart the influx and impacts of invaders.
Box 1.1
Unknown legend
Charles Sutherland Elton (1900–1991) produced the foundational book on biological invasions, The Ecology of Invasion by Animals and Plants, in 1958, just less than ten years before his retirement from Oxford University in 1967. This classic book grew out of a series of three radio lectures Elton made for the British Broadcasting Corporation (BBC) under the title of “Balance and Barrier.” Within the book, Elton summarized a relatively obscure literature on the impact and spread of non-native species that, at that time, was largely confined to the disciplines of entomology and plant pathology (Southwood & Clarke 1999). Perhaps because of the original intent of the lectures to reach a larger non-academic audience, The Ecology of Invasion by Animals and Plants is a pleasure to read. Elton manages the nearly impossible task of melding scientific rigor with an engaging and often witty writing style. As Dan Simberloff relates in his Foreword to the 2000 reprinted edition, “A writer who can describe oysters as “a kind of sessile sheep” and characterize advances in quarantine methods by the proposition that “no one is likely to get into New Zealand again accompanied by a live red deer” is more than just a scientist pointing out an unrecognized problem.” The predominant feeling one gets after reading The Ecology of Invasion by Animals and Plants is that Elton presaged nearly all the arguments within the field, and indeed there are very few topics that we will cover here that were not originally discussed by him. In fact, one could argue that invasion ecologists are simply putting numbers and assigning mechanisms to the patterns Elton noticed nearly half a century ago. Of course, one could also make that argument for the majority of community and population ecology theories. Elton also wrote Animal Ecology (1927), Voles, Mice and Lemmings: Problems in Population Dynamics (1942) and The Pattern of Animal Communities (1966). Within these volumes Elton lays out the foundation for population cycles, food chains, pyramids of numbers, and the structure of communities. The first one of these books was written in 85 days when
An Introduction to Invasion Ecology
Box 1.1
5
Continued
Elton was in his late 20s (Southwood & Clarke 1999). If only we could all be so productive so early in our careers! Also presaging where ecologists stand today (especially as our work relates to biological invasions), Elton embraced the socio-political implications of his work and often steered his research agenda toward solving fundamental societal problems. He founded the Bureau of Animal Populations in Oxford during World War II in part to satisfy the need to reduce the loss of stored grains from over-abundant rodent populations (Southwood & Clarke 1999). He helped found the Nature Conservancy and sat on its Scientific Advisory Board until 1957 despite the fact that he was “allergic to committees” (Nicholson 1987 as quoted in Southwood & Clarke 1999). Elton made the critical connection between population cycles of mammals and human health concerns such as the plague, and, in an early example of cross-disciplinary collaboration, spent several years exploring the population cycles and parasites of voles and mice. Thus, perhaps his most lasting contribution is recognition that ecology was a distinct discipline from biology, and that ecologists had a large role to play in how society dealt with the problems it faced. In this sense, and in several other ways, the book you are reading is infused with the spirit of Charles Elton.
A brief history of invasion ecology As Mark Davis points out in his recent work on the history of invasion ecology, the topic “. . . would be a dream dissertation. . . for some history of science graduate student” (Davis 2005). The rapid ascendancy of the discipline over the last decade continues a rather long history of scientific interest in non-native species that goes back at least to the publication of Elton’s book in 1958. In addition, the list of scientists that have dabbled in invasion ecology over the decades reads like an all-star roster, including such ecologically famous names as Ernst Mayer, E.O. Wilson, and Rachael Carson. Much like other modern disciplines (or sub-disciplines, depending on your predilection), invasion ecology grew out of a variety of much older research foci including agriculture, forestry, entomology, zoology, botany, and pathology. Within each of these disciplines there were scientists grappling with the effects of non-native species. Foresters were concerned about non-native species that decimated natural and managed forested lands. Agricultural scientists (e.g., plant pathologists) were concerned about non-native species that reduced crop yields. Animal scientists were concerned about non-native species that killed or caused disease in livestock. Botanists were concerned about non-native species that transformed native plant communities. Other scientists found the growing number of non-native species in their home countries noteworthy and documented their presence and sometimes their interactions with native species. Because of this fragmented history, interest in non-native species is not tied to a particular country or region. There was substantial and somewhat simultaneous work on
Chapter 1
6
350
Number of citations
300 250 200 150 100 50 0 1975
1980
1985
1990
1995
2000
2005
Year
Figure 1.1 Number of citations returned from a search of the Science Citation Index between the years 1975 and 2005 that included the terms “invas*” and “ecolog*.”
non-native species occurring amongst European, North American, South African, and Australian biologists and natural resource managers throughout the 1950s and 1960s (Davis 2005). Since English has grown to be the language of science over the decades, much of the early work not published in English has escaped notice by contemporary biologists (Davis 2005). A rising interest in invasion ecology is manifest in the increasing number of contemporary publications related to invasive species. For a quick view of how interest in invasion ecology has changed through time we did a search in the Science Citation Index® for articles published between 1975 and 2004 that included as keywords “invas*” and “ecolog*.” The results are shown in Figure 1.1. From 1975 to 1985 there were no articles that utilized variations of “invasion ecology” in their keywords. This almost certainly reflects the inconsistent use of the term “invasion” prior to the mid 1980s, but it also indicates a somewhat scattered and diffuse interest in the field during this time (Davis 2005). After 1990 there was an exponential growth in such articles. Indeed an exponential line fits the data depicted in Figure 1.1 nearly perfectly (R2 = 0.9545), thus generating the best statistical fit to empirical data any of the authors have ever attained. There is a telling blip in Figure 1.1 around the mid-1980s that corresponds with the publication of the Scientific Committee on Problems of the Environment (SCOPE) volumes. This committee, founded by the International Council for Science in 1969, is an amalgamation of environmental scientists that seeks to develop syntheses and reviews on various environmental issues. SCOPE pulled together a large group of ecologists from around the world between 1982 and 1989 to document the problems that invasive species posed (Mooney 2005). A whole series of books and journal articles on invasive species were produced, as well as a generation of newly minted PhDs who saw biological invasions as a significant component to global change (Vitousek et al. 1996). Arguably, the SCOPE volumes signal the gelling of the study of non-native species into a coherent field. The language used in these volumes marks a move toward viewing invasions as a conservation concern, much like the viewpoint of Elton, and such term usage continues
An Introduction to Invasion Ecology
7
to be reflected in today’s literature. We count ourselves among these academic “children” of SCOPE and suggest that the need to produce this textbook comes from the expansion of research in invasive species generated by these early SCOPE publications.
The wicked terminological web we weave In many ways the “grassroots” growth of invasion ecology is one of the joys of working within the field, as it is inherently an interdisciplinary science and one that focuses squarely on applied problems. On the other hand, early scientists working on non-native species developed terminology largely unique to their disciplines. When these scientists finally began talking to one another, the terminological milieu that resulted was baffling. With the growing interest in invasions, inconsistent terminology has become a real problem. On the surface, much of the discussion over the use of terms is a problem in semantics. However, there are important distinctions in how different fields use the various terms and these distinctions can cause confusion when attempting to collate and synthesize results (Richardson et al. 2000b) or produce management recommendations (Shrader-Frechette 2001). We take a close look at terminology here so readers can clearly identify the limits to our use of various terms and can place the semantic traditions of their respective fields within our framework. Table 1.1 is a list of terms in the English literature on invasion ecology and follows Colautti and MacIssac (2004). Some of these terms are used interchangeably to describe the same concept. For example, “non-indigenous,” “exotic,” and “alien” are often used to signify that a species is not native to a particular location. In other cases, one term is used to represent a variety of slightly different things. For example, Richardson et al. (2000b) identified four unique applications for the term “naturalized.” These are (i) non-native species that are reproducing outside of human cultivation, (ii) the set of non-native species that are reproducing in natural or semi-natural settings, (iii) species that are found outside their native range but may or may not be successfully reproducing, and (iv) non-native species that have expanded their geographic ranges. The lack of a consistent meaning for the term “naturalized” represents a real problem when trying to assemble lists of species that demonstrate a particular trait (i.e., established but not expanding its range vs. established and spreading), and thus for subsequent analyses based on this trait (Richardson et al. 2000b; Daehler 2001). Another issue related to terminology is that a few of the words in Table 1.1 evoke anthropocentric concepts that are militaristic and pejorative. For example, a standard thesaurus gives the following synonyms for “invasion”: attack, assault, incursion, raid, foray, and offense. Of course it is unwise to assume that words co-opted by scientists have the same meaning in and out of the profession (Daehler 2001). Nevertheless, the connotations of such terms as “invasion” have opened the field to criticism (Simberloff 2003a; Brown & Sax 2004). It is possible to find overzealous interpretations of invasion terminology in the broader academic and public literature; however, we would recommend Simberloff (2003a), Brown and Sax (2004), Slobodkin (2001) and Cassey et al. (2005) for thoughtful discussions of both the scientific applications and the sociological implications. A number of ecologists have responded to criticisms of
8
Chapter 1
language by adopting neutral terms such as non-indigenous and non-native. We adopt this approach here with the use of “non-native” to describe species that were moved outside their normal geographic ranges via human actions regardless of their eventual impact on native ecosystems. It is hard to escape the fact that some non-native species do cause serious ecological and economic damage (Simberloff 2003a; Cassey et al. 2005), and it is thus helpful to label such species with a unique term. Thus, we use the term “invasive” to describe species that have a demonstrable ecological or economic impact. On the odd occasion where we vary from this terminology, we do so for ease of communication. There are only so many times you can write the word “nonnative” or “invasive” in one sentence before it becomes incomprehensible and sleep inducing. Others have argued that the word “invasive” should only be used to describe non-native species that have become abundant and expanded their geographic range (Richardson et al. 2000b; Daehler 2001; Colautti & MacIsaac 2004). This usage ties the term to a more easily recognizable biological phenomenon, geographic range expansion, as opposed to the subjective and sometimes nebulous notion of “impact.” Mark Davis and colleagues have argued that biological invasion is an extension of normal colonization processes such as succession, and thus the terminology used should reflect this fact (Davis & Thompson 2000; Davis et al. 2001). We recognize and discuss the problem of defining invasive species impacts in Chapter 10. However, an over-riding recognition in this text is that an “invasion” is really the end product of a series of stages, where each stage serves as a barrier to the passage of a species onto the next stage. We take the approach that by explicitly recognizing biological invasion as a process instead of a dichotomous classification (i.e., invasive or not) it becomes clear where and when standard ecological principles come into play and how society can most effectively manage invaders (see also Colautti & MacIsaac 2004). A final note on definition relates to the spatial scale over which individuals of a species are moved via human activities. It is often convenient to classify non-native species as only those that have been transported from one continent and introduced into another. This definition can be expanded to include other relatively discreet biogeographical regions such as oceanic islands. However, over time many species expand beyond the limits of their native range via human actions yet never leave their native continent or island. These “home grown” non-natives (Cox 1999) can cause substantial harm to ecosystems they enter (e.g., rainbow trout in the high mountain lakes of the Sierras; Moyle 2002) yet there is some reluctance to label them as “invasive” since they are still in some sense “native” to the regional species pool. Our definition of “non-native” includes species that have clearly been introduced outside of their native range via human actions even if they are still within their native continent (e.g., house finches introduced within North America). Some subset of these home grown nonnatives causes ecological and economic harm, and thus we label them as “invasive” (Chapters 9 & 10). We see no difference in the ecological patterns and mechanisms behind the expansion of these species as compared to those that were introduced further afield. Indeed, several species that may be considered home grown non-natives in their local region also have introduced populations outside of their native continent or island.
An Introduction to Invasion Ecology
9
Invasion stages
Transport
Transport Death or captivity
Introduction Establishment Fail
Establish Spread Remain
Spread
local Impact Impact Low
Human High perception
Figure 1.2 Invasion process model depicting the discrete stages an invasive species passes through as well as alternative outcomes at each stage.
The invasion process If we start from the beginning and work our way forward, we can readily see that non-native species must pass through at least three stages before they are able to inflict ecological or economic harm (Fig. 1.2, Box 1.2). All non-native species originally began as individuals that were picked up from their native range, transported to a new area, and released into the wild (Transport in Fig. 1.2). These individuals must then establish a self-sustaining population within their new non-native range, or else the population becomes extinct (Establishment in Fig. 1.2). An established non-native population may then grow in abundance and expand its geographic range, or else it remains small in numbers and local in distribution (Spread in Fig. 1.2). Typically it is only when the non-native population is widespread and abundant that it will cause some sort of ecological or economic harm, and thus earn the name “invasive.” Acknowledging the stages of the invasion process allows for the explicit recognition of the actions of humans either as facilitators or inhibitors of transitioning a stage (e.g., Kolar & Lodge 2001). For these reasons we adopt this simple model as the framework for the rest of the book. Although not quite as distinct as we portray them in Figure 1.2 (as we shall see later), these stages represent biologically identifiable steps along the path to becoming an invader. Passing between each of these stages requires overcoming several ecological barriers. Williamson (1996) was one of the first ecologists to portray biological invasion as a series of barriers. One of his key insights was the recognition that most species that have the chance to transit an invasion stage do not successfully do so. He and his colleagues collated several examples of this phenomenon and generally found that
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Box 1.2 In the navy Guam is the largest island of Micronesia (54,000 ha) and, like many islands, harbored a fantastic assemblage of native endemic birds, reptiles, and mammals. The vast majority of these species are now considered extinct, and the remainder are threatened with extinction. Although the colonization of the islands by Polynesians contributed to the loss of species, as did massive habitat conversion and pesticide use following World War II, the principal culprit in this extinction drama is the invasive brown treesnake (Boiga irregularis). The following account of the brown tree snake in Guam is based heavily on the excellent summary of Fritts and Rodda (1998). The modern era for Guam began with the US Navy’s use of the island as a staging ground for a planned invasion of Japan during World War II. Guam’s human population grew more than 10-fold between 1944 and 1945. With the end of World War II in 1945, such a large naval base proved useful for salvaging derelict war materials from the region. In particular, the US Navy transported vehicles, aircraft, and other supplies from New Guinea where these items may have sat unused for some time. The brown treesnake is native to New Guinea and other regions of Australasia. It is typically nocturnal and, during the day, rests within crevices and holes that provide good cover. It is commonly found in wheel-wells on airplanes, under the hoods of cars, and in boxes of cargo. For this reason, most biologists think the brown treesnake was a stow-a-way within surplus Navy equipment that was brought to Guam from New Guinea in the post-war years of 1945–1950. The brown treesnake was first noticed along Guam’s southern town of Apra Harbor. It then spread, somewhat slowly, until it occupied the entire island by 1970. Guam has no native snakes and no native predators that could have controlled treesnake numbers. Instead, the native vertebrates were all small species vulnerable to predation by the generalist treesnake. Because the native animals were evolutionarily naïve to such predators, they were easy targets. The decline of most native forest animals was immediate and dramatic. Guam’s Division of Aquatic and Wildlife Resources discontinued surveys of native birds and bats in the 1970s, as there were no more individuals of these species to count. The three native birds and mammals that persisted the longest were the Mariana fruit bat (Pteropus mariannus mariannus), Guam rail (Rallus owstoni), and island swiftlet (Aerodramus vanikorensis bartschi). The fruit bat and rail were relatively long-lived species that likely persisted simply because some individuals escaped predation and lived out the remainder of their lives; however neither species had any reproductive success in the presence of the tree snake. The island swiftlet persists to this day but is confined to one cave on Guam. The swiftlet builds nests on the walls of caves using its own saliva and mud to adhere the nest to the cave wall. Despite this unusual habit, swiftlets are still vulnerable to treesnake predation as treenakes can easily capture prey in total darkness and climb most surfaces. The one cave where swiftlets persist is unique in that the cave walls are not textured enough for snakes to support themselves. This does not inhibit the swiftlets’ ability to build nests, but it does prohibit treesnakes from reaching those nests.
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Box 1.2
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Continued
There have been many other losses of native forest animals on Guam. Consequently, the food web of Guam’s forests (the dominant native habitat) has shifted dramatically through time (Figs 1.3, 1.4, 1.5). The most striking change is the reorganization of the web from one in which most components are native (plain type in figures) to one in which most components are non-native (italics in figures). Beyond the brown treesnake, the curious skink (Carlia fusca), Polynesian rat (Rattus exulans), house mouse (Mus domesticus), Phillipine turtle-dove (Streptopelia bitorquaata) and mutilating gecko (Gehyra mutilata) have all successfully established. Three native skinks survive in Guam’s forests, the blue-tailed skink (Emoia caeruleocauda), morning gecko (Lepidodactylus lugubris) and house gecko (Hemidactylus frenatus). All of the skinks are small and diurnal. One would assume that, with the destruction of the native foodweb, the brown treesnake would suffer from a lack of food resources,
Carnivores
1945 Nocturnal arboreal-carnlvore
Piscivore Fairy tern
Diurnal omnivore-carnivore Mangrove monitor
Diurnal insectivore-carnivore Micronesian kingfisher
Nocturnal arboreal insectivore-carnivore
Oceanic gecko
Insectivores
Nocturnal terrestrial insectivore-carnivore
Nocturnal terrestrial insectivore
Pelagic gecko
Terrestrial omnivore *Polynesian rat
Herbivores
*House mouse
Nocturnal arboreal insectivore
Diurnal terrestrial insectivore
Mutilating gecko
Blue-tailed skink Mariana skink
Nocturnal omnivore
Black rat
Spotted-belly gecko Mourning gecko
Diurnal arboreal insectivore
Moth skink Rufous fantail
House gecko Diurnal omnivore Guam rail Mariana crow
Herbivore
Diurnal arboreal omnivore Bridled white-eye Micronesian honeyeater
Micronesian starling *Philippine turtle-dove
Mariana fruit bat White-throated ground-dove Mariana fruit-dove
Figure 1.3 Typical vertebrate food web for Guam pre-1945. Species in italics = introduced species, asterisks = historic introduction. Type size indicates relative abundance (biomass) using order of magnitude increments. Boxes are labeled with predominant trophic role for the included taxa. Arrows indicate known trophic pathways. Reprinted with permission from Fritts, T. H. and G. H. Rodda, 1998. The role of introduced species in the degradation of island ecosystems: a case history of Guam. Annual Review of Ecology and Systematics 29: 113–140.
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Box 1.2
Continued
but it survives because the three native skinks and the non-native animals maintain populations despite the predation pressure of the treesnake. The available pool of invasive prey and native skinks maintains treesnake densities and keeps predation pressure high on the susceptible natives. Consequently natives quickly went extinct with no corresponding negative effect on the treesnake. The treesnake even seems to be adapting to the diurnal habits of the remaining native skinks. Biologists have documented a shift in the treesnake’s activity patterns toward one of solely nocturnal movements to one that increasingly includes activity during daylight hours. The reduction in the complexity of the Guam food web has had consequences beyond the loss of the native animals. The loss of mammalian and avian insectivores presumably increased insect abundances at some cost to agricultural crops 1965 Carnivores
Nocturnal arboreal carnivore *Brown treesnake
Piscivore Fairy tern
Diurnal omnivore-carnivore Mangrove monitor Nocturnal arboreal Insectivore-carnivore
Diurnal insectivore-carnivore Micronesian kingfisher
Oceanic gecko
Insectivores
Nocturnal terrestrial insectivore-carnivore *Musk shrew Nocturnal terrestrial insectivore
Spotted-belly gecko Mourning gecko
Diurnal terrestrial insectivore
Pelagic gecko
Terrestrial omnivore
*Polynesian rat *House mouse
Nocturnal arboreal insectivore
Blue-tailed skink Mariana skink
Nocturnal omnivore
Black rat
Multilating gecko House gecko
Diurnal omnivore Guam rail Mariana crow Herbivore
Diurnal arboreal Insectivore
*Green anole Moth skink Rufous fantail *Black drongo
Diurnal arboreal omnivore Bridled white-eye Micronesian honeyeater
Herbivores
Micronesian starling *Philippine turtle-dove White-throated ground-dove Mariana fruit bat Mariana fruit dove
Figure 1.4 Typical vertebrate food web for Guam in 1965 after the brown tree snake was introduced. Note the large number of taxa preyed upon by the tree snake. Species in italics = introduced species, asterisks = historic introductions. Type size indicates relative abundance (biomass) using order of magnitude increments. Boxes are labeled with predominant trophic role for the included taxa. Arrows indicate known trophic pathways. Reprinted with permission from Fritts. T.H. and G.H. Rodda. 1998. The role of introduced species in the degradation of island ecosystems: a case history of Guam. Annual Review of Ecology and Systematics 29: 113–140.
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Continued
and to crop production. Newly introduced non-native species may find it easy to invade Guam given the many “open niches” left by the loss of native species. These newly arriving natives must have some evolutionary experience with generalist predators to survive, however. Post-treesnake invasion, Guam seems to have much higher densities of web-building spiders than nearby islands free of treesnakes. Spiders that place obvious filaments in their webs to avoid bird damage to the web do so less often on Guam than on nearby islands that still have native birds. Also, pollination and seed dispersal services of native birds and mammals to plants have been lost. Similarly, we should expect to see reduced frugivory and herbivory of native (and invasive)
Carnivores
1995
Nocturnal arboreal carnivore
*Brown treesnake
Piscivore
Diurnal omnivore-carnivore Mangrove monitor
Diurnal insectivore-carnivore
Nocturnal arboreal insectivore-carnivore
Insectivores
Nocturnal terrestrial insectivore-carnivore
Nocturnal arboreal insectivore
Diurnal terrestrial insectivore Nocturnal terrestrial insectivore
*Curious skink
Diurnal arboreal insectivore
Multilating gecko
Moth skink
Blue-tailed skink
Nocturnal omnivore Terrestrial omnivore
*Polynesian rat *House mouse Herbivores
Mourning gecko House gecko
Diurnal omnivore
Black rat
Diurnal arboreal omnivore Herbivore
*Philippine turtle-dove
Figure 1.5 Typical vertebrate food web for northern Guam in 1995. Note the large reduction in both biomass and trophic functional roles among the remaining taxa. Species in italics = introduced species, asterisks = historic introductions. Type size indicates relative abundance (biomass) using order of magnitude increments. Boxes are labeled with predominant trophic role for the included taxa. Arrows indicate known trophic pathways. Reprinted with permission from Fritts. T.H. and G.H. Rodda. 1998. The role of introduced species in the degradation of island ecosystems: a case history of Guam. Annual Review of Ecology and Systematics 29: 113–140.
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Box 1.2
Continued
plants on Guam in the absence of most avian and mammalian herbivores. Such indirect effects of the brown treesnake invasion still require full documentation and exploration. There are now massive efforts to control or eradicate the brown treesnake from Guam. Some of these efforts, including killing the snakes by implanting aspirin in mice carcasses, have been successful, but they do not restore the native food web. Nor do they help to control the spread of the brown treesnake to other susceptible islands. Ships and airplanes from Guam require careful scrutiny to guard against introduction of the brown treesnake into other susceptible ports-of-call. Although the brown treesnake may not cause such damage in other areas, we should be especially cautious of this species becoming established on other islands rich in endemic species that are largely naïve to predators (e.g., Hawaii, Caroline Islands).
between 5 and 20% of species successfully transit an invasion stage, with the average being 10%. Williamson (1996) dubbed this pattern the Tens Rule, and it suggests two insights. First, the barriers between invasion stages are formidable, and thus there is a lot to learn about why some species successfully negotiate these barriers and others do not. Second, the nested structure of the invasion process ensures that only a small fraction of all the species introduced outside their native range will eventually cause some ecological or economic harm. Thus, the task set before ecologists is to find a way to sift through all the non-natives that will not cause harm to find the few that will. As we will see, these two insights simultaneously make the study of invasion ecology rewarding and frustrating. Beyond these insights, viewing the invasion process as a series of stages has a couple of key advantages. First is the recognition that it is something of a misnomer to label a species as invasive or not. Individuals of particular species are transported and released into new environments, and it is this set of individuals that must survive and reproduce if the non-native population is to persist (Chapters 2 & 3). In cases where individuals of one species have been transported to a variety of different places, it is possible that some of these incipient populations established, became widespread, and affected native biota (i.e., became invaders), while other populations did not. Thus, one species may be considered “invasive” in one location but not in another. Of course, the individuals of a species share a set of common traits that may allow all non-native populations of this species, on average, to become invasive. However, a shift in focus from species to populations opens doors to our understanding of invasions that would otherwise be closed (Colautti & MacIsaac 2004; Lockwood et al. 2005). Second, using the invasion process model of Figure 1.2 allows us to pull away from simplifying dichotomies contained in questions such as, “Do species traits or community properties determine the success of an invader?” There is no simple answer to that question as we will see. Non-native populations confront a host of extrinsic forces that determine whether they will persist into the future (Chapters 4, 5 & 6). These forces
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include competition, predation, parasitism and other interactions between individuals of two (or more) species. They also include stochastic physical forces such as flooding, freezing, and fire. How non-native populations respond to these forces depends on their life history. One set of extrinsic forces modifies the effect of the other, and the species themselves can modify the magnitude and extent of the extrinsic forces. This interaction makes it nearly impossible to gain a clear picture of what determines success and failure of a non-native population by focusing on one side of the dichotomy or the other. Furthermore, properties of the community, environment, and species often play a role in the successful transition of non-native populations across all invasion stages although they may take different forms and perhaps vary in their importance (Kolar & Lodge 2001; Cassey et al. 2004b). The waxing and waning in importance of various ecological mechanisms across invasion stages adds yet more confusion when trying to answer simple dichotomous questions (Lockwood et al. 2005). Third, the invasion process model allows us to see more clearly where the actions of humans interact with the fate of non-native populations. By our definition, humans begin the invasion process by purposefully or inadvertently transporting individuals beyond their natural range limits and then releasing them into the wild (Chapters 2, 3 & 4). This aspect of the invasion process has often been divorced from the rest of the pathway, despite the growth in our knowledge of this stage (Puth & Post 2005). Humans also enter the picture through their effects on the ecosystem in which the non-native population is attempting to establish (e.g., causing disturbance, Chapter 5). They can facilitate the geographic range expansion of an established non-native population via their commerce or travel although this is not necessary for a non-native population to spread (Chapters 7 & 8) or to cause some type of harm (Chapters 9 & 10). Finally, human actions serve to strengthen the barriers between invasion stages through the use of direct eradication techniques or through socio-political actions that slow the transportation of non-native individuals (Chapter 12). A disadvantage to using the invasion process model as a structure for this book is that we divvy up discussions of key ecological processes such as facilitation and competition between chapters. There are insights to be gained from concentrating on one ecological force, such as competition, and considering its overall effects throughout the invasion process (Shea & Chesson 2002), but, ultimately, we found it more compelling to follow the invasion process model in order to look at how several forces might interact at any one stage. There are many papers that examine the role of single forces in the success of an invasive species, and we reference these papers in the hope that readers will benefit from both approaches.
How to use this book As Figure 1.1 suggests, it is increasingly difficult to encapsulate the state of our understanding of biological invasions. Halfway through 2005, there were already 257 articles in the Science Citation Index® that included some form of “invasion ecology” in their keywords. Thus, in writing this book we strove to provide the reader with a clear overview of the variety of topics being discussed within invasion ecology today and
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to tie them together in a logical and informative way. We hope this provides interested scientists with a type of roadmap to the literature and to the inter-relationship of various concepts discussed within invasion ecology. To further this effort, at the end of each chapter we provide a set of suggested articles for use as companion readings. These “Companion Papers” are articles that delve more deeply into the issues we present, or are key syntheses of relevant literature on a particular topic. We suggest that those using this book in a graduate or undergraduate class incorporate these articles into their reading assignments. These suggested companion papers should make student discussions of the topics we present in each chapter a bit more lively. We also provide a set of articles for those readers that would like to conduct research on a topic we present. These “Further Complexities” are the books and articles that laid the foundation for various aspects of invasion topics. These are the works we suggest as “must reads” to anyone interested in pursing invasion ecology as part of their graduate education. Others may disagree about which articles we included in these lists, but at the very least they provide a starting point for the student just entering the field. We made a few key decisions concerning the breadth of material we cover that may leave some readers frustrated. Invasion ecology interfaces with a variety of fairly distinct subdisciplines such as biocontrol, agro-ecology, weed science, epidemiology, ecological economics and environmental law. We do not cover these particular topics in any real depth here, but we do point the way for interested readers. Our decision not to extensively cover the area of biocontrol is likely the most open to criticism. We do consider some examples of biocontrol in detail; however, it quickly became apparent that the traditions and experience of biocontrol practitioners were somewhat specific and therefore differed from those of many more broadly defined ecologists, and we simply could not do justice to this history within the page limits of this book. There is so much to explore in the crossover between invasions, biocontrol, and conservation biology that it deserves its own consideration. Finally, we made the decision to pull two topics out of our framework and present them as cross-stage syntheses. These concern the evolution (Chapter 11) and management (Chapter 12) of non-native species. This was a tough decision on both counts. These topics could easily be incorporated into each of the invasion stages we use as the structure for the book. For example, management options change drastically as one moves from the transport and release stage to the post-establishment stages. Ultimately we felt there was more to be gained from presenting these topics as synthetic wholechapters as there were key threads within each subject that were hard to maintain in other organizational structures. We have endeavored not to break too much new ground in this book but instead hopefully strike a balance between breadth and depth in coverage of invasion ecology. If we have broken new ground at all, it was because it was necessary in order for us to present a coherent picture of the overwhelming variety of research results in one area. We have also tried to incorporate a variety of examples that cross taxonomic groups and geopolitical boundaries. However, most readers will quickly recognize the following biases: (i) we predominately study fish, birds and plants; (ii) we live in North America; and (iii) at some point in our lives we have each lived and worked in California. Such
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biases are perhaps unavoidable, and we welcome input from our colleagues around the world just in case a second edition of this book is warranted.
Further complexities Brown, J. H., and D. F. Sax. 2004. An essay on some topics concerning invasive species. Austral Ecology 29: 530–536. Cassey, P., T. M. Blackburn, R. P. Duncan, and S. L. Chown. 2005. Concerning invasive species: reply to Brown and Sax. Austral Ecology 30: 475–480. Crosby, A. 1993. Ecological Imperialsim: The Biological Expansion of Europe, 900–1900. Cambridge University Press, Cambridge, UK. Richardson, D. M., P. Pysek, M. Rejmanek, M. G. Barbour, F. D. Panetta, and C. J. West. 2000. Naturalization and invasion of alien plants: concepts and definitions. Diversity and Distributions 6: 93–107. Shrader-Frechette, K. 2001. Non-indigenous species and ecological explanation. Biology and Philosophy 16: 507–519.
Companion papers Colautti, R. I., and H. J. MacIsaac. 2004. A neutral terminology to define “invasive” species. Diversity and Distributions 10: 135–141. Davis, M. A., K. Thompson, and J. P. Grime. 2001. Charles S. Elton and the dissociation of invasion ecology from the rest of ecology. Diversity and Distributions 7: 97–102. Shea, K., and P. Chesson. 2002. Community ecology theory as a framework for biological invasions. Trends in Ecology and Evolution 17: 170–176. Simberloff, D. 2003. Confronting introduced species: a form of xenophobia? Biological Invasions 5: 179–192.
2 Transport Vectors and Pathways
Overview All invasions begin with a set of individuals (sometimes as few as one) being taken from their native range, transported to a novel location, and released. A transport vector is the conveyance that moves a nonnative propagule to its novel location, and a transport pathway is the route between the source and release location. From the few available estimates of the numbers of species moved around the world via human-mediated transport vectors, we know the numbers are huge. This chapter endeavors to characterize these early events in the invasion process. We tackle the question of the similarity between natural and human-assisted invasions. We then characterize the types of transport vectors, making an initial distinction between species purposefully transported and those transported inadvertently. We also take a look at the dynamics of transport pathways, which highlights emergent patterns of interest such as the hub-and-spoke model of species transport. Understanding these vectors and pathways is critical for all subsequent stages of invasion ecology research.
What’s the difference between a vector and a pathway? The terminology of this invasion stage is still evolving; however, we have adopted the term usage of Mack (2004). We define a transport vector as the manner in which species are carried along a pathway, and a pathway as the route between the source region of a non-native species and its location of release. To understand the difference, consider the triangular commercial sailing route of the late 1800s between Europe, West Africa, and the Caribbean (Fig. 2.1). This pathway developed, in part, because the ships could sail on the trade winds, which were themselves produced by the Gulf Stream, Canary Current, and South Equatorial Current (Mack 2004). This triangular trade route carried a phenomenal diversity of goods between the three locations, including a wide variety of plants, animals, insects, and diseases (Crosby 1986). The vectors in these cases are the sailing ships themselves; whereas the pathway is the route that these ships traversed. Ships, as vectors, carried non-native species across many pathways beyond these three
Transport Vectors and Pathways
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Britain
m Raw
ater
ial s
Manufactured goods
North America
The Caribbean
N W
Sl a
ve
s
E
West Africa
S
Figure 2.1 Triangular trade routes between Europe, West Africa, and the Caribbean. These routes transported a large variety of goods including non-native species.
regions. For this reason invasion vectors are often classified into broad groups, as we do below. The pathways of this triangle were traversed by vectors other than sailing ships, like steam powered vessels and airplanes, within the span of another 50 years. Thus pathways can wax and wane in their importance through time following changes in commercial trade patterns or in the technological innovations of transport vectors. Although the two aspects of transport are clearly linked to one another, by disconnecting them we can explore the influence of each on subsequent invasion patterns.
Does human-mediated transport differ from natural colonization? Before we move to a discussion of transport vectors and pathways, we take the opportunity to consider how human-assisted transport vectors differ from natural vectors. Like species extinction, species invasion is a natural event. Throughout the earth’s history species have consistently found opportunities to expand their ranges, and sometimes this has happened over a relatively short period of time (Vermeij 2005). Even today, we know that species can traverse seemingly large distances by “riding” prevailing winds or ocean currents (Bush 1991; Mack & Lonsdale 2001; Mack 2004). The principal question with invasions is whether human-mediated transport is qualitatively different from natural colonization events. The rates of species arrival to oceanic islands are helpful in answering this question. In these isolated locations we can calculate the historical rate of invasion necessary to account for the indigenous species richness, and compare this rate to known modern
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rates of introduction. For example, the National Research Council (NRC 2002), based on information from Carlquist (1974) and Fosberg (1948), calculated that the Hawaiian native plant flora could have arisen from a natural dispersal rate of one species every 100,000 years. Hence, the entire native Hawaiian flora of 1094 species could have originated from about 280 successful natural colonization events with subsequent adaptive radiation (Wagner et al. 1990). The rate of addition of plant species to Hawaii accelerated to one new plant species every 50 years after the arrival of Polynesians, and to one new species every 22 years after the arrival of Europeans. The end result is a near doubling of the number of plant species persisting on the Hawaiian Islands as compared to what may have naturally dispersed and evolved there (Sax et al. 2002). Such an increase in the rate of plant invasions to Hawaii is driven by profound changes in the vectors that delivered these species. Prehistorically, plants principally found their way to Hawaii via wind dispersal or by floating to the island on ocean currents. Plants began arriving in Hawaii via boats after the Polynesians arrived. These boats came from other Pacific islands that the Polynesians occupied. Since such travel required sailing or rowing, relatively few Polynesians or boats made the trip. Today, plants arrive in Hawaii via ships and airplanes that visit the islands many times a day. These more recent vectors originate from all over the world, but most frequently from Japan, other Pacific Islands, and North America. We see a similar pattern if we examine winged insects on Gough Island, which is located in a very remote section of the southern Atlantic Ocean (Gaston et al. 2003). There is no evidence that Gough Island was occupied by humans prior to the landing of a whaling ship there in 1675. Since this initial landing, Gough Island has been a regular seal hunting site. South Africa placed a permanent, and still active, weather station on the island in 1956. Using an unusually complete list of insects introduced to the island since the late 1600s, Gaston et al. (2003) calculated that winged insects have been introduced to Gough Island via human-mediated vectors at a rate of 218 species per thousand years. In contrast, these authors calculate that one successful colonization event every 95,000 years sufficed to account for the indigenous insect fauna. The transport vector for the indigenous insects was probably westerly wind currents. The primary vector for the introduced insects is vessels that have visited the island to transport sealers and scientists and all their equipment. By no means is Gough Island a transportation hub. Even with this small amount of human traffic, however, the colonization rate for the island was raised by two to three orders of magnitude over the calculated historical rate (Gaston et al. 2003). Both the pathways and vectors by which non-native species reach their destination appear to be much more diverse and dynamic than natural colonization processes. In the few cases where ecologists have been privy to the re-colonization of an area after a serious destructive event (e.g., volcano eruption), the species that colonize are generally all from nearby source populations on islands or mainlands and can be predicted based on prevailing wind and water currents, the life-history characteristics of the species themselves, and the successional dynamics of the recovering flora (e.g., Thorton 1996). In contrast, the geographic scale of modern transport pathways is much larger, allowing species from one side of the globe to establish on the other side. These human-derived transport pathways may fade in and out of existence much faster than natural pathways
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following the whims of human desires and economies (Mack 2004). In addition, human vectors allow these globe-trekking species to move from one location to the other in a matter of hours instead of months or years. The set of species that globe-treks via human transport vectors is also ecologically more varied than the set of species that disperses long distances naturally. For example, for a plant to be transported to New Zealand via a natural vector it is necessary that it possess a suite of characteristics that allow it to float, withstand salt water, and remain viable after long periods of unfavorable conditions. A plant introduced to New Zealand via airplane traffic need not possess any of these traits. In summary then, we can think of human-mediated transport as being faster, more dynamic, and often encompassing a larger geographical scale than natural colonization processes. Not all ecologists agree that human-mediated invasions differ fundamentally from natural colonization events (see Brown & Sax 2004; Vermeij 2005; cf. Cassey et al. 2005). Although we think there is much to be learned from the study of natural colonization processes and paleontological biotic interchanges, the rates, scales, and mechanisms of human-mediated transportation of species differs enough from natural events to warrant careful scrutiny. Even if we discount the differences between natural and human-mediated colonization events, we may still find human-mediated transportation patterns of interest because of the large negative impacts associated with species introduced by humans (Chapters 9 & 10).
Transport vectors A basic dichotomy There are many potential ways to categorize transport vectors of non-native organisms: taxonomically, by transportation scheme, or by geopolitical boundaries. We have chosen a basic dichotomy of intentional versus unintentional introductions because it cuts through all of these categories. Intentional transport refers to the movement of organisms to a new area for an express purpose. This intentional movement of species includes a multitude of vector categories like the importation of ornamental plants, crop species, pets, biocontrol, and so on. Because intentional introductions reflect human desires, we typically have written records of when and where they have occurred. In contrast, unintentional transport refers to the movement of non-native species as a secondary result of the intentional transfer of another product, good, or service. We typically have a much less detailed knowledge of unintentional introductions.
Intentional transport Perhaps it is easiest to understand the intentional transport of non-native species. Intentional movement of organisms by humans has been occurring since ancient times. For example, plants have been moved and cultivated since the dawn of civilization, perhaps as long ago as 8000 years before the common era (bce) (Reichard & White 2001). China for example, has a long historical written record of the transport and
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introduction of plants and animals along the ancient trade routes (Yan et al. 2001). As early as the fourth century bce, the Shu-Yuan Du trade route linked India and China, and there are written accounts of tamarind (Tamarindus indica) being introduced and grown in Chinese territory (Yan et al. 2001). There are as many reasons for intentionally moving organisms as there are people to do the moving. Thus, it is often helpful to place these reasons into broad, inter-taxonomic categories that may ultimately allow us to make useful generalizations.
Food and game Some of the earliest movement of species occurred with the beginning of managed cropland and agriculture. The majority of the fruits, vegetables, and seasonings people consume (Reichard & White 2001), as well as meat and dairy products, are derived from taxa that have been intentionally introduced from other areas. Included in this transport category are cattle, sheep, chickens, wheat, tomatoes, maize, potatoes, bees, fish, shellfish, bullfrogs, and hundreds of other species. Many of these taxa rely on humans for their continued propagation. Tomatoes for example do not seem to become invasive species on their own. Nonetheless, some of these species have escaped from the confines of the farm, ranch, or aquaculture facility and established free-living, non-native populations. For example the first European settlers of the Americas brought with them the crops, medicinals, herbs, and domesticated animals they felt were necessary for their immediate survival in an unknown, hostile environment (Mack 2003) and subsequently gave very little thought to the fate of these organisms outside the confines of their settlements. Relatively quickly some of these species became established outside of the immediate settlement environment. As early as 1634, William Wood, an early North American colonist in Plymouth Bay, Massachusetts, recorded that at least two introduced species (sorrel, Rumex acetosella, and yarrow, Achillea millefolium) had escaped cultivation and grew freely in the surrounding forest (Mack 2003). Similarly, in the early nineteenth century, St John’s Wort (Hypericum perforatum) was introduced for medicinal purposes into New England and the mid-Atlantic states of the United States and quickly became a major pest of pastures and fields (Mack 2003). Today St John’s Wort is listed as a noxious weed in several western US states and is considered a serious pest around the world. This type of movement also includes forage organisms (plants and animals) intentionally transported to provide food for farm animals. For example, kudzu (Pueraria montana var lobata) was introduced into the United States in 1876 at the Philadelphia Centennial Exposition, where it was promoted as a forage crop and as an ornamental plant (Miller & Edwards 1983). Private individuals and state officials subsequently planted it over much of the southeastern United States. Kudzu was eventually removed from the list of acceptable cover crops in 1953, and by 1970 it was recognized as a weed by the US Department of Agriculture (Miller & Edwards 1983). Food-related transportation vectors are not limited to plants and mammals. The aquaculture and seafood industry purposefully imports live animals and breeds them for sale. Some of these species escape the confines of the facility or are released by facility owners. For example, the red swamp crayfish (Procambarus clarki) was first introduced
Transport Vectors and Pathways
23
into California in 1932 by a frog farmer in San Diego County as forage for his frogs and was later moved around the state to be grown for food (Riegel 1959). A few individuals escaped these early confines and the species is now extremely abundant and widely distributed in the western United States and Hawaii. Similarly, Siguan (2004) reports that 64 species of marine plants associated with aquaculture have been introduced outside their native range. Cowie and Robinson (2004) describe several examples of freshwater snails establishing non-native populations throughout the Pacific Islands and the Americas after having been intentionally brought to these localities as food. Far from being an “old” or defunct vector, the lure of establishing local populations of rare or expensive food items is still quite strong. There is evidence that Chinese mitten crabs (Eriocheir sinensis) were intentionally introduced into the San Francisco Bay by local citizens who wanted to establish a cheap and readily available supply of this favored Asian food item (Weigle et al. 2005). Cowie and Robinson (2004) describe the transport of European snails (Helicidae) used as escargot within personal baggage or sewn into the linings of clothing. These snail-smugglers intended to release the prized individuals in their backyards or in vacant lots, allow the populations to increase, and then harvest them for consumption or sale. Mack (2004) laments the resurgence of backyard herb gardens because the burgeoning trade in seeds includes the sale of species known to cause ecological and economic harm. Another important yet closely related category of intentional transport is game species. These are species that are introduced for the sole purpose of providing fishing or hunting recreational opportunities. Historically this type of introduction was thought of as wildlife enhancement, an improvement on the existing natural fauna of an area. For example, the European population of the western United States decided they did not want to fish for the native “rough” fish, which were primarily minnows in the family Cyprinidae. Instead they wanted to fish for species that were more familiar to them, namely eastern United States game species such as brook trout (Salvelinus fontinalis), the black basses (Micropterus spp.) and the catfishes (Ictalurus spp.) (Moyle 2002). Perhaps the best example of this type of movement involves the now virtually cosmopolitan game fish, the rainbow trout (Oncorhynchus mykiss). This species of salmonid is native to the streams and rivers of the western United States. By 1996, the species had been transported to every continent around the globe and more than 100 countries and oceanic islands primarily for game purposes (Lever 1996). Along with the importation of game fish species, came the importation of species used to catch these fish. Many freshwater fish and invertebrates are transported to new areas as bait for sportfishing and subsequently released at the end of the day or season. In her summary of aquatic introductions into North America, Fuller (2004) indicates that species such as rudd (Scardinius erythrophthalmus), red shiner (Cyprinella lutrensis), golden shiner (Notemigonus chrysoleucas), and fathead minnow (Pimephales promelas) all have very broad distributions in the United States because they are used as baitfish. Fuller (2004) also reports that salamanders (Plethodontidae) endemic to the eastern United States may be used as bait within the western states. This so-called “bait-bucket” transport is very common and can act to rapidly spread a species over a wide geographic area.
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Chapter 2
Similar patterns exist for the importation and release of birds and mammals to enhance game opportunities. For example, mallards (Anas platyrhynchos) were transported out of the mainland of North America and released into places as far away as Hawaii, New Zealand, Australia, and Tahiti in an effort to establish readily accessible game populations (Long 1981). Dozens of species of quail and pheasants have been shipped across oceans to be released as game species (Lockwood 1999). The same is true for the importation of various species of deer (Cervidae) (Lever 1985). Some of these populations failed to establish, and those responsible for the transportation and release of them gave up. Other populations persisted in the face of heavy harvesting only because of repeated releases by interested private citizens and governmental agencies (Pimm 1991). Yet other species managed to establish self-sustaining populations despite being harvested (Long 1981).
Environmental enhancement Many species have been moved intentionally to enhance natural environments. We define enhancement as efforts to stock local environments with species that are aesthetically or culturally satisfying to some portion of the human population. This vector category is quite broad, including ornamental and horticultural plants, the importation or breeding of pets, and the activities of acclimatization societies (see Box 2.1).
Box 2.1 Fly Robin fly According to the Oxford English Dictionary, “acclimatize” means to habituate to a new climate or environment. Acclimatization societies, then, were groups of individuals that endeavored to import species into their home countries in an effort to have these species in their vicinity. Such societies reached their zenith in the early twentieth century, perhaps in part due to the publication of Darwin’s On the Origin of Species only a couple of decades before. If species were changeable objects, then conceivably they could be moved well outside of their known geographical range and adapt to the conditions provided there. At that time the pace of evolution and adaptation were far less clear than they are now, and the world was a scarier place with more threats for humans. Introducing organisms to test the limits of survival was adventurous, and introducing organisms to make the world more comfortable seemed like common sense. Examining whether an Arabian camel could survive in London was not a trivial question back when we were still learning the geographical limits of most animals on Earth. In fact, current introductions suggest that it still may not be trivial to ask if introduced camels could form a free-roaming population in a London-esque climate. These kinds of questions motivated ordinary, although often wealthy, citizens to form groups devoted to acclimatization. The first well-documented
Transport Vectors and Pathways
Box 2.1
25
Continued
acclimatization society was formed in Paris, France in 1854 (La Société Zoologique D’Acclimatation), and according to Lever (1992), the mission of this society was “the introduction, acclimatization, and domestication of animals, whether useful or ornamental” for the greater good of the French economy and food supply. From this it is clear that the motivation for acclimatizing species had utilitarian roots that went beyond the purely speculative motivations described above. For example, Lever (1992) quotes Geoffroy Saint-Hilaire, the Société’s first chairman, as declaring “we have given the sheep to Australia; why have we not taken in exchange the kangaroo – a most edible and productive creature.” The work of the Société was so popular that within 10 years nearly 40 regional societies were formed, and several French colonies established their own acclimatization societies (e.g., Algeria, Reunion, Martinique). Not to be outdone by the French, the British established the Acclimatization Society of the United Kingdom in London in 1861. As in France, the UK society quickly grew to encompass several regional groups, and, importantly, many of its colonies founded their own societies. Although these societies did introduce plants and insects from time to time, they were mainly concerned with vertebrate animals. The UK society became largely a bird acclimatization society when it merged with the Ornithological Society of London in 1865. These original societies died out relatively quickly. The Acclimatization Society of the United Kingdom, for example, petered out of existence sometime around 1868 (Lever 1992). The societies were criticized for being “non-scientific” in their approach to acclimatization, which at the time meant they did not keep good records of how many individuals of what species were released or where these individuals were released. In addition, the efforts of the various regional organizations often were not synchronized, leading to confusion in the success or failure of any particular release effort. This left most academic biologists lukewarm to downright cold toward the societies (Lever 1992). There were also a few individuals who questioned the wisdom of releasing so many species since it was apparent that a few “successes” were causing economic and ecological problems. Nevertheless, the spin-off societies persisted long after the demise of their putative “mother” institutions, including very active groups in Australia, Algeria, New Zealand, Hawaii, and the United States (Lever 1992). The Hui Manu (Hawaiian for “Bird Society”), for example, was formed in 1930 and persisted in releasing non-native birds into the Hawaiian Islands until 1968. Although short-lived at times, acclimatization societies left a profound biological legacy. Each group had a slightly different focus that followed the interests of their membership, and these interests are reflected in the current list of nonnative species found on each country or island. For example, New Zealand has, to a great extent, become a biological replica of England since the members of its acclimatization societies found the climate of the islands to be hospitable but the “streams almost destitute of fish, with shy song birds and few game birds” and
26
Chapter 2
Box 2.1
Continued
thus “it seemed to them that it [New Zealand] only wanted the best of the plants and animals [from England] to make it a terrestrial paradise” (Thomson 1922). Similarly, North America and surrounding islands (i.e., the Caribbean) now have the European Starling because Eugene Schieffelin, a Shakespeare fanatic and native New Yorker, wanted to introduce all the birds mentioned by the Bard into Central Park (Fig. 2.2). The utilitarian motivation for introducing species persists to this day among biocontrol advocates, albeit quite tamed compared to acclimatization societies (see Box 4.2). Of course, the human desire to be surrounded by aesthetically pleasing species is still manifest in the variety of horticultural and landscaping plants available at your nearest garden market. The few acclimatization societies that are still in existence, however, shifted their focus to the rearing and re-introduction of severely threatened native species (e.g., New Zealand).
1970
1960 1950
1940
New York 1890 1920 1930
Starling
Figure 2.2 Westward range expansion of the European starling (Sturnus vulgaris) across North America following intentional release into Central Park, New York, in 1890. [Redrawn from Krebs, C. J. 1988. The message of ecology. Harper & Row Publishers, New York with permission from Pearson Education Inc.]
Transport Vectors and Pathways
27
The number of species and the number of individuals transported via these vectors is immense. For example, Radonski et al. (1984) estimate that up to 1500 species of freshwater fish for the aquarium trade were regularly imported into the United States in the early 1980s. Fuller et al. (1999) suspects this figure is a severe underestimate of the numbers moving into the United States today. Siguan (2004) reports that a single commercial aquarium company offered 102 seaweed species for sale, including some of the most damaging invasive species known (e.g., Caulerpa taxifolia and Sargassum muticum). Mack (2004) reports on the results of a similar survey indicating that 25,000 taxa of angiosperm plants are available for purchase in New Zealand, including many that are known to be harmful. Kraus (2004) indicates that the pet trade is consistently the largest vector for the importation and release of non-native reptiles, amphibians, and birds worldwide. For woody plants in the United States, there are as many as 200 horticultural species established outside of areas where they were cultivated (Reichard & White 2001). This problem is not confined to the US, as estimates suggest between 57% and 65% of the introduced woody and herbaceous species in Australia were introduced for horticultural purposes (Reichard & White 2001).
Biocontrol A group of intentional introductions often overlooked are the species introduced for biocontrol. The idea is to introduce a “natural enemy” of an unwanted pest species with the expectation that the enemy will serve to reduce the population of the pest to a level that is not harmful (Hoddle 2004). The “pest” has historically included native species considered harmful to agriculture or human health but now is typically restricted to non-native species that cause some economic or environmental harm (see Chapter 12). Biocontrol organisms can include insect predators, parasites, or diseases that attack the pest reducing its survivorship or fecundity. Introductions of insect and pathogenic biocontrol agents are an active scientific discipline today. In the past, however, biocontrol organisms were often vertebrates. For example, the western mosquito fish (Gambusia affinis) was used for mosquito control in much of the western United States and around the world (Lever 1996). Insectivorous birds were introduced to many oceanic islands in order to protect horticultural crops from insect pests. The small Indian mongoose (Herpestes javanicus) was transported to, and released on, several islands in the hopes that it would reduce populations of rats in sugarcane fields (Simberloff et al. 2000).
Conservation and scientific pursuits In certain circumstances organisms are transported to new areas for the express purpose of conserving or protecting the future of the species. This type of transport includes the many plants brought to botanical gardens or arboreta and animals brought to zoos or wildlife preserves as well as endangered species moved to novel natural environments for their immediate protection. For example a population of the Devil’s Hole pupfish (Cyprinodon diabolis) was introduced to the Amargosa Pupfish Station in
28
Chapter 2
1983 (Sigler & Sigler 1987) because the only natural population of the fish was being threatened by groundwater pumping (Soltz & Naiman 1978). While transport for such conservation purposes is unlikely to result in invasions, we cannot be certain. Scientific studies have transported species that become invasive. For example, Cowie and Robinson (2004) relate the story of 10 species of Cerion snails native to the Bahamas, Puerto Rico, and Cuba that were brought to the Florida Keys in order to investigate the controlling mechanism for shell morphology. At least one of these species established a self-sustaining population in Florida. A particularly devastating introduction that falls under this category is the release of the marine alga, Caulerpa taxifolia, outside of the Oceanographic Museum of Monaco (Meinesz 1999). This hardy green alga is native to tropical oceans and was cultivated in the museum’s aquaria and used as a decorative addition to presentations of tropical fish. It was first found growing in the Mediterranean Sea outside the museum’s window in 1984. Since then it has steadily spread across the seafloor and now carpets thousands of hectares much to the detriment of native marine species (Meinesz 1999). It has shown up in harbors outside the Mediterranean more recently, perhaps transported there on fishing gear or anchors of boats docked in previously invaded harbors. Caulerpa is also still used in the aquarium industry and is regularly transported outside its native range in the dumping and cleansing of private aquaria.
Unintentional transportation Some species are transported unintentionally as the byproduct of the movement of other goods. Species can be transported as hitchhikers on or within other non-native species, or they can be transported within cargo holds and in packing material. By their very nature, unintentional introductions are difficult to evaluate in terms of their overall importance, but recent evidence suggests they are quite prevalent (Carlton 2000). Perhaps the best-known example of unintentional introduction involves movement of species within ship ballast. Ocean-going vessels must balance their cargo load in order to stabilize the ship and have it move efficiently through the water. Early cargo ships used solid ballast of rubble, gravel, stones, and soil (Mack 2004). This ballast was collected indiscriminately at the port of departure and off-loaded, equally indiscriminately, at the port of call. As you can imagine, anything small enough to escape the notice of the ballast-loading workmen was scooped into and out of the ship, usually ending up stacked in a heap near the boatyard. The species carried within solid ballast included insects, plants, earthworms, and many other small organisms. Mack (2004), citing Rhoades and Klein (1993), indicated that 81 non-native plants recorded in the US state of Pennsylvania have only been found on discarded ballast heaps around the port city of Philadelphia. Most of these plants are native to the Mediterranean, Africa, and South America, and were first recorded in the flora in the mid- to late- 1800s. Vazquez and Simberloff (2001) indicate that most insects introduced into the United States around this same time (mid-1800s) were also transported via ballast soil.
Transport Vectors and Pathways
29
Widespread recognition of the importance of ballast water transport came from the work of James T. Carlton and colleagues (see Box 4.1 for more on ballast). The significance of this vector in providing sources of non-native marine invaders has grown as free trade has increased and the size and speed of ocean-going vessels have grown (Carlton 2000). Carlton (1999) estimated that 10,000 or more species may be transported within ship ballast water on any given day. If this estimate is anywhere close to correct, ballast water rivals the intentional importation of plants and pets as the most important invasion vector operating today. Another important source of non-native species is the transport of “contaminants” within commercial products. These contaminants are individuals of non-native species that are difficult for people to detect and that hitchhike their way to a non-native locale. Common hitchhiker species include ticks, mites, fleas, or other parasites carried on domestic animals or livestock (Vazquez & Simberloff 2001), small plants growing within the pots of horticultural species, and seeds mixed in with desired crop seeds (Mack 2004). In a study of horticultural plants and perennial stock from a sample of seven plant nurseries in Sheffield UK, Hodkinson and Thompson (1997) found that 51% of all pots examined contained between one and two unintended plants per pot. Massive amounts of hay were imported into Australia for livestock during the drought of 1980. Thomas et al. (1984) found that each bale of imported hay contained an average of nearly 70,000 seeds, and a survey of 38 bales yielded 105 non-native hitchhiker species. It has even been shown that horse dung can be a significant source of viable seeds along equine trail corridors (Campbell & Gibson 2001). In other cases, species are transported within ship or airplane holds (e.g., brown tree snake, Box 1.2), within packing material, attached to the side of ships (see Box 3.1), or within columns of water diverted for human use (Vazquez & Simberloff 2001; Fofonoff et al. 2004). Movements of live oyster clusters for commercial oyster fisheries often bring many additional marine organisms, including algae, found in and among the clusters (Carlton 2001; Wasson et al. 2001). Hendrix and Bohlen (2002) reported 45 species of non-native earthworms moved with soil products into the postglaciation oligochete-free soils of the northern United States and Canada. Similarly the Shimofuri goby (Tridentiger bifasciatus), a small Japanese estuarine fish, was released into the San Francisco Bay Estuary in the mid-1980s and was subsequently transported over 500 km south across a mountain range to the Los Angeles basin by water export through the California State Water Project (Matern 2001). Some of the most devastating impacts to native ecosystems to date have been through the invasion of non-native plant fungi (Palm 2004). Fungi are transported primarily in association with horticultural plants, wood products, and non-propagative plant material (i.e., fruits, vegetables, cut flowers, dried herbs) (Palm 2004). These hitchhikers are very hard to detect and control (see Box 2.2 for a detailed example).
Which species are transported via what vector group? Although the sheer variety of transportation vectors leaves one feeling that the process is random, there are clear associations between species groups and vector categories.
30
Chapter 2
Box 2.2
Knocking on heaven’s door
Forest pathogens (e.g., fungi, bacteria, and viruses) play a large role in determining plant species composition and forest structure (Holdenrieder et al. 2004). Thus, when a pathogen finds its way into a forest to which it is not native, it can cause substantial changes. Well-known examples include chestnut blight in North America (see Box 9.2), jarrah dieback in Australia (Weste & Marks 1987), and alder dieback in Europe (see Chapter 12, Brasier 2001). In many of these cases, we can only see the end result of the infestation (e.g., shift in tree dominance), or the invasion is so widespread we can only control continued population expansion around the margins of the non-native pathogen’s distribution. These infestations can cause permanent changes to communities and ecosystems with ripple effects on plants, insects, and vertebrates. Control or eradication is much more effective if we can catch the invasion early when the population is small and relatively restricted. Even if we fail to immediately control the invader, early detection allows us to learn more about subsequent transport and spread of non-native pathogens. Such information is useful for monitoring and control measures as well as for developing efficient and effective quarantine laws and early-warning systems. A recent example of an emerging forest disease is Sudden Oak Death, which infects at least 23 plant species from 12 families within California and Oregon coastal forests (Meentemeyer et al. 2004). The pathogen in this case is Phytophthora ramorum, which was only recently discovered and first isolated from rhododendron (Rhododendron spp.) and viburnum (Viburnum spp.) nursery stock in Europe (Rizzo & Garbelotto 2003). The pathogen causes two types of disease, one of which tends to be lethal. The lethal form infects main branches and stems and kills several trees including tanoak (Lithocarpus densiflora), coast live oak (Quercus agrifolia), and California black oak (Quercus kellogii). The loss of these trees has broad implications for the future of forest communities along the western coast of North America (Rizzo & Garbelotto 2003). The disease got its name because groups of infected trees appear to die quickly, sometimes within a span of weeks (Rizzo & Garbelotto 2003). Not all trees die this quickly, though. In fact, it is the species that do not show the lethal form that are more important in understanding the spread of P. ramorum. The trees that develop the lethal disease usually die long before the production of sporangia and chlamydospores, the dispersal propagules of P. ramorum. On the other hand species with some resistance develop non-lethal infections on leaves and twigs and persist long enough for P. ramorum to produce dispersal propagules. In particular, California bay laurel (Umbellularia californica) may be a principal “reservoir” for the disease because it never shows lethal effects but maintains P. ramorum in forest systems. There is some evidence that oaks growing near bay laurel are more likely to be infected and to die of the pathogen. Spores of P. ramorum naturally disperse via wind and potentially rain splash (Meentemeyer et al. 2004). Thus, well-timed rainstorms may be critical to the species’ geographical range expansion. This also fits with the spatial distribution of infected trees because there is a clear clumping
Transport Vectors and Pathways
Box 2.2
31
Continued
pattern of infection within forests. There is also some evidence that spores spread over longer distances via the disposal of green waste or via soil encrusted on the hooves of deer or other large mammals, including hikers. Together these mechanisms of dispersal suggest that the spread of P. ramorum may be usefully modeled using principles of landscape ecology and metapopulation theory. The global transport of this pathogen, though, requires human action. Ornamental plants such as rhodendron, viburnum, and camellia regularly show leaf (non-lethal) infections. The warm, moist conditions within plant nurseries are favorable for the production of spores, and the physical proximity of plants may allow rapid infection of much of a nursery’s stock. If infected individuals are sold outside of the infested area, they can introduce the disease well outside of the current distributional boundaries. This route is probably how P. ramorum originally came to California although it is still unclear where the species is native. Through genetic analyses, researchers have confirmed that the European population of P. ramorum was not the source for the population in California (Rizzo & Garbelotto 2003). This has led to some speculation that P. ramorum is native to someplace else besides Europe and North America and was introduced onto each continent on separate occasions, or that it has a recent hybrid origin between two other Phytophthora species (see Chapter 11 for additional examples of hybridization). In either case the potential for long-range dispersal via the ornamental plant industry is key to controlling the additional spread of P. ramorum. The California Department of Food and Agriculture has recognized the potential for California nurseries to act as a vector to other regions of North America and has instituted quarantines of nurseries in 12 counties that represent the center of the current infestation (Stokstad 2004). Nevertheless, in 2004 officials found six varieties of camellias infected at a nursery with national distribution outside of Los Angeles. Several nurseries surrounding this one showed evidence of infection soon thereafter (Stokstad 2004), and there were confirmed shipments of infected plants to the eastern United States. These shipments are worrisome because laboratory tests have demonstrated that two native, dominant oaks of eastern deciduous forests are susceptible to P. ramorum (red oaks, Quercus rubra, and pin oaks, Quercus palustris) (Rizzo & Garbelotto 2003). The US Department of Agriculture immediately tried to trace where the infected plants were shipped and have had some success in ensuring that infected plants were not subsequently sold to consumers for planting in their gardens. At the time of writing, however, it is unknown if P. ramorum has made its way into eastern forests. Several US states have taken the precaution of banning the importation of ornamental plants from California growers (Stokstad 2004). The case of Sudden Oak Death illustrates the complexity of dispersal of nonnative organisms, especially when their geographical range expansion is helped along by continued human actions. Understanding how natural and humanderived dispersal mechanisms operate is critical to controlling harmful non-native species before they place native ecosystems at death’s door.
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Chapter 2
If we intentionally leave out the cryptic and hard to evaluate taxa (e.g., viruses, bacteria, protozoans, and fungi) we can find data for several commonly transported taxonomic groups (Table 2.1) and examine typical vectors of transport for each group. Specificity of the data necessitates limiting the geographic ranges for the taxa under scrutiny; however for comparison we include data at a number of different geographic and geopolitical scales. We can see some striking patterns in the vectors of particular species. For example, conservation and biocontrol are not common modes of transport for any of the taxa in the table. This makes sense given that movement of organisms for conservation purposes is a relatively recent phenomenon caused by growing concern for biodiversity and conservation biology. It is unlikely that conservation movements will ever be a major source of non-native species. Biocontrol movements on the other hand have been attempted a number of times, but there have also been some spectacular failures (see Chapter 9) that may have limited this mode of introduction. This pattern would almost certainly change if we had included data for terrestrial invertebrates, as they are commonly moved to new areas for biocontrol purposes today, and there are many examples of subsequent establishment and use of native hosts. A large vector category in Table 2.1 is environmental enhancement. People intentionally bring species with them when they move in order to make their home space more “livable.” Many deliberate introductions of this sort are related to the interest humans have in maintaining species that are helpful to them (McNeely 2001). In fact some scientists have gone so far as to suggest that the frenetic pace of globalization has increased people’s disconnection from the land and its ecological processes so that they are currently more likely to desire common or familiar species near them than during other times in history (Staples 2001). While this psychological reasoning is hard to prove, it is clear that the pace of introductions is steadily increasing through time (see Chapter 3) and that many of these species are transported expressly to satisfy human desires for ornamental species. Exceptions to this pattern occur in the freshwater fishes of California and the marine fishes of the world. Here it seems that the number of species moved for environmental enhancement is relatively low when compared with reasons having to do with food and game (California) and hitchhiking (marine fishes). This is also true to a lesser extent for the freshwater fishes of the United States. It seems that intentional introduction of fish species often has very utilitarian purposes. A remarkable result coming from Table 2.1 is the lack of successful introductions of marine fishes. Why would marine fish be excluded from transport as non-natives? Of the nearly 25,000 fish species in the world, a little over 14,000 of them are marine species, yet from one of the best databases on fishes in the world (www.fishbase.org) we could find only 56 marine fishes that had been successfully introduced into new areas. This lack of marine fish introductions may be due to a number of different factors. The first and simplest is that marine fishes may be transported as often as other organisms, but because of the less complete taxonomic inventories for marine systems, it is difficult to tell if a newly discovered marine fish is non-native or a previously un-described native (Carlton & Geller 1993). Second, fish are typically present at lower densities than planktonic organisms, for example, and therefore are numerically less
Table 2.1 Percentage of non-native organisms transported by particular vectors out of the total number moved per group. Actual numbers of organisms moved provided within parentheses. Global marine fishes†
United States aquatic invertebrates∗
United States freshwater fish∗
United States amphibians∗
United States reptiles∗
Global birds‡
Australia birds‡
California freshwater§ fish†
California exotic plants
Food & Game
17 (9)
11 (8)
42 (205)
14 (8)
13 (8)
25 (103)
22 (20)
69 (49)
13 (?)
Environmental enhancement
0 (0)
20 (14)
20 (99)
39 (22)
69 (44)
13 (53)
17 (15)
7 (5)
40 (?)
Bio-control
0 (0)
0 (0)
3 (14)
11 (6)
0 (0)
4 (18)
6 (5)
3 (2)
Conservation & Science
0 (0)
0 (0)
4 (21)
0 (0)
0 (0)
3 (12)
2 (2)
3 (2)
Unintentional
73 (41)
41 (29)
25 (132)
11 (6)
0 (0)
28 (117)
22 (20)
18 (13)
5 (?)
Unknown
11 (6)
27 (19)
3 (15)
25 (14)
19 (12)
27 (113)
31 (28)
0 (0)
42 (?)
Totals
56
70
486
56
64
416
90
71
1152
Notes: ∗ based on data in Benson (1999) † based on data from FishBase (www.fishbase.org/home.htm) ‡ based on Cassey (2001) reporting common modes of transport § based on data in Moyle (2002) based on data from J. DiTomaso pers. com.
Transport Vectors and Pathways
Mode of transport
33
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Chapter 2
likely to be inadvertently moved within ballast water (Carlton & Geller 1993; Smith et al. 1999). Third, many marine species have very large native ranges. Thus, in order to move a marine fish out of its native region and into another novel non-native location the transport vector must move individuals over quite a long distance. Very few vectors are likely to do so. In fact, 55% of all the recorded marine fish invaders achieved their non-native distributions through the opening of the Suez Canal and the transfer of fishes from the Red Sea into the Mediterranean. The growth of salmon aquaculture may serve to bring the number of marine fish invaders up in the future (Naylor et al. 2001).
Dynamics of transport pathways Transport pathways are defined as routes between the source of non-native species and the recipient location for these species. Ecologists that study transport pathways tend to focus on the strength of these pathways, defined as the number of species moved along them and the viability of individuals of these species once they are released (Mack 2004). The stronger the pathway the more likely it is to lead to the establishment of non-native populations for a variety of reasons detailed in Chapter 4. Here, we want to illustrate the spatial and temporal dynamics of these pathways as they set the stage for patterns we will revisit in subsequent chapters. The spatial scale of movement of non-native species follows the developing technologies of commercial intercontinental traffic. Mack (2004) reminds us that early sailing ships were confined to traveling oceanic routes with favorable wind patterns. With the advent of sophisticated navigation systems and steam-powered ships, the number of pathways increased dramatically. These changes in technology connected nearly all cities with ports deep enough to harbor a large boat. In addition, the speed with which these vessels could traverse large expanses of ocean meant that the transported nonnative individuals arrived at their destination in much better health than could have been possible in wind-powered ships. The rise in airplane transport of people and cargo ratcheted up the number of connecting pathways and the viability of the individuals transported yet further. Airline hubs are not restricted to coastal cities as are shipping hubs, thus allowing non-native species to penetrate directly into the interior of continents. Airplane travel also served to revolutionize island-state economies, allowing “flown-in fresh” products, and their associated hitchhikers, to be delivered daily. In New Zealand the introduction of the Mediterranean fruit fly (Ceratitis capitata) is postulated to have originated from larvae in a fruit brought in by a passenger on an overseas flight and thrown out the window of a car between the Auckland Airport and the city. The cost of measures to control the outbreak has been over $5,000,000 and the cost from lost exports over $10,000,000. The spatial connection between port cities also closely follows geopolitical events. For example, the Polynesians were one of the earliest societies to explore and colonize new parts of the world. They brought with them an array of species to the islands they colonized, and several of those species persist on these islands today (Crosby 1986). However, the scale of the Polynesian expansion was dwarfed by the period of European
Transport Vectors and Pathways
35
Communications : Principal steamer routes and coaling stations, 1889
6
20˚–
0˚–
Figure 2.3 British steamer routes and large coaling stations as of 1899. Notice the number of connections between locations around the world that were established well before modern global trade and airplane transport developed. [Reprinted from Mack, R.N. 2004. Global plant dispersal, naturalization, and invasion: pathways, modes and circumstances. Pages 3–30 in Ruiz, G.M. and Carlton, J.T., eds. Invasive species: vectors and management strategies. Island Press, Washington, DC.]
expansion (Crosby 1986). The connections between Britain and other parts of the world in the mid-1880s are particularly impressive (Fig. 2.3). A close examination of Fig. 2.3 explains many of the patterns in numbers of established non-native species across locations detailed in Chapter 3. A similar insight comes from understanding the rapidly changing relations between nations during, and just after, wars (Cowie & Robinson 2004; Mack 2004, chapter 3). The power that understanding transportation pathways provides is evident in their use in predicting where invasive species are likely to come from next (Chapter 12). For example, an analysis of transport patterns into the North American Great Lakes has demonstrated that non-native amphipods, zooplankton, mussels and fishes have been transported primarily along the major shipping routes between the Great Lakes and northern and western Europe (MacIsaac et al. 2002). In addition, heavily invaded regions or ports may subsequently serve as “hubs” for the transport of non-native species to nearby “spoke” locations (Carlton 1987, Fig. 2.4). For example, in a study on intra-coastal transport of non-native species in ballast water, Lavoie et al. (1999) found that millions of surviving organisms (dominated by dinoflagellates, diatoms, and copepods) were released into the receiving harbors located near “hub” cities. A recent survey of an embayment that is not connected to a large international harbor found 56 exotic species with only modest effort (Wasson et al. 2001), suggesting that the effect of unintentional introductions is not limited to the immediate areas of release. In a similar manner, proximity to urban areas often correlates with numbers of non-native plant and
36
Chapter 2 Site of secondary invasion
Site of secondary invasion
Site of secondary invasion
Hub port
Export route
Site of secondary invasion
Figure 2.4 Hub and spoke diagram illustrating that some ports can accumulate many nonnatives and serve as a significant source of propagules to smaller ports in the surrounding area. These ports would not normally receive these species since they cannot accommodate large ships.
animal species established there due to the likelihood of unintentional introductions from urban centers (Reichard & White 2001). One of the more remarkable aspects of transportation pathways is that they tend to function as sources of non-native species in only one direction. We detail these inequalities in non-native species transfer in Chapter 3.
Summary Our review of the transport vectors and pathways that result in the release of nonnative species may have left your head swirling. There is a huge diversity of vectors, and the magnitude of their influence on the distribution of species on Earth is not yet known. It is very difficult to provide an in-depth review of the variety of ways in which species become non-natives without resorting to a “hit and run” survey. We encourage those students interested in the interplay of society, technology, and invasive species to pursue the suggested readings below or, better yet, their own research agenda. Remarkably, even though this is the foundational step in the invasion process, relatively few of these vectors and pathways have been studied extensively. In particular, we simply
Transport Vectors and Pathways
37
do not know how many species are being transported around the world on any given day, month, or year. Nor do we know the likely origin and destination of these species or if the pathway has changed dramatically in recent years. This is unfortunate since, as you will see, there are many facets of the invasion process that are influenced by factors expressed during the transport stage. Indeed, toward the end of this textbook, you will find that targeting pathways that regularly deliver non-native species is one of the principle ways in which society attempts to stem the flow of invasive species.
Further complexities Ruiz, G.M. and J.T. Carlton. 2004. Invasive species: vectors and management strategies. Island Press, Washington, DC. Crosby, A. 1986. Ecological Imperialism: the biological expansion of Europe, 900–1900. Cambridge University Press, Cambridge.
Companion papers Carlton, J.T. 1987. Patterns of transoceanic marine biological invasions in the Pacific Ocean. Bulletin of Marine Science 41: 452–465. Mack, R. and W.M. Lonsdale. 2001. Humans as global plant dispersers: getting more than we bargained for. BioScience 51: 95–101. Reichard, S.H. and P. White. 2001. Horticulture as a pathway of invasive plant introductions in the United States. BioScience 51: 103–113. Vermeij, G.J. 1991. When biotas meet: understanding biotic interchange. Science 253: 1099–1104.
3 Trends in Numbers of Invaders
Overview In Chapter 2 we outlined the common modes of transporting nonnative species and discovered that these modes can be both taxon and site specific. In this chapter, we will explore the effect these trends have on the rate at which non-native species are detected, and the spatial discontinuities in numbers of non-native species that can often result. Because we often do not know how many, or which, species are being transported and released, this chapter relies heavily on patterns in the relative number of non-native species that have successfully established (for at least long enough to be recorded) across decades or locations. As the movement of goods and people has increased over the decades, the number of non-native species recorded in many locations has also increased. This is true whether or not we consider by-product or intentional introductions. Temporal trends in the origins of nonnative species are just beginning to be documented. However, the results of such studies reveal complex interactions between changes in vector importance and kinds of non-native species introduced. Using standard ecological tools and principles, such as species-accumulation functions, we can make predictions about how many non-native species any one location will collect through time. These tools may provide a first approximation as to the speed at which locations will accumulate non-natives; however, as we will see, there is much room for further investigation. There are marked differences in the number and origin of non-native species across geographic locations. Spatial trends in the number of species introduced explain much of this spatial variation.
Invasion rates through time Despite the knowledge that invaders often cause substantial ecological and economic problems (Chapters 9 & 10), we are woefully short of good analyses examining temporal
Trends in Numbers of Invaders 150
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Figure 3.1 Cumulative number of non-native species established within the San Francisco Bay Estuary. The raw number of invasions (A), and the modified number of invasions (B), increase through time. Reprinted with permission from Cohen, A. N., and J. T. Carlton. 1998. Accelerating invasion rate in a highly invaded estuary. Science 279: 555–558. Copyright 1998 AAAS.
patterns in overall numbers of non-native species introduced (Cohen & Carlton 1998). This is because of a marked lack of data on the timing, or even sequence, of species introductions (Pysek et al. 2003). Nevertheless, given this limitation, there are enough analyses to discern two general patterns regarding invasion rates. First, the number of non-native species recorded in any one location has tended to increase through time. Second, this increase is caused by the sometimes idiosyncratic temporal trends in how and why species are transported. The most widely cited study of temporal trends in number of non-native species is that of Cohen and Carlton (1998). These authors assembled information on aquatic organisms (plants, protists, invertebrates, and vertebrates) found within the San Francisco Bay estuary using published accounts coupled with their own sampling efforts. Their results indicated that nearly 250 non-native species have been recorded in the estuary (Fig. 3.1A). Cohen and Carlton then whittled down this list to include only those species for which they could assign a date of first observation, collection, or planting. This left approximately 150 species that they could arrange in temporal order to create a cumulative invasion curve that spans the years from 1850 to 1995 (Fig. 3.1B). The cumulative curve for San Francisco Bay estuary is a non-linearly increasing function clearly indicating that non-native species have been arriving and establishing at an accelerating rate. About half of all non-native species in the estuary arrived during the last 35 years of Cohen and Carlton’s 145-year record. The rate of invasion increased from an average of one new species every 55 weeks between 1851 and 1960 to an average of one new species every 14 weeks between 1961 and 1995. Because many of the species included in the work of Cohen and Carlton (1998) were by-product introductions, it is possible that this result is exclusive to this class of transportation vectors. Thus, we may fail to see an accelerating rate of invasion for species that are intentionally introduced. This would assume that society has recently and effectively halted the number of purposeful introductions in response to the growing
Chapter 3
40
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Releases
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0 <1850 1850 1860 1870 1880 1890 1900 1910 1920 1930 1940 1950 1960 1970 1980 1990 Decade
Figure 3.2 The number of biocontrol agents released by decade (bars) in Florida, USA, with the upper line being the cumulative number of species released, and the lower line being the cumulative number of releases that actually established. Notice that this latter line saturates, whereas the cumulative release line does not. Also notice that only cumulative counts are “accelerating” as in Figure 3.1. From Frank and McCoy (1993).
recognition of the ill-effects these species sometimes perpetuate. The available evidence for this assumption is equivocal at best. There is some limited empirical data on the rate of purposeful introductions in the literature. Frank and McCoy (1993) collected records of insect biological control agents released in Florida between the 1850s and 1990s and found an accelerating cumulative invasion rate (Fig. 3.2). However, when Frank and McCoy (1993) examined cumulative pest species arrivals (the species the biocontrol agents were brought in to control), the rate was nearly linear (Fig. 3.3), suggesting a fairly constant pace of arrival through time. Moulton et al. (2001) performed a similar analysis for the non-native birds of the Hawaiian Islands. The non-native birds introduced to Hawaii were often, and sometimes exclusively, imported and released purposefully. Moulton and colleagues assigned a date of introduction for all the birds introduced to the six main Hawaiian Islands and found an accelerating cumulative rate of invasion (Fig. 3.4). Notable in these two datasets is a drop in the number of recorded non-natives occurring after 1990 (Fig. 3.3 & 3.4). Thus, it is possible we may be finally approaching an expected drop-off in the number of intentionally introduced species at these two locations. However, this drop is currently only a “blip” in the overall trend toward more
Trends in Numbers of Invaders
41
80
Species
60
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0 <1850 1850 1860 187018801890 19001910 1920 19301940 19501960 19701980 1990 Decade
Figure 3.3 Number of non-native insects arriving in Florida per decade (boxes) and cumulative numbers over time (line) (Frank and McCoy 1998). Notice that the cumulative rate (line) at which species are arriving (irrespective of their eventual fate) is also increasing, although at a steady rate instead of an accelerating rate.
and more non-natives being recorded. It remains to be seen whether this persists, and whether it depends on the next pattern described. The manner in which Moulton et al. (2001) present their data suggests recognition of our second general pattern: idiosyncratic temporal trends in how and why species are released combine to produce observed invasion rates. The numbers of birds introduced from the three bird families Moulton and his colleagues consider change through time (Fig. 3.4). Galliforms (game birds) were steadily introduced onto Hawaii from 1901 until the early 1960s, after which introductions declined to zero. The peak in introductions for Columbiforms (pigeons and doves) occurred in the 1920s, with only two species from this family introduced since the 1960s. Passerines (perching birds) came next, with an apparent peak in introductions occurring between 1950 and 1980. There has been only one new passerine recorded on the islands since 1980. Although not shown, the Psittacines (parrots) have experienced a recent surge in introductions onto the islands, and this may continue the increasing trend into the future. The non-native birds of Hawaii were introduced for a variety of reasons mainly including: (i) game and hunting purposes; (ii) the efforts of active acclimatization societies; or (iii) as escaped pets. These introduction modes are closely associated with particular families of birds and particular decades of history. Thus, the rise and fall of introductions among families mirrors the rise and fall of human interest in releasing
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Chapter 3
Number of species introduced
25
20
Galliformes Columbiformes Passeriformes
15
10
5
0 pre−1876 1876−1900 1901−25
1926−50
1951−75
post−1975
Year
Figure 3.4 The number of birds introduced to Hawaii by decade. Notice in this figure how the different bird families fall in and out of favor through time, each perhaps showing a humped distribution in numbers introduced. Reprinted from Moulton, M. P., K. E. Miller, and E. A. Tillman. 2001. Patterns of success among introduced birds in the Hawaiian Islands. Pages 31–47 in J. M. Scott, S. Conant, and C. Van Riper, III, editors. Evolution, ecology, conservation, and management of Hawaiian Birds: a vanishing avifauna. Studies in Avian Ecology No. 22, Cooper Ornithological Society, Camarillo, CA, with permission.
game and importing aesthetically pleasing birds, or keeping birds as pets (with pet species being the main source of non-natives in Hawaii since the 1960s). This effect of human interest on temporal invasion patterns can also be seen among the non-native plants recorded within the Czech Republic (Pysek et al. 2003). Pysek and his colleagues used the remarkably well-documented Czech flora to build a timeline for the introduction of more than 1000 plants brought in after the year 1500. Pysek and colleagues document a steady increase in the cumulative numbers of nonnative plants recorded (Fig. 3.5), but these increases vary when you examine the continent of origin (Fig. 3.6). Plants native to Europe reached the Czech Republic earlier than those from America, Asia, and Australia. American species were introduced earlier than those from Australia. These authors also documented significant interactions between the timing of introduction and the traits of individual species (Pysek et al. 2003). Deliberately introduced species arrived earlier than unintentional introductions. Among those species introduced deliberately, taxa cultivated for utilitarian purposes arrived earlier than those cultivated as ornamentals. Finally, species that flower early in the growing season and have more generalized life histories were introduced earlier than those with later flowering times or ones that had more specialized life histories. This latter pattern probably resulted from different suites of plants being commonly introduced as ornamentals when compared to those introduced as unintentionals.
Cumulative number of species (years)
Trends in Numbers of Invaders
43
600 500 400 300 200 100 0 1740
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Figure 3.5 Cumulative number of established non-native plants found within the Czech Republic. As with other estimates of the rate at which non-natives accumulate in one location, the curve is non-linear and shows an acceleration in the latter part of the twentieth century. Reprinted from Pysek, P., J. Sadlo, B. Mandak, and V. Jarosik. 2003. Czech alien flora and the historical pattern of its formation: what came first to Central Europe. Oecologia 135: 122–130, with permission from Springer-Verlag.
Given the complex way in which transportation vectors have changed through time, it should not be surprising that accelerating rates of invasion are not always apparent. A good example is the temporal trend in numbers of non-native marine species recorded in Pearl Harbor, Oahu, Hawaii. Coles et al. (1999) collected information on the nonnative benthic invertebrates and fishes found over time in Pearl Harbor using their own sampling, historical published accounts, and museum records. They found 95 non-native species in the harbor for which they could assign dates of first collection. Coles and colleagues then plotted the ratio of the number of new non-native taxa to the total number of newly identified taxa by decade (Fig. 3.7). Increased collecting efforts should increase the number of detected non-native species and thus skew raw numbers upward during these sampling events. However, if these newly detected non-natives are scaled to the detection of any new species (including natives) then this effect should be reduced (see below for a more detailed discussion of the influence of collecting effort). The results of Coles et al. (1999) suggest that the number of non-native species found in Pearl Harbor is increasing through time, although it is not accelerating. Instead, there have been a number of distinct peaks in the introduction of non-natives; of particular interest are a peak in the 1910s and a second in the 1940s (Fig. 3.7). These two peaks make logical sense when you review world history. The principle use of Pearl Harbor over the last 120 years has been for the docking and maintenance of naval ships belonging to the United States. The harbor was first dredged to allow large Navy ships to enter in 1911, after which the harbor was used heavily while the United States participated in World War I (1917–1918). The first influx of non-native species occurred during this decade and probably was a response to the changed physical nature of the harbor (i.e., greater tidal influence and depth) and the increase in numbers of species released in the harbor from temporarily docked warships. Pearl Harbor also played a critical role during the years of US involvement in World War II (1941–1945),
44
Chapter 3
Cumulative number of species
(a) 350 300
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Figure 3.6 The number of established non-native plants accumulating in the Czech Republic through time, divided by continent of origin (a). Panel (b) documents the average year of introduction (with standard error bars) divided by the species’ continent of origin. Just as the type of birds introduced to Hawaii changes through time, so does the origin of the plants introduced to the Czech Republic. Horizontal lines in panel (b) show groups that are not significantly different. Reprinted from Pysek, P., J. Sadlo, B. Mandak, and V. Jarosik. 2003. Czech alien flora and the historical pattern of its formation: what came first to Central Europe. Oecologia 135: 122–130, with permission from Springer-Verlag.
serving as the headquarters for all Pacific operations for the Allies and thus was the receiving locus for thousands of ships every year from all over the Pacific. This spike in ship traffic undoubtedly increased the numbers of non-native species released in the harbor. In between the world wars, there is a steady background rate of invasions that results in about 5 to 10% of all newly described species being non-native. Yet this rate increases dramatically during wartimes, to 30 to 40% of new species being non-native.
Trends in Numbers of Invaders
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Figure 3.7 First reports of non-native species in Pearl Harbor, Hawaii from Coles et al. 1999. Although the number of non-natives is generally increasing, there is a clear “peak” in the numbers of non-native species found in the harbor, one around World War I, one at World War II, and two more recently. Reprinted from Coles, S. L., R. C. DeFelice, L. G. Eldredge, and J. T. Carlton. 1999. Historical and recent introductions of non-indigenous marine species into Pearl Harbor, Oahu, Hawaiian Islands. Marine Biology, 135: 147–158, with kind permission from Springer Science and Business Media.
Should we expect invasion rates always to increase after a transport pathway has been established, even if the increase in numbers of non-natives is not at an accelerated pace? Levine and D’Antonio (2003) provide a compelling attempt to answer this question, and in the process develop forecasts for the number of species introduced with increasing trade levels. The answer to this question is complicated by the fact that every new container ship does not bring with it a whole new suite of species. Each ship brings “samples” from donor regions already “sampled” by previous ships. Levine and D’Antonio suggest that this sampling process is similar to the dilemma faced by ecologists when they estimate species richness from a series of sampling plots. Each new plot added to the area already sampled contains some species that were present in previous samples, plus the possibility for new species. Thus, cumulative species richness measurements scale to the number of plots sampled in a non-linear way. If transportation vector traffic behaves in the same manner, we should expect the accelerating rates of non-native species detections shown above to eventually level off. Ecologists have come up with a number of species-accumulation functions that relate sampling intensity to species richness. Using these functions, we can extrapolate the number of species in larger samples from the numbers obtained in smaller ones. Levine and D’Antonio use log–log, log–linear, and Michaelis–Menten models to
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describe the number of new species introduced via a transportation vector that traverses a constant geographic pathway (i.e., the vector does not switch from “sampling” within one donor region to “sampling” within another). As empirical data they used historical information on numbers of established non-native insects, plant pathogens, and mollusks found within the United States. This information was paired with statistics of merchandise imports into the United States calculated from 1800 onward. Levine and D’Antonio converted merchandise numbers into 1999-US $ amounts, and then calculated the cumulative number of imports (in US $ value) per decade up to 1990. In other words they used an economic measure (increasing import trade dollars) as a surrogate for a measure of time that is typically used in species-accumulation curves. They found that all three species-accumulation models fit the observed relationship between number of US imports and numbers of non-natives recorded in the United States through time reasonably well (Fig. 3.8). The Michaelis–Menton model fit the observed data the best of the three, although the log-log model does not assume an upper bound to the number of invasions which has been a criticism leveled at speciesaccumulation relationships (Williamson et al. 2001). For all species, the relationships seen in Fig. 3.8 support the hypothesis that the per-ship probability of transporting and releasing new non-native species declines as imports increase. When they use their models to forecast further increases in numbers of exotic species, their results suggest that by 2020 the United States should expect to see the introduction of at least 3, 5, and 115 new mollusks, plant pathogens, and insects respectively (based on the more conservative Michaelis-Menton model), indicating that the United States has not reached the upper limit of non-native species introductions just yet. These results have several key assumptions that should be made explicit: (i) trade volume does not increase such that the transport vector begins depositing higher numbers of individuals (i.e., propagules, see Chapter 4); (ii) the types of goods being traded does not change through time as this would alter the set of species moved as by-products; and (iii) whatever inspections and regulations that were in place during the collection of the historical data continue unaltered into the future. The work of Pysek et al. (2003), Moulton et al. (2001), Coles et al. (1999) and others suggests that these assumptions are often violated, especially assumptions (i) and (ii). When they are violated, we should expect the number of non-native species recorded to continue to increase, potentially in a dramatic fashion, as we are essentially re-starting the sampling process from the beginning of another curve (i.e., the origin in Fig. 3.8). In fact, the use of a speciesaccumulation function assumes that there exists an upper-bound to the number of non-native species one site can receive. This may not be true if the transport pathways and vectors into a receiving site are dynamic enough to badly violate assumption (ii). Taken in total, the evidence presented above suggests that temporal patterns in numbers of non-native species recorded in any one location are strongly influenced by transport vectors. We should expect high numbers of non-native species to be recorded when transport vectors are active or recently established and to be low when they are not active or have traversed a constant pathway through time. Accelerating rates in the arrival of non-natives may be limited to locations receiving foreign species through transport pathways that have experienced rapid increases in their importance due to changes in global trade. These sites of increased importance may also have experienced
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Figure 3.8 Relationship between US imports and the accumulation of non-native mollusks, plant pathogens, and insects since 1920. The lines within each panel represent the fit of different species-accumulation models (dotted, log-log species-area; dashed, log-linear species-area; solid, Michaelis–Menten). Each relationship begins to level out through time indicating that rates of non-native introduction eventually flatten out. Reprinted from Levine, J. M., and C. M. D’Antonio. 2003. Forecasting biological invasions with increasing international trade. Conservation Biology 17: 322–326, with permission from Blackwell Publishing.
multiple switches in transport vectors or vector origins over time. Such locations are likely to be found along newly formed trade routes or ports-of-call for shipping, airline, freight or rail traffic (Box 3.1). Yet this is not true for all recipient locations. Places that have seen massive but short-lived spurts in vector traffic, that play host to vectors that do not change through time, or that have not experienced upsurges in their trade importance, will most likely evince more complicated patterns. We should note that several of the above authors took great pains to remove the effects of sampling effort from their analysis. This is worth examining as it was done for
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Box 3.1 Come sail away Regional lists of non-native species strongly reflect the history of commerce and exploration. It is not hard to envision how these forces shape the distribution of species in today’s world. Most of us are experienced travelers who understand the varied connections between continents via ships and airplanes. We have a hard time, however, envisioning how world travel looked several centuries ago. Thus, it is difficult to know how the trade in goods via wooden sailing ships contributed to the movement of species outside their native ranges. These ships went on very long voyages stopping regularly in ports along the way for extended stays of up to several months, so they could have been responsible for transporting quite a few non-native species. It must have been a rough journey, though. An airline trip that now takes us 10 hours took travelers in the 1700s several months, and maybe more than a year to make. How many species survived such long journeys? What sorts of species were likely to hitch a ride on such vessels? How much of our current flora and fauna is made up of species that were moved well before a naturalist’s inventory? In an effort to provide an answer to these questions, at least for marine invertebrates, Carlton and Hodder (1995) sampled the species that settled on the keel, hull, rudder post, and waterline of a replica of a 1500s sailing vessel as it made its way down 800 km of the Pacific Coast of North America (Fig. 3.9). The ship was the Golden Hinde II, which was launched in 1973 as a reproduction of the original Golden Hinde. The Golden Hinde was a small three-masted English carrack-galleon ship that is probably best known as the vessel that took Sir Francis Drake around the world in 1577. Thus, although small and probably frightful in which to sail across the waters of the Pacific and south Atlantic, this type of boat could make incredibly long voyages and thus easily carry a variety of marine species from one port to another. The likelihood of picking up and dropping off such species was enhanced by the long residence times of these ships at various ports en route. The Golden Hinde II traveled at an average speed of 4 knots when at sea, spending from 1 to 3 days traveling between three ports and 30 days dockside within each port (Fig. 3.9). At least a few individuals representing all common marine fouling species (i.e., species that grow on, or in, the wood of a ship) survived the open sea voyages between protected harbors. In addition, most of the species attached to the ship had stable relative abundances during the trip. Thus, wooden sailing ships most likely carried with them a largely unchanging collection of marine invertebrates to their every port-of-call. The species found on the Golden Hinde II also included several benthic organisms that were picked up from the estuarine substrates (i.e., mud and sand) when the boat settled on the harbor floor. These species had virtually no chance of moving between harbors under their own powers. Instead, the work of Carlton and Hodder (1995) suggests that they were regularly moved between locations via shipping traffic even though they are not normally found in the water column or in ballast. Such findings led Carlton and Hodder to suggest that slow-moving, fouled vessels significantly altered the distribution of marine and estuarine organisms globally and along continental
Trends in Numbers of Invaders
Box 3.1
49
Continued
margins. Given the time period in which these vessels operated (1500s to 1700s), they further suggest that many marine and estuarine species that we think of today as native may in fact be non-natives deposited well before anyone cataloged the species of that area. Such species are often called cryptic invaders (Carlton 1996a). Columbia River Tillamook Bay Yaquina Bay
Coos Bay
Oregon
Pacific Ocean
Humboldt Bay California
San Francisco Bay N
Figure 3.9 Ports of call of the voyage of the Golden Hinde II in Oregon and California in the 1995 study. Carlton J. T. and J. Hodder. 1995. Biogeography and dispersal of coastal marine organisms: experimental studies on a replica of a sixteenth century sailing vessel. Marine Biology 121: 721–730.
a good reason. More species will be recorded with more sampling effort, and because these authors relied on historical documents and museum records, it is safest to assume that sampling effort has not been constant through time. If differences in sampling effort were randomly spread through time such that we had no reason to expect relatively more effort to have been expended early versus late, this may not be a problem (e.g., among purposefully introduced species such as birds or biocontrol insects). However, there is good reason to believe that as we have become more aware of invasion problems, there
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has been a greater effort expended to try and record the presence of non-native species. The more ecologists sample for non-native species the more they find. This effect, by itself, will produce an accelerating rate of non-native species detection and clearly needs to be taken into account when examining historical data (Costello & Solow 2003).
Geographic patterns in numbers of invaders One of the first things ecologists did after they began to seriously contemplate invasions was compare the numbers of non-native species across geographic locations. Not surprisingly, they found differences (Box 3.2). This bred the term “invasibility.” A geographic location that holds comparatively more non-native species than another is said to be more invasible, or to have a higher invasibility. As we will show here this term subsumes and often confounds several mechanisms that can produce differences in numbers of non-native species between locations (Lonsdale 1999). Charles Elton was one of the first ecologists to compare the numbers of non-native species between locations (Elton 1958). Elton noted the differences in numbers of
Box 3.2
Everybody wants to rule the world
With a quick glance at a list of the non-native species of, say, Australia, one can immediately discern the strong geopolitical pattern of the flow of non-native species. Most of Australia’s non-native birds, for example (Table 3.1), originated from the Paleartic, which includes all of Europe, most of Asia, and parts of North Africa. The next most common donor ecoregions are the Indo-Malaysian and Austro-Pacific. Perhaps we can understand the latter two places of origin for Table 3.1 Non-native birds in Australia come from a variety of locations. The number of birds coming from each ecogeographic donor region is provided below, with the percentage of total introductions from each ecoregion given in parentheses. Place of origin
Number of species (%)
Afro-tropical Austro-Pacific Indo-Malayan Nearctic Neotropical Palearctic
4 (4) 22 (20) 17 (15) 7 (6) 4 (4) 46 (41)
Total
90
Source: Cassey 2001.
Trends in Numbers of Invaders
Box 3.2
51
Continued
non-native Australian birds by noting their proximity to Australia itself (Australia is within the Austro-Pacific ecoregion). People tended to move regularly around these areas and thus probably moved lots of species; however, the influx of birds from the Paleartic ecoregion is due more to Australia’s former status as a British colony. The early residents of Australia were apt to import and release birds that they knew, and these residents knew the birds of Britain (see Box 2.1). This pattern is repeated the world over, with the “home country” supplying its colonies or protectorates with the vast majority of the non-native species they harbor today (Table 3.2). This pattern led Alfred Crosby (1986) to coin the term “ecological imperialism.” He ascribes more sinister motives for such importations of species into colonies besides being homesick (namely that introduction of “your” species helps to overcome native resistance to “your” presence), but it is not clear why the reverse pattern did not occur since there was much two-way commerce between the colonies and their home countries. That kind of trade was one of the principle reasons for establishing colonies in the first place. A closer look at the place of origin for non-native species evinces much more complex patterns than commonly recognized. For example, for any one Table 3.2 Percentages, with absolute number in parentheses, of non-native species introduced from various biogeographic regions to the United States, California, and Pacific Islands. Place of origin
United States∗
United States∗
United States∗
United States∗
United States∗
California† Pacific Islands‡
Aquatic Freshwater Amphibians Reptiles Aquatic Freshwater invertebrates fish plants fish
Land snails
N. Am. S. Am. Europe Asia Africa Australia Atlantic Pacific Unknown
29 (20) 4 (3) 18 (12) 18 (12) 4 (3) 0 (0) 7 (5) 6 (4) 13 (9)
Total
68
66 (330) 13 (66) 4 (20) 10 (51) 6 (31) 0.4 (2) 0 (0) 0.6 (3) 0 (0) 503
71 (25) 6 (2) 0 (0) 9 (3) 6 (2) 0 (0) 0 (0) 0 (0) 9 (3)
62 (33) 17 (9) 0 (0) 17 (9) 2 (1) 0 (0) 0 (0) 0 (0) 2 (1)
9 (14) 12 (20) 15 (24) 48 (78) 8 (12) 4 (6) 4 (7) 0 (0) 0 (0)
85 (60) 0 (0) 3 (2) 8 (6) 4 (3) 0 (0) 0 (0) 0 (0) 0 (0)
22 (18)§ 0 (0)§ 15 (12) 22(18) 12 (10) 0 (0) 0 (0) 10 (8) 20 (16)
35
53
161
71
82
Notes: ∗ data adapted from (Benson 1999). † data adapted from (Moyle 2002). ‡ data adapted from (Cowie 2001). § data for North and South America combined in original source. data for Asia and Australia combined in original source.
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Box 3.2
Continued
country or island, successful non-native taxonomic groups tend to differ in their place of origin. Table 3.2 shows that the majority of non-native aquatic invertebrates, freshwater fish, amphibians, and reptiles introduced to the United States originated in North America, while the majority of non-native aquatic plants originated in Asia. Furthermore, a change in vector or pathway can change the geographic pattern or taxonomic composition of the flow of non-natives. For example, Jeschke and Strayer (2005) show that the rate of introduction of species from Europe to North America peaked in the nineteenth century, but the rate of introduction of species from North America to Europe is still rising. A few more years down the road, and North America may have contributed as many, if not more, non-native species to Europe as Europe contributed to North America.
non-native species between islands and equivalently sized pieces of mainland. Several authors have provided quantitative support for the idea (e.g., Lonsdale 1999; Sol 2000; Blackburn & Duncan 2001b; Cassey 2003). Elton also noted that temperate ecosystems tended to have more non-natives than tropical systems. Sax (2001) provides the best recent support for this latter observation. Sax examined published latitudinal extents (to the nearest 1◦ of latitude) for the geographic ranges of non-native species. He then tabulated the number of non-native species present in 5◦ increment latitudinal bands (Fig. 3.10). Sax’s results suggest there are fairly distinct breaks in the number of established non-native species at 23.5◦ north and south latitude (Tropics of Cancer and Capricorn). Another often cited geographic pattern is the difference in numbers of non-native species found in the Old World versus the New (di Castri 1989). di Castri compared the flora of the Mediterranean Basin to the flora of California, which is also a Mediterranean environment. According to di Castri’s calculations, the Mediterranean Basin has no more than 400 “important” non-native species recorded within it out of a total of 20,000 species. In contrast, California has over 1000 non-native species out of a total of about 5200 species. Thus, the proportion of non-native species in California is about ten times higher than that found in the Mediterranean Basin. The numbers di Castri used are now dated, but the pattern still holds. Similar to this result, Lindroth (1957) estimates that 10 times more species have been moved from Europe to North America than vice versa even though there has been an active two-way commerce connection between these locations since at least 1492. As efforts to document non-native species grow, there are many more examples of one-way transfers of species between two locations that are connected via regular commerce. For example, 70% of the non-native species recorded in the North American Great Lakes are native to the Ponto-Caspian Sea. However, relatively few Great Lakes species have been recorded in the Ponto-Caspian Sea (Ricciardi et al. 2000). Similarly, Carlton (1987) recorded six incoming routes of non-native marine species to the Hawaiian Islands, four of which have transported nearly 30 species from the western Pacific to Hawaiian waters. However, Hawaii has no exporting routes. That is,
Trends in Numbers of Invaders
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Figure 3.10 Species richness of non-native species in 5◦ latitudinal bands on the American and European continents. The dashed lines indicate the Tropic of Cancer or Capricorn, dividing tropical and temperate latitudes. Notice the drastic drop in numbers of established non-natives within the tropics as compared to the temperate zone. Reprinted from Sax, D. F. 2001. Latitudinal gradients and geographic ranges of exotic species: implications for biogeography. Journal of Biogeography 28(1): 139–150, with permission from Blackwell Publishing.
no marine taxa believed to be endemic to the Hawaiian Islands have been recorded outside of Hawaii. In fact, of all 14 marine species transport pathways documented by Carlton (Fig. 3.11), only one is significantly bi-directional. The Pacific coast of North America and the Pacific coast of Asia seem to swap species regularly. The natural response of an ecologist after observing these patterns is to ask “why?” What is it about Old World species that makes them such good invaders of the New World? Why do islands tend to have many non-native species, but contribute so few? Current answers to these questions have focused on a variety of mechanisms, almost all
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Asia 12B
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Figure 3.11 Map of transportation routes across the Pacific Ocean based on origin of nonnative species found in each port. Notice that Hawaii has no exporting routes (all arrows point outward) and that only a small fraction are two-way exchanges (arrows on both ends of line). Reprinted from Carlton, J. T. 1987. Patterns of transoceanic marine biological invasions in the Pacific Ocean. Bulletin of Marine Science 41: 452–465, with permission.
operating at the establishment stage. Potential answers include differences between two regions in degree of human disturbance (Chapter 5), in the frequency of natural disturbance (Chapter 5), in numbers of native species, and in propagule pressure (Chapter 4). Sax (2001) suggests two possible reasons for low non-native richness in the tropics: (i) low numbers of introduction attempts, or (ii) a contemporary ecological force excludes them. He quickly dismisses the first explanation as being unlikely given available evidence. For example, tropical regions such as Mexico have had a long history of European occupation and thus have probably had many species introduced there (as does the United States); however Mexico has a relatively low number of non-natives as compared to more northerly locations. Sax remains equivocal regarding the second possibility, and indeed future research should address this question. A third possible reason for this pattern may have to do with historically low sampling effort put forth in the tropics suggesting the more we look around in the tropics for non-native species, the more we may find them. di Castri (1989) explains the difference between the invasibility of the Old World versus the New by noting the long history of human settlement in the Old World. This longer association with humans, and the dynamic physical nature of the Old World
Trends in Numbers of Invaders
55
in general, has served to expose these species to frequent disturbance events. Those species that could not survive under these conditions went extinct long ago, and those that could found their way across the Old World either on their own or with the help of ancient human transportation. di Castri (1989) expresses the conclusion he draws from this observation like this: Expressed in a slightly caricatural way, the fact of having already been an invader and of still being submitted to high spatial and temporal variability, as well as to varied [human] impacts, makes it easier for [Old World] species to continue to have an “invader destiny” when transported by man – intentionally or accidentally – to colonize new territories.
In essence, di Castri is arguing that modern Old World species have already successfully passed through the “crucible” of humankind. Those that remain thus have all the necessary traits to survive in the more recently disturbed habitats of the New World. The same cannot be said for modern New World species. Presumably there are still quite a few “hothouse flowers” among their ranks that will never do well if transported and released outside of their native environment. We will explore the role of disturbance in species establishment more thoroughly in Chapter 5; for now it is enough to observe these patterns and recognize that transport patterns may significantly contribute to their formation, as we will investigate next. The conclusions of di Castri, Sax and others are based on the premise, whether stated or not, that the two regions under comparison have had equal numbers of non-native species transported to them and released on them (Lonsdale 1999). As can be seen in Chapter 2 and above, this is a highly unlikely scenario. Transportation vectors deposit particular taxa into particular locations. The taxa transported and released may change through time, as can the location from which a region receives its non-native species (i.e., the donor region). Any use of historical information to show differences in invasibility between regions necessarily incorporates all of these transportation quirks and difficulties, and these studies should be interpreted with this in mind (Blackburn & Duncan 2001a). The problem has long been that these quirks were invisible to ecologists (and everyone else), and thus they tended to ignore them. Those who have not turned a blind eye routinely find that the overall number of non-native species released plays just as strong a role in determining patterns in invasibility than other potential explanations.
Transport patterns formalized Lonsdale (1999) succinctly defined the problem posed by transport patterns in a simple model, E = I × S, where E is the number of exotic (non-native) species, and is a product of the number of species introduced (I ) and the survival rate (or rate of establishment) of non-native species in their new range (S). Lonsdale further divides I and S into component parts. For I , I = Ia + Ii , where Ia is the number of species accidentally introduced (by-product releases) and Ii is the number of species released intentionally. For S, S = Sv Sh Sc Sm , where the S-variables are the rate of species survival (establishment) after accounting for competition (Sv ), herbivory and pathogens (Sh ),
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chance events (Sc ), and maladapation (Sm , e.g., terrestrial species being released in an aquatic environment). When ecologists have noted that one region is more invasible than another (i.e., E varies across regions), Lonsdale suggests they are self-evidently focusing on S. What his equation makes clear is that to understand the underlying causes for these differences we must know both I and S. Furthermore, to understand the true influence of the survival rate (S) on the number of exotics (E), we must control the number of introductions (I ). In a wide-ranging comparison of numbers of established non-native species between regions (E), Lonsdale finds support for geographical patterns in invasibility such as observing higher values of E for islands as compared to mainlands, and for the New World as compared to the Old. However, he suggests that I (number of species introduced) could account for as much as 56% of the variance in E within his comparisons. In addition, his comparisons produced little support for the notion that native species richness was associated with fewer non-natives. In fact, he found the opposite pattern (Fig. 3.12; see also Box 6.2). Blackburn and Duncan (2001b) use information on birds purposefully released at various locations around the world to produce a more direct evaluation of the importance of I on differences in E between mainlands and islands. Blackburn and Duncan assembled information for 389 species of birds. Several of these species had been introduced to more than one location, thus their information included 1466 introduction events. Based on previous research, the authors suspected that the outcomes (i.e., successful or failed establishment) of these introduction events were inter-correlated. For example, the outcomes of introducing the same species into two locations are probably interrelated as that species brings the same set of life history traits to bear on its establishment in both places. If these life history traits increase the species’ chance of establishing, then it is more likely to be classed as “successful” in both places. Also, the authors had previously shown that islands had been the recipient of more bird introductions than mainlands. After statistically controlling for these effects, Blackburn and Duncan found no relationship between the number of established non-native birds and island/mainland status. They conclude that the “traditional perception, that species-poor islands . . . are easy to invade, probably reflects the greater total number of avian introductions to these locations.” This result has similarly been supported by subsequent analyses (Sol 2000; Cassey 2003).
Summary The number of non-native species recorded at any one location appears to at least increase steadily through time, and in some instances the rate of introduction is accelerating through time. This simple result hides a complex relationship among transportation vectors, the non-native species they deposit, and sampling effort. Changes in transportation vectors that follow increases (or decreases) in trade, commercial interest in particular products, and wars serve to continually introduce new non-native species into regions that are already heavily invaded. In many ways this process is similar to standard ecological problems related to sampling intensity, and thus may lend itself
Trends in Numbers of Invaders
57
Number of exotic species
103
102 Islands
101
Mainland 100 101
102 103 Number of native species
104
Figure 3.12 The relationship between the number of non-native species and the number of native species across island reserves (◦), island non-reserves (), mainland reserves (•), and mainland non-reserves (). Notice that non-native and native richness are positively correlated no matter the location or the protection status of the area under consideration. Also notice that, on the whole, islands hold more non-natives than mainland sites. Reprinted from Lonsdale, W. M. 1999. Global patterns of plant invasions, and the concept of invasibility. Ecology 80: 1522–1536, with permission from the Ecological Society of America.
well to the use of mathematical functions that can project introduction rates into the future. If we compare the number of non-native species found across geographic locations, we often see large differences. Recognition of these patterns has formed the basis for many theories about the ease with which species should invade various habitats. The proportion of non-native species established in any given location is often referred to as that site’s invasibility. For example, islands are thought to be more invasible than mainlands since there are comparatively more non-native species found on islands than mainlands. However, such patterns may be adequately explained by differences in the number of species introduced across regions without the need to invoke mechanisms related to the establishment stage (e.g., disturbance, competition). This suggests that many long-standing theories of invasibility need to be re-visited and the role of transport explicitly considered.
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Further complexities Carlton, J. T. 1987. Patterns of transoceanic marine biological invasions in the Pacific Ocean. Bulletin of Marine Science 41: 452–465. di Castri, F. 1989. History of biological invasions with special emphasis on the Old World. Pages 1–30 in J. A. Drake, H. A. Mooney, F. di Castri, R. H. Groves, F. J. Kruger, M. Rejmanek, and M. Williamson, editors. Biological invasions: a global perspective. John Wiley & Sons, New York. Costello, C. J. and A. R. Solow. 2003. On the pattern of discovery of introduced species. Proceedings of the National Academy of Sciences of the United States of America 100: 3321–3323.
Companion papers Blackburn, T. M. and R. P. Duncan. 2001b. Establishment patterns of exotic birds are constrained by non-random patterns in introduction. Journal of Biogeography 28: 927–939. Cohen, A. N. and J. T. Carlton. 1998. Accelerating invasion rate in a highly invaded estuary. Science 279: 555–558. Lonsdale, W. M. 1999. Global patterns of plant invasions and the concept of invasibility. Ecology 80: 1522–1536. Ricciardi, A. and H. J. Maclsaac. 2000. Recent mass invasion of the North American Great Lakes by Ponto-Caspian species. Trends in Ecology and Evolution 15: 62–65.
4 Propagules
Overview This chapter explores how the circumstances surrounding transport and release influence establishment. The set of individuals released in the non-native environment are called propagules. The more individuals released, the more locations to which individuals are released, and the healthier the released individuals, the more likely the incipient population is to establish. We review the concepts behind, and detail experimental and historical studies, that show the importance of propagule pressure in establishment success. We also show that the life history of the species being released has a strong influence on the propagule size necessary for establishment. Also, the status of the native community plays a role in determining the propagule size necessary for a non-native population to establish. Finally, we explain how propagule pressure interacts with other mechanisms of establishment (e.g., disturbance, life history) often leading to confusion over the relative importance of these various processes.
What are propagules? Throughout this chapter the set of individuals released in the non-native environment are referred to as propagules. The larger the propagule size is, the more individuals are released. The more release events there are, the higher the propagule number. The healthier the propagule is, the better the physiological condition of the individuals released. Together these three elements describe propagule pressure. As we will show below, high propagule pressure leads to increased probability of establishment. This chapter provides a critical link between conditions in the species’ native range (or from wherever individuals were entrained within a transport vector, see Chapter 2), and conditions in the non-native range. To clarify this relationship we will refer to the place of origin for these individuals as the donor region and the place of release as the recipient region (e.g., Ruiz et al. 2000). Putting it all together then, propagules originate from a donor region and are released within a recipient region, and it is within this
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recipient region that the individuals in this propagule either establish a population or fail to do so. Note that there is a slight semantic discrepancy among plant and animal ecologists in describing propagules. Often when plant ecologists speak of increased propagule size they mean the physical size of the dispersing propagule itself (e.g., seed size). In plants it is often true that increased physical size of a propagule indicates increased health. For the sake of simplicity, however, we will use the definition given in the above paragraph throughout this book.
Donor region and propagule condition At the most fundamental level, propagules will not survive in a recipient region where their basic physiological requirements are not met. For example, crustaceans and mollusks that require low saline water will not survive being released into the open ocean, and the seeds of tropical flowering plants are not likely to survive after burial under desert sands. These examples seem too obvious to be worth considering. However, a mismatch between the physiological requirements of the released individuals and the non-native environment is undoubtedly a principal reason why many incipient populations never establish. This scenario seems especially relevant for those species that are transported as by-products of other activities. Donor regions may also directly or indirectly affect the physiological condition of the individuals they provide as propagules. The condition of individuals may vary substantially across their geographical range. Thus, a transportation vector that pulls propagules from a donor region with marginal resources will deposit individuals in poor health. More commonly, the donor region will affect the condition of propagules simply by its proximity to the recipient region. Most individuals that become entrained within a transport vector experience living conditions en route that are far from ideal (Box 4.1). The longer the travel time, the more likely it is that the individuals being moved will die or drift toward poor health. Less healthy individuals generally will have less chance of establishing a viable population. The locality from which propagules are taken can greatly influence propagule size. As a simple example, for many species density is often higher at the core of a species’ native range than at the edge of this range (Brown 1995a). Provided individuals are becoming entrained in the transportation vector randomly, transport vectors that originate at the core of a species’ native range will likely move more propagules into the recipient region than those that originated at the edge of the species’ range. More generally, any factor that causes density to vary across a species’ native (and non-native) range can produce differences in propagule sizes between vector sources. As we will see below, propagule size can be a strong predictor of establishment success. Thus, a change in the donor region that increases the density of the transported species may also increase the chances that this species will successfully establish within the recipient region. The influence of the donor region on propagule pressure suggests that a switch in transport vector can suddenly allow a non-native species to establish in a recipient region where it never could have in the past. The prevalence of such unexpectedly
Propagules
Box 4.1
61
Staying alive
Ballast is ambient material loaded into the tanks of commercial vessels (Wonham et al. 2001). It is crucial to the transport of goods because it is used to correct imbalances in a ship’s stability (Tamburri et al. 2002) while at the same time increasing the efficiency and safety of vessels (Wiley & Claudi 1999). In addition, ballast can improve the maneuverability of lightly loaded vessels, reduce the effects of wind, and allow enough of the hull to be submerged for proper draft and propeller function (Wiley & Claudi 1999). Currently water is the main form of ballast for large ships. Prior to water, ships used dry or solid ballast such as rocks, soil, sand, lead ingots, or other cheap, easily available material that was manually loaded into the hold and dumped at the destination (Wiley & Claudi 1999). This process not only required massive amounts of manual labor but also transported and released many arthropods, mollusks, and plants worldwide (Carlton 2001). Water has loading and unloading advantages as a ballast material, but it, too, can transport hitchhiking organisms. In the typical scenario water is pumped into ballast tanks at the port of origin and discharged at a receiving port. There are more than 45,000 commercial oceangoing cargo vessels (Carlton 2001) transporting approximately 80% of the world’s commodities (Tamburri et al. 2002). The ballast water in these ships can contain a wide range of living organisms, such as drifting plankton (including bacteria, viruses, and larval dispersal stages of many marine organisms), nekton (free swimming organisms), sessile or attached organisms, and benthic organisms. Carlton (2001) estimated that over 7000 different species of marine life are transported and released every day around the world. In the United States alone over 2 million gallons of ballast water per hour are released into coastal systems (Carlton 2001). The sheer number of potential propagules in these releases is staggering. Wonham et al. (2001) suggest that for organisms to be successfully introduced via ballast water they must pass through a series of filters. They must (i) survive being uploaded into the ballast tanks of the ship, (ii) survive the voyage while in the ballast tanks, and (iii) survive being discharged into the recipient port. The sources of mortality within these stages include the physical harm that may accompany being taken up or discharged from ballast tanks, the change in physical conditions within the tanks (e.g., lack of sunlight, change in water temperature), predation in a tank with few predator avoidance possibilities, and starvation from the lack of available food resources. These sources of mortality may dramatically reduce the number of living individuals released at the recipient port (i.e., reduce propagule size), and the released survivors may be in poor health. Nevertheless, the absolute number of non-native marine organisms introduced via ballast water suggests that too many individuals thrive during the journey. Given this background, the most pertinent questions are which species typically survive ballast transport, why, and how to decrease the number of survivors. Recent regulations mandate openocean ballast water exchange to increase mortality for transported organisms (see Chapter 12), and it would be interesting to know if this exchange is useful.
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Box 4.1
Continued
In an attempt to answer these questions Wonham et al. (2001) sampled the ballast tanks of a ship that made a trans-Atlantic voyage between Hadera, Israel, and Baltimore, US. The ship completed an empty-refill (as opposed to a flow-through) ballast water exchange in the open ocean along the way. Wonham and colleagues sampled the tanks before and after the exchange occurred to judge how completely the water in the tank was replaced by ocean water and what happened to the species identity and diversity of the tank waters post-exchange. They found a minimum of 50 live taxa in the ballast tanks after taking on water in Hadera. However, over half of these taxa were absent in the samples taken of the ballast water when it was off-loaded in Baltimore. Only a very small percentage of the total number of individual organisms that were originally taken onto the ship in Hadera made it to Baltimore (Fig. 4.1). And none of the individuals released into Baltimore Harbor were capable of living in the low salinity conditions there. Some of the high mortality observed was due to ballast water exchange. However, plankton density and diversity increased after exchange and several coastal organisms survived fine in the ballast water post-exchange. In addition, some (b)
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Figure 4.1 Ballast-water plankton density and diversity decreased over time during a voyage in (a) the cargo hold and (b) deck tanks on an ocean going ship. Panel (c) relates the total taxon richness in the cargo hold and deck tanks. Reprinted from Wonham, M. J., W. C. Walton, G. M. Ruiz, A. M. Frese, and B. S. Galil, 2001. Going to the source: role of the invasion pathway in determining potential invaders. Marine Ecology-Progress Series 215: 1–12, with permission from Inter-Research Science Center.
Propagules
Box 4.1
63
Continued
species were more vulnerable to mortality en route than others. For example, copepods showed significantly higher mortality than the larvae of gastropods, bivalves, and other taxa that tend to live only part of their lives as plankton (i.e., meroplankton). This may be because meroplankton maintain food reserves throughout their larval stages, and these reserves buffer them against long periods of low prey in ballast tanks. All told, these results suggest that the conditions within ballast tanks have a strong influence on the types of species likely to be released at recipient ports and on the number of individuals of any one species released. Even if there are high numbers of individuals released, however, these organisms may not be in good health. Thus, the conditions organisms experience en route probably attenuate the positive influence of large propagule size on establishment success. The degree to which the results of Wonham et al. (2001) transfer to invasion vectors beyond ballast water is unknown but certainly worthy of study. “sudden” invasions spurred Jim Carlton to ask, “Why, if a [pathway] is in place for a given period of time (for example, 100 years), would a species newly appear at the end of that [pathway] in year 101? Why does the native western fouling sea squirt (ascidian) Ascidiella aspersa, transportable by ships’ hulls, successfully colonize New England in the 1980s and not the 1780s?” (Carlton 1996b). Part of the explanation may be that the transport vector or pathway changed, perhaps imperceptibly, in recent years. An altered vector could be depositing larger, healthier or better-adapted propagules to the recipient region than the former vector. A change in pathway could involve a geographic switch from one donor region to another. No matter, the end result is the deposition of propagules that are much more likely to establish themselves.
Propagule size and number Conceptual basis Numbers A central tenant of population biology is that the more individuals there are in a population, the less likely the population is to become extinct. At its most basic level, this reflects a numbers game. The more individuals there are, the better the population can absorb sudden and relatively short-term drops in survival or birth rates. External events that can produce these sudden drops in vital rates include cold snaps, heat waves, flooding, and other stochastic weather events (i.e., environmental stochasticity). Birth and death rates can also drop due to random production of a surplus of one gender over the other, or simply through the effect of having discrete numbers of individuals either dying or reproducing (demographic stochasticity). Despite the often complex
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mathematical modeling that accompanies discussions of this topic, the central observation is that the higher the population size, the longer it will persist given occasional random drops in birth or survival rates. The smaller the population size is, the more likely these events are to drive the population to extinction. In the context of invasion ecology, this suggests that the higher the number of individuals released (i.e., the higher the propagule size), the more likely the incipient population will be to survive inevitable stochastic shocks. If only a few individuals are released, it may take only one small weather event to kill all of them, or to drive the population so low that it will become extinct simply by chance demographic processes. Unlike the extinction of a species (i.e., the loss of the last remaining population), the extinction of a non-native population may not be forever. In natural settings, individuals from neighboring populations may re-colonize the location as suggested by metapopulation biology (McCullough 1996). In terms of invasions, the extinction of an incipient non-native population may be reversed through the influx of newly arriving propagules via the same vector that the original propagule arrived, or occasionally through an alternative vector. The spatial distribution of propagule releases has an effect on the probability of establishment as well. Environmental conditions are often spatially structured such that a propagule released in one location may experience very different environmental conditions than if released just a few hundred meters down the road. Spatially spreading out the points at which propagules are released should serve to deposit non-native individuals in at least one location where they can thrive (Lockwood et al. 2005).
Behavior Relatively high numbers of individuals also confer advantages that are related to behavior or other group dynamics. Several species realize peak reproductive success and low death rates through their interactions with each other. For example, seabirds will often nest in very large colonies where they fight for space and nesting materials. One advantage to putting up with all this congestion is a reduction in the probability that a predator will eat any one individual. Nesting within a large colony, much like flying in a flock or swimming in a school, serves to reduce individual predation probability by (i) becoming one of many potential prey instead of one of a few prey, (ii) confusing the predator through the sheer number of potential prey or their collective movements, and (iii) having many more eyes, ears, and chemoreceptors tuned to detecting the predator so that all members of the group can take evasive action before the predator can make a kill. Reduction in population growth rates that accompany low population density is called an Allee Effect after Warder Allee, a behavioral ecologist who first suggested behavioral interactions of vertebrates were important components to understanding population dynamics (Allee 1931). Allee Effects have since been generalized outside of animal behavior, and examples can be found among plants (Courchamp et al. 1999; Dennis 2002). Examples of Allee Effects include the failure to find a mate at low densities, a reduction in foraging efficiency in the absence of cooperating hunters, and the failure to attract pollinators when there are few conspecifics nearby to create an appealing place to forage. Any non-native species that require groups to attain even minimal
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reproductive and survival rates may find themselves at a severe disadvantage when they are released in very small numbers. These small incipient populations simply will not realize enough positive population growth to establish a self-sustaining population. Increasing propagule size or number may serve to overcome Allee Effects.
Empirical evidence Given the conceptual importance of propagule size and number on establishment success, one may naturally assume there is a great deal of empirical information to back up these expectations. However, this is one of the least documented aspects of invasion ecology for a number of reasons. A major impediment to examining propagule pressure for historical invasions lies in the difficulty of finding credible information on the number of individuals released or the number of release events even for intentional introductions. The few cases where we do have reliable information on propagule pressure come from the intentional release of vertebrates as game or from purposeful releases of biocontrol agents (Box 4.2). Below we provide three such examples. In 1975, B.P. Beirne published a review of previous efforts to establish insect biological control agents within Canada. His effort was revolutionary in that it was one of the first to comprehensively test various explanations as to why insects released for biological control often failed to establish permanent populations. By his estimate over 90% of the biocontrol introductions that occurred in Canada between the 1880s and 1970s failed. Beirne (1975) whittled down the total number of release events he could analyze from the 208 total species released to the 159 species in which he could positively determine successful or failed establishment. Of these, 115 species failed and 44 succeeded. When Beirne (1975) totaled the number of individuals released across all release events, the median number of individuals released was 5000. The median for the 44 species that
Box 4.2 Reunited The regulation of plant and animal populations by natural enemies such as predators, parasites, parasitoids, and pathogens is a common and natural phenomenon. Invasive species are often more widespread and abundant in their introduced range than in their native range. There is evidence that this asymmetry in numbers is a consequence of the invader “escaping” its natural enemies within its non-native range (see Box 6.1 for more on this notion). Reuniting the invader and its natural enemies may reduce the invader’s population size to a level that does not allow the invasive species to continue inflicting economic or ecological harm. This idea is the basis of biological control (Hajek 2004). The term “biocontrol” can refer to a natural phenomenon, a field of study, or an applied pest management technique (van den Bosch et al. 1982). At its most inclusive, the definition of biocontrol is as follows: “The action of parasites, predators and pathogens in maintaining another organism’s density at a lower average than
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Box 4.2
Continued
would occur in their absence” (p. 2, DeBach 1964 as quoted in van den Bosch et al. 1982). Classical biocontrol is the control of non-native pests through the introduction of natural enemies from the pest’s native range. Conservation biocontrol is the control of non-native pests through the conservation and enhancement of natural enemies native to the location that received the pest. The key difference between the two types of biocontrol is that classical biocontrol deliberately introduces a second non-native species while conservation biocontrol relies on native species for control of non-native pests. Ecologists that practice classical biocontrol have scored some major victories against invasive species, including early successes that catapulted biocontrol into the scientific mainstream in the nineteenth century (van den Bosch et al. 1982). In 1868, the cottony-cushion scale insect (Icerya purchasi) was noticed within horticultural plantings of acacia in Menlo Park, California. In 1876, the scale was inadvertently moved 400 miles south to the Los Angeles area, probably via the shipment of infested lemons destined for market. Cottony-cushion scale quickly spread and became a serious pest of the burgeoning citrus industry in the Los Angeles basin. An alliance of entomologists in the region pinpointed the scale insect as being native to the Austral-Asian region and then successfully implemented the release of two insects (a parasitic fly, Cryptochetum iceryae, and predatory beetle, Rodolia cardinalis or the lady bird beetle) that were observed attacking cottonycushion scale in Australia. Both introduced biocontrol agents established after their release in California and quickly spread from their initial release sites. The lady bird beetle was seen attacking cottony-cushion scale almost immediately. The final result was a dramatic reduction in cottony-cushion scale infestations in California, and cottony-cushion scale has largely remained at non-infestation levels there since the biocontrol release. Nevertheless, via the citrus trade and horticultural products, cottony-cushion scale has spread to other locations around the world where it often causes damage to native trees and citrus crops. Introducing lady bird beetle as a biocontrol agent for the scale has been successful in many of these invaded areas but has also led to the widespread presence of this non-native predator. There are reasons why spreading a predator may not be wise. Despite biocontrol successes, controversy has arisen over the use of classical biological control agents because of non-target effects on native species (e.g., Louda & O’Brien 2002). Using a non-native species to control an invasive species has the pitfall of potentially introducing an even worse invader in order to control one already present (Simberloff & Stiling 1996; Mack et al. 2000). Some of the most infamous examples of biocontrol gone wrong are the introduction of the mongoose (Herpestes auropunctatus) in the West Indies to control rats (Simberloff & Stiling 1996), the introduction of the cane toad (Bufo marinus) in Australia and Hawaii to control beetles in sugar cane (Boland 2004), and the introduction of the rosy wolf snail (Euglandina rosea) to control the giant African snail (Achatina fulica) on tropical islands worldwide (Hadfield et al. 1993). In each of these cases,
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the introduced agent has had very strong effects on native species and has required control measures of its own. The mongoose attacked native bird species but had little effect on rat densities. The cane toad is a nocturnal omnivore that exudes a toxin through its skin and has had effects both up and down the food chain. It is a very large toad and a very undiscriminating generalist – almost anything that fits within its gape becomes its prey. A recent study shows that it preys on eggs and hatchlings of the rainbow bee-eater, an endemic, Australian, ground-nesting bird (Boland 2004). Not only does it preferentially target nests with hatchlings present, but it takes over the nest for diurnal cover, thereby reducing reproduction as well as survival for adults (Boland 2004). Its effect on other ground-nesting species has not been studied, but it is likely that the toad has similar effects on other native bird species. In fact, one of the few arenas in which the cane toad has had little effect is in the control of cane beetles. The rosy wolf snail had very little effect on the giant African snail for which it was introduced, but it has had a huge impact on native, terrestrial snail species (Civeyrel & Simberloff 1996). We discuss more of these effects in Chapter 9. Despite these mistakes, biological control is still considered an important tool in reducing the spread of non-native species, particularly those with strong impacts (Carruthers & Onsager 1993). Indeed, the growth in interest of bioncontrol stems in large measure from the increasing inflow of non-native species and their occasional harmful impacts (Hoddle 2004).
successfully established was 31,200; a remarkable number in its absolute size. Of the species released in totals under 5000, 9% (9 of 98) established. In contrast, for species released in totals from 5000 to 31,200, 39% (13 of 33) established, and for species released in totals over 31,200, 78% (22 of 28) established. Beirne realized that his result confounded two independent measures of propagule pressure, propagule size and propagule number, so he split his data into the component parts for a subset of information on insects released into forest environments up to 1958. Sixty percent of biocontrol insect species that averaged over 800 individuals released per event successfully established but only 15% of those that averaged less than 800 individuals established. In terms of the effect of propagule number, 70% of the species that were released 20 or more times (in separate release events) became established as compared to only 10% of those species released fewer than 10 times. Beirne’s (1975) results offer a clear relationship between propagule pressure and establishment. The more individuals released overall, the greater the probability that the incipient population grew and became self-sustaining. There is a noticeable effect of both propagule size and number, so that more individuals in any one event or more release events increased the probability of establishment. Beirne (1975) points out one important caveat to using historical records to examine the role of propagule pressure on invasion success. He notes that early successes or
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failures of a biocontrol release could generate their own feedback. Species that were released early and were obviously successful may have encouraged additional and perhaps redundant releases of the same species by managers. Species that were released early and were obviously failures may have discouraged additional releases. This effect could easily inflate the total number of individuals released, the number of individuals released per event, and the number of release events for successful species while simultaneously reducing it for failed species. Another example demonstrating the positive relationship between propagule pressure and establishment is found with New Zealand birds. In the mid-1990s three papers were published on the effect of propagule pressure on the success of birds released onto New Zealand (Veltman et al. 1996; Duncan 1997; Green 1997). All three studies used the same core dataset. In 1922 G. Thomson published records of attempts to establish plants and animals in New Zealand (Thomson 1922). Included in this account is information on propagule size and number for over half of the 133 bird species released. Although there is similar information for other bird release programs around the world (Long 1981), the record for New Zealand is unusually complete due to the relative gusto with which several acclimatization societies kept records and imported birds onto the islands in the early 1900s (see Box 2.1). Each author made decisions about which species and release events to include based on their individual objectives. All included in their analyses variables other than propagule pressure that had been hypothesized to influence establishment success such as body size, clutch size, size of native range, diet breadth, or migratory ability. Each study reached the same broad conclusion; increased propagule pressure resulted in increased probability of establishment success (Fig. 4.2). Green (1997) tallied all individuals released across events and found that, among the 47 species he analyzed, only 6% of those with fewer than 10 individuals released successfully established. In contrast, 36% of species with 11 to 100 individuals released established, and 83% of species with over 100 individuals released established. Duncan (1997) and Veltman et al. (1996) broke down this measure of propagule pressure into propagule size and number. Each found that increasing propagule size and number had independent positive effects on probability of establishment. Furthermore, when the effect of propagule pressure was included in potential explanatory models of establishment success with the other variables (e.g., body size, diet breadth, range size), only the measure of propagule pressure was retained. This indicates that propagule pressure has an over-riding effect on establishment probability, and thus information related to propagule pressure holds more potential to predict establishment success than any of the other explanatory variables they considered. A third example of the effect of propagule size on establishment comes from a remarkably comprehensive experiment on an insect biocontrol agent in New Zealand (Memmott et al. 2005). Here a field experiment was used to manipulate the propagule size of Arytainilla spartiophila, a psyllid (Hemipteran) phloem feeder native to Europe that feeds exclusively on broom (Cytisus scoparius; not native to New Zealand). Fifty five individual releases of the insect were made along a 135 km linear transect. Numbers of individuals released (propagule size) at each site were replicated in the following manner: 10 releases each of two, four, ten, thirty, and ninety individuals, and five
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40 35
Percent of species
30 25 20 15 10 5 0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 Number of release attempts
Figure 4.2 Frequency distribution of propagule sizes of birds released on New Zealand for successfully established (shaded bars) and unsuccessful (open bars) species. Most unsuccessful introductions involved initial releases of only a handful of individuals (i.e., very low propagule pressure). Reprinted from Veltman, C. J., S. Nee, and M. J. Crawley, 1996. Correlates of introduction success in exotic New Zealand birds. The American Naturalist 147: 542–557, with permission from University of Chicago Press.
releases of 270 individuals (55 total initial populations). These incipient populations were all followed for an amazing six years after the initial release. Of the original 55 founding populations, 22 were still present six years after the initial introduction. Using a logistic regression model, the authors demonstrated the probability of establishment was significantly and positively related to propagule size, but only during the first year in the field. Populations surviving after the initial year were not significantly related to the initial propagule size. Although less likely to survive initially, some successful populations were established from very small propagule sizes (2 and 4 individuals; Fig. 4.3). Populations that survived their first year had a 96% chance of surviving to the sixth year and final population growth was not significantly related to initial propagule size. These results are interesting for two reasons. First, records of successful establishment from very small propagule sizes are rare (Memmott et al. 2005). Yet in this experiment, 20% of releases of two individuals and 40% of releases of four individuals successfully established populations. This suggests that small propagule sizes may have a higher probability of establishment than assumed, although this result may depend on the life history characteristics of the species (Lockwood et al. 2005). Second, given that under natural conditions repeated invasions with small propagule sizes are more likely than single large introduction events (Memmott et al. 2005) and the fact that these small releases may have a relatively high probability of success, invasion ecologists should
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Figure 4.3 Population growth rate of biocontrol insects resulting from releases of varying propagule sizes. Note that three very small founding populations established. Reprinted from Memmott, J. P. G. Craze, H. M. Harman, P. Syrett, and S. V. Fowler, 2005. The effect of propagule size on the invasion of an alien insect. Journal of Animal Ecology 74: 50–62, with permission from Blackwell Publishing.
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not overlook vectors that transport small numbers of propagules as these may also be important sources of establishment.
The hidden influence of propagule pressure Perhaps because the condition and number of individuals released is so difficult to document in any one invasion attempt, we have only recently begun seriously to consider how propagule pressure interplays with other ecological mechanisms to influence establishment success. In this section we review available evidence that propagule pressure strongly interacts with other determinants of establishment success, such as species-specific traits and community interactions (Chapter 6).
Propagule pressure and life history In an effort to construct a more complete vision of how propagule size influences establishment success, Cassey et al. (2004c) trawled through all the available information on propagule sizes of bird introductions and combined this with several other variables purported to increase establishment success (Table 4.1). They built multivariate statistical models that accounted for the most variance in establishment outcome (i.e. success or failure) using the fewest possible predictor variables. Like the results obtained for Table 4.1 Correlation between biological traits and propagule size using information on birds released worldwide. Most traits are correlated to some degree with propagule size, making the interpretation of studies that show a relationship between the traits and establishment success problematic. Across species comparisons
Body mass Geographical range Annual fecundity Diet generalism Habitat generalism Dichromatism Migratory tendency Latitudinal difference Island or mainland
Controlling for phylogenetic relatedness
Estimate
s.e.
X2
Estimate
s.e.
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0.17 0.30 1.48 0.18 0.06 0.23 −0.19 −0.42 −0.79
0.06 0.08 0.16 0.04 0.04 0.08 0.04 0.06 0.08
9.46** 14.18*** 90.15*** 25.60*** 1.85 8.22** 23.70*** 42.11*** 109.56***
−0.12 0.19 0.44 0.03 0.06 0.01 0.06 −0.14 −0.30
0.06 0.06 0.14 0.03 0.03 0.07 0.03 0.03 0.04
4.66* 10.91*** 10.39** 1.22 3.49 0.01 4.42* 20.17** 67.88***
Note: X 2 significance test results are statistically significant at the following levels: *p < 0.05, **p < 0.01, ***p < 0.001. Source: Reprinted with permission from Cassey, P., T. M. Blackburn, D. Sol, R. P. Duncan, and J. L. Lockwood. 2004c. Global patterns of introduction effort and establishment success in birds. Proceedings Of The Royal Society Of London Series B-Biological Sciences 271: S405–S408.
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New Zealand birds alone, the principal determinant of successful establishment was propagule size, regardless of the geographic location of the recipient region or the taxonomic group to which the birds belonged. Interestingly, Cassey et al. (2004c) also found that propagule size was highly correlated with almost all the other potential explanatory variables considered including, among others, a positive relationship between propagule size and (i) large body mass, (ii) high annual fecundity (iii) large geographic range and (iv) being released on islands. The results of Cassey et al. (2004c) suggest that any studies that have shown a correlation between these variables and establishment success may simply be reflecting the hidden effect of propagule size and not the proposed effect of the variable itself. Why these correlations exist is open to debate. However, it seems likely that human bias in choice of release organism and location drives the correlations documented by Cassey et al. (2004c). For example, acclimatization societies formed on island-states seemed to be particularly deliberate in their efforts and thus may have released comparatively more individuals than their mainland counterparts. In addition, several governmental agencies (e.g., New Zealand and Australia) have been positively singleminded about introducing several large-bodied game birds in unusually high numbers, especially as compared to the occasional releases of small seed-eating passerines. Similar patterns to this are seen in California fishes, where deliberately stocked fish species have different habitat associations and establishment success than other introduced species (Moyle 2002; Marchetti et al. 2004c). Such influences on release patterns may be ubiquitous across taxa; however, they are effectively hidden by the lack of information on propagule sizes.
Propagule pressure and biotic resistance Propagule pressure also interacts with biotic and abiotic resistance to create what are, at least initially, confusing differences in the number of non-native species found across community types. Ecologists have suggested several reasons why these differences exist, including differences in disturbance levels (Chapter 5) and native species richness (Chapter 6). However, recently ecologists have suggested that these differences are a product of variation in propagule pressure (e.g., Tilman 1997; Levine 2000; Blackburn & Duncan 2001a, b). In this case, communities with many invaders are simply those that have had many individuals of several species released, and vice versa. As D’Antonio et al. (2001a) point out, the debate over whether propagule pressure or within-community factors determine composition is not new within ecology. Plant ecologists have debated for years over whether communities are limited by seed supply or the availability of safe sites. Similarly, marine ecologists battle over whether marine assemblages are structured mostly by interspecific interactions (i.e., competition or predation) or dispersal between communities (i.e., supply-side assembly). However, D’Antonio et al. (2001a) go a step beyond this realization and describe a conceptual model of how propagule pressure and community-level interactions can combine to determine the number of invaders across sites.
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Probability of community invasion
High propagule supply Medium propagule supply Low propagule supply
Increasing abiotic and biotic resistance
Figure 4.4 Interaction of propagule supply (different lines within a graph) and biotic resistance to invasions (x-axis). Lower propagule pressure is required to successfully establish within less resistant communities. Reprinted from D’Antonio, C. M., J. Levine, and M. Thomsen, 2001a. Ecosystem resistance to invasion and the role of propagule supply: a California perspective. Journal of Mediterranean Ecology 2: 233–245, with permission from JME.
Instead of depicting communities as either invasion resistant or not, D’Antonio et al. (2001a) portray resistance as a continuum (Fig. 4.4). Thus, some communities are more resistant to invasion than others based on a suite of biotic and abiotic factors such as disturbance and species interactions. Furthermore, communities can vary between more or less resistant states through time depending on weather patterns, management, or other exogenous forces (Fig. 4.5). Where a community lies on this continuum dictates the number of released individuals required to establish a self-sustaining population. Where resistance is high, a comparatively large number of individuals must be released for the non-native population to establish. Where resistance is low, many fewer individuals must be released. Any community can be invaded given the correct timing and sufficient propagule pressure. Furthermore, once the non-native population is established it may weaken resistance (Chapters 6, 9 & 10) making it easier for future invaders to enter the community by lowering the threshold propagule size necessary for their establishment. The conceptual model of D’Antonio et al. (2001a) is appealing as it is consistent with several independently derived invasion models, and it is also consistent with several empirical results that at first seemed confusing and idiosyncratic. For example, Cassey et al. (2004c) showed that the same propagule size led to the successful establishment of 70% of birds in the family Sturnidae (starlings and mynahs), but only 30% in Muscicapidae (thrushes). Similarly, the same propagule size led to 56% success among birds released in Hawaii but only 35% among those released in New Zealand. Beirne (1975) reported that the median total number of biocontrol individuals released was 31,500 for species introduced into forests, and 23% of these species were successful. However,
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Medium propagule supply
Increasing abiotic and biotic resistance
Figure 4.5 Biotic resistance can be influenced by previously successful invaders (or other forces), and this will decrease the propagule supply necessary for later arrivals to establish. This is an example of synergistic (facilitative) interactions between early and later arriving species that shift downward the propagule size necessary to establish. Reprinted from D’Antonio, C. M., J. Levine, and M. Thomsen, 2001a. Ecosystem resistance to invasion and the role of propagule supply: a California perspective. Journal of Mediterranean Ecology 2: 233–245, with permission from JME.
the median total number of individuals released in ornamental shrub situations was 6600, and 43% of these species were successful. It is possible that these patterns reflect basic differences in the resistance of these communities to invasion, and once propagule pressure is controlled for, we can uncover more fundamental patterns in the role of community-level forces.
Summary This chapter explores how the individuals (propagules) transported and released at a site interact with the establishment phase of the invasion process. Propagules arrive from a donor region, which itself can play a significant role in establishment probability. All individuals do not arrive at a recipient site in equal physiological condition, which is potentially influenced by the conditions present in the donor region. Likewise, the number of propagules transported can be influenced by the specific geographic location where they were entrained within a transport vector. There is fairly good support for the idea that the total number of individuals transported (propagule size and number) has a significant and positive effect on establishment success. In certain circumstances, though, a population threshold (Allee Effect) may place constraints on establishment. The effect of propagule pressure can and has been confounded with
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other variables used to explain establishment success, suggesting that propagule pressure by itself is not the unifying answer for why some species establish and others fail. In fact, propagule pressure may often be confounded with simple human interest or persistence of release activities as is often the case for deliberate or purposeful introductions. Propagule pressure may also interact with biotic and abiotic forces of resistance, and the combination of these factors may help explain some of the contrasting patterns we see in establishment success.
Further complexities D’Antonio, C., J. Levine, and M. Thomsen. 2001a. Ecosystem resistance to invasion and the role of propagule supply: A California perspective. Journal of Mediterranean Ecology 2: 233–235. Veltman, C. J., S. Nee, and M. J. Crawley. 1996. Correlates of introduction success in exotic New Zealand birds. American Naturalist 147: 542–557. Levine, J. M. 2000. Species diversity and biological invasions: Relating local process to community pattern. Science 288: 852–854.
Companion papers Beirne, B. P. 1975. Biological control attempts by introductions against pest insects in the field in Canada. The Canadian Entomologist 107: 225–236. Memmott, J., P. G. Craze, H. M. Harman, P. Syrett, and S. V. Fowler. 2005. The effect of propagule size on the invasion of an alien insect. Journal of Animal Ecology 74:50-62. Cassey, P., T. M. Blackburn, R. P. Duncan, and J. L. Lockwood. 2005. Lessons from the establishment of exotic species: A meta-analytical case study using birds. Journal of Animal Ecology 74: 250–258. Marchetti, M. P., P. B. Moyle, and R. Levine. 2004a. Alien fishes in California watersheds: Characteristics of successful and failed invaders. Ecological Applications 14: 587–596. Lockwood, J.L., P. Cassey, and T.M. Blackburn. 2005. The role of propagule pressure in explaining species invasions. Trends in Ecology and Evolution 20: 223–228.
5 Disturbance
Overview Disturbance facilitates invasions. This statement, or variations of it, is perhaps one of the most commonly accepted truisms in the field of invasion ecology. The role of disturbance in community structure and the maintenance of species diversity is one of the fundamental tenets of modern ecological theory (Connell 1979). Elton (1958) commented on this relationship when he suggested that the majority of our non-native species are found in humandisturbed areas. At first glance this seems like a perfectly reasonable statement. Disturbance by its nature creates openings or “invasion windows” (Johnstone 1986) that non-native species can occupy. Because disturbance creates vacant space, releases resources, and alters species interactions, ecologists have often viewed it as a necessary precursor to the establishment of many non-native species. Indeed the literature is replete with examples of this phenomenon, especially for plant systems (for a review see Hobbs & Huenneke 1992). Nevertheless, the more ecologists examine the relationship between disturbance and the establishment of non-native species, the more complicated we recognize it to be. We start this chapter by giving a historical perspective of the role of disturbance in ecology and discussing what constitutes a disturbance, both in terms of definitions and sources. As we discuss how disturbances differ from one another, we uncover some limitations in our definitions. Readers with a strong ecological background will find some review in this first section. The central section of this chapter explores in detail the biotic and abiotic factors that change after a disturbance and that allow non-native species to establish or increase their performance. By examining our initial generalization, that disturbance promotes invasion, we hope to illustrate the complexity of the disturbance–invasion relationship. The remainder of the chapter focuses on the interplay between the study of disturbance, invasions, and restoration ecology; the effects of urbanization and agriculture on non-native species richness; how invasions themselves can act as disturbances; and, finally, how global climate change may affect the prevalence of non-native species.
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History and definition of disturbance A brief history of disturbance in ecology Ideas about disturbance and its effects have a long and varied history within the field of ecology. Early ecologists, including Lyell in 1832 and Darwin in 1859, commented on the ways in which disturbances altered the composition of communities (Wilkinson 2002). By the twentieth century, disturbance dynamics were viewed as occasional hiccups in the development of communities and ecosystems. For example, Fredric Clements (1936) noticed that disturbance could slow or alter the process of forest succession, but he generally dismissed these effects as small steps on the inevitable path to an unchanging forest community composition. This dismissive attitude was typical during the early decades of the twentieth century and was reinforced by the equilibrium-based population and community models developed from the 1930s through the 1970s. Based on these models all variation was directional, toward a stable equilibrial abundance of interacting taxa. Despite empirical information, such as the careful work by Andrewartha and Birch (1954) on variability in insect abundances, it took decades for non-equilibrium views to gain acceptance and for the every day importance of disturbance in community dynamics to gain recognition (McIntosh 1985). In reaction to the static and stability driven ideas of the past decades, many ecology texts of the 1980s emphasized non-equilibrium theories and models, patch dynamics, and disturbance regimes. As a result, much of what we know about the role of disturbance comes from relatively recent research.
Definition of disturbance One of the difficulties with a general topic such as disturbance lies in finding a workable yet comprehensive definition. The most commonly cited definition is, “any relatively discrete event in time that disrupts ecosystem, community or population structure and changes resources, substrate availability, or the physical environment” (White & Pickett 1985, p. 7). This definition will work for our purposes, as would any definition that includes a change in habitat and resource availability. Within this definition there is a broad range of events and processes that will lead to these sorts of changes. This range can be described by the intensity, frequency, duration, predictability, distribution, and synergism of any specific disturbance regime (Sousa 1984; White & Pickett 1985; Petraitis et al. 1989). The intensity of disturbance describes its magnitude and severity. Frequency is the mean number of disturbances in a unit of time. Duration addresses the temporal extent of the disturbance. An important distinction between disturbances comes with recognition that some anthropogenic disturbances last a long time, whereas disturbances that we might call “natural” are more shortlived. Predictability describes the regularity of the occurrence and offers insight into how organisms may adapt to the disturbance. Distribution of disturbances is significant in invasion biology because it describes the size and spatial extent of the disturbance and will be important in Chapters 7 and 8 in discussions of geographic range expansion.
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Finally, synergism describes whether a disturbance precipitates or exacerbates further disturbances. Largely based on these disturbance characteristics, ecologists have devised several dichotomies for classifying disturbances. Disturbances can be natural or humancaused, biotic or abiotic, and originate externally or internally (Sousa 1984; Pickett et al. 1989). Natural disturbances are ones without anthropogenic roots, including such things as hurricanes, floods, and some forms of herbivory or predation. Human-caused disturbances are seen as categorically different from natural disturbances and include things like land use change, toxic spills, and alteration of flooding regimes. Although we may react differently to anthropogenic disturbances than to natural ones, it is not clear that they have fundamentally different effects (either facilitative or inhibitory) on non-native species establishment as we will see later. There is the possibility that human disturbances register further on all scales (intensity, frequency, duration, predictability, distribution, and synergism) than natural disturbances, but as yet there is no concrete evidence for this. The biotic and abiotic dichotomy distinguishes between physical and biological forms of disturbance. Abiotic disturbances are those forces that act outside of the biological realm, for example wind, volcanism, drought, asteroids, and hurricanes. Biological disturbances on the other hand are mediated by living organisms and include predation, herbivory, rooting behavior, plus all the anthropogenic forces mentioned above. A final dichotomy separates disturbances into either endogenous or exogenous, meaning those that originate from within the disturbed system and those from outside (White & Pickett 1985; Simberloff & Von Holle 1999). Endogenous disturbances have a historical or evolutionary presence in the system and produce change from within (such as successional forces or predation). In contrast, exogenous disturbances are those that lie outside the evolutionary range the system has experienced (such as asteroid impacts or volcanic eruptions). This feature of disturbances can be very important in invasions, particularly if an exogenous disturbance mimics forces within the evolutionary history of a non-native species but outside the selective forces that have shaped the native community. This sort of exogenous disturbance puts non-native species at an advantage. In order to understand the role of disturbance in invasions, it is critical, therefore, that we keep in mind the evolutionary history of both the disturbed community and any introduced species. Unfortunately the latter dichotomy is often hard to assess for natural systems (White & Pickett 1985) and becomes even more problematic when we consider nonnative species. For example, is the introduction of a predator an endogenous disturbance (due to predation) or exogenous (due to the predator being introduced from outside the system)? Perhaps this dichotomy might be better approached as a continuum with endogenous/exogenous as two end points (White & Pickett 1985), which would allow us to avoid becoming mired in a morass of categorization yet would retain the useful observation that all disturbances are not equal in terms of their effects on the establishment of non-native species. Any discussion of how to define a disturbance must include recognition of temporal scale. We often think of a disturbance as a discrete event, but discreteness depends
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on life span. For example, some definitions insist that disturbance is a process rather than a discrete event (Grime 1979; Petraitis et al. 1989). Events and processes differ primarily in temporal scale. For example, is a hurricane an event or a process? It depends on perspective and life span. Within a forest ecosystem, a hurricane is an event that removes some trees and creates substrate and increased light availability for young or newly arriving trees to establish. The trees that establish in these clearings may depend on seed banks and dispersal frequencies between habitat patches. For the entire community of forest trees, the hurricane is, therefore, an event. For seeds moved through hurricane winds and waters to a newly cleared patch of land, it is perhaps more of a process. How about floods or the passage of a herbivore that eats plant material and leaves behind hoof prints? Are these events or processes? The central point is that, regardless of scale, a disturbance is an ecological disruption that leads to some type of open opportunity or vacant area in a community (Box 5.1). It is these specifics of any disturbance and disturbance regime that are important in determining the effect of disturbance on the establishment of non-native species. What is clear is that there are potentially as many types of disturbance as there are organisms to interact with the disturbance. To add a further layer of complexity, disturbance of a system can result from the removal or alteration of an existing disturbance. For example in flowing water (lotic) systems there is a characteristic flow regime for every geographic region (referred to as a natural hydrological regime, see Poff et al. 1997). The flow regime often contains seasonal periods of flooding. In these systems, the normal bouts of flooding act as a disturbance within the stream communities, yet if the flow regime is altered and normal flooding is modified or removed the change in flooding can also act as a disturbance to the system (Michener & Haeuber 1998). This additional level of intricacy is common and can be found with many seasonal disturbance forces. These many facets of disturbance make it clear that there is no single qualifier to describe how disturbances affect ecosystem processes or community structure. The above list of definitions, though, should highlight the challenging and complex nature of ecological disturbance and remind us that making generalizations and predictions about the role of disturbance in invasion dynamics may be inherently fraught with difficulty. Despite this difficulty, we forge on and examine the evidence for disturbance promoting invasions.
Does disturbance promote invasion? Our approach to this tangled subject is a little different from what we have done in other chapters. Instead of supplying lists of generalizations about disturbance and invasion or burying you under a plethora of case studies, we have decided to present the first sentence of this chapter as a disturbance-invasion “straw man.” As Webster’s Dictionary defines it, a “straw man” is a flimsy argument set up for the purpose of being easily refuted. In our case, we will show that disturbance does promote invasion, but that as an ecological axiom the above statement is a gross oversimplification. From our definition(s) of disturbance, this complexity should be obvious, but the field of
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Box 5.1
Sooner or later
Most definitions of disturbance include as a component the fluctuation of resources. Mark Davis and colleagues suggest that, at least for plant communities, this fluctuation is the key piece to understanding why and when plants invade (Davis et al. 2000). Their theory, dubbed the fluctuating resource hypothesis, states “communities are more susceptible to invasion whenever there is an increase in the amount of unused resources” (Davis et al. 2000, p. 529). The theory is rooted in ecologists’ understanding of vegetative succession, and relies on the assumption that competition for available resources is the primary factor limiting newly arriving species from establishing self-sustaining populations within the community. If resources such as light, nutrients, water and space are available, a new arrival can establish. If these resources are not available, the newly arriving species cannot establish. An increase in the availability of resources can happen in two ways (Fig. 5.1). First, there can be an influx of resources beyond the ability of the existing species in the community to exploit them. This can happen for a variety of natural reasons including experiencing an unusually rainy year, or the loss of one or more individuals of a dominant species from wind storms or trampling by herbivores (i.e., creating light gaps). It can also happen because humans have added resources to the system at a relatively high rate. Examples of this include run-off of nitrogen or other minerals from farm fields into waterways or lower lying ground,
e→
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Figure 5.1 The theory of fluctuating resource availability suggests that a community’s susceptibility to invasion increases as resource availability increases. In the plot, resource availability (and therefore invasibility) increases as the trajectory moves further to the right and/or below the supply/uptake isocline. Resources in this model can increase for a number of reasons: a pulse in resource supply (a → b), a decline in resource uptake (a → c) or both (a → d). Reprinted with permission from Davis, M. A., J. P. Grime, and J. Thompson. 2000. Fluctuating resources in plant communities: a general theory of invasibility. Journal of Ecology 88: 528–534.
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adding fertilizer to the soil, dumping garbage, or clear-cutting a forest. The second way for resources to become available is through a reduction in the use of resources by resident members of the plant community. Any reduction in the health or number of individuals within the community will result in an increase in available resources. Disease, herbivory, or the physical removal of plants via human actions will each reduce resource uptake by the residents. Of course, any one of these actions can result in an upward pulse in resources while at the same time causing the resident community to reduce its resource uptake. Thus, these modes of creating available resources are not mutually exclusive. In fact, disturbances simultaneously affect both modes by decreasing the residents of the community and increasing the available resources for organisms entering the community. The fluctuating resource hypothesis provides a mechanism by which disturbance can facilitate the establishment of non-native plants. In doing so, it fills a gap in the invasion ecology literature and provides testable hypotheses for those ecologists interested in experimental approaches to understanding invasion. Importantly, it is also a dynamic theory in that it does not assume a community’s susceptibility to invasion is a static property; invasibility changes with time and circumstances. Instead, the hypothesis suggests that any community can be invaded, or any community can repel invaders, depending on the ratio of gross resource supplies versus resource uptake by resident species. This ratio will shift within and between years, providing yet another reason why predicting which species will invade what communities is difficult (Davis et al. 2000). Although not specifically designed to apply to animal (or other) communities, it makes sense that the theory would be to some degree applicable outside the plant realm. Testing this hypothesis with a range of species and communities could be very fruitful for invasions theory.
invasion ecology is overloaded with well-intentioned aphorisms, which unfortunately do not hold up across taxa and environments (Williamson 1996). It is important to show the source, the utility, and then the limits of these maxims. There are plenty of examples of disturbance facilitating the establishment of nonnative species. For example, in a simple but elegant experiment Minchinton (2002) investigated the effect of disturbance by wrack (dead mats of plant material moved by tides, waves, and storms) on the spread of the invasive common reed (Phragmites australis) in coastal marshes in New England, USA. In these northern marshes, the main agents of disturbance for the coastal marsh community are ice, sediments, and wrack. Minchinton (2002) manipulated the presence and timing of wrack deposition by adding wrack in the fall of 1997 and then removing it the following spring, before seedling shoots emerged. Addition of wrack in the fall resulted in more non-native seedlings and taller adults with more inflorescences (Minchinton 2002, and see Fig. 5.2). With this timing, therefore, wrack disturbance facilitates seedling establishment of the non-native P. australis by creating bare patches favorable for germination.
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Figure 5.2 The effects of wrack on the establishment of common reed vary considerably based on the timing of wrack deposition. Mean (plus one standard error) of density, height, and number of inflorescences of new shoots of common reed at the end of the growing season after experimental addition and then removal (striped bars) or no manipulation of wrack deposition (solid bars). Reprinted from Minchinton, T. E. 2002. Disturbance by wrack facilitates spread of Phragmites australis in a coastal marsh. Journal of Experimental Marine Biology and Ecology 281: 89–107, with permission from Elsevier.
Similar findings have been observed in many other systems, involving many types of disturbance, particularly fire (Hobbs & Huenneke 1992; D’Antonio 2000, and see Table 5.1). For example, in a study of non-native plant invasion across vernal pools in Northern California, Gerhardt and Collinge (2003) found that a history of fire increased non-native species richness (predominantly grasses and wildflowers) in natural pools. Increased invasion was likely due to abiotic soil changes following fire, as their results indicate that non-native species richness also increased with decreasing soil organic matter, pH and cation exchange capacity. In an earlier similar study Zedler and Scheid (1988) found that non-native ice plant (Carpobrotus edulis) increased from 0% to 26% of total vegetative cover in the three years following a fire amongst coastal chaparral vegetation near Santa Barbara, California. This dramatic change was attributed to
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Table 5.1 Examples illustrating how disturbance by fire can play a suite of different roles in the invasion process across plant communities worldwide. Study region (habitat)
Invader (common name)
Role of fire in the community
California (riparian)
Arundo donax (giant reed)
Invasion occurs without the fire but fire promotes thickening
California (maritime chaparral)
Carpobrotus edulis (highway iceplant)
Invades slowly without fire; germination reduced by fire but invasion enhanced by it
California (Mojave desert)
Bromus tectorum (cheatgrass)
Slow invasion without fire; thickening and dominance promoted by fire
Nevada (Great Basin desert)
Bromus rubens (red brome)
Invades without fire; enhanced by fire
Arizona (Sonora desert grassland)
Eragrostis lehmanniana (Lehmann’s lovegrass)
Can spread without fire but may be enhanced by fire
Hawaii (seasonal sub-montane woodland)
Schizachyrium condensatum (beard grass) & Melinis minutiflora (molasses grass)
Invades without fire, but M. minutiflora promoted by fire
Venezuela (llanos)
Hyparrhenia rufa (thatching grass/ jaragua)
Invades without fire but is promoted by fire
Brazil (cerrado)
Melinis minutiflora (molasses grass)
Can invade without fire but also promoted by fire
South Africa (mountain fynbos)
Pinus radiata (Monterey pine), Pinus halepensis
Invasion slow without fire but enhanced by fire
Source: Adapted from D’Antonio 2000.
extremely successful seedling establishment (700 seeds/hectare with a 70% survivorship) likely due to the evolutionary history of C. edulis with fire in its native range in South Africa and Australia (Zedler & Scheid 1988). Evidence for a positive invasion effect by disturbance is not limited to plants or terrestrial systems. For example, Marchetti et al. (2004a) examined patterns of non-native fish establishment across all the watersheds in California to determine if there were general landscape-level factors associated with fish invasion. They found that pervasive landscape-level disturbances, particularly urban development, water diversions, aqueducts, and agriculture, influenced the likelihood of non-native fish establishment. The more disturbed a watershed was, the greater the number of non-native species present (Fig. 5.3). In general, the alteration of the natural flow regime (or hydrograph) of a river or stream can be seen as a disturbance and is typically caused by the presence of a dam or as a result of water diversion for agricultural irrigation or urban consumption (Michener & Haeuber 1998). This type of disturbance has been implicated in the establishment of many non-native aquatic and riparian associated species. For example, Schreiber et al.
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Figure 5.3 Biplots of the species scores for the first two canonical axes from canonical correspondence analysis (CCA) showing the influence of environmental variables and forces of disturbance on individual fish species across the watersheds of California. The biplots have been separated into (a) native species and (b) non-native species for ease of interpretation. Arrows represent the correlation of the environmental and disturbance variables with the first two canonical axes. Points represent individual species’ relative association with the environmental and disturbance variables. For full explanation of the variables and species present see original text. Adapted from Marchetti, M. P., T. Light, P. B. Moyle, and J. Viers, 2004a. Fish invasions in California watersheds: testing hypotheses using landscape patterns. Ecological Applications 14: 1507–1525, with permission from the Ecological Society of America.
(2003) found a positive association between altered flow regimes and the invasion of the New Zealand mud snail (Potamopyrgus antipodarum) in Australia. Similarly, Brown and Ford (2002) found that an altered flow regime tended to favor non-native fish species in the Tuolumne River in California, USA. Altered flow regimes have also
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been implicated in the extensive salt cedar (Tamarix spp.) invasion of arid regions in both the southwestern United States and Australia (Griffin et al. 1989; Busch & Smith 1995). Once it has established, salt cedar further alters the flow regime by removing vast quantities of water through rapid evapotranspiration and by changing flooding because of the density of salt cedar stands (Busch & Smith 1995; Di Tomaso 1998). In a similar manner, Wilcox et al. (1985) found evidence that anthropogenic stabilization of seasonal water levels in the Cowles Bog Wetland Complex, Indiana, USA is correlated with the areal extent of cattail (Typha spp.) invasion over a fifty-four year period. Unfortunately, despite the large body of evidence reviewed above, the statement “disturbance promotes invasion” is not likely to become a fundamental law in ecology. There are numerous examples of situations where disturbance tends to prevent the establishment of non-native species (e.g., Smith & Knapp 1999). Furthermore, we suspect the number of such cases is underreported in the literature for a variety of reasons. One of the more important reasons relates to the configuration of most invasion ecology studies. Because information on non-native populations that failed to establish is difficult to come by, we tend to ignore them in discussions on the role of disturbance in establishment. Instead, the vast majority of invasion research related to disturbance looks at the effects of disturbance on non-native species already in the system. To explore the inherent problems associated with the aphorism that began this section we will re-examine three of the examples we used above. First, we go back to Minchinton’s study (2002) on the conditions that favor the establishment of the non-native P. australis in New England, USA. This study investigated three aspects of the ecology of wrack disturbance by (i) simulating natural disturbance by adding wrack for one growing season and monitoring P. australis emergence (discussed above); (ii) removing natural wrack and comparing the establishment, growth, and seed set of P. australis to control sites; and (iii) quantifying patterns of P. australis emergence where natural wrack was present. Despite the facilitative effect of wrack disturbance on P. australis establishment that we discussed above, in two of Minchinton’s experiments wrack actually inhibited the growth and establishment of the non-native P. australis. Adding wrack to plots where new shoots of P. australis had already emerged reduced the number of reed shoots surviving by the end of the growing season (Minchinton 2002). Additionally, the presence of naturally stranded wrack suppressed the emergence of new P. australis shoots (Fig. 5.2, Minchinton 2002). These results are in stark contrast to those presented above which showed that simulating the natural pattern of wrack disturbance enhanced density, size, and reproductive potential for the invader. Which is the dominant result? The situation is complicated as disturbance by wrack is temporally and seasonally variable. Seasonal storms can produce large piles of wrack, which tides and wind move around. There is also a positive feedback mechanism because P. australis can contribute to the total amount of wrack deposition. An overall answer is not clear, but due to the continued spread of P. australis, it is likely that any inhibition that is happening within New England, USA is not inordinately limiting establishment. We can also turn our attention back to the invasion of vernal pools in California (Gerhardt & Collinge 2003). In this system we find that soil disruption is an additional disturbance agent with the opposite effect of fire. Recall that fire in this system tends to
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promote non-native species establishment. One of the goals of this study was to compare non-native plant richness in both natural and restored vernal pools. Restored vernal pools were created by grading shallow depressions in the soil and removing the existing vegetation and seed bank (Gerhardt & Collinge 2003). Restored pools were found to contain lower numbers of eight non-native plant species and lower levels of overall nonnative plant cover (Gerhardt & Collinge 2003). The authors suggest that the creation of the pools may have reduced the existing soil resources as well as the non-native seed bank, and that this disturbance limitation may be a temporary phenomenon (Gerhardt & Collinge 2003). The authors also mention that such conflicting effects of disturbance on non-native species establishment (in this case, between soil disruption and fire) are consistent with other studies, implying that disturbances often have unpredictable effects on invasion (Gerhardt & Collinge 2003). Indeed if we examine just the effects of fire on invasion, we find there are many examples of both facilitation and inhibition of non-natives (Table 5.1; D’Antonio 2000). Finally, returning to the influence of disturbance and urbanization on the number of non-native fish species established across California, Marchetti et al. (2004a) found that not all non-native fish species have the same response to disturbance (Fig. 5.3). In fact eight of the forty-five non-native fish species in the state were negatively associated with disturbance (i.e., the species found below the zero line on the y-axis). The reasons for these negative associations are varied, but in general, a subset of non-native fishes released in California waters are better suited to undisturbed habitats. Three of them, all salmonids, require relatively pristine alpine habitat and therefore have only successfully established in mid- to high-elevation lakes and streams of the Sierra Nevada mountains, most of which were fishless prior to these releases and currently experience little disturbance. The other five non-native species that are negatively associated with non-native species have restricted distributions in northern California. Their negative association with disturbance is probably a result of historical factors because they were only released into these few, relatively unaltered watersheds. This example serves to highlight the often hidden role of history in our perception of the link between disturbance and non-native species establishment (e.g., McKinney 2001).
Straw man synthesis The key to understanding why this general statement is often wrong lies in acknowledging two fundamental, but not mutually exclusive, principles: (i) disturbance creates opportunities for natives and non-natives alike, and (ii) for disturbance to increase the prevalence of non-native species, there must be a source of non-native propagules (Chapter 4). Neither of these principles is a particularly novel insight. Indeed, you may have predicted one or more of them based on your readings elsewhere or in other chapters of this book. Yet these principles do warrant explicit recognition here because they have often been overlooked within the context of discussing the effect of disturbance on invasion. The first principle, that both natives and non-natives often respond positively to the resource fluxes associated with disturbance, was expanded upon by Daehler (2003). Daehler noted that there is often an implicit assumption within invasion ecology that
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non-natives tend to outperform native species especially in the face of disturbance. To test the generality of this assumption, he collated seventy-nine independent studies that compared performance measures (i.e., growth rates, competitive ability, and fecundity) of native and non-native plants. He found no statistical evidence that non-native plants outperformed natives (Daehler 2003). Instead, when performance was evaluated over more than one type of growing condition (55 examples), 94% of the studies showed that the natives tended to outperform the non-natives given the right growing conditions. The non-natives did have, on average, higher phenotypic plasticity than the natives indicating that they could respond to more varied conditions than the natives (Fig. 5.4). Daehler (2003) suggests this may be the more important piece to the puzzle in terms of the positive effect of disturbance (natural or anthropogenic) on the successful establishment and spread of plant invaders. The second principle has been elaborated upon by several recent articles (e.g., Higgins & Richardson 1998; Daehler 2003; Lake & Leishman 2004). Successful introductions require not just a ready source of propagules but a ready source of the right species of propagule. Too often ecologists tend to lump non-native species into one monolithic group when each species clearly has its own evolutionary history. Thus the response to disturbance we observe among non-native species is very much dependent on the “fit” between the disturbance characteristics and the species available to take advantage of that disturbance (Horvitz et al. 1998). For example, in a modeling exercise to mechanistically explain the invasion of pine trees in the southern hemisphere, Higgins and Richardson (1998) teased apart the influence of the disturbance regimes (fire and treefalls), the traits of the non-native trees available to colonize disturbed sites, and the environmental context of grassland and forest habitats (anthropogenic changes to the natural disturbance regimes). They found that increasing disturbance enhanced the success of the non-native pine trees as predicted; however, the traits of these species and the environmental context were strong interacting factors. Disturbances killed non-native individuals depending on the environmental context and the traits possessed by the non-native species (Higgins & Richardson 1998). Other disturbances enhanced the performance of non-native pines, again depending on the context and the species’ traits (Higgins & Richardson 1998). These results mirror several of the examples presented above. Higgins and Richardson (1998) concluded that it was useless to debate whether disturbance facilitated or impeded invasion because the effects of disturbance are context-specific. It is probably already clear from our approach to this chapter that we agree strongly with this conclusion. It is important to recognize, though, that human-assisted transport mechanisms tend to introduce more species that are well adapted to human disturbances or to ruderal environments than those that require pristine habitats. This bias may mean that anthropogenic disturbances promote the establishment of most non-native species.
Restoration and disturbance The above examples of human disturbance inhibiting non-native species establishment illustrate a principle point of crossover between the disciplines of restoration
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Figure 5.4 Summary of native species versus non-native species performance, according to whether the native performed better than (first column) or as well as (second column) the non-native under some conditions, or whether the non-native plant always performed better than the native (third column). Asterisk indicates significant differences using a chi-square exact test (P < 0.05). The column to the far right (n) is the number of independent comparisons for each measure of performance. The performance of non-native and native plants differ considerably depending on the environmental context in which they are growing. There is no consistent evidence that non-native species perform better after a disturbance. Reprinted from Daehler, C. C. 2003. Performance comparisons of co-occurring native and alien invasive plants: implications for conservation and restoration. Annual Review of Ecology and Systematics 34: 183–211, with permission from Annual Reviews, www.annualreviews.org.
Departure from natural disturbance regime
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Invader > native
Native ≥ invader
Departure from natural level of resource availability
Figure 5.5 Conceptual model predicting relative performance of native species versus nonnatives. The checkerboard region represents possible unnatural disturbance regimes that might favor the native species if resource availability is not extraordinarily high. Restoration ecologists are often (implicitly) exploring this relationship between disturbance and species composition (between natives and non-natives). Their task can be thought of as promoting a disturbance (or disturbance regime) that tips the balance back towards natives. Reprinted from Daehler, C. C. 2003. Performance comparisons of co-occurring native and alien invasive plants: implications for conservation and restoration. Annual Review of Ecology and Systematics 34: 183–211, with permission from Annual Reviews, www.annualreviews.org.
and invasion ecology. Anthropogenic disturbance, or alteration of natural disturbance regimes, is often the culprit in driving ecosystems and species to the brink of extinction. Restoration is the profession and science that endeavors to recover natural ecosystems after the disturbance is halted, or the disturbance regime re-created. Also, as anyone who has ever participated in a restoration project will attest, the practice of restoration can easily be viewed as a disturbance event in itself (as in Gerhardt & Collinge 2003). The physical and chemical removal of unwanted species (most of them non-natives), the addition of chemicals or minerals to offset the effects of prolonged agricultural application of fertilizer, or the disruption of soil necessary to create more natural water infiltration all fit our definition of a disturbance. Restoration professionals re-create natural disturbance regimes that favor native species over non-natives, thus tipping the balance toward desirable species. The results they obtain are important to understanding how disturbance regimes favor or inhibit the establishment of non-natives and vice versa (Fig. 5.5; D’Antonio & Meyerson 2002; Daehler 2003). For example, the effect of human introduced grazers (cattle in this case) can be a positive force to prevent invasions at least in certain environments and situations. Jaymee Marty (2005) examined the effect of cattle grazing on non-native plant richness within vernal pools in the Central Valley of California. Vernal pools in California are a priority for conservation efforts as they contain over 100 endemic vascular plant species and are rapidly being lost to agriculture and development. Marty (2005) examined the species composition of seventy-two vernal pools over a 3-year time period. Her experimental design involved four cattle grazing treatments: (i) ungrazed, (ii) dry season grazing, (iii) wet season grazing, and (iv) continuous grazing. The ungrazed plots contained between 60% and 130% more non-native species cover than all of the grazing treatments. In contrast, the continuously grazed plots contained between 20% and 40%
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more native grass cover than the seasonally grazed or ungrazed plots and had native species richness that was equal to or greater than all of the ungrazed plots (Marty 2005). Marty attributes this result to two major factors. First, the extensive evolutionary history of herbivory in these systems by herds of tule elk and pronghorn antelope, both of which are now absent from the landscape, left the system with many native species that evolved with browsing pressure and may require disturbance to persist. Although cattle graze differently than these native herbivores, they may create breaks in plant cover that are important for native community dynamics. Second, and perhaps most importantly, cattle selectively graze on grasses, which in this case are almost exclusively non-native species. The non-native grasses in the system access water earlier in the season than native plants and thereby change the hydrology of the vernal pools. Because vernal pool vegetation is dependent on periods of inundation, these hydroperiod changes fundamentally alter the plant communities. The introduction of cattle tended to favor natives and encourage the natural hydrology while disfavoring the non-natives. (We discuss grazing examples that differ from this one later in this chapter and in Chapter 6.) As Marty suggests, the exact reaction of natives and non-natives to grazing depends on the particular grazers and the evolutionary history of the plants that they eat. In a similar finding, flow manipulation that increased river flushing in the spring to mimic a more natural flow pattern downstream of dams on the San Juan River in the desert southwest of North America has tended to inhibit the establishment of non-native fishes (Fig. 5.6; Propst & Gido 2004). Increased spring flow disturbance was shown to decrease the abundance of the non-native western mosquitofish (Gambusia affinis) while also positively affecting three species of native fish (Propst & Gido 2004). At the same time, Propst and Gido (2004) found that the abundance of three species of summer-spawning non-native fish were positively related to the number of days that the summer mean daily discharge was less than 14 m3 /second. This result suggests that higher summer flows (a further disturbance of the natural flow pattern) would act to decrease the abundance of non-native fishes. In this situation, alterations of the flow pattern (disturbance) at different times of the year can act to prevent the invasion of non-native species. Indeed, altered flow regimes can also have surprising effects on non-native species. Marchetti and Moyle (2001) found that the flow regime below California dams in dry years favored a suite of non-native fish species over native species. Conversely when wet years occurred and the dams’ releases inadvertently mimicked a more natural flow regime, the abundance of native fishes increased with a corresponding decrease in the non-native fish abundance (Marchetti & Moyle 2001). Thus a human disturbance (the presence of a dam and the corresponding alteration of the flow patterns) was actually favoring native species during certain periods and under certain environmental conditions.
Agriculture and urbanization as disturbance Sometimes anthropogenic disturbances are short in duration and occur frequently through time, as is the case with small-scale toxic chemical spills, release of water
Disturbance Native fishes
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Figure 5.6 The left series of graphs shows correlations between annual density of all San Juan River native fishes and the three most common natives with mean daily spring discharge. The right series of graphs shows correlations between annual density of all nonnatives and three most common non-natives with mean daily spring discharge. The bottom two plots are correlations with a common native (left) and non-native (right) species with mean daily summer discharge <14 m3 /sec. Reprinted from Propst, D. L., and K. B. Gido. 2004. Responses of native and nonnative fishes to natural flow regime mimicry in the San Juan River. Transactions of the American Fisheries Society 133: 922–931, with permission from the Ecological Society of America.
across a dam, or the mowing of grasslands. As we have seen above, these disturbances often have unpredictable effects on the establishment of non-native species. However, most anthropogenic disturbances have lingering, long-term effects. Good examples include (i) the more persistent components of agriculture such as fertilizer additions or chemical pesticide use, (ii) dams and roads, and (iii) houses or buildings. These disturbances are often labeled “exogenous” because they lie outside the realm of the
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usual set of natural disturbance events. Thus, there are typically few native species that have evolved to respond to the opportunity these disturbances create. Long-term effects of such anthropogenic disturbances can substantially alter the ecological playing field, often tipping competitive interactions away from natives and toward any non-native species that can survive in this “brave new world.” There are a variety of anthropogenic disturbances that can create these new worlds. Two of the best studied are the lingering effects of agriculture and the multidimensional effects of urbanization.
Agriculture and invasion Myster (2004) describes four “crop signatures” that differentiate abandoned agricultural land from other land that has been disturbed. First, crops with extensive root systems and large root-to-shoot ratios often persist long after the farming of these crops has ended. This is especially common for abandoned hayfields, where hay grasses are simply left unmown and persist for many years (Myster & Pickett 1994). It is also common within abandoned citrus or nut orchards as many of these trees persist in residential backyards, as anyone living in a recently urbanized Mediterranean climate can attest. These species will have strong effects on soil nutrient content, soil stability, and other factors that would normally become “available” after a disturbance event. Second, root exudates from crops can remain long after the crops are gone. These exudates can change litter chemistry and alter the fungal, soil microbe, decomposer, and seed predator community composition. Third, most agricultural weeds will persist after the farmer has abandoned the land. Most, if not all, of these weeds are non-native species. Thus, abandoned agricultural land comes with an assortment of non-native plants already well established. Fourth, the continual application of fertilizers and herbicides produces residual abiotic and biotic effects. Abiotic effects include relatively high levels of phosphorus or nitrogen. Biotic effects include low abundances of some microbes and insects. Any of these crop signatures may represent enough of a shift of the baseline environment that most native species cannot take advantage of the conditions created. Given that many large areas were converted to agricultural production and are now being abandoned, the positive effect of agriculture on non-native species establishment is often observed across entire landscapes. For example, the island of Puerto Rico was nearly completely forested at the time of its colonization by Europeans in 1492 (Lugo 2004). Widespread deforestation followed European colonization, with many of those cleared lands being converted for agricultural production. By the 1940s only 6% of the original forest cover remained. In the latter half of the twentieth century, most of this agricultural land was abandoned (Fig. 5.7). These lands were characterized by Lugo (2004) as being hot, dry sites with compacted soils and low fertility. Although tree species native to Puerto Rico are adapted to take advantage of the flux in resources produced by hurricanes, floods, and fires, none of them are capable of surviving in the conditions created within former agricultural fields. Imported non-native trees created a source pool of various non-native species that were capable of colonizing these crop lands as agriculture waned. Even as successional processes have changed the composition of these sites through time, these non-native trees have persisted and remain
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Figure 5.7 Change in forest cover and active agriculture in Puerto Rico between 1700 and 2000. Forest cover declined dramatically as agriculture grew on the island. As the human population soared post-1900, agriculture slowly declined. In its place, forests are returning; however these forests are dominated by non-native trees. Reprinted from Lugo, A. L. 2004. The outcome of alien tree invasions in Puerto Rico. Frontiers in Ecology and the Environment 2: 265–273, with permission from the Ecological Society of America.
dominant today. Interestingly, these forests have a healthy understory composed of many native herbaceous plants. Thus, in many ways the ability of the non-natives to reclaim abandoned agricultural sites in Puerto Rico acts to preserve several native forest understory species. This is not necessarily true in other similar situations, even on Puerto Rico. The non-native cayeput tree (Melaleuca quinquenerva, also called Australian paperbark) is capable of colonizing abandoned agricultural lands that flood and burn. In fact, it is the only tree species in Puerto Rico, native or non-native, that can tolerate these conditions. Thus, it tends to form monocultures where these three disturbances interact. The same thing happens in south Florida, USA (Serbesoff-King 2003). Another non-native invasive tree in Florida, the Brazilian pepper (Schinus terebinthifolius), also forms dense monocultures within abandoned agricultural lands of the Everglades, an ecologically prized wetland ecosystem. The Everglades ecological playing field was altered by farmers rock-plowing soil (i.e., crushing the limestone bedrock to increase soil depth) in the early part of the twentieth century, and by the subsequent persistent addition of fertilizer to aid in the growth of vegetables. Available nutrients in the soil of rock-plowed lands increased soil concentrations of phosphorus 3-fold, zinc 6-fold, and copper 10-fold over unplowed soil (Li & Norland 2001). These conditions heavily favored species that could quickly take advantage of such a nutrient boon. Most plants native to the Everglades can grow in these conditions; however, the Brazilian pepper outcompetes them in this enriched environment (Li & Norland 2001). The end result is a near monoculture of Brazilian pepper in these abandoned agricultural areas within the Everglades.
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Urbanization and invasion Urbanization has a similar signature as agriculture in terms of its effects on nonnative species. There are a multitude of papers showing that non-native species richness increases with level of urbanization, or is correlated with human population density (e.g., Holway 1995; Vila & Pujadas 2001; Kennard et al. 2002). For example, the results illustrated in Figure 5.2 strongly suggest that non-native fish are associated with urbanization in California (Marchetti et al. 2004a). In a more broad-scale survey, McKinney (2001) found that non-native plant and fish diversity were higher in US states that are relatively more densely populated. The correlation was much stronger for plants than for fish, reflecting the fact that non-native fish were often transported and released into US waters for sport or game enhancement. These releases did not necessarily occur within more populated states. However, the transport and release of many plant species closely tracks increases in human population size and subsequent increases in the prevalence of horticultural or ornamental landscaping around houses and buildings. Furthermore, McKinney (2001) found that the length of European human settlement within a US state was correlated with non-native plant richness but not fish richness for largely the same reasons. In an interesting twist on biogeographic patterns in number of non-native species, McKinney (2001) found smaller US states tended to have more non-native plants than larger states (i.e., state area did not correlate with non-native plant richness). In the United States, smaller states have a longer history of human settlement, and the effect of time of settlement on non-native richness overshadows the effect of area (McKinney 2001). McKinney’s results (2001) raise an important caveat to the oft-found correlation between human population density and non-native species richness. Human settlements certainly create novel disturbance regimes, often compact and enrich the soil, and alter hydrologic regimes. Any of these conditions can be defined as a disturbance, and it is likely that they will facilitate the establishment of non-native species. However, humans increase propagule pressure (Chapter 4) at the same time that they alter disturbance regimes. It is difficult to know which of these forces is driving the correlation between population density and non-native species establishment. The same issue plagues our understanding of how roads, railways, trails, and other forms of habitat fragmentation increase non-native species richness (e.g., Forman & Deblinger 2000; Jesson et al. 2000; Gelbard & Belnap 2003). This same dynamic plays a strong role in agricultural disturbances, as well. As we discussed in Chapters 2, 3, and 4, many widespread invasive plants are transported around the globe as contaminants in crop seeds. Similarly, many insect invaders arrive as pests on crop plants and then spread to native systems. In all these cases, human disturbance interacts with propagule pressure to increase invasion success.
Biotic disturbance Another perspective on the interplay between disturbance and invasions is to consider the non-native species itself as the “disturbance” or as the agent that alters a
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disturbance regime. This perspective has been somewhat overlooked in the invasion ecology literature, often because it is difficult to tell if the establishment of a non-native species caused the disturbance (or change in regime) or was a response to the disturbance (or change) (Mack & D’Antonio 1998). This problem has been aptly summarized by MacDougall and Turkington (2005) as asking whether non-native species are the “drivers” or “passengers” of changes in ecosystems. In the above section, non-native species were generally the passengers. Here, we consider them as the drivers. This point of view requires us to think about the effects of non-native species from an ecosystem perspective. It is generally believed that some species can exert a particularly strong or controlling influence over community or ecosystem properties (Power et al. 1996). When this influence is out of proportion to the biomass of the species in the system, ecologists label these influential species “keystone species” (although this term engenders some controversy because of its ambiguous use in conservation) because they act in a pivotal role to hold up or support an entire suite of species (D’Antonio & Vitousek 1992; Power et al. 1996). When the influence is proportional to the abundance or biomass of the species but fundamentally alters ecosystem or community properties, these species are often called ecosystem engineers (Jones et al. 1994; Brown 1995b; Jones et al. 1997). We think of invasive species more frequently as ecosystem engineers because of the implication of alteration, but when either of these types of species is added to or removed from a community, the result is a fundamental change in species composition. These types of invasion are a significant threat to natural ecosystems (D’Antonio & Vitousek 1992). Indeed, if non-native species can themselves act as a disturbance, they may potentially facilitate other non-natives in a self-perpetuating system of positive feedback loops. This can result in an invasional meltdown scenario (Simberloff & Von Holle 1999) in which positive interactions among invaders enhance the probability of establishment, spread, or impact for other invaders (see Chapters 6, 8, & 9). Mack and D’Antonio (1998) found fifty-eight studies that clearly labeled the nonnative species as the “driver” of ecological change. These studies were drawn from around the world and include plants and animals, and aquatic and terrestrial systems. From this body of literature these authors divided effects of non-natives into two categories: (i) those that changed physical or biological aspects of the disturbance and, (ii) those that changed the response of the biota to the disturbance. We examine each of these briefly below.
Non-natives that alter disturbance regimes Some invaders can alter the physical forces or physical agents of disturbance. For example, grass invasions can severely alter a region’s fire regime, by changing the fire intensity, the fire frequency, or both (see Box 5.2). As we mentioned before, tamarix or salt cedar (Tamarix spp.) can change water availability and soil salinity, but it can also alter flooding regimes by dessicating water ways (Vitousek 1990) or by crowding water banks and reducing the spread of water across the landscape (Busch & Smith 1995; Di Tomaso 1998). Alteration of physical disturbance forces can also occur through geomorphological processes. Acacia mearnsii, an invasive riparian tree in South Africa, increases
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Box 5.2 Fields of fire One of the more well-studied interactions between disturbance and invasions is the connection between particular non-native plants and altered fire regimes (D’Antonio & Vitousek 1992). Grasses have very high surface area to volume ratios compared with woody tissue; their life form supports standing dead flammable material; and consequently they burn under a broader range of conditions than woody tissue (D’Antonio & Vitousek 1992). The establishment of non-native grasses in a fire-prone area can set in motion a grass/fire cycle that can increase the frequency of the natural fire regime, and thus by default decrease the intensity of these fires. Often the non-native grasses are able to recover from fire more rapidly than the native species, further reinforcing the cycle (D’Antonio & Vitousek 1992; Mack & D’Antonio 1998). These effects have been demonstrated in Hawaii, western North America, Central and South America, Australia, and South Africa (D’Antonio & Vitousek 1992). In a recent far-ranging review of the effects of non-native species on fire regimes, Brooks et al. (2004) provide many examples of non-native species that change the frequency, intensity, spatial scale, and seasonality of fire (Table 5.2). Non-native species that have large effects on fire regimes often represent elements novel to the ecosystem they enter. Traits of particular importance in relation to fire are the moisture content and chemical composition of the plant tissues (Brooks et al. 2004). These two factors often serve to increase fire fuel properties and lead to alterations in the frequency, intensity, or seasonality of fires. The manner in which non-native plants grow through space can also change fuel properties. Non-natives that grow vertically into the vegetation canopy (e.g., vines) often provide fire “ladders” where none formerly existed. This link between canopy levels serves to spread fire into the canopy where it would rarely go under natural conditions. In a similar manner, non-native plants that occupy large expanses of habitat and grow in a uniform density will often drastically increase the spatial extent of fires. Each of these factors can act in the reverse manner as well, decreasing fuel loads or the continuity of fuels (Fig. 5.8). The reduction in fire frequency, intensity, and spatial extent can have just as large effects on the native species as the alternative (Brooks et al. 2004). While it is more common to focus on non-native plants that increase fire frequency, extent, or intensity, there are invaders that decrease these variables. These changes in fire regime may disfavor native plants and favor the regeneration or establishment of a fire-suppressive invader. With either type of change to the system, the invader-controlled fire cycle may favor the establishment of other nonnative species, thus leading to a cycle of invasion and fire (Brooks et al. 2004). Breaking this cycle is a land manager’s nightmare. Where does one start? Remove the plants, manage the fires, or both? Brooks et al. (2004) suggest evaluating the situation based on three criteria. First, assess whether the established non-native plant (or plants) are fire promoters or fire suppressors. Second, determine whether there is still native habitat that can be protected by breaking the invasive plant-fire cycle. Third, determine the extent to
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which the native species pool still exists. This pool of species can be tapped as a source for restoration or regeneration of the natural plant community, including its natural fire regime. Management then proceeds through initiating eradication or reduction of the non-native plants that alter the fuel structure, establishing species that can either suppress or enhance the fuel bed and that can successfully compete with the non-native, altering fuel loads by either mechanically creating fuel breaks
Table 5.2 Methods by which non-native plants can change fuel and fire regime properties. Fuel property changed Intrinsic fuel properties Increased plant tissue flammability
Fire regime property
Examples
Increased fire frequency and intensity, and increased annual window of fire activity
Eucalyptus spp. in North America
Decreased fire frequency and intensity, and decreased annual window of fire activity
Opuntia spp. in Europe, Acacia saligna in Africa
Extrinsic fuel properties Increased fuel load
Increased fire intensity
Andropogon virginicus in South Pacific, Melaleuca quinquenervia in North America
Increased horizontal continuity
Increased fire frequency and extent
Cenchrus ciliaris in Australia, Hyparrhenia rufa in South America
Decreased horizontal continuity
Decreased fire frequency and extent
Mimosa pigra in Australia, Schinus terebinthifolius in North America
Increased vertical continuity
Surface fire to crown fire
Cryptostegia grandiflora in Australia, Chromolaena odorata in Africa
Decreased vertical continuity
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Bromus spp. and Avena spp. in North America
Change in packing ratio
Change in fire frequency, intensity, and annual window of fire activity
Acacia saligna and Hakea sericea in Africa, Sapium sebiferum in North America
Decreased plant tissue flammability
Source: From Brooks et al. 2004.
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or increasing the fuel’s flammability. Often these methods will need to be used in tandem with the use of prescribed burns. For example, it is commonly necessary to manually (or chemically) remove fire suppressors, manage fuel loads for increased flammability, and reintroduce fire into the system at a regular return interval to “push” the plant community back toward a more natural composition. In some instances, the invasive plant-fire cycle has so greatly altered the plant community that a return to the pre-invasion state is not possible. D’Antonio et al. (2000) found that the native tropical dry forests of Hawaii have been so altered by an invasive grass-fire cycle that, in some places, the forest species have completely disappeared. In these instances there are two options. First, managers can create Fire regime Frequency Intensity Extent Type Seasonality
Other ecosystem properties
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Alien plants Impact Impact=range x abundance x effect Invasive Spread beyond the sites of introduction Must overcome regional dispersal Environmental barriers Naturalized Establish self-perpetuating populations at the site of introduction. Must overcome local environmental and reproductive barriers
Introduction Propagules arrive in a new region beyond their previous geographical range. Must overcome major geographical barriers to dispersal
Figure 5.8 The non-native plant–fire regime cycle. Notice how the invasion process model used in this text plays into the cycle. Non-native species that tend to add novel traits relative to the response to fire must be introduced into the system before the fire regime can be altered. Once in, these species affect fire cycles through a variety of mechanisms. Reprinted from Brooks et al. (2004) Effects of invasive alien plants on fire regimes. BioScience 54: 677–688, with permission from American Institute of Biological Sciences.
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a plant community that has as many native species present as possible while accepting that non-natives will persist and the combination of native species may be quite novel (Brooks et al. 2004). Second, managers can introduce another non-native plant that counteracts some aspect of the invasive plant-fire cycle such as fuel continuity or flammability (Brooks et al. 2004). Clearly this last option should be approached with caution as well-intentioned management may lead to the replacement of one invasive plant problem with another. Perhaps the most effective way to deal with the ill-effects of the invasive-plant fire cycle is to prevent the introduction of species that are likely to alter fire regimes in the first place. The costs of the non-native and the probability that management will be successful change dramatically as the invasion process proceeds from transportation to establishment to spread (Fig. 5.8). We review several approaches to identifying problem non-natives at the transportation and release stage within Chapter 12, but suffice it to say here that most of these approaches do not expressly consider the possibility that a non-native plant will alter fire regimes. If prevention is not likely to be the answer, then certainly early detection and eradication of known problem non-native plants will prove less costly and more successful for land managers. However, the development of information networks that land managers can access and contribute to is still in its infancy.
stream bank erosion by uprooting during high flows. This disturbance leads to rapid colonization of exposed stream banks by other non-native species (Mack & D’Antonio 1998). Non-native species can also stabilize geomorphologic processes and prevent the normal course of disturbance. Grasses with extensive root systems can stabilize hillsides, mining sites, disturbed areas, and sand dunes. For example, invasive European beach grass (Ammophila arenaria) has been shown to alter dune formation in Australia, New Zealand, and North America (Mack & D’Antonio 1998).
Non-natives that are the disturbance Invaders themselves can sometimes act as intense agents of disturbance. This often occurs with grazing or rooting mammals. For example, large populations of feral animals (goats, sheep, cattle, waterbuffalo) can contribute to soil disturbance, bouts of intense herbivory, and altered fire cycles (Mack & D’Antonio 1998). Often these effects are relatively easy to see through the use of either enclosures or exclosures. Kotanen (1995) used these methods to examine the effects of feral pigs (Sus scrofa) in California grasslands. Feral pigs were introduced into North America by European colonists and have altered California grasslands and coastal scrub communities by overturning soil and digging up roots. Kotanen (1995) found that pigs turned over about 8% of the soil annually creating a mosaic of excavations and soil mounds. When Kotanen (1997)
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documented the effects of natural and experimental pig rootings he noted that these disturbances altered soil temperatures and increased the availability of soil nutrients. The suite of species that colonized disturbed plots varied with disturbance type because of the effects on soil quality and resource availability. Overall, disturbances increased the abundance and diversity of annual species, both native and non-native, with a bias toward non-native annual grasses. A very common scenario whereby the addition of a non-native species introduces a novel disturbance agent is grazing. Altering a grazing regime can occur in many ways: (i) addition of animals to a previously ungrazed area, (ii) addition of novel grazers to a system, and (iii) a change in the level of grazing pressure or removal of grazing from a system with a long grazing history (Hobbs & Huenneke 1992). For example, grazing animals are an important component of many grassland ecosystems, yet grazing domesticated livestock on grasslands is believed to alter the natural grazing regime and has been implicated in increases in non-native plant richness as well as the loss of native plant species (Hobbs & Huenneke 1992). Kimball and Schiffman (2003) examined this process in a recent study of the grasslands of California’s Carrizo Plain. These authors compared plant cover estimates (native and non-native) in grazed versus ungrazed areas and created an experimental array (field and laboratory) in which they simulated grazing by clipping. Their comparison of ungrazed and grazed areas indicated a significant increase in nonnative species cover with grazing (57% to 79% respectively) as well as a decrease in native species diversity (Kimball & Schiffman 2003). The results of their simulated grazing experiment suggested that the native species were negatively affected by clipping while the non-natives (primarily the European grasses Erodium cicutarium and Bromus madritensis) were largely unaffected. The fact that grazing in this situation favors the non-native species at the expense of the natives is believed to be due to the absence of native grazing herds in the recent evolutionary history of the community (Kimball & Schiffman 2003). Note that this result and explanation differ substantially from the example we mentioned previously in the discussion of grazing in the section on Restoration and Disturbance (Marty 2005). This difference could stem from the different evolutionary pressures from herbivory by browsers versus grazers.
Global climate change and non-natives For most ecologists, there is little doubt that the climate is warming in response to the addition of greenhouse gases to the atmosphere, in particular due to elevated levels of CO2 (Hansen et al. 2005; Root et al. 2005). The amount of CO2 in the atmosphere has increased exponentially through time, from about 280 parts per million (ppm) in pre-industrial times to 370 ppm in 1997 (Fig 5.9; Weltzin et al. 2003a). The amount of CO2 is predicted to increase well into the future, with some estimates placing CO2 concentrations at 540–970 ppm by the end of the twenty-first century (Weltzin et al. 2003a). Although ecologists have largely explored climate change and invasion independently (Weltzin et al. 2003a), there are clear overlaps with our discussion of the role
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Figure 5.9 Levels of global atmospheric CO2 for the last 1000 years, derived from measurement of CO2 in air bubble layers of ice extracted from a core drilled in Antarctica (beginning of line up to 1900), and from atmospheric measurements at Mauna Loa, Hawaii since 1958. Reprinted from Weltzin, J. F., R. T. Belote, and N. J. Sanders. 2003. Biological invaders in a greenhouse world: will elevated CO2 fuel plant invasions? Frontiers in Ecology and the Environment 1: 146–153, with permission from the Ecological Society of America.
of disturbance in invasion. The disturbances we have reviewed above alter the environmental context in which species are growing and reproducing. Climate change also alters the environmental context of non-natives. In the same way that agriculture leaves the soil enriched, increased atmospheric CO2 and nitrogen produce fundamental increases in the availability of carbon and nitrogen and alter temperature and water availability (Dukes & Mooney 1999; Dukes 2000). All of these elements are, of course, critical to plant growth, and alterations to plant communities will alter other species, such as animals and fungi. Atmospheric CO2 is the primary source of carbon for terrestrial plants and is fixed via photosynthesis. Weltzin et al. (2003a) suggest that all plants should respond strongly to increased CO2 ; however, not all plants respond in the same way. Since atmospheric CO2 is the primary source of carbon for terrestrial plants and is fixed via photosynthesis, the response of plants to increased CO2 should be mediated by their photosynthetic pathway and physiology (Weltzin et al. 2003a). In general, C3 photosynthetic pathways are the more common among terrestrial plants, and these plants tend to respond positively to increased availability of carbon. Those plants that fix carbon via the C4 photosynthetic pathway tend to respond in more varied and unpredictable ways to elevated
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carbon (Reynolds 1996). This is because the C4 pathway is more efficient in capturing CO2 ; thus it is not clear that the performance of these species is currently limited by a lack of available C. In one of the few studies to directly compare the performance (mean weight gain) of native and non-native plants grown under conditions of elevated CO2 , Dukes (2000) found both groups responded positively, and there were no statistical differences between the magnitude of the boost in growth rate between natives and non-natives. Echoing the results of Daehler (2003), Dukes (2000) found that there was no evidence to suggest that non-native plants as a group will necessarily respond more strongly to elevated CO2 ; all plants will respond regardless of the “invasion” status we give them. Elevated atmospheric CO2 has the indirect effect of altering global air and water temperatures (i.e., global warming); however the expected changes in temperature are much more spatially variable than the direct effects of increased available carbon and nitrogen. The effects of warming on non-native species prevalence has been studied mostly through the examination of long-term datasets, and the results indicate that the warming climate has already precipitated changes in non-native species’ distribution. For example, marine waters off the coast of New England, USA have increased in mean temperature over the past 25 years. Importantly, the annual maximum and minimum temperatures have also changed. Stachowicz et al. (2002b) examined sessile marine invertebrate species composition on panels that were submerged in the water in coastal New England, USA for an entire year. They repeated this process over the span of 10 years and matched this information to the mean and range of water temperatures for those years. The effect of an increasing minimum temperature seems to be the more important element in determining which species (native and non-native) can survive in these waters (Stachowicz et al. 2002b). Winter temperatures off New England normally can go below −1◦ C and such an extreme in temperature is a strong biological filter on which species can survive. Stachowicz et al. (2002b) reported that a group of three common non-native species of sea squirts (ascidians) tended to recruit (i.e., settle on the panels) earlier than any of the native species during warmer years (Fig. 5.10). The authors point out that several factors could affect how well these species do beyond the initial timing of recruitment. To test for a longer-term effect of water temperature they calculated total annual recruitment (i.e., total abundance of each species on the panel each year) for natives and non-natives. Total annual recruitment was positively correlated to mean winter temperature for the non-natives but negatively correlated for the natives (Fig. 5.11). This suggests that the marine communities of New England could shift toward a dominance of non-native species if annual winter minimum water temperatures continue to rise. This same effect can be seen in the Antarctic (Frenot et al. 2005). Over the last two centuries, non-native microbes, fungi, plants, and animals have become established in the Antarctic. Most of these species were vectored via the movement of people and cargo for industrial, scientific, and tourist operations. The vast majority of the non-natives here are native to Europe, and these species have a number of important impacts on the native species and ecosystems of the Antarctic. Frenot et al. (2005) suggest that, with rapid warming of the climate clearly occurring in this quite cold and inhospitable
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Figure 5.10 Panel (a) shows mean winter water temperature for areas off the coast of New England (USA) as measured over the past 30 years. Annual minimum water temperatures have steadily risen through time. Panels (b–d) show March 1st water temperature is strongly negatively correlated with the timing of initial recruitment for three non-native sea squirts. The dotted line box in Panel (a) indicates the time frame included in the data used for Panels (b–d). Reprinted from Stachowicz, J.J. et al. 2002b. Linking climate change and biological invasions: ocean warming facilitates nonindigenous species invasion. PNAS 99: 15497–15500, with permission from National Academy of Sciences, USA.
region, there are likely to be many more non-natives that are capable of establishing selfsustaining populations. For this to happen, the current non-native transport vectors need not change. They probably bring many more propagules of non-native species than can currently establish there for the obvious reason that most species in the world are not adapted to the cold and dark climate of the Antarctic (Chapter 3). However, this stasis in establishment is not likely to last, thus highlighting the need to impose strong controls on the cargo and gear being brought into the Antarctic. If the air temperatures of this region increase in the future at the same time as transportation of non-natives increases, there is the potential for a profound change in the composition and function of this ecosystem (Frenot et al. 2005). Together these two studies suggest that very high and low latitude communities may be particularly susceptible to invaders from warmer climes if global warming
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Figure 5.11 Total annual recruitment of non-native sea squirts is positively correlated to mean water temperature during the preceding winter, whereas native species recruitment is negatively correlated with winter temperatures. This indicates that marine communities in New England (USA) may shift toward a dominance of non-natives if water temperatures continue to increase. Reprinted from Stachowicz, J.J. et al. 2002. Linking climate change and biological invasions: ocean warming facilitates nonindigenous species invasion. PNAS 99: 15497–15500, with permission from National Academy of Sciences, USA.
continues. This could prove quite overwhelming to these locations since there are many, many more species that could shift their range toward colder climates than the reverse. Far northern or southern communities are relatively depauperate in species number, mimicking in many ways the situation for very isolated oceanic islands. It is not hard to imagine that, as the climate warms, these formerly extreme environments will increase their species richness several fold (e.g., Sax et al. 2002). Another approach to investigating how altered climate will affect the distribution of non-native species is to model the environmental controls on species’ current range boundaries and then project how these boundaries might change under future altered climate scenarios. Most efforts to date are based on the niche modeling approach of Peterson et al. (2003; see Chapter 12 for a full description of this approach). The underlying concept is to model the ecological niche of a species using a series of climate variables measured in its native range. The suite of variables that correlate strongly with the range limits are then projected onto a map of future climatic conditions or the conditions of an area where the species is regularly introduced (or has become marginally established). The result is a visual representation of where this species can be found in the future. For example, Roura-Pascual et al. (2004) projected the worldwide distribution of the highly invasive Argentine ant (Linepithema humile), a native of South America. The models these authors developed did a good job of describing the ant’s native and current non-native range. When the ant’s geographical range was projected under future altered climates (into the 2050s), Roura-Pascual et al. (2004) found that
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the Argentine ant is most likely to establish more non-native populations in more northern latitudes, but suffer range contractions in tropical regions, including within its native range. Although compelling, there are two missing pieces to research on the role of climate change on non-native species. First, most research to date has focused on predicting if a currently harmful invader will substantially expand its range under a future altered climate, thus making the invader an ecological or economic problem for currently uninvaded locations. This is a sensible place to start, since this approach has very practical implications for the management of these harmful invaders (Chapter 12). However, to fully understand the complete suite of effects that altered climate imposes on non-native species, we must investigate the influence of climate change on nonnatives that are not yet problems or are not yet established in the system (e.g., Weltzin et al. 2003a). Second, there is a glaring need to examine the role of altered CO2 , nitrogen, and temperature on species embedded within a community (Dukes & Mooney 1999; Weltzin et al. 2003a). Most experiments on the effects of altered climate examine the growth of plants in monoculture or individually, but growth under such controlled conditions may tell us very little about responses in field situations (Dukes & Mooney 1999; Dukes 2002).
Summary It is clear that disturbance and invasion can be strongly linked but that the links are not always easy to see a priori. Indeed a good amount of solid natural history work on the system of interest and potential invaders needs to be in place before making predictions about the influence of disturbance on the fate of particular suites of non-native species. The dependence on context leaves the interaction between disturbance and invasions resistant to generalizations. There is very little evidence to suggest that non-native species outperform natives after disturbance. Instead, situations where non-natives dominate are characterized by the system experiencing a disturbance event that is well outside the evolutionary history of the natives. This is common when considering the widespread and multidimensional impacts of urbanization and agriculture. Non-native species can themselves be the drivers of ecosystem change either in the form of creating a disturbance or altering a disturbance regime. Even from this perspective, we see strong evidence that the effect these non-natives have is dependent on how their inherent traits interact with the community of native species and physical ecosystem properties. Finally, there is so much ground to cover in order to understand how global climate change affects non-native species that our review reads like a horoscope, compelling but so general as to be nearly always true. We encourage readers to explore climate change and other aspects of disturbance, taking great care to (i) include sites that are disturbed but currently un-invaded, (ii) find and utilize information on failed introductions in disturbed environments, (iii) include a range of non-native species including those that currently do not appear to be well established or are not expected a priori to respond to the disturbance, and (iv) consider the effects of disturbance, including global climate change, within a community ecology framework.
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Further complexities Brooks, M. L., C. M. D’Antonio, D. M. Richardson, J. B. Grace, J. E. Keeley, J. M. DiTomaso, R. J. Hobbs, M. Pellant, and D. Pyke. 2004. Effects of invasive alien plants on fire regimes. BioScience 54: 677–688. Daehler, C. C. 2003. Performance comparisons of co-occurring native and alien invasive plants: implications for conservation and restoration. Annual Review of Ecology and Systematics 34: 183–211. Frenot, Y., S. L. Chown, J. Whinam, P. M. Selkirk, P. Convey, M. Skotnicki, and D. M. Bergstrom. 2005. Biological invasions in the Antarctic: extent, impacts and implications. Biological Reviews 80: 45–72. Hobbs, R. J., and L. F. Huenneke. 1992. Disturbance, diversity, and invasion: implications for conservation. Conservation Biology 6: 324–337. D’Antonio, C. M., and L. A. Meyerson. 2002. Problems and solutions in ecological restoration: a synthesis. Restoration Ecology 10: 703–713.
Companion papers Higgins, S. I., and D. M. Richardson. 1998. Pine invasions in the southern hemisphere: modeling interactions between organism, environment and disturbance. Plant Ecology 135: 79–93. Lake, J. C., and M. R. Leishman. 2004. Invasion success of exotic plants in natural ecosystems: the role of disturbance, plant attributes and freedom from herbivores. Biological Conservation 117: 215–226. Stachowicz, J. J., J. R. Terwin, R. B. Whitlach, and R. W. Osman. 2002. Linking climate change and biological invasions: ocean warming facilitates nonindigenous species invasions. Proceedings of the National Academy of Sciences 99: 15497–15500. Lugo, A. E. 2004. The outcome of alien tree invasions in Puerto Rico. Frontiers in Ecology and the Environment 2: 265–273. Weltzin, J. F., R. T. Belote, and N. J. Sanders. 2003. Biological invaders in a greenhouse world: will elevated CO2 fuel plant invasions? Frontiers in Ecology and the Environment 1: 146–153.
6 Establishment Success: The Influence of Biotic Interactions Overview One of the more persistent conceptual paradigms within invasion ecology is the notion of biological resistance (Elton 1958; Simberloff 1986). This paradigm suggests that native species can thwart the establishment of nonnatives provided there are enough natives present to commandeer available resources. The most commonly explored mechanism for the relationship between invasion resistance and species diversity is competitive exclusion, whereby two species cannot share the same niche space (e.g., food resources) without one of the two eventually going extinct (Hutchinson 1959). We will explore this possibility in some depth in this chapter. In terms of the successful establishment of non-natives, though, we can imagine that other negative interactions such as predation, parasitism, or herbivory could play a large role. In addition, positive mutualistic interactions may determine whether a non-native species establishes (Bertness & Calloway 1994; Bronstein 1994; Stachowicz 2001). Simply by growing, many species alter the local environment by providing shade, cover, or substrate that allows other individuals to flourish (Bertness & Hacker 1994). We will explore both sides of this ecological coin. We will start by exploring interactions that impede establishment of non-native species and then look at interactions that facilitate them. Within each section we will follow convention by exploring negative interactions first, then positive. Note that it is important not to confuse the sign (positive or negative) of the biotic interaction with its role in establishment outcomes. For example, mutualisms benefit both partners in the interaction, but mutualisms can have either a facilitative or inhibitory influence on the establishment by non-native species. Obviously, the successful creation of a mutualism with a common species – native or non-native – may facilitate establishment (Table 6.1), but mutualisms can also impede establishment. The failure to create a mutualism with a native species may prohibit the establishment of a non-native, and mutualisms between native species may give the critical edge that allows natives to outcompete non-natives and exclude them from particular environments. The same is true for predation and competition (Table 6.1). We suggest students take a moment and decipher Table 6.1 before proceeding as it forms the outline for the chapter.
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Conceptual issues There are three complicating factors to this discussion that warrant some attention before we proceed. First, it is extremely difficult to observe establishment events directly. Our inability to predict when and where new non-native species will enter an environment makes it hard for us to collect natural, unbiased data, and ethics make it difficult to carry out large-scale, experimental introductions. Second, it is important to keep in mind the difference between establishment and spread when considering the topics in this chapter. The different factors that promote establishment and spread involve a subtle difference in population growth but a fundamental difference in biology. Establishment is determined by whether or not a population can survive at all in a specific location, a difference between some reproduction and none. Spread is governed by the much more luxurious difference between positive population growth and stasis, and spread must also always involve movement of individuals into at least one new location. The difference may appear subtle on paper; nevertheless, it is a clear conceptual difference that profoundly influences the ubiquity and impact of the non-native. For this reason, we will find ourselves revisiting many of the same themes covered here in the chapters on spread (Chapters 7 & 8) and impacts (Chapters 9 & 10). Third, there is some confusion in the literature around the term “biotic resistance.” While some researchers look at any negative influence on establishment as the fundamental component of resistance, others have interpreted resistance more strictly,
Table 6.1 Native species, and previously established non-natives, can either inhibit or facilitate the successful establishment of a newly arriving non-native species. The biotic interactions involved in each case include both positive and negative associations. This table forms an outline of this chapter, and should help you differentiate between the sign of the interaction and its effect on probability of non-native species establishment. Effect on establishment
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Native pine trees repelling non-native pines on Isla Victoria, Argentina
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Non-native fig trees failing to establish in Florida, USA, in the absence of their obligate mutualist fig wasps
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Non-native annual grass establishing underneath native shrub in Australia
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Non-native bullfrogs establishing in ponds in Oregon, USA, in the presence of their predators’ predator
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Native mycorrhizae establishing an association with non-native plants allowing the non-native to survive
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limiting the term to describe repelling non-natives completely. We do not use “biotic resistance” per se very frequently in this chapter, partially to avoid the semantic discussions, but also because we feel that this division creates a false dichotomy. In general we use the term “resistance” broadly, to imply a reduction in establishment probability. In this chapter we are most interested in interactions that directly prohibit establishment, although we recognize that such resistance may be place and time dependent, and we do not think that the inability to restrict establishment implies an inability to limit population growth or the spread of a species after establishment.
Resistance to invasion Competition A large proportion of ecological studies over the past half century have focused on competitive interactions, and Elton (1958) expected competitive interactions to act as the frontline in defense against invasion. With more native species in a community, there are more roles in the environment (i.e., niches) that are already occupied by a native species. Any invader is, therefore, more likely to overlap with a native species in habitat and resource use and is less likely to gain a foothold. Simple Lotka–Volterra equations support the theory behind this hypothesis by suggesting that superior competitors (or equal competitors that are already present) will exclude newly arriving non-native species. We see exactly this sort of interaction in many empirical examples of nonnatives that establish in habitats with fewer species but that fail to establish in adjacent habitas with more species. The most common examples of competitive resistance come from studies on plant communities, not because these interactions are rare elsewhere but because interactions between plants are less likely to cross into predation or parasitism than interactions between animals. A good example of competitive resistance comes from Simberloff and colleagues (2002) who examined forests on Isla Victoria, Patagonia, Argentina. These ecologists looked for non-native pine species in the native forest and in adjacent meadows that harbored no pine competitors. Isla Victoria is an island in Lake Nahuel Huapi in the Nahuel Huapi National Park. The park was established in 1934, but a plant nursery continued on the island into the 1960s, with hundreds of thousands of non-native trees planted in the center of the island. In 2000 and 2001 Simberloff and his colleagues looked for indications that nine of these non-native pine species had established in the nearby native forest. They chose these nine species based on a study by Rejmanek and Richardson (1996) indicating that these particular species were some of the most common pine invaders globally. They found almost no adult trees of these nine non-native pines in intact forest despite ample evidence that the trees could reproduce in the island’s harsh climate (i.e., there were plenty of seedlings and established adults in cleared or cultivated meadows on the island). Similar results have been found in intact grasslands (Petryna et al. 2002), shrublands (Keeley et al. 2003), and other forest communities. For example, Green et al. (2004) found the same results in a tropical forest on Christmas Island in the Indian Ocean.
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Figure 6.1 Percent survival of (a) Leucaena, (b) Muntingia, (c) Adenanthera, and (d) Clausena – all non-natives – within the forest edge (open circles) and forest interior (closed circles) on Christmas Island. Some seedlings within each habitat were also caged (solid line) to protect them from herbivory by the red land crab (dashed lines indicate uncaged seedlings). Nearly all seedlings of Leucaena and Muntingia died within forest interior due to competition with the natives for light. Herbivory by the red land crab reduced the survivorship of seedlings of all species, although this effect was stronger in the forest edge habitat. Reprinted from Green, P. T., P. S. Lake, and D. J. O’Dowd. 2004. Resistance of island rainforest to invasion by alien plants: influence of microhabitat and herbivory on seedling performance. Biological Invasions 6: 1–9 with kind permission from Springer Science and Business Media.
They placed seedlings of four different non-native tree species (Leucaena leucocephela, Muntingia calabura, Adenanthera pavonia, and Clausena excavata) in cages (to protect them from herbivores) along roads, forest edges, and in intact forest interior and found that for two of these species (Muntingia calabura and Leucaena leucocephela) none of the individuals placed in the competitor-rich forest interior survived (Fig. 6.1, Green et al. 2004). Because the bulk of the biomass in forest communities comes from long-lived species with many chances to replace themselves with their own offspring, recruitment into canopy stage classes can take a long time. This slow process of change is sometimes called “biotic inertia” and can contribute to invasion resistance (Von Holle et al. 2003). On the other hand, the data from Simberloff et al. (2002) and Green et al. (2004) reflect almost half a century of potential invasion time and suggest that competition with native
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species within intact forest helps to prohibit the establishment of the non-natives. As we will see below, other factors may also play into these examples of resistance. In general plant communities where the native species occupy nearly all bare ground or capture almost all light at the soil surface limit the establishment success of nonnative species (Hobbs & Huenneke 1992). This result is so strong that many researchers state that disturbance of native plant communities, with its consequent removal of competitors that resist establishment, is one of the most important factors in establishment of non-native plants (Hobbs & Huenneke 1992; Parker et al. 1993; Rejmanek 1996). However, as we discussed in Chapter 5, this generalization can be an oversimplification because there are plenty of examples where non-native species have established in undisturbed habitats (Almasi 2000; Fine 2002; Ellsworth et al. 2004). Elton’s original conjecture was that more species tend to use more of the resources. That does not necessarily mean that a non-native species will completely overlap with one particular species in that community, and thus fail to establish. It is possible that several members of the native community will each use a portion of the resources necessary for the non-native species to establish (Davis et al. 2000). Similarly, the number of species in the community may be less important in preventing the establishment of the non-native species than the number of functional groups, or ways of using resources (Cowling et al. 1994; Prieur-Richard et al. 2000; Symstad 2000). A huge debate has focused on the roles of species richness, species function, and resource use in invasibility (Tilman 1997; Levine & D’Antonio 1999; Naeem 2002). We cover this debate in more detail in Box 6.1. Clearly, field tests of all of the potential ways in which diversity might affect nonnative species establishment is warranted; however such investigations have proven monumentally daunting to those ecologists who have tried. Empiricists have, instead, examined the effects of diversity on establishment in laboratory communities (Drake 1991; Holyoak & Lawler 1996b, 1996a). These “bottle experiments” offer a good way to expand invasion theory without introducing new non-native species, but results have been hard to generalize, partially because they are dependent on the particular lab conditions without really being dependent on critical features of natural ecological communities. Importantly for invasion ecology, these bottle experiments have demonstrated that the order in which organisms enter a community can determine final community structure and the strength of competitive interactions (Drake 1991; Price & Morin 2004). This result indicates that biotic resistance can be critical, but that it is not universal, is dependent on species identities, and is not necessarily dependent on species richness. Mathematical modeling also offers a way to explore a diversity of ecological outcomes without actually introducing non-native species into pristine habitats and without confounding lab effects. Many theoreticians have examined Elton’s hypothesis to see if communities with more species are more likely to repel a newly arriving non-native species than communities with fewer species (May 1972; Goodman 1975; Pimm 1979). Most of these models have assembled communities and then looked at how adding a “non-native” species affects the stability of the community as a whole. That is, what happens to the abundances of native and non-native species rather than just whether or not the non-native can establish a self-sustaining population. In general these models
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Box 6.1 Land of confusion One of Charles Elton’s arguments in The Ecology of Invasions by Animals and Plants was that places with many native species were more likely to repel non-native species than places with few native species. Elton’s hypothesis has been embraced, potentially for two reasons. First, it fits well with basic competition models that suggest that more competitors are more likely to use up resource space and leave few empty niches. Second, it is reassuring to conservation biologists to think that hotspots of native biodiversity may be less vulnerable to invasion than coldspots, those places with few native species (Kennedy et al. 2002). On the other hand, there is an inherent paradox in the idea that areas with high diversity of native species are unlikely to have high diversity of non-native species. It makes more sense that the factors that promote high diversity for native species may promote high diversity for non-native species, as well. Not surprisingly, this topic has support on both sides. Numerous small-scale experiments on plant communities (e.g., Tilman 1999; Naeem et al. 2000; Kennedy et al. 2002) and a few mathematical models (e.g., Post & Pimm 1983; Case 1990) have supported Elton’s hypothesis, but several large-scale observational studies and more mathematical models have found the complete opposite relationship (Fig. 6.2, Lonsdale 1999; Stohlgren et al. 1999, 2003). Because of these conflicting results, this topic has generated more debate than any other facet of invasion ecology. Here we will attempt to shed a little light on the debate and the drivers behind the different results and to rephrase the set of experimental questions to increase our insight into the ecology of invasions. There are many potential reasons for these differing results, but let us focus first on confounding factors at each scale. For example, proponents of the diversity– invasibility hypothesis, as Elton’s argument has been called, point out that patterns in observations can be influenced by a wide variety of factors, while controlled experiments truly test the relationships of interest. It is true that observational studies may combine the effects of propagule pressure with the effects of richness in ways that controlled experiments probably do not. Plus, environmental changes and the role of habitat heterogeneity may play a role in the observational studies but not in the experiments. Neither approach is perfect, though. Although small-scale studies generally do find that high diversity plots are less likely to be invaded than low diversity plots, several researchers have pointed out that this result may be driven by the particular species in the high and low diversity plots rather than by species richness per se. First of all, by including more species, high diversity plots are more likely to include a species that has a large positive or negative effect (Huston 1997). David Wardle (2001) examined recent experiments that found a positive relationship between species richness and resistance to the establishment of non-native species. He found that for each experiment the result was driven by a “sampling effect,” in which the most competitive species or species combination was more likely to be included in the high diversity than the low diversity treatments. Without these superior competitors, the results might have been different. Melinda Smith and her coauthors (2004b) found that in a tall grass
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prairie system, where the number of successfully established non-native species initially seemed tied to species richness (Smith & Knapp 2001), it was actually the presence of the single dominant competitor species that was the most important factor in the success or failure of establishment. Several researchers NH 0.90
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Figure 6.2 Map showing whether the relationship is positive (+) or negative (−) and correlation coefficients (r) of native to non-native plant species richness in counties for each state. The graph below the map shows the correlation for Alabama as an example. States in bold show a significant positive relationship, those not in bold show a non-significant relationship. Hawaii shows the single negative (non-significant) relationship. Figure reprinted with permission from Stohlgren, T. J., D. T. Barnett, and J. T. Kartesz. 2003. The rich get richer: patterns of plant invasions in the United States. Frontiers in Ecology and Environment 1: 11–14.
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have also suggested that, if the filling of niches is really the deterrent to establishment, then the diversity of species functional groups may be more important than the number of species (Cowling et al. 1994; Prieur-Richard et al. 2000; Symstad 2000). Finally, what one superior species can do well, many inferior individuals may also do well; consequently the density of competitors may play a role in limiting establishment success (Weltzin et al. 2003b). The theme here is basically that resource acquisition and monopolization limit establishment. The above paragraph condenses decades of debate into (i) problems with largescale studies and (ii) a handful of studies that isolate the factors that drive results at small scales. Although such discussion has increased our understanding of diversity relationships and experimental design, it does not satisfactorily explain the difference in results at large and small scales. There have been several recent attempts to explain the contrasting patterns at different scales (e.g., Levine & D’Antonio 1999; Shea & Chesson 2002). The main thread that unravels the paradox is the explicit acknowledgment of how conditions and the role of competition change with spatial scale (Huston 1999). At small scales, interspecific interactions such as competition have a stronger role to play in determining community membership. At this “neighborhood” scale, individual plants directly compete for resources such as water, light, and nutrients. At larger spatial scales, the importance of competition for these resources diminishes, and in its place is the over-riding influence of extrinsic forces on species diversity, such as propagule pressure, primary productivity, or resource heterogeneity. This explanation fits with the contention that large-scale correlations in native and non-native species richness patterns come from similar factors promoting diversity for both groups. Kendi Davies and her co-authors (2005) pinpointed the predominant role of spatial heterogeneity in producing the paradox between scales in California grasslands. At small spatial scales, physical variables such as soil depth, aspect, and soil nitrogen content do not vary. Since these are variables that commonly determine which plant species establish locally, there is a corresponding lack of beta, or across-space, species diversity. Under these conditions, individual plants are likely to compete for these resources heavily, and thus there is a negative correlation between native and non-native richness. At even moderately large spatial scales, however, there can be much spatial variability in these physical conditions. Davies et al. (2005) show that, at least in California grasslands, variation in physical conditions can produce higher levels of beta diversity, and higher levels of beta diversity in their study were correlated with a positive relationship between native and non-native richness. The species that coexisted at these larger spatial scales were not competing for a homogeneous set of resources. Instead they were segregating themselves along resource gradients, and the higher the differences between sites, the more native and non-native species could coexist. The work of Davies and her colleagues puts a finer point on our understanding of the role of scale in explaining the paradox. The paradox is not produced by
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an elevated mean level of available nutrients or other such resources, but an elevated variance in these resources. Furthermore, the paradox is not produced by an elevated temporal variance in resources, as suggested by the resource fluctuation hypothesis (Davis et al. 2000), but elevated spatial variance in resources. We have discussed the similar effects of variation in time and variation in space previously in establishment, and we will discuss them further in the chapters on the spread of invasive species (Chapters 7 & 8). Readers with some familiarity with community ecology theory and competitive coexistence theory will recognize the temporal storage effect (Chesson & Warner 1981; Warner & Chesson 1985) in the Resource Fluctuation Hypothesis and the spatial storage effect (Chesson 1985) in Davies et al.’s (2005) emphasis on the role of spatial variation in resource availability. This result remains to be tested within other ecosystems, but it is likely to find support just as both temporal and spatial storage effects have been demonstrated to affect coexistence in ecological communities. One of the interesting predictions from this emphasis on spatial variation is the suggestion that at regional scales (i.e., at relatively large scales) homogeneous conditions should show a negative correlation between native and non-native richness (Stachowicz et al. 2002a).
have not supported Elton’s hypotheses, perhaps because they begin with randomly assembled communities that may not interact stably even before the introduction of another component. Real world communities are the result of a multitude of feedback loops between species, and over time they develop into persistent if not stable interactions. Completely random mathematical communities, therefore, do not test how well a functioning real community might repel a non-native species (Doak et al. 1998). Ted Case (1990) addressed this issue by examining how successfully he could add another participant to “communities” made up of a range of species already interacting stably. His equations represented pairwise negative interactions with indirect interactions spread through shared competitors. He found that a “non-native” species that used a subset of the resources that a resident “native” species utilized was less likely to establish when species richness was high. In more diverse communities, there were more examples of priority effects in which a native species was not the superior competitor to the non-native but simply gained an advantage by being present in the community first. Case (1990) also found an increase in indirect interactions in communities with more members, so that several native species in concert were responsible for the exclusion of a single non-native.
Predation The topic of “predation” covers the gamut of predator-prey, host-parasite, plantherbivore, and other interactions in which one species eats another. Predation in any of these forms can play just as strong a role as competition in native community resistance
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to the establishment of non-natives. Because predators often (with the exception of herbivores and some parasites) have much stronger effects on survival than competitors, the contributions of predation in inhibiting non-native species establishment are more readily observable. A good example is the establishment of the European green crab (Carcinus maenas) in marine habitats all over the world. Around 1900 the green crab became established in the western Atlantic Ocean on the east coast of North America, where it has had large impacts on native mussels, urchins, and barnacles (Leonard et al. 1999). More recently it has established off the coast of South Africa, Tasmania, and western North America (LeRoux et al. 1990; Cohen et al. 1995; Grosholz & Ruiz 1996). The European green crab has failed to establish, though, in some of the more diverse habitats along the Pacific coast of North America, specifically in northern sites occupied by the native red rock crab (Cancer productus), a larger crustacean (Hunt & Behrens Yamada 2003). Hunt and Behrens Yamada (2003) noticed that catch records did not record the European green crab in cooler, more saline areas occupied by the native crab in both California and Oregon, USA. When they caged the invader near the mouth of an estuary in Oregon, the crabs survived, suggesting that abiotic factors (e.g., water temperature or salinity) did not explain the failure to establish. They then collected small, medium, and large European green crabs and red rock crabs and paired them to see how smaller individuals fared in the presence of their own species, and against the other species. Small and medium individuals of both species experienced about 14% mortality from predation when paired with larger European green crabs. For the native crab, the predation rate was about the same or a little bit lower if they were paired with larger conspecifics, but, for the European green crab, the predation rate rose to between 52 and 76% mortality from predation in the presence of a single larger native red rock crab (Fig. 6.3). The native crab, therefore, most likely prohibits the European green crab from establishing in cooler habitats along the Pacific Coast of North America through intense predation. Predators have similar effects in terrestrial systems. In fact, a terrestrial crab, the red land crab (Gecarcoidea natalis), repels the introduced giant African snail (Achatina fulica) from tropical forests in the system on Christmas Island described above by preferentially eating them to local extinction (Lake & O’Dowd 1991). The giant African snail has successfully established non-native populations in a wide assortment of tropical island systems (see Chapter 9), so this example of resistance is potentially critical. The red land crab is an omnivorous species, so it acts not only as a predator, but also as an herbivore. Although herbivores may eat only part of their prey, they can still play a strong role in excluding non-native species from novel environments. In the seedling experiment on Christmas Island, in which researchers placed seedlings of four nonnative trees in tropical forest and along roads and forest edge, the plants were placed in cages to protect them from the red land crab (Green et al. 2004). Although protected seedlings of two non-native species died in intact forest because competition reduced access to light, unprotected seedlings of all four species fared worse in all environments – including along roads – because of herbivory by the red land crab (Green et al. 2004, and see Fig. 6.1). Not only native species can impede the establishment of non-natives. On Isla Victoria, in Argentina, where Simberloff and his colleagues (2002) examined the establishment
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Figure 6.3 Percent predation experienced by European green crabs (Carcinus maenas) when paired with other green crab individuals, and with native red rock crabs (Cancer productus) of various sizes (S = small, M = medium, L = large). All crabs experienced some predation when paired with individuals larger than themselves. Green crabs experienced very high predation, however, when paired with larger red rock crabs. Reprinted from Hunt, C. E. and S. Behrens Yamada. 2003. Biotic resistance experienced by an invasive crustacean in a temperate estuary. Biological Invasions 5: 33–43 with kind permission from Springer Science and Business Media.
of pine seedlings in the vicinity of the former nursery, they found that introduced deer have high impacts on seedlings of the non-native pine species. The deer reduced the presence of the non-native trees at forest edges, and in meadows where deer prefer to browse and where the competitive effects on the non-native trees from native forest species are minimal.
Mutualisms Because multualisms are generally facilitative interactions that assist species, it is difficult to envision how the presence of a native mutualist might repel a colonizing, non-native species. Instead, the lack of a mutualist partner may keep a non-native from entering a system. In fact, if you had to place a bet on the one biotic interaction that would most effectively prohibit the establishment of a non-native species, you might bet on a missing obligate mutualist. By definition an obligate mutualism involves two species that have co-evolved such that the presence of one is required for the persistence of the other, and vice versa. Thus, if a non-native species were introduced without the obligate mutualist from its native range, chances are high that it would not establish a self-sustaining population.
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Pollinating female finds receptive fig Fig plant (Ficus sp.) with immature figs
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Figure 6.4 The figure shows a generalized lifecycle for a fig and fig wasp mutualism. Figure modified from artwork by Simon van Noort (http://www.figweb.org).
An oft-cited example is the obligate mutualism between fig trees and fig wasps. The tree genus Ficus (Moraceae) has about 750 species distributed across the tropics and subtropics. With only a few exceptions, each fig species is pollinated by a wasp in the family Agaonidae (Hymenoptera). Figs reproduce almost exclusively through seeds and therefore are extremely reliant on visits to their flowers by pollen-laden wasps. Fig wasps, in turn, lay their eggs exclusively within the flowers of fig trees. Fig wasps form galls within the floral ovaries of fig flowers where the young wasps will develop, and immediately after hatching, mate with one another (Fig. 6.4). Successful reproduction by either the fig trees or wasps is, therefore, completely dependent on their interaction (Nadel et al. 1992). Given such a tight association between these two species, one may naturally assume that a particular fig tree introduced outside of the range of its mutualist fig wasp would not reproduce successfully and thus would not establish a self-sustaining non-native population. This has been the assumption of horticulturalists and landscape architects in Florida for decades. Over 60 species of fig trees have been imported into Florida. Figs are popular ornamental trees because they propagate easily from cuttings. Florida does sport two native fig trees and two associated species of fig wasps. There is no evidence, however, that these native fig wasps can breed in the 60 or more non-native species of fig trees. Thus it should be no surprise that none of the non-native fig trees in Florida successfully reproduced for decades after their introduction. As you
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can imagine, though, and as we will discuss in the next section, this situation is not necessarily a very persistent one because the barrier to establishment depends on the absence of the non-native fig wasps. The tight association between species observed in obligate mutualisms, like that of the figs and fig wasps, is relatively unusual. Facultative mutualisms, which involve generalist species that interact to their mutual benefit, are much more common. In these interactions the loss of any one partner does not automatically doom one of the others to extinction. Facultative mutualists may require the service of a partner to complete some portion of their lifecycle, but that role may be filled by many different species. Common facultative mutualisms include pollination of flowers, and the increase in nutrient and mineral uptake by plants due to an association with mycorrhizal fungi or bacteria. In both of these examples the other partner also benefits. Pollinators get a nectar or edible pollen reward, and mycorrhizae exchange phosphorus and minerals for carbon sugars from plants. Although it is easy to envision how facultative mutualisms might be critical to establishment, it is difficult to definitively pin the failure of a nonnative species to establish on the lack of a facultative mutualist (Richardson et al. 2000a; Levine et al. 2004). Despite this lack of empirical evidence, ecologists have developed theories about the types of non-native species that are more likely to successfully establish. In general these theories suggest that species that require mutualists are less likely to establish, and instead species that can complete their life cycle independently of the involvement of other species will tend to be successful. Based on this logic, many invasion ecologists have suggested that a limitation of pollination service is a principal facet in identifying the establishment potential of plants (Baker 1965). For an animal-pollinated plant to establish a self-sustaining population it must make do with native resident pollinators or bring its pollinator with it. Plants that are entirely or mostly self-fertilizing and require unspecialized pollination services therefore may be more likely to establish non-native populations (Baker 1965; Marshall & Brown 1981). Although there are many examples of dramatically successful wind-pollinated or selfing non-native plants, including wildly successful establishment of garlic mustard (Alliaria petiolata) and cheatgrass (Bromus tectorum) in North America (Richardson et al. 2000a), there are also wildly successful non-natives that require insect pollinators. Purple loosestrife (Lythrum salicaria) in North America is one excellent example. There is, in fact, very little evidence that pollinator mutualisms limit the establishment of non-native species. Part of this result is due to the difference between establishment and spread that we discussed at the outset of this chapter. In order to stop establishment, it is necessary to keep a species from surviving and reproducing in a particular area at all. Facultative mutualisms, such as pollination services, are much more likely to play a role in how well a species performs once it has established. This pattern is repeated in mycorrhizal or bacterial associations with plants. Richardson and colleagues (2000a) found that non-native plants are rarely inhibited in their survival by a failure to establish an association with soil fungal symbionts. These interactions can be quite plastic (in the evolutionary sense of plasticity, meaning the response can vary with conditions), and non-native plants will often form associations with mycorrhizae and bacteria fairly quickly. Even if they do not form these associations
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immediately upon arrival, it is unlikely that the lack of a soil symbiont alone will restrict their ability to survive in an environment although it may reduce their ability to survive pressure from either a predator or, even more likely, a competitor (see below). An interesting side-note to this story, however, is the observation that in current horticultural practice slurries of microscopic symbionts are often sprayed around nonnative species when they are initially planted (Richardson et al. 2000a). These slurries are very likely made up of species that are not native to the area. Thus, human actions in this regard may be creating a cosmopolitan alliance of plant rooting symbionts that are purposefully introduced wherever non-native plants are utilized for horticultural or ornamental reasons.
Indirect effects As much of art and literature can attest, life perhaps gets more interesting but definitely gets more complicated when you add a third member to any party. The same is true when ecologists consider interactions between more than two species. In such cases, it is possible that the effect of one species on another, either positive or negative, may be mediated through the third species. Here we consider the influence of such indirect effects on resistance to the establishment of non-native species. Note that most examples of species facilitating the establishment of non-natives are indirect, rather than direct, effects. We consider these in the facilitation section of this chapter. Despite the fact that facultative mutualisms alone do not seem to preclude establishment of non-natives, these mutualisms can change the balance of power between species to assist in establishment failure. One example is mycorrhizal associations that increase nutrient capture and competitive ability for native plants but do not “work” for non-native plants. Some fungus–plant associations are quite specific, and it is thus possible that non-native species simply cannot take advantage of the native mycorrhizal community (Smith & Read 1997). This leaves these non-native plants at a severe disadvantage as they have far lower access to nutrients than natives (Richardson et al. 2000a). Simberloff and colleagues (2002) hypothesized that this process might be instrumental in the failure of pine trees to establish in native forests on Isla Victoria in Argentina because the native tree species tended to form associations with specialist mycorrhizae that were unavailable to the non-natives. Competition and predation can also act in concert without mutualisms. You perhaps noticed in the last two examples in the predation section that competition and herbivory both played roles in limiting establishment success of trees on Christmas Island and Isla Victoria. In both cases the non-native species was more vulnerable to herbivory than the native species, and the herbivores helped to skew competition even more in favor of the native species (Simberloff et al. 2002; Green et al. 2004). Often these two biotic interactions combine, with herbivores changing the balance of competition between native and non-native species. For example, Case and Crawley (2000) found that Conyza sumatrensis, a newly arriving non-native plant in Great Britain, does not readily establish in undisturbed grasslands. They examined the effects of competition as well as herbivory from insects, snails, and rabbits. They found that insect herbivores had
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Figure 6.5 Interaction between (a) competition and (b) herbivory by rabbits in the number of established Conyza sumatrensis, a non-native plant in Great Britain. Cult + MB is the low competition treatment. Cult is a treatment with moderate competition, and Grass is a treatment with high competition. −R represents a treatment where rabbits were excluded, and +R represents a treatment where rabbits were allowed free access to the plants. The differently shaded bars represent results for 1996 (dark) and 1997 (light). Competition decreased the number of Conyza present as did rabbit herbivory. The treatment that precluded any Conyza establishing, however, was rabbit herbivory combined with high competition. Reprinted from Case, C. M. and M. J. Crawley. 2000. Effect of interspecific competition and herbivory on recruitment of an invasive alien plant: Conyza sumatrensi. Biological Invasions 2: 103–110, with kind permission from Springer Science and Business Media.
little to no effect and that snail herbivory is relatively unimportant in intact grassland because competition usually precludes establishment. When native competitors were removed, however, snails had a strong negative effect and could preclude C. sumatrensis establishment. The single combination that always excluded the invader, though, was rabbit herbivory in intact native vegetation (Fig. 6.5).
Resistance synthesis Which of these interactions plays the most dominant role in limiting establishment of non-native species? Levine et al. (2004) addressed this issue by collating published
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research on the role of competition, herbivory, species diversity, and soil fungal communities for inhibition of non-native species establishment and their subsequent performance (i.e., survival, growth, and reproduction) once established. They used meta-analytical statistical methods to gauge the importance of each form of resistance (Gurevitch & Hedges 1999). Their results indicated that competition, herbivory, and species richness each had strong inhibitory effects on the performance of established non-native species. Levine et al. (2004) argued, however, that none of these factors alone was capable of prohibiting non-native species from establishing. In general, the conclusions of Levine et al. (2004) match our review above. Their findings are particularly relevant to how students should view this chapter in relation to the information presented in Chapters 7 and 8. These authors are arguing that biotic interactions can be quite influential at the spread stage of invasions, but they are not responsible for single-handedly keeping a newly arriving non-native species from surviving or reproducing. They suggest that the ill-effects of competitors, herbivores, and, by association, species richness are not sufficiently consistent and widespread to drive incipient non-native population sizes to zero. This combination may be necessary for native species to fully prevent establishment by a non-native species. This topic certainly deserves further attention, but with a more careful eye toward the difference between establishment of non-natives and their subsequent geographical spread.
Facilitation of establishment In this section we will explore the flip side of of the interactions from above, and even some of the same examples that we have discussed. Note that in this section each of these interactions facilitates establishment, which, at least for the competition and predation subsections, requires indirect effects. The main role for the facilitator in all of these interactions is to mitigate the effects of a stress, which may or may not be a biotic stress. Competitors and predators that facilitate establishment generally behave similarly to a mutualist in a direct interaction; their facilitative effect is often inadvertent. The motivation for the action is their own benefit. The invading species often sacrifices something to the facilitating species, but this sacrifice is balanced by the reduction in another stress. What we have is an ecological repetition of the old maxim, an enemy’s enemy is a friend.
Competition The main role for competitors in facilitating the establishment of non-native species is to reduce predation (Box 6.2) although there are also many cases, particularly in the plant literature, in which a competitor mitigates abiotic stress for some portion of the lifecycle of a non-native species. Predation reduction generally is not due to the native competitor being preferred by predators over the non-native (which would merely serve to increase predator densities), but is often because the competitor offers actual physical shelter from the herbivore. For plants, herbivory may be lower under adult shrubs because the shrubs repel their grazers with thorns, or by existing at really
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Box 6.2 Escape One of the most widely cited explanations for the successful establishment and spread of non-native species is the escape from natural enemies hypothesis, which suggests that release from predation, parasitism, and herbivory allows non-native species to survive, grow, or reproduce at higher rates than in their native range (Elton 1958; Blossey & Notzold 1995; Maron & Vila 2001). This idea is not new to invasion ecology but was a fundamental part of Darwin’s original thoughts on the origin of species. The new insight comes from specific hypotheses concerning how escaping enemies creates advantages for non-native species. Do invaders really have fewer natural enemies in their non-native range than in their native range? Does this loss of enemies lead to a loss of traits geared toward defense or resistance? In the absence of their enemies do invaders reallocate energy from defenses to growth or reproduction? If so, how does this reallocation affect competition? The first piece of the puzzle is to establish whether, on average, non-native species experience lower pressure from predators, herbivores, and parasites in their nonnative range as compared to in their native range. The evidence in favor of a lasting escape from enemies is equivocal (e.g., Agrawal & Kotanen 2003), but non-native species frequently experience at least a temporary release from predatory pressure before establishment of their enemies from the native range or discovery by hosts and predators native to the introduced range (e.g., Di Tomaso 1998; Maron & Vila 2001; Vila et al. 2005). This release, even though temporary, may be enough to explain why some species establish while others do not. Optimal defense theory comes into play after establishment. This theory suggests that, without enemies, non-native species will reallocate defensive energy toward whatever most increases fitness (Coley et al. 1985; Bazzaz et al. 1987). This reallocation might take the form of increased growth or higher reproduction, depending on the competitive environment. Blossey and Nötzold (1995) formulated these ideas into the evolution of increased competitive ability hypothesis (EICA), which suggests that non-native species evolve to compete better in the absence of their natural enemies. They tested their hypothesis on purple loosestrife (Lythrum salicaria), which is a wetland plant that is native to Europe but invasive in much of North America. In its native range, purple loosestrife is a common wetland plant that suffers herbivory from leaf and root feeding insects. This herbivory controls loosestrife densities and keeps the species from attaining dominance within its native wetland communities. In North America purple loosestrife suffers little to no herbivory from native species and takes over wetland areas, creating monocultures that persist for generations. Blossey and Nötzold (1995) grew plants from seeds collected in Ithaca, New York, USA (non-native range with no herbivores) and from Lucelle, Switzerland (native range with herbivores) in a common garden and looked at biomass and herbivore performance. They found support for the EICA hypothesis in that Ithaca plants attained significantly higher biomass than the Lucelle plants, but root-feeding herbivores also survived better on the Ithaca plants than on Lucelle plants. These results
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Box 6.2
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suggest that the non-native population of purple loosestrife in Ithaca had lost herbivore resistance and had reallocated resistance energy toward faster growth rates. These results are not universal, though. Other researchers have found that non-native plants are not more competitive in their introduced range than in their native range despite a loss of herbivores (Maron & Vila 2001; Vila et al. 2003). In an attempt to explain these varied results, Müller-Schärer et al. (2004) explored the ways that plants defend themselves against herbivores. They suggested that qualitative defenses (i.e., those that occur in low concentrations and are usually effective against generalist enemies) and quantitative defenses (i.e., those that occur at high concentrations, are based on digestibility-reducing compounds, and are usually effective against specialist enemies) affect post-establishment dynamics differently. Non-native plants that have escaped specialist herbivores may lose their quantitative defenses without losing their qualitative defenses. Müller-Schärer and his colleagues (2004) thus hypothesized that the EICA hypothesis will hold only if non-native plants lose their specialist herbivores and experience low generalist herbivory in their introduced range. On the other hand, non-native plants that have escaped their specialist enemies may evolve higher levels of qualitative defense in their introduced range if they experience high levels of generalist herbivory there. In this case, any “saved” resources will not be re-allocated to growth or reproduction but instead will go toward the production of toxins, such as alkaloids and glucosinolates. Leger and Forister (2005) have found some support for this hypothesis in studies of the California poppy (Eschscholzia californica) in its native range in California, USA and its introduced range in Chile. They found that generalist herbivores grew and survived better on plants from California (native range) than on plants grown from seeds from Chile (introduced range), suggesting that the non-native populations had evolved more defensive compounds. They did not find a complete trade-off between competitive ability and defensive traits, though. Plants from the introduced range also grew larger than plants from the native range (Leger & Rice 2003), suggesting that plants may be able to allocate additional resources from specialist defenses to both growth and increased generalist defense.
low densities. For example, Facelli and Temby (2002) examined the factors affecting the establishment of Carrichtera annua, a non-native annual plant that has become the dominant species in an Acacia woodland in South Australia. The understory is dominated by native chenopod shrubs, which are well-adapted to the arid climate of the region and have many structural defenses against herbivory. They found that herbivory on the non-native C. annua is lower beneath these native shrubs than in the more open understory areas, allowing the non-native to create source populations beneath them. The non-native also germinates better beneath native shrubs because the deep roots of the shrubs bring water and nutrients to the soil surface through
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hydraulic lift, creating a less arid, more hospitable microclimate for the non-native to germinate and grow. The favor is not returned, however. The non-native annual plant then removes nutrients and water from the soil around the native shrub and competes with it. Similar effects are found in grassland systems in the Central Valley of California where long-lived, native perennial grasses alter the soil and facilitate germination and survival of non-native Mediterranean annual grasses (Hoopes 1999; Hoopes & Hall 2002). California’s grasslands are dominated by winter annual non-native species from the Mediterranean region of Europe, but there are tiny isolated remnants of the native perennial grassland with low levels of non-native establishment. In the southern portion of the Central Valley, Sporobolus airoides, a C4 native perennial bunchgrass, exists in a matrix of vernal pools and uplands dominated by annual grasses, and saline sears dominated by S. airoides. Saline sears are areas with extremely high salt concentrations where natural and human-derived salts accumulate from winter pooling and drying. Non-native annual grasses can only establish in the salt sears on native bunchgrass litter or directly beneath bunchgrass canopy. The native bunchgrass S. airoides is weakly halophytic in that it draws salts up into its tissues, exudes them in its vegetative and reproductive tissues, and thereby reduces the salinity in the soil in its immediate vicinity. This allows the salt-intolerant non-native annuals to establish themselves. Just as in the case with the native Australian shrub that facilitates the establishment of non-native annuals, the long-term effect on the native S. airoides is not good. Non-native annual grasses that establish on or around the native bunchgrass reduce vegetative growth, seed set, and seedling recruitment of the native (Hoopes 1999). Zebra mussels (Dreissena polymorpha) have a similar effect on many aquatic ecosystems (Simberloff & Von Holle 1999). Although zebra mussels have had strong negative effects on native Great Lake communities (Ricciardi et al. 1998), they also filter more plankton than native filter feeders and reproduce extremely rapidly (Leppäkoski 1984). These characteristics increase organic material and settling substrate for other benthic macroinvertebrates and lead to increases – both native and non-native – in macroinvertebrate densities, thus, facilitating the establishment and recruitment of non-native competitors (Ricciardi et al. 1997). One of the non-native species that has benefited from the presence of zebra mussels is the Eurasian faucet snail, Bithynia tentaculata (Ricciardi et al. 1997). Zebra mussels have an even wider circle of effects, though, because they reduce plankton and increase water clarity, which alters habitat quality and facilitates the growth and establishment of several non-native aquatic plant species, including Eurasian watermilfoil (Myriophyllum spicatum) (MacIsaac 1996). In all of the above examples, the competitor changes the environment in a way that allows a non-native species to establish in an otherwise inhospitable habitat. Many of these changes are to the fundamental abiotic conditions that limit species establishment, and at times the facilitators are themselves non-native species. Notice that the subsequent negative effects on the facilitator may lead to a negative feedback and impose limits to the facilitation, but these limits are unimportant in the scheme of establishment. If the facilitator allows a non-native to establish despite subsequent competition, the negative feedback cannot bring the system to the original community state but instead brings it to a new state that incorporates the invader.
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Predation The most common way for predators to facilitate establishment is for a non-native predator to remove a species that has a strong negative effect on the establishment of a non-native species. This sort of dynamic is common in grazed systems around the world, particularly those with native bunchgrasses (Kimball & Schiffman 2003). European landscapes have been managed with domesticated grazers for millennia, leading to structural and physiological adaptations in the associated grassland plant species (Mack & Thompson 1982). Outside of Europe, many grasslands have not coevolved with congregating grazers, and thus, when grazers are introduced, non-native plants from Europe tend to flourish (Mack & Thompson 1982; Wagner 1989). For example, cattle grazing within grasslands in Australia, South America, and North America bring seeds of grazing-adapted European species on their hides and hooves (Mack 1981, 1986). The cattle then create germination sites for these non-native grasses by opening “holes” in the vegetation with their hooves, and they also disturb native symbiotic relationships by destroying soil crusts (Mack 1986, and see Chapter 4 and Box 7.2). Asian water buffaloes have had the same effect on flood plains in Australia (Braithwaite et al. 1989); introduced sheep have had similar effects on grasslands in Tasmania (Leonard & Kirkpatrick 2004); and a variety of herbivores increase the establishment of purple loosestrife in wetlands in North America (Rachich & Reader 1999). In general a non-native herbivore or predator that removes native vegetation has a double-whammy effect in facilitating additional invaders: they create a disturbance or bare site that allows a non-native species to reach exposed soil and available nutrients, and they reduce the ability of native plants to compete with the non-native by imparting physiological damage from grazing. Such facilitation effects are also possible at higher trophic levels. A well-documented example is the facilitation of non-native frogs in Oregon, USA, by non-native fishes, which is mediated through the predatory effects of native dragonflies. Bullfrogs (Rana catesbeiana) are native to eastern North America, but have been invading western North America, Mexico, Hawaii, Japan, and several other areas through aquaculture vectors. Bullfrogs are voracious predators that often eat other amphibians (Moyle 1973). Bullfrogs in their native range have an indirect positive interaction with centrarchid sunfish (Werner & McPeek 1994). The sunfish eat the bullfrog’s main predators: dragonfly nymphs and other predaceous insects. Dragonfly nymphs are particularly fond of eating bullfrog tadpoles. Several species of eastern North American sunfish have also been introduced into western North America, and beyond. Adams et al. (2003) wanted to know if this indirect facilitative interaction within Oregon (where the sunfish and bullfrogs are non-native) is fully functional and contributing to the success of non-native bullfrogs there. Through a series of field observations and experimental enclosures, Adams et al. (2003) showed that bluegill sunfish increase the survival of bullfrog tadpoles by reducing the population size of the bullfrog’s predators, dragonfly nymphs. Survival of bullfrog tadpoles decreased from 57% in enclosures with no dragonflies to 10% with them. Dragonfly survival decreased from 85% in enclosures with no bluegill sunfish to 13% with them. Within enclosures that had both the dragonflies and the sunfish, bullfrog
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Figure 6.6 Percentage survival of bullfrogs (Rana catesbeiana) and dragonflies (Aeshnid sp.) within experimental ponds that did (left) and did not (right) include bluegill sunfish (Lepomis macrochirus). When dragonflies were excluded (open bars) most bullfrogs survived, however when dragonflies were present (dark bars) bullfrogs faired much worse. When bluegill sunfish are added, they prey on the dragonflies and thus allow bullfrogs to survive at percentages equal to when dragonflies were excluded. [Reprinted from Adams, M. J., C. A. Pearl and R. B. Bury. 2003. Indirect facilitation of an anuran invasion by non-native fishes. Ecology Letters 6: 343–351, with permission from Blackwell Publishing.]
survival was as high as within enclosures that had neither of the other two species (Fig. 6.6). These results suggest that within the bullfrog’s non-native range, in any ponds in which it co-exists with a species of sunfish, it will experience high survivorship and thus reach high abundance. In ponds that do not have sunfish, the bullfrogs may be prohibited from establishing due to the predation pressure from the native dragonfly nymphs. Finally, we explore facilitation of establishment through predation by one of the most devastating invading groups globally: non-native ants (Holway et al. 2002a). These social insects can become extremely abundant, prey on a diverse array of native species including plants (both herbaceous material and seeds), arthropods, and occasionally even vertebrates. They also act as seed dispersers and frequently farm or tend
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other species in order to create a consistent supply of resources. These actions substantially alter native community composition and function and can lead to a variety of facilitative effects on other non-native species. In the Christmas Island system that we examined earlier, the invading yellow crazy ant (Anoplolepis gracilipes) attacks the native herbivorous red land crab (Gecarcoidea natalis) (O’Dowd et al. 2003). Because this native crustacean is instrumental in excluding non-native plants (Green et al. 2004) and snails (Lake & O’Dowd 1991) from the forest ecosystem on Christmas Island, the yellow crazy ant facilitates both of these non-natives by removing their principle predator. Additionally, the ant is associated with a honeydew producing scale that it tends within the forest habitat (O’Dowd et al. 2003). The increase in honey dew that follows ant establishment also increases a sooty fungus, leading to canopy dieback, penetration of light to the forest floor, and the removal of competitive effects that previously served to reduced non-native tree establishment (O’Dowd et al. 2003). Through multiple avenues, therefore, this ant is assisting an onslaught of non-native species invasion on Christmas Island. As we will see in later chapters, predators can have large impacts on native species in the communities they invade. That their indirect potential facilitative effects can also be large should not be surprising. Rather than the slow, incremental changes that competitors may impose, predators can remove entire populations in a single generation or even a single season. Generalist predators in particular can have widespread, far-reaching effects. When a prey species is a keystone species, such as the land crab on Christmas Island, the system may undergo rapid, catastrophic shifts in species composition from mostly native species to mostly non-native species.
Mutualism Finally, we examine mutualisms, the interaction where we most expect to see facilitative effects on non-native species establishment. If we begin by examining obligate mutualisms, we would expect the presence of an obligate mutualistic partner to facilitate establishment. However, with modern travel and the constant flow of non-native hitchhikers in nursery, agricultural, pet, and tourist traffic, it is not surprising that mutualistic partners might arrive at a site separately yet form a successful association. Establishment is most likely when at least one partner is long-lived and the mutualism acts on a single, short stage in the life cycle. In this way, one mutualist can survive, without reproducing for a long time, and wait for its partner to arrive. Exactly this scenario has allowed several (but not all) non-native fig trees in Florida to begin to reproduce. Their successful reproduction in turn has allowed further generations of pollinating fig wasps to be produced. The “happy” reunion between fig trees and fig wasps in Florida is a consequence of accidental introductions of wasps in fig fruits and nursery imports, often decades after the first imports of the fig trees. Fig wasps are extremely small and cryptic, making them difficult to detect and remove from imported fig products (Fig. 6.4). As we pointed out before, it is extremely difficult to demonstrate that the lack of a facultative mutualism will exclude a non-native species from establishing a self-sustaining
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population, but it is much simpler to see how a successful mutualism facilitates establishment. A review by Richardson et al. (2000a) indicates that non-native species often form facultative mutualisms with resident native species in their new range. Indeed, several mutualism examples involve only non-native species, each from a different place. These species are not reassembling the jigsaw puzzle from their native range. Instead, they are inventing their own community configurations and relationships. A fascinating example comes from one of the many ant invaders, an African ant (Pheidole megacephala) that tends a Brazillian green scale (Coccus viridis) that lives on an Indian fleabane (Pluchea indica) in Hawaii (Bach 1991). As mentioned before, ants are known to “tend” (or farm) homopteran insects so that they may feed on secreted honeydew. The scale insect benefits from this interaction through reduced predation and competition when the ants eat other, potentially predatory, insects. Removing the honeydew benefits the plant by reducing its susceptibility to a variety of sooty molds that can build up in the presence of the sweet honeydew. (Note that in systems without such sooty molds prior to the invasion – such as the Christmas Island example above – this benefit is somewhat of a mixed blessing.) Such interactions have fascinated ecologists for nearly a century because they seemed so unusual and dependent on such complex coevolution. Nevertheless, we presented two examples of intricate facilitative webs assembled de novo in a recipient community (Hawaii and Christmas Island) from species with no prior evolutionary history with one another. How common are such facilitative interactions between non-native species? Simberloff and Von Holle (1999) reviewed 254 published articles dealing with non-native species invasions. They found that the vast majority of studies described predator– prey or phytophagus insect–plant interactions between non-natives (156). Looking at establishment, it appeared that facilitation was important almost twice as often as competition – despite the predominance of competition in ecological studies (see above). Their review focused on interactions between non-native species, but there are also many instances of native species facilitating the establishment of a non-native species. Finally, as Richardson et al. (2000a) point out, by definition all non-native species have benefited from their association with humans in the sense that humans aid in their dispersal and very often provide for their basic needs once released (e.g., water, food, cover). We have ignored facilitation between humans and non-native species because, to a certain extent, Chapter 5 covers the indirect facilitation of humans on non-native species establishment by discussing the role of human disturbance in establishment.
Summary Despite the expectation that native species will often put up significant resistance to the establishment of non-native species, especially if there are many natives, the evidence for resistance is somewhat equivocal. There are some compelling field examples showing that predators, herbivores, and competitors have adverse effects on non-native species,
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but rarely are these effects strong and consistent enough to single-handedly prevent a non-native species from establishing. There is also no clear pattern showing a negative relationship between species richness and non-native species establishment although there is the suggestion that this pattern operates at small spatial scales, particularly if all other factors (propagule pressure, resource availability, disturbance regime, etc.) are equal (Hooper et al. 2005). Similarly, the failure to create mutualistic interactions is clearly not stopping the establishment of non-native species. So, what does stop a non-native species from establishing? Perhaps it takes two or more negative blows to the non-native for it not to establish. This could come from any combination of competition, predation, herbivory, parasitism, or lack of mutualism. It could also include an interaction between one of these biotic associations and a less-than-welcoming suite of abiotic and climatic variables. Despite the fact that the number of non-native species is increasing rapidly (Chapter 3), ecologists have had a difficult time constructing experiments to uncover what exactly produces resistance to establishment. It is very difficult to observe a failed attempt at establishment. On the other side of the ecological coin, there are some fascinating examples of how direct positive, and indirect negative, interactions have facilitated the establishment of non-natives. Maybe ecologists have more compelling evidence for the influence of these interactions on establishment because they are easier to observe. It is worth noting that recent biodiversity surveys have shown that many locations have increased in species richness dramatically in the face of accelerating invasion rates (Sax et al. 2002). This result may be because we have not had time to see the loss of native or non-native species after an invasion, but it could also suggest that there are resources to accommodate nonnatives. Perhaps this stands to reason if non-natives facilitate new ways of exploiting the environment. As Davis et al. (2000) suggest, all biological systems can be invaded; some just may incorporate non-natives more quickly than others. The real ecological action in terms of biotic resistance may materialize only at the spread stage (Levine et al. 2004).
Further complexities Davis, M. A., J. P. Grime, and K. Thompson. 2000. Fluctuating resources in plant communities: a general theory of invasibility. Journal of Ecology 88: 528–534. Hooper, D. U., F. S. Chapin, J. J. Ewel, A. Hector, P. Inchausti, S. Lavorel, J. H. Lawton, D. M. Lodge, M. Loreau, S. Naeem, B. Schmid, H. Setala, A. J. Symstad, J. Vandermeer, and D. A. Wardle. 2005. Effects of biodiversity on ecosystem functioning: a consensus of current knowledge. Ecological Monographs 75: 3–35. Levine, J. M., and C. M. D’Antonio. 1999. Elton revisited: a review of evidence linking diversity and invasibility. Oikos 87: 15–26. Rejmanek, M. 2000. Invasive plants: some generalisations. Australian Journal of Ecology 25: 497–506. Richardson, D. M., N. J. Allsopp, C. M. D’Antonio, S. J. Milton, and M. Rejmanek. 2000. Plant invasions – the role of mutualisms. Biological Reviews 75: 65–93.
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Companion papers Adams, M. J., C. A. Pearl, and R. B. Bury. 2003. Indirect facilitation of an anuran invasion by non-native fishes. Ecology Letters 6: 343–351. Levine, J. M., P. B. Adler, and S. G. Yelenik. 2004. A meta-analysis of biotic resistance to exotic plant invasions. Ecology Letters 7: 975–989. O’Dowd, D. J., P. T. Green, and P. S. Lake. 2003. Invasional “meltdown” on an oceanic island. Ecology Letters 6: 812–817. Simberloff, D., and B. Von Holle. 1999. Positive interactions of nonindigenous species: invasional meltdown? Biological Invasions 1: 21–32. Stohlgren, T. J., D. T. Barnett, and J. Kartesz. 2003. The rich get richer: patterns of plant invasions in the United States. Frontiers in Ecology and the Environment 1: 11–14.
7 Modeling the Geographical Spread of Invasive Species Overview Once a non-native species has established a self-sustaining population, it may begin expanding its range beyond the point of initial introduction. Notice that a self-sustaining population automatically takes us into the realm of positive population growth. This chapter and the next one focus on this process of geographical spread, or range expansion. Because non-native populations become very noticeable when they cause impacts or attain a large geographical range, this is a stage that has captured many ecologists’ attention. Elton (1958) devotes nearly half of his book to describing the range expansion and impacts of non-native species on continents and islands. As you will read below, some of the earliest studies of invasive species concentrated on modeling the spread of non-native species. Because the ecology of non-native species’ geographical spread is so rich, we have divided the topic into two chapters. In this first chapter, we will review mathematical models of spread and their application to the invasion process. In the next chapter, we will explore the ecology of range expansions and discuss the factors that affect spread and the patterns that emerge. Because non-native species inflict the most ecological or economic damage when they are widespread (i.e., have expanded their range to some extent), we save a detailed discussion of the impacts of these species until Chapters 9 and 10.
What exactly is geographical spread? Geographical spread is one of the most conspicuous stages of the invasion process because suddenly these species appear in new locations and interact more with native species. Although transport moves species outside their native ranges, spread allows non-native populations to expand and even to establish new populations within the non-native range (Fig. 7.1). One of the simplest ways to think about spread is to envision the original introduction point as a focal point surrounded by concentric circles, with each successive circle representing the amount of space across which the non-native
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Spread Transport Establishment Spread Spread
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Figure 7.1 Spread incorporates elements from the transport and establishment stages. Dark circles represent high density populations and lighter areas represent lower densities. The non-native can spread out from an initial establishment site in a diffusive way (solid lines from the established population) that leads to lower initial densities at the leading edge. The non-native can also spread to new distant sites without colonizing the intervening area (dashed lines). Such jump dispersal might occur because of dispersal vectors such as wind, water, or organisms that eat or carry the non-native. It might also occur via human vectors operating within the non-native range.
population spreads during a finite period of time (Fig. 7.2). The circles continue to expand outward through time, slowly but surely encompassing all suitable habitats, just as “The Blob” would slime its way outward to engulf screaming victims in the classic 1950s horror movie. Surprisingly, this simplistic vision of the range expansion process is often not far from what we really observe in nature. A classic example of such blob-like range dynamics is the spread of muskrats (Ondatra zibethica) in central Europe from 1909 to 1927 (Fig. 7.3 & Box 7.1). Figure 7.3 has been reproduced in nearly every book written on the dynamics of invasive species’ ranges and was one of seven examples used by Elton (1958) to presage his arguments on why society should care about invasive species. Elton drew his original figure from data first published by Ulbrich (1930). J.G. Skellam (1951) also used the muskrat example in his classic 1951 work on the mathematical form of species’ range expansion. As you can see from the figure, the idea of an amorphous blob (i.e., muskrat invasion) that expands and engulfs a large portion of central Europe is not far off the mark.
Why do we want to model geographical spread? Before we launch into the mathematics of modeling invasive species’ range expansion, it is worth reviewing why we would want to do such a thing. For most students interested in biological invasions, the math associated with modeling geographic spread is daunting to say the least (ourselves included!). Thus, it is helpful to jump ahead and see what it is that ecologists gain from formalizing processes such as geographic range expansion into mathematical expressions or computer algorithms.
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Figure 7.2 Conceptual model of the geographical spread of non-native species. The original location of release is represented by the center star. The concentric lines represent the limits of the species’ range through time (t). This map of range expansion includes range limits for three time steps (t = 1, 2, 3).
In many ways, models serve as a test of our knowledge of any system of interest. Like the process of formalizing an ecological supposition into a falsifiable hypothesis, the process of model building requires linking observations into a logical expression of how an ecological pattern is produced. As anyone who has built a model can attest, the more realism you try to include in a model, the more unwieldy it becomes. Thus, in the context of understanding invasive species’ range expansion, a model serves as a kind of filter whereby the modeler chooses which biological processes are of paramount importance in determining how fast a population will occupy a given area. This in itself is a useful exercise, but it is just the beginning. The primary utility of a model lies in its ability to focus our attention on the components of the model that make it troublesome, and as a general rule of thumb, all models are troublesome to some degree. Particularly troublesome models are those that do not do a good job of forecasting future “states,” which in this case would mean they do not succeed in describing the speed at which the non-native species expands its range. The troublesome components of a model typically point toward a key knowledge deficit that requires attention to fully understand the system in which we are interested. Alternatively these troublesome pieces may be too simplistic to represent real world dynamics, in which case we need to capture the essence of these dynamics in future incarnations of the model. Finally, troublesome pieces may be those for which we have a hard time collecting enough field data to fully parameterize our model (i.e., provide real numbers for a variable). In these cases, we may attempt to build a new model that relies more heavily on data that are easier to collect from field observations. In some cases, we may build models that are so troublesome that they should be discarded (see below), but, when we hit upon a simple, easily parameterized model that does a good job of forecasting future system states, we feel comfortable using it to help
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Figure 7.3 Map indicating spread of muskrats from the initial locus of introduction (the city of Prague) over the years 1907–1963.
us visualize the outcome of various management options. In the context of invasions, models that forecast the rate of geographical spread allow us to meet a number of goals: (i) anticipate which non-native species will quickly spread over considerable areas and which will not, (ii) pinpoint the life history or ecological conditions that are most important in allowing rapid range expansion, (iii) prioritize control efforts, and (iv) target management efforts towards halting or slowing the rate of spread in an efficient and effective manner. With these goals in mind, we will now delve into the details involved in modeling the spread of invasive species.
The reaction–diffusion model In its simplest form, the expansion of a non-native species’ range can be viewed as the coupling of dispersal and reproduction. Individuals will reproduce, and then some subset of their young will disperse away from the parent(s). If the organism is mobile, the group of dispersing individuals may also include a few adults. If the organism is sessile (attached to its substrate) in its adult form, then there is usually a life stage
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Box 7.1 Muskrat love The muskrat is a large, stout, semi-aquatic rodent native to North America. Muskrats live in or near water for nearly all of their lives. They dig dens or lodges along the banks of swamps, slow-moving rivers, and wetlands. Their fur is thick and water-resistant, which has made them attractive as fur-bearing game animals because their fur can be made into a nice weatherproof coat. For this reason, the muskrat was brought to Europe and bred extensively in “fur farms” starting around the turn of the nineteenth century. Some muskrats inevitably escaped the farm confines, and others were purposefully released by hunters or fur-enthusiasts (Hengeveld 1989). The muskrat inhabited the entire European continent within 50 years of its initial introduction (it has been successfully eradicated in Britain and Scotland). One focus of the muskrat invasion of Europe began when five muskrats escaped a farm located near Prague within the Czech Republic in 1909. Muskrats have a high reproductive rate with up to four litters per year and four to six young per litter. They have a polygamous mating system where one male may mate with many different females. In addition, muskrats are omnivorous with a diet that includes aquatic vegetation and many different kinds of small aquatic animals. Given these life history characteristics, perhaps it should not be surprising that this small population of five escapees quickly grew to several hundred individuals. Because muskrats destroy roads and dikes through their den-building behavior, they quickly came to be considered invasive pests in Europe. Eradication efforts began relatively quickly after muskrats began to expand their range in the Czech Republic (Ulbrich 1930). Many of the agencies responsible for eradication kept track of the number of muskrats captured per province (Ulbrich 1930). These quantitative data contributed to the construction of maps such as that depicted in Figure 7.3 and allowed Skellam (1951) to mathematically describe the rate at which the muskrat spread out of Prague. Skellam (1951) calculated the area that the muskrat inhabited (its range limit) each year based on the map depicted in Figure 7.3. He then took the square root of these area calculations and plotted them against their corresponding year to produce Figure 7.4. The most remarkable thing about Figure 7.4 is that the data describe √ a near perfect straight line. This suggests that the square root of the range area ( area) corresponds to the radii of the expanding circles that describe the annual geographic limits of the muskrat. Because the relationship described in Figure 7.4 is linear, that is, it increases at a constant rate through time, Skellam also suggested that the speed of range expansion for the muskrat, and perhaps other invasive species, is constant through time. We explore the validity of this notion throughout the rest of Chapter 7. A close look at Figure 7.3 indicates that the expanding concentric circles Skellam produced to make Figure 7.4 are an oversimplification. Instead, the muskrat population expanded quickly along water courses and in the wet season, as we would expect given its reliance on aquatic habitats (Hengeveld 1989). Note that
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Continued
Figure 7.3 is slightly elongated to the northwest and southeast where there is more water. Andow et al. (1990) estimated that the rate of population expansion was approximately 4 times as fast to the southeast as in other directions. Clearly the concentric circles of Figure 7.3 are something of an idealized version of diffusive spread for a real organism. Nonetheless, we will see that the mathematical model Skellam (1951) derived to describe the spread of the muskrat is quite good at predicting average rates of range expansion under a variety of conditions despite not perfectly predicting the exact dimensions of the expanding occupied habitat. For this reason, Skellam’s work has become foundational for the study of ecological range expansion.
Area
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Figure 7.4 Geographical range expansion for the muskrat, starting from the original site of release in Prague, Czech Republic in 1909. The square root of the area of the muskrat’s range increases at a constant rate through time. Adapted from Skellam, J. G. 1951. Random dispersal in theoretical populations. Biometrika 38:196–218 with permission from Oxford University Press.
that moves passively. For example, seeds move by wind, water, or animal vectors, and many marine invertebrates that are buried in the sand or attached to rocks as adults have current-borne larvae that can move long-distances. For mobile organisms, we generally only see individuals dispersing if the population is growing and resources are limited. This should bring to mind two concepts from your basic ecology class: intrinsic rate of population increase and density dependence. A growing population spreads in order to acquire new resources, thereby avoiding density dependence. The other necessary component is some idea of how far and fast individuals disperse away from their location of origin. We can summarize this component as where they settle relative to their population of origin and the speed with which they get there.
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The most general form of the reaction diffusion equation examines the change in density at any point 2 ∂N ∂ 2N ∂ N + = f (N )N + D (7.1) ∂t ∂x 2 ∂y 2 where x and y indicate a point (x, y) in space, N is the density of the population at that point, t is time, f (N ) is a function describing the per capita growth rate of N , and D is the diffusion coefficient. Because density depends on how far we are from the original release point, we must always examine density at a particular point, N (x, y). In order to reduce the clutter in Equation 7.1 we have dropped the (x, y) from the N term. The left hand side of the equation describes the change in the population density at the point (x, y) during an infinitesimally small period of time, or a rate of change in population density. The right hand side of the equation can be divided into two sections, one that describes population growth (per capita growth times density) and a second that describes movement. The first term is the reaction to the population size, and the second is diffusion. Hereafter we will refer to Equation 7.1 as the reaction–diffusion model. The population growth part of the reaction–diffusion model is fairly simple to grasp as long as you have had a basic ecology class. The only difference is that we have used f (N ) to represent the function for per capita population growth. For geometric growth, f (N ) would be a constant (equivalent to r), and, for logistic growth, f (N ) would be r(1 − N /K ). Notice that the growth rate is dependent on the density of individuals, N , in the population at each time step. Without density dependence, the more individuals there are to reproduce at a particular location, the faster the population will grow. This aspect of Equation 7.1 mirrors the simple expressions of geometric and exponential population growth familiar to most ecology students. For most of us the new aspect of the reaction–diffusion model is the diffusion term, D(∂ 2 N /∂x 2 + ∂ 2 N /∂y 2 ). Here D is the diffusion coefficient, a measure of the mean squared displacement per unit time, or how far individuals disperse in a given period of time. The rest of Equation 7.1, (∂ 2 N /∂x 2 + ∂ 2 N /∂y 2 ) is the Diffusion Expression (expressed in two-dimensional x, y space), which describes how movement affects density at a particular distance from the original release point. The diffusion expression has been used in a variety of other contexts to describe such things as the Brownian motion of molecules, turbulent flow of fluids during mixing, and “random walks” (Okubo 1980). This expression is also known as the “dispersal kernel,” or the portion of the equation that describes the distribution of dispersal distances. Diffusion assumes that individuals disperse away from their place of origin in a random direction and settle at a random distance from the origin and that all this randomness can be described by a normal distribution. This means that the solution to Equation 7.1, or the description of density at any point is, N0 −((x 2 +y 2 )/4Dt ) (7.2) e 4π Dt where N0 is the initial release density at the origin. Although that equation may look messy and scary, ecologists actually use it all the time. It describes a two-dimensional N =
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normal distribution. In the context of non-native species range expansion, this equation says that a species’ geographic range is expanding in perfect concentric circles emanating from the original site of release. Because the underlying probability of dispersal is twodimensional normal (meaning that it looks like a bell sitting on the landscape with the height of the bell at any point along its curve describing the proportion of individuals that disperse to that location), we can calculate the equivalent of confidence intervals for individuals in geographic space. From the dispersal kernel in Equation 7.1 within a finite equal √ time frame (t ), 68% of all individuals will be within a circle with a radius√ to 4Dt , and 95% of all individuals can be found within a circle with radius 2 4Dt (Shigesada & Kawasaki 1997). Thus, the diffusion expression specifies that population density falls off as you move away from the origin. However, an important property of diffusion is that it is continuous and infinite. This means that it distributes individuals in a continuous fashion (which may lead to fractions of individuals) and that a small fraction of the dispersers settle infinitely far from the origin in any period of time. The probability of such super-long-distance dispersal is very low, but it is not zero, which differs fundamentally from the real world probability of infinite (or extremely far) dispersal. We will revisit this piece of the diffusion model below as it turns out to be a critical assumption. The one remaining element to the diffusion model that we should explore is the diffusion coefficient, D. This parameter and the intrinsic rate of increase, f (N ), are the two terms in the diffusion model that must be estimated using field observations. The intrinsic rate of increase, f (N ), can be calculated in a straightforward manner using demographic or population census information (Gotelli 2004 and other general ecology texts offer thorough explanations). Typically, the dispersal coefficient (D) is determined from mark-recapture data, the collection of which is described in detail in Sutherland (1996) and Begon (1979), among others. The basic methodology is to (i) capture individuals, (ii) mark them using a method that is relatively long-lasting and unique to each individual or to each initial capture site, (iii) release them at the initial capture site, and (iv) recapture individuals at specific distances and times from the initial capture. Given this information, D can be estimated using the equation, 2 M n=1 x/M (7.3) D= πt where M is the total number of marked individuals recaptured, x is the distance from the initial capture site, and t is the time since initial capture. Kareiva (1983) noted that D varies between species, and that individuals of the same species can demonstrate various values of D depending on the habitat they occupy. This seems an especially relevant point to consider when forecasting the rate of range expansion for non-native species that are released at multiple locations. Unfortunately it may be necessary to estimate D for each of these particular geographic locations, a potentially time- and labor-intensive approach. Or it may be that D reflects innate life history constraints that allow one to estimate it with some confidence using information from previous invasions of the same, or closely related, species. Because multiple modes of dispersal frequently characterize the spread of a species across a landscape, it is unlikely that a
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single estimate of D for a species will allow us to describe its spread across all terrains (Andow et al. 1990; van den Bosch et al. 1992; Holmes 1993). We will discuss some of the methods to deal with variation in dispersal distances below. Intuitively we understand that the value of D will determine, at least in part, the rate at which the non-native population will expand its range. If the value of D is large, the non-native species will expand its range quickly, and, if it is small, the range will expand slowly. The other parameter to consider when evaluating the speed of an invasion is f (N ), the intrinsic rate of population increase. When values of f (N ) are high, there are many more individuals available to disperse, and, when values are low, there are few individuals available to disperse. We can define a new term, c, that combines D and f (N ) and describes the ultimate speed of an invasion. If we assume no intraspecific competition (i.e., f (N ) = √ r), then we can estimate c from Equation 7.1 (Shigesada & Kawasaki 1997), as c = 2 rD. This is the “asymptotic” speed of invasion, the speed that it approaches but may never ultimately reach. There is often a lag phase to most range expansions, and c describes the rate of range expansion after this lag phase is no longer apparent. We will delve into lag phases in non-native species’ range expansions in the next chapter, as it can be explained by several independent biological mechanisms. Even if we assume intraspecific competition (i.e., density dependence), c can still accurately describe the rate of expansion at the edge of the range. This is because there are relatively few individuals at the range front (edge); thus intraspecific competition is low, and the group of individuals at the edge are experiencing exponential population growth. This property is characteristic of reaction–diffusion models, and many other mathematical expressions of range expansion, and leads us to our final insight about range expansions before we evaluate how good this model actually is at describing reality. We can describe the density of individuals (N ) at any point in space (x, y) at time (t ) as N (x, y, t ). If we express the distance between the origin site and each point by a single parameter, d, then we could substitute N (d, t ) for the density of organisms at a given distance from the origin. If we had a series of locations (d values) in which we knew the number of individuals (N ) at a particular time (t ), we could graph the distribution of individuals across space. This distribution is called the probability density function. We used√a probability density function earlier to describe 68% of the individuals being 4Dt of the origin at a particular time or 95% of the individuals being within within √ 2 4Dt of the origin. The fact that we could describe these distributions so cleanly suggests that there must be some constancy to them, and there is. If we plotted a probability density function using the reaction–diffusion model and assumed that the time since initial release was short such that there was no intraspecific competition, the graph would look like Figure 7.5(a) and show a sigmoid relationship between N (d, t ) and d with most individuals close to the origin and a few further away. This graph simply restates what we already know about range expansion dynamics under the reaction–diffusion model; density is high at the point of initial release and drops off as we move away from the origin (in the graph, the point of initial release is conveniently the origin). If we assume that at the range edge there are no effects of intraspecific competition, the distribution of individuals around the edge should look very much like that in Figure 7.5(a) through time. In fact, when we let the diffusion model “run”
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for many time steps into the future, and then we plot the probability density function at the range front at each time step, we see this sigmoid relationship repeated through time (Fig. 7.5(b)). Also note that the distance between the sigmoid lines is constant in Figure 7.5(b), indicating that the front of the invasion is advancing at a constant speed, which (also conveniently) is equal to c (i.e., the speed of invasion). This pattern, of a solution that advances at a constant speed and maintains a constant shape, is called a “traveling wave front.” It is characteristic of the reaction–diffusion model and similar equations (Kot et al. 1996).
Muskrat Redux Now we get to put the reaction–diffusion model to the real test and see if it does a good job of describing the rate of muskrat range expansion and similar real-life examples of non-native species’ range expansions. Skellam (1951) took the map of muskrat spread (Fig. 7.3) and calculated the total area occupied by the muskrat. When he plotted the square root of the area against the year of that range, he found a linear relationship, which he suggested was an indication of a constant rate of range expansion (Box 7.1).
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He went on to mathematically demonstrate that Equation 7.1 could reproduce this relationship, thus suggesting that the reaction–diffusion model captured enough biological reality to be of some use. However, Skellam (1951) did not estimate D or f (N ) and thus we cannot derive c, using Equation 7.1. To really test the utility of the reaction– diffusion model in forecasting the rate √ of geographic range expansions, we want to compare values of c derived from c = 2 rD to the value derived from real field data like that which produced Figure 7.3. If the model and observed estimates agree, we can be more confident of our ability to use the reaction–diffusion model to forecast future rates of range expansion. Using field-derived information on survivorship and fecundity of the muskrat in Europe, Williamson and Brown (1986) found f (N ) to be 1.39 individuals/year, and Andow et al. (1990) estimated it to be 0.2 to 1.1 individuals/year. Andow et al. (1990) used mark-recapture information on the muskrat in Finland and Denmark to estimate D as 51 km2 /year to 230 km2 /year. Notice that this method (i.e., estimating dispersal and vital rates from field data) has the problem that we mentioned above of using information from one site to parameterize dispersal at another site when we know that both growth rates and D can vary with habitat and topography. Nevertheless, if we √ use the range of field-derived values of D and f (N ) to calculate c = 2 rD, we can estimate muskrat range expansion speeds at 6 to 32 km per year. Remember that r in this expression is the intrinsic rate of increase in a density independent population. Range expansion speeds estimated directly from maps such as that in Figure 7.3, suggest c = 1–25 km per year. This is not a bad match between predicted range expansion using the reaction–diffusion model and observed spread in the 1900s, but we would like to see how well the model performs for a variety of species under a variety of circumstances. There have been several efforts to estimate the rate of range expansion for non-native species using the reaction–diffusion model (Equation 7.1). Shigesada and Kawasaki (1997) produced a table, which we emulated in Table 7.1, that summarizes how the reaction–diffusion model has performed in predicting other population range expansions. The key aspect of Table 7.1 is the final two columns where we compare the observed rate of spread, c, to that derived from range maps produced from field data. The two estimates of c match up fairly well although the reaction–diffusion model sometimes underestimates the observed rate of geographic range expansion and occasionally does so by a large margin (e.g., Table 7.1 cereal leaf beetle). We can conclude that the reaction–diffusion model offers a reasonable description of reality, in as much as it offers an estimate of the rate of range expansion within an order of magnitude of the rate seen in the real world. The reaction–diffusion model summarizes the pertinent factors involved in geographic range expansion and puts these pieces together in a logical enough way to capture this aspect of non-native species spread. Despite the high range that Andow et al. (1990) found (Table 7.1), when the reaction–diffusion model fails to mimic real spread rates, it generally underestimates rates of spread, at least in terrestrial situations (c.f. Grosholz 1996). This consistent bias strongly suggests that, at times, we are missing something of importance in our modeling of spread. What is this “missing something?” Recent studies have isolated it as being the prevalence of long-distance dispersal events.
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Table 7.1 Comparison of rates of spread, c, estimated from the diffusion model (Equation 7.1) and field derived range expansion maps (modified from Shigesada & Kawasaki 1997). Species
Muskrat∗ Cabbage butterfly∗ Cereal leaf beetle∗ Gypsy moth† Sea otter (expansion North)‡ Plague§ Rabies
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6–32 9.3–90 1.6–1.7 <2.5 1.74 720 70
1–25 15–170 27–90 3–20 1.4 320–650 30–60
∗ Andow et al. (1990). † Liebhold et al. (1992). ‡ Lubina and Levin (1988). § Nobel (1974). Yachi et al. (1989).
Long-distance dispersal Ecologists have documented dispersal patterns for a variety of species and found that the distribution of dispersing individuals is not well-described by a normal distribution, as assumed by the reaction–diffusion model (Lonsdale 1993; Weber 1998; Buchan & Padilla 1999). Instead dispersal often follows a pattern described as leptokurtic, meaning that actual dispersal is skewed to further distances than predicted by a normal distribution. A leptokurtic dispersal curve suggests there are more extreme movement events, or long-distance dispersal events, than we would expect, based on the reaction– diffusion model. There are two types of leptokurtic dispersal curves. The first still fits many of our assumptions because it still aproaches a constant invasion speed, c. If there is enough long-distance dispersal, though, that the distribution approaches zero at the tails of the distribution more slowly than a negative exponential, then the dispersal curve is described as having “fat tails” (Kot et al. 1996, Fig. 7.6). These fat-tailed distributions are extremely leptokurtic and can lead to accelerating invasion speeds. Mark Kot and colleagues (1996) suggest that in most documented examples of dispersal, these leptokurtic and even fat-tailed dispersal events are not rare, which could explain why the reaction–diffusion model sometimes drastically underestimates the speed of range expansion. Clark (1998) includes an excellent discussion of how normal dispersal kernels, leptokurtic dispersal, and extremely leptokurtic or fat-tailed dispersal kernels can change estimates of spatial spread. Several authors have explored the role of long-distance dispersal in a variety of models. The simplest approach is to assume that there are two types of dispersal, long and short. Hengeveld (1989) termed such dispersal “stratified dispersal.” We could imagine that such dispersal could come from a directional bias (e.g., wind or stream
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Figure 7.6 A normal and leptokurtic distribution sketched in the same graph. A leptokurtic distribution is one that has many more rare observations than the normal distribution. These rare events are “captured” in the tails of the distribution, thus making them look “fat” as compared to the normal distribution. In the context of non-native species’ range expansion, the fat tails are a product of having relatively many individuals disperse over long distances.
flow) or from two different modes of dispersal that require two diffusion expressions (e.g., seed dispersal via wind or via animals). We can also imagine dispersal that does not fit our assumptions of normality (i.e., perhaps produces fat tails). For example, many seed shadows (the density of seeds as one moves away from the parent plant) are best described by a Weibull distribution (Fig. 7.7), which can be far more skewed than a normal distribution. In order to include this sort of long-distance dispersal, it may be necessary to use a different dispersal kernel than the normally distributed diffusion term in the reaction–diffusion model. Below we detail approaches to directional bias, stratified diffusion, and continuous and discrete time methods for including alternative dispersal kernels.
Directional dispersal A key assumption of the reaction diffusion model is that dispersal is isotropic, meaning dispersal (theoretically) occurs at equal speeds and covers equal distances in all directions. This is clearly not the case for many types of species, including freshwater and marine species that disperse via water currents and plants that disperse seeds via wind currents. In these organisms, the direction and speed of the wind or water current drives the direction, distance, and speed of dispersal. In these cases, the speed of range expansion will be much higher in the direction favored by the wind and water current and much lower in the direction opposite these forces. The muskrat, for example (Box 7.1), moved faster along waterways than in other directions. Shigesada and Kawasaki (1997) offer a simple mathematical solution to this problem by adding an advection term, υ, to the reaction–diffusion model of Skellam (1951). Thus, the maximum rate of range expansion in the direction of the wind or water
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Figure 7.7 The Weibull distribution offers considerable flexibility, particularly in producing skewed distributions. Notice that changing only the parameter β offers a broad range of distributions. Reprinted from (http://www.statisticalengineering.com/extreme_value.htm) with permission of Charles Annis.
√ current is 2 rD + υ, and the √ minimum speed of range expansion in the opposite direction of the current is 2 rD − υ. Okubo and Levin (1989) and van den Bosch et al. (1990) have derived more sophisticated models of this process. The key to how well these sophisticated models predict the rate of range expansion lies in their ability to capture the physical dynamics of wind and water as they relate to the dispersal of individual seeds, larvae, or juveniles. For example, the distance that wind will carry seeds from trees depends on height of seed release, wind speed and turbulence, and the critical wind speed at which the seed will settle to the ground. In a unique comparison of rates of geographic spread between terrestrial and marine species, Grosholz (1996) observed a few key differences in how we should think about directional dispersal. He compared the rate of range expansion, c, for 10 marine invaders (e.g., barnacles, crabs, and snails) versus 10 terrestrial invaders (e.g., gypsy moths, cabbage butterflies, and muskrats) using the reaction–diffusion model. He found that the speed of range expansion was consistently slower for marine invaders as compared to terrestrial ones and that the reaction–diffusion model consistently overestimated the rate of range expansion for the marine species. Compare this to the conclusion
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we came to above where the reaction–diffusion model often underestimated speed of range expansion for terrestrial organisms. This result is interesting by itself as we might expect marine species to have a comparatively high rate of range expansion since their larvae can be carried over very large distances via ocean currents. Grosholz (1996) noticed, however, that there is a clear difference in modes of dispersal between the marine and terrestrial species. The marine species all disperse during larval stages via advective sea-surface currents, whereas all of the terrestrial examples disperse via non-directional means (e.g., juvenile movements away from parent, human contact, or animal vectors). Grosholz (1996) suggests that ocean currents have a wider range of effects on dispersal of marine organisms because currents vary in speed and sometimes even direction with season and decadal weather patterns. Thus, in some years, marine larvae are dispersed over long distances, and in others they are essentially not dispersed at all. Over many years, and given sufficiently infrequent dispersal opportunities, range expansion will be quite slow. The reaction–diffusion model does not explicitly include this type of variation in dispersal distances between years, and thus it tends to overestimate range expansion rates of marine species that disperse via directional currents. There are likely several other such examples involving dispersal mechanisms that rely on water or wind currents, however, as of now this topic has been relatively less studied (although see Caswell et al. 2003 for an example of a model that incorporates stochastic variation in the dispersal coefficient, D).
Stratified dispersal Classifying dispersal into two distance categories, short and long, results in “stratified” dispersal (Hengeveld 1989). The speed of range expansion in these models may accelerate through time (up to the speed predicted by the longer of the two dispersal distances), but stratified dispersal models can also produce a constant speed of range expansion depending on how the two dispersal modes behave in combination (Hengeveld 1989). To understand stratified dispersal, we start with a primary colony of the non-native species that is expanding its range at a constant rate, c, following the reaction–diffusion model (Fig. 7.8). A few individuals from this primary colony disperse well outside the current range and start their own satellite colony. Just like the primary colony, the satellite colony is expanding its geographic range at a constant rate, c, following the reaction–diffusion model. This satellite colony is of sufficient distance from the primary colony that the two are growing and expanding independently of one another. Eventually, the two will come into contact, at which point we simply re-draw the boundary of this merged colony to encompass both. This feature of a stratified diffusion model makes it a “spatial contact model,” in which movement connects processes at different scales (Mollison 1977). The merged colony continues to expand at a constant velocity, and it may produce another satellite colony given enough time and the right conditions. This model of non-native species range expansion will produce accelerating speeds of range expansion if (i) satellite colonies are sufficiently far from the primary
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Satellite C
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Figure 7.8 Graphic of stratified dispersal showing the relationship between the primary colony of the non-native species (i.e., range limit achieved from initial release) and a satellite colony (i.e., range limit achieved by a secondary expansion after long-distance dispersal). Dispersal can be categorized into two modes, short (solid arrows) and long (dashed arrow), that are co-occurring in time. The primary and satellite colonies are expanding their ranges at a constant rate, c, according to the diffusion model. When the two colonies come into contact, they will merge to form one very large colony. Redrawn from Shigesada and Kawasaki 1997.
colony to grow substantially before merging, or (ii) a large number of satellite colonies appear through time (Shigesada & Kawasaki 1997). If these conditions are not met, the non-native species’ range will expand at a constant rate, c. Several of the examples of stratified dispersal include two distinct modes of dispersal, and nearly all of them show an accelerating rate of range expansion. For example, Mack (1981) described two modes of dispersal for cheatgrass (Bromus tectorum). Normally cheatgrass disperses by attaching its seeds to the fur of animals that trot over to a relatively close location where the seeds are shed. In North America (especially in the early twentieth century) many of these seed-carrying animals trotted onto livestock trains and trucks to be transported across hundreds of miles. Seeds shed at the other end of train trips were certainly long-distance dispersers. Propagules that survived in distant locations created satellite populations of cheatgrass. Shigesada and Kawasaki (1997) suggest that such stratified dispersal produced the accelerating rate of range expansion observed in cheatgrass in North America (Fig. 7.9, Box 7.2). Other examples of stratified diffusion, in which the long-range dispersal stratum is produced by human actions and the short-range by normal dispersal mechanisms, include the movement of the zebra mussel (Bossenbroek et al. 2001), the Argentine ant (Suarez et al. 2001), and the house finch (Veit & Lewis 1996). Indeed, we can think about all non-native species as having (at least) two strata/modes of dispersal if we broaden the definition of dispersal to include the rare, but important transport out of the native range and release into a non-native locale (e.g., Suarez et al. 2001), but such a definition would conflate the transport and spread stages. Finding a way to incorporate the many different geographic scales of dispersal into models of spread is a continuing challenge for ecologists.
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Figure 7.9 Accelerating expansion of the geographical range of cheatgrass across North America. Shigesada and Kawasaki (1997) suggest that stratified dispersal produced the observed relationship, with long-distance dispersal mediated by human actions and shortdistance dispersal following normal ecological processes. Data originally published in Mack (1981).
Variable dispersal kernels We have discussed constant dispersal and two distinct modes of dispersal, but how should we deal with dispersal that varies in more than two distinct “stratified” ways? Dispersal could vary across habitats, according to weather patterns, according to different interactions with other species, or for a variety of other factors. For these broader ranges of dispersal types, we need variable dispersal kernels that allow the dispersal distance to change based on some other factor. In the reaction–diffusion model, the only included factor was the density of the dispersing organism and distance from the release point. With variable kernels we can include any other factor we choose. Variable dispersal kernels have generally relied on integrodifferential (continuous time) and integrodifference (discrete time) equations (Lee et al. 2001). Most ecology students are familiar with differential equations that look at small changes in density over infinitely small time steps, dN /dt = rN , for example. The reaction–diffusion models that we have discussed so far are differential equations. Integrodifferential equations add an integral into the dispersal kernel to sum over all the dispersal distances. A very general equation in one dimensional space (i.e., a linear system) would be ∂N (x1 , t ) (7.4) = g [Nt (x2 )]k(x1 , x2 )dx2 − g [Nt (x1 )] k(x2 , x1 )dx2 ∂t where Nt (xi ) is the density of organisms at a particular point and a particular time, k is the dispersal kernel, and g is the expression for population growth. Translated into words, this equation says that population change at a point in space and an instant in time comes from redistribution, described by k(x1 , x2 ), and population growth, described by g [Nt (xi )], both added up over space. Because this equation describes
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Box 7.2 Brilliant disguise Often we assume the current distribution of a non-native species is the product of one initial release of individuals, or of a series of localized releases. Usually this initial release site is considered to be the site of first record for the species in its non-native range. It is this information (site of initial release and range boundaries through time) that we feed into the calculations of the rate of geographical range expansion. It is not hard to imagine, however, more than one initial release site for a non-native species that is now widespread. If these release sites are far enough apart, the calculations of range expansion could be way off the mark. In essence, we would be lumping the secondary release points in as “spread” rather than as unique invasion foci. This error will lead to badly overestimated rates of range expansion. Until recently, the only way to disentangle separate releases from spread was to unearth as much written, historical information as possible. Often these written accounts just muddied the water further, since we could not be sure that the author of these accounts relayed information without error. With the advent of genetic analysis, however, we have gained a powerful tool to determine the source in the native range of populations of a widespread nonnative species. Because species in their native range will often show some degree of genetic divergence between populations, it is often possible to match up an individual collected in the non-native range to its native source population. By collecting individuals across the non-native range and matching them to a location or population in the native source, we can begin to piece together the history of the invasion. One of the better examples in this regard comes from Stephen Novak and Richard Mack’s study of cheatgrass (Bromus tectorum) invasion in North America. Cheatgrass is an annual grass with a wide native distribution that includes Eurasia and North Africa. It has been introduced to several locations besides North America, including Argentina, New Zealand, and Hawaii. Cheatgrass is now distributed across North America (Fig. 7.10). How did this species manage to spread across North America, and at what rate did this occur? In an effort to answer this question, Novak and Mack (1993, 2001) took genetic samples of individuals across the native and non-native range. Based on multilocus genotypes, they found that individuals now living in North America were drawn exclusively from central Europe and the western Mediterranean. This is not too surprising since the United States regularly imported most of its crop seed from this region up until the 1900s, and “waste seed” (e.g., cheatgrass seed) was purposely packed in with these valuable crop seeds to bulk up the weight of the bag and thus make an extra profit (Novak & Mack 2005). The more remarkable finding was that the invasion of western North America stemmed from at least six independent introduction events (Novak & Mack 1993, 2001). However, in eastern North America, the genetic markers indicate that perhaps only two independent introductions occurred. The individuals in western North America were for many years thought to have been introduced from eastern populations, and to some extent the genetic analysis of Novak and Mack (2001)
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back this up. However, they found ample evidence that most of the western North American individuals were derived from introductions directly from Europe and the Mediterranean. All told, the work of Novak and Mack indicates a much more complex and dynamic history of range expansion of cheatgrass in North America than anyone previously suspected. It is not clear how often we should expect to see such complex invasion histories; however with the continued use of genetic analyses in invasion ecology, we should soon find out.
Figure 7.10 Map showing allelic variation of cheatgrass (Bromus tectorum) across North America. Open circles are populations fixed for the b allele, open triangles are populations fixed for the c allele, open squares are polymorphic populations with a high frequency of the b allele, and filled squares are polymorphic populations with a high frequency of the c allele. Reprinted with permission from Novak, S.J. and R.N. Mack. 2001. Tracing plant introduction and spread into naturalized ranges: genetic evidence from Bromus tectorum (cheatgrass). BioScience 51: 114–122.
only the change in population distribution, we add up growth and movement into the current spot from all other spots and subtract movement away from the point x1 . These equations offer a more precise way of describing dispersal in that they are flexible enough to handle any dispersal probability distribution, including distributions that result from real field data. The summations within integrodifferential equations can be very simple or somewhat complex. Not all kernels lead to analytically tractable
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systems (i.e., sometimes the equations do not produce a simple set of solutions). In general, tractable equations assume a homogeneous environment where the actual starting and ending points are unimportant; instead only the distance between these points affects the likelihood of getting from point a to point b. The diffusion expression is an example of such a kernel in which the starting point has no effect on the probability of where an individual actually ends up. Notice in Equation 7.4 that we simplified by expressing a point simply as x1 rather than using x–y coordinates. In general these equations have been formulated in one dimension, meaning that organisms can only move along a single axis. The equations are complicated enough without multiple dimensions, and the results in one dimension are easy to extrapolate to a real world in which organisms might move in two dimensions as long as movement is symmetrical. Remember also that the only reason to specify specific points (since the distance between points, or x1 − x2 , is the important factor in determining movement probability) is to relate the movement to the density of the organisms at a given point. Clearly, these equations can be much more complex than the simple reaction–diffusion equations described earlier. van den Bosch et al. (1992) used continuous time integrodifferential equations to examine the influence of variable dispersal kernels in rates of range expansion. The particular model that van den Bosch and his colleagues examined also allowed them to look at age structure so that dispersal and reproduction could occur at particular stages in the life cycle of an organism. Their results, therefore, reflect the effects of age structure as much as that of their choice of dispersal kernel. They used this approach to look at the range expansion of collared doves (Streptopelia decaocto), starlings (Sturnus vulgaris), and cattle egrets (Bubulcus ibis), which all disperse primarily in the juvenile phase or their first year. They also reanalyzed the muskrat data, including in their model the fact that muskrats disperse both in their first year and as adults but at different rates. They found that variation in dispersal affected spread more than variation in age of dispersal and that dispersal kernels with more long-distance dispersal led to faster rates of spread. Kot et al. (1996) explored this issue more generally with a discrete time integrodifference equation of the general form: Nt +1 (x1 ) =
k(x1 , x2 )g [Nt (x2 )]dx2
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where Nt (xi ) is the density of organisms at a particular point and a particular time, k is the dispersal kernel, and g is the expression for population growth. In words, this equation says that the population at a particular point in space and time is determined by population growth and redistribution in the previous time interval, added up over all space. Notice the similarities to Equation 7.4. Basically, the discrete time model looks only at the first half of Equation 7.4. There is no need to subtract movement out from a particular point, because the first integral gives us the distribution over all space (not the change in the distribution, but the actual distribution). Basically, Equation 7.4 subtracted current distribution from the future distribution. In general ecologists use discrete time formulations like Equation 7.5 more than continuous time formulations
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Figure 7.11 When leptokurtotic dispersal patterns are included within integrodifference equations, the non-native species’ geographical range expands at an accelerating rate through time. This is depicted here using the probability density function. The distance between the lines in this graph increases through time as compared to a situation with a constant speed of range expansion.
like Equation 7.4. Discrete time equations work particularly nicely for organisms that have breeding seasons or times of year when dispersal is more likely. Kot and his colleagues (1996) found that including dispersal kernels that produce fat tails (i.e., more individuals disperse long distances than with a negative exponential dispersal kernel) resulted in an accelerating rate of range expansion (Fig. 7.11) just as stratified dispersal did. Thus, unlike the reaction–diffusion model, integrodifference models frequently predict that the speed of invasion is not constant through time. Instead it increases through time. Higgins and Richardson (1999) show that increases in the probability of long-range dispersal events of even a minor amount over what is dictated by the reaction–diffusion model can lead to profound underestimates of the rate of range expansion. Although these variable dispersal kernel models allow more latitude in dispersal assumptions, they must make other simplifying assumptions in order to be analytically tractable. A number of problems arise from these assumptions. First, these models sometimes predict continuously accelerating rates of spread with no eventual deceleration or stabilization. Clearly it is not realistic for acceleration to continue indefinitely, just as it is not possible for populations to grow exponentially forever. Sooner or later, something has to slow the growth and geographic spread of a population. More recent incarnations of integrodifference equations have bounded the rate of spread by looking at maximum dispersal distances (e.g., Clark et al. 2001), and we refer interested readers to this work for more details. It is important to point out, though, that, within the time frame of interest to invasion ecologists, it is possible for spread to accelerate (Hastings et al. 2004). An invading organism can enter a habitat and spread with accelerating speed until it fills an entire continent or the entire geographical area that fits its habitat tolerances. A second problem lies in gathering the field data needed to parameterize longdistance dispersal kernels. It is hard enough to obtain field data on any dispersal events, but good information on relatively rare long-distance dispersal is almost impossible to gather. This issue is much harder to address, and Clark et al. (2003) suggest that the
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quality of dispersal data may limit our ability to predict the speed of range expansion when long-distance dispersal is common.
Other forms of heterogeneity There are several other factors that may affect the rate at which a non-native species expands its geographic range. Clearly the presence of other species that affect the growth or dispersal of an invading species will influence the rate of geographic spread. Ecologists have only just begun to model these sorts of biotic interactions in spread. The difficulties in examining species interactions across space make generalization difficult (Holt & Hoopes 2005; Hoopes et al. 2005). Some progress has been made with models of interspecific competition. Yuzo Hosono (1998, 2003) has incorporated Lotka–Volterra competition in a diffusion model. He took two equations of the form of Equation 7.1, but for f (N )N he substituted the standard Lotka–Volterra competition equations. He found that competition generally slows the rate of spread, but it can also lead to nonlinearities in spread rates, so that rates switch between increasing and decreasing as densities of the two species change (Hosono 1998). This result is robust to other ways of including competition. Using an integrodifference model, Hart and Gardner (1997) found that even a very weak competitor could reduce the rate of expansion of an invader as long as the competitor was established before the invasion. This makes intuitive sense because the condition of prior establishment means that the native competitor is always present, regardless of how fast the invader spreads. In this scenario the invader will always encounter the native competitor and will always experience a reduction in population growth. Unfortunately, these results are from models and have not been tested empirically or applied to empirical data. It seems unlikely that a single native competitor will be present in all locations or that competition coefficients will remain constant across space. Empirical tests of these sorts of models could greatly improve not only our ability to predict spread, but also our ability to predict impacts on specific native species that frequently compete with an invader.
Spatial heterogeneity Some spatial predator–prey models have also examined the spread of invasive species. Fagan et al. (2002) used a reaction–diffusion model to examine the effects of specialist and generalist predators on an invading species. This scenario is useful because predators are often released as biocontrol agents to control spreading pest species. Fagan et al. (2002) found that specialist predators were likely to reduce the density of the invader but not to reduce the spread, while generalist predators might actually slow the rate of spread. These results might seem backwards, but the specialist predator cannot survive without the invader, so it cannot advance beyond its host to meet the invasion with high densities of predators the way the generalist can. On the other hand, the specialist does reduce the density, but, because Fagan et al. (2002) looked at a density-dependent
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scenario, the reduction in density keeps the population at maximum reproductive capacity and, therefore, does not slow down the spread. The only scenario in which the predators could successfully reduce the geographic range of the invader was if a specialist predator attacked an invader that had Allee dynamics at low density (Fagan et al. 2002). Again, these results are interesting, particularly in light of biocontrol efforts to manage invasive species. Tests of these results are ongoing, but further investigations can only improve our understanding of the interactions between native and non-native species and how these interactions affect the spread of invaders.
Temporal heterogeneity Temporal heterogeneity can also alter the rate of spread. The integrodifferential model that van den Bosch et al. (1992) used to examine the effect of age structure showed that dispersal or reproduction that changed with age generally decreased the rate of spread. This result makes sense if we think about reproduction and dispersal potential not being constant throughout a lifetime, but instead coming in concentrated periods. As variance increases in reproduction or dispersal over time, actual population growth decreases (Hastings et al. 2004). Similar results have been found using integrodifference equations where the rate of spread decreased as temporal variability increased (Neubert et al. 2000; Clark et al. 2001). Clark et al. (2001) found that variance made the biggest difference in the spread of organisms with fat-tailed dispersal kernels, which paradoxically helps to explain how such dispersal can accelerate over short time periods. Once there is less temporal heterogeneity, the invasions speeds up markedly. Results with spatial variation depend somewhat on how variation compares to the most commonly encountered or sampled habitat. Because it is more difficult for us to sample long-distance dispersal, we tend to underestimate potential rates of dispersal as discussed above. Thus, variation in the environment often increases rates of spread in models that examine fits to real data (Shigesada & Kawasaki 1997; Bossenbroek et al. 2001; Hastings et al. 2004). Unfortunately, it is difficult to create generalizations yet about the role of spatial variation per se on range expansion of invading organisms, as more work remains to be done.
Models with implicit space Thus far, we have focused on the methods used most frequently to model spatial spread in invasion dynamics, and we have gone into some detail on them because they are not the models that we all encounter in our initial ecology classes. Some of the more familiar models also have invasion applications, particularly when examining heterogeneity. The most common spatial model in ecology is the metapopulation model (Levins 1969). Most of the models that we have discussed in this chapter are simple models in which space is explicit, as opposed to implicit. Patch models, like the Levins metapopulation model, are implicit spatial models, which assume that separations exist between habitat patches without quantifying the arrangement of the patches. Instead, these implicit spatial models assume set amounts of movement between patches. Despite their lack
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of realistic features, these models simplify the world in ways that help to distill very basic mechanisms. They are useful for examining a single species in several patches of different habitat quality, different predator–prey dynamics, or the coexistence of competitors. Basically the models are useful in any situation in which trade-offs between movement and ecological interactions assist in the maintenance of species across landscapes. These models have implications for invasion ecology if there are specific interactions that we want to examine, such as the interaction between an invader and a suite of native species. They are less useful for looking at spread, however, because they assume that all organisms can reach all patches and, therefore, that invasion has already occurred and that spread is universal. The exception lies in looking at thresholds for invasion and at examining how temporal variation may alter those thresholds. For example, disease organisms are invaders in new hosts. Disease models are generally patch models that describe the fraction of “patches” or hosts as susceptible, infected, or recovered/immune. Understanding what makes a host susceptible or immune is similar to understanding a threshold that makes a habitat susceptible or resistant to invasion. This threshold may depend on many of the factors described in the establishment section, for example the presence or absence of a specialist predator or resource, maximum or minimum temperatures or water availability, toxins, or a variety of other resources or conditions that set absolute limits on the survival of a particular non-native species. Metapopulation models, like the Hanski incidence function model (Hanski 1994, 2004; Hanski et al. 1996), that refer more explicitly to space by making colonization and dispersal contingent on explicit spatial dimensions fit better with some of our discussion of spatial spread. There are several excellent texts on metapopulation and metacommunity dynamics that examine these implicit space models (e.g., Hanski & Gilpin 1997; Hanski 1999; Holyoak et al. 2005), so we will not go into more detail here.
Models with explicit discrete space In order to look at spread and predict the movement of a non-native organism into real habitats across a real landscape, it is important to have a modeling framework that acknowledges real distances between habitats as well as the challenges of crossing those distances. Any model that quantifies spatial relationships between points or between habitats is spatially explicit. All the continuous spatial models that we discussed above are spatially explicit models. They are also extremely computationally intense. One alternative to these computationally intense continuous space models is to build models based on discrete sections of space by dividing the landscape into cells and looking at movement across cells. Lattice models and cellular automata are two such approaches that you may have encountered. They are often used in theoretical examinations of two species interactions, particularly to look at predator–prey dynamics or other sorts of dynamics in which distinct behaviors can happen in different populations and change with the immigration of new organisms. Individual based models (IBMs) are another example of these sorts of approaches. Lattice models, cellular automata, and
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IBMs are all used more for theoretical explorations than for applications that predict the spread of non-native species. For predictive applications, geographic information system (GIS) models allow invasions ecologists to match dispersal and population growth information with specific habitat features. These landscape–scale models rely less on ecological interactions and more on thresholds defined by previous experiments. One example is research examining the spread of the hemlock woolly adelgid (Adelges tsugae), an insect native to Asia that is currently leading to massive mortality in hemlock trees (Tsuga canadensis and Tsuga caroliniana) in the northeastern United States. Maps of hemlock distribution indicate the potential range of the invader. By examining how other factors overlay the potential range and the actual range, researchers can identify factors that may impede spread. Thus far, only minimum temperature seems critical. By allowing dispersal and establishment to proceed according to probabilities based on habitat features, these models allow us to examine more statistical risk assessments. They are species-specific and are limited by the decision rules used to establish movement probabilities, but they are an excellent way of examining potential spread or looking for correlations between past spread and habitat features. They are extremely data-intensive and used primarily with species with large economic impacts. Generalizations from these sorts of models frequently relate to management in actual geographic locations. They are the least generalizable spread models, but also the most realistic. They, therefore, are also the most prone to misinterpretations due to data biases. Remember that all modeling efforts are limited by the data and assumptions used to build them. Although broad, simplistic models cannot point to specific habitat features that impede or encourage spread and cannot predict exact rates of spread, they also do not pretend to offer this information.
Summary The study of non-native species’ range expansion relies in large part on mathematical modeling because the biological processes of spread are difficult to observe or manipulate in natural settings. We have introduced you to the basics of these models, starting with the reaction–diffusion model and highlighting more recent advances that explore a broader range of dispersal types, including incorporating stratified and directional dispersal, and the use of integrodifference equations (i.e., recognizing the commonness of long-distance dispersal events). We did not delve into the mathematics of more recent models as it would require significantly more space to do so, and many users of this text may not have the mathematical background or inclination to tackle the derivation of the models with élan. For those of you who found this chapter too light in detail, we refer you to several recent articles on non-natives species’ range expansion given in the last half of this chapter, and to the delightful work of Shigesada and Kawasaki (1997). The next chapter should satisfy those of you who are happy the mathematical descriptions are over and would instead like to delve more deeply into the concepts and ecology behind dispersal and range expansions.
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Further complexities Clark, J. S., M. Lewis, and L. Horvath. 2001. Invasion by extremes: population spread with variation in dispersal and reproduction. American Naturalist 157: 537–554. Clark, J. S., M. Lewis, J. S. McLachlan, and J. HilleRisLambers. 2003. Estimating population spread: What can we forecast and how well? Ecology 84: 1979–1988. Fagan, W. F., M. A. Lewis, M. G. Neubert, and P. van den Driessche. 2002. Invasion theory and biological control. Ecology Letters 5: 148–157. Hosono, Y. 1998. The minimal speed of traveling fronts for a diffusive Lotka–Volterra competition model. Bulletin of Mathematical Biology 60: 435–448. Weinberger, H. F., M. A. Lewis, and B. T. Li. 2002. Analysis of linear determinacy for spread in cooperative models. Journal of Mathematical Biology 45: 183–218.
Companion papers Andow, D., P. Kareiva, S. Levin, and A. Okubo. 1990. Spread of invading organisms. Landscape Ecology 4: 177–188. Clark, J. S. 1998. Why trees migrate so fast: confronting theory with dispersal biology and the paleorecord. The American Naturalist 152: 204–224. Hastings, A., K. Cuddington, K. F. Davies, C. J. Dugaw, S. Elmendorf, A. Freestone, S. Harrison, M. Holland, J. Lambrinos, U. Malvadkar, B. A. Melbourne, K. Moore, C. Taylor, and D. Thomson. 2004. The spatial spread of invasions: new developments in theory and evidence. Ecology Letters 8: 91–101. Holmes, E. E. 1993. Are diffusion models too simple? A comparison with telegraph models of invasion. American Naturalist 142: 779–795. Lonsdale, W. M. 1993. Rates of spread of an invading species: Mimosa pigra in northern Australia. Journal of Ecology 81: 513–521. Skellam, J. G. 1951. Random dispersal in theoretical populations. Biometrika 38: 196–218.
8 Ecological Processes and the Spread of Non-Native Species Overview The spread of non-native species entails many of the same mechanisms involved in the earlier stages of invasion. Individuals of the non-native species must disperse into a new habitat, mimicking the transport phase, and these individuals must then establish a selfsustaining population in this new habitat, just as in the establishment phase. Spread is, therefore, a microcosm of the entire invasion process. We discuss the roles of population growth, dispersal, biotic interactions, and temporal or environmental heterogeneity in the spread of non-native organisms. You may note that all of these topics arose in Chapter 7 as important considerations in models of geographical spread. In this chapter we will look at their ecological expressions rather than their mathematical ones although many of our references and lines of evidence will still come from modeling exercises. By examining both empirical patterns and model results, we also discuss how ecological factors can produce complex spread patterns, such as long lag times between establishment and spread and boom and bust population dynamics.
Population growth Population growth is one of the most basic requirements for the spread of an invasive species. So basic, in fact, that we often overlook its importance in promoting geographic range expansion. All the models discussed in Chapter 7 combined population growth with different formulations for the movement of organisms. In most models population growth is depicted as density-independent, showing exponential (or geometric) patterns of increase. In such scenarios, numbers of individuals increase slowly at first, and after some point, the growth momentum of the population leads to a sudden “eruption” of numbers. Such eruptions may bring a cryptic species above some threshold of detection (Carey 1996) so that, to us, the species seems to be appearing at new locations from out of the blue when instead it is merely increasing in numbers at these locations to the point that we begin to actually see them.
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Although population growth above such thresholds is not spread, it is often difficult to distinguish from spread. Thus before we try to control or eradicate incipient populations of known invasive species, it is important to think about the species’ life history characteristics, population growth rates, and the error inherent in all methods of detecting small populations. For example, the Mediterranean fruit fly, also known as the medfly, appeared in numerous locations in California following particularly mild seasons (Fig. 8.1). Although many people in these localities claimed that these San
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Figure 8.1 Mediterranean fruit fly (Ceratitis capitata) occurrence records for the Los Angeles Basin (USA) from 1975 to 1994. Each time the medfly was detected it was assumed to be an independent introduction event. However, Carey argues persuasively that these detections followed good medfly years and thus local populations simply rose above the detection threshold and were counted. The medfly had become established after the initial sightings in the early 1970s and spread throughout the basin, undetected. Reprinted from Carey, J. R. 1996. The incipient Mediterranean fruit fly population in California: implications for invasion biology. Ecology 77: 1690–1696, with permission from the Ecological Society of America.
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emergent populations were recent imports, Carey (1996) makes an excellent argument for sudden population growth as the explanation for the new sightings. Carey (1996) suggests that the medfly successfully established a self-sustaining population after the initial “outbreak” recorded in the 1970s. The medfly then expanded its range across much of the Los Angeles Basin, but did so while maintaining such low population sizes that no one detected them. In good medfly years (i.e., years with a mild climate), these populations increased to sizes that were suddenly noticeable. The spread itself, however, had happened years before. Density-dependent population growth, in which the per capita survival or reproduction changes as density increases, can also have a very large effect on population spread. Remember from the simplest reaction–diffusion model from Chapter 7, ∂N = f (N ) + D ∂t
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that the square root of the area occupied by an invader is linearly related to the time since initial expansion out of the point of establishment site (i.e., square root of the area is proportional to time). The square root part of the equation comes from the constant distance moved per unit time, which requires a constant growth rate, rather than a changeable one. If the growth rate decreases with population size, then the spread rate will also decrease. On the other hand, reverse density dependence, in which the population growth rate increases with population size, will lead to accelerating spread. This sort of dynamic has been examined in a number of models (Lewis & Kareiva 1993; Keitt et al. 2001). With an Allee effect, in which population growth decreases at low densities (see Chapter 4), individuals in small populations may have difficulty reproducing, let alone dispersing their offspring. The Allee effect leads to decelerating spread rates for small populations, which can keep a population from ever spreading beyond its original establishment site. When populations do overcome Allee effects, spread rates may accelerate suddenly (Lewis & Kareiva 1993; Keitt et al. 2001). Veit and Lewis (1996) found that an integrodifference model (see Chapter 7) with an Allee effect incorporated into the population growth segment of the model was a good fit for the spread of house finches (Carpodacus mexicanus) across eastern North America, where they are non-native. If this example sounds too much like the math of the previous chapter, remember to think of the mathematical models as experiments, ways of looking at the effects of particular factors. Intuitively it makes sense that any species that fares better with a threshold number of conspecifics will be more likely to spread if that threshold has been surpassed. It also makes sense that, once the species attempts to spread, that threshold will need to be surpassed at each new site. Such hurdles will limit population range expansion. It would be easy to misinterpret this simple demographic or behavioral constraint as a limit imposed by physiology or biotic interactions. Thus, it is not surprising that nonnative species with Allee effects have often been interpreted as innocuous, naturalized species with little tendency to spread and low danger of impacts on native species until, like the house finch, they achieve a sufficient population size and suddenly erupt across
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Figure 8.2 The graphs indicate alterations of the per capita growth rate with changes in population density. The first panel (a) indicates no Allee effect and the second (b) a weak Allee effect. A critical threshold below which a population declines to zero is only possible with a strong Allee effect (c). At environmental boundaries, this critical threshold may lead to “range pinning,” or the confinement of a non-native to one environment, because it can not produce a large enough population of dispersers to survive in the new environment. Weak Allee effects can contribute to lag times in the spread of an invasive species. Reprinted from Taylor, C. M. and A. Hastings. 2005. Allee effects in biological invasions. Ecology Letters 8: 895–908, with permission from the Ecological Society of America.
the habitat. Keitt and his colleagues (2001) use the terms “invasion pinning” or “range pinning” to describe a species that is pinned to a particular location by an inability to overcome a population threshold (Fig. 8.2). This sort of dynamic is more common along environmental boundaries where the change in habitat quality on either side of the boundary leads to distinct differences in population growth. Spreading across the boundary requires a growth pulse as opposed to the more incremental growth possible in a homogeneous environment (Keitt et al. 2001). Because population growth can determine whether or not a non-native species ever moves from an initial establishment site, it is critical to the spread of invasive organisms. Once population sizes and growth rates are sufficient to allow dispersal, their relative
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importance depends on the particular species. In general, long-distance dispersal will be the most important factor in determining the speed of range expansion so long as growth rates remain high enough to assure a constant supply of dispersing individuals. If growth rates hover close to equilibrium (i.e., no growth or decline), with lots of variability, then growth rate rather than dispersal distance may have the largest effect on the spread of invading organisms. Caswell and his colleagues (2003) used sensitivity analysis to examine the relative importance of population growth and dispersal in the spread of three different bird species (pied flycatcher Ficedula hypoleuca, starling Sturnus vulgaris, and sparrowhawk Accipiter nisus) after introduction to a new country. Although dispersal was more important for the pied flycatcher in the Netherlands and the starling in the United States, demography and population growth were critical to all three. In fact, for the sparrowhawk in England and the Netherlands, population growth and demography explained 86% of the observed spread rate (Caswell et al. 2003).
Dispersal It makes intuitive sense that dispersal is critical for the spread of invasive species and that the distribution of dispersal distances among members of a population can greatly affect the rate of spread of invading species. This topic was covered in detail in Chapter 7, and we will, therefore, only summarize the previous comments. We will focus most of our discussion on the practicalities of using dispersal data to predict the spread of non-native organisms as well as the ecological factors that contribute to dispersal. The most important factor in dispersal is the long-distance portion. If we look at discrete dispersal distances, with just long or short alternatives, models show that increasing jump dispersal (or long-distance dispersal) increases the spread of the species (Holmes 1993; Shigesada et al. 1995; Shigesada & Kawasaki 1997). With more variation in dispersal distances, increasing the proportion of the population that moves further alters the speed of the invasion front (van den Bosch et al. 1992; Kot et al. 1996). If the proportions change so that more individuals are further from the origin than in a negative exponential distribution, then the wave front may actually accelerate (Kot et al. 1996). Caswell and his colleagues (2003) found that the 10% of individuals who moved the furthest had the largest single effect on the spread of the three bird species mentioned in the population growth section above, regardless of the relative importance of population growth or dispersal. Unfortunately for invasion ecologists, it is very difficult for us to use dispersal data with the models of spread from Chapter 7 to produce accurate predictions of range expansion because we rarely have accurate measurements of long-distance dispersal (Andow et al. 1990; Liebhold et al. 1992; Lonsdale 1993). When we do have longdistance dispersal estimates, we tend not to have accurate short-distance dispersal data. This can also throw off our estimates by changing the density and, therefore, the population growth aspects of the equations (Clark et al. 1999). Methods for measuring dispersal vary with the taxon of interest, and we have more reliable methods for some taxa than others. For plant species, seed traps are
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helpful, and for animals radio collars and hair traps or track pads may help identify dispersing juveniles. Sampling effort must increase with distance from the source population, however, in order to maintain the same degree of data reliability as at short distances. This increased effort may be prohibitively expensive while increased distances may also require higher levels of cooperation across jurisdictional and property boundaries. New genetic techniques (Cain et al. 2000) as well as new statistical approaches (Clark et al. 2003) are improving our ability to determine the proportion of individuals that move different distances, but increased long-distance sampling will always be useful for understanding how invasive species spread across new habitats. In measuring dispersal it is important to consider the different types of natural dispersal vectors. Different seed predators or fruit eaters might move seeds different distances from their parent plants. Many edible fruits are ingested by a wide range of omnivores and may have highly variable seed shadows depending on the abundance of different animal species in the habitat. Even when animals do not ingest the seeds of a fruit, they often will remove fruit (and the seed that the flesh covers) for eating and will, thus, move the resultant offspring further from the parent plant than wind or gravity would if the fruit merely dropped. Dispersal by wind or water does not necessarily reduce individual variation in dispersal distances. For example, seeds may travel further along sloughs or rivers than across dry plains. This dynamic explains some of the variability in dispersal distances between individuals of Mimosa pigra, catclaw mimosa, an aggressive woody shrub that forms impenetrable, prickly thickets up to four to five meters high (Lonsdale 1993). This fast growing shrub makes infested areas inaccessible to animals and people and interferes with stock watering, irrigation, and recreational use of waterways. Because its seeds float and are frequently water-borne, it invades watercourses and seasonally flooded wetlands in tropical and sub-tropical regions. In a comparison of spread across a single wetland and across the entire western portion of the Northern Territory of Australia, Lonsdale (1993) found that the plant spread primarily along water ways and that the spread was far greater than that predicted by a reaction–diffusion model because the individual variability in spread rates was higher than the normal Gaussian distribution assumed by Skellam’s (1951) model (see Chapter 7; Fig. 8.3). Temporal variability, primarily in rainfall and water availability, also had strong effects on the rate of spread of the woody shrub (Lonsdale 1993). Marine organisms may experience similar passive dispersal. Many marine animals and plants have pelagic larvae that float in currents, and therefore individuals disperse across distances that vary according to the variation of currents. The pelagic dispersing stage of marine organisms lasts usually a few weeks, with most species settling within 5 weeks (Hedgpeth 1994), but some are able to survive up to 12 weeks (Leppäkoski & Olenin 2000). Although movement of pelagic individuals via ocean currents seems as though it would lead to long-distance dispersal along fairly constant marine currents, there is some argument about whether these long-distance events are common. Grosholz (1996) found that the reaction–diffusion model, despite overestimating dispersal for marine organisms, actually fit marine dispersal patterns better than it did terrestrial dispersal patterns, which the model consistently underestimates (Fig. 8.4).
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Figure 8.3 Dispersal distribution of Mimosa pigra in Australia derived from aerial photographs taken from 1980 to 1986. The distribution of dispersal distances includes several long-distance moves, which makes the reaction-diffusion equation underestimate spread rates. These long-distance dispersals were mediated by water and not wind movement of seeds. Reprinted from Lonsdale, W. M. 1993. Rates of spread of an invading species: Mimosa pigra in northern Australia. Journal of Ecology 81: 513–521, with permission from Blackwell Publishing.
He hypothesized that the very steadiness of marine currents led to fewer unexpected events with larvae settling out of the current in fairly predictable ways (Grosholz 1996). This is merely one study, though, and it is possible that more studies of marine invasions will reveal different results. For example, Leppäkoski and Olenin (2000) examined the spread of seven non-native species in the Baltic Sea and found average spread rates comparable to Grosholz (1996) but also found dispersal from the introduction site at >9 times the average rate. Organisms that do not require natural vectors for dispersal may hitchhike for some portion of their life cycle or may have altered dispersal because of weather or habitat variables. Pathogens and parasites spread to distant locations when their host moves, often leading to two or more types of dispersal curves: one for dispersal under their own power and one for dispersal as a hitchhiker within their hosts. Many pathogens
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Figure 8.4 Difference in spread rate between marine and terrestrial species as calculated by Grosholz. The difference between the two may lie in the consistency in (natural) marine dispersal vectors that rarely include really long-distance dispersal events. In terrestrial invasions, such long-distance events are sometimes possible. Reprinted from Grosholz, E. D. 1996. Contrasting rates of spread for introduced species in terrestrial and marine systems. Ecology 77: 1680–1686, with permission from the Ecological Society of America.
and most parasites also colonize more than one species of host, leading to even more individual variation in dispersal distances. Similarly variable individual dispersal distances can occur for birds and insects. Unusual weather may change migration patterns and blow birds and insects off course. Although birds will frequently right their course before settling in a habitat for breeding, there are sufficient examples of birds nesting outside of their established ranges for bird watchers the world over to add non-local species to their life lists merely by looking out their back windows. Insects blown into new areas tend to remain there and create new colonies, as long as the habitat supports them. Natural dispersal from an establishment site may be only one method of spread. Human transport of non-native species rarely stops once establishment occurs. Instead, the original transport vectors generally remain in place, and additional movements of individuals out of established populations may add new, within-country transport pathways. We see this happening for species purposefully transported in the horticultural industry (Mack 1991) as well as with the spread of many animals used in particular trades (e.g., fur and pet trade; see Box 7.1). Humans also frequently help to disperse non-native organisms unintentionally (Box 8.1). This sort of movement is found in the dispersal of non-native plant species within agricultural products (Mack 1981; Mack & Erneberg 2002), seeds of non-native plants on people’s shoes or car tires (Lonsdale & Lane 1994), and the movement of aquatic species as bait and on the bottoms of recreational boats (Buchan & Padilla 1999; Leppäkoski & Olenin 2000). Undoubtedly, dispersal ability also strongly affects how quickly any organism spreads across the landscape. Individuals that can move long distances not only increase the average dispersal distance but also may lead to satellite populations, and these satellites greatly increase the rate of spread by creating new sources of colonists. Simulations that
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Box 8.1 If I had a boat Non-native species that live within flowing aquatic ecosystems present a unique problem when attempting to determine how, and to where, their populations will spread. Streams and rivers have discrete boundaries and connections (e.g., canals, confluences; Johnson & Carlton 1996). Thus, unlike terrestrial invaders, nonnative stream or river-dwelling organisms, if left to their own dispersal devices, will expand their ranges in a highly unidirectional manner. Individuals of such species disperse passively via water currents, or by attaching to something (or another species) that itself moves downstream in the watershed. Only very occasionally do individuals have the opportunity to disperse against the water current, and they very rarely disperse over land into adjacent, but unconnected, watersheds. Although there has been less focus in invasion biology on models of unidirectional spread, these models are simpler than models that describe the spread of terrestrial invaders (see Chapter 7). The interesting aspect of these models lies in their failure to predict spread because these failures force us to examine our assumptions about life history limitations and modes of dispersal. A well-documented case of an aquatic invader that has a perplexing pattern of range expansion is the zebra mussel (Dreissena polymorpha) within North America (Johnson & Carlton 1996). The zebra mussel, a native of Eurasian waters, first appeared in Lake St. Clair within the North American Great Lakes in the mid1980s. The founding individuals were thought to have been transported within ship ballast (see Box 2.1 for more on ship ballast). The zebra mussel population spread downstream and quickly came to occupy much of the Hudson and the St. Lawrence Rivers (Fig. 8.5). Because nearly all reaches of these watersheds contained zebra mussels by the mid-1990s, most spread was believed to have occurred through natural processes, which in this case is the passive movement of zebra mussel larvae via water currents. However, there were continuing reports of zebra mussels in watersheds very far from this initial point of entry such as locations far south of the Great Lakes within coastal Lousiana (near the Gulf of Mexico), and in isolated glacial lakes surrounding the Great Lakes (Fig. 8.5). There was no way zebra mussels could have colonized these places on their own power. Good ecological sleuthing by Ladd Johnson and his colleagues (Johnson & Carlton 1996; Johnson & Padilla 1996) established that zebra mussels were being transported in high numbers up, down, and between watersheds via recreational boaters, typically anglers. From surveys conducted at recreational boat launches around Michigan, USA, Johnson and colleagues identified seven different mechanisms by which boats could transport zebra mussels overland: (i) adults could attach to the exterior of a boat or a motor, (ii) adults could attach to anchors or to material snagged by the anchor, (iii) adults could attach to aquatic plants and algae that themselves became entangled on the boat trailer, (iv) larvae could passively become trapped in bilge water or in (v) engine cooling water, and (vi) larvae could get picked up within bait buckets or in (vii) live wells (Johnson & Carlton 1996; Johnson & Padilla 1996). Of these mechanisms, adults were most likely to be moved
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Figure 8.5 Map showing known distribution of the zebra mussel (Dreissena polymorpha) in North America as of December 31, 1993. Xs represent adult populations thought to be established by overland dispersal. Open circles represent sightings (larval or adult) during a detection program in Michigan. Black circles and thick lines represent other sightings in the Great Lakes and connected waterways. Reprinted with permission from Johnson, L.E. and J.T. Carlton. 1996. Post-establishment spread in large-scale invasions: dispersal mechanisms of the zebra mussel Dreissena polymorpha. Ecology 77: 1686–1690.
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via vegetation entangled on trailers, and larvae were most likely to be moved via water in live wells. The numbers of individual zebra mussels found on any one infested boat sometimes reached several thousand (Johnson & Carlton 1996; Johnson & Padilla 1996). Recreational boats clearly have the potential to move large numbers of zebra mussels even if the per boat likelihood of being “infested” with zebra mussels is quite low. Building models that incorporate both the natural and boat-facilitated dispersal of zebra mussels has been difficult since each mechanism operates at a particular spatial scale. However, Bossenbroek et al. (2001) and Leung et al. (2004) have made considerable headway using gravity and maximum likelihood models, respectively. These models also take into account the life history of the zebra mussel and the conditions it requires for positive population growth. For example, zebra mussels are dioecious with external fertilization. Thus, for zebra mussels to reproduce, males and females must be in close proximity so that sperm will “find” eggs relatively easily. This suggests that zebra mussels are prone to Allee effects whereby small numbers of individuals will not found a growing population as long as they are sufficiently spaced apart such that reproduction is low or nil (Leung et al. 2004). To some extent, transport of thousands of zebra mussels via one recreational boat ensures that this is unlikely to happen. However, the incorporation of Allee effects is critical to predicting the rate and location of geographical spread. On a much larger spatial scale, there is yet another conundrum to zebra mussel distributions in North America. Once the population hit about the 100o meridian across North America, its geographical spread slowed down tremendously. Was this just a strange twist of fate, or does the zebra mussel possess life history traits that limit its range to the relatively humid and wet portions of North America (i.e., everywhere east of the 100◦ meridian)? Drake and Bossenbroek (2004) utilized genetic algorithms for rule production (GARP; see Chapter 12 for a description) to determine the limits to zebra mussel distribution. Their work confirms that zebra mussels are unlikely to establish in western North America, except right along the Pacific coast. The more troublesome areas of future range expansion are within southeastern watersheds where there are hundreds of native endemic unionid mussels that are an attractive substrate for zebra mussels and are thus highly susceptible to negative effects from zebra mussels. The results of these modeling exercises provide a powerful set of tools for directing eradication and control efforts and should prove useful for many other aquatic invasive species.
contrast the effect of removing either the expanding edge of large populations or the more easily controlled small, nascent populations, generally find that control of these satellites is a much better way to manage the spread of invasive species (Moody & Mack 1988; Shigesada & Kawasaki 1997; Taylor & Hastings 2004). Allee effects can change this result, but the basic message is that small, new populations can be dangerous.
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Biotic interactions In Chapter 6 we discussed many of the ways in which biological interactions could either facilitate or impede the establishment of a non-native species. In that chapter we restricted ourselves to talking about interactions that moved the non-native over (or under) a population survival threshold. Here we focus not only on the biotic interactions that increase dispersal of non-native individuals but also those that increase (or decrease) population growth rates and thereby alter the rate of geographic spread of the invasion.
Predation Predators may reduce or increase the spread of invading organisms. Many predators act as dispersers and carry propagules to new locations. This is particularly true for seed predators that cache seeds, such as squirrels and birds, but it can also be true of predators who consume whole plants or seeds without necessarily damaging the viability of those seeds. Many seeds, in fact, require some sort of scarification such as that produced from passing through a gut or gizzard, before they will germinate. On the other hand, predation is usually a negative interaction that reduces population growth, thereby reducing the number of individuals available to disperse. In addition, dispersing individuals are often far more vulnerable to predation than individuals residing on an established territory. Predation can affect spread, therefore, not just by removing individuals but by removing the very individuals that would otherwise disperse. One of the most obvious examples of predation that reduces spread is biocontrol efforts. By reducing population growth, biocontrol agents reduce the spread of the species. Pathogenic or parasitoid biocontrol agents often move with their host, so that the biocontrol agent is spread to new host populations as the host itself spreads, thereby reducing the chance of establishment of these new populations. This sort of effect is seen with the pathogenic fungus, Entomophaga maimaiga, which attacks the gypsy moth, Lymantria dispar in moist, cool environments (Dwyer et al. 1998; Weseloh 2003). Because gypsy moths sometimes experience Allee effects (Liebhold & Bascompte 2003), reducing population size and growth in small, dispersing populations is an excellent way to reduce the spread of the species (Fig. 8.6) Many predators also disperse over longer distances than their prey, an adaptation that allows them to aggregate where their prey are abundant. This larger dispersal step may help predators, both native and non-native, reduce the spread of invasive species because the invader cannot outrun the predator. Better dispersal by predators also may stabilize the abundance of both the predator and prey and help to keep the predator species extant when local prey densities decrease (Murdoch & Briggs 1996; Briggs & Hoopes 2004; Hoopes et al. 2005). The flip side of this stabilization and persistence is that the invasive prey species also survives in low density patches not worthy of the predator’s attention. It is this dynamic that makes it close to impossible for a biocontrol agent to eradicate an invasive species. As described in Chapter 7, only a specialist attacking an invader with Allee dynamics at low densities could reduce the range of the invader (Fagan et al. 2002). It is this theoretical ideal that motivates biocontrol introductions.
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Figure 8.6 Relationship between spread rate of gypsy moths and original site of invasion in 1900. Each circle represents one county or census district. Filled circles represent counties with a mean minimum temperature >7◦ C, and open circles represent counties where mean minimum temperature <7◦ C. Note that spread rate is lower with lower mean temperatures. Reprinted from Liebhold, A.M., et al. 1992. Gypsy moth invasion in North America: a quantitative analysis. Journal of Biogeography 19(5): 513–520, with permission from Blackwell Publishing.
Remember from the establishment chapters (Chapter 6) that we came across several examples where predators restricted the areas in which non-natives could establish. This exclusion from certain habitats can have a strong effect on spread by reducing the available habitat across which individuals are moving. When a predator keeps a non-native out of one habitat, it reduces the source habitats for further invasion. This is somewhat like a landscape level Allee effect in which isolated populations cannot find each other, and, therefore, cannot meld so that the geographical range does not grow as fast as it would otherwise. Many of our examples in this section rely on biocontrol, which often involves a non-native species that is introduced to control an invader. Native predators can also restrict the spread of invaders, by restricting establishment in new locations. See Chapter 6 for more discussion of such resistance or see Box 4.2 and Box 6.2 for more on biocontrol. Finally, remember that a predator may have positive effects on its prey if it also kills the prey’s main competitors. This scenario may have distressing consequences if the predator is a biocontrol agent and the main competitors are native species, particularly rare native species. In the midwest United States, a weevil, Rhinocyllus conicus, was introduced as a biocontrol agent for the control of non-native populations of musk and Canada thistle (Carduus nutans and Cirsium arvense). In the sand dunes of Nebraska, the weevil crossed over to preying on several native thistles, including the uncommon Platte thistle, Cirsium canescens, which subsequently declined significantly (Louda et al. 1997; Rose et al. 2005). Because thistles use similar habitats and resources, the decline of the native thistles may increase habitat and resource availability for the invasive thistles.
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In addition, the presence of R. conicus on Platte thistle (C. canescens) led researchers to worry about potential impacts on the closely related Pitcher’s thistle (Cirsium pitcheri), which is federally listed as threatened in the US and endangered in Canada. Lab and garden trials have demonstrated that the weevil performs similarly on C. canescens and C. pitcheri and that the weevil’s active period overlaps significantly with flowering in the rare C. pitcheri (Louda et al. 2005). At the time of publication, the weevil had not yet dispersed to natural populations of the rare thistle, but the weevil’s introduction has increased rather than decreased ecologists’ concern about the spread of invasive species.
Competition Competitors can also slow the rate of range expansion. By reducing access to resources, competitors slow population growth rates and often particularly reduce reproduction amongst their rivals. Because juveniles are frequently the stage most likely to disperse (e.g., seeds, planktonic larvae, and young birds), reduced reproduction, even if it does not come with a strong reduction in population size or growth, can lead to slower rates of geographical range expansion. Competitive effects can be hard to document, particularly for mobile organisms because it is difficult to disentangle the effects of competitors from the effects of a particular environment or location. Most published competition results, therefore, are from plant species where the presence of competitors definitely reduces the germination, growth, survival, or reproduction of non-natives. All of these factors contribute to reductions in spread rates. In Chapter 6, we discussed examples in which the presence of competitors excluded non-native species from establishing at a site. We gave examples in which the presence of a canopy combined with competition for light excluded pine species from intact forest on Isla Victoria in Argentina (Simberloff et al. 2002) or excluded vines from forest on Christmas Island (Green et al. 2004). These examples are unusual because invaders are more likely to slowdown the rate of spread than to block establishment. Disturbances that remove vegetation often allow invasive species to enter habitats and then spread from a single entry point. For example, Kama Almasi (2000) examined the spread of Tibouchina herbacea, a small South American shrub invading the Hawaiian Islands. On the Big Island of Hawaii she found that T. herbacea did very well in open, disturbed habitats but could establish in almost any sort of habitat, including intact forest (Fig. 8.7). However, the plant’s growth, reproduction, and dispersal were all reduced in intact forests, so that forest communities slowed spread the most (Almasi 2000). Competition need not always be with native species. Hoopes and Hall (2002) examined plant growth and reproduction in a California grassland and found that the abundance of Hordeum marinum gussoneanum, an annual foxtail grass from the Mediterranean region of Europe, reduced the growth, survival, reproduction, and spread of Bromus diandrus, a second Mediterranean annual grass species. Similarly, Holway and Case (2001) found non-native ant species competed with each other, reducing foraging success and survival as well as the creation of new colonies for each species. And, finally, introduced crayfish in Wisconsin lakes responded not only to the presence of native crayfish but also to the presence of other non-native crayfish (Garvey et al. 2003).
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0.08
Germinated (%)
0.07 0.06 0.05 0.04 0.03 0.02 0.01 0 Open
Canopy only
Ground cover only
Closed
Habitat
Figure 8.7 Performance of Tibouchina herbacea on the island of Hawaii. This non-native species can germinate in all habitats, including closed intact forest. Dark bars represent information from intact forest in 1995, grey bars from intact forest in 1996, and open bars from open forest in 1996. However, it does not do as well in intact forest as in other sites and this slows the non-native’s spread across the island. Reprinted from Almasi, K. N. 2000. A non-native perennial invades a native forest. Biological Invasions 2: 219–230, with kind permission from Springer Science and Business Media.
Mutualism Recall from Chapter 6 that mutualisms are associations between individuals of two species where both participants benefit in the form of increased survival, reproduction, or dispersal. Mutualisms can be established between two non-native species, or between natives and non-natives. No matter the players, the establishment of a mutualism can speed the rate of geographical range expansion for the non-native through many mechanisms. When a mutualist increases the abundance of any life stage of the non-native, they increase the number of individuals available to disperse. This effect is most pronounced, however, when a mutualist increases the production of juveniles who are more likely to disperse. Mutualists may also make nutrients more available to non-native individuals in certain environments, thus allowing the establishment of satellite populations in otherwise uninhabitable areas and assisting non-native individuals in traversing hostile environments. Most importantly for our current topic, though, mutualists may actively spread invasive organisms by assisting in their dispersal, particularly for plant invaders. Richardson et al. (2000a) reviewed the role of vectored seed dispersal after invasion. Their results suggest that most non-native plants are not limited in their initial establishment by a lack of seed dispersal. Sometimes seeds are dispersed by native species (bats, birds, or ants), and sometimes it appears non-native animals are mostly
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responsible. The availability of dispersal mutualists may reflect the lack of tightly coevolved seed dispersal relationships in general, or the results of this first review may be overturned with more detailed study of such facilitative associations. There are several examples, however, of invasive species that were naturalized – established but not spreading and with little impact – until the introduction of a particularly effective nonnative disperser. For example, the red whiskered bulbul (Pycnonotus jocosus), native to India, disperses the seeds of several invasive plants on islands in the Indian Ocean (Macdonald et al. 1991). In Florida, where the bulbul is also established, there are over 24 non-native species whose fruit it eats and whose seeds it spreads (Simberloff & Von Holle 1999). These include the invasive Brazilian pepper (Schinus terebinthifolius) and several vine species (Simberloff & Von Holle 1999). Similarly, the Japanese white-eye (Zosterops japonicus) has had a singularly devastating effect with its dispersal of Morella faya, the Hawaiian fire tree, a nitrogen-fixing tree that is changing plant community structure and ecosystem processes across the Hawaiian Islands (Woodward et al. 1990). In general, non-native plants that produce fleshy fruit, particularly late hanging berries, tend to experience high frugivory and, therefore, high dispersal rates and fast spread. Japanese barberry (Berberis thunbergii) and European privet (Ligustrum vulgare) are excellent examples of such non-natives on the east coast of the United States. Although pollination mutualisms may be somewhat more specialized, the majority of invasive species are pollinated by generalist pollinators. The most common generalist pollinators in North America are the common honey bee (Apis mellifera) and several native and European bumble bee species (Bombus spp.). Non-native bees have been spread by humans looking for reliable pollination services for orchards and other agricultural uses, but they also have managed to spread on their own largely because of their dietary generalism. In addition, they provide the majority of the pollination for many non-native species such as yellow starthistle (Centaurea solstitialis) and purple loosestrife (Lythrum salicaria) in North America (Simberloff & Von Holle 1999). Although sharing generalist pollinators could create competition among non-native species for limiting pollinators, it seems that pollen limitation is not severely exacerbated by the presence of other non-native plant species (Richardson et al. 2000a; Parker & Haubensak 2002). Instead a larger pool of flowering plants pulls in more pollinators so that showy plants may actually facilitate pollination for their neighbors (Groom 1998). Finally all of these biotic interactions can interact, as we saw in Chapter 6. The presence of a predator can change the outcome of competition between non-native plant species and make it so that both can invade an area. Similarly, the predator may spread the seeds of both species while removing the fruiting bodies before seed set on a native species. All of these outcomes depend on very specific interactions and timing, but such synergistic effects can increase rates of establishment, spread, and impacts on native species (Box 8.2).
The role of heterogeneity Both temporal and spatial heterogeneity can have large effects on the spread of invasive organisms. As we have noted repeatedly in the previous sections of this chapter, as well
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Box 8.2 Melt with you In keeping with the predominant themes in ecology over the past half century, most invasion ecologists have focused on negative interactions, such as competition and predation, between non-native and native species. This makes a certain amount of practical sense because interest in non-native species is, in large measure, driven by concern over the possibility that non-natives will drive native species extinct. This focus has led to two dominant themes in invasion ecology. First, native species should be able to repel non-natives so long as competition for resources is strong and native species richness is high (i.e., biotic resistance hypothesis; see Box 6.1). Second, non-native species may often drive natives toward extinction by competing with, or eating, native species. In the late 1980s, as the study of positive species interactions (i.e., mutualism, facilitation, commensalism) began to emerge as a relevant ecological course of study (Bronstein 1994), invasion biologists also began to consider the role of these interactions in invasions. Simberloff and Von Holle (1999) were the first to comprehensively discuss the influence of positive interactions in biological invasions. Their survey of the invasion ecology literature illustrated that positive biological interactions between non-natives and other species (native and non-native) were at least as common as competitive interactions. They further suggested that positive interactions could lead to two distinct patterns: (i) non-native species could facilitate each other’s establishment or geographical spread either directly or indirectly, and thereby produce an accelerating rate of invasion for any one site, and (ii) the cumulative impact of several positively interacting non-natives on native species may be larger than each non-native species’ individual impact. Simberloff and Von Holle (1999) suggested that these types of interactions should be important to practical-minded invasion ecologists because these interactions could create “invasional meltdowns” in which an entire native ecosystem could collapse under the weight of a growing, biologically integrated set of invaders. Since the publication of this paper, several more detailed examples of meltdowns or components to the meltdown scenario have emerged. For example, Ricciardi (2001) showed that the rate of accumulation of non-native aquatic species in the North American Great Lakes has accelerated through time. Since this pattern could also be produced by an acceleration in the transport and release of non-natives, Ricciardi systematically surveyed the literature categorizing the types of interactions observed between non-native species. Like Simberloff and Von Holle (1999), Ricciardi (2001) found direct positive interactions (facilitation) more common than direct negative interactions (competition), thus suggesting that at least some of the acceleration in invasion rates is due to facilitation between non-natives. Illustrating that indirect interactions can have just as profound an impact on the success of a non-native, Grosholz (2005) found that in Bodega Bay, California, the invasion of European green crab (Carcinus maenas) has led to the rapid range expansion of a previously innocuous clam (Gemma gemma) by preferentially preying on its major competitor, the native Nutricola spp. clams.
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Finally, Floerl et al. (2004) detail how positive biotic interactions can lead to increased numbers of non-natives transported on ship hulls. The bryozoan, Watersipora subtorquata, is tolerant of several antifouling paints commonly used on boat hulls. These paints keep boat hulls clean of encrusting organisms so they move more efficiently in the water, and since most fouling species are non-natives, they reduce the likelihood that a boater will inadvertently spread a non-native from one water body to another. Since W. subtorquata forms colonies, Floerl et al. (2004) suspected its presence provided a refuge for other non-native organisms that could not tolerate the antifouling paint. They found that twenty-two taxa were found exclusively on W. subtorquata colonies that themselves had colonized a painted experimental tile. Some of these non-native species were 248 times more abundant on W. subtorquata than on surfaces not colonized by the bryozoan. Thus, this one toxin-resistant non-native fouling organism serves as a foundational species for a whole community of non-natives that adhere to a ship’s hull. This translates into a many-fold increase in the number of non-native species W. subtorquata-infested boats may transport into new harbors.
as in much of Chapter 7, heterogeneity in dispersal distances can slow or speed the rate of spread in models of range expansion. We have also noted that these models serve best as a kind of “experiment” than as accurate tools that predict how fast invasive species will spread or where they will arrive next. This lack of accuracy in spread models comes primarily from the way that organisms react to variation in the landscape in both biotic and abiotic factors as well as temporal variation in factors that speed or impede the movement of organisms across a hostile landscape. It is the reaction to this spatial and temporal variation that leads to the distribution of dispersal distances and to biases in the direction of spread. Many of these sources of variation have been examined in models. Table 8.1 follows Hastings et al. (2004) to summarize the resultant spread dynamics from certain types of variation incorporated in spread models. To think about heterogeneity without the framework of models, simply think of a landscape with patches that support a variety of different plant and animal species. This sort of spatial heterogeneity will lead to variation in the presence of biotic facilitators or inhibitors of non-native species spread. As a non-native species encounters a habitat boundary, it comes in contact with a new species assemblage that may impede or assist dispersal in all the ways listed above. What we did not mention above, though, was the underlying abiotic heterogeneity that may assist this transition. Certain habitats are more amenable to invasion, especially if they have experienced significant disturbance or alterations to disturbance regimes (Chapter 5). For example, in grasslands in the Central Valley of California Hoopes and Hall (2002) found that populations of the Mediterranean annual grass, B. diandrus, could expand swiftly across high nutrient environments regardless of competitors. This species grows poorly in low nutrient or saline environments, though, and the species does not spread into these areas (Hoopes
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Table 8.1 The resultant spread dynamics from certain types of variation incorporated in spread models, following Hastings et al. (2004). Source of heterogeneity
Spread dynamics result
Key references
Deterministic diffusion without density-dependence
Linear rate of spread
Skellam 1951
Deterministic diffusion with density-dependence
Linear rate of spread under most conditions
Fisher 1937
Stochastic diffusion
Same spread rates as deterministic diffusion
Mollison 1977
Stage structured reproduction or dispersal
Generally slows spread rate
van den Bosch et al. 1992
Allee effects in population growth rates
Increased spread rates as the population passes the critical threshold density
Lewis & Kareiva 1993
Stratified or jump dispersal (i.e., occasional long-distance dispersal)
Jump dispersal increases spread rate
Shigesada et al. 1995
Discrete time with variable, deterministic distributions of dispersal distance and reproductive mechanisms
Fat-tailed dispersal kernels lead to accelerating rates of spread
Kot et al. 1996
Competition
Nonlinear, often slows rates of spread
Hosono 1998
Stochastic reproduction and dispersal distances
Finite expectations for the position of the wave front based on furthest dispersal and reduced spread rates for variable reproduction versus constant reproduction
Clark et al. 2001
Variable dispersal and reproductive kernels
Conditions for linear and nonlinear rates of spread
Weinberger et al. 2002
& Hall 2002). Similarly, Marchetti et al. (2004b) found that the best indicators of spread for non-native fish were number of propagules, as we would expect, and physiological tolerance. The same sort of result was found with Argentine ants (Linepithema humile) whose movement was much faster across moist, cool environments than across dry, hot ones in California (Holway et al. 2002b). Although these conditions are the exact opposite conditions in which native ants thrive, the Argentine ant was definitely reacting to abiotic rather than biotic conditions because Argentine ants exhibited no change in activity in the presence of native ants in either environment. Predicting rates of spread gets extremely difficult when multiple sources of variation change dispersal distances, population growth, and population interactions. Because
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many of these factors can interact, there are also unexpected non-linearities and indirect effects that can speed or slow the geographic expansion of a non-native species across a new environment. The next two sections will detail two such unexpected patterns in geographic spread; lag times and boom and bust dynamics.
Lag times One of the main sources of confusion in measuring the spread of invasive species across a landscape comes from the frequently long periods of time between when a non-native population becomes established and when it begins to rapidly expand its range. The time from initial establishment to subsequent spread is called the “lag time.” Sometimes lag times are very short, lasting only a couple of generations; just enough time to be noticeable. There are other examples of lag times that span 100 or more years. Kowarik (1995) surveyed the woody invasive plants of Germany, computing lag times for 184 of these species. He found an average lag time of 131 years for woody shrubs and 170 years for trees. If non-native species find conditions in their new range suitable enough to establish a self-sustaining population, why then do they often take so long to begin spreading? These lag times may have multiple sources, many of which tie into the ideas that we have outlined above or venture into in Chapter 11. Here, we will briefly review the possible ecological mechanisms producing lag times, and we have divided the sections below along the same themes for which we have already given you some background. Note, however, that ecologists really do not have a good feel for which of these mechanisms is most important in the production of lag times, nor do we really have a good handle on how common lag times are in real systems. This is a dilemma since, as you will read, lag times can lure us into thinking that a non-native species is innocuous (or eradicated) when in reality it is well on its way to becoming a significant problem.
Population growth There are three potential sources for lag times within the process of population growth: (i) the nature of population growth trajectories, (ii) thresholds to detection, and (iii) Allee effects. The first two explanations share many factors but are not completely overlapping. The first and simplest explanation for lag times is an effect of numbers of individuals and the way they accumulate in a geometrically or exponentially growing population. Even large rates of population growth if expressed within very small populations will lead to very small absolute changes in population size from one year to the next. These same large rates of growth will eventually lead to very large absolute changes in population sizes, but this switch in magnitude may take a long time, and the non-native species may not spread until the population becomes large enough to tax available resources. Second, between the time of establishment of a non-native species and its subsequent spread there may be a period during which the individuals of this species simply are
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Threshold to detection
Time since invasion
Figure 8.8 Illustration of how geometric/exponential population growth can lead to a lag time. The time between arrival and the surpassing of the detection threshold is the lag. Notice that a constant per capita rate of increase may take a long time to reach noticeable proportions (threshold population size) and then appears to explode. The dynamics are the same at all points, but the abundance of organisms makes per capita growth much more noticeable after many years.
not obvious to resource managers and ecologists (see Medfly example above). The population size above which we regularly encounter and detect non-native individuals is called the “detection threshold” (Fig. 8.8). The existence of the detection threshold may also be due to the cryptic nature of the dispersing stage. Cryptic dispersing stages are particularly common in animal-dispersed species, such as seeds in animal hides and parasitoid eggs in insect hosts. Trees on a landscape are much more noticeable than seeds in the soil, and emerged parasitoids are much more noticeable than larvae hidden inside the body of another insect. There is not necessarily an ecological change that occurs between this cryptic stage of dispersal and eventual geographic spread, but the population growth at low densities is less apparent than at high densities even if the growth rate is the same. The third potential source of a lag time is an Allee effect at low densities. We discussed Allee effects in establishment in Chapter 4, particularly in the role of a critical threshold population density above which a population grows and below which the population shrinks. For an introduction to succeed, the initial population size must be above this critical density. Allee effects do not have to be this strong, though. They are simply a reduction in per capita population growth with reduced population density. There is no reason why Allee effects must be so strong that per capita population growth is actually negative at low densities (Fig 8.2). Weak Allee effects may allow population growth at low densities with much faster population growth at higher densities. These weak effects slow the rate of spread of invasive species, but also contribute to lag times in spread. As we mentioned earlier in this chapter, Veit and Lewis (1996) found that the sudden spread of the house finch in North America could be partially explained by an Allee effect. Similarly, the initial slow spread of Spartina alterniflora on the west coast
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of North America may be partially explained by reduced pollination at low densities, leading to low seed set (Davis et al. 2004; Taylor et al. 2004). Once the populations reached higher densities, both population growth and spread increased. For S. alterniflora that story is somewhat more complicated due to clonal growth, genetics, and herbivores, but Allee effects do play enough of a role to make a distinct difference in optimal management strategies (Taylor & Hastings 2004). Taylor and Hastings (2005) point out that a weak Allee effect can be very hard to detect, making it very difficult to predict which non-native species may suddenly explode across a landscape once they reach higher densities. If the Allee effect interacts with environmental conditions, then altered rates of population growth on either side of environmental boundaries may keep non-native populations confined to a small range in the better habitat until a large enough propagule is introduced into the lower-growth environment (Keitt et al. 2001). Such “range pinning” may also lead to lag times in the spread of invasive species.
Dispersal In a slightly counterintuitive way, the early stages of geographical range expansion make detection of spread more difficult. When a non-native population establishes itself, it begins to accumulate individuals through successful reproduction. After the accumulation of a certain number of individuals, the detection threshold may be crossed and we humans are suddenly aware that the non-native population exists. If dispersal is related to density (e.g., territorial birds), the population will begin to “lose” individuals to dispersal only when a particular density is attained. For a short while, dispersal may remove enough individuals from the source population to bring it back down below the detection threshold (Fig. 8.9). This increase, and then decrease, in population size may make it appear to us that the source population is unable to produce enough offspring to maintain itself (Kean & Barlow 2000). This is an illusion that makes the subsequent wave of invasion more surprising. The drain on the population in the initial phases of population spread may also keep the source population balanced on the edge of an Allee effect, which will serve to further lengthen the lag period.
Biotic interactions After years of stagnant population growth, non-native populations will begin experiencing significant geographic range expansion when interactions with other species favor high population growth or long-distance dispersal. Conceptually the most straight forward way for this to happen is through the establishment of mutualisms. Figs and fig wasps provide an excellent example of this. Recall that fig trees can survive for extremely long periods (decades to centuries), thereby holding open a window for invasion if an appropriate propagule of fig wasps appears. The arrival of the fig wasp in Florida allowed a sudden shift in fig trees from being non-reproductive to reproductive, and from showing no spatial population expansion to becoming widespread, within
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Dispersal events Population growth
Threshold to detection Time since invasion
Figure 8.9 Dispersal can bring an incipient population back down below the detection threshold, producing a lag time. This can happen multiple times, as shown. Similar dips beneath thresholds for Allee effects may stop or slow population growth while dips beneath the detection threshold will only affect our ability to see the populations.
the span of a few years. This same mechanism can operate with dispersal mutualisms. Although the shift will not be as extreme (from no reproduction to population spread), the change to the landscape may be equivalent. That is, a stable local population of adults will suddenly shift into a wide spread invasion. There are many non-native plant species that have spread in this manner after years without population growth (e.g., Japanese barberry [B. thunbergii], privets [L. vulgare and other species of the same genus], and burning bush [Euonymus spp.]). These species seem innocuous and hence are commonly planted as ornamentals. Once the dispersal mutualism is established, however, these innocuous species begin to “jump the fence” and grow well outside of cultivated environs.
Spatial and temporal heterogeneity Propagules of non-native species are released and must survive where the transport vector leaves them. These locations may provide sufficient biotic and abiotic conditions for the population to become self-sustaining; however, it may not be optimal habitat. In optimal habitat, the population will grow in numbers and expand through space very quickly. In less optimal habitat, the population grows and expands slowly. If these two types of habitat are located nearby one another, the non-native population will slowly spread across sub-optimal habitat until it hits optimal ground. At this point, the species begins to rapidly expand its range. We see this dynamic very clearly with range-pinning and an Allee effect, but less extreme examples are also possible. A similar scenario can occur with changes in habitat suitability through time. Often populations grow more slowly during periods of unfavorable weather (e.g., frost, heavy rainfall, and drought). When these conditions change, the population may begin to substantially expand its range. Temporal variations in weather conditions may interact with other mechanisms described above (e.g., Allee effects) such that a series of good
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years may push the population size above a critical threshold beyond which range expansion is consistently fast.
Evolution and genetic augmentation Ellstrand and Schierenbeck (2000) suggest that lag times can result from either postestablishment evolution or increases in genetic heterozygosity. We will explore the interplay between invasion dynamics and these factors in depth in Chapter 11. The gist of these arguments for lag times, however, is that non-native species experience a fundamental shift in their ability to cope with conditions in the novel range, and this shift enables range expansion. Non-native populations may undergo microevolutionary shifts in their morphology or physiology in response to strong selection pressures such as interactions with novel parasites, herbivores, or diseases. These negative interactions keep the non-native population size low for long periods; however, they do not drive the population extinct (see Chapter 6). Once the evolutionary shift occurs, the population is released from these constraints and may begin to expand its range accordingly. A similar role can be envisioned for the effects of genetic variability. Often resistance to disease, parasitism, or herbivory is conferred by a suite of genes, or occasionally by one rare gene. If these genes are not present in the non-native population, or if the gene complex is disrupted through genetic drift (see Chapter 11), this population will be susceptible to the ill-effects of negative biotic interactions. If the original invasion vector is still intact, new individuals may arrive at a later date and join the earlier arriving individuals. These late arrivals may bring with them much needed genetic diversity, including the gene(s) necessary to insulate the population from harmful biotic interactions. As soon as the resistance gene(s) are re-introduced into the population, most individuals are capable of mitigating the negative effects of herbivores, parasites, and the like. Henceforth, the population may expand its range at a relatively high rate.
Boom and bust Finally we tackle the subject of boom and bust population dynamics among non-native species. As the name implies, sometimes non-native populations will increase dramatically in their numbers and geographic extent after initial establishment. And then, just as dramatically, these populations decline in numbers and extent although not often down to extinction. Williamson (1996) suggests that boom and bust dynamics are a relatively common feature of biological invasions. When Simberloff and Gibbons (2004) endeavored to collate examples of this phenomenon, however, they ran into a wall of ignorance. These authors could not find any published articles that explicitly detailed boom and bust dynamics for a non-native population. They had to beat the academic bushes to retrieve seventeen examples that they could analyze at all. The examples they did find spanned across a wide range of species, including insects, mammals, birds, plants, and fishes. Their examples also spanned a wide spatial scale, including a boom and bust dynamic across Great Britain (Elodea canadensis), within
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the greater metropolitan area of Vancouver, Canada (Acridotheres cristatellus), and on several Pacific islands (Achatina fulica). When Simberloff and Gibbons (2004) attempted to collate the ecological reasons for these boom and bust scenarios they found that the reasons for the cycles were often mysterious. They could find no reason for the population crash for ten of the 17 examples they detailed. Of the seven for which they could assign at least partial explanations, competition with another non-native species played a role in four of the population crashes. The other reasons given included exhaustion of resources, parasitism by a subsequently introduced non-native species, and adaptation by a native herbivore so that it could feed on the non-native in question. None of the examples really nailed-down an ultimate cause for the population crash. Even the seven best examples, therefore, are still open to further investigation. Boom and bust dynamics may be quite common, but the ecological mechanisms definitely are not clear. Basic ecological theory may offer the most complete (but still unsatisying) explanation. A basic concept of population growth and life history models is that non-equilibrium dynamics are somewhat common for species with rapid growth rates and high dispersal. If we think about basic r- and K -strategies for organisms, we can imagine that non-native species that establish in a new area may be r-strategists that are well-adapted to moving and colonizing new habitats but not to maintaining populations in highly competitive environments. This set of adaptations may lead to boom and bust cycles, but it does not offer us specific mechanisms. More research is definitely necessary.
Summary The effects of population growth, dispersal, biotic interactions, and spatial and temporal heterogeneity all ensure that the models we build to explain the spread of non-native species will forever be rough approximations of reality. In fact, in practice we frequently find it very difficult to detect the source of particular limits to spread or sudden increases in spread. For example, it can be very difficult to distinguish between the roles of biotic facilitators and weak Allee effects that have been surpassed. We clearly need better studies of dispersal dynamics, even though these efforts are sure to be difficult. As suggested in Chapter 6, the influence of biotic interactions on the rate and extent of geographical spread is quite large. Even weak competitive and predatory forces can substantially slow the rate of range expansion for non-native species. This inhibitory force can come from native species and non-native species that previously established, or that are released on purpose to control the invader. The establishment of mutualisms can also greatly speed the rate of non-native species invasion. Indeed, there are so many possibilities for interacting influences on rates of spread that we often see non-native populations exhibiting strange population dynamics. We explored two of these sets of dynamics, lags and boom–bust cycles, and found good examples of the pattern but so many potential mechanisms that generalizations are currently difficult. Given the multitude of ways in which this stage of the invasion process interacts with the others (i.e., transport and impact), future study of these mechanisms should prove enormously enlightening.
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Further complexities Hastings, A., K. Cuddington, K. F. Davies, C. J. Dugaw, S. Elmendorf, A. Freestone, S. Harrison, M. Holland, J. Lambrinos, U. Malvadkar, B. A. Melbourne, K. Moore, C. Taylor, and D. Thomson. 2004. The spatial spread of invasions: new developments in theory and evidence. Ecology Letters 8: 91–101. Kot, M., M. A. Lewis, and P. van den Driessche. 1996. Dispersal data and the spread of invading organisms. Ecology 77: 2027–2042. Mack, R. N. 1981. Invasion of Bromus tectorum L into western North-America: an ecological chronicle. Agro-Ecosystems 7: 145–165. Murdoch, W. W., and C. J. Briggs. 1996. Theory for biological control: recent developments. Ecology 77: 2001–2013. Shigesada, N., and K. Kawasaki. 1997. Biological invasions: theory and practice. Oxford University Press, Oxford, UK. Weinberger, H. F., M. A. Lewis, and B. T. Li. 2002. Analysis of linear determinacy for spread in cooperative models. Journal of Mathematical Biology 45: 183–218.
Companion papers Andow, D., P. Kareiva, S. Levin, and A. Okubo. 1990. Spread of invading organisms. Landscape Ecology 4: 177–188. Carey, J. R. 1996. The incipient Mediterranean fruit fly population in California: implications for invasion biology. Ecology 77: 1690–1696. Grosholz, E. D. 1996. Contrasting rates of spread for introduced species in terrestrial and marine systems. Ecology 77: 1680–1686. Liebhold, A., and J. Bascompte. 2003. The Allee effect, stochastic dynamics and the eradication of alien species. Ecology Letters 6: 133–140. Louda, S. M., A. E. Arnett, T. A. Rand, and F. L. Russell. 2003. Invasiveness of some biological control insects and adequacy of their ecological risk assessment and regulation. Conservation Biology 17: 73–82. Veit, R. R., and M. A. Lewis. 1996. Dispersal, population growth, and the Allee effect: dynamics of the house finch invasion of eastern North America. American Naturalist 148: 255–274.
9 Ecological Impacts of Invasive Species
Overview In this section we consider the range of ecological impacts that non-native species have on native species. In accordance with our invasion process model, impacts will generally only result from species that have moved through all previous stages. Although there may be some effects of non-successful introductions (Case 1995), such release impacts are often due to anthropogenic disturbances more than specific interactions between the non-native species and the native biota. Given the variety of ways that a non-native species could affect native organisms, this chapter endeavors simply to review the range of impacts already documented rather than to provide an overall synthesis of where, when, and how impacts occur. We emphasize that this review surely does not capture all possible ecological impacts; there are many conceivable mechanisms of impact that have yet to be fully explored. We have divided this range of impacts according to biological level of organization, but remember that, as in all ecological fields, what affects one level of organization will often affect others. Thus we begin at the gene level and go through individuals, populations, and communities to ecosystems and effects on a global scale.
Genetic impacts The impacts of invasive species are most insidious when they affect the genetic integrity of native species. These impacts are difficult to document as they require specialized skills (e.g., molecular genetic techniques) or difficult to gather data (e.g., genetic material). This may explain why SCOPE participants, or classic authors such as Elton, left discussion of genetic impacts of invasive species untouched. Despite this inattention, when such impacts are documented, they are often profound (Rhymer & Simberloff 1996). Genetic effects of invasive species are defined as alterations to the gene pools of native species usually through hybridization or introgression (Krueger & May 1991). Hybridization is the interbreeding of individuals from distinct species (or occasionally from genetically distinct populations). The young produced through hybridization
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may or may not be sterile. If they are sterile, then they waste native gametes and reduce native reproduction. If they are not sterile, then they may compete with the native parent species and reduce its survival or reproduction. Non-sterile hybrids also may backcross, leading to introgression, which essentially produces gene flow between species. The impact of most concern comes from a combination of hybrid vigor and preferential backcrossing of the hybrid young with the invasive parent species. In this instance, not only does the hybrid have increased fitness over the native (i.e., vigor), but the slow trickle of reproduction that the native does manage is increasingly tainted by genes from the invader. This can lead to the loss of genetic integrity and the extinction of the native genotype one gene at time. There are several examples of hybridization between invasive and native populations. A well-documented example is that of the mallard duck (Anas platyrhynchos) and several of its narrowly distributed congeneric relatives. The mallard has a wide native distribution, which includes much of the Northern Hemisphere. Because the mallard is a traditional food item for humans and is considered an excellent game species, it has also been purposefully transported and released into many other parts of the world including New Zealand, Hawaii, and Australia. Within these non-native localities the mallard comes into contact with endemic congeneric ducks such as New Zealand’s grey duck (A. superciliosa superciliosa), the Australian black duck (A. superciliosa rogersi), and the Hawaiian duck (A. wyvillana). The native populations and the introduced mallards interbreed, producing fertile young. These F1 (i.e. first generation) hybrid young mate with either of the parental populations, but there are so many mallards compared to the natives in these locations that the hybrid offspring are most likely to mate with mallards (Rhymer et al. 1994). For already rare endemics like the grey duck and Hawaiian duck, this decreased reproduction combined with the slow infusion of mallard genes has led many to wonder about the long-term persistence of these taxa as distinct species. Among fishes, the rainbow trout (Oncorhynchus mykiss), a salmonid native to North America, shows similar impacts on its congeners. Rainbow trout have been purposefully introduced as game species to every continent excluding Antartica (MacCrimmon 1971). They have also been widely released outside of their native drainage basins within North America. As with the mallard, the widespread introduction of the rainbow trout has allowed for several of its populations to come into contact with narrowly distributed, endemic congeners. For example, the rainbow trout hybridizes with California golden trout (O. aguabonita aguabonita) and Paiute cutthroat trout (O. clarki seleniris). Each of these species is endemic to specific watersheds within California. The Paiute cutthroat trout has legal status in the United States as a threatened species, and the California golden trout has been under petition for listing on the United States endangered species list since 2000. The hybrid offspring produced by the rainbow trout and the native congeneric trout are fertile and will mate with either parental population, leading to substantial introgression. This same pattern is repeated within other isolated freshwaters of the southwestern United States where the rainbow trout hybridizes with the Apache trout (O. apache) and Gila trout (O. gilae), but in these cases the F1 hybrids preferentially mate with the non-native parental species. Dowling and Childs (1992) indicate that of all extant Apache trout populations, 65% of them contain individuals with at least some rainbow trout alleles. However, there have been no rainbow trout
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individuals found with Apache alleles suggesting introgression is not an issue for the non-native species’ genetic integrity. Even in cases where introgression is not a problem, hybridization can represent a threat to the native population whether the hybrid offspring have reduced or enhanced fitness. As an example of the former, invasive brook trout (Salvelinus fontinalis) populations are displacing native bull trout (S. confluentus) across the native species’ range in the northwestern section of North America. Using mitochondrial DNA evidence, Leary et al. (1993) found that males from each species will mate with females of the other species. However, the hybrid F1 offspring from these matings apparently have very low fitness, as there is little evidence that these hybrids breed at all. This may be the result of their being sterile, having low survivorship, or poor mating success (Rhymer & Simberloff 1996). Low hybrid survivorship may not be a problem for the numerous invasive brook trout; however “wasting” gametes on unsuccessful hybrid young may prove detrimental to the rare native bull trout in the future. One of the best-known examples of the opposite effect, whereby the F1 hybrid offspring enjoy a fitness advantage over both the parental populations, involves hybridization between the invasive smooth cordgrass (Spartina alterniflora) and the native California cordgrass (S. foliosa). S. alterniflora is native to the eastern seaboard of North America and for decades has been an instrumental component of saltmarsh reclamation programs and erosion control efforts. Because of this success, S. alterniflora was introduced into saltmarshes of San Francisco Bay in the 1970s. This brought S. alterniflora into contact with its endemic congener S. foliosa, and it quickly became apparent that they could interbreed and produce hybrids (Daehler & Strong 1997). Because males of the invasive S. alterniflora have a staggering 28-fold increase in reproductive output over their native counterparts, there is substantial genetic introgression that threatens the native species (Anttila et al. 1998). To make matters worse, crosses between these species produce “super hybrids” that have greater growth rates and higher reproductive vigor than either of the parental species (Fig. 9.1; Ayers et al. 2004). Further, the hybrids are capable of withstanding a broader array of salinity levels than either parent, giving them a broader potential niche. As a result, the normally open mudflats of San Francisco Bay are being rapidly colonized by S. alterniflora × S. foliosa hybrids (as well as by “pure” S. alterniflora), which then spread from these colonization foci via vegetative reproduction (Ayers et al. 2004). The loss of open mudflats has cascading effects on the physical dynamics of these marshes and thus the rest of the native foodweb (e.g., the elimination of shorebird habitat).
Individual impacts Individual-level impacts alter the fitness or traits of individuals and are frequently easy to measure and to extrapolate to higher-level impacts, as we will discuss later in the chapter. Obvious examples of individual-level impacts are changes in the morphology, behavior, or demographic rates of natives in response to introduced predators or competitors (Parker et al. 1999).
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Figure 9.1 Measures of fitness for Spartina hybrids collected in San Francisco Bay. Hybrid category is defined by the proportion of the genetic marker states consistent with S. alterniflora (i.e., 0% = S. foliosa; 100% = S. alterniflora). Notice that most measures peak at mid to upper hybrid categories indicating that S. alterniflora × foliosa hybrids are more fit than their parental taxa. Reprinted from Ayers, D. R., D.L. Smith, K. Zaremba, S. Klohr, and D. R. Strong. 2004. Spread of exotic cordgrasses and hybrids (Spartina spp.) in tidal marshes of San Francisco Bay, California, USA. Biological Invasions 6: 221–231, with kind permission of Springer Science and Business Media.
One good example of a behavioral shift comes from the use of an alternative host plant by monarch butterflies (Danaus plexippus). Monarchs are fascinating insects with some forms that migrate from eastern North America to breeding and overwintering grounds in Mexico, or from western North America to overwintering grounds in California. They specialize on milkweed plants (Asclepias syriaca) for oviposition and larval feeding. With the introduction of the non-native pale swallowort or dog-strangler vine (Vincetoxicum rossicum), a European vine in the same family as milkweed, in the midwestern United States and Canada there have been anecdotal observations of monarchs ovipositing on the invading European species. In a field study, Heather Mattila and Gard Otis (2003) examined the use of each host plant alone and in the presence of the other plant. They found that monarch butterflies overwhelmingly preferred the native milkweed as a host but would use the V. rossicum when milkweed was not available. Monarch butterfly populations have been decreasing over the past decade due to anthropogenic disturbances to their habitat and toxin effects from agricultural pesticide applications. An alternative host plant (besides milkweed) could offer the possibility of population stabilization and recovery if reproduction and survival were similar on the native and non-native hosts. Unfortunately, larvae fed only negligibly on the nonnative vine and thus young oviposted there never survived beyond the first instar (first larval stage) (Mattila & Otis 2003). Any eggs that monarch butterflies lay on the invader, therefore, are doomed and reduce the fitness of the parent through wasted reproductive effort. Box 9.1 discusses very different effects from non-native pathogens that shift to new hosts.
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Box 9.1 Suddenly last summer Most of the invasive species we have discussed so far have had ecological consequences. Quite a few non-native species directly affect our every day lives and our economies. Although rarely discussed within ecological literature, most emerging infectious diseases we hear about on the TV news are non-native in the sense that they were first isolated on one continent and were then transported around the world via human travel or commerce. A recent example in the New World is West Nile virus, and its negative impact on human health and wildlife has been quick and dramatic. West Nile virus was first isolated in 1937 from a woman living in the West Nile district of Uganda (Marra et al. 2004). West Nile is an arthropod-vectored flavivirus that causes flu-like symptoms in humans and sometimes manifests as fatal encephalitis (Zeller & Schuffenecker 2004). Humans and horses are considered “dead end” hosts since they usually do not harbor enough of the virus to infect a vector arthropod, such as a mosquito or tick (Marra et al. 2004). West Nile virus was never very common after its initial isolation but started showing up regularly in humans and horses in Europe in 1996 (Zeller & Schuffenecker 2004). A particularly virulent strain was identified in Israel in 1998 from dying migrating geese (Anser spp.), and shortly thereafter (in 1999) the same strain was found in New York. This jump into the New World was a first, and, although no one is sure how the virus made its way to New York, mosquitoes (the primary vector) could have easily made their way between Israel and New York in ships and airplanes used to move goods and people. Since 1999 the virus has spread across North America, down to the Caribbean, and into South America (Zeller & Schuffenecker 2004). This strain of West Nile virus can be lethal to humans. Mosquitoes differ between species in their competence as vectors (i.e., how easily they become infected, replicate the virus, and then pass it on; Marra et al. 2004). However, there are a group of common mosquitoes (e.g., Aedes albopictus, Culex salinarius, Anopheles atropos) that are opportunistic feeders, readily taking bloodmeals from mammals, birds, and reptiles. It is likely that these mosquito species play a key role in the transmission of the virus, and a detailed look at their ecology and behavior seems more than appropriate (Marra et al. 2004). Since humans and horses do not serve as reservoirs for the virus, it is an open question as to which species do. Marra et al. (2004) report that as of September of 2003, at least 208 species of birds, 29 species of mammals, 1 species of lizard (Varanus salvadorii), and 1 species of alligator (Alligator mississippiensis) had exhibited symptoms of West Nile infections. This is probably an underestimate since there are few wildlife monitoring programs regularly determining the cause of death for wild-living individuals. Given the variety of bird species that show infection, there has been some focus on them as potential reservoirs. However, not all bird species are equally susceptible to the virus. Regardless of the ability of birds to maintain population levels of West Nile virus that pose a risk to human health, the virus itself poses a risk to
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wild bird populations. Marked American crow (Corvus brachyrynchus) populations suffered around 30% mortality from the virus (Marra et al. 2004). There is some evidence that nestling birds are more vulnerable then adults, and adults can pass the virus to their young via feeding. The real worry is the possibility that West Nile virus will find its way into populations of critically threatened species, where a loss of 30% of individuals in a short time span could spell the difference between persistence and extinction. It is especially important to keep the virus out of particularly susceptible populations like those that evolved in the absence of such viruses on isolated islands (e.g., the Hawaiian honeycreepers and Galapagos finches). There is a lot yet to be learned about the ecology of infectious diseases, and how these diseases affect human and wildlife populations. The fact that many of these diseases are of non-native origin calls the experiences and skills of invasion ecologists to the table. A second example of a shift in behavior comes from research on red imported fire ants (Solenopsis invicta), which were putatively introduced from Argentina into the southern United States in the 1940s. As we will discuss in the next section, these ants have signficant community-level effects, but they can also alter the behavior of native species they encounter. Researchers have found that several rodents will alter their foraging behavior in the presence of red imported fire ants (Pedersen et al. 2003; Orrock & Danielson 2004). For example, old field mice (Peromyscus polionotus) showed a stronger aversion to areas with red imported fire ants compared to their level of aversion to areas with both direct and indirect predator risks (Orrock & Danielson 2004). Indirect risks included foraging in open moonlit patches, which small rodents will avoid due to the risk of owl predation. Direct predator risks involved evidence of the presence of ground predators, such as the application of predator urine to areas (Fig. 9.2). In other words, old field mice would rather risk being eaten by predators than come into contact with the red imported fire ant. In another study, hispid cotton rats (Sigmodon hispidus) altered their habitat use to avoid areas occupied by red imported fire ants (Pedersen et al. 2003).
Population impacts Invasive species impacts are most commonly studied at the population level (Parker et al. 1999). This concentration may reflect the bias of ecologists toward species as the primary units of ecology, but it also most likely results from the relative ease of measuring population numbers, distribution, and age structure. The individual impacts we described above often translate into declines in rates of reproduction and survival, which affect population growth. If the non-native species has negative effects on a native species’ abundance across its entire range (e.g., when the native is rare or endangered)
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Figure 9.2 Mean proportion of seeds remaining in foraging trays after visitation by old field mice (Peromyscus polionotus) when fire ants (Solenopsis invicta) were absent (open bars) and present (gray bars) on nights with rain (a) or with no rain (b). Rain status is an indicator of the mice’s visibility to aerial predators like owls (rain = less visible). Trays were placed underneath shelter or in the open. Typically we would expect the mice to eat many more seeds when there is rain and the food is under shelter (i.e. mice unlikely to be preyed upon). However, in the presence of ants, mice substantially reduce their foraging rate because ants present as much, if not more, of a threat to their survival as other predators such as owls. Reprinted from Orrock, J. L., and B. J. Danielson. 2004. Rodents balancing a variety of risks: invasive fire ants and indirect and direct indicators of predation risk. Oecologia 140: 662–667, with kind permission of Springer Science and Business Media.
and if these effects persist for several years, the native species may be driven to extinction. However, extinction is the most extreme form of population-level impact. In this section we focus on the ecological mechanisms by which the invader changes a native species’ population structure or abundance. Such impacts reflect the multitude of ways in which two species may interact in general, but we will concentrate on cases where the invader competes with the native for critical resources, preys upon the native species, or physically inhibits the growth or reproduction of native species.
Examples of competition impacts In Chapters 6 and 8 we discussed the negative effects of competitors on the establishment and spread of non-native species. Here we reverse the coin and present two examples where the non-native has negative competitive effects on native species. Spotted knapweed (Centaurea maculosa) is a perennial composite (Asteraceae) that is a native of Eurasia but was introduced into North America in the early 1900s. It
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was apparently transported to North America accidentally as seed contaminant within crop seed imports. It now infests millions of acres of semi-arid grazing land in the northwestern states and provinces of North America. Spotted knapweed produces an allelopathic chemical that retards the growth of natives and improves the invader’s relative ability to capture resources (Kelsey & Locken 1987). In Montana, spotted knapweed is encroaching on the range of the endemic Mt. Sapphire rockcress, Arabis fecunda (Brassicaceae). This native perennial species is short-lived and confined to highly calcareous soils in the foothills of the Rocky Mountains in southwest Montana, USA. Aside from competition from spotted knapweed, A. fecunda faces threats from trampling by livestock and mining activities (Lesica & Shelly 1996). In an effort to identify the role of the invader in the decline of A. fecunda, Peter Lesica and Stephen Shelly conducted experiments in which they hand-removed spotted knapweed individuals from around A. fecunda individuals (within replicated 1 m2 plots) and then documented the growth, fecundity, survivorship and recruitment (i.e. the number of seeds that germinated) of A. fecunda. They then combined these field-derived estimates of demographic rates into a matrix-projection model to determine how spotted knapweed affected the long-term population dynamics of A. fecunda. Notice that these approaches combine data from both the individual and population levels. Within a year after removal of spotted knapweed, the average annual population growth rate (λ) for A. fecunda had gone from negative to positive, thus from a value indicating population decline to one indicating increase. This change resulted mostly from a significant increase in recruitment of A. fecunda after removal of spotted knapweed (Fig. 9.3), and to a lesser extent, from an increase in the number of fruits per A. fecunda plant (Lesica & Shelly 1996). The removal of spotted knapweed probably reduced or removed allelopathic toxins while also increasing the availability of limiting resources such as water, minerals, or light. Lesica and Shelly (1996) conclude that
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continued increases in the density of spotted knapweed will drive population sizes of A. fecunda down and increase the risk of extinction of this native species. They emphasize the need to institute concerted control of spotted knapweed within the range of this threatened plant, if not across the geographic extent of knapweed’s distribution in North America. We see a similar effect of competition between two native parrot species in Australia and two invasive birds there. Most of Australia’s native parrots use hollows in old trees for nesting. Conversion of natural woodlands into tree plantations (where old or dead trees are removed) or into other land use types (e.g., agriculture or suburban housing) has left the landscape short of good nesting sites. Like most species that rely on such natural hollows, many of Australia’s parrots have been declining due to the large drop in the number of natural hollows available. Birds compete heavily for the few natural hollows that remain as well as for the artificial nest boxes placed in woodlands to help alleviate the nest site shortages. This situation is made worse by the presence of behaviorally aggressive invasive birds, such as the common myna (Acridotheres tristis) and the European starling (Sturnus vulgaris), that readily use nest boxes and natural hollows for nesting. Pell and Tidemann (1997) endeavored to document the extent of this problem in two nature reserves near Canberra, Australia. These Eucalyptus woodland reserves held resident populations of two native cavity-nesting parrots, the crimson rosella (Platycercus elegans) and the eastern rosella (P. eximus). The authors estimated the density of each species within the reserves and documented the number of boxes or natural cavities used for nesting, the species using the boxes, the success of clutches laid in the boxes, and the species that won any aggressive interactions near the nest boxes. Pell and Tidemann’s (1997) results show that population density of the invasives was higher than that for the natives in all areas, except for deep within the woodlands (i.e., well away from the suburban edge of the reserves). Among the limited number of natural cavities available, the vast majority was used either by the common myna or the starling. Although all four species inspected the nest boxes as potential nest sites, the common myna ultimately occupied the majority of the boxes. Of particular interest is Pell and Tidemman’s (1997) observation of mynas harassing and often evicting breeding pairs of the native parrots from the nest boxes. The rosellas could not enter the box to incubate their eggs or feed their nestlings. Often the mynas pecked at the resident chicks directly killing them. This aggressive behavior greatly decreases reproductive success for the resident parrots. Notably, the starling did not show similar behaviors to the myna in this study although starlings have been implicated in the declines of hollow-nesters in other locations where they have been introduced (Koenig 2003).
Example of predation impacts The impact of invasive predators on native species is by far the most frequently discussed impact category, probably because of the often-devastating effects of invasive predators on island systems where the native fauna are evolutionarily naïve to their effects (e.g., Box 1.2). We will review one case study here in reference to the effects of
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invasive rats on the endangered tuatara (Sphendon punctatus). However, we will revisit this topic in the next section (community effects) since invasive predators can often decimate entire suites of native species. The tuatara is a lizard-like reptile endemic to New Zealand. Over the past 200 years, populations of the tuatara on the main islands of New Zealand have disappeared because of habitat change and predation by humans and invasive mammals. There are still tuataras on 25 islands off the northern coast of New Zealand, but some of these populations are also declining. Eight of these islands also have the non-native Pacific rat (Rattus exulans), locally called kiore (Atkinson & Moller 1990). On kiore-inhabited islands, populations of tuatara have precipitously declined, leading several conservationists to suspect a direct link between the presence of kiore and the tuatara decline. Alison Cree and her colleagues (1995) used field observations to document the impact of kiore on tuatara populations. They found that that on islands with kiore, tuatara populations were half the size of populations on kiore-free islands. Kiore apparently prevent or greatly inhibit the recruitment of young tuatara through a combination of competition for limited food resources and direct predation on the young and eggs. On kiore-free islands nearly 24% of all tuatara individuals caught were juveniles, whereas on kiore-inhabited islands this percentage dropped to about 8% (Fig. 9.4). On the kiore-inhabited islands with the lowest densities of tuatara, no juveniles were found 175˚E
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(Cree et al. 1995). Tuatara may also experience Allee effects at low densities. On islands with kiore the distance between adult tuatara and their nearest neighbor was frequently as high as 200–300 m; quite a long distance for such sedentary creatures (Cree et al. 1995). Cree and her colleagues hypothesized that at these densities low encounter rates between adults may further limit reproduction. Cree and her colleagues (1995) recommend eradication of kiore from the islands where tuatara still exist and the implementation of strong measures to prevent the spread of kiore to other islands, but the tuatara may be on its way to extinction. Even if protected from kiore, it is extremely vulnerable to any other egg predator introduced to these islands.
Examples of physical impacts Formidable population-level consequences can come from physical attributes of the invader. For example, there are several mussels in the family Mytilidae that attach to any hard substrate, including the hard shell of native bivalves. The most notable of these invasive fouling bivalves are the zebra mussel (Dreissena polymorpha; Box 8.1), which is native to the Black and Caspian seas, and the golden mussel (Limnoperna fortunei), which is native to China and southeast Asian freshwater rivers. The zebra mussel was originally introduced into the Great Lakes region of North America and has now spread to many other drainages in the middle of North America (Ricciardi 2003). The golden mussel was introduced into la Plata Basin in Argentina and has since spread into Uruguay, Paraguay, and Brazil (Darrigran & de Drago 2000). Both invaders lead to declines in unionid (Family Unionidae) mussel densities by physically overgrowing the natives. Hebert et al. (1991) documented more than 10 layers of zebra mussels growing on a single unionid mussel in Lake St. Clair of the North American Great Lakes. The golden mussel has similar layering capacities due to byssal threads that allow it to attach to hard substrates at a variety of depths (Montalto & de Drago 2003). The layering effect of both species reduces access to nutrients for natives on which they settle and often smothers the native bivalves (Ricciardi 2003).
Community impacts The population effects reviewed above can lead to substantial changes in community composition if the invader adversely impacts more than one native species. In some cases, the invader’s impacts are so great and so ubiquitous that entire assemblages of native species become extinct. We will begin by exploring instances of mass extinction due to an invader and then go on to explore more common examples in which attributes of the community are altered in the presence of an invader.
Examples of mass extinction Mass extinctions caused by invaders occur most often when an evolutionarily isolated system receives a novel predator or herbivore. In these cases, the long isolation that has
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allowed impressive adaptive radiation within one lineage has frequently taken place in the absence of predators or herbivores or at least in the absence of specific types of predators (i.e., large or omnivorous species). Predation is an extremely strong selective pressure, and naïve native species are particularly vulnerable to novel predation. A now-classic example involves the introduction of the piscivorous Nile perch (Lates niloticus) into Lake Victoria in the 1950s by British colonials and the subsequent loss of native cichlids. Lake Victoria lies near the African Rift Lakes, birthplace of nearly 400 native cichlid fish species. The isolation of the lake from others nearby, the morphologically “plastic” jaw structures of the cichlids, and the spatial and temporal variability of ecological conditions within the lake set the stage for a massive adaptive radiation of cichlids. The cichlids of Lake Victoria have been the subject of many studies on evolution, and their diversity led Tijs Goldsmith (1998) to refer to the lake as “Darwin’s Dreampond.” Nile perch were released with several other omnivorous fishes (that were to provide food sources for the perch) in an effort to boost the human food resources of the region and provide a source of economic revenue for the local human communities around the lake. Before the introduction of these fish, Lake Victoria fisheries had relied primarily on two native cichlid species (Oreochromus esculentus and O. variabilis), but these native species provided small harvests (Goudswaard et al. 2002). Because the Nile perch is much bigger than any of the cichlids and eats such little fish, the total harvested fish biomass was expected to increase with the perch’s introduction. After introduction of the non-native fish, the native cichlids declined due to competition with the introduced omnivorous fish and predation from the Nile perch (Goudswaard et al. 2002). Although initially the Nile perch did not constitute a large fraction of the fishermen’s catch, the population boomed around 1980. According to long-term records kept by local management agencies, Nile perch went from being less than 1% of the annual harvest (in biomass) in 1978 to being 97% of the catch in 1987 (Witte et al. 2000). The biomass of the native cichlids showed the opposite trend, going from 92% of the stock to barely present (Witte et al. 2000). During the 1980s, approximately 200 previously recorded native cichlid species disappeared completely (Witte et al. 1992). Of the 110 species collected in nets between 1978 and 1982, only 10 were found in 1987 (Witte et al. 2000). The Nile perch is not a particularly voracious piscivore, but it is has predatory behavior unlike any encountered by cichlids in Lake Victoria before its introduction. Nile perch grow in size to be several orders of magnitude larger than native cichlids, and they feed exclusively on smaller fish. Thus, each individual Nile perch can consume many cichlids in a single day. The rarest of the native cichlids and those that shared more of their habitat with the Nile perch probably went extinct first, while those species that use shallow habitats unavailable to the Nile perch were more immune to its effects. In these shallow waters, native cichlids are more susceptible, though, to human predation, leaving them between sharp teeth and fast nets. Careful management of the fishery has led to reductions in the average size of Nile perch (i.e., preferentially harvesting the really big perch) and a reduction in the density of the perch population. As a result, a few of the native cichlid species have recovered in the last few years. Nonetheless, at least half the native cichlids are now considered
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extinct. Another 40 cichlid species are being kept from extinction by a massive captive rearing effort sponsored by the International Union for the Conservation of Nature (IUCN) (Baskin 1992). The final result is a suite of problems for the local economy and the loss of a large fraction of this impressive assemblage of fishes (Baskin 1992). In another example of good intentions gone bad, the introduction of the predatory rosy wolf snail (Euglandina rosea) to the Pacific Islands for biocontrol led to the extinction of almost all of the area’s endemic snails (Partula spp.). The rosy wolf snail, which is native to Florida and Central America, was purposefully introduced by governmental agencies to Tahiti on 16 March 1977 as a biocontrol agent for the giant African snail (Achatina fulica), which is a well-known non-native agricultural pest that eats garden plants and crops (Clarke et al. 1984). A well-meaning citizen introduced the giant African snail onto Tahiti as a food crop in 1967. There is little evidence that Euglandina rosea (the rosy wolf snail) played any role in controlling the African snail’s population (Civeyrel & Simberloff 1996; Gerlach 2001). In fact, recent research has demonstrated that giant African snail declines are most likely due to disease dynamics (Civeyrel & Simberloff 1996; Gerlach 2001). Nonetheless, the decline in the giant African snail led to the declaration of the rosy wolf snail as a successful biocontrol agent. Unfortunately, there is unequivocal evidence that the rosy wolf snail readily consumes most native snails, and specifically the Partula species endemic to Polynesia (Coote & Loeve 2003). Clarke et al. (1984) documented the rapid decline of the Partula group as the rosy wolf snail expanded its range to cover the entire island of Tahiti, and it is this group of researchers that reported Partula extinctions in 1988 (Murray et al. 1988). Unfortunately, the impacts were not confined to Tahiti. Since the rosy wolf snail was considered a “success” as a biocontrol agent, it was subsequently introduced on several other Pacific Islands, each with their own endemic assemblages of land snails. In almost all cases, the native land snails are currently considered highly imperiled. In fact, a recent survey of the Society Islands, the archipelago in French Polynesia that contains Tahiti, found no remaining Partula species on any island except Tahiti (Coote & Loeve 2003).
Examples of community alterations We commonly see invasive mammalian herbivores alter plant communities, especially on islands. Because mammalian herbivores are also one of our principal sources of dietary protein, transportation, and clothing fiber, a few of these herbivores are distributed worldwide. Examples of invasive mammalian herbivores that have had a considerable impact on native plant communities include rabbits (Oryctolagus cuniculus), goats (Capra hircus), pigs (Sus domestica), horses (Equus caballus), and sheep (Ovis aries). The impact of these non-native herbivores is concentrated on the base of the trophic web (autotrophs), and consequently their effects on the native community can be wideranging (Courchamp et al. 2003). Feral sheep, for example, preferentially eat native shrubs and grasses endemic to Santa Cruz Island, California (Van Vuren & Coblentz 1987). The sheep on Santa Cruz Island show a more generalized diet than domestic sheep found on the mainland of California and have taken to standing on
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their hind legs to reach the foliage of several endemic native shrubs. The end result is physically simplified vegetation structure, high amounts of open bare ground, and the absence of several endemic plants within sites with large feral sheep populations. This effect “cascades” up the trophic web and virtually eliminates habitat for ground nesting birds (Van Vuren & Coblentz 1987). Within mainland communities, extinctions due to the effects of invaders are less likely and tend to be fewer when they do happen. A good example is the impact of introduced ants on native ants. For example, red imported fire ants (Solenopsis invicta), which we mentioned in the section on individual impacts, have caused severe changes to native ant assemblages in the southeastern United States. After the ant’s initial introduction into Mobile, Alabama (USA), it spread quickly mostly as a hitchhiker on grass and sod stock used to landscape the growing suburban areas of the southern United States. The red imported fire ant has two social varieties, one that forms single-queen colonies (monogyne), and a second that forms multiple-queen colonies (polygyne). It is the polygyne form, which first arrived in the United States in the 1970s, that seems to damage native ant communities the most (Porter & Savignano 1990). Polygyne red imported fire ants form supercolonies that have 50–100 times the density of monogyne ant colonies. Porter and Savignano (1990) examined ant abundance and diversity at the University of Texas Brackenridge Field Laboratory after fire ants arrived in 1980. They found that native ant species richness in areas with the fire ant (16 species) was half that in areas without the fire ant (32 species) and only 30% what it had been (57 species) before fire ant invasion (Porter & Savignano 1990). This effect carried over to native arthropods other than ants, as well. Species richness of non-ant arthropods dropped by 30% in the presence of polygyne fire ants, and the numbers of individuals dropped by 75% (Porter & Savignano 1990). Mainland and island systems may differ in the naivete and vulnerability of the native species, the availability of refuges, and the number of potential competitors or predators on the invasive species. For example, native ant species in Texas appear to have refuges from red imported fire ant effects. Recent research has shown that native ant species diversity has potentially rebounded at the Brackenridge Field Laboratory site (Fig. 9.5; Morrison 2002). Part of this putative recovery is due to decreases in the abundance of fire ants (Fig. 9.6), but recovery would not be possible at all if the native ants had gone extinct. Some of the hypotheses for imported fire ant declines involve a drop in resource abundance (Morrison 2002), recent prolonged drought (D. Holway, pers. com.), introduced parasitoids (Mehdiabadi et al. 2004), and a pathogen that followed the ants from South America (Williams et al. 1998; Oi & Williams 2002). In the wake of red imported fire ant invasion outside of Texas, several native species, across trophic levels, have declined. In the presence of red imported fire ants, many plant species have seen reductions in seed dispersal (Zettler et al. 2001; Ness & Bronstein 2004); native arthropod richness has declined at several sites (Eubanks et al. 2002; Morrison 2002); rodents have shown behavioral changes with implications for foraging and reproduction, as mentioned in the individual impacts section (Pedersen et al. 2003; Orrock & Danielson 2004); bird species have shown reduced fledgling success and population growth (Allen et al. 2001b; Smith et al. 2004a); and loggerhead turtles have shown reduced hatchling success (Allen et al. 2001a). From this example it is clear that
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invasive organisms can have large community impacts for mainland systems as well as islands. It is increasingly apparent that individual and community-level impacts can combine. When they do, the effects frequently appear at multiple trophic levels as in the red imported fire ant case, or alter human interactions with the environment, as discussed in Box 9.2. For example, the Sacramento pikeminnow (Ptychocheilus grandis), a predatory cyprinid that is native to the San Joaquin and Sacramento watersheds in California, was illegally stocked in California’s Eel River around 1980 (Brown & Moyle 1981). In the absence of the introduced pikeminnow, the native fishes were distributed across the width of the stream except at the shallow edges. In the presence of the large, predatory pikeminnow, the fish native to the Eel River shifted away from the deeper parts of the stream toward the shallow edges they had previously avoided. Overall, this behavioral shift in habitat choice resulted in much less overlap in the spatial use of habitats among the native species. Where they formerly coexisted across many habitat types, in the presence of the pikeminnow the natives began “specializing” on particular habitats (Brown & Moyle 1991). Similar effects have been seen in New Zealand with the introduction of the brown trout (Salmo trutta), which is native to Asia, Europe, and parts of northern Africa. Beyond just altering fish habitat use and abundance, the brown trout indirectly regulates the abundance and composition of algae in streams either by altering (through direct predation) invertebrate grazer abundance or foraging behavior (Flecker & Townsend 1994; McIntosh & Townsend 1995).
Ecosystem impacts Changes in population sizes, individual behavior, or community structure can alter both the flow of materials through an ecosystem and the disturbance regime that is typical of a native ecosystem. In relation to the brown trout introduction into New Zealand streams, for example, Huryn (1998) showed that the presence of brown trout translates into highly altered nutrient budgets with primary production six times higher in streams with brown trout than those without. Similarly, Simon and Townsend (2003) report that New Zealand streams with brown trout retain more nitrogen compared to trout-free streams. In general, an invader that uses a wide variety of resources will frequently alter ecosystem processes (e.g., Leavitt et al. 1994; Schindler et al. 2001). For example, zebra mussels alter the availability of nitrogen, phosphorus, and carbon to other species by filtering vast quantities of water through their digestive tract. This increases available nitrogen and phosphorus in the water column and “flushes” carbon from high in the water column down to the benthos (Gardner et al. 1995). Another example is salt cedar (Tamarix spp.), which we discussed briefly in Chapter 6. This shrubby tree sends roots into the ground water and does not actively regulate its transpiration rates. Its massive water consumption and ability to sequester and release salts (with falling leaves) have led to increased aridity and salinity in riparian zones of the semi-arid southwestern United States (Di Tomaso 1998). These changes have led to declines in native plant
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Box 9.2 Long time gone The impact of invasive species sometimes transcends ecological and economic concerns, especially when the invader greatly alters cultural and social systems. Here we review the consequences of an invasive fungus, chestnut blight (Cryphonectria parasitica), on American chestnuts (Castanea dentata) and the human communities that depended on these trees. The American chestnut was one of the dominant hardwood trees of deciduous forests of eastern North America particularly on relatively dry, north-facing slopes of the Appalachian mountains, which run through the eastern United States from Maine to Georgia (Smith 2000). Although not a strong component of the understory of these forests, American chestnut often comprised the canopy, regularly reaching 100 feet in height and 10 feet in diameter. Native wildlife and local farmers depended on the abundant crop of nuts the tree produced. Local people regularly gathered chestnuts and stored them for winter food supplies, an important goal in the pre-industrial lives of most American farmers. People especially liked to roast the nuts in the winter, giving rise to lyrics in standard Christmas tunes. Landowners in the southern Appalachians sold chestnuts to urban areas to the north (Philadelphia and New York City), and this was one of their principal cash crops. One of their other “cash crops” was the production of moonshine (i.e., home-brewed grain alcohol), the boilers for which were often heated by American chestnut wood (Youngs 2000). The primary use of the American chestnut, however, was for the production of timber. American chestnut grew straight and tall, its wood was easy to work and dried without warping, and the heartwood was extremely resistant to rot. The timber was used for such diverse things as telegraph poles, railroad ties, furniture, paneling, and pianos (Youngs 2000). It was a key element to the lives of people all along eastern North America. Given the importance of American chestnut to the economy of nineteenth and early twentieth century Americans, it was considered a “right smart little jolt” (Walter Hopkins quote from Youngs 2000) when the tree rapidly declined in abundance in the 1920s and 1930s and became so rare in the 1940s as to become economically “extinct.” The cause of the decline was the 1904 introduction of chestnut blight. Chestnut blight is a fungus native to Asia brought into North America on infected nursery stock. It was first detected at the Bronx Zoological Park in New York City, and within 40 years it had spread across all areas of the natural range of the American chestnut (Griffin 2000). Chestnut blight girdles large trees by producing cankers that kill the vascular cambium, resulting in the loss of the tree’s foliage. The fungus is transmitted to nearby trees via spores that are winddispersed. It persists in North American deciduous forests to this day and regularly infects and kills American chestnut saplings once they reach about 15 m in height and 20 cm diameter (Paillet 2002). This is long before these trees make it into the canopy and produce flowers or buds (American chestnuts are not self-fertile). Thus, even though American chestnut trees can still be found in these forests, they are not reproducing. Apparently, all “new” saplings are actually very old saplings
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that keep regrowing from sprouts, a fortuitous adaptation to living in forests in which tree canopies are not often opened by disturbance (Paillet 2002). The loss of American chestnut trees as forest dominants did not lead to massive deforestation in the eastern United States. Instead, various oaks such as red oak (Quercus rubra) and white oak (Q. alba) filled in the spaces left by the large chestnut trees. Thus, it is not entirely clear what impact the loss of American chestnut trees had on native wildlife, except to note that their demise left many snags for woodpecker nesting and resulted in a certain decline in available food stores for such species as squirrels and chipmunks. We may never know the ecological impact of the loss of these trees, unless forest scientists manage to develop blight-resistant individuals that can regain their place in the forest (Griffin 2000). However, clearly the loss of American chestnuts had an intense cultural impact on the local people. species and changes in arthropod and bird communities (Di Tomaso 1998; Ellingson & Andersen 2002). Notice that the examples above not only alter their environment by using more resources and expelling byproducts, but they also have a somewhat novel approach to resource use. All of them can capture resources despite abiotic stresses that exclude many native species. This ability to use novel resources or to extract resources where other species cannot also often leads to ecosystem changes. For example, nitrogen-fixing plants are critical to early succession because they help build soil and introduce limiting nutrients. As invaders, nitrogen-fixers can fundamentally alter resource availability and the balance of competition between native and non-native plant species with consequences for community structure and ecosystem processes. An oft-cited example of this is the fire tree, (Morella faya, formerly known as Myrica faya), which was introduced to the Hawaiian Islands in the late 1800s. M. faya forms actinorrhizae, woody nodules of symbiotic nitrogen fixers on its roots. Peter Vitousek and colleagues have been studying the influence of M. faya on the primary succession of young volcanic substrates within Kilauea Volcano (Hawaii Island) since the early 1980s. Ecosystems developing on volcanic substrates are nitrogen limited, giving nitrogenfixing species a clear advantage in the early stages of the successional process that will ultimately result in a rainforest. However, Hawaiian rainforests have no native nitrogenfixing species that are capable of colonizing these young volcanic substrates (there are nitrogen-fixing plants that show up in later stages of succession). Thus, the presence of the invasive M. faya adds a novel element to succession, ultimately quadrupling inputs of fixed nitrogen as compared to “native” inputs from rainfall, native nitrogen-fixers, and native decomposers (Vitousek 1990). Further, these inputs of nitrogen are available for uptake by other plants. In the early development of volcanic soils, these plants are more often other non-natives than natives. Ultimately, the presence of M. faya transforms ecosystem dynamics of Hawaiian volcanic substrates by setting the successional process on a novel trajectory that leads to the dominance of non-native species.
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In a comprehensive review of the literature on soil nutrient cycling in the presence of invaders, Ehrenfeld (2003) produced a tentative summary of patterns. Invasive plants tend to (i) have higher standing crop biomass and net primary productivity resulting from faster growth rates compared to natives, (ii) produce litter that decays faster than that of co-occurring natives, (iii) leave more extractable nitrogen in underlying soils than natives, (iv) produce higher levels of nitrogen mineralization and nitrification than natives, (v) have a large effect on nitrogen cycling if they are associated with symbiotic nitrogen-fixers and (vi) change the timing of nutrient fluxes and the spatial distribution of soil pools even when there are no changes to the absolute quantities involved. In addition, invasive non-native plants that do not fix nitrogen themselves can affect nitrogen fixation rates through interactions with nitrogen-fixing natives. Overall, soil pools of nitrogen, carbon, and water will respond to the presence of an invasive plant, but the direction of change is fairly unpredictable.
Landscape, regional and global impacts Finally, it is important to think about the global implications of invasive species impacts. Although we have given many examples of species loss in this chapter, many local areas have experienced an increase in species richness with the introduction of non-native species (Sax et al. 2002; Sax & Gaines 2003). However, gains and losses in species richness are not evenly distributed across taxa. Species that coexist with humans and that thrive on human modifications to the environment are more likely to be transported and more likely to survive in novel habitats. On the other side of the biodiversity ledger, endemic species that are naïve to human impacts or to the impacts of novel competitors, predators, or pathogens are likely to decline towards extinction. This leads to a group of species that is slowly losing ground (“losers”) and a much smaller group that is becoming ever more numerous (“winners,” McKinney & Lockwood 1999). The loss of endemic species and the gain of already widespread invaders serves to increase similarity in flora and fauna across previously distinct regions, a phenomenon dubbed homogenization by McKinney and Lockwood (1999). Homogenization is not a fast process, and thus it is not always readily apparent (Marchetti et al. 2006). The time scale for detecting invasions is much faster than the time scale of extinctions. Increased mortality and decreased reproduction can lead to slow population declines that may take generations to detect. Even if these changes are so extreme as to preclude reproduction, long-lived species may give the appearance of healthy populations if we collect abundance data rather than detailed demographic information. Thus, while introductions can happen instantaneously, extinction usually requires several generations. On what time scale, then, should we examine homogenization? With the increasing pace of introductions, should we ever expect to witness declines in net species richness (i.e. including both gains from non-natives and losses from natives), and should we allow increases in richness to lull us into dismissing the threat of homogenization? These questions have engendered a fair amount of debate. Julian Olden and LeRoy Poff (2003) have pointed out that simple diversity comparisons are extremely sensitive
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to time scales (because of the difference in rates between invasion and extinction) and to initial native diversity (because of comparison statistics and biological resistance). They suggest that a more useful model might be to examine the ratio of winners to losers in any given environment (Fig. 9.7; Olden et al. 2004). However, even if we remove the effects of time scales and original species richness, we do not gain a useful understanding of homogenization purely by looking at taxa. Species perform individual functions within a community and ecosystem. The process of invasion introduces new selective agents that then change the distribution of successful traits or characters within a community. A more productive line of inquiry may be to look at genetic and functional homogenization in addition to taxonomic homogenization (Fig. 9.8; Olden et al. 2004). For example, hybridization and introgression decrease novel genotypes and phenotypes in a community. If there are additional selective pressures placed on that community that conspire to make the traits of the invader more successful, then the overall functional diversity of the community may not only shift but may decrease.
Summary The range of impacts of invasive, non-native species is large, encompassing nearly every ecological relationship imaginable. However, there is a noticeable deficit in the
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Figure 9.8 Communities can start with several different initial distributions of traits. In each diagram above, the solid line indicates the current trait space distribution, and the dashed line indicates the historic distribution. All three communities display genetic homogenization removing the tails of the distribution, but communities B and C also display functional homogenization. Reprinted from Olden, J. D., N. L. Poff, M. R. Douglas, M. E. Douglas, and K. D. Fausch. 2004. Ecological and evolutionary consequences of biotic homogenization. Trends In Ecology & Evolution 19: 18–24, with permission from Elsevier.
information we have on impacts across biological levels of organization (e.g., Vitousek 1990; Simon & Townsend 2003). For example, how does genetic introgression translate into changes in demographic rates and, ultimately, to population declines? How does the behavior of invasive fishes influence nutrient cycles and, ultimately, contribute to the establishment of other non-native species? As is often the case with the study of nonnative species, these links seem obvious, but they have remained largely undocumented. This obscurity and the need to simply confirm several of the patterns we have reviewed here should make it clear to students that there is considerable research left to do on the impacts of invasive, non-native species. The impact of invasive non-natives can be substantial even though only small fractions of non-native species successfully move through all prior invasion stages. However,
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is there a pattern to when impacts are high versus when they are not? Indeed, is there a single metric of impact severity that we can use to compare across situations and that we can potentially use to help us determine where to put scarce management money to its best use? These are topics that we tackle in the next chapter.
Further complexities Courchamp, F., J. L. Chapuis, and M. Pascal. 2003. Mammal invaders on islands: impact, control and control impact. Biological Reviews 78: 347–383. D’Antonio, C. M. and P. M. Vitousek. 1992. Biological invasions by exotic grasses, the grass/fire cycle, and global change. Annual Review of Ecology and Systematics 23: 63–87. Flecker, A. S., and C. R. Townsend. 1994. Community-wide consequences of trout introduction in New Zealand streams. Ecological Applications 4: 798–807. Morrison, L. W. 2002. Long-term impacts of an arthropod-community invasion by the imported fire ant, Solenopsis invicta. Ecology 83: 2337–2345. Rahel, F. J. 2000. Homogenization of fish faunas across the United States. Science 288: 854–856. Vitousek, P. M. 1990. Biological invasions and ecosystem processes: towards an integration of population biology and ecosystem studies. Oikos 57: 7–13.
Companion papers Ehrenfeld, J. 2003. Effects of exotic plant invasions on soil nutrient cycling processes. Ecosystems 6: 503–523. Ness, J. H., and J. L. Bronstein. 2004. The effects of invasive ants on prospective ant mutualists. Biological Invasions 6: 445–461. Olden, J. D., and N. L. Poff. 2003. Toward a mechanistic understanding and prediction of biotic homogenization. American Naturalist 162: 442–460. Rhymer, J. M., and D. Simberloff. 1996. Extinction by hybridization and introgression. Annual Review of Ecology and Systematics 27: 83–109. Ricciardi, A. 2003. Predicting the impacts of an introduced species from its invasion history: an empirical approach applied to zebra mussel invasions. Freshwater Biology 48: 972–981. Sax, D. F., S. D. Gaines, and J. H. Brown. 2002. Species invasions exceed extinctions on islands worldwide: a comparative study of plants and birds. American Naturalist 160: 766–783.
10 Impact Synthesis
Overview In the preceding chapter we discussed the range of potential invasive species impacts and gave examples at different biological levels, but we did not try to synthesize patterns or compare their severity. Here we examine some of the general issues surrounding recognition of invasive species impacts, including how scientific and human perceptions influence our analysis of impact. This discussion leads us to two graphical models that highlight the range of ecological, historical, and societal factors needed to assess invasive species impacts. To measure impacts across hierarchical levels from genes to communities and compare effects across such different levels we need a common currency. We discuss economic currencies and a couple of different approaches to developing ecological currencies. We find that quantitatively assessing the degree of impact within one ecological currency is difficult given the extent of complexity surrounding ecological interactions and the multistage nature of the invasion process.
Perception and recognition of impact This final and often dramatic stage of the invasion process has been the impetus for much of the interest in invasion ecology, yet the definition of an impact is somewhat problematic for two reasons. First, there are scientific and societal influences on the perception of impact. Second, after impact is perceived, there are a variety of ecological factors that determine the level of impact produced. We intuitively understand that species that substantially alter community structure or cause significant declines in native population density have an impact, but where do we draw the line for non-native species having no impact? In fact, is it possible for a species to invade a new environment but truly have no impact? By establishing in a new community, a non-native species creates interactions with natives that must have some effect. There are, nonetheless, numerous examples of non-native species establishing with little to no detectable effect on native species (Williamson 1996; Davis 2003). Is this because measured effects rely on both our powers of observation and our valuation of the impact, or because there
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really is no effect? Here we review the factors that influence our perception of invasive species impacts, and then in the next section we integrate these influences with the ecological context of the invasion process. Human perception and valuation are an integral part of this stage of the invasion process (Williamson 1996). For example, society actively encourages many non-native species through agriculture, horticulture, pisciculture, and other economic ventures. Society clears space and disrupts native communities for these enterprises and expects these non-native species to provide consistent sources of food, medicines, clothing, labor, and general profits. Society sometimes encourages not just particular taxa but also particular ecological processes. As reviewed in Chapter 5, restoration ecologists use cattle to mimic the positive effects of herbivores on native plant communities, and they plant non-native species to push fire cycles toward more natural regimes. The entire biocontrol industry is based on our desire to use non-native species to control negative ecological effects, particularly the population sizes of non-native pest species that decrease our agricultural yields. As we discussed in Chapter 2, there is a whole range of aesthetic, medicinal, cultural, and economic reasons for the transport and introduction of non-native species. Sometimes these species produce positive impacts, and society encourages them as long as they can be controlled. Indeed it is possible that a non-native species that establishes but causes no decline in native diversity may actually offer ecosystem enhancements by allowing greater diversity, more functional repetition with associated increased ecosystem stability, and more resources for native species. These are all examples of “impacts,” but they are considered positive and are often ignored within the invasion ecology literature. It is important to remember, however, that the way society values the effects of nonnative species changes over time so that a positive impact in one decade may be of little interest to society in a later decade. For example, many of the positive emotional benefits of ornamentals that we valued in the middle part of the twentieth century are now less important to us than the loss of native diversity that may accompany these introductions. In addition, if we cannot detect an impact now, we often assume that there will never be one. There are many reasons why non-native species may be slow to establish or spread, and impediments experienced in these earlier invasion stages can lead to time lags in the onset of ecological impacts. For example, the non-native clam (Gemma gemma) began to expand its range within San Francisco Bay only after the recent introduction of the European green crab (Carcinus maenas) (Grosholz 2005). Because the European green crab prefers native clams (Nutricola spp.) that compete with G. gemma, their establishment in the bay led to competitive release for the invasive clam. G. gemma is now spreading across San Francisco Bay, changing native community structure in its wake, after more than fifty years of restriction to low salinity areas where it produced little (or restricted) ecological impacts. In other cases, however, a time lag occurs because of a delay in detecting the impact rather than in the onset of the actual effects themselves. Inherent limits to our ability to detect ecological changes and the vagaries of academic and practical interest in particular species often blind us to invasive species impacts (Box 10.1). For example, once a non-native species has established and spread to new sites, its ecological effects
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Box 10.1 I can see clearly now The analysis of the impact stage is clearly a complex and challenging endeavor involving many features of species and natural systems and many levels at which we can assess effects. To highlight an example of the complexity we turn to a study of competition between a native California grass (Sporobolus airoides) and two separate invasive species (Hordeum marinum gussoneanum and Bromus diandrus) that have very different impacts (Hoopes & Hall 2002). Many native species in California grasslands have either been replaced by non-native annual grasses or have been largely confined to poor quality refuge areas. In the S. airoides system Hoopes and Hall noticed that the native grass was either not found with the nonnative species or performed significantly worse in the presence of the non-natives (Fig. 10.1; Hoopes 1999). Hoopes and Hall collected soil from areas dominated by either one of these two non-native grass species or the native perennial grass S. airoides. They then examined soil characteristics and conducted field and laboratory competition experiments to see how all three species performed in different soil types and how well S. airoides performed when competing with either of the non-native grass species. B. diandrus could invade high nutrient soils and exclude S. airoides from these sites, while H. marinum gussoneanum performed well in all soil types where S. airoides was found and was the most common neighbor for S. aroides. This pattern initially suggested that B. diandrus had larger effects than H. marinum gussoneanum on the population viability of S. airoides. By looking at non-lethal effects, though, Hoopes and Hall (2002) found that H. marinum gussoneanum had potentially stronger and more widespread impacts on the native perennial grass
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species. H. marinum gussoneanum limited seedling survival, adult growth, and adult reproduction over a far wider range of habitats than B. diandrus. Based on these results, Hoopes and Hall (2002) identified two different categorical types of non-native species. The first type they named a “selective” invader (B. diandrus in their example) and defined as a species that only successfully invades specific habitats (high nutrient habitats in their example) but that is a superior competitor in these habitats and can therefore exclude native species (Hoopes & Hall 2002; see Fig. 10.2). The second type of invader they called a “twin” invader (H. marinum gussoneanum in their example) and defined as a species that has broad niche overlap with the native species (Hoopes & Hall 2002). When examining the impact of both types of non-native species on native species, it is often the “selective” invaders that produce the obvious or strong effects that monitoring and research quickly identify (Hoopes & Hall 2002). The showy easily identified impacts of the “selective” invaders may allow us to overlook “twin” invaders that produce much less obvious immediate effects but in the long run may be more damaging to native systems and species due to the high potential for long-term negative effects (Hoopes & Hall 2002). Selective invaders may be species that only invade disturbed areas, while twin invaders have the ability to invade relatively intact systems. Information on these two types of invaders may allow us to increase our predictions regarding the impacts of non-native species. In addition what this example demonstrates is that the impact of non-native species is not always obvious a priori. There are temporal, spatial, ecological and, perceptual issues involved at this stage.
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selective Local habitats
Figure 10.2 The dashed line represents a wide-spread invader with smaller or less apparent effects, a twin invader, and the dotted line represents an invader with localized, strong effects, a selective invader. The area under the curves represents the regional effect of each invasive species on a focal native species. The twin invader is more likely to seem innocuous, but has a larger regional effect. It is interesting to consider the conditions under which either invader can become a super invader (solid line), with strong effects in all habitats (Hoopes and Hall, 2002).
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on co-occuring short-lived native species often persist for many years before ecologists notice them (Hoopes & Hall 2002). Imagine how long it might take ecologists to detect genetic changes or changes in the age structure of long-lived native species. These changes can lead to very slow declines in population abundances and to subtle shifts in community structure (Holway & Suarez 1999) that may take decades, if not longer, to detect. At an even more basic level, we have to identify organisms that are producing impacts and recognize these taxa as being non-native in order for these species’ impacts to be noticed. Unfortunately our taxonomic knowledge of most groups of organisms (e.g., insects, arachnids, viruses, etc.) is far from complete (Wilson 1990). Although we believe that we have identified and classified more than 80% of vertebrates and flowering plants, for most other groups we have identified less than 50% and often less than 1% of the world’s species (Heywood 1995). Without full knowledge of what was, or is, present at any location, it is very difficult to identify non-native species introductions. New genetic techniques are helping, but there are still heated debates about some of their interpretations. The invasive status of Littorina littorea, a marine gastropod, is an excellent example of this taxonomic deficit. This intertidal gastropod was first found in New England (USA) in 1840 and has subsequently become common along most of the Atlantic coast of North America. This pattern of discovery made the species appear to be a nineteenthcentury introduction, perhaps brought with the increase of whaling and fishing in New England during that time. Recent mitochondrial and nuclear DNA sequence data suggest that this initial assumption was wrong and that L. littorea has been present in North America for more than 8000 years (Wares et al. 2002), but these results are still under debate. Similarly, we do not have complete species level descriptions for the aquatic insects in most continental freshwater systems (Cummins & Merritt 1996). It is quite likely that large numbers of freshwater aquatic invertebrates were introduced along with the many fish taxa society has transported, but we cannot identify these aquatic invaders without more information about native species. It is, therefore, entirely likely that there are “quiet” or “hidden” invaders species in our midst of which we are unaware.
Integrating perception with ecological determinants of impact The above section illustrates the multivariate way in which impact is perceived and thus defined. How do we put all of this together with our knowledge of the ecological determinants of impact? There is really no easy way to accomplish this goal; however, we offer two simple graphical models that help us envision the range of interactions that lead to labeling a non-native species as having an impact. Figure 10.3 places the ecological factors that determine level of impact within the larger framework of perception, both human and scientific, so that we can more easily recognize the subjective interpretation we bring to the topic. Notice that the overlapping
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Invasive species impacts Human and Scientific Perception Species characteristics
Ecological interactions
History
Environmental site characteristics
Figure 10.3 Venn diagram of potential areas of invasive species impact. Each sphere represents an area of potential impact. Notice that all impacts are dependent on perception, both human and scientific.
spheres pull together the roles of site characteristics, invader characteristics, ecological interactions, and history to determine when an invasive species is causing an impact. We also need to consider how the importance of these factors varies across invasion stages. For example, Kolar and Lodge (2001) show that species-specific traits that increase the probability of transport and release can go on to influence (positively or negatively) the likelihood of successful establishment and spread for a non-native species. If we add this across-stage perspective to the ideas portrayed in Figure 10.3, we can imagine that each factor has a role to play within one or more invasion stages (Fig. 10.4). Each sphere in Figures 10.3 and 10.4 represents topics we covered up to this point in the book. Below, we review these ideas in the context of our graphical models of impact.
Environmental site characteristics Suites of environmental characteristics may make recipient sites more or less susceptible to impacts. Several researchers have examined how productivity affects invasibility and eventual impacts on native diversity (Tilman et al. 1997; Stohlgren et al. 1999; Mouquet et al. 2002) with varying results (see Chapter 6). Spatial and temporal variability affect both establishment and spread, as we discussed in Chapters 6, 7, and 8, and these effects can carry over into impacts. For example, Marchetti et al. (2001) show that disturbance regime (natural or anthropogenic) influences the degree of impact that invasive fish impose within California watersheds. The overall biodiversity of a site can also exert
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Environmental site characteristics
Species characteristics
Species characteristics
Ecological Interactions
Scientific and Human perception
History
Transport
Establishment
Spread
Impact
Figure 10.4 Success at each stage may depend on several types of factors, may affect subsequent stages, and will affect impact. Final impact is also dependent on our perception. This figure combines the Venn diagram of Figure 10.1 and the concepts from Equation 7.3.
both positive and negative influences on non-native species establishment and impact as reviewed in Box 6.2 (Levine et al. 2003; Wearne & Morgan 2004; MacDougall & Turkington 2005).
Species characteristics The inherent traits of species, both native and non-native, can play a role in the impacts of non-native species. These traits can be standard life history characters such as number of seeds/propagules/offspring produced per generation, overall body size, degree of polyploidy and more (Marchetti et al. 2004a). Species characteristics can also include ecological traits such as size of home range, degree of physiological tolerance, diet breadth, trophic category, ability to fix nitrogen, and adaptation to fire. The evolutionary legacy of native species can also play a role at this stage. For example, if native species are island endemics or if they have a long history of isolation, they may be more vulnerable to invasive species effects (Kitchell et al. 1997; Fritts & Rodda 1998).
Ecological interactions Non-native species may affect natives by directly interacting with them, but their impacts are unlikely to be limited to direct interactions. In Chapters 6 and 8 we reviewed the most commonly studied ecological interactions (i.e., mutualism, competition, and predation). Clearly non-native species that have any of these direct interactions with natives will have impacts, but food web effects and indirect interactions may strongly influence the eventual impact of non-native species (Vander Zanden et al. 1999; Vander
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Zanden et al. 2003). For example, non-native species may carry significantly lower parasite loads within their introduced range, and release from the effects of parasites may allow invaders to behave differently, accumulate resources more aggressively, or simply increase in numbers so that they have stronger direct effects on natives (Torchin & Mitchell 2004). Non-native species may vector diseases or parasites that not only affect their own performance but also alter the native community in which they become established (Torchin & Mitchell 2004). Non-natives may also have indirect effects by altering interactions between native species.
History History can influence invasive species impacts through past natural events, such as weather, climate, rainfall, earthquakes and the like, as well through anthropogenic antecedents, such as acclimatization societies, mode of transport, propagule pressure, and land use (Marchetti et al. 2004a; Moffatt et al. 2004). For example, cheatgrass may have produced a higher degree of impact on North American rangelands and ranchers because it was introduced into North America several different times and across several distinct locations (see Box 8.2).
Perception The importance of this aspect of defining impact cannot be overstated because it is critical to our understanding of the overall context of effects of non-native species. Perception is characterized by scientific detection and by human bias. Our tools, whether they are microscopes, genetic assays, or species descriptions, limit our ability to detect non-native species and their effects. The impact of a non-native species is also a value judgment based on human criteria such as economics, aesthetics, ecological function, ecosystem processes, and human health or well-being, and these values can change with the shifting sands of social mores.
Finding common currencies In order to generalize about the impacts of non-native species we need ways to define, assess, and compare impacts across ecological communities and biological levels. A common currency for measuring impacts from genes to ecosystems may help us decide in a consistent fashion what constitutes a minor or a major impact.
Economic currency One fairly obvious common currency would be the economic cost of non-native species (Pimentel et al. 2000). Pimentel et al. (2000), for example, estimate that the cost of nonnative species to the US economy is nearly $137 billion per year. This estimate includes
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losses and damages of goods and services due to non-native species and the cost of controlling invasive species. Despite the fact that an economic currency seems fairly consistent, many of these impacts do not translate easily into economic terms. Pimentel and his colleagues use a variety of methods to place dollar values on the ill-effects of non-native species, including pricing intangibles, such as changes in biodiversity (see Box 10.2). Many conservation biologists have criticized these approaches, however, as misguided comparisons of apples and oranges, or even apples and orangutans. Others believe that such valuations are conceptually sound but are not complete without a full cost–benefit analysis. Pimentel and his colleagues did not include the benefits that nonnative species can sometimes produce (e.g., through crops, biocontrol, ornamental plantings), but they point out that these benefits do not diminish the large costs of non-native species.
Ecological currency Parker formula Despite the (relative) ease and commonality of economic valuation, this approach has very little relevance to biological or ecological issues. Several researchers have proposed alternative ecologically rooted formulas, based on per capita effect sizes, abundance, and other factors. The most influential has been Parker et al. (1999), who proposed the following simple linear equation, I =R∗A∗E
(10.1)
where I is overall impact, R describes the range of the non-native species (m2 ), A is the average abundance of the non-native species (per m2 ), and E is the effect of the non-native species (per individual or per unit biomass of the non-native species). This proposed relationship allows us to compare the impact of non-native species across taxa and across geographic space (Parker et al. 1999), and it is a useful heuristic tool that identifies some of the basic interacting components of invasive species impacts. Parker and her colleagues point out that many species have large impacts because of their particular “quirks” or characteristics, such as the effects of the nitrogen fixer, Morella faya in Hawaii (see Chapters 6 & 9), while others have large impacts because of their vast numbers, such as zebra mussels (Dreissena polymorpha) in the North American Great Lakes (see Chapters 6 & 9), and still others have large impacts because they are extremely widespread, such as the chytrid fungus (Batrachochytrium dendrobatidis), which is affecting numerous amphibian species worldwide. The difference between these types of impacts is basically a debate over whether effects per individual invader (E), overall abundance (A), or range (R) are most important. Clearly though, a widespread, abundant species with strong per individual effects will have the most powerful impacts. Despite its simplicity and appeal, the Parker formula presents some minor difficulties that are worth examining. For example, and as we mentioned above, we frequently have trouble simply identifying non-native species much less acquiring accurate range (R)
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Box 10.2 Material world We decline to step into the debate over the appropriateness of economic valuations such as that done by Pimentel and his colleagues (2000). Because we believe it is important that ecologists understand how economists approach questions of costs and benefits, we instead offer some insight into these approaches so that our readers can judge their applicability themselves. Following closely the excellent overview of the economics of non-native species by Naylor (2000), we describe methods of economic valuation and how they relate to the ecological (and other) impacts we describe in Chapter 9. Several types of values can be calculated with respect to non-native species, and these values should reflect both the economic and ecological dimensions of the issue. Although there are a variety of methods for carrying out valuations, there are three main categories: (i) direct consumptive and non-consumptive use values, (ii) indirect-use values, and (iii) non-use values. Direct consumptive use is probably the easiest method of valuation as it reflects the monetary worth of goods and services affected (positively or negatively) by non-native species. Current market prices for these goods and services are calculated and used directly in cost–benefit analyses although sometimes these prices must be corrected to reflect government subsidies or other factors that alter market values. For example, it is easy enough to calculate the market value of livestock, crops, water supplies, pets, and ornamental plants simply by doing a survey of retailers or markets dealing in these goods. Often these goods are non-native species allowing a quick way to calculate the benefit of the species, at least to the company or individual that sold the species. In other instances, when the good or service is harmed by a non-native species, this loss can be directly calculated based on what the supplier could have made had its product not been compromised by the non-native. The calculation of non-consumptive direct values is a bit trickier. Examples of non-consumptive use include bird and whale watching, catch-and-release fishing, scuba diving reefs, and ecotourism. Since the species being non-consumptively “used” do not have a direct market price, we have to find an indirect means of placing a value on their use. Typically, economists use travel costs to value these experiences, which are the total costs incurred by individuals who wish to participate in an activity such as bird and whale watching. Thus, in the case of bird watching, we would need to add up the cost of purchasing binoculars and field guides, arranging travel to a location of choice, paying a guide to help you find the birds, lodging costs, fuel costs, and any other related expenses. Although not usually the case, non-native species could be the species that people desire to see, thus providing a way to calculate their non-consumptive use benefit. In most cases, non-native species are seen to degrade these experiences (if not wipe them out) and thus we can calculate the reduction in travel expenditures due to the presence of the non-native species. Indirect-use values are generally thought of as the services provided by ecosystems such as providing clean and abundant water, sequestering carbon, or controlling erosion. These costs and benefits are much more difficult to estimate
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Box 10.2
Continued
than direct values. Most often ecosystem services are valued by estimating the cost of replacing the service they provide. For example, New York City utilizes the Hudson river watershed as a source of clean water. The value of that watershed could be estimated by calculating what it would cost to provide the millions of city residents with clean water in the absence of the watershed. When thinking about the costs of non-native species, we want to know how much less (or more) water the Hudson river watershed produces for New York citizens after the non-native(s) establish there. This can also work the other way around, where we ask how much more of a service society could enjoy if a non-native species were removed. In this instance, we are placing a value on the opportunity we are missing by not removing the non-native(s). Finally we come to the most contentious form of valuation, and the one that most ecologists would consider the critical one for inclusion in a full cost–benefit analysis: non-use values. These are the values we give to species and ecosystems simply because they exist (i.e., existence values). It also may include the value we give to species because we think they may be important in the future (e.g., provide the cure for cancer) and thus we would rather not lose them just yet. Non-use values clearly reflect each individual’s ethical relationship to nature, their belief that technical solutions are possible for most human problems, and their appreciation of simply knowing that unique places exist in the world. For example, some individuals value unique assemblages of species highly, and others do not. Usually, non-use values are estimated by surveying people and asking them how much more they would be willing to pay to ensure that particular natural amenities remain in existence. These price premiums could be attached to a particular good or service that the survey participant regularly enjoys (e.g., a cup of coffee), or it can be a more free-flowing notion of cost (e.g., asking how much would you pay to know that ivory-billed woodpeckers are ensured of not going extinct). As you can guess, this form of valuation can be skewed by the manner in which the survey is administered, and this variability makes such valuations contentious. There are many places where economics and invasion ecology cross paths. Valuing the costs and benefits of non-native species is only one. Economists often consider the effects of natural systems on societal problems such as famine and drought. Increasingly, these events are caused or exacerbated by non-native species. In Chapter 12 we highlight some of the other ways that economics enters into invasion ecology.
and abundance (A) data for them. Therefore, the Parker formula may only be useful for well-studied organisms. Similarly, it is not inherently obvious how we should measure per individual effects (E) or whether the per individual effect of a given non-native organism is constant within and among populations. Parker and her colleagues (1999) recognize this issue and break per individual effects down into genetic, individual,
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population, community, and ecosystem level impacts as we did in Chapter 9. They then explore how per individual effects have been measured in the past and how to model them in the future. Unfortunately, Parker et al. (1999) do not address whether per individual effects are comparable among widely differing taxa nor do they develop a satisfactory way of calculating effects when they cross several ecological levels of organization. As we saw in Chapter 9, many non-native species have effects at multiple organizational levels from genetic through behavioral to ecosystem structure. These problems impede the application of the Parker formula but do not decrease its value for recognizing the pertinent ecological variables that determine the severity of impacts.
Ricciardi formula A similar definition of impact comes from Anthony Ricciardi (2003), who proposes that impacts come from how a non-native species interacts with both its environment and species in the native community. The Ricciardi formula represents impacts as, I =A∗F ∗C
(10.2)
where I is impact, and A is abundance of the invasive species, while F describes the invader’s ecological function, and C describes the composition of the recipient community. This approach highlights different components of impact from the Parker formula and highlights the importance of an invasive species’ ecological role within the community it is invading. The best example of this approach comes from Ricciardi and Atkinson (2004) who conducted a meta-analysis on invasive species across a number of aquatic systems to see if species that had high impacts tended to be more taxonomically novel in their recipient community than those non-natives that showed little to no detectable impact. Clearly this approach conflates taxonomic distinctiveness with functional distinctiveness (F ), but it is still a useful first approximation as to whether invaders that are very different from the recipient community have larger impacts. In general, Ricciardi and Atkinson found that non-native species with high impacts were more likely to be from genera that were novel in the invaded community than nonnatives with no detectable impact (Kolar & Lodge 2001; Lockwood et al. 2001; Shea & Chesson 2002). Ricciardi follows many other researchers in highlighting the importance of ecological attributes and functions of non-native species (Moller 1996; Simberloff & Von Holle 1999; Williams et al. 2001) as well as attributes of the recipient community and the environment (Davis et al. 2000; Kolar & Lodge 2001; Huston 2004) in determining invasive species impacts. These researchers have suggested that non-native predators are more likely to produce native species extinctions than are non-native competitors. Similarly, non-native species that greatly alter their environment to favor their own survivorship and reproduction often produce more impacts because they change the ecological playing field for co-occuring natives. For example, beavers (Castor canadensis) on Tierra del Fuego have rapidly altered forest stand dynamics in a biome where woody accumulation is naturally extremely slow. Their dams have also changed water relations that subsequently altered vegetative communities with rippling effects on nesting bird communities and small mammals.
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The behavioral and social structure of invaders relative to native communities can also be important determinants of degree of impact. Moller (1996) points out that social organisms have an advantage as invaders because they have individual-level and colony-level responses to their environment that occasionally allow them to swamp native biotic resistance and thus achieve rapid rates of geographical range expansion. Because quantification of ecological function (F ) and composition of the recipient community (C) may be difficult to compare across systems, Ricciardi suggests that invasion theory can profit from combining models that predict impact from abundance with models that predict abundance from environmental or abiotic factors. In other words, by looking across a variety of habitat types and ecosystems and examining how environmental variation affects the abundance of invasive species and how invasive species affect native biota, we can use environmental variables or gradients to predict invasive species impacts. Huston (2004) has focused on this interaction by looking at how environment affects species establishment and reproduction and then looking at how the environment interacts with a species’ ability to affect its neighbors. He calls his approach a “dynamic equilibrium model of species diversity” (Huston 1979) and applies it to invasions to address both establishment and impacts. He finds that establishment of plant invaders is expected to be highest in low productivity, undisturbed sites (which is also true for native species), but that impacts on native species should be highest in productive environments (Huston 2004). Perhaps the largest issue with both the Ricciardi (2003) and the Parker et al. (1999) formulations lies in their oversimplification of impacts. The linear form of the equations implies a straightforward one to one relationship among the variables, which, as both sets of authors acknowledge, is unlikely to be true. For example, density dependence among taxa may lead to negative correlations between abundance (A) and per individual effect (E), resulting in a non-linear dynamic (Parker et al. 1999). A linear equation is also unlikely to capture metapopulation dynamics that may lead to positive correlations between the range of a non-native species (R) and its per individual effect (E) (Parker et al. 1999). Similarly, positive interactions between species resulting in synergisms (Box 8.2) will lead to non-linearities in Ricciardi’s (2003) formula. One side effect of these non-linearities is that, as more species are introduced with more resultant interspecific interactions, the severity of impacts become harder to predict and measure (Ricciardi 2003). One of the main values of these formulas, though, is that, despite their simplicity, they have forced the ecological community into a new awareness of the complexity of invasive species impacts.
The Tens Rule This complexity is in marked contrast to the most common impact formula, Mark Williamson’s (1996; Williamson & Fitter 1996a) oft-repeated Tens Rule, which emphasizes the low incidence of high impact invaders. The Tens Rule states that only 10% of species transition through each stage of the invasions process. In other words, we should expect approximately 10% of the introduced species to establish in the new environment, 10% of the established non-natives to spread, and only 10% of the spreading invaders to have impacts on native biota. Williamson’s (1996) transition
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stages are somewhat different from ours. They are (i) appearing in the wild, (ii) establishing, and (iii) becoming invasive (defined as having economic effects, or becoming “pests”). In both his scheme and ours, though, the impact stage requires species to make three transitions. According to the Tens Rule, after these three transitions, only one introduced species in a thousand should cause problems in the recipient region. Williamson found that invasive angiosperms (1993) and pines (1996a) in the United Kingdom and a variety of other organisms globally (1996b) fit the Tens Rule. A number of other researchers have found similar results (Lonsdale 1994; Boag & Yeates 2001). The Tens Rule thus emphasizes that few transported taxa should go on to have large impacts. Unfortunately, the Tens Rule is potentially misleading. The actual percentage of species that transitions between each invasion stage is unlikely to be exactly 10%, and error on either side of this rule of thumb can lead to very different final predictions about the number of non-native species that produce impacts. In fact, allowing the margin of error around 10% to encompass everything from approximately 5% to 20%, as Williamson and many others have done, means that at each transition the error can be a factor of four, or a 64-fold difference in the final fraction of introduced species that will go on to produce negative impacts (4 × 4 × 4; Manchester & Bullock 2000). A recent exploration of the Tens Rule is illustrative here because it demonstrates that the rule sometimes poorly fits empirical data, and that this fit becomes increasingly poor as we progress through invasion stages towards impact. Jeschke and Strayer (2005) examined data for all vertebrate introductions between North America and Europe and found that far more than the expected percentage of species established and spread. Categorizing by taxon, they looked at the total percentage of European fauna to be introduced into North America and the percentage of North American fauna introduced to Europe. They found that between 2% and 11% of the taxa from each continent had been transported and released in the other continent, as predicted by the Tens Rule. However between one third and three quarters of these introductions successfully established. Once a non-native species was established, its likelihood of spread differed even further from the Tens Rule than in the previous stage, with spread probabilities over 50% for all groups except North American birds in Europe, which had approximately a 25% chance of spreading. Although Jeschke and Strayer (2005) did not examine impacts, large success rates at these earlier stages suggests that the number of non-native species with impacts may also be far higher than predicted by the Tens Rule. The two largest problems with the Tens Rule come from misapplications. As several researchers have pointed out, the Tens Rule was never meant to be used as a quantitative tool for prediction or risk analysis (Ehler 1998; Manchester & Bullock 2000; Williams et al. 2001). In addition, the rule focuses on the number of invasive species as opposed to the magnitude and character of the invasive species. It is not possible to use the Tens Rule like the Parker or Ricciardi formulas to identify important invaders or predict impacts. Instead, what the Tens Rule says most clearly is that we do not know which species will have large impacts, thus highlighting the idiosyncratic and unpredictable nature of invasions.
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A cross-stage impact formula We would like to build on the Parker et al. (1999) and Ricciardi (2003) formulas by adding some of the complexity implicit in our approach to this book, specifically an acknowledgment of the stage-based nature of the invasion process (Lodge 1993; Kolar & Lodge 2001). The factors that allow an invasive species to succeed during transport, establishment, or spread are likely to influence the impact stage, as well. We could think of success factors (Fi ) within each invasion stage that might affect impact. We can thus envision an alternative formulation for impacts where, I = Ft ∗ F e ∗ F s ∗ E
(10.3)
where I is the per capita impact, Fi represents the invader’s success at the ith stage of invasion with the lower case subscripts representing transport (t ), establishment (e) and spread (s), and E is a measured effect per individual invader. For example, consider the eventual impact of a hypothetical non-native plant. Imagine that this invading plant possesses a long seed stage that allows propagules to survive transport (Ft ), plasticity in germination conditions that enhances its establishment across multiple sites (Fe ), and high reproductive output combined with hooked seeds that facilitates its rapid spread (Fs ). Before the impact stage, each of these factors plays a role in allowing the invader to rapidly occupy a large total area and attain high densities, and as the Parker formula suggests, these are key elements to the degree of impact produced. Note that we have basically expanded on Ricciardi’s (2003) ecological function factor (F ) by breaking a single factor down into particular features that act at each stage of the invasion process. Notice also that Equation 10.3 has both an effect (E) and an impact (I ), and that both of these are per capita measurements. Parker et al.’s (1999) formula uses abundance (A) and range (R) primarily as conversion factors to multiply a single individual’s effect (E) into the effect of an entire invasive species (I ). Our effect (E) is somewhat different. Instead it encompasses conversion factors that translate the units of success factors at each stage into effects on the native biota at any biological level of interest. Basically, we are trying to translate the features that lead to success in a particular stage into their relevance for impact. Notice that we could also include range (R) and abundance (A) from the Parker formula within this Equation 10.3, in which case our impact (I ) would also be a measure of species-level effects. A potentially more interesting offshoot of this formula is the way it makes us focus on the invasion stages and the factors that facilitate transition through those stages. Equation 10.3 is extremely simple with linear effects across stages. It is unclear how often the latter stages of invasion are facilitated (or inhibited) by the circumstances surrounding earlier stages; however, reviews by Kolar and Lodge (2002) and work by Cassey et al. (2004b) suggest that the importance of various life history traits varies substantially across invasion stages, and that the same trait can often work for a species in one stage and against it in another or can work for a species across more than one stage. Thus across-stage interactions add non-linearities in the degree of impact that go beyond synergistic interactions between species or metapopulation dynamics, as described above in the context of the Parker and Ricciardi formulas.
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As a worked example of this cross-stage synergism, we review the introduction of Argentine ants (Linepithema humile) to North America. Argentine ants (L. humile) are native to South America and have established many non-native populations throughout the world via their tendency to associate with humans and thus be transported accidentally in ship cargo (Suarez et al. 2001). They do best in Mediterranean climates and have been implicated in the decline of several native plants and animals within their non-native range. In their native range, Argentine ants are highly social in that they live in colonies and collectively defend the colony against all non-member individuals. Several distinct colonies may exist in proximity to one another; however, the boundaries between these colonies are very distinct (i.e., multicoloniality). This behavior is thought to limit population density in ants (and other animals) since the energy spent on territorial defense is not being allocated to growth or reproduction. In California, Argentine ants are much less aggressive toward each other. Workers and queens move between colonies regularly, so that these ants essentially live in gigantic supercolonies (i.e., they are unicolonial). Argentine ants live in much higher density in California than they do in their native range. It is this high population density, rather than their super-competitive or predatory ability as individuals, that allows them to expand their geographic range in California and leads to their intense impacts on native species (Holway et al. 1998). Tsutsui and colleagues (2000, 2001, 2003) present a convincing argument that the loss of intraspecific aggression within the California population is due to the significant loss of genetic diversity this population experienced at its founding. Tsutsui et al. (2000) compared the California population to the native one and found that, in the California population, nearly half the native alleles were missing, and heterozygosity was reduced by over 60%. Genetically based cues let Argentine ants of one colony recognize their nest mates. An encounter between two individuals that are not closely related will often lead to combat and death. This correlation between aggression and genetics has lead to a great deal of genetic structure and variability in the Argentine ant native range. Because only a small fraction of this genetic variability made it to California via the founding individuals, the majority of ants in California recognize one another as nest mates even though they may physically be living quite far from one another. This allows the formation of supercolonies, the achievement of high population densities, rapid geographic spread through colony budding, and a high level of impact on native species. In this case, the successful spread and high impact of Argentine ants in California is a function of the circumstances surrounding the transport of founding individuals such that: I = Ft ∗ Fe ∗ Fs (Ft ) ∗ E(Ft )
(10.4)
where Ft and Fe describe transport and establishment success, while spread (Fs ) and per individual effect (E) are both functions of the circumstances surrounding transport (Ft ). The main point we wish to stress is that there is a very broad spectrum of ways in which an invader’s ecological or genetic characteristics may interact with novel
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communities to produce negative impacts. A single linear relationship, therefore, will never completely encompass the range of potential invasive species impacts. Instead, allowing ourselves to recognize a wide range of potential interactions and functions that affect impacts allows us to formulate testable hypotheses about these impacts. We thus recommend the inclusion of cross-stage factors in our quest to find ways of understanding and measuring ecological impact from invasions.
Summary As we have detailed in many different ways, recognizing invasive species impacts is not straightforward, and comparing the magnitude of these impacts is even more complex. Because impacts are so varied and influenced by such a broad range of factors, it is important to formulate clear, quantitative relationships that help us to develop testable hypotheses. This is potentially a very rich area for invasive species research, but it is critical that the measurement of impacts be transparent in all discussions both because of the way that different ecological levels interact to increase the degree of impact and because lags in effects may change results in future years. Reporting invasive species impacts in the scientific literature will also allow us to accumulate enough information for meta-analyses at specific spatial scales, hierarchical levels, habitats, and trophic levels. All of these types of information will improve our ability to generate invasion theory and manage invasive species.
Further complexities Huston, M. A. 2004. Management strategies for plant invasions: manipulating productivity, disturbance, and competition. Diversity and Distributions 10: 167–178. Parker, I. M., D. Simberloff, M. W. Lonsdale, K. Goodell, M. J. Wonham, P. M. Kareiva, M. H. Williamson, B. Von Holle, P. B. Moyle, J. E. Byers, and L. Goldwasser. 1999. Impact: toward a framework for understanding the ecological effects of invaders. Biological Invasions 1: 3–19. Ricciardi, A. 2003. Predicting the impacts of an introduced species from its invasion history: an empirical approach applied to zebra mussel invasions. Freshwater Biology 48: 972–981.
Companion papers Davis, M. A. 2003. Biotic globalization: does competition from introduced species threaten biodiversity? Bioscience 53: 481–489. Hoopes, M. F., and L. M. Hall. 2002. Edaphic factors and competition affect pattern formation and invasion in a California grassland. Ecological Applications 12: 24–39. Levine, J. M., M. Vila, C. M. D’Antonio, J. S. Dukes, K. Grigulis, and S. Lavorel. 2003. Mechanisms underlying the impacts of exotic plant invasions. Proceedings of the Royal Society of Sciences, Series B 270: 775–781.
11 Evolution of Invaders
Overview One of the defining processes of invasion is the initial entrainment of individuals into a transportation vector and the subsequent release of these individuals into a locale to which they are nonnative (Chapter 2). The release of these individuals (propagules) in the non-native range and their subsequent struggle to realize positive population growth defines the founding process. Depending on the vector by which individuals are moved, the founding population can be large or small and may be supplemented by later independent release events. As we reviewed in Chapter 4, high propagule pressure is a principal determinant of establishment success because it allows the incipient population to overcome behavioral limitations to population growth and to avoid the ill-effects of stochastic environmental forces (e.g., harsh weather events). The size of the propagule and the subsequent rate at which the incipient population grows (if at all) also determine the genetic diversity of the non-native population. The genetic changes that occur at the time of founding may influence establishment success, but they may also determine the rate of population expansion, the level of impact on native species, and the probability that the non-native species will evolve in response to its novel biotic and abiotic environment. This chapter will introduce case studies of how genetic changes influence various stages of the invasion process. If you are unfamiliar with population genetics, keep your finger on Table 11.1 as it provides definitions for the various terms used.
Founding process The role of the founding process in evolution has received a great deal of attention over the decades. Debate around the relative importance of selection and drift and around the predominance of additive or higher-order genetic interactions in evolution
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Table 11.1 Glossary of terms used in this chapter. Term
Definition
Additive genetic variance
The proportion of genetic variation that results from the sum of all individual genes influencing a trait.
Allele
Any one of a number of alternative forms of the same gene occupying a given locus on a chromosome.
Dominance
The situation where one allele is preferentially expressed when in the presence of other alleles.
Drift
The random loss of alleles that arises through the mating and production of offspring.
Effective population size
The number of individuals in a randomly mating population that would produce the same amount of genetic exchange as is found in the actual population.
Epistasis
The control of a phenotype by two or more interacting genes.
Fitness
A measure of an individual’s ability to propagate its genes either through high survival or reproduction relative to other individuals in the population.
Genotype
An allele combination that differs among organisms and that, in combination with the environment, produces a particular trait or phenotype.
Geographic genetic structure
Genetic differences among populations of the same species across a physical region.
Heterozygosity
The fraction of individuals in a population that are heterozygous for a particular locus; heterozygotes have different alleles at a single locus.
Inbreeding
Breeding between close relatives likely to share alleles.
Locus
The position of a gene on a chromosome.
Phenotype
The expression of a genotype often in physical appearance, trait, or behavior of an organism in a way that varies between individuals. Selection acts on phenotypes rather than on genotypes.
Definitions follow Hartl (1988).
goes all the way back to Sir Ronald Fisher (1930) and Sewall Wright (1932). Fisher promoted the theory that evolution is driven primarily by natural selection, mostly through its effect on additive genetic variation. Wright thought that changes in allele frequencies associated with population bottlenecks provided a critical mechanism for evolutionary change. He also recognized that founding events reduce genetic diversity. Thus, even though bottlenecks may be necessary for evolutionary change, the attendant reduction in diversity curbs the evolutionary process because there are fewer genes and gene combinations for natural selection to act upon after a bottleneck. The debate on
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which theory is correct persists to this day although most evolutionary biologists agree that both processes play a role in evolution (Brodie III 2000). What does any of this have to do with invasion ecology? One of the motivations for the seminal Baker and Stebbins (1965; Chapter 1, this volume) volume on invasive species was to consider how non-native species could answer such basic evolutionary questions as posed by Fisher and Wright (Vellend et al. 2005). Indeed, the study of nonnative species is increasingly seen as key to exploring evolutionary theories (Rice & Sax 2005; Wares et al. 2005). From the perspective of those interested in managing invasions, however, the genetic changes that accompany the founding process set the stage for the fate and impact of non-native populations. The key questions in this regard are (i) how often does genetic variability decline at the time of initial population establishment? (ii) Does low genetic variability prohibit incipient populations from becoming self-sustaining (i.e., establishing)? (iii) Is there enough genetic variability left after the founding event to allow key phenotypic traits to evolve that allow the population to spread and perhaps cause ecological impact?
Losses and gains in genetic variability via transport mechanisms Recall that, by definition, non-native populations are founded by a set of individuals that were transported out of their native range and released in a new non-native locale. As each of these transported individuals carries a unique sub-set of the genes present in the entire native population, the set of individuals that found a non-native population usually contains only a fraction of the total genetic diversity (i.e., total set of alleles) of the donor population (Nei et al. 1975; Lee 2002). Thus, each non-native population may harbor many fewer rare alleles than their donor sources because the more rare alleles in the source population were not “captured” by the subset of transported founder individuals. The degree to which genetic diversity is eroded via such a population bottleneck depends on propagule size (i.e., number of founder individuals), the propagule(s) source location, and how often the transportation vector deposits new propagules into the incipient population. Since we know from Chapters 2 and 3 that there is wide temporal and spatial variation in these qualities across transport vectors, it is worthwhile asking how common the loss of genetic diversity is in founding non-native populations. There have been several studies on the genetic diversity of non-native populations relative to their donor sources (reviewed in Cox 2004). To understand the process, it is easiest to go first through some of the research documenting a reduction in genetic variability (Baker & Moeed 1987; Holway et al. 1998; Bartlett et al. 2002). For example, Baker and Moeed (1987) compared the genetic diversity of non-native populations of the common myna (Acridotheres tristis) with its native source population in southeast Asia. The common myna is a large omnivorous bird that has been intentionally introduced to New Zealand, South Africa, Hawaii, Australia, and Fiji (among several other places). These populations were all introduced in the mid- to late-1800s, and all, except the introduction to South Africa, involved individuals taken directly from the native
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Table 11.2 Comparison of genetic variability in native and introduced bird populations. Species
Area
ηp
ηl
Genetic parameters
Introduction history
A
P
HE
Intro
Abun.
n
# times
Sturnus vulgaris
Native NZ
6 6
24.0 24.0
1.42 1.35
28.5 33.8
0.033 0.043
– 1867
– 1878
– 653
– 13
Acridontheres tristis
Native NZ AU HA FI SA
7 5 2 1 2 2
28.4 57.4 43.9 37.6 25.9 38.3
1.43 1.24 1.30 1.20 1.30 1.15
31.5 19.5 23.1 18.0 20.6 12.9
0.060 0.042 0.060 0.060 0.050 0.030
– 1862 1862 1865 1888
– 1912 1883 1879 – 1930
– 100 252 – – 6
– 2 5 – – 2
Fringilla coelebs
Native NZ
10 8
28.7 26.1
1.34 1.39
24.5 28.3
0.048 0.066
– 1867
– 1870
– 377
– 13
Carpodacus mexicanus
Native E-USA HA
6 6 4
40.6 38.5 20.0
1.48 1.37 1.33
27.4 23.7 21.0
0.045 0.042 0.055
– 1940 1860s
– – –
– – –
– – –
Carduelis chloris
Native NZ
14 7
33.6 35.1
1.19 1.22
14.1 16.8
0.025 0.025
– 1862
– 1868
– 64
– 5
Passer montanus
Native USA AU
2 3 1
55 83 15
1.45 1.33 1.3
33.3 23.9 23.1
0.098 0.078 0.074
– 1870 1863
– – –
– 20 –
– 1 –
Passer domesticus
Native NZ AU
14 4 3
60.7 51.4 36.7
1.63 1.42 1.50
36.9 36.2 33.3
0.097 0.080 0.103
– 1859 1863
– 1869 1906
– 418 494
– 7 8
Note: ηp = number of populations, ηl = average number of individuals scored per locus, A = average number of alleles per locus, P = percentage polymorphic loci, HE = average expected heterozygosity, n = minimum number of introduced individuals, # times = minimum number of introductions. AU = Australia, E − USA = eastern USA, FI = Fiji, HA = Hawaii, NZ = New Zealand, SA = South Africa. Source: Reprinted from Merila, J., M. Bjorklund, and A. J. Baker. 1996. The successful founder: genetics of introduced Carduelis chloris (greenfinch) populations in New Zealand. Heredity 77: 410–422, with permission from McMillan Publishers Ltd.
donor range. The common myna individuals that founded the South African population were transported from another location of introduction (Mauritius) that itself was founded by individuals from southeast Asia. Thus, the South African population most likely went through two population bottlenecks before establishing. Because the release of mynas was a planned event, we have good records of the number of founding individuals initially released (Table 11.2). Baker and Moeed (1987) collected individual mynas from New Zealand, South Africa, Australia, Hawaii, Fiji, and several native locations within India. Using isozyme
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electrophoresis of 39 loci, they compared the average heterozygosity, number of alleles per locus, and the percentage of polymorphic loci between all populations. They found that the average heterozygosity and the percentage of polymorphic loci were lower in each of the non-native populations compared to the source populations. Also, the non-native populations showed an 18% reduction in the number of alleles per locus. All of the “lost” alleles were rare in the native donor populations. As expected, based on our knowledge of the origin of the founders, the South African population showed the greatest reduction in genetic diversity according to all the above metrics. Additionally, when Baker and Moeed (1987) compared genetic similarity between the non-native populations, they found substantial differentiation. In some cases, the differences in the genetic make-up between two non-native populations rivaled the degree of genetic difference observed between recognized subspecies of birds. Baker and Moeed (1987) suggested that this degree of between-(non-native) population diversity was due to the effects of genetic drift and random subsampling of the available donor genetic diversity. In other words, each non-native population had a set of founding individuals that carried with them a unique sub-set of the alleles available in the source gene pool. Subsequent to release, some of these alleles were lost due to drift. The end result is a set of non-native populations that share few alleles between them. This seems to be a relatively common phenomenon when one species generates several successfully established non-native populations (Merila et al. 1996; Lambrinos 2004), and it may set the stage for future alternate adaptive pathways (Huey et al. 2000; Box 11.1). Because of the detailed records often kept on bird introductions, there have been several studies conducted on possible reductions in genetic diversity similar to that of Baker and Moeed (1987). Merila et al. (1996) reviewed these studies and showed that other non-native bird populations show a similar pattern to that seen in the common myna (Baker & Moeed 1987). Contrary to the myna invasion and to expectations provided by classical theory, though, there are also several cases where the genetic diversity in the non-native populations is equal to or higher than that in the native sources. Merila et al. (1996) make sense of these mixed results by relating them to the size of the founder population, the rate at which these populations increased after initial release, and the number of documented release events. Figure 11.1 depicts the relationship between these three factors and measures of genetic diversity of non-native bird populations. As suggested by Nei et al. (1975), the smaller the founding population and the slower the rate of population increase after initial release, the greater is the loss of genetic diversity. In addition, loss of genetic diversity is minimal when individuals continue to enter the founding population through time, as might be the case for repeated introductions for biocontrol, aesthetics, or similar reasons. A broader survey of the literature on the genetic diversity of non-native populations suggests that in most cases non-native populations lose very little genetic diversity (Novak & Mack 2005; Wares et al. 2005). Based on these reviews, we suggest the following three “rules of thumb” about the relationship between genetic diversity and the circumstances surrounding initial introduction: (i) genetic diversity is likely to be lower than that in source populations if transport vectors select only a small fraction of
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228 0.04 0.03
A1
B1
C1
A2
B2
C2
A3
B3
C3
Diff H
0.02 0.01 0 0.01 –0.02 20 15
Diff P
10 5 0 –5 –10 0.3
Diff A
0.2 5 0 –10 0
200
400
600
800 0
Number of introduced individuals
10
20
30
40
50
60 0
Time before common (years)
5
10
15
Number of introductions
Figure 11.1 The effect of propagule size, time since introduction, and propagule number on genetic variation. Each graph represents the difference between variation in the native environment and in the non-native environment. A, P, and H refer to alleles per a locus, polymorphic loci, and heterozygosity as in Table 11.1. Reprinted from Merila, J., M. Bjorklund, and A. J. Baker. 1996. The successful founder: genetics of introduced Carduelis chloris (greenfinch) populations in New Zealand. Heredity 77: 410–422, with permission from Nature Publishing Group.
the native variation with limited sampling in both time and space; (ii) genetic diversity may be higher in the non-native range if the introductions come from a broad sample across native donor populations that show high degrees of geographic genetic structure; and (iii) if the circumstances encountered after initial release keep the population size low for long periods, even the release of many individuals will not forestall the loss of genetic variation. We elaborate on these generalizations below.
Lower genetic diversity from limited propagule sources Genetic diversity in the non-native population is likely to be reduced when the transport vector is sporadic in time and space. Only with the most severe of population bottlenecks (<10 individuals released) and only if there is virtually no subsequent release of nonnative individuals into the founding population, should we expect to see genetic erosion within a non-native population (Nei et al. 1975). Although we have already given examples that follow this pattern, we suggest that only a fraction of current transport vectors will result in the release of so few individuals that lowered genetic variation would be expected. Examples of transport vectors likely to result in losses of genetic
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Box 11.1 Superfly One of the more astonishing and informative examples of rapid evolution of a nonnative species comes from studies of the fruit fly, Drosophila subobscura, within its non-native range of North and South America (Huey et al. 2000; Huey et al. 2005). D. subobscura is native to the Old World, and within its native range the species shows a marked cline in wing length with larger wings associated with colder climates. In the 1970s D. subobscura was introduced into South America, initially showing up in Chile. In the 1980s D. subobscura was introduced into North America, initially showing up in Oregon. Huey and colleagues wanted to know if wing clines had evolved within these two non-native regions, and, if so, what aspect of wing morphology changed to produce overall larger or smaller wings. Through genetic analyses they found that both non-native populations (i.e., in North and South America) were founded by individuals from the same Mediterranean parent population (southern part of their native range); thus the two non-native populations were founded by small-winged individuals and could be considered replicates of one another (Huey et al. 2005). Despite each non-native population of D. subobscura having been in existence for less than 30 years at the time of Huey’s study (2005), he found that wing size increased as latitude increased in both the North and South American populations (Huey et al. 2005 and references therein). Furthermore, the slope of the relationship between wing size and latitude was the same across non-native populations, at least amongst females (Fig. 11.2). There are other examples of morphological clines evolving within a species’ non-native range (e.g., Johnston & Selander 1973), indicating that such macroevolutionary patterns may be re-created over a remarkably short period of time. The particularly interesting part of Huey’s work, though, lies in his noticing that larger overall wing size was achieved through changes in different aspects of wing shape in North versus South American populations. For example the basal portion of the wing does not increase with latitude among North American males and females; however, this aspect of wing morphology shows substantial increases along latitudinal clines in the South American population and within D. subobscura’s native range. In North American populations, large wing size was achieved through increases in other wing dimensions. This led Huey et al. (2005) to conclude that, although the evolution of larger wing sizes with increasing latitude is largely predictable, the way that cline is developmentally achieved is decidedly unpredictable. Such insights into the mechanics of evolution are hard to achieve without having multiple populations of the same species introduced into distinct areas, such as is the case when one species establishes non-native populations on several continents. It is still largely unknown, however, how such subtle differences in evolution translate into differing rates of geographical range expansion or ecological impact across these non-native populations. This is yet another place where students of invasion ecology have much to contribute.
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Box 11.1
Continued
Wing size (PC1)
0.1
0.0
–0.1
–0.2 30
35 40 45 50 Latitude (degrees North or South)
55
Figure 11.2 Changes in wing size of fruit flies (Drosophila subobscura) introduced to North and South America. Wing size increases with latitude in Old World (circles) populations, and in non-native North American (triangles) and South American (inverted triangles) populations. Solid and open symbols represent females and males respectively. PC1 refers to the first Principle Component of three wing dimensions. Reprinted with permission from Huey, R.B. G.W. Gilchrist, and A.P. Hendry. 2005. Using invasive species to study evolution: case studies with Drosophila and salmon. Pages 139–164 in D.F. Sax, J.J. Stachowicz, and S.D. Gaines, eds. Species Invasions: Insights into Ecology, Evolution, and Biogeography. Sinauer Press, Sunderland, MA.
diversity are the release of species that are aesthetically pleasing to one (or a few) people, species that are released as a trial (with no follow-up releases) for biocontrol or game augmentation, or species that are very sporadically transported as by-products (Chapter 2). These criteria are not mutually exclusive.
Elevated diversity from intermixing genetically distinct propagules We should expect genetic variation to be higher, not lower, in the non-native range if the following two conditions hold true. First, the source population must show a relatively high level of geographic genetic structure. Second, the transport vector must effectively “sample” across the range of this variation, probably through several independent
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transportation and release events. If these conditions are met, novel genetic material will continually arrive in the non-native population. In essence, human-mediated transport of non-native individuals brings formerly distinct lineages together within the non-native range. Within the academic field of population genetics the resultant population is called an “admixture” (Wares et al. 2005). Because this result differs from the predictions of classic population genetics (see above), we will work through an example of an admixture, relying heavily on Kolbe et al.’s (2004) excellent research on the brown anole (Anolis sagrei). The brown anole is native to the Caribbean islands and has well-established nonnative population in Florida, as well as in other areas of the world. Many of these non-native populations of the brown anole have high rates of population growth and geographic spread. They also compete with and prey on native lizards. Brown anoles first appeared in the Florida Keys (extreme southern tip of Florida) in the late 1800s, and, like many non-natives, they did not appreciably expand their range for another 50 years (see Chapter 8 on lag times). Between the 1940s and the 1970s, the brown anole appeared all over Florida. These new populations were possibly a result of spread northward from the Florida Keys combined with more recent introductions directly into northern Florida. Kolbe et al. (2004) compared the genetic variation of brown anoles in Florida to the variation seen in the native Caribbean range. Their assays aimed to identify the source locations for the non-native populations in Florida and to assess the loss of genetic variation. They found that the genetic diversity of brown anoles in the native range was highly geographically structured. Most variation occurred across populations with little variation within each population. This made it relatively easy to assign a source location for the non-native populations. They found that Floridian brown anoles originated from eight native source populations, which probably means that the anole was introduced into Florida on (at least) eight independent occasions. Furthermore, individual anoles that originated from very different native locations were regularly found mixing together within the non-native populations of Florida. Although the individuals in Florida were quite different genetically, they still readily interbred. Thus, the genetically mixed non-native population shows much more withinpopulation variation than can be seen within any of the source populations. There are other examples of such increased variation in non-native populations (e.g., Novak et al. 1993; Martel et al. 2004), and they tend to come from species that have been introduced multiple times from multiple native sources with high degrees of genetic structure across the source range. Given that geographic genetic structure is fairly common among species that are widespread in their native range and that individuals from these species are often transported and released outside their native range (Blackburn & Duncan 2001b), seeing increased diversity within the non-native range may be much more common than previously suspected.
Post-release conditions and genetic diversity The conditions encountered after establishment will influence the amount of genetic diversity observed in the non-native range in two ways. First, even when large numbers
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of individuals are released, a loss of genetic diversity can occur if only a fraction of these individuals successfully reproduces and thus passes on genes to the next generation. This bias in mating creates differences between the evolutionarily important portion of the population and the actual population size. The “effective population,” which is the number of individuals in a randomly mating population that would produce the same amount of genetic exchange as is found in the actual population, may be a fraction of the actual population size (Hartl 1988). Effective population size is determined by a variety of factors including Allee effects (see Chapter 4), an unequal sex-ratio, or a mating system that allows only a few individuals to parent most of the offspring (e.g., harems or selfing plants, Hartl 1988). Thus, we should expect to see some variation in the degree of genetic erosion across non-native species according to life histories (Novak and Mack 2005). Second, a loss of genetic diversity can occur if the population grows slowly over a long period of time, or experiences a sharp reduction in numbers of individuals postrelease (Hufbauer et al. 2004). The key here is understanding that genetic diversity can erode well after the date of initial propagule release via the same process that (potentially) limited genetic diversity in the first place; only a few individuals are alive to reproduce and these individuals carry a subset of all available genetic information. If low population sizes persist for several generations, genetic diversity will erode still further through inbreeding and drift. Non-native populations can be kept at low numbers for a variety of reasons, including competition with individuals of other species (native or non-native), heavy predation pressure, or maladaptation to local conditions. Sudden environmental catastrophes (e.g., flood, freeze, or fire) within the non-native range can also drive population size down. We discuss the mechanics of these forces within Chapters 4, 5, and 6.
Genetics and post-release success From the above we can broadly conclude that genetic diversity of non-native populations is not necessarily lower than that of the native source population, and indeed diversity sometimes is higher. Do increases or decreases in genetic diversity have any bearing on the fate of the non-native population after release?
Evidence that loss of genetic diversity influences invasion success There is plenty of reason to suspect that losses of genetic diversity, when they occur, will determine the fate of non-native populations. Population bottlenecks and genetic drift can result in the fixation of deleterious alleles and a reduced capacity to adapt to changing conditions. The fixation of deleterious alleles may lead to increased mortality or decreased fecundity within the non-native population, which in turn should lead to declining rates of population increase. The observation that many non-native populations experience lag times in population growth and spread (Chapter 8) has led many ecologists to speculate that introduced populations may lack key traits that must arise
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through evolution or arrive in subsequent introductions (Ellstrand & Schierenbeck 2000). Although there are examples of the deleterious effects of low genetic diversity relative to threatened native populations, there are very few studies that tie reduced genetic diversity to the failure of a non-native population to establish a self-sustaining population (although see Ahlroth et al. 2003 for an exception). We should point out, as we did in the establishment section, that it is very difficult to pinpoint the cause of negative results. It is just as problematic to determine whether genetic diversity plays a role in failed introductions as it is to identify any factor that keeps a non-native population from establishing. There is some evidence that a lack of genetic diversity will inhibit the geographic expansion of a non-native population. For example, Goodwin et al. (1994) and Saltonstall (2002) provide examples of how the introduction of a novel genetic strain allowed Phytophthora infestans (the cause of potato late blight) and Phragmites australis (common reed) respectively to expand their non-native ranges. In both cases the novel genetic strain was introduced long after populations of the non-natives had established but failed to spread in their new ranges. Similarly, the red imported fire ant did not spread widely and cause much ecological damage until the polygyne (multi-queen) form arrived in the United States in the 1970s (see Chapter 9). Krieger and Ross (2002) showed that the difference between the polygyne and monogyne forms is the presence of one allele at a single locus. The allele that confers polygyny apparently did not arrive in the United States until a secondary release of ants occurred in the 1970s. These examples suggest that a lack of diversity, and more specifically the lack of a rare allele, may heavily influence the ultimate geographic range and impact of a non-native species.
Evidence that hybridization influences invasion success Although our understanding of the role of genetic diversity in invasion success is still developing, we have a good deal of insight into the influence of hybridization on the success of invading species (Williamson 1996). In Chapter 9 we reviewed cases where non-natives hybridize with natives. We pointed out that these interactions often result in negative impacts on the native species through reduced fitness or introgression. Here we shift focus slightly and discuss evidence that interbreeding between formerly distinct populations (i.e., admixtures) or between species, can give non-native populations the evolutionary momentum to become invasive. This evolutionary momentum is represented by the novel recombination of genetic elements, or an increase in overall genetic or phenotypic variability. This genetic change often provides the trait needed for rapid population growth and range expansion within the non-native habitat, or for increased efficiency when interacting with other species (i.e. competitors, predators, prey, or disease). Ellstrand and Schierenbeck (2000) found 18 examples representing 9 plant families where there was molecular evidence that invasiveness was preceded by hybridization. All of these examples represented spontaneous hybridization. They did not include
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species that escaped from cultivation. They also found an additional 24 examples for which morphological evidence (but no additional evidence) suggested that invasion was preceded by hybridization. The majority of these plants were herbaceous, perennial, or biennial plants, and all of them grew in areas characterized by human disturbances (Ellstrand & Schierenbeck 2000). The pattern of perenniality is consistent with reviews of spontaneous hybridization events in general (i.e., without regard to invasions, Ellstrand et al. 1996) and likely reflects the longer persistence times of perennials (as opposed to annuals) and their energetic investment in vegetative rather than reproductive growth (Ellstrand & Schierenbeck 2000). The association with disturbance almost certainly reflects the fact that human activities were responsible for transporting and releasing the non-native parental taxa, which makes novel contact and hybridization more likely in the vicinity of human activity. It may also reflect the ability of hybrids to adapt to the often quite novel conditions created by human activities, whereas the two parental taxa could not (Ellstrand & Schierenbeck 2000). Why do hybrids become invasive? Ellstrand and Schierenbeck (2000) suggest that hybridization can stimulate invasiveness through (i) the creation of novel genotypes, (ii) high genetic variation that permits survival under a broad range of conditions, and (iii) hybrid vigor through fixed heterogeneity or the dumping or masking of deleterious alleles. The most common hypothesis for a positive influence of hybridization on invasiveness is the generation of novel genotypes that lead to increased fitness or adaptive ability. Hybridization (of any kind) results in genetic recombination in the progeny, which itself often leads to increases in the gene pool of the population (e.g., brown anoles in Florida). Some (perhaps most) of the resulting “novel” genotypes will not be well adapted to local conditions. However, some of them may possess traits that are better suited to the local environment than the genotypes found in the parental taxa. This seems to be the case for the hybrid offspring of Spartina alterniflora and Spartina foliosa, invading San Francisco Bay as reviewed in Chapter 9. Additionally, even if hybridization does not create a superior adaptation, it may simply leave more possibilities for future adaptation. If genetic variation is positively linked to rates of adaptive evolution, then the increase in genetic variation from hybridization may, by itself, lead to invasiveness. Examples of this mechanism are harder to find because they rely on post-hybridization change and the identification of two genetic steps in the invasion process, hybridization and subsequent evolution made possible by hybridization. Finally, hybrid populations often show higher levels of heterozygosity along with hybrid vigor associated with heterozygosity. This effect is most likely at an intermediate level of differentiation between parent taxa (Fig. 11.3; Ellstrand & Schierenbeck 2000). If parent taxa are evolutionarily closely related, then hybridization does not add enough new material to lead to invasiveness. If parent taxa are too disparate, then hybridization may break up important gene complexes. Like Goldilocks, the intermediate that is “just right” may lead to higher invasiveness. In addition to the plant examples in Ellstrand and Schierenbeck (2000), there are examples of increased invasiveness with hybridization for pathogens, fish, amphibians, and birds. Brasier (2001) reviews several cases of hybridization between plant pathogens where the resultant hybrid causes more damage than either of the parental
Progeny fitness relative to that of their parents
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High
Equal
Low
Genetic distance among mating colonists
Figure 11.3 As genetic distance between mating colonists increases, so too should heterosis in their progeny – up to a point. Then progeny fitness declines as outbreeding depression becomes important. Reprinted from Ellstrand, N. C., and K. A. Schierenbeck. 2000. Hybridization as a stimulus for the evolution of invasiveness in plants? PNAS 97: 7043–7050 with permission from National Academy of Sciences, USA.
taxa. For example, a new hybrid Phytophthora species attacks alders in Europe that neither parental species could attack (Brasier et al. 1999). Similarly, Newcombe et al. (2000) suggest that Melampsora rust hybrids combine the host ranges (i.e., set of species the rust will attack) of both parental species, thus creating a type of “mega-rust” that can attack several kinds of poplars in the North American Pacific Northwest. Indeed, the intentional production of hybrid animals (e.g., salmon) for commercial purposes is the subject of intense debate (Box 11.2). Although there are examples of fish, crayfish, salamanders, and parrots hybridizing, it is not clear if these hybrids are more invasive then their parental taxa. In almost all cases when hybridization is documented, we lack direct evidence that the genetic changes led to increased fitness because we have no control data for these natural experiments. This is yet another area ripe for future research. This research will require extensive synthesis between molecular genetics and population ecology.
Local adaptation and life history evolution Individuals in non-native populations are undoubtedly constrained in their growth, reproduction, and survival by a different set of factors than individuals in the source population, and each non-native population is likely to experience a different environment. If any of these factors act in a consistent manner to cull particular phenotypes from the non-native population, we should expect to see the population evolve towards the favored phenotypic state. Although we can easily imagine this evolutionary shift will result in greater geographic range expansion and perhaps higher incidence of ecological impact, thus far most evolutionary ecologists interested in invasions have focused strictly on questions of how and why non-native populations have evolved.
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Box 11.2 Handle with care A key point of cross-over between evolution and invasion ecology is concern over the production of genetically modified organisms (GMOs). These organisms carry genes from other species (transgenes) to increase their body size, resistance to herbivory or disease, and other traits that confer increased yield. There are several concerns over the widespread production of GMOs, including the possibility that genes conferring resistance to herbivory will harm native herbivores (e.g., Losey et al. 1999). Insights from invasion ecology are called into play when considering whether transgenes may “invade” native (non-modified) populations. Such invasions involve gene flow between transgenic and wild individuals with implications for mating success and survivorship of both groups. In a series of papers William Muir and his colleagues (Muir & Howard 1999; Howard et al. 2004) provide evidence that genetically modified fish have the potential to drive wild populations extinct. Large bodied individuals are prized in commercial and recreational fishing. Thus, several genetically modified fish species have been produced with transgenic individuals growing to be several times larger than their wild counterparts. The risk that these transgenic individuals will harm wild populations is often considered low since transgenic individuals are considered “evolutionary novelties” and often show reduced survivorship in the wild (e.g., they swim more slowly or die younger). These individuals could negatively impact wild populations, anyway, if their genetic make-up confers a fitness advantage in terms of reproduction. Animals that are modified to become larger in body size are a prime candidate for such effects because male body size is frequently a sexually selected trait (i.e., bigger males are preferentially chosen as mates by females). Muir and Howard (1999) and Hedrick (2001) illustrate through simple population models that a simultaneous increase in mating success by transgenic individuals and a reduction in their survivorship can quickly lead to extinction. Essentially, females preferentially choose to mate with males that confer a reduced lifespan on their offspring. As long as large body size and reduced viability remain genetically coupled, the population steadily decreases in abundance through time. In a series of lab experiments Howard et al. (2004) confirmed the mating preference for large males and the reduced viability of young that carry the transgene in a model fish species, Japanese medaka (Oryzias latipes). They call this effect the “trojan gene hypothesis.” It is not clear how common this situation is among existing genetically modified fish or other organisms. However, it does raise the troubling possibility that tinkering with the genetic make-up of economically important species can have harmful effects on their native counterparts. Invasion ecologists were first broadly introduced to the notion that non-native populations could evolve quite rapidly through the work of Richard Johnston and Robert Selander (1964, 1971, 1973). These ecologists, based on earlier smaller-scale efforts, endeavored to test whether house sparrows (Passer domesticus) in North America had
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diverged from their native source populations in England and Germany. They found distinct clines in plumage coloration and body size across populations in North America. Individual house sparrows that lived in relatively cool and humid climates tended to be darker in plumage; whereas individuals that lived in relatively hot and dry climates tended to be paler. Similarly, house sparrows that lived in boreal climates with severe winters were larger in body size as compared to sparrows that lived in mild austral climates. These patterns conform to plumage and body size clines found among native North American birds and follow from well-accepted biogeographical patterns (i.e., Gloger’s and Bergmann’s rules, respectively). Each is based on thermoregulatory mechanisms, and the recreation of these clines within a non-native population speaks to the selective force of climatic variables (Johnston & Selander 1973). Johnston and Selander (1964, 1971) demonstrate that the degree of observed divergence between house sparrow populations in North American are of the magnitude seen between recognized native subspecies. This level of divergence was achieved within the span of just over 100 generations and perhaps less in some locations. Since the publication of their work, there have been other attempts to demonstrate the formation of geographical clines among non-native populations (Baker & Moeed 1987; Huey et al. 2000). Huey et al. (2000) present a fascinating example of a rapid generation of geographic clines among fruitflies (Box 11.1). This research indicates that such adaptive shifts in morphology are predictable based on known ecophysiological principles, but the morphological changes that produce these clines are contingent on historical conditions. There is growing evidence that non-native populations will also evolve in response to the presence, or absence, of predators and competitors. For example, Lorne Wolfe and colleagues explored the evolutionary story in Silene latifolia, a small herbaceous plant native to Europe but introduced to North America about 200 years ago where it is now a common agricultural weed (Blair & Wolfe 2004; Wolfe et al. 2004). Wolfe and colleagues found that North American (non-native) populations of S. latifolia had fewer natural enemies (i.e., fungi, herbivores, and diseases) than European (native) populations. They hypothesized that S. latifolia populations in North America would divert resources from defense to growth and reproduction as described by the EICA hypothesis (see Box 6.2; Blossey & Nötzold 1995). This change in resource allocation could make them superior competitors in their non-native range. Common garden and reciprocal transplant experiments illustrated that North American plants grew faster, produced more flowers, germinated earlier and had a higher survival rate than European plants (Blair & Wolfe 2004; Wolfe et al. 2004). The North American plants also devoted less energy toward defensive traits, such as thick fruit walls, and the production of defensive chemicals. Although there is a growing list of other studies that show a loss of defenses, there are also several studies that suggest that changes may be more complex than a simple reallocation of resources from defense to growth. For example, Siemann and Rogers (2003) compared non-native populations of the Chinese tallow tree (Sapium sebiferum) in Texas and Hawaii. The Texas populations had lost all native herbivores and grew the fastest; the Hawaiian population, which had acquired an Asian herbivore (Adoratus sinecus, the Chinese rose beetle), had intermediate growth; and the Chinese population grew the slowest (Siemann & Rogers 2003). However, the Texas
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populations had not escaped from herbivory. Instead, faster growth allowed plants from the Texas populations to compensate for root herbivory (Rogers & Siemann 2004). While plants from the Chinese population defended leaf tissues from herbivory, they were unable to recover from damage that severed roots (Rogers & Siemann 2004). Although the invader lost phytochemical defenses, the compensatory growth offered more appropriate defenses, leading to increased fitness. Such examples of non-native population divergence are not restricted to plants. Divergence between non-native insect species has proven relatively common (e.g., Via et al. 2000), and there are several examples of both morphological and behavioral shifts in vertebrate species. One of the most interesting is an example of character displacement in an introduced mongoose (Simberloff et al. 2000). In the western part of its native range on the Indian subcontinent, the small Indian mongoose (Herpestes javanicus) has two larger congeners, the ruddy mongoose (H. smithii) and the grey mongoose (H. edwardsii). In this sympatric range the small Indian mongoose is smaller, with shorter canines and skull dimensions, than in the eastern part of its range where it has no close competitors. The small Indian mongoose has been introduced to several islands in the West Indies for snake control, and the propagules for these non-native populations came from the western, smaller body-sized, sympatric portion of the native range. In the 100–200 generations since introduction, these non-native island populations have shifted to larger male size, longer canines, and larger skulls, making them more like the eastern native populations with no close competitors (Simberloff et al. 2000). This is a stunning example of character displacement, a process that is difficult to demonstrate due to the impossibility of disentangling cause and effect in stable communities in which character differences limit current competition but could have been driven by competition in previous generations. Shifts in reproductive behavior are more common but no less effective in altering interactions between species or even potentially in causing speciation. For example, Hendry et al. (2000) found that sockeye salmon (Oncorhynchus nerka) introduced to rivers and lakes in the state of Washington (USA) show habitat dimorphisms that restrict mating between genotypes that breed in rivers and those that breed in the beach habitat at river mouths and along lakes. This separation persists despite the frequent movement of riverine individuals into beach habitat. Body dimensions of adults reflect the breeding conditions of each habitat, and male body size is sexually selected. Because the beach habitat has fewer predators, less current, and deeper water, males of the beach-breeding genotype are large in comparison with river-dwelling males. On the other hand, females that breed in the beach habitat are generally smaller than their river-dwelling counterparts. The beach habitat floods less frequently than the river, and beach-breeding females do not need to be long to dig deep nests to protect their young from disturbance during flooding. These differences in body size between habitats result in low reproductive success for individuals that originate in one habitat and attempt to breed in the other. The end product is virtually no genetic mixing between the two non-native populations, thus producing reproductive isolation despite habitat overlap (Hendry et al. 2000). This separation has occurred over the span of only 56 years, or 13 generations, at a maximum. Although not tested, it would be interesting to see if similar levels of reproductive isolation have occurred between the other examples of divergence between
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non-native populations described above. Collectively, these studies demonstrate that post-establishment evolution is common and that we have not given enough credence to the notion that rapid rates of geographic range expansion and invasiveness may evolve once the non-native population has established, although there is much to explore in this regard.
Summary The genetic reshuffling that occurs when individuals are transported and released within a non-native locale can be quite extreme. We have only offered a glimpse into the range of possible outcomes of such founding events, despite the fact that much evolutionary theory could be tested using non-native populations. This genetic reshuffling can lead to higher rates of geographical spread and ecological impact. There is also good evidence that non-native populations, once well-established, can diverge from one another in morphology and life history. At least occasionally, these changes can lead to increased invasiveness in the non-native range and ecological speciation. There is much to be learned from non-native species concerning the interplay between founding events, directional selection, and ecological conditions. In return, such studies should help invasion biologists target circumstances where post-establishment evolution poses a risk to economic and ecological systems (e.g., van Klinken & Edwards 2002).
Further complexities Daehler, C. C., and D. R. Strong. 1997. Reduced herbivore resistance in introduced smooth cordgrass (Spartina alterniflora) after a century of herbivore-free growth. Oecologia 110: 99–108. Novak, S. J., and R. N. Mack. 2005. Genetic bottlenecks in alien plant species: influence of mating systems and introduction dynamics. Pages 201–228 in D. F. Sax, J. J. Stachowicz, and S. D. Gaines, editors. Species Invasions: Insights into Ecology, Evolution and Biogeography. Sinauer Associates, Sunderland, MA. Kolbe, J. J., R. E. Glor, L. R. G. Schettino, A. C. Lara, A. Larson, and J. B. Losos. 2004. Genetic variation increases during biological invasion by a Cuban lizard. Nature 431: 177–181. Lee, C. E. 2002. Evolutionary genetics of invasive species. Trends in Ecology and Evolution 17: 386–391. Wares, J. P., A. R. Hughes, and R. K. Grosberg. 2005. Mechanisms that drive evolutionary change: insights from species introductions and invasions. Pages 229–258 in D. F. Sax, J. J. Stachowicz, and S. D. Gaines, editors. Species Invasions: Insights into Ecology, Evolution, and Biogeography. Sinauer Associates, Sunderland, MA.
Companion papers Anttila, C. K., C. C. Daehler, N. E. Rank, and D. R. Strong. 1998. Greater male fitness of a rare invader (Spartina alterniflora, Poaceae) threatens a common native (Spartina foliosa) with hybridization. American Journal of Botany 85: 1597–1601.
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Ellstrand, N. C., and K. A. Schierenbeck. 2000. Hybridization as a stimulus for the evolution of invasiveness in plants? Proceedings of the National Academy of Sciences of the United States of America 97: 7043–7050. Lambrinos, J. G. 2004. How interactions between ecology and evolution influence contemporary invasion dynamics. Ecology 85: 2061–2070. van Klinken, R. D., and O. R. Edwards. 2002. Is host-specificity of weed biological control agents likely to evolve rapidly following establishment? Ecology Letters 5: 590–596. Krieger, M. J. B., and K. G. Ross. 2002. Identification of a major gene regulating complex social behavior. Science 295: 328–332.
12 Prediction, Risk Assessment, and Management of Species Invasions
Overview Throughout this book we have attempted to ground the larger field of invasion ecology within a heuristic framework while simultaneously providing an introduction to the major research arenas. Two of the larger themes we have neglected until now are whether ecologists can predict species that are likely to become invasive and whether we can adequately assess a species’ risk of causing economic or ecological harm. We have also largely ignored how science can inform on-theground management of invasive species. We address these issues here by first exploring the difference between explanation, prediction, and risk assessment. In large measure, the studies we have highlighted are efforts at explanation. Using a variety of investigative approaches, they seek to explain the variance we see within a particular aspect of invasion ecology (e.g., the role of competitors in limiting species’ establishment). In the past, results from these explorations were used with mixed success to predict which species were likely to become invasive. Risk assessment is distinct from prediction in that it seeks to assess the probability of “undesirable outcomes,” which in invasion ecology is defined as non-native species inflicting ecological or economic harm. There is much debate surrounding the ability of ecologists to provide robust predictions or risk assessments. Ecological science informs policy and management in distinct ways depending on the stage of invasion considered. In an effort to prevent harmful species from being transported and released, ecologists have tried to identify risky transportation vectors and risky species. Once non-native individuals have been released and have established incipient populations, ecologists have focused on building early detection systems and prioritizing on-the-ground management. Once an invasive species is well-established and causing ecological or economic harm, ecologists have developed models that expedite control and potential eradication of these species.
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Explanation vs. prediction vs. risk assessment Many questions posed in invasion ecology are problems of explanation (Williamson 1999). For example, from retrospective analyses, we can identify a suite of explanations for why the brown treesnake was so wildly successful on Pacific islands, why the European starling spread so easily across the North American continent, and why the zebra mussel has had such a large impact on the North American Great Lakes. Typically this type of work is accomplished through experimentation, population studies, and statistical analysis or modeling, and requires a detailed understanding of the natural and social history of a taxon and environment. It is important to remember, however, that explanation is not prediction (Williamson 1999). Prediction is “the generation of statements about the likelihood that events will occur” (NRC 2002). In contrast, explanation generates statements about what has already occurred. In a comprehensive review of the state of invasion prediction, the US National Research Council (NRC) listed three criteria that all predictive systems must have. These are (i) systems must be transparent and open to peer review and evaluation; (ii) systems must have a logical framework that includes independent factors important to the invasion process; and (iii) the use of this system must be repeatable and lead to the same outcome regardless of who uses it (NRC 2002). The NRC came to the conclusion that there are very few predictive systems currently in place that fulfill all of these criteria. Even more basic to the goal of prediction, however, is whether or not ecologists can produce sufficiently robust predictive systems to satisfy any criteria, much less satisfy the NRC criteria. Before we move on to describing the nature of risk assessment, it is worth delving into this controversy in a bit more depth.
The quest to predict invaders The desire to predict the fate of non-native populations goes back to the earliest movements of species to satisfy human needs and desires (NRC 2002). Prediction in its modern form began in earnest with the publication of the SCOPE volumes (Chapter 1). In the preface to Drake et al. (1989), the editors laid out a series of three questions for invasion ecologists with the first one being “What are the factors that will determine whether a species will be an invader or not?” The first attempts at prediction were essentially lists of traits reasoned or hypothesized to be associated with non-native species. For example, Paul Ehrlich (1986) tackled this question for animals. He listed possible traits associated with successful invaders and those associated with unsuccessful invaders (Ehrlich 1986). Examples of his traits favoring invaders included high abundance in the native range, short generation times, high genetic variability, association with Homo sapiens, and wide physiological tolerance. This sort of general list has been compiled for other taxa (Baker & Stebbins 1965; Lodge 1993; Ricciardi & Rasmussen 1998), and, although useful as a starting point, the traits listed are often too broad to be of much predictive use and have been criticized as such (Rejmanek & Richardson 1996; Kolar & Lodge 2001; Marchetti et al. 2004b).
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Much of the early work on prediction suffered from a suite of general problems (Lodge 1993; Williamson 1999; D’Antonio et al. 2001b; Marchetti et al. 2004b). The search for predictive characteristics needed more quantitative analysis and statistical rigor, as vote counting and professional opinion were not sufficient. In addition, few studies compared the traits of successful with unsuccessful invaders because the data on failed invasion attempts were often lacking or difficult to acquire. Studies rarely addressed taxonomic association (and therefore inherent trait similarity) among nonnative species (Harvey & Pagel 1991), or the human component by which species are introduced (e.g., shipping patterns or human interest in a species). Purposeful introductions of native species outside of their range but still within a geopolitical unit were often ignored. Finally invasion is a process with discrete stages, and these early studies often failed to recognize that different biological characteristics (or suites of characteristics) may be important at each stage. As we have seen throughout this book, each stage of the invasion process is characterized by unique ecological and social factors. It is highly unlikely that one set of species’ traits would predict transition through all stages (Kolar & Lodge 2001; Lockwood 2005). In the following section we will review efforts at prediction and risk assessment that endeavor to overcome these limitations.
Inherent limitations to prediction There are other more insidious problems with prediction that are difficult to overcome. Williamson (1999) describes five non-mutually exclusive statistical reasons why prediction may fail. First, it is relatively common to get a statistical significance (i.e., P < 0.05) that explains very little in the way of the observed variance (i.e., R2 is very small). In other words, one can get a solid statistical explanation that is a very poor predictor. Second, extrapolating beyond the conditions that produce a prediction is often dangerous. Williamson (1999) suggests this sort of prediction is particularly worrisome when extrapolating from multivariate, transformed, or non-linear functions. Third, statistical shrinkage means the predictor variable (e.g., seed size predicting establishment success) nearly always fits independent empirical data worse than the original information used to produce the predictor (Copas 1983). Fourth, it is possible to use irrelevant information in the production of a predictive system, or what Williamson (1999) refers to as base-rate effects. Irrelevant information has no power to directly predict the outcome of an event (e.g., successful or failed establishment), but it is kept in a predictive model because it is unknowingly correlated to information that does have direct predictive value. This error commonly manifests as the over-production of false positives, which in this case is identifying a species as a potential invader when it is not (Cassey et al. 2004a). Finally, even though predictive systems strive to include all relevant variables, sometimes species previously unidentified as likely to be introduced are in fact released and established in regions of concern. These species may be well outside “the norm” in terms of their traits thus, by default, the predictive system will fail to predict the fate of this “novel” invader very well. There are a host of biological reasons why predictive systems can fail, as well. One need only read the previous chapters to understand that ecological outcomes are difficult
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Knowledge and understanding of invasion biology
Information needs
Likelihood of introduction, establishment or invasion
Characteristics of species and environments Algorithm
Parameter assignments
Species and environment parameters
Figure 12.1 General structure of a predictive system for biological invaders from the NRC. Ovals represent processes in system, and boxes represent state variables. The width of arrows represent levels of certainty, with thicker arrows meaning higher levels of confidence. Reprinted from NRC, 2002. Predicting Invasions of Nonindigenous Plants and Plant Pests, with permission from The National Academy Press, Washington DC.
to summarize in a few variables that may have some predictive value. Mack (1996) suggests that in the face of this uncertainty, small-scale experiments can test the ability of a potential non-native to do well in a given region. However, Kareiva et al. (1996) suggest that, when confronted with environmental heterogeneity, even the powerful combination of controlled experiments and modeling may be insufficient to provide reasonably accurate predictions on how non-native organisms will perform. To borrow a metaphor from Williamson (1999), perhaps invasions are likely to be predictable only in the way earthquakes and tsunamis are predictable; individual successful invasion events are inherently random but general patterns of risk both taxonomically and geographically may be identified. As a summary of this problem, we reproduce the NRC (2002) synopsis of the types of information needed to produce a predictive system within Figure 12.1. The NRC suggests that ecologists know enough about invasions to broadly outline the types of information they need to know; however, obtaining this information is much more difficult, as is understanding how to interrelate this information within a predictive algorithm.
Risk assessment Given this level of uncertainty in developing predictions, perhaps the task is better approached, at least initially, as a process of risk minimization. Risk is the product of the likelihood of an event and its consequences (NRC 2002). In this case we are referring to the “event” as the likelihood of a species transiting a particular invasion
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stage. We may be concerned about any one of the stages, or the “event” may be the joint likelihood of all stages being successfully transited; that is, that the species becomes an invader. Clearly the highest risk scenarios unfold when the event is likely to occur, and the consequences are significant. Conversely, low risk is judged when the event is unlikely, and the consequences are minor. To produce more informed regulatory policies, decision-makers need a range of consequences of a potential invasion. For example, if the likelihood of severe consequences is relatively high, even though we are uncertain about the event actually occurring, decision-makers may opt for a more cautious approach to regulation. Lest we leave the reader with the impression all is lost on this front, we first turn our attention to reviewing fruitful recent efforts at tackling the problem of prediction and risk assessment. In our opinion one of the key reasons for any improvement in prediction and risk assessment is the recognition that each stage of the invasion process presents a unique set of controlling mechanisms (Kolar & Lodge 2001). Breaking down the problem into invasion stages also explicitly recognizes that our policy and management options change drastically as species move along the invasion pathway (Fig. 12.2;
Prevention
Native elsewhere
Survival in transport
Eradication
Establish in new areas
Lag period
Control or restoration
Spread
Ecological impact
Human impact
Figure 12.2 Generalized steps in the invasion process showing the dynamic relationship with management as the invasion progresses. Note that some stages are more relevant for prevention; others are more relevant for issues of control and restoration. Feedback may occur among many of these steps. Reprinted with permission from Sakai, A. K., F. W. Allendorf, J. S. Holt, D. M. Lodge, J. Molofsky, K. A. With, S. Baughman, R. J. Cabin, J. E. Cohen, N. C. Ellstrand, D. E. McCauley, P. O’Neil, I.M. Parker, J. N. Thompson, and S.G. Weller. 2001. The population biology of invasive species. Annual Review of Ecology and Systematics 32: 305–332.
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Richardson & van Wilgen 2004; Simberloff et al. 2005). Thus we address the ecology of incipient invasions and how this knowledge interfaces with efforts to eradicate these populations. Finally, we tackle the question of how fundamental ecological knowledge informs on-the-ground efforts to control harmful invaders.
Screening risky species Risk assessment tools One area of agreement among many researchers is the need to tackle the problem of invasions up front by prevention rather than by management, control, or eradication (Hobbs & Humphries 1995; Reichard & White 2001). This line of thinking makes great sense, as prevention is often much easier and significantly less costly than finding a cure (Bax et al. 2001). Preventing the transport and release of species thought to have a high probability of causing damage is an approach that builds directly from early attempts to identify characteristics of invaders. Once identified these signature or indicator characteristics could be used to develop filters or screening profiles that would allow us to prevent invasions before they begin. There are currently a number of government-funded risk assessment protocols for invasive species (e.g., Weed Risk Assessment of Australia, the Ecological Risk Assessment Framework, and the Generic Non-indigenous Aquatic Organisms Risk Analysis Review Process of the United States). Most risk assessment protocols (sometimes called screening protocols) that are in active use by governments today share a basic structure (NRC 2002). Using a variety of species traits and environmental characteristics, these protocols evaluate the likelihood that a species will transit all invasion stages. Usually the protocols call for “yes” or “no” answers to a series of questions related to these traits and characteristics. The consequences of the invasion are generated by qualitative judgments made by experts in the field. Consequences are considered high if there are several possible costs, and low otherwise. The resulting assessment of the likelihood of the event and the level of negative consequences are combined into a categorical pronouncement of risk (e.g., high, moderate, low). For example, a widely used approach to screening potential invaders is the weed risk assessment (WRA) system developed for the Australian quarantine authority to help prevent importation of unwanted plants (Pheloung et al. 1999). The WRA is designed to evaluate the risk of importing plant species by answering a series of 49 questions based on known attributes and impacts of plant species. The questions are scored and result in one of three recommendations: accept the plant, evaluate the plant further, or reject the plant from entering the country. The questions cover a wide range of “weedy” attributes in three areas including (i) biogeography – distribution, climate, domestication, and previous success at establishing a non-native population; (ii) undesirable attributes – toxicity, unpalatability and ability to establish within native plant communities; and (iii) biology/ecology – attributes that would allow a plant to reproduce, spread, and persist within Australia. Not all questions have to be answered for the WRA to provide a recommendation (there is a minimum number in each category), as information
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about new species is often limited. The WRA for Australia was calibrated using 370 species of successfully introduced species placed into one of three categories: serious weeds, minor weeds, and non-weeds. The acceptance and rejection criteria were then “set” at appropriate places so that none of the plants in the “serious weed” category were given an accept score. This protocol has been widely utilized in Australia and New Zealand and has recently been modified to screen plants for tropical islands such as Hawaii (Daehler et al. 2004). This modified Pacific Island WRA correctly identified 95% of major pests and 85% of non-pests thus showing considerable promise for use in Hawaii and elsewhere (Daehler et al. 2004). The WRA is not flawless, however (McDowall 2004). The protocol involves investing considerable time to amass data to answer the questions (5–8 hours per species). The inability to find data for some species and the occasional difficulty of identifying plants to the species level both severely restrict the number of species that can reasonably be assessed (Daehler et al. 2004). In addition, risk assessment systems such as the WRA and its offshoots sometimes require that users make subjective decisions about how to answer each of the questions posed. Recognizing the limitations of current systems, the NRC (2002) considered risk assessment procedures as an interim step toward a fully predictive system. Importantly, these risk assessment tools allow society to minimize the risks associated with various commercial and private activities while accepting some level of predictive uncertainty. In this sense, they should be held to the same list of criteria by which predictive systems are judged (see above). The NRC (2002) review concludes that risk assessments currently employed by regulatory entities do not fit these criteria. In particular, they call for more quantitative tools for assessing the probability of invasion.
Quantitative tools One of the first manuscripts to apply quantitative statistical analysis to the question of prediction was published by Rejmanek and Richardson (1996). In this study the authors looked at a group of invaders, pines in the genus Pinus, across the Northern Hemisphere. They examined two groups of pines: 12 species that a priori could be classified as invasive due to known instances of successful introduction and spread on two or more continents, and 12 species that could be classified as non-invasive based on plantings in three continents with no reported spread (Rejmanek & Richardson 1996). For each of these species, they gathered information from the literature on 10 life history characteristics including the following: average height, maximum height, minimum juvenile period, mean longevity, average seed mass, and mean interval between large seed crops. They then used a simple discriminant function analysis (DFA) with the characteristics to assign each species to membership in either of the two groups (invaders and non-invaders). The DFA correctly classified 100% (24 out of 24) of the species, and results suggested three characteristics of importance: mean seed mass (mg), minimum juvenile period (years), and interval between large seed crops (years). This simple tool based on easily obtainable demographic parameters represents one of the better ecological schemes for predicting invasive woody seed plants.
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Alternatively, Ricciardi and Rasmussen (1998) took a more heuristic approach and developed a three-step “low-cost screening profile” for aquatic invaders. Their first step was to identify potential geographic donor regions. A potential donor region could be identified in three ways: (i) it could have a history of disproportionate taxonomic contribution to the recipient biota; (ii) it could have climatic patterns similar to the recipient region; or (iii) it could be a region linked to the recipient region by patterns of transport (i.e., shipping patterns). Their second step was to select potential invaders using biological criteria. They used general characteristics and attributes of invaders (i.e., lists of traits) that had been previously put together (see above), suggesting that these traits apply to a wide variety of organisms and are thus useful for prediction. Finally, the last step in the process was to use prior invasion success to inform and narrow the list of potential invaders. The authors demonstrated the efficacy of their screening program by applying it to the North American Great Lakes as an example. They identified the Ponto-Caspian basin (which includes the Black, Caspian, and Azov seas) as the most likely donor region given the extensive shipping links between the basin and the Great Lakes and because the basin had already contributed eight non-native organisms to this region. Using species lists for the Ponto-Caspian basin, they evaluated each species’ probability of invasion based on whether the species were euryhaline (i.e., tolerate brackish water), had life history stages that would survive oceanic transport in ballast tanks, and had a documented history of prior success outside their native range. Their results suggested 17 taxa (14 invertebrates and three fish) with a high potential for success in the Great Lakes and implied that simple heuristic models such as theirs may have wide applicability. A few studies in the late 1990s directly addressed whether taxonomic affiliation was a good predictor of invasive species. Taxonomic patterning may be important when looking for biological characteristics because species-specific traits, including those that enhance invasiveness, are often distributed according to phylogenetic association (Lockwood 1999). In one of the first studies to address taxonomic issues, Daehler (1998) examined phylogenetic patterns among agricultural weeds and non-native plants that invaded natural areas (e.g., nature parks and preserves) using a global angiosperm data set. He used re-sampling statistics to determine if any taxonomic groups were overrepresented among successful invaders and then correlated ecological (species-specific) characteristics with the global richness of each taxon. His results suggested different suites of characters for agricultural and natural area invaders. The most successful groups of angiosperms at invading natural areas included aquatic or semi-aquatic plants (four families), grasses (Poaceae), nitrogen fixers (Fabaceae), climbing vine-like plants and clonal trees (e.g., some Salicaceae and Myrtaceae) (Daehler 1998). Lockwood (1999) conducted a similar analysis for non-native avian species around the world. She examined 396 successfully introduced species of birds and looked for non-random patterns of success among families at the transport and establishment stages using binomial probabilities. Seven families were more successfully established than expected by chance: ducks/geese (Anatidae), pheasants (Phasianidae), sparrows/finches/starlings (Passeridae), parrots (Psittacidae), pigeons/doves (Columbidae), rheas (Rheidae), and New World quails (Odontophoridae). Using contingency tests,
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Table 12.1 Relationships between characteristics of release events and completion of an invasion transition for species characteristics tested in at least two studies. Characteristics
Number of individuals released Number of introduction attempts Year of introduction
Transitions examined Plants
Animals (birds)
Invasive/not
Establish/fail
Invasive/not
+
+, +, +, +, +, +, +, + +, +, +, +, +, NS, NS +, −, NS
+ + NS
Notes: Statistically significant relationships are indicated by + (positively related to making transition) or – (minus) (negatively related to making transition). NS = not significant. Blank boxes indicate that no suitable data were available. Establish/fail = establishing birds have been compared with birds that fail to establish; Invasive/not = invasive plants have been compared with noninvasive plants. Source: Reprinted from Kolar, C. S., and D. M. Lodge. 2001. Progress in invasion biology: predicting invaders. Trends in Ecology & Evolution 16:199–204 with permission from Elsevier.
Lockwood (1999) also examined the influence of human interest on the probability of transport by comparing two modes of introduction for avian species, pet species, and exploited wildlife species (hunting or game). The results suggested that some families (ducks/geese, New World quails, and pheasants) had larger than expected numbers of species transported for wildlife exploitation, while other families (starlings and mynahs – the Sturnidae family) had more pet species transported than expected. Taken together these two studies (Daehler 1998; Lockwood 1999) support the assertion that taxonomic patterns may be important for predicting species that successfully transit a particular invasion stage, especially when combined with other attributes that have some predictive power (Reichard & Hamilton 1997). Kolar and Lodge (2001) synthesized a number of quantitative investigations into the usefulness of species’ traits for prediction and found that different traits are important at different stages of the invasion process. They examined 16 studies published between 1986 and 1999 that were quantitative, empirical, and that examined at least 20 species (Tables 12.1 & 12.2). Their meta-analysis is interesting for three reasons. First, they were only able to find data for avian and botanical studies, which points strongly to a need for further research across a wider variety of taxonomic groups. Second, their results clearly identified large gaps in our knowledge across invasion stages, particularly for establishment in plants and spread in birds, again indicating directions for future inquiry. Third, and perhaps most important, there are very few consistent trends in the published data either across taxonomic groups (between birds and plants) or across invasion stages within taxonomic groups. Some traits are useful predictors for one stage of the invasion process, but are useless in another. Even more frustrating for management and prediction, their results suggest that a single trait can work for a species at one stage and against it in another.
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Table 12.2 Relationships between species characteristics and transition between invasion stages for botanical and avian taxa. Characteristics
Transitions examined Plants Establish/Fail
Animals (birds) Invasive/Not Establish/Fail Invasive/Not
a. Number of seeds or eggs
+
b. Fire tolerant c. Migrating
+, −, NS
d. Body mass e. Origin f. Taxon g. History of invasion h. Family or genus invasive i. Vegetative reproduction j. Variability in seed crop k. Dispersal mechanism
∗ ∗
l. Seed or egg mass
m. Length of juvenile period n. Temp./ habitat match o. Broods per season p. Flowering period length q. Annual (versus perennial) r. Range area s. Diet breadth or type t. Inhabits diverse climates u. Height or body length v. Longevity w. Reproductive system
+ + + NS
∗, ∗, NS ∗, ∗ +, +, +, + +, +, +, + +, +, + −,− +, +, NS, NS, NS − , −, −, NS, NS, NS, NS −, −, −, NS, NS
+, −, NS, NS
(+/−)
− , −, NS, NS, NS +, +, NS, NS, NS ∗, ∗ ∗, NS
+
− − + +, NS
NS, NS NS, NS +, NS NS +, NS, NS NS ∗, NS
−
+ +
NS, NS NS, NS NS NS +
Notes: Statistically significant correlations are indicated by + (positively related) or − (negatively related). NS indicates no significant relationship. For categorical variables, significant relationships are indicated by ∗. For additional information regarding the specifics of each characteristic and the list of references used in this analysis see Kolar and Lodge (2001). Source: Reprinted from Kolar, C. S., and D. M. Lodge. 2001. Progress in invasion biology: predicting invaders. Trends in Ecology & Evolution 16: 199–204.
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Stage-specific approaches There are a growing number of investigations that have taken some of the lessons of Kolar and Lodge (2001) to heart and have evaluated characteristics across more than one invasion stage (e.g., Cassey et al. 2004b). For example, Marchetti et al. (2004b) examined how biological traits of non-native fishes contributed to success at transiting three invasion stages (i.e., establishment, spread, and integration/impact) in California. Marchetti and colleagues (2004b) documented the presence and abundance of every non-native fish species inhabiting each watershed across California, including native species established in watersheds outside their native range. They also collected information on attempted but failed introductions. Based on information in the literature, eight species-level characteristics were recorded for all fishes. These characteristics included trophic status, size of native range, type of parental care, maximum adult size, physiological tolerance, distance from native source, prior invasion success, and propagule pressure. The authors then employed an information theoretic model selection approach (Burnham & Anderson 2002) to directly evaluate a set of a priori invasion hypotheses (regression models) and measure the relative importance of each of the eight variables on invasion success. The results suggested different suites of characteristics were important at each stage of the invasion process. For species establishment, all eight traits were equally important, suggesting a largely stochastic process at this stage. For spread (measured as the number of watersheds a species invaded) physiological tolerance and propagule pressure were important, while for integration/impact (measured as the average abundance of each non-native species) prior invasion success and propagule pressure were important. In a different vein, Kolar and Lodge (2002) developed a fairly sophisticated risk assessment protocol for non-native fishes in the North American Great Lakes over three invasion stages (establishment, spread, and impact) by using the Rejmanek and Richardson (1996) approach as a starting point. First the authors identified 24 established and 21 introduced but not established species of fish in the Great Lakes. They then identified those established fishes that spread quickly through the lakes and those that had negative impacts (i.e., nuisance species). For each of these species they collected data on 24 species-specific traits (life-history, habitat needs, and invasion history). They then conducted both DFA and categorical and regression tree analysis (CART) on the data for establishment but only DFA on the data for spread and impact. Kolar and Lodge (2002) had an impressive success rate as they were able to discriminate between successful and failed species with 87% accuracy using DFA and 94% with CART (Fig. 12.3). When evaluating the other invasion stages (spread and impact), the authors could distinguish fast spreading fishes with 94% accuracy and nuisance species (high impact) with 89% accuracy. They next used their models to predict the risk of future introductions from the Ponto-Caspian basin through transport in ballast water. They found sufficient data on 66 species from the Ponto-Caspian and ran them all through both their DFA and CART models for the establishment phase. The DFA predicted 24 species could establish while the CART predicted 36 species, with a Jaccard similarity (similar to overlap) of 57% between the two lists. Using a conservative approach, they suggest that the 22 fishes predicted by both procedures may pose a serious risk to establishment
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Relative growth @ 2 yrs
< 68.5
> 68.5 Minimum temperature °C
Number taxa in diet < 4.5? < 4.5
> 4.5
< 5.5
> 5.5 Fail
Success Number taxa in diet < 1.5? < 1.5
Relative growth @ 1 yr(%)
3 Success/1 Fail > 1.5
0 Success/5 Fail
< 26.5
> 26.5
Success
Fail
Fail
Success
1 Success/0 Fail
0 Success/13 Fail
0 Success/1 Fail
20 Success/1 Fail
Figure 12.3 Predictive decision tree (CART) of successful and failed non-native fishes in the Great Lakes, USA. Ovals represent decision points and rectangles are terminal points in the tree resulting in classification of invasion risk. The numbers of known successful and failed non-native species categorized into each terminus are given, illustrating that 2 of 45 were misclassified. Reprinted with permission from Kolar, C. S., and D. M. Lodge. 2002. Ecological predictions and risk assessment for alien fishes in North America. Science 298: 1233–1236. Copyright 2002 AAAS.
in the Great Lakes and urge that caution should be used so as not to transport them to the region. Several authors have also attempted to use information about a species’ ecological niche or its eco-physiological constraints to predict which species will pose a significant establishment risk within a given recipient region. This approach to prediction looks for matches between the physical environment of the donor and recipient regions and examines how the species in question responds to these conditions (e.g., Duncan et al. 2001). The greater the match in conditions is, the more likely a propagule is to establish. For example, Peterson and Vieglais (2001) utilize a complex informatics algorithm (genetic algorithm for rule-set prediction, or GARP) to quantify a species’ ecological niche based on information gathered in its native range. This type of ecological niche modeling involves documenting the occurrence patterns in the species’ native range, developing a model that accurately describes this distribution using biophysical properties (e.g., elevation, rainfall, temperatures), and then projecting that model onto other landscapes to identify areas of potentially suitable non-native habitat (Peterson 2003; see also Fig. 12.4). The usefulness of this approach is illustrated with the North American invasive plant Himalayan bushclover, Lespedeza cuneata (formerly Sericea lespedeza; Peterson et al. 2003), for which models were derived from the geographic distribution in its native range (Asia) and then projected on to North America to look for areas coincident with known invasions (Fig. 12.5 (a) & (b)). The models predict the North American distribution invasion very well, suggesting that ecological niche modeling has a high potential to predict geographic patterns of invasion. Unfortunately,
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Ecological space
Geographic space
Ecological niche modeling
Model of niche in ecological dimensions Precipitation
Occurrence points on native distribution
Temperature
Projection back onto geography Native range prediction
Invaded range prediction
Figure 12.4 Conceptualization of the ecological niche models of Peterson that are used to predict the geographic range of species invasion. The figure shows the process of characterizing the distribution of a species in its native geographic range, modeling the ecological niche of this species, and projecting the geographic expected range of the same species in its non-native range. Reprinted from Peterson, A. T. 2003. Predicting the geography of species’ invasions via ecological niche modeling. The Quarterly Review of Biology 78: 419–433, with permission from the University of Chicago Press.
this method also has some drawbacks, namely the extensive time necessary to gather data and make the models (approximately 2+ months), as well as a general lack of accessibility of species occurrence data within their native ranges (Peterson et al. 2003).
Synthesis A quick synopsis of the state of predicting invaders strongly suggests that there is no “magic bullet” character (or set of characters) that predicts the successful transit of a species through all invasion stages or across all groups of species (Cassey et al. 2004a).
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(a)
Figure 12.5 Predictions of native (a) and potential non-native distribution (b) of Lespedeza cuneata using GARP. White symbols on the native distribution maps indicate occurrence records used to build ecological niche models. Black polygons on the non-native distribution maps indicate known occurrences at the level of USA counties or hydrologic units. Increasingly dark shading indicates greater confidence in the prediction of the species’ presence in the non-native range. Reprinted from Peterson, A. T., M. Papes, and D. A. Kluza. 2003. Predicting the potential invasive distributions of four alien plant species in North America. Weed Science 51: 863–868, with permission from WSSA.
However, ecologists have made good headway by breaking down the process into its constituent stages and utilizing quantitative approaches. This realization does leave us with a difficult problem when considering the rate at which species are finding their way into non-native territory. How can we ever hope to evaluate all of these species in the quantitative manner of, for example, Kolar and Lodge (2002)? The burden of quantitatively evaluating the invasive potential of all species transported as non-natives is so large that alternative, and hopefully complementary, approaches are in order. One option is to shut down transport vectors that seem to add an inordinate number of harmful non-natives to a region. We will discuss this approach next. Before that, however, we want to explore a wrinkle to the standard screening procedure that may help in this regard: the regulatory approach of “guilty until proven innocent” (Mack et al. 2000; Holway et al. 2002a). The regulatory structure within which most species risk assessment protocols operate today lays the burden of proof on the agency deciding whether to prohibit introduction and not on the party interested in importing the non-native or importing the
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(b)
Figure 12.5
Continued
commodity that often carries non-native species. If the agency fails to prove that the species is risky, the species is imported. Thus, this approach is often called “innocent until proven guilty.” In contrast, “guilty until proven innocent” approaches assume that all non-native species pose a high enough risk to ecological or economic systems to be prohibited until proven otherwise. The risk assessment system is then charged with dividing lists of species into those that are considered high risk and those considered low risk. Low risk species can be introduced at will and the high risk species remain prohibited. The allure of the “guilty until proven innocent” approach is the shifting of the burden of proof away from the regulatory agency and toward the advocating party. Under this regulatory approach, the inherent uncertainty in classifying risk is sufficient cause for withholding permission to import and release the non-native. If the advocating party would like to continue pursuing importation they would have to work toward reducing that uncertainty. “Innocent until proven guilty” regulations function the other way around. There are two overriding problems with the “guilty until proven innocent” approach to regulation. First, private individuals often demand access to non-native species such as garden and landscaping plants, pets, and aquarium species. This demand is so broad that it will always be difficult to sharply curtail these types of non-native species imports without a widespread change in how society views the risks associated with these species (Simberloff 2003a). Second, international treaties largely govern the restriction of movement of biological invaders across international borders, and the enforcement
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of these treaties often gives environmental concerns short shrift. One of the most important treaties related to screening invasive species is among members of the World Trade Organization (WTO), called the Agreement on the Application of Sanitary and Phytosanitary Measures (SPS), which allows members to restrict the transport of species that threaten human, animal, or plant life (Mack et al. 2000). If a nation in the WTO attempts to ban the importation of a particular species, its effort can fail if the regulatory portion of the WTO decides the ban is illegal or is an attempt at a protectionist trade barrier rather than a legitimate attempt to prevent a species invasion (Mack et al. 2000). Despite this impediment, New Zealand has successfully implemented a “guilty until proven innocent” policy for intentional importations of non-native species under their Hazardous Substances and New Organisms Act of 1996 (Williams & West 2000). This illustrates that there is plenty of room left to explore in the relationship between WTO free-trade regulations and the need to secure each nation’s environment from invasive species.
Screening risky transportation vectors Transportation is not a biological or ecological process; it is a social, economic, and political entity, something wholly human in construct. We have a better understanding of methods of transport than we do of ecological systems, and we can perhaps more effectively limit or screen the movement of goods, vehicles, animals, and people into novel areas because of this.
Ballast water exchange We have seen that ballast water is widely regarded as the leading vector of marine invasions. Yet ballast water regulation is the one case for which there is measurable success in the adoption of regulatory policies. In certain ways this success is due to the nature of the transport mechanism itself. Ballast water is not cryptic, like seeds mixed with soil or found hitching on ornamental plants. It is clear how and where ballast is being off-loaded. Ballast water is contained within a ship’s hull, and there are detailed records of water transfers due to shipping records and logbooks that allow us to monitor and track past and future movements. Since the early 1990s a complex set of regulations and guidelines, both international and national, has governed the management of ballast water. In 1991 the United Nations International Maritime Organization instituted voluntary international guidelines calling for mid-ocean ballast water exchange (BWE) (Weigle et al. 2005). The idea behind BWE is that coastal and estuarine species would not survive a mid-ocean transfer of water and that mid-ocean organisms would not survive in coastal waters. In principle BWE should be successful, but a number of factors limit its effectiveness. Exchange mechanisms often do not remove 100% of the ballast water as some ships are not designed with this type of ballast release. In addition, BWE is time consuming, and bad weather sometimes prevents exchange from occurring. Even if BWE does exchange close to 100% of the ballast water, recent studies indicate that the diversity of organisms
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in ballast tanks actually increases following BWE (see Box 4.1) and that mid-ocean water is not lethal to some coastal organisms (Wonham et al. 2001). As a result of these and other problems associated with BWE, a large number of alternative treatments to ballast water have been proposed, including intensive filtration, thermal treatment, biocide treatment, or ultraviolet light control (Tamburri et al. 2002). Many of these suggestions are problematic due either to high economic costs or an indirect negative influence of released toxins on ship-workers’ health or the receiving environment. One promising suggestion, however, that is both relatively inexpensive and completely inert is the use of nitrogen gas to deoxygenate ballast water (Tamburri et al. 2002). On a national level, the United States passed the Nonindigenous Aquatic Nuisance Prevention and Control Act in 1990 in response to the cost and difficulties of the zebra mussel (Dreissena polymorpha) invasion of the North American Great Lakes. This act was a significant piece of legislation for the control of non-native species but was limited in its scope. The act provided that ships entering the Great Lakes must exchange or treat their ballast water. Unfortunately the language in this legislation was limited to aquatic organisms, ballast water transfers, and the Great Lakes. In 1996 the US Congress reauthorized the 1990 act and changed its name to the National Invasive Species Act (NISA). Although the focus of the new act remained BWE, its scope expanded slightly. The NISA authorized research on water treatment methods, physical barriers for ballasting water, and provided funds for a database on ballast water. There are similar governmental policies for various other countries (e.g., Williams & West 2000). However, as yet, no one has done a comparative policy analysis of these regulatory approaches to managing risky invasion vectors (Simberloff et al. 2005). In addition, in most countries the authority over regulating the movement of non-native species is fragmented across several governmental levels (i.e., national, provincial, state, municipality). It seems critical to efforts to stem the movement of non-native species to evaluate which of these policies have proven effective, which ineffective, and what determined the difference.
Regulating purposeful introductions A substantial proportion of most nation’s invasive species found their way into the country through purposeful transportation vectors (Mack & Lonsdale 2001). These include vectors such as the horticultural plant trade, aquaculture facilities, and the pet trade (Holcik 1991; Naylor et al. 2001; Reichard & White 2001). Reducing the risk from these transportation vectors requires political, economic, and social research. Aquaculture facilities, for example, may adopt more secure housing for the individuals they grow (e.g., salmon, trout, shrimp) if it is clearly necessary and proves economically viable. Horticultural enterprises may cease selling known (much less suspected) harmful non-native plants if the public responds positively to their decision to take some plants off the market. And the pet industry may take some responsibility for the release of unwanted pets into natural settings by their owners if required by the appropriate regulating agency. The most likely form for these controls on risky transportation vectors is through voluntary best-management practices adopted by professional societies.
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Education of the pet-buying and gardening public may also be effective (Simberloff et al. 2005). Top-down regulation probably will not work for such diverse enterprises that have large markets, substantial financial profits, and loyal customers. An alluring, although circumscribed, quantitative approach to assessing the risk of commonly transported species comes from Bartell and Nair (2003). These researchers evaluate the risk of establishment of known invasive species using demographic models linked to continuing transportation and release vectors. These models differ from those of Kolar and Lodge (2002), for example, in that they assess the value of regulatory or control policies on the probability of establishment of a single species. Bartell and Nair (2003) illustrate the mechanics of their model using the probability that the Asian longhorn beetle (Anoplophora glabripennis) will establish in the eastern United States after arriving at a port of entry. The key innovation of their approach is linking a standard population projection matrix to the probability that individuals of each beetle stage class will arrive via regular commercial shipments. They also include the prevalence of suitable hosts near the port of entry. Thus, Bartell and Nair (2003) are linking propagule pressure (Chapter 4) with the population dynamics (Chapter 7) of the individuals already released. The result is the production of “exposure–response” functions. Their methodology includes an explicit recognition of various control methodologies that can be employed to reduce the number of beetles escaping from shipments containing wood packing material at the port of entry. These include microwave treatment of the packing material, fumigation of the material using insecticides, or heating the material to temperatures beyond the lethal level for the beetles. The result is a modified exposure function that reflects the propagule reduction effects of instituting control mechanisms at time of arrival (Fig. 12.6). Such approaches allow quantitative analysis of the reduction in risk associated with various management options. It could easily be extended to reflect different population dynamics of the already-released individuals around the port of entry and multiple spatially structured ports of entry. No matter what approach we use to reduce the risks associated with particular transportation vectors, a detailed understanding of the rate at which species are introduced and the spatial patterns of source regions can hone the effectiveness of policy and management tools. For example, the few studies that have been able to document temporal invasion trends suggest that as vector traffic increases, or if the geographic origin of the vector changes, the number of recorded non-natives will increase (Chapter 3). Thus, any temporal or spatial change in vector traffic should trigger both a reassessment of threats from non-native species and the forecasted effectiveness of the regulatory mechanisms currently in place. For this reason, the collection of detailed information on global patterns in transportation routes is of particular importance. Such an “Invasion Atlas” (Carlton 1996b) should contain information on the geographic origin of all recorded non-native species, the approximate year in which these species were first recorded, and the vector that they probably utilized for initial entry. From such an atlas, managers could identify significant changes in vector traffic, particularly those that risk exposure to donor regions with a history of exporting large numbers or particularly harmful non-native species. Unfortunately, we are a long way from realizing this goal. In the meantime, it should
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1
p(N(T) > viable size)
p Risk reduction
Exposure–response function
p
RI,i RI,i 0 0
1
10
100
1,000
Rate of entry, RI
Figure 12.6 Exposure–response curve generated for the Asian longhorn beetle. The Y-axis is probability of establishment. Risk reduction comes from the differences between the baseline (pβ ) and treatment (pτ ) after applying treatments at the port to control the size of persistent propagules. The model combines propagule pressure with the population dynamics of established individuals and thus provides estimates of risk reduction that can be derived from management. Reprinted from Bartell, S. M., and S. K. Nair. 2004. Establishment risks for invasive species. Risk Analysis 24: 833–845, with permission from Blackwell Publishing.
prove useful to model how spatial changes in vector traffic will influence temporal patterns in non-native species introductions (e.g., Drake & Lodge 2004). Such efforts will allow a more thorough understanding of when agencies should act to reinforce (or enforce) regulatory policies, create new policies, and what price society will have to pay when they fail to do either.
Early warning systems and eradication of invasion foci Once a non-native population has been detected within a region of concern, the policy and management options shift dramatically (Richardson & van Wilgen 2004; Simberloff et al. 2005). Control options move into traditional methods of invasive species removal such as chemical treatment, manual removal, or biocontrol. Unlike the situation described in the next section, however, total population eradication at this stage is still a reasonable possibility (Myers et al. 2000; Simberloff 2004). Perhaps the most efficient way to prevent ecological or economic harm is to eradicate the non-native population while it is still relatively small. Rejmanek (2000a) illustrates this point nicely using eradication information on two non-native plants in California (Fig. 12.7). The key is realizing that populations build growth momentum through time, sometimes expanding their geographical ranges exponentially under highly favorable conditions. Thus, quick action to eliminate the population while it is still small will be far more efficient economically than waiting until the population expands or waiting to see if the species causes any impact. These small populations are often called nascent invasion
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Eradication effort (work h)
100000
10000
1000
100
10 0.001 0.01 0.1
1
10
100 1000 10000
Initial infestation (ha)
Figure 12.7 The relationship between ease of eradication (measured in US dollars) versus initial population size when control began for Hydrilla verticillata (open squares = eradicated; closed squares = ongoing) and Onopordum acanthium (open circles = eradicated; closed circles = ongoing). Concentrating on invasion foci is always a better option than controlling after much geographical range expansion and population increase. Reprinted from Rejmanek, M. 2000a. Invasive plants: approaches and predictions. Austral Ecology, 25: 497–506, with permission from Blackwell Publishing.
foci, and detecting such foci is the lynchpin to the success of this approach. Even when detection rates are high, however, managers often still face the challenge of deciding which foci to attack first as the funding available for invasive species control is currently very limited and inconsistently delivered (Myers et al. 2000).
Early detection systems Although many early botanists haunted ports and railways because of their exciting suite of novel species, there are very few comprehensive efforts to systematically sample common ports of entry for the presence of non-native species. Examples include efforts to detect non-native insects or pathogens at common ports of entry that are known to cause severe agricultural damage (e.g., Carey 1996) and systematically surveying for non-natives within marinas that regularly have ballast water released into them (Myers et al. 2000). Although to our knowledge there are no systematic lists of instances where incipient non-native populations were found, our impression is that most examples of the detection of nascent invasion foci result from fortuitous finds by knowledgeable biologists and citizens. The void in our ability to detect nascent invasion foci is well recognized amongst invasion ecologists, and it prompted Ricciardi et al. (2000) to
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suggest the implementation of a network of early-warning systems (Wotton & Hewitt 2004). Efforts to list the most troublesome invaders within states, provinces, or regions serve to fill this void in early detection to some extent. One of the best predictors of a species’ likelihood of causing economic or ecological problems is its past performance after having been introduced into another non-native locale (Reichard & Hamilton 1997; Ricciardi 2003). Thus, astute managers look to neighboring states, provinces, or countries to see what invasive species are on their lists and prepare for introductions. Because several organizations (governmental and otherwise) regularly update such lists and the practice is widespread (e.g., Bossard et al. 2000), it may only take cross-agency coordination for this to function as an effective early warning system. Currently, IPANE (Invasive Plant Atlas of New England) serves this function in the northeast US and trains botanists and volunteers to recognize potential invaders, flagged by invasiveness in neighboring states (Mehrhoff et al. 2003). There are obvious limits to relying on lists of existing invasive species as an earlywarning system. The most important is that the approach ignores the arrival of a new non-native species whose impacts cannot be known. Such species have no invasion history and by default will not show up on regional lists. The magnitude of this problem probably varies from region to region; however its extent is of sufficient size to be of concern. In New Zealand for example, Rejmanek (2000a) reported that 20% of nonnative plants collected since 1950 were never previously reported as invasive outside of New Zealand. In addition, it is not clear how well lists of known invaders would work as an early warning system for areas that lie geographically far apart. Nevertheless, the principle has worked well enough to be incorporated into several screening protocols (Reichard & Hamilton 1997).
Eradication of invasion foci Once a non-native species has been detected, and this species is likely to be invasive, the next move may be to eradicate it. Although there is no systematic accounting of how often nascent invasions have been eradicated, there are enough examples of successful eradications to suggest that it is a viable option (Simberloff 2003b). Eradication efforts often require very little ecological knowledge and instead have been successful due to fast action by regulators and managers. In most cases, simple “brute force” approaches to eradication (e.g., use of chemicals, manual removal, or trapping) are all that are needed to stop the invasion before it starts (Simberloff 2003c). When ecological knowledge has come into play, it provides a key insight into the species’ life history or behavior that leads to more efficient removal strategies. The examples of successful eradication provided by Simberloff (2003b,c) point to three insights. First, whether an incipient non-native population is eradicated or allowed to expand is complicated by a series of political and social dynamics. This is yet another place where comparative policy analysis would prove informative. For example, it may be that successful eradication results from clear lines of authority for implementing control or the persistence of local citizens and biologists in attaining
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funding (and permission) for implementing control. Second, our understanding of eradication techniques comes almost exclusively from on-the-ground experience and not necessarily from ecological science. Simberloff (2003b) goes so far as to suggest that the study of eradication is the “step-child” of invasion ecology. There is a clear need to collate and make accessible successful methods of eradication so that managers do not continually “reinvent the wheel” or make easily avoided mistakes. Although detailed ecological science may not be necessary for successful eradication (and indeed may be counterproductive), basic ecological knowledge often plays a key role in the efficiency of eradication. Ecologists need to devise methods for rapid evaluation of newly arriving non-natives so that these insights are filtered to managers quickly and actively contribute to the adaptive management of eradication efforts. Third, and following from the last point, ecologists have much to contribute in efforts to limit non-target (or collateral) damage that accompanies brute force approaches to eradication. One of the principal obstacles to quick eradication is the effect of the treatment on native species and human populations (Simberloff 2003b). There is a glaring need for invasion ecologists to collaborate with ecotoxicologists so that eradication methods can be made less brutish but remain forceful. The above discussion assumes that the incipient invasion could be narrowed to one small non-native population. This is not always the case. Often a known harmful non-native species is arriving via vectors that deposit individuals across a huge area. This often happens when a non-native is a common but difficult to notice hitchhiker in commercial imports. These individuals are thus arriving at several ports across a country, at which point the question becomes one of prioritizing limited eradication dollars across the several invasion foci. For example, niche modeling efforts for the Asian longhorn beetle in North America highlighted the importance of the Great Lakes region as the start of most invasion scenarios that led to widespread occurrence of the beetle (Peterson & Vieglais 2001). Nascent invasion foci that were located near the North American ports along the Pacific coast, where the species is more regularly found within solid wood packing material, are much less likely to result in rapid colonization of the entire continent. Thus, niche modeling appears poised to allow managers to prioritize control efforts across several independent nascent invasion foci. In this case, the models suggest the counterintuitive action of expending greater effort at controlling invasion foci around ports in the Great Lakes region that receive relatively few shipments infested with the invasive beetle, relative to ports along the Pacific coast. Although this is an information intensive endeavor, the harm inflicted to North American forests by the beetle is so high that the cost of devising the model is well worth the effort.
Controlling populations of widespread invasive species The final phase of management is controlling, and occasionally eradicating, wellestablished populations of invasive species. Things become much more complicated when attempting to implement effective management of such populations. These species are geographically widespread, their population numbers are high, and their integration into local biological and cultural systems may be complete. The task is
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so daunting that some have argued that management directed towards such invasive species is unwise. Limited environmental resources may be better applied to problems that have a better prognosis of success. Nevertheless, there are enough examples of successful control and eradication of well-established invasive species to warrant serious consideration (Myers et al. 2000; Simberloff et al. 2005). Indeed, the discipline of weed science has historically been dedicated to developing herbicidal management of plants that cause agricultural and environmental problems (Jordan 2003). Similarly, the science of biocontrol has found a calling in the control of well-established invasive plants and insects (Hoddle 2004). Ecologists have jumped into this arena more recently, and they bring to the table sophisticated approaches to targeting the most vulnerable aspect of an invasive species’ life history. A good example is the use of modeling techniques to prioritize alternative control methods for the invasive plants Heracleum mantegazzianum and Impatiens glandulifera in the British Isles (Wadsworth et al. 2000). These two species were introduced to the United Kingdom from Asia as ornamentals around the turn of the last century (1900). Both species are widespread in the United Kingdom and cause a variety of problems including altering stream function and harming human health (Wadsworth et al. 2000). Each species has been controlled with some success using herbicides, manual removal, or grazing regimes. Each option is effective if applied and timed properly; however, each option has its drawbacks. Wadsworth et al. (2000) approached the problem of prioritizing control options by combining models describing the rate of spread for each species with information on the effectiveness of the various control options. Model simulations were run for 25 “years,” and control was applied annually at varying intensities represented by the spatial extent over which non-native individuals were killed. Control mechanisms were applied randomly across space, or were applied preferentially to areas that had high population sizes of the invasive species, had been occupied the longest, had been occupied the most recently, were upstream from uninvaded areas, or were near human population centers. The results indicated that, in the absence of control, both species would experience a short lag period and then spread across the entire area of concern, although H. mantegazzianum had the longer lag time and the slower rate of geographical spread. As one might expect, the larger the area treated each year, the greater the reduction in each species’ range and the more rapidly total eradication was achieved. No scenario resulted in total eradication of I. glandulifera unless at least 14% of the total area was treated. H. mantegazzianum spread so much more slowly that much less of the total area could be treated annually, and the population could be eradicated. Prioritizing control efforts on sites that have been occupied for long periods or relatively short periods were both about as effective as randomizing control efforts. The more efficient control methods either targeted areas near human population centers or areas upstream of uninfested sites. The strategy of targeting human population centers seemed to work well because urban centers were particularly susceptible to invasion and because control targeted around human centers resulted in a clustered pattern of eradication. Also, the information required to target such areas was easily obtained from census records, whereas the information needed to target areas already supporting each invasive species was more difficult to come by (i.e., added a significant secondary cost). Given the costs
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and controversy over the use of any of the control options available, the results of Wadsworth et al. (2000) provide managers with a clear prescription for the regional eradication of these two harmful invasive species. Another viable option for management of very widespread invasive species is biocontrol (Box 4.2). For example, Shea and Kelly (1998) used demographic projection matrices to analyze the efficacy of controlling nodding thistle (C. nutans), a noxious rangeland weed in New Zealand, using an insect natural enemy. Simulations of the attack of a biocontrol agent, the nodding thistle receptacle weevil (R. conicus), suggested that it must kill nearly 70% of all C. nutans seeds to be effective. Field trials of the biocontrol agent rarely showed more than 40% loss of seeds. Thus, biocontrol using this natural enemy will probably not, by itself, control nodding thistle (Box 12.1).
Box 12.1
Dead flowers
As described in Box 4.2, classical biocontrol is an important tool in our arsenal for control of invasive species. Although there is often a steep learning curve associated with finding the right natural enemy to release as a biocontrol agent, in cases where the same species has become invasive across many regions, what we learn in one place may be of tremendous help in deciding what to do in another. One such invader is Carduus nutans (nodding or musk thistle), a member of the family Asteraceae that is native to Eurasia. It has become a nasty pasture weed in North and South America, southern Africa, Australia, and New Zealand. In response, several insect species have been released as biocontrol agents. The three most prominent are the thistle receptacle weevil (Rhinocyllus conicus), the receptacle gallfly (Urophora solstitialis), and the root-crown weevil (Trichosirocalus mortadelo, formerly T. horridus). The first two weevils lay eggs within flowers, thus preventing seed formation. T. mortadelo reduces survivorship and growth of nodding thistle by burrowing into the meristem (Shea et al. 2005). Like some other biocontrol agents introduced into separate invasive populations, these weevils (in particular R. conicus) have shown different levels of nodding thistle control across continents (Shea & Kelly 1998). For C. nutans the effects differ so much that biocontrol agents that have worked on one population of the invader have shown no effect on other populations. There could be several reasons why this is so, and Katriona Shea and her colleagues endeavored to figure out what was going on. Using detailed information on the demography of nodding thistle populations in New Zealand and Australia, they noticed that, in New Zealand, nodding thistle population increases were driven by early life history transitions that resulted in high fecundity but shorter life spans per individual. However, in Australia, nodding thistle population growth was driven by the opposite pattern, meaning longer life spans but lower fecundity. It is not entirely clear why these differences exist between populations. The Australian nodding thistles came from New Zealand, and thus they share a common evolutionary history. Through the use of matrix models, Shea et al. (2005) showed that these differences in life history
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between New Zealand and Australia could easily translate into differences in the effectiveness of weevils introduced as biocontrol. As expected by Shea and colleagues, of the three biocontrol weevils released in Australia, T. mortadelo appears to control nodding thistle the best. This is the weevil that strongly influences survival, and in Australian populations of nodding thistle this is the life history stage that is most responsible for observed rates of population growth. However, this is exactly the wrong biocontrol weevil to do the job in New Zealand since this population of nodding thistles is increasing in size due to high fecundity. Reducing survivorship is meaningless to a population that already dies young. Instead, Shea et al. (2005) suggest that U. solstitialis is expected to have the largest impact on nodding thistle in New Zealand since it has large negative impacts on thistle fecundity. The release of all three weevils in New Zealand has happened too recently to confirm this prediction. Nevertheless, the work of Shea et al. (2005) suggests that subtle differences in the performance of invasive species across their non-native range can have huge impacts on how managers should implement control measures. What works in one place, may not work in another. Although there is plenty left to explore in this context, Shea et al. (2005) suggest that the ecological context (i.e., physical environment, number, and types of competitors) in New Zealand versus Australia may be a strong factor in determining these across-population differences in nodding thistle demography.
Finally, one of the most comprehensive analyses of control options comes from Odom et al. (2003). These authors address the options for controlling Scotch broom (Cytisus scoparius) in Barrings Top National Park, Australia. The key innovation of this effort is the explicit incorporation of the monetary costs of control options and valuing the losses of recreational opportunities and biodiversity in the absence of control. They utilize a dynamic optimization framework that includes a series of equations depicting economic costs, value losses under no control, and standard population equations (i.e., difference equations). Like Wadsworth et al. (2000) these authors consider a variety of control options, which they depict as a control matrix. They also evaluate real-life policy scenarios whereby monetary income is very limited, and thus control options cannot be utilized to their full potential, or are haphazardly applied across years (Fig. 12.8). The results indicate a $21 (Australian) dollar per hectare marginal benefit to implementing control measures versus not controlling the invader and losing recreational opportunities and the indirect benefits of biodiversity preservation. Also, a combination of control options that target population density and reduction of the seed bank are most critical to effective control. This modeling exercise produces a package of control measures that can be used to attack the problem each year, allowing managers to respond efficiently to erratic funding situations (see Box 12.2 for the results of a similar cost–benefit analysis).
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Box 12.2 Cruel to be kind The Cape flora of South Africa is the hottest of the world’s biodiversity hot spots (Myers 1990) with over 8500 plant species, nearly 70% of which are endemic to the region (Cowling et al. 1992). The region’s Mediterranean climate gives rise to fynbos (distinctive South African Mediterranean scrub vegetation) but also attracts both people and their associated non-native species. The principle threat to the Cape’s biodiversity is the impact of non-native species. Non-native woody plants and trees displace native fynbos vegetation and increase biomass by between 50 and 1000% (van Wilgen & Cowling 1996). The rise in biomass increases the frequency and size of fires much to the detriment of native species (see Box 5.2). The invasion of watersheds by Australian, European, and American species (e.g., Acacia, Eucalyptus, Pinus, and Prosopis species) also reduces water runoff substantially. This has a direct effect on South Africa’s largest cities in the region (Cape Town and Port Elizabeth) as people rely on runoff from the nearby catchments for freshwater (van Wilgen & Cowling 1996). Le Maitre et al. (1996) deduced that the presence of non-native vegetation in Cape watersheds reduces the amount of water available to Cape Town residents by 30%. Using economic valuation techniques described in Box 10.2, van Wilgen and Cowling (1996) weighed the costs of physically removing non-native trees from Cape watersheds to increase streamflow versus developing alternative water supplies (e.g., desalination plants, dam construction, and reuse of sewage water). The results suggest that developing alternative water sources is 1.8 to 6.7 times more expensive than removal of invasive trees in the watershed. When these figures were presented to the newly elected government of South Africa, officials opted to institute a program of watershed management that actively removed invasive woody plants. The program, called Working for Water, employed disadvantaged citizens as work crews that cut down particularly water-thirsty non-native trees within habitats prioritized as primary producers of freshwater (van Wilgen et al. 1998). Working for Water has proven quite successful and has managed to leverage private funds and international support that have allowed it to expand its purview over the years (Richardson & van Wilgen 2004). Working for Water perhaps represents the largest ecosystem management program in Africa. Its benefits include the preservation of a unique assemblage of species, enhancing the delivery of ecosystem services, and contributing to improving the lives of people in the region. van Wilgen et al. (1998) make the strong argument that the thorough cost–benefit analysis was a “major stimulus” for launching Working for Water. They further suggest that in cash-strapped countries in the developing world, such analyses are absolutely necessary for the implementation of environmentally driven policies.
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(b)
Weed density
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Figure 12.8 Population growth projections for scotch broom in a nature park in Australia. Included in the population model was the effectiveness of control measures given budget constraints. This gives a direct measure of the cost of such constraints on the reduction of the invasive and thus the benefits that such reduction brings about. Reprinted from Odom, D. I. S., O. J. Cacho, J. A. Sinden, and G. R. Griffith. 2003. Policies for the management of weeds in natural ecosystems: the case of scotch broom (Cytisus scoparius, L.) in an Australian national park. Ecological Economics 44: 119–135, with permission from Elsevier.
We have highlighted only a few of the more recent attempts to combine ecological knowledge with the social and economic issues surrounding control or eradication of well-established invasive species. There are many manuals and volumes available that detail specific chemical application procedures and manual removal options (e.g., Luken & Thieret 1996; Holloran 2004). As the Odom et al. (2003) example highlights, there is much room for interdisciplinary efforts here. Several authors have noted the similarity between modeling the control of invasive species and the control of infectious disease (e.g., Drake 2005). Fruitful cross-collaboration between these largely distinct disciplines should serve to reinvigorate each (Shea et al. 2000). In addition, the need to engage economists and environmental professionals in discussions of how to deal with invasive species is critical to achieving any lasting success. There are a variety of costs to be considered, and various social groups view the consequences of invasions quite differently. For example, non-native species may be considered useful in controlling erosion, nutrient cycling, as hosts for valued native species, and for jump-starting restoration of severely degraded ecosystems (Altieri 1995; Shapiro 2002). Balancing these competing views on invasive species, especially when simultaneously considering the controversies and costs associated with control options can be more than daunting. Invasion ecologists are thus called upon to embrace a variety of approaches within their own discipline and collaborate with others well outside the discipline. This is a tall order to be sure.
Summary Ecologists have traditionally concerned themselves with explaining patterns and mechanisms of invasions. Although there is a long history of trying to predict invaders, it is
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only very recently that measurable progress has been made. This progress is a direct result of developing more quantitative approaches to risk assessment and prediction, and integrating social and economic factors into these systems. There are clear needs for more accurate understanding of invasion vectors, and how the risk of invasion changes following shifts in vector traffic. There is an analogous need to erect early-warning systems that are capable of identifying nascent invasion foci of known or suspected, harmful invaders. Eradication efforts are often successful if implemented while the nonnative population is small; however there are a variety of social and economic forces that inhibit fast response. Finally, ecologists have much to offer those engaged in controlling, or eradicating, well-established invasive species. There are several good examples of how standard ecological modeling can inform on-the-ground efforts at structuring control activities. Indeed, there is considerable room for interdisciplinary-minded students to contribute to these issues.
Further complexities Drake, J. M. 2005. Risk analysis for invasive species and emerging infectious diseases: concepts and applications. American Midland Naturalist 153: 4–19. Hoddle, M. S. 2004. Restoring balance: using exotic species to control invasive exotic species. Conservation Biology 18: 38–49. Kolar, C. S. and D. M. Lodge. 2001. Progress in invasion biology: predicting invaders. Trends in Ecology & Evolution 16: 199–204. Louda, S. M. and C. W. O’Brien. 2002. Unexpected ecological effects of distributing the exotic weevil, Larinus planus (F.), for the biological control of Canada thistle. Conservation Biology 16: 717–727. Pheloung, P. C., P. A. Williams, and S. R. Halloy. 1999. A weed risk assessment model for use as a biosecurity tool evaluating plant introductions. Journal of Environmental Management 57: 239–251. Peterson, A. T. 2003. Predicting the geography of species’ invasions via ecological niche modeling. Quarterly Review of Biology 78: 419–433.
Companion papers Daehler, C. C., J. S. Denslow, S. Ansari, and H. Kuo. 2004. A risk-assessment system for screening out invasive pest plants from Hawaii and other Pacific Islands. Conservation Biology 18: 360–368. Drake, J. M. and D. M. Lodge. 2004. Global hot spots of biological invasions: evaluating options for ballast-water management. Proceedings of the Royal Society of London, Series B Biological Sciences 271: 575–580. Odom, D. I. S., O. J. Cacho, J. A. Sinden, and G. R. Griffith. 2003. Policies for the management of weeds in natural ecosystems: the case of scotch broom (Cytisus scoparius, L.) in an Australian national park. Ecological Economics 44: 119–135. Wadsworth, R. A., Y. C. Collingham, S. G. Willis, B. Huntley, and P. E. Hulme. 2000. Simulating the spread and management of alien riparian weeds: are they out of control? Journal of Applied Ecology 37: 28–38.
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Index
Acacia mearnsii, 95 Accipiter nisus, 162 acclimatization societies, 24–26, 41, 72 Achantia fulica, 196 Achantina fulica, 66–67 Achatina fulica, 116 Acridotheres tristis, 192, 225–227 Adelges tsugae, 156 Adenanthera pavonia, 110 admixture, 231, 233 Adoratus sinecus, 237–238 advection term, 144 adventive, 3 African ant, 129 African giant snail, 196 Agaonidae, 118 alien, 7 Allee effect, 64–65, 154, 160, 168, 169–170, 178–180 allelopathy, 191 American chestnut, 200–201 Anas platyrhynchos, 24, 185 Anas superciliosa rogersi, 185 Anas superciliosa superciliosa, 185 Anas wyvillana, 185 Anolis sagrei, 231 Anoplolepis gracilipes, 128 Anoplophora glabripennis, 258 Antarctic, 102–103 Apis mellifera, 173 Arabis fecunda, 190–192 Argentina, 109, 116–117, 120, 194 Argentine ant, 104–105, 147, 176, 221 Arytainilla spartiophila, 68–69 Asclepias syriaca, 187 Asian longhorn beetle, 258, 262 atmospheric carbon dioxide, 100–105 Australia, 50–51, 66, 72, 124, 163, 185, 192, 246–247, 264–265, 266
Australian black duck, 185 Australian paperbark tree, 93 bait-bucket introduction, 23 ballast ballast water exchange (BWE), 61–62 definition, 61 exchange, 256–257 regulations, 61 soil, 28 solid, 61 water, 28–29, 32, 36, 61–62, 166, 256–257 Batrachochytrium dendrobatidis, 214 beaver, 217 biocontrol, 16, 27, 65–69, 153, 169, 170–171, 196, 207, 230, 263, 264–265, 266 biotic inertia, 110 biotic resistance, 72–74, 108–122 Boiga irregularis, 10–14 Bombus spp., 173 boom and bust, 181–182 Brazilian pepper tree, 93, 173 Brazilian green scale, 129 Bromus diandrus, 171, 175, 208–209 Bromus tectorum, 119, 147, 149–150 brook trout, 23, 186 broom, 68–69 brown anole, 231 brown treesnake, 10–14, 29 brown trout, 199 Brownian motion, 138 Bubulcus ibis, 151 Bufo marinus, 66–67 bull trout, 186 bullfrog, 126–127 California cordgrass, 186 California golden trout, 185 California poppy, 124
300
Index
Canada thistle, 170–171 Cancer productus, 116 cane toad, 66–67 Capra hircus, 196 Carcinus meanas, 116, 174, 207 Carduus nutans, 170–171, 264–265 Carpobrotus edulis, 82–83 Carpodacus mexicanus, 160 Carrichtera annua, 124 Castanea dentata, 200–201 Castor canadensis, 217 casual, 3 catclaw mimosa, 163 cattle egret, 151 Caulerpa taxifolia, 27, 28 cayeput tree, 93 Centaurea maculosa, 190–192 Centaurea solstitialis, 173 Cerion spp., 28 chaparral, 82–83 cheatgrass, 119, 147, 149–150 chestnut blight, 30, 200–201 Chile, 229 Chinese mitten crabs, 23 Chinese rose beetle, 237–238 Chinese tallow tree, 237–238 Christmas Island, 116, 120, 128 chytrid fungus, 214 Cirsium arvense, 170–171 Cirsium canescens, 171 Clausena excavata, 110 Clements, F., 77 cline, 229–230, 237 Coccus viridis, 129 collard dove, 151 colonization rate, 20 colonizing, 3 Columbiforms, 41 common honey bee, 173 common myna, 192, 225–227 common reed, 81, 85, 233 community impacts, 194 companion papers (defined), 16 competition, 109–115, 122–125, 153, 171, 190–192, 208–209, 217, 232, 237 Conyza sumatrensis, 120 cottony-cushion scale insect, 66 crimson rosella, 192 crop signatures, 92 cross-stage formula, 220–222 Cryphonectria parasitica, 200–201 cryptic invader, 49 cryptogenic, 3 Cytisus scoparius, 68–69, 265 Czech Republic, 42, 136
Danaus plexippus, 187 Darwin, C., 24, 77, 123 density dependence, 137, 138, 140, 153–154, 160 density independent, 158 detection, 260–261 detection threshold, 178, 179 disease model, 155 dispersal hitchhike, 164 human assisted, 20, 165 natural, 20 passive, 163 unidirectional, 166 kernel, 138, 143–144, 148–153 disturbance agriculture, 90–93 alter physical forces, 95–99, 111 biotic, 94–100 characteristics, 77–79 definition, 77 facilitation, 79–81 fire, 82–83, 85–86, 95, 96–99 invaders as disturbance, 99–100 scale, 78–79 straw man, 79–87 urbanization, 90–92, 94 divergence, 237–238 diversity, 111 diversity/invasibility debate, 112–115 dog-strangler vine, 187 donor region, 59 Dreissena polymorpha, 125, 166–168, 194, 214, 257 Drosophila subobscura, 229–230
eastern rosella, 192 ecological imperialism, 51 ecological niche modeling, 104 ecosystem engineer, 95 ecosystem impacts, 199–202 effective population size, 232 Elton, C. S., 3–5, 50–52, 109, 111, 112, 133 Entomophaga maimaiga, 169 Equus caballus, 196 eradication, 259–260, 261–262 Eriocheir sinensis, 23 escape from natural enemies hypothesis, 123–124 escaped, 3 Eschscholzia californica, 124 establishment, 53, 55 Euglandina rosea, 66–67, 196 Eurasian watermilfoil, 125
Index European green crab, 116, 174, 207 European starling, 26, 151, 162, 192 Everglades, 93 evolution, 16, 24, 123–124, 181, 223–240 evolutionary momentum, 233 exotic, 7 exponential growth, 158 extinction, 64, 190, 194–196, 200, 202 facilitation, 120, 122–129, 174 fat tails, 143, 152, 154 feral pig, 99–100 Ficedula hypoleuca, 162 Ficus, 118 fig tree, 118–119, 128, 179 fire cycle, 96–98 fire tree, 201 flow regime, 83–85, 90, 95 fluctuating resource hypothesis, 80–81, 115 food web effects, 212–213 foreign, 3 functional group, 111, 114 fungi, 29–31 further complexity readings (defined), 16 fynbos, 266 Galliforms, 41 Gambusia affinis, 90 Gecarcoidea natalis, 116, 128 Gemma gemma, 207 genetic algorithms for rule production (GARP), 168, 252–253 analysis, 149–150 diversity, 224–225, 227, 231–235 drift, 223, 227 effects, 184–186, 203, 210, 221 modified organism (GMO), 236 variability, 225, 230–231, 233–234 geographic information system, 156 geographic patterns islands vs. mainland, 53, 56 latitudinal, 52, 54 Old vs. New World, 52–56 one-way transfer, 52 place of origin, 50–52 giant African snail, 66–67, 116 global climate change, 100–105 goat, 196 Golden Hinde, 48–49 golden mussel, 194 Gough Island, 20 grasses, 96 grazing, 89–90, 100, 126
301
great American interchange, 2 Great Lakes, 35, 52–53, 125, 166–168, 174, 194, 248, 251, 257, 262 Guam, 10–14 gypsy moth, 169 Hawaii, 20, 25, 40–44, 52–53, 98, 129, 171, 185, 201, 237–238, 247 Hawaiian duck, 185 Hawaiian fire tree, 173 hemlock, 156 hemlock woolly adelgid, 156 Heracleum mantegazzianum, 263 herbivores, 196 Herpestes auropuntatus, 66–67 Herpestes javanicus, 27, 238 heterogeneity, 173–177, 180 heterozygosity, 227, 234 Himalayan bushclover, 252–253 hitchhike, 29 home-grown non-natives, 8 homogenization, 202–203 Hordeum marinum gussoneanum, 171, 208–209 horses, 196 house finch, 147, 160 house sparrows, 236–237 hub and spoke, 35 human perception, 206–207, 210–213 hybridization, 31, 233–234 Hypericum perforatum, 22 ice plant, 82–83 Icerya purchasi, 66 immigrant, 3 impact definition, 206 ecological currency, 214–222 economic currency, 213–214, 215–216 models, 210–218, 220–222 Impatiens glandulifera, 263 imported, 3 Indian fleabane, 129 Indian mongoose, 27 indirect effects, 14, 120–121 individual impacts, 186–189 invasibility, 50, 57 invasion atlas, 258 invasion process establishment, 8–17 impact, 8–17 spread, 8–17 stages, 8–17 transport, 8–17
302
Index
invasion rate accumulation functions, 45–47 cumulative curve, 39–40, 45 forecasts, 45–47 idiosyncratic trends, 41 temporal trends, 39–40 invasional meltdown scenario, 95, 174–175 Japanese white-eye, 173 keystone species, 95 kudzu, 22 La Société Zoologique D’Acclimatation, 25 lady bird beetle, 66 lag time, 177–181, 232, 263 Lake Victoria cichlids, 195–196 Lates niloticus, 195–196 leptokurtic, 143 Lespedeza cuneata, 252–253 Leucena leucocephela, 110 Limnoperna fortunei, 194 Linepithema humile, 104–105, 176, 221 Littorina littorea, 210 Lotka-Volterra, 109, 153 Lymantria dispar, 169 Lythrum salicaria, 119, 123, 126, 173 mallard duck, 24, 185 management, 16, 242 marine fishes, 32 marine invertebrates, 48–49 mark-recapture, 139 mathematical models definition, 111–115 directional dispersal, 144–146 integrodifference, 148, 151, 152, 153, 154, 160 integrodifferential, 148–151, 154 intrinsic rate of increase, 137, 139, 140 long distance dispersal, 142–144, 152–153, 154, 162–163 maximum likelihood, 168 metapopulation, 154–155 reaction–diffusion, 138–141, 142, 145–146, 148, 151, 152, 160, 163 spatially explicit, 154–156 stratified dispersal, 146–147 time lag, 140, 177–181 utility of, 133–135, 175 variable dispersal kernel, 148–153 Mediterranean fruit fly, 159–160 Mediterranean Sea, 28 Melaleuca quinquenerva, 93
metapopulation(s), 31, 64, 154–155 Mexico, 54 Michaelis–Menton model, 45–46 milkweed, 187 Mimosa pigra, 163 monarch butterfly, 187 mongoose, 66–67 moonshine, 200 Morella faya, 173, 214 Mt. Sapphire rockcress, 190–192 Muntingia calabura, 110 muskrat, 133, 136–137, 141–142, 144, 151 mutualism(s), 117–120, 128–129, 172–173, 179–180 facultative, 117, 119, 128–129 obligate, 117–119, 128 mycorrhizae, 119–120 Myrica faya, 201 Myriophyllum spicatum, 125 nascent invasion foci, 259–260, 262 National Invasive Species Act, 257 natural enemy, 27 naturalized, 7 New Zealand, 25–26, 34, 68, 72, 73, 185, 193–194, 199, 256, 261, 264–265, 266 New Zealand grey duck, 185 New Zealand mud snail, 84 nile perch, 195–196 nitrogen, 201–202 nodding thistle, 264–265 non-indigenous, 7 normal distribution, 139, 143 novel genotype(s), 234 noxious, 3 nuisance, 3 ocean temperatures, 102 old field mice, 189 Oncorhynchus aguabonita aguabonita, 185 Oncorhynchus clarki seleniris, 185 Oncorhynchus mykiss, 23, 185 Oncorhynchus nerka, 238 Ondatra zibethica, 133, 136–137, 141–142, 144, 151 open niche, 13 optimal defense theory, 123 Oryctolagus cuniculus, 196 outbreak(s), 160 Ovis aries, 196 Pacific rat, 193–194 Paiute cutthroat trout, 185
Index paleontological record, 2, 21 Passer domesticus, 236–237 Passerines, 41 pathogen, 30–31 Pearl Harbor, 43 Peromyscus polionotus, 189 pest, 3, 27 Pheidole megacephala, 129 phenotypic plasticity, 87 Phragmites australis, 81, 85, 233 physical impacts, 194 Phytophthora infestans, 233 Phytophthora ramorum, 30–31 pied flycatcher, 162 pig, 196 Platte thistle, 171 Platycercus elegans, 192 Platycercus eximus, 192 Pluchea indica, 129 pollination, 118–119 Polynesia(ns), 20, 34–35 Ponto-Caspian Sea, 52–53, 248, 251 population bottleneck, 225–226, 228 density, 94 growth, 158–162, 169, 171, 177–179, 189 growth rate, 138 impacts, 189–194 positive interactions, 174–175 Potamopyrgus antipodarum, 84 predation, 10–14, 64, 115–117, 122, 123, 126–128, 153, 169–171, 189, 192–194, 195, 199, 217, 232, 237 prediction, 242–244 primary productivity, 114 priority effect(s), 115 probability density function, 140–141 propagule pressure, 94, 112, 114, 176, 180, 251, 258 propagule(s) behavior, 64–65 condition, 60 definition, 59 life history, 71–74 location, 60 pressure, 59, 65, 67–74 size, 59–60, 63–71 Psittacines, 41 Ptychocheilus grandis, 199 Pueraria montana var lobata, 22 Puerto Rico, 92–93 purple loosestrife, 173, 119, 123, 126 Pycnonotus jocosus, 173 quiet invader, 210
303
rabbits, 196 rainbow trout, 23, 185 Rana catesbeiana, 126–127 random walks, 138 range expansion, 2 range pinning, 161, 179, 180 Rattus exulans, 193–194 recipient region, 59 red imported fire ant, 189, 197, 233 red land crab, 116, 128 red rock crab, 116 red whiskered bulbul, 173 restoration, 86, 87–90 Rhinocyllus conicus, 170, 264–265 risk assessment, 156 ecological niche, 252–253, 262 phylogenetic association, 248 quantitative tools, 247–250, 254 regulation, 254–256 screening protocols, 246–247 stage-specific, 251–252 Rodolia cardinalis, 66 rosy wolf snail, 66–67, 196 ruderal, 3
Sacramento pikeminnow, 199 Salmo trutta, 199 salt cedar, 85, 95, 199 Salvelinus confluentus, 186 Salvelinus fontinalis, 23, 186 Sapium sebiferum, 237–238 Sargassum muticum, 27 satellite colony, 146–147, 165–168, 172 Schinus terebinthifolius, 93, 173 Science Citation Index, 6, 15 Scientific Committee on Problems of the Environment(SCOPE), 6–7, 186, 242 Scotch broom, 265 selection, 223 selective invader, 209 sheep, 196 Silene latifolia, 237 small Indian mongoose, 238 smooth cordgrass, 186 sockeye salmon, 238 Solenopsis invicta, 189 South Africa, 225–227, 266 sparrowhawk, 162 Spartina alterniflora, 179, 186 Spartina foliosa, 186 spatial heterogeneity, 114, 115 species characteristics, 212 species richness, 45, 111
304
Index
Sphendon punctatus, 193–194 Sporobolus airoides, 125, 208–209 spotted knapweed, 190–192 St. John’s Wort, 22 stochasticity, 63–64 Streptopelia decaocto, 151 Sturnus vulgaris, 26, 151, 162, 192 succession, 77, 92, 201 Sudden Oak Death, 30–31 Sus domestica, 196 Sus scrofa, 99 tamarind, 22 Tamarindus indica, 22 Tamarix spp., 85, 95, 199 tens rule, 14, 218–219 terminology, 7 thistle receptacle weevil, 264–265 Tibouchina herbacea, 171 tramp, 3 transformer, 3 transient, 3 translocated, 3 transplanted, 3 transport biocontrol, 27, 32, 40 conservation, 27–28, 32 environmental enhancement, 24–27, 32, 41, 94 food and game, 22–23, 32, 41, 94 human assisted, 19–21, 165 intentional, 21–22, 39–40, 42, 55–56, 257–259 natural, 19–20 pathway, 18–20, 34–36 pattern model, 55–56
unintentional, 21–29, 39, 42, 55, 165, 166–168 vector, 18–21, 29–34, 46–47, 55, 60–64, 225, 228–231, 254, 256–259 wooden ships, 48–49 traveling wave front, 141 triangular trade route, 18–19 Tsuga cannadensis, 156 Tsuga carolinensis, 156 tuatara, 193–194 twin invader, 209 value, 206–207, 213, 215–216 vernal pool(s), 82, 85–86, 89 Vincetoxicum rossicum, 187 waif, 3 Watersipora subtorquata, 175 weed, 3 weed risk assessment system (WRA), 246–247 West Nile virus, 188–189 western mosquitofish, 90 World Trade Organization, 256 World War I, 43–44 World War II, 10, 43–44 wrack, 81, 85 yellow crazy ant, 128 yellow starthistle, 173 zebra mussel, 125, 147, 166–168, 194, 199, 214, 257 Zosterops japonicus, 173