Seaweed Invasions

Seaweed Invasions Edited by Craig R. Johnson

Seaweed Invasions A Synthesis of Ecological, Economic and Legal Imperatives Reprinted from Botanica Marina Vol. 50 (2007) Double Issue 5/6

Edited by Craig R. Johnson

de Gruyter

Berlin · New York

Professor Craig R. Johnson School of Zoology and Tasmanian Aquaculture and Fisheries Institute University of Tasmania GPO Box 252-05 Hobart, Tasmania 7001 Australia E-mail: [email protected] http://fcms.its.utas.edu.au/scieng/zoo/ Cover images Front cover: Undaria pinnatifida on sea urchin barrens off the east coast of Tasmania, Australia. The photograph was taken in September 2000 at Lords Bluff within Mercury Passage. Courtesy of Dr. Hugh Pederson, Tasmanian Aquaculture and Fisheries Institute, University of Tasmania, Hobart, Australia. Back cover: Caulerpa taxifolia on a 3-m deep submarine cliff off the French Mediterranean coast. The photograph was taken in September 1991 at The Lavandou, Département Var. Courtesy of Professor Alexander Meinesz, Laboratoire Environnement Marin Littoral, Université de Nice-Sophia Antipolis, Nice, France. This work contains 18 figures and 13 tables. ISBN: 978-3-11-019534-7

Library of Congress Cataloging-in-Publication Data Seaweed invasions : a synthesis of ecological, economic, and legal imperatives / edited by Craig R. Johnson. p. cm. “Reprinted from Botanica marina, vol. 50 (2007), double issue 5/6.” Includes index. ISBN 978-3-11-019534-7 (pbk. : alk. paper) 1. Marine algae--Ecology. 2. Marine algae--Control. 3. Marine algae--Harvesting. 4. Invasive plants. I. Johnson, Craig R. (Craig Richard) QK570.2.S42 2007 363.7’8--dc22 2007044827 Bibliographic information published by the Deutsche Nationalbibliothek The Deutsche Nationalbibliothek lists this publication in the Deutsche Nationalbibliografie; detailed bibliographic data are available on the Internet at http://dnb.d-nb.de. ∞ Printed on acid-free paper, which falls within the guidelines of the ANSI to ensure permanence and durability. © Copyright 2007 by Walter de Gruyter GmbH & Co. KG, 10785 Berlin All rights reserved, including those of translation into foreign languages. No part of this book may be reproduced or transmitted in any form or by any means, electronic or mechanic, including photocopy, recording, or any information storage retrieval system, without permission in writing from the publisher. Printed in Germany. Typesetting: Compuscript Ltd., Shannon, Ireland; Printing and binding: druckhaus köthen GmbH, Köthen, Germany; Cover design: Martin Zech, Bremen, Germany.

Contents

Introduction Seaweed invasions: introduction and scope Craig R. Johnson and Anthony R.O. Chapman

1

Reviews Introductions of seaweeds: accidental transfer pathways and mechanisms Chad L. Hewitt, Marnie L. Campbell and Britta Schaffelke Intentional introductions of commercially harvested alien seaweeds Timothy D. Pickering, Posa Skelton and Reuben J. Sulu Mechanisms of invasion: establishment, spread and persistence of introduced seaweed populations Joseph P. Valentine, Regina H. Magierowski and Craig R. Johnson

6

18

Methods for identifying and tracking seaweed invasions Alexandre Meinesz

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Molecular approaches to the study of invasive seaweeds David Booth, Jim Provan and Christine A. Maggs

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Impacts of introduced seaweeds Britta Schaffelke and Chad L. Hewitt

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Control of invasive seaweeds Lars W.J. Anderson

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Invasive seaweeds: global and regional law and policy responses Meinhard Doelle, Moira L. McConnell and David L. VanderZwaag

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Conclusion

31

Mechanisms of invasion: can the recipient community influence invasion rates? Piers K. Dunstan and Craig R. Johnson 41

Seaweed invasions: conclusions and future directions Craig R. Johnson

131

Author information

139

Subject index

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Botanica Marina 50 (2007): 321–325

2007 by Walter de Gruyter • Berlin • New York. DOI 10.1515/BOT.2007.037

Introduction

Seaweed invasions: introduction and scope

Craig R. Johnson1,* and Anthony R.O. Chapman2 School of Zoology and Tasmanian Aquaculture and Fisheries Institute, University of Tasmania, GPO Box 252-05, Hobart, Tasmania 7001, Australia, e-mail: [email protected] 2 Department of Biology, Dalhousie University, Halifax, NS, Canada, B3H 4J1 1

* Corresponding author

Background and motivation At no time in human history has the need to understand invasions of alien species – the process of invasion, impacts of invasion and meaningful options to respond to invasion – been so urgent. The rate of anthropogenically-mediated translocation of species to regions outside native ranges has never been greater. This is particularly true for marine species, among which it is estimated that, at any point in time, several thousand are being transported between biogeographic regions in ballast water alone (Carlton and Geller 1993, Carlton 1999). Furthermore, there is good evidence to suggest that in some areas establishment of alien marine species originating from hull fouling exceeds that attributable to transport in ballast water (Hewitt et al. 2004). Not surprisingly, the rate of establishment of alien marine species, including invasives, appears to be increasing (e.g., Cohen and Carlton 1998, Ruiz et al. 2000), and some marine bays realise a newly established species every 30–40 weeks (Hewitt 2003). These trends are evident in marine, freshwater and terrestrial environments alike, and have raised considerable angst about the ecological, economic and social consequences (e.g., Pimentel et al. 1999, Mack et al. 2000, Sala et al. 2000, Lodge 2001, Bax et al. 2003). Seaweeds are a significant component of those marine organisms that have established as alien species in new bioregions, in some regions comprising ;5% of the total flora (Ribera and Boudouresque 1995) and ;10–40% of the total alien species (Schaffelke et al. 2006), and several species have been invasive (e.g., Nyberg and Wallentius 2005, this special issue). However, as is the case in studying many other kinds of invasive marine species (Grosholz 2002), investigation of the seaweed component has been dominated by case studies that are often strongly idiographic, focusing on high profile taxa that have, or might have, large ecological or economic effects. There has been little attempt to synthesise this body of work, either in the context of seaweed biology and ecology or more general invasion ecological theory. Our intention here is to go beyond the case studies and

the sui generis in search of patterns and commonalities. Of course, the case studies must be included for reference, for they comprise the knowledge base on the species and communities invaded. A deep understanding of the invasion process, impacts and options to manage invasions can only come from integrating observation of natural systems across a variety of scales with results of controlled experiments, and with ecological theory. A lack of integration of this kind and a focus on case studies and particular invasion events is arguably part of the reason for the historical disconnection between invasion ecology and mainstream ecological theory (Davis et al. 2001, see also Cadotte 2006). Notwithstanding Davis’ (2006) harsh criticism that there has been little change in the questions and answers about invasion ecology in four decades, we suggest that recent syntheses working towards a confluence of survey, experiment and theory have contributed important advances in the understanding of invasions (e.g., Levine and D’Antonio 1999, Shea and Chesson 2002, Bruno et al. 2005, Stachowicz and Tilman 2005, Fridley et al. 2007). It is now both opportune and necessary to attempt to consider invasive seaweeds in this concourse. Deep understanding also requires critical analysis of data and evidence. This sounds self-evident, but several authors have pointed out that, for example, putative claims of the impacts of alien invasive species are often unsupported by data or critical analysis (Gurevitch and Padilla 2004, Didham et al. 2005, MacDougall and Turkington 2005), and a recent review concludes there is little evidence that many alien invasive species cause the impacts and problems attributed to them (Bruno et al. 2005). While it is clear that some invasive marine species do have large impacts on the structure and dynamics of the systems in which they proliferate (e.g., Nichols et al. 1990, Carlton 1996, Shiganova 1998, Daskalov 2002, 2003, Ross et al. 2003), this review provides an opportunity to carefully examine available evidence of invasion processes and impacts for invasive seaweeds. The overall aim of this collection is thus to synthesise current information about invasive seaweeds and human responses to them, and attempt to consider seaweed invasions in the context of a broader thinking about invasion ecology. We consider the means, both accidental and intentional, by which seaweeds are introduced to new biogeographic domains, mechanisms of their invasion and impact, and practical approaches to tracking and controlling seaweed invasions. Because practical responses to seaweed invasions invariably take place within a regulatory framework, a review of legal and policy responses is also included as a fundamental element of the interaction between society and invasive seaweeds. Inevitably, this work is also about identifying gaps, and, therefore, challenges and priorities for the future. w1x

322 C.R. Johnson and A.R.O. Chapman: Seaweed invasions: introduction and scope

Important questions that are not unique to seaweed invasions provide a structure for examining whether generalisations can be drawn from the case studies, and these questions have framed the approach to this topic. They include: • What are the major modes of introduction of invasive seaweeds? • Is there tangible pressure for ongoing intentional introductions? • What are sensible approaches to reducing risk of further introductions? • Is it possible to predict the ‘‘next pest’’ seaweed? • Are there common life-history or genetic traits of successful invaders? • Why do some species become invasive while others do not? • Are there common mechanisms underpinning seaweed invasions? • Why do some communities appear to be more susceptible to invasion than others? Do the traits of the recipient community influence invasion rates? • How have seaweed invasions been tracked, and can existing approaches be improved? • Is it possible to predict the course of an invasion? • What are the ecological, genetic and economic consequences of seaweed invasions? • Can we expect that existing and, in particular, emerging techniques in genetics and genomics will provide a much deeper understanding of seaweed invasions? • How should humans respond to seaweed invasions? • Is the global regulatory framework in which responses to actual and potential seaweed invasions are determined adequate? We do not expect all of these questions to be answered with equal conviction, but we should be sanguine in establishing them as useful, if only to identify important areas of deficit in knowledge. The approach to these questions by most of the authors is strongly empirical, although the interface of some of these issues with ecological theory is considered where appropriate. Dunstan and Johnson (2007) in particular address the question of the properties of receiving communities in influencing invasion rates from a theoretical perspective, in part because of the dearth of meaningful empirical observations that contribute to the issue. Indeed, the extent of empirical knowledge of seaweed invasions is limited and highly skewed towards particular species. For example, 260 or so alien seaweed species have been identified (Schaffelke et al. 2006) but for only 17 have ecological impacts been considered at all and, arguably, for only four is there a solid empirical and experimental basis (Schaffelke and Hewitt 2007). Thus, we recognise that this synthesis, driven by questions relevant to applied and theoretical ecology, may be evanescent. Nonetheless it is overdue.

Modes of introduction In recognising that -3% of introductions of alien seaweed species are intentional, Hewitt et al. (2007) focus w2x

on reviewing modes of accidental introductions and identify hull fouling as the most significant, but also the most poorly managed, transport mechanism for seaweeds. They emphasise that while eliminating risk is rarely possible, there are several options for risk mitigation. They also address the challenge of identifying potential ‘‘next pests’’, and argue that this is best tackled based on assessment of risk at the three main stages of the invasion process (uptake and transport, establishment, spread) and not on particular properties of species. Later in the issue, Valentine et al. (2007) corroborate this stance in concluding that there is no evidence of a common suite of traits of invasive seaweeds, in line with suggestions two decades ago (e.g., Crawley 1987). But there is also pressure for further intentional introductions. This is driven by ongoing demand for seaweeds and their products, and perceptions that seaweed based industry offers an alternative and sustainable livelihood to coastal populations, particularly in developing nations (Pickering et al. 2007). Only a small number of seaweed species have been introduced intentionally, and rarely have these become particularly problematic wthe introduction of Undaria pinnatifida (Harvey) Suringar in Brittany may be a notable exceptionx. Importantly however, Pickering et al. (op. cit.) find that intentionally introduced seaweeds are no more or less risk prone than unintentionally introduced seaweeds. Not surprisingly then, they too consider a careful risk assessment essential when there are plans to introduce species for aquaculture purposes. Clearly, species being considered for aquaculture should not be on watch lists of invasive species maintained by government agencies, NGOs or intergovernmental organizations, or figure prominently in scientific literature on invasive seaweeds. Moreover, as Pickering et al. (op. cit.) acknowledge, impacts realised in one area may not be good at predicting those in another (see also Grosholz 1996, Schaffelke and Hewitt 2007). Another risk is that intentionally introduced specimens may harbour ‘‘hitchhikers’’, a problem that can only be dealt with by proper quarantine procedures. Notably, there are only two published reports of quarantining seaweeds.

Mechanisms of invasion and tracking invasions In addressing mechanisms of invasion (Dunstan and Johnson 2007, Valentine et al. 2007), it is useful to consider why some species become invasive while others do not. For example, of the many taxa of the genus Codium off Japan, only one has become a worldwide pest wCodium fragile ssp. tomentosoides (van Goor) P.C. Silva; Trowbridge 1998x. Historically, seaweed ecologists sought explanations for species occurrences in physiological attributes. Species tolerances to light, temperature and salinity, for example, were thought to explain patterns seen in nature (reviewed by Lu¨ning 1990), although by the 1970s interactions among species (e.g., competition, predation, facilitation) were recognised as major structuring agents of seaweed communities (Chapman 1986). In the same way, the first studies of invasive seaweeds concentrated on properties of the

C.R. Johnson and A.R.O. Chapman: Seaweed invasions: introduction and scope 323

invaders in predicting consequences of introduction at new sites, sometimes with disastrous consequences as occurred with the introduction of Undaria pinnatifida in Brittany (Meinesz 2007). However, as outlined earlier, Valentine et al. (2007) show that life history characteristics are of little value in predicting invasion. Rather they show that, in many cases, disturbances to native assemblages free resources and pave the way for alien species to establish at high densities. These facilitative disturbances may be grazers, storms or other invasive species. In Tasmania, for example, disturbance patterns can account for observations that some patches are invaded by Undaria pinnatifida at high densities (Valentine and Johnson 2003, 2004) while others nearby are not. Invasion patterns are also highly variable at much larger spatial scales. Codium fragile ssp. tomentosoides does not reach nuisance proportions in all of the communities in which it has been introduced. This subspecies is quite rare in subtidal waters of the eastern North Atlantic Ocean, whereas it forms meadows that can replace kelp forests in the western Atlantic Ocean (Chapman et al. 2002). In fact several species that are invasive elsewhere in the world are not pests in their native communities (Trowbridge 1998). These observations suggest that properties of the invaded community also determine the success of the invader. This topic is explored by both Valentine et al. (2007) and Dunstan and Johnson (2007), who suggest that this kind of variability can be explained by patterns of differential resource availability. Dunstan and Johnson’s (op. cit.) work is largely theoretical, but they argue that seaweed communities, which often manifest a dynamic mosaic of patches in space and time, have properties likely to show stronger responses to resource variability than to species richness or diversity of the recipient community per se. Means of tracking seaweed invasions have, on the whole, been notable for the simplicity of the technologies employed, with the possible exception of some kinds of remote sensing (Meinesz 2007). Exotic species have been tracked along the coasts of several countries, and these positive results have underscored the importance of public education programs and community involvement in initial detection. Indeed, the first occurrence of Caulerpa taxifolia (M. Vahl) C. Agardh in Tunisia was reported by a fisherman responding to a public awareness campaign that distributed 300,000 brochures across eight Mediterranean countries. Cartographical data from an informed public, along with expert sampling, has allowed tracking of the invasion pathways of two Caulerpa species. Sophisticated genetic techniques can potentially be helpful in tracking invasions, but have largely been employed, with considerable success, in identifying the initial source(s) of invasions, including cryptic ones (Booth et al. 2007).

Consequences of invasions Schaffelke and Hewitt (2007) review the ecological impacts of seaweeds on recipient communities. They catalogue a variety of impacts, but they also emphasise the limited scope of extant work, which covers remark-

ably few of the total number of known alien seaweed species, and as few as four invasive species in any detail. With the exception of studies in Tasmania (on Undaria pinnatifida), Nova Scotia (on Codium fragile ssp. tomentosoides) and Tuscany (on two invasive Caulerpa species co-existing with two introduced red turfing algae from Australia), there have been few comprehensive experimental works on seaweed invasion ecology. In most cases the mechanisms of observed ecological effects are unknown (Schaffelke et al. op. cit.). However, even the relatively limited amount of work to date shows that a given species might have very different impacts in different locales. Along with the rather limited information on ecological impacts, Schaffelke et al. (op. cit.) point out that there is surprisingly little known of economic impacts of seaweed invasions. Nevertheless, applied science studies have received government funding, resulting in considerable research effort, for example, in the Mediterranean Sea. There are even fewer studies of genetic consequences of invasive seaweeds, with most genetic investigations focusing on identifying source populations (Booth et al. 2007). The work that has been done reveals the complexity of underlying colonisation patterns and genetic impacts, which Booth et al. (op. cit.) categorise broadly as changes in population genetic structure and changes in genomic structure, for example, through hybridisation. In one example, molecular analyses revealed considerable genetic diversity within invader populations of Undaria pinnatifida, probably reflecting multiple introductions from different sources. This species is a vigorous invader, first found outside its native Japanese range in the Mediterranean Sea in 1971 (Meinesz 2007), but with subsequent invasion of the northeastern Atlantic Ocean, New Zealand, Australia, Argentina and the northeastern Pacific Ocean. However, it seems that its vigour as an invader cannot be related to its genetic diversity, in part because most of the other successful invasive seaweeds have experienced genetic bottlenecks and manifest greatly reduced genetic variation. Indeed, the highly invasive strain of Caulerpa taxifolia consists of male thalli that reproduce vegetatively. Clearly, invasive virulence does not depend on genetic diversification. At a genomic level, there is only a single unequivocal example of hybridisation involving an invasive seaweed. Fucus evanescens C. Agardh was introduced to the Oslofjord but it migrated south into the Baltic Sea where it hybridised with F. serratus L.; interestingly, the hybrids occur in a restricted hybrid zone on the shore. Although hybridisation involving invasive seaweeds is likely to be rare, the future for genomic-level research is nonetheless a bright one, with the prospect of revealing adaptive traits and genotypephenotype-environment interactions (Booth et al. 2007).

Human responses to seaweed invasions Unfortunately, about 97% of seaweed incursions are accidental (Hewitt et al. 2007) and usually occur in regions not subject to monitoring, so they escape early detection. In these cases, eradication is not likely to be successful and management measures may need to be w3x

324 C.R. Johnson and A.R.O. Chapman: Seaweed invasions: introduction and scope

invoked. However, efforts to manage invasive seaweed populations have not been very successful (Anderson 2007), in part because of the massive reproductive potential of many seaweeds (e.g., Chapman 1984, Schaffelke et al. 2005) and their capacity for relatively long distance dispersal (Reed et al. 1988, Kinlan and Gaines 2003). In contrast, the few eradication programs that aimed to completely extirpate an invader population have been highly successful, as occurred in response to invasion of the California coast by Caulerpa taxifolia and Ascophyllum nodosum (L.) Le Jolis (Anderson 2007). Success stories like this are rare because of long latency periods during which detection can be difficult, and because they are expensive. Even when an invasion is discovered early, immediate, coordinated and massive action using highly developed methodologies is usually necessary, and even the sampling design to detect every single invading individual usually requires research development effort. Nonetheless, eradication in concert with early detection emerges as a cost effective response (Anderson op. cit.). No consideration of human responses to invasive species would be complete without consideration of the regulatory environment that defines limits to movement, use and handling of alien species and, in some cases, responses to invasive aliens. Ultimately it is governments that determine responses to invasive species, not scientists or environmental agencies, but coordination among governments at a global scale is poor. For example, only Australia, New Zealand, USA, Canada, Switzerland and Germany have legislation controlling introductions for aquaculture. This legislation has developed from both global and regional policies including the Law of Sea, the Convention on Biodiversity, the International Convention on Wetlands, and the Convention on Migratory Species of Wild Animals (Doelle et al. 2007). These policies have led to global initiatives such as the development of a Code of Conduct for Responsible Fisheries by the Food and Agriculture Organization. The International Maritime Organisation recognises ships’ ballast water as a major vector for invasives, and there are guidelines for handling waste water and sediment in ships. However, fouling of ship hulls is a much more important vector for seaweeds than ballast, and changes in the composition of anti-fouling substances on ships’ hulls (mandated by the International Convention on Control of Harmful Anti-Fouling systems on Ships) will promote increased fouling by seaweeds as the use of toxic compounds in antifoulants such as tributyltin (TBT) are phased out. Importantly, none of the international conventions is self-implementing, even those issued by a close-knit political alliance like the European Union, and so national legislation and enforcement are required. Australia and New Zealand have taken this more seriously than other nation states but, in general, the development of effective legal and policy responses to invasive seaweeds is fragmented at both regional and global levels, and at an early stage of development (Doelle et al. op. cit.). In the end, control of seaweed or any other invasions should be a component of an integrated multi-faceted approach dealing with all problems in the marine enviw4x

ronment including, for example, overfishing, climate change, marine debris and habitat modification. Doelle et al. (op. cit.) point out that there are a raft of powerful regulatory tools available to employ, but there is yet a great deal to do. It will not be possible to turn the clock back 500 years to the more pristine conditions that once existed, but it is necessary, at least, to stop things getting worse.

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Pickering, T.D., P. Skelton and R.J. Sulu. 2007. Intentional introductions of commercially harvested alien seaweeds. Bot. Mar. 50: 338–350. Pimentel, D., L. Lach, R. Zuniga and D. Morrison. 1999. Environmental and economic costs of nonindigenous species in the United States. Bioscience 50: 53–65. Reed, D.C., D.R. Laur and A.W. Ebeling. 1988. Variation in algal dispersal and recruitment: the importance of episodic events. Ecol. Monogr. 58: 321–335. Ross, D.J., C.R. Johnson and C.L. Hewitt. 2003. Assessing the ecological impacts of an introduced seastar: the importance of multiple methods. Biol. Invasions 5: 3–21. Ribera, M.A. and C.-F. Boudouresque. 1995. Introduced marine plants, with special reference to macroalgae: mechanisms and impact. Prog. Phycol. Res. 11: 187–268. Ruiz, G.M., P. Fofonoff, J.T. Carlton, M.J. Wonham and A.H. Hines. 2000. Invasions of coastal marine communities in North America: apparent patterns, processes and biases. Annu. Rev. Ecol. Syst. 31: 481–531. Sala, O.E., F.S. Chapin III, J.J. Armesto, E. Berlow, J. Bloomfield, R. Dirzo, et al. 2000. Biodiversity scenarios for the year 2100. Science 287: 1770–1774. Schaffelke, B. and C.L. Hewitt. 2007. Impacts of introduced seaweeds. Bot. Mar. 50: 397–417. Schaffelke, B., M.L. Campbell and C.L. Hewitt. 2005. Reproductive phenology of the introduced kelp Undaria pinnatifida (Phaeophyceae, Laminariales) in Tasmania, Australia. Phycologia 44: 84–94. Schaffelke, B., J.E. Smith and C.L. Hewitt. 2006. Introduced macroalgae – a growing concern. J. Appl. Phycol. 18: 529–541. Shea, K. and P. Chesson. 2002. Community ecology theory as a framework for biological invasions. Trends Ecol. Evol. 17: 170–176. Shiganova, T.A. 1998. Invasion of the Black Sea by the ctenophore Mnemiopsis leidyi and recent changes in pelagic community structure. Fish. Oceanogr. 7: 305–310. Stachowicz, J.J. and D. Tilman. 2005. Species invasions and the relationships between species diversity, community saturation, and ecosystem functioning. In: (D.F. Sax, J.J. Stachowicz and S.D. Gaines, eds) Species invasions: insights into ecology, evolution and biogeography. Sinauer Associates, Sunderland. pp. 41–64. Trowbridge, C. 1998. Ecology of the green macroalga Codium fragile (Suringar) Hariot 1889: invasive and non-invasive subspecies. Oceanogr. Mar. Biol. Ann. Rev. 36: 1–65. Valentine, J.P. and C.R. Johnson. 2003. Establishment of the introduced kelp Undaria pinnatifida in Tasmania depends on disturbance to native algal assemblages. J. Exp. Mar. Biol. Ecol. 295: 63–90. Valentine, J.P. and C.R. Johnson. 2004. Establishment of the introduced kelp Undaria pinnatifida following dieback of the native macroalga Phyllospora comosa in Tasmania, Australia. Mar. Freshw. Res. 55: 223–230. Valentine, J.P., R.H. Magierowski and C.R. Johnson. 2007. Mechanisms of invasion: establishment, spread and persistence of introduced seaweed populations. Bot. Mar. 50: 351– 360.

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Botanica Marina 50 (2007): 326–337

2007 by Walter de Gruyter • Berlin • New York. DOI 10.1515/BOT.2007.038

Review

Introductions of seaweeds: accidental transfer pathways and mechanisms Chad L. Hewitt1,*, Marnie L. Campbell1 and Britta Schaffelke2,a 1 National Centre for Marine and Coastal Conservation, Australian Maritime College, PMB 10, Rosebud, Victoria 3939, Australia, e-mail: [email protected] 2 CRC Reef Research, PO Box 772, Townsville, QLD 4810, Australia

* Corresponding author

Abstract Macroalgae are a significant component of historic and modern invasions, with association to a wide variety of transport mechanisms. These transport mechanisms pose specific constraints on the ways by which species can be taken up, transported and released into a new environment. Currently operating transport mechanisms for marine macroalgae are either associations with intentional introductions (translocations for aquaculture, aquarium or live seafood trade) or accidental introductions (mainly as hull-fouling). A number of potential management options exist, including the development of international instruments and regional agreements. The development of treatment options for hull fouling, the most significant and poorly managed transport mechanism for macroalgae, is of urgent need. Our current ability to identify which species are likely to invade next is limited. However, an examination of the synergies between species’ functional traits, transport constraints, and recipient community attributes will likely provide possible options in the future. Keywords: aquaculture; ballast water; hull fouling; introduced macroalgae; packing material; risk management; risk mitigation; vectors.

Introduction The global transfer of marine species by human-mediated means both within and between non-contiguous biotic provinces is of significant concern for biodiversity conservation and the sustainable development of coastal and oceanic resources (e.g., Lubchenco et al. 1991, Carlton 2001, Ruiz and Carlton 2003). While this issue was recognised by early workers (Ostenfeld 1908, Elton 1958), significant progress on identifying patterns and processes has been made only in recent decades (e.g., Current address: Australian Institute of Marine Science, PMB 3, Townsville, QLD 4810, Australia. a

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Carlton 1985, 1996, 2001, Williamson 1996, Ruiz et al. 2000, Hewitt et al. 2004, Castilla et al. 2005). We can now say with certainty that no region of the world’s oceans has remained free of alien (non-indigenous) marine species (e.g., Carlton 1979, Cranfield et al. 1998, Cohen et al. 1998, Coles et al. 1999, Hewitt et al. 1999, 2004, Coles and Eldredge 2002, Hewitt 2002, Oresanz et al. 2002, Leppa¨koski et al. 2002, Lewis et al. 2003, Castilla et al. 2005, Wyatt et al. 2005). Similarly, alien marine species from all major animal, plant and algal phyla have been detected. Of these, macroalgae not only represent a large component of the globally introduced biota (e.g., Ribera and Boudouresque 1995, Lewis 1999, Ribera Siguan 2002, Schaffelke et al. 2006), but also represent significant economic and environmental risks for which we have limited post-incursion control and management options (e.g., Ribera and Boudouresque 1995, Thresher 1999, McEnnulty et al. 2001, Anderson 2007, Schaffelke and Hewitt 2007). Between 163 (Ribera Siguan 2002) and 260 (J. Smith, J. Schaffelke and C.L. Hewitt, unpublished data) macroalgal species are recognised introductions, with representatives from seven out of nine orders in the Chlorophyta, 16 out of 19 orders in the Rhodophyta, and eight out of 12 orders in the Phaeophyceae (Schaffelke et al. 2006). Of these, -3% of macroalgal introductions have been intentional releases. Several authors have reviewed the current status of macroalgal introductions, either as regional assessments of marine invasions (e.g., Lewis 1999, Ruiz et al. 2000, Verlaque 2001, Leppa¨koski et al. 2002, Oresanz et al. 2002, Hewitt et al. 2004, Castilla et al. 2005, see Figure 1), or as assessments of vectors associated with introduced macroalgae (e.g., Ribera and Boudouresque 1995, Wallentinus 1999, 2002, Ribera Siguan 2002, 2003, Schaffelke et al. 2006). Despite significant efforts, our understanding of the processes determining invasion success remains limited, specifically for the prediction of the most likely ‘‘next pests’’ (but see Chapman 1999, Hayes and Sliwa 2003, Nyberg and Wallentinus 2005, Dunstan and Johnson 2007). The scope of this review paper is restricted to accidental introductions of macroalgae; numerous intentional introductions have occurred for aquaculture purposes, which are covered elsewhere in this issue (Pickering et al. 2007). We will i) identify and discuss the constraints posed by individual transport vectors to the successful transport and inoculation of macroalgae; ii) discuss the life history characteristics of successfully introduced macroalgae in relation to the current transport mechanisms; and, iii) review relevant biosecurity (biological security) mitigation measures at both national and global scales. This review paper draws heavily on previous work, specifically on case histories that have provided indepth evaluations of the invasion process.

C.L. Hewitt et al.: Accidental introduction pathways of seaweeds 327

Figure 1 Number of recorded accidentally introduced taxa of macroalgae for each IUCN bioregion (see Figure 3; Kelleher et al. 1995); data from J. Smith, unpublished data).

Transport mechanisms Humans have undoubtedly transported species intentionally and accidentally for several thousand years (di Castri 1989). However, these movements are likely to have been spatially restricted and of relatively low frequency. The modern era of European expansion (post 1500 AD) has resulted in the massive transport and inoculation of species between non-contiguous biotic provinces (Crosby 1986, di Castri 1989). Transport mechanisms in the marine environment are largely associated with commerce and exploration. These include: woodenhulled vessel boring, fouling, dry and semi-dry ballast; steel-hulled vessel fouling and the transport of planktonic organisms and fragments in ballast water; the intentional

transfers of mariculture organisms (specifically oyster introductions) including the unintentional movement of associated organisms (e.g., Elton 1958, Carlton 1989, 1996, Ribera and Boudouresque 1995); the transfer of live, frozen and dried food products and the aquarium trade (e.g., Weigle et al. 2005); the use of biological material for packing (e.g., Ribera Siguan 2002, Miller et al. 2004); and scientific research. Many of these vectors have not been limited to single species movements but have often resulted in entire assemblages or communities of tens to hundreds of species being transported between disparate bioregions. These vectors of transport typically result in the unidirectional movements of species over long periods, inoculating new individuals or propagules for multiple generations (e.g., Carlton and Geller 1993, Ruiz et al. 2000, Hewitt et al. 2004). Several transport mechanisms have ceased to exist as significant vectors (e.g., wooden hull boring, dry and semi-dry ballast, accidental mariculture introductions), while others have become more apparent (e.g., ballast water) (see Figure 2). Similarly, some linkages or transport corridors between donor and recipient regions have ceased to exist, with new linkages developing between bioregions as new trade routes became established (Carlton 1985, 1989, 1996, Crosby 1986, di Castri 1989, Ribera and Boudouresque 1995, Campbell and Hewitt 1999, Hewitt et al. 2004). Each of these transport mechanisms has a unique set of constraints that act as selection criteria, influencing a species’ ability to successfully enter and survive the invasion process (Table 1). While it is virtually impossible to establish with a high level of confidence the link between an introduction and the mechanism of transport, examining the range of potential transport mechanisms will prove useful in detecting common patterns. Hull boring During the age of sail, organisms could bore into wood (e.g., teredinid bivalves) creating deep ‘‘galleries’’ in the hulls of vessels (Carlton 1985). These galleries created protected habitats in which encrusting, nestling, and motile species could have been found. These areas of the ship hull were completely subtidal and exposed to ambient seawater conditions of temperature and salinity, but lacked light, were protected from wave action and the influence of vessel speed (Table 1). Copper cladding or sheeting was frequently used to prevent the settlement

Figure 2 Number of introduced macroalgal species attributed to specific vectors. Note – some species are represented in multiple vectors; data from J. Smith, unpublished data).

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Table 1 Specific constraints associated with identified transport mechanisms. Transport mechanism

Uptake

Transport

Exposure Planktonic Association Shear (settlement/ phase with target stress uptake) period species or habitat Hull boring ● Hull fouling ● Dry and semi-dry ballast Water ballast Ballast tank sediments Aquaculture/mariculture Aquarium and live seafood trade Bait and packing material Scientific research Maritime equipment ●

● ●

●

● ● ● ● ●

of boring and fouling organisms. Copper is toxic to marine invertebrates, fish and algae (Anderson et al. 1991, Gnassia-Barelli et al. 1995, Diannelidis and Delivopoulos 1997) and continues to be used as an active compound in many antifouling paints. Since the advent of effective anti-fouling paints and the replacement of wooden hulls with steel, aluminium, and fibreglass, hull boring has virtually ceased to exist as a transport mechanism for merchant vessels. However, timber hulls are still common in recreational vessels and regional trading and fishing vessels, particularly in artisanal fisheries and developing countries. We have little indication as to which organisms were transported by hull boring, with the exception of shipworms (teredinid bivalves) and limnoriid isopods (Turner 1966). However, it is highly unlikely that this mechanism would have provided significant opportunities to macroalgae given the protected, dark nature of the galleries. In one of the few modern studies of wooden hull fouling, Carlton and Hodder (1995) observed the inadvertent collection and transport of crevicolous (species found in crevices or narrow spaces) species when the replica vessel Golden Hinde II settled into the mud of Humboldt Bay. However, no macroalgae were recorded. While some potential for hull boring continues to exist, this mechanism will not be considered further in this evaluation. Hull fouling Fouling organisms include both plants and animals that attach to the hulls (including the rudder, propeller, water intakes such as sea-chests and internal piping etc.) of vessels. During a journey, these species are exposed to variations in ambient sea temperature and salinity. They are also exposed to wave action and shear forces as a result of vessel speed. Historically, this vector was believed to have been significantly reduced for merchant vessels by the use of copper cladding, and, subsequently, with the development of anti-fouling paints and increased vessel speeds (Ribera and Boudouresque 1995, Carlton 1996). Recent studies, however, indicate that this vector remains an active and significant mechanism of transport for a variety of species including w8x

●

Desiccation Darkness Crushing Exposure to and changing physical environment stress ● ●

● ● ●

● ● ● ●

●

●

●

●

● ●

●

macroalgae (e.g., Lewis 1982, 1999, Womersley 1990, Carlton and Hodder 1995, Hay and Dodgshun 1997, Coutts 1999, Hewitt and Campbell 2001, Lewis et al. 2003, 2004, Hewitt et al. 2004). Virtually all macroalgae are capable of fouling hulls (Schaffelke et al. 2006). Hull fouling can occur either through recruitment onto the hull from planktonic (albeit short-lived) life history stages, direct attachment from adjacent surfaces or as attached drift (see Lewis et al. 2004). The constraints of the transport process (vessel speed, exposure to changing temperature and salinity regimes depending on the voyage) control the successful transport of species by this vector (Table 1). Additionally, the use of anti-fouling compounds has created hardselection pressures for species and ecotypes that demonstrate resistance to the active substances (e.g., copper, organotins). Several macroalgae are common members of fouling communities (Fletcher 1980). Ribera Siguan (2003) reported that of 189 taxa of alien marine algae and angiosperms reported worldwide, 39 were attributed to hull fouling as a likely vector of introduction and, of these, 31 were red algae. While the majority of macroalgae associated with hull fouling are small or have crustose or filamentous growth forms, several large species have been collected from hulls. Undaria pinnatifida (Harvey) Suringar (Phaeophyta) sporophytes )1 m have been observed attached to vessels, having survived extensive sea voyages (Hay 1990), and its gametophytes have survived hull cleaning by scraping (Hay 1990, but see Wotton et al. 2004 for successful hull-cleaning by heat treatment). Larger macroalgae can also be transported during the microscopic phase of their life history (see Peters and Breeman 1992). Coutts (1999) detected Phloiocaulon species (Phaeophyta) and microscopic life history stages of macroalgae, plethysmothalli of Punctaria species (Phaeophyta), attached to hulls of international commercial merchant vessels visiting the port of Launceston (Tasmania, Australia). Of significant interest was the detection of Phloiocaulon species, which are typically found in non-port habitats (i.e., deep water or shaded pools in shallow coastal waters) of Australia and South Africa (Coutts 1999).

C.L. Hewitt et al.: Accidental introduction pathways of seaweeds 329

Dry and semi-dry ballast Both dry and semi-dry ballast were used extensively prior to the twentieth century and can be considered to exist in the transport of dredged material as cargo for landfill or as ballast within hopper barges and dredges. This mechanism typically collected sand, cobble, and rock from intertidal and supra-tidal environments. Only species present in these habitats were available to this transport mechanism. Little is known about which species were transported by this vector. However, the cosmopolitan nature of many beach and shore fauna may be attributable to the large quantities of shoreline transported around the world via this vector. This transport vector imposed several constraints on macroalgal transport including desiccation, darkness, and physical abrasion and/or crushing (Table 1). However, it also favours species that use vegetative fragments to establish. Discharge of species was accomplished during the discharge of the ballast itself by shovelling material out of the holds. While many crustaceans and molluscs may have been introduced via this mechanism, only one brackish-water macroalga is suspected to have been introduced by this mechanism: Chara connivens Salzmann ex A. Braun (Luther 1979, Wallentinus 2002) but see also the discussion of Lewis (1999) for Gymnogrongus crenulatus (Turner) J. Agardh. This vector will not be considered further in this evaluation. Ballast water Water ballast was first proposed as a vector for the transport of species nearly 100 years ago (Ostenfeld 1908). Since then, numerous studies have detected thousands of taxa within the ballast tank environment (see Carlton and Geller 1993, Gollasch et al. 2002). This may well be the most generic vector of species transport, providing a mechanism for holo-, mero- and tycho-planktonic organisms, as well as demersal species (e.g., Carlton and Geller 1993, Lavoie et al. 1999). Williams et al. (1988) found over 69 taxa in five animal phyla wArthropoda, Chordata (ascidians and fish), Cnidaria, Mollusca, and Platyhelminthesx while Carlton and Geller (1993) reported 367 taxa, including taxa from three macrophytic phyla (including the angiosperms) – representing all of the major trophic groups from infaunal soft and hard bottom epifaunal, epibiotic, and planktonic habitats. The constraints limiting transport by ballast water have been reviewed extensively elsewhere (Carlton 1985, 1996, Carlton and Geller 1993, Ribera and Boudouresque 1995) and include the availability of planktonic (including tychoplanktonic) material at the time of uptake; the crushing shear stresses of the ballast pump, and the ability to survive long periods in darkness as unattached material (Table 1). Several macroalgae are capable of meeting these criteria, particularly their microscopic stages, propagules or vegetative fragments. Ballast tank sediments (includes sea sediments) The uptake of sediment-laden (turbid) water coupled with the natural mortality of plankton in ballast tanks, leads to the development of sediment layers in the base of ballast

tanks and on horizontal surfaces. These sediments can contain benthic organisms including the resting stages of toxic dinoflagellate species and benthic diatoms (e.g., Gollasch et al. 2002). Additionally, ballast tanks and seachests can have an established and reproductive resident benthic fauna (Coutts et al. 2003). While this mechanism is active for microalgal species, no macroalgal species have been identified from ballast tank or sea-chest sediments. This transport mechanism will not be considered further in this evaluation. Aquaculture associates Several authors have suggested that this transport mechanism, particularly associated with oyster culture, is responsible for the majority of global macroalgal introductions (Elton 1958, Neushul et al. 1992, Wallentinus 2002, Ribera Siguan 2003). Numerous unequivocal examples exist: the introduction of Pacific oysters wCrassostrea gigas (Thunberg 1793)x into the Northeast Pacific Ocean resulted in the establishment of Sargassum muticum (Yendo) Fensholt (Scagel 1956, Critchley et al. 1990); the transport of C. gigas into the Thau Lagoon (France, Mediterranean Sea) resulted in at least nine macroalgal introductions (Verlaque 2001, Verlaque et al. 2005) including Saccharina japonica (J.E. Areschoug) C.E. Lane, C. Mayes, Druehl et G.W. Saunders, Sargassum muticum, Undaria pinnatifida, Porphyra yezoensis Ueda, and Grateloupia. Species which are intentionally released for stocking or aquaculture purposes are pre-selected for likely establishment in the new environment and are typically released at high inoculation densities, often with multiple releases through time to ensure establishment (see Pickering et al. 2007). The accidentally transported species growing attached to, boring in, or living inside the target organisms, or associated with the transported substrata, may experience some transport constraints. However, these are likely to be ameliorated due to the strong commercial incentive to keep the target species alive during transport. Uptake of associated species would be predicated on ecological association with the target species and the absence of management or quarantine measures to prevent transport and release of associated species (see risk mitigation below). The ability to survive transport would be related to the similarity in physiological tolerances between the target species and the associated species (Table 1). As a result, this transport vector is unlikely to have a consistent suite of constraints. Aquarium trade The aquarium trade, either for small-scale hobbyists or for large-scale commercial aquaria, is a significant commercial industry with well-prescribed quarantine practices. In Australia and New Zealand, import is based on risk assessments that identify a suite of ‘‘approved’’ species that must comply with agreed quarantine standards (through Import Health Standards wIHSx, Australia) including specific quarantine periods and containment requirements. To be exempted, non-approved species must undergo an individual risk assessment process. However, these measures have been developed to prevent the w9x

330 C.L. Hewitt et al.: Accidental introduction pathways of seaweeds

inadvertent introduction of fish pathogens and parasites and, consequently, do not deal effectively with macroalgae, especially if these are imported as target species. Despite quarantine and IHS efforts, a number of high profile macroalgal introductions, many through accidental escapes, have been associated with this transport mechanism. The most sensational has been the introduction and spread of Caulerpa taxifolia (Vahl) C. Agardh in the Mediterranean Sea (Meinesz et al. 1995) and subsequent introductions in California (Williams and Grosholz 2002), and the southern waters of Australia (Schaffelke et al. 2002). Aquarium species can be also purchased via the Internet and enter many countries by post or courier mail. This import pathway is difficult to control, and, consequently, numerous non-approved species bypass mandatory import requirements. For example, C. taxifolia has been intercepted entering New Zealand illegally in courier mail after being purchased via the Internet. There are few uptake and transport constraints placed on the intentional transport of allowed macroalgae in the aquarium trade (Table 1), as the commercial incentive is to ensure that the target species arrive in good condition. Most importations consist of species listed by common names (rather than scientific binomials; Weigle et al. 2005) possibly allowing numerous species to be imported under a single common name. Similarly, the currently allowed import of ‘‘live rock’’ (i.e., rock that has not been sterilised and contains numerous species) poses a significant risk (Fossa˚ and Nilsen 1996, Wallentinus 2002). Several species of the genera Caulerpa, Ceramium and Gigartina have been identified from live rock (Frisch and Murray 2002). The aquarium industry itself is generally regulated. However, once a species is sold to a private person it is under less stringent control, and in some countries, not controlled by regulations. This becomes an issue when aquarium species are discarded (‘‘released’’) into the local environment when no longer wanted. This ostensibly compassionate gesture will, at its best, see the species released into an unsuitable environment resulting in quick mortality; at its worst, it will result in release into a suitable environment where the species can thrive to become invasive. Consequently, this mechanism of transport must be considered as a high risk for the transport of macroalgae. Packing material Marine macrophytes (both macroalgae and seagrasses) are commonly used as packing material for the transport of live bait and live seafood (e.g., abalone, clams, lobsters, and oysters). Packing material for bait is the most likely to be discarded in the marine environment (Wallentinus 1999, Miller et al. 2004). However, packing material used for transporting live seafood has also resulted in the establishment of species in the marine environment. Typically, packing material is not covered under IHS and therefore slips through biosecurity or quarantine detection. Numerous packing material-associated invasions have been recognised in the Mediterranean Sea (Wallentinus 1999, 2002, Ribera Siguan 2002), the NE Pacific Ocean (Zostera japonica Aschers. et Graebner: Harrison and Bigley 1982, Ascophyllum nodosum (L.) Le Jolie: Milw10x

ler et al. 2004), and the NW Atlantic Ocean wCodium fragile (Suringar) Hariot ssp. tomentosoides (Van Goor) Silva: Carlton and Scanlon 1985x. This mechanism of transport continues to exist, especially in domestic transport between biogeographic regions we.g., in France between the NE Atlantic Ocean and Mediterranean Sea; in North America (USA and Canada) between the NW Atlantic and NE Pacific Oceans; in Australia between Pacific and Indian Oceans; bioregions after Kelleher et al. 1995x. Much like the transport of aquaculture species, this mechanism is intended to keep the target species (bait or live seafood) alive, hence, the species used for packing material is likely to survive the transport process (Table 1). Association with marine and maritime equipment Observations and anecdotal evidence indicate that introduced species often become entangled in fishing gear such as nets and ropes, anchor ropes and chains (e.g., Carlton and Scanlon 1985, Trowbridge 1995, 1996, 1998, Relini et al. 2000), possibly leading to further spread of these species. Macroalgal species tolerant to emersion could be successfully transported by this vector. For example, Caulerpa taxifolia and Codium fragile ssp. tomentosoides survive emersion in high humidity for up to 10 and 90 days, respectively (Sant et al. 1996, Schaffelke and Deane 2005). Freshwater diatoms can be transported attached to equipment such as fishing gear, and spores or gametes of macroalgae could be also transported in a similar fashion, especially from areas with high abundance of introduced macroalgae and during periods of high reproductive output. Rapid settlement of Undaria pinnatifida zoospores was shown on glass slides suspended from ropes in an infested Tasmanian bay (B. Schaffelke, unpublished data). Constraints for this pathway are desiccation and, possibly, freshwater stress, e.g., by washing gear before re-immersion in salt water. Scientific research Several intentional introductions of macroalgae for scientific research, especially aquaculture research, have been recognised (see Pickering et al. 2007), many with subsequent escapes (e.g., Ribera and Boudouresque 1995, Wallentinus 1999, 2002, Ribera Siguan 2002, Sulu et al. 2004) and documented impacts (Schaffelke and Hewitt 2007). This vector includes translocation of target and associated species (Table 1). For example, Russell (1983) identified four macroalgae that were transported from Kaneohe Bay, Hawaii, USA to Fanning Island, Kiribati with a commercially cultivated species, Kappaphycus alvarezii (Doty) Doty ex P.C. Silva. Like aquaculture translocations, this mechanism is likely to ensure high survivorship because effort is made so that the target species arrive in good condition.

Risk mitigation and management Macroalgal introductions continue to occur through a variety of transport vectors and pathways in most regions

C.L. Hewitt et al.: Accidental introduction pathways of seaweeds 331

of the world (e.g., Ribera and Boudouresque 1995, Lewis 1999, Ribera Siguan 2002, Schaffelke et al. 2006). Several transport mechanisms are either under current management or regulation at national or global scales (e.g., Carlton 2001, Bax et al. 2003, Hewitt 2003, Weigle et al. 2005, Doelle et al. 2007), or are in the final stages of negotiation we.g., International Convention for the Management and Control of Ships’ Ballast Water and Sediments (BWM) Convention, IMO 2004, see Doelle et al. 2007x. These efforts are largely focused on a few transport mechanisms, with the establishment of agreed quarantine standards (through IHS) and treatment options. Risk mitigation addresses either the uptake of propagules at the source location (e.g., prevention of hull fouling by anti-fouling paints), or the release of viable material into the recipient environment (management of ballast water discharge, treatment of aquaculture stock). We discuss various management actions separately for intentional and unintentional introductions. Intentional introductions Achieving zero-risk is rarely possible, consequently, risk mitigation is the only acceptable management option. Intentional transport and introduction of species for aquaculture, aquarium trade or scientific research can result in the accidental release of target species through escapes from growing facilities, or in the accidental transport of non-target (associated) species. Intentional introductions, however, also provide opportunities for maximum control through appropriate management either prior to introduction or in the establishment and operation of the growing facility. The International Council for the Exploration of the Seas (ICES) has developed a Code of Practice (ICES CoP) for Introductions and Transfers of Marine Organisms (ICES 2003), which advises on how to evaluate the risk of introducing a marine organism, including determining the likelihood and consequences of also introducing associated species. The ICES CoP recommends a risk assessment procedure, a decision tree that includes feedback loops for seeking additional information, and recommends decision criteria. It has been applied successfully in a variety of circumstances and in several regions (Mediterranean Sea; NE Atlantic Ocean; NW Atlantic Ocean; SE Pacific Ocean). However, this requires a significant investment in resources and remains fundamentally reliant on the availability of expertise and political ‘‘good will’’. As has been noted, the main difficulty with the ICES CoP is that it is not a legally binding instrument. In the case of species associated with intentional movements of target species (aquaculture, aquarium trade, scientific research), control can be aimed at a variety of points in the invasion process. The probability of ‘‘uptake’’ of associated species can be controlled by maintaining specimens in a quarantine or containment facility, perhaps combined with cleaning methods (e.g., Sulu et al. 2004). This would be equivalent to a stagegate approach whereby the target species could not be transported prior to demonstration that there are no associated macroscopic species (note that disease is controlled under other mechanisms).

Similarly, the selection of the life history stage for transport of the target species is a significant control point. For invertebrates such as the Pacific oyster, Crassostrea gigas Thunberg, transport as D-stage larvae is less expensive, more effective, and significantly minimises the risk of transporting associated organisms. In contrast, the transport of adults has demonstrably introduced significant numbers of associated species, despite significant efforts to ‘‘clean’’ the valves prior to transport (Wallentinus 1999). Lastly, consideration should be given to the aquaculture (or mariculture) of non-indigenous target species in controlled, land-based facilities with appropriate filtration and sterilisation measures prior to effluent discharge into the ocean. In situations where open ocean release is deemed acceptable, the target species transported as adults should be maintained in a quarantine facility as brood stock, with the release of F1 generation material allowed only after demonstration of no infection by diseases, pathogens or other associated species, such as macroalgae. Caution must be taken to ensure that quarantine facilities are sited where the impact of natural disasters that possibly breach containment (e.g., cyclones, earthquakes, floods, tsunamis) is reduced (E. Gonzalez personal communication). Similarly, adequate filter systems are required for facilities that culture micro- and macroalgae. Such conditions should be included in IHSs but unfortunately this is often considered too prescriptive. Unintentional introductions In contrast to the intentional movements of target species, management of species translocations associated with commercial shipping (including slow-moving oil platforms, ocean going barges and tugs) and recreational vessels is more problematic. At present, international opinion has identified ballast water as the transport pathway with the highest chance for successful management in the immediate future. This mechanism has been perceived by the international community as the greatest threat (e.g., Carlton 1985, 2001, Ruiz et al. 2000, Hewitt 2003), leading to a significant effort to establish concerted national, regional and global management regimes including the adoption of a new International Convention for the Management and Control of Ships’ Ballast Water and Sediments (BWM, IMO 2004, see Doelle et al. 2007). In contrast to ballast water, hull fouling is currently a largely unmanaged transport mechanism (Minchin 2004) that has been demonstrated to be a significant transport pathway of large numbers of invertebrates and macroalgae (e.g., Rainer 1995, Ribera and Boudouresque 1995, Cranfield et al. 1998, Hewitt et al. 1999, 2004, Wallentinus 1999, 2002, Ruiz et al. 2000, Gollasch 2002, Minchin and Gollasch 2002, Ribera Siguan 2002, 2003). Currently available treatment options are limited to preventative measures (anti-fouling paints) or reduction measures (in water hull cleaning or dry docking). The use of anti-fouling paints with copper and organotin (e.g., tributyl tin; TBT) as active ingredients to prevent the settlement and growth of organisms has been effective at reducing the economic costs of fouling (i.e., increased fuel costs associated with increased drag). w11x

332 C.L. Hewitt et al.: Accidental introduction pathways of seaweeds

Unfortunately, this has not led to a concomitant elimination of biological risks. In addition, as discussed by Doelle et al. (2007), the International Convention on the Control of Harmful Anti-fouling Systems on Ships (AFS; IMO 2001), which aims to phase-out the use of organotin-based paints (specifically TBT), may result in an increase in hull fouling associated introductions, because alternative paints with equivalent efficacy are likely to be more expensive or require more stringent application procedures (Minchin 2004). Alternative paint technologies with appropriate application and renewal guidelines must be developed as a matter of urgency, with large effort being expended to date, but with little progress so far. In-water cleaning is a common mechanism for reducing fouling; diver-operated brush carts or physical abrasion using scrapers or brushes are common methods of reducing fouling between dry docking periods. This method is typically the only hull-cleaning treatment for small recreational vessels between anti-fouling applications. The removal of fouling organisms in-water often occurs in ports or in marinas, where the physical removal of organisms can lead to vegetative fragments or gametes/spores being released in areas conducive to their establishment. Current research on the development of biologically safe in-water cleaning techniques is underway in numerous countries, e.g., New Zealand, Australia and the USA, and Australia has adopted a ban on inwater cleaning in several ports. It should be noted, however, that removal of species that originate in a port or marina prior to departure would act to reduce the transfer of species between regions. Cleaning while in dry dock is the most effective procedure for removing organisms, and uses methods such as physical abrasion (scraping, water pressure), freshwater, steam cleaning, and desiccation. Dry dock cleaning is extremely effective, particularly when followed by the application of anti-fouling coatings. Unfortunately, many dry-docking facilities (particularly those suited to handling small vessels) do not have appropriate provisions to hold and treat removed materials, including viable hull fouling organisms, and often discharge removed materials into coastal waters. The need to develop standards for dry-docking facilities with guidelines on appropriate treatment is urgent (see the ANZECC Code of Practice).

Prediction of new invaders The accurate prediction of likely invaders will continue to be a driving force in academic and applied invasion ecology (e.g., Hayes and Hewitt 1998, Hayes and Sliwa 2003, Nyberg and Wallentinus 2005, Johnson and Chapman 2007). The factors that contribute to the successful introduction of a species include species’ functional traits, characteristics of transport vectors and attributes of the recipient community. Most previous discussions concerning the identification of potential invaders have revolved around the development of species-specific invasion attributes (e.g., Baker 1965, Beddington et al. 1978, Carlton 1979, 1996, Arthington and Mitchell 1986, Bazzaz 1986, Lodge 1993, Ricciardi and Rasmussen w12x

1998). However, traits believed to be essential for a species to become a successful invader are also found in non-invasive conspecifics and congenerics (Paula and Eston 1987 for Sargassum species; Trowbridge 1995, 1996, 1998 for Codium fragile subspecies). As Crawley (1987, 1989) and Valentine et al. (2007) conclude, of the myriad factors which ultimately influence success and failure of an invasion, individual species’ characteristics contribute only a small amount and are unlikely to be useful predictors alone. The absence of knowledge about failed introductions hampers the accurate assessment of controlling factors (see Miller et al. 2007). We suggest that the three main stages of the invasion process (uptake/transport, establishment, spread) act as filters, with species’ attributes either being beneficial or detrimental to successfully pass the filter (Carlton 1985, 1996). These filters may either completely prevent certain characteristics of species or ecotypes from passing through, and/or partially prevent other characteristics from passing through. For example, a stenohaline species that is restricted to habitats of less than 15 psu would likely experience a ‘‘complete’’ filter in transiting the open ocean with salinity )30 psu. For the primary modern transport vectors, hull fouling and ballast water, it is crucial that we begin to understand these constraints. The implications of understanding these critical points for developing cost-effective and appropriate management activities are great (Hayes and Hewitt 1998). In this evaluation, we undertook to delineate the constraints placed on uptake and transport of species, but not to determine the attributes for successful establishment and spread of populations. Uptake constraints consist of presence of the species and availability of suitable life history stages in the donor location, and exposure to the transport vector (Table 1). Several transport mechanisms are restricted to a limited set of the life history stages of a species. For example, ballast water is largely restricted to planktonic stages, with limited availability to benthic stages that have been ‘‘swept’’ off the substratum (tychoplankton). In many macroalgae however, detached individuals without reproductive structures are generally unable to develop sporogenous tissue while drifting walthough the invasive Sargassum muticum is a notable exception (Norton 1977, Deysher and Norton 1982)x. In contrast, transport constraints consist of physical and physiological constraints (Table 1). Physical constraints include shear stress (hull fouling) and crushing (dry ballast, gear transport, aquaculture). Physiological constraints include desiccation (dry ballast, high water hull fouling, gear transport, aquaculture, packing material), light limitation (dry ballast, water ballast, aquaculture, packing material) and exposure to varying environments such as changing temperature and salinity (hull fouling). Lastly, the role the recipient environment and community plays in determining the success or failure of invasions must be critically examined in the marine environment (see Dunstan and Johnson 2007). Theoretical (e.g., Case 1990, 1991, Drake 1990) and empirical studies in terrestrial and freshwater environments (e.g., Drake 1991, Burns and Sauer 1992, Robinson et al. 1995) suggest that ‘‘ecological resistance’’ may indeed

C.L. Hewitt et al.: Accidental introduction pathways of seaweeds 333

of authors (e.g., Stachowicz et al. 1999, Dunstan and Johnson 2004, 2007, Britton-Simmons 2006) have identified resource use and subsequent preemption as an important process in restricting invasibility of systems. Evaluations of habitat invasibility need to incorporate information about native biodiversity, trophic structures, competition strengths and niche occupancies. In a simple analysis based on recognised introductions of marine macroalgae (e.g., Ribera and Boudouresque 1995, Wallentinus 1999, 2002, Ribera Siguan 2002, 2003, Schaffelke et al. 2006; J. Smith, unpublished data), coupled with indications of their associations with currently active transport mechanisms, and their propensity for introduction (number of successful invasions into the IUCN large scale bioregions after Kelleher et al. 1995, see Figure 3), we identify twelve species that are likely to be of ongoing concern (Table 2). These are by no means the only species likely to be introduced in the future, nor does this simple analysis take into account the likelihood of causing harm. This merely identifies those species with a demonstrated propensity to surmount the constraints of the invasion process. Other analyses aimed at predicting future invaders list a similar suite of macroalgal species (e.g., Hayes and Sliwa 2003, Nyberg and Wallentinus 2005). An improved prediction of future invaders must take into account the synergy between the three aspects of invasions (vector strength, species’ traits, attributes of the receiving environment). In the face of an overwhelming suite of successful invading marine species with disparate traits and life history patterns, we must evaluate a species’ success or failure to invade in the light of

Figure 3 IUCN large scale bioregions after Kelleher et al. (1995). (1) Antarctica. (2) Arctic. (3) Mediterranean. (4) NW Atlantic Ocean. (5) NE Atlantic Ocean. (6) Baltic Sea. (7) Wider Caribbean Sea. (8) W Africa. (9) S Atlantic Ocean. (10) Central Indian Ocean. (11) Arabian Seas. (12) E Africa. (13) E Asian Seas. (14a and 14b) S Pacific Ocean. (15) NE Pacific Ocean. (16) NW Pacific Ocean. (17) SE Pacific Ocean. (18) Australia and New Zealand. (19) Great Lakes.

occur. Davis et al. (2000) propose that invasibility of terrestrial plant communities is a function of fluctuations in resource availability, coupled with sufficient inoculation. For macroalgae, critical resources would include substratum, light and nutrients, while fluctuations would be caused e.g., by physical disturbance, higher nutrient availability and changing herbivore pressure. A number

Table 2 Identified macroalgae that have a high likelihood of invasion based on prior history wG3 invaded IUCN Bioregions (bioregions after Kelleher et al. 1995, see Figure 3)x and a continued association with active vectors deemed to pose a continuing global threat of invasion*. Evaluation of intentional transport is limited to those species that are transported as associates of other intentionally transported species. Taxa

Chlorophyceae Caulerpales Codiales Phaeophyceae Dictyosiphonales Laminariales

Species and authority

HF

BW

E&G

Intentional transport (associates) Number of invaded IUCN A LT PM SR bioregions

Caulerpa taxifolia Codium fragile ssp. tomentosoides

● ●

● ●

● ●

● ●

● ●

Striaria attenuata (C. Ag.) Grev. Undaria pinnatifida

● ●

● ●

● ●

●

●

Rhodophyceae Bonnemaisoniales Bonnemaisonia hamifera Hariot Ceramiales Antithamnionella spirographidis (Schiffner) E.M. Wollaston Ceramiales Antithamnionella ternifolia (J.D. Hooker et Harvey) Lyle Ceramiales Heterosiphonia japonica Yendo Ceramiales Polysiphonia brodiaei (Dillwyn) Sprengel Ceramiales Polysiphonia harveyi J. Bailey Halymeniales Grateloupia subpectinata Holmes Halymeniales Grateloupia turuturu Yamada Rhodymeniales Lomentaria hakodatensis Yendo

Unintentional transport

●

●

3 6

5 5

● ●

● ●

4 3

●

●

3

● ●

●

3 4

● ● ●

● ●

●

● ● ●

3 3 3 3

*Note that this does not infer risk, as no evaluation of impact is incorporated into the analysis. Transport mechanisms are abbreviated as: HF: Hull Fouling; BW: Ballast Water; E & G: Equipment and Gear; A: Aquaculture; LT: Live Trade (Aquarium and Live Seafood); PM: Packing Material; SR: Scientific Research.

w13x

334 C.L. Hewitt et al.: Accidental introduction pathways of seaweeds

transport mechanisms acting as constraints and of other characteristics of the invasion process, such as recipient community attributes (see Dunstan and Johnson 2007). Within this framework, the field of potential invaders may be narrowed and ultimately predicted with increasing success.

Acknowledgements We thank Drs Craig Johnson and Tony Chapman for inviting our participation in this special issue. We would also like to thank John Lewis, Stephan Gollasch and Inger Wallentinus for fruitful discussions in preparation of this manuscript.

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2007 by Walter de Gruyter • Berlin • New York. DOI 10.1515/BOT.2007.039

Review

Intentional introductions of commercially harvested alien seaweeds Timothy D. Pickering1,*, Posa Skelton2 and Reuben J. Sulu3 School of Marine Studies, The University of the South Pacific, Private Mail Bag, Suva, Fiji Islands, e-mail: [email protected] 2 International Ocean Institute Regional Center for Australia and the Western Pacific, P.O. Box 1539, Townsville, Queensland 4810, Australia 3 Solomon Islands Center, The University of the South Pacific, P.O. Box 460, Honiara, Solomon Islands 1

* Corresponding author

Abstract The two main drivers for intentional introductions of commercial macroalgae are (1) increasing global demand for macroalgae and macroalgal products, and (2) increasing need for alternative and sustainable livelihoods among coastal communities in less-developed countries (particularly to reduce degradation of coral reefs) and in the less-developed rural areas of more-developed countries. The macroalgal species that form the basis for commercial aquaculture (mainly Saccharina japonica (J.E. Areschoug) C.E. Lane, C. Mayes, Druehl et G.W. Saunders, Porphyra species, Undaria pinnatifida, Kappaphycus alvarezii, and Gracilaria species) are thus the ones most likely to be intentionally introduced to other places. The highest-profile cases of ‘‘invasive’’ macroalgae have mainly resulted from unintentional introductions, particularly via shipping. Two cases are species important commercially for aquaculture, U. pinnatifida and K. alvarezii, although the global spread of U. pinnatifida beyond Asia has been caused mainly by shipping. K. alvarezii has been intentionally introduced to many countries for aquaculture and has been reported as invasive in one locality in Hawaii; however, more recently it has emerged that Eucheuma denticulatum is in fact the main culprit at this locality. While environmental problems from intentional introductions have been few compared with those from unintentional introductions, it does not appear that commercial species are, as a group, inherently any more or less risk-prone than most unintentionally introduced species. Only a minority of alien species may ever become invasive, but it is difficult to predict which will become pests. In principle, international norms allow states to intentionally introduce exotic species for commercial purposes, provided that environmental threats can be avoided. In practice, the burden of proof and duty of care about environmental threats and protection of biodiversity is nowadays much higher than before. States cannot take it for granted that alien species may be introduced; new proposals should follow formal risk assessment and w18x

monitoring processes that are science-based, and should be strongly justifiable in terms of ability to provide expected economic benefits. Keywords: commercial macroalgae; introductions; invasive; translocations.

Introduction We review intentional introductions of commercially harvested macroalgae for stock enhancement or cultivation in a geographical location that is outside the native range. This includes both introductions from one country to another (sometimes called ‘‘exotic introductions’’) and introductions to a location outside the native range but within the same country (sometimes called ‘‘translocation’’; Stickney 2002). Aquaculture and shipping are the two main vectors for introduction of alien marine species generally, and this also applies to macroalgal introductions (Schaffelke et al. 2006, Trowbridge 2006). Introductions of macroalgae via shipping can usually be regarded as unintentional. In the case of aquaculture, introductions of macroalgae can be either intentional or unintentional. Unintentional introductions attributable to aquaculture may result from escape of target species from ‘‘secure’’ culture facilities, or from introduction of non-target species (‘‘hitchhikers’’ or ‘‘stowaways’’) associated with the target species or with gear and equipment used for culture or transportation of the target species (Maggs and Stegenga 1999, Reise et al. 1999). Almost half of the unintentional macroalgal introductions documented in Europe are thought to be a result of transfer of shellfish for aquaculture (Wallentinus 2002).

Economic and social imperatives for commercial-macroalgal introductions Two main drivers can be identified for intentional introductions of macroalgae, viz. increasing global demand for commercially important macroalgae and macroalgal products, and increasing need for alternative and sustainable livelihoods among coastal communities in rural and less-developed areas. While attention nowadays is often focused more upon the adverse impacts of alien species, introductions have been regarded historically as a valid means to improve production and economic benefits from fisheries and aquaculture. Large industries generating thousands of jobs and millions of dollars in income are now based upon aquaculture of alien species, for which relevant communities perceive their advantages as outweighing any disadvantages. These introductions took place, how-

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ever, in an era when awareness about possible environmental risks, and the unpredictability of risk, was much lower. An overview of the history and recent status of global seaweed utilization is provided by a range of books and reviews, such as Bird and Benson (1987), Critchley and Ohno (1998), McHugh (2002, 2003), FAO (2004), and Critchley et al. (2006). The main macroalgal species now being commercially cultivated (and hence likely to be intentionally introduced to other places) are Saccharina japonica J.G. Areschoug, Porphyra species, Undaria pinnatifida (Harvey) Suringar, Kappaphycus alvarezii (Doty) Doty ex P.C. Silva, and Gracilaria species. Macroalgae form a large component of global aquaculture production in terms of tonnage but, since they are of lower value than other aquaculture products such as crustaceans or molluscs, a disproportionately smaller component of value. It might then be expected that the economic imperatives behind efforts to introduce macroalgae beyond their native range will also be lower than for other aquaculture species, nevertheless there is unsatisfied demand globally for macroalgae and macroalgal products. This is one basis for continued interest to introduce or expand macroalgal cultivation in a number of places around the world. The demand for macroalgae and their products stems from their uses as food or food constituents, and as pharmaceuticals or nutraceuticals. The great majority of seaweed production is for food (as a sea vegetable) or for food additives like carrageenan, agar or alginates (phycocolloids), and these are the main traditional uses of macroalgae. In FAO (2004) the situation is summarized thus: ‘‘The seaweed industry provides a wide variety of products that have an estimated total annual production value of US$ 5.5–6 billion. Food products for human consumption contribute about US$ 5 billion to this figure. Substances that are extracted from seaweeds – hydrocolloids – account for a large part of the remaining billion dollars, while smaller, miscellaneous uses, such as fertilizers and animal feed additives, make up the rest. The industry uses 7.5–8 million t of wet seaweed annually, harvested either from naturally growing (wild) seaweed or from cultivated (farmed) crops. The farming of seaweed has expanded rapidly as demand has outstripped the supply available from natural resources. Commercial harvesting occurs in about 35 countries, spread between the northern and southern hemispheres, in waters ranging from cold, through temperate, to tropical.’’ Only in the cases of species like Kappaphycus alvarezii or Gracilaria where large-scale propagation can be vegetative (so life cycles do not need to be manipulated in shore-based facilities) does aquaculture of macroalgae as sources of industrial raw materials seem justified (McHugh 2002). Even then, very low-wage economies need to be chosen. For example, socio-economic conditions appear right for K. alvarezii aquaculture in remote outer islands of the Fiji Group or Kiribati, but not in the main centers of these countries where there are a range of other competing livelihoods (Namudu and Pickering 2006). Even in China, aquaculture of macroalgae for alginates does not appear economical compared with

mechanical harvesting of natural resources (Jensen 1993). Emerging new uses for macroalgae are as pharmaceuticals and nutraceuticals. A pharmaceutical is a substance used in the treatment of disease, whereas a nutraceutical is a product isolated or purified from plants or other foods that is taken in a dosage (non-food) form in order to provide a physiological benefit and/or protection against disease. Also known as ‘‘dietary supplements’’ or ‘‘nutritional supplements’’, nutraceuticals represent a middle ground between pharmaceuticals (‘‘drugs’’) and so-called ‘‘functional foods’’. These are lower-volume but potentially high-value uses of macroalgae. Macroalgal species which are now known to contain either pharmaceutical or nutraceutical agents are very widely represented among the Chlorophyta, Phaeophyta and Rhodophyta. Just a few of the species listed in reviews by Baker (1984) or Smit (2004) include Saccharina japonica, Undaria pinnatifida, Gelidium species, Agardhiella tenera (J. Agardh) F. Schmitz, Gigartina skottsbergii Setchell et N.L. Gardner, Gracilaria species, Caulerpa species, Hypnea species, Hizikia fusiformis (Harvey) Okamura, Sargassum species, Kappaphycus alvarezii, and Codium fragile (Suringar) Hariot. This wide range of species (far wider than the range of popular edible species, and many of wide geographical range) plus the still-emerging state of research and actual products, makes it very difficult to discern the extent to which pharmaceutical or nutraceutical uses of macroalgae may become a driver toward further intentional introductions of macroalgae. Meanwhile, demand for the main food and food-constituent macroalgal species is large and continuing to increase. Of all the macroalgal biodiversity potentially available however, very few species are used commercially for these purposes, and even fewer are cultivated. Commercial demand is met by the handful of macroalgal species for which large markets already exist, and for which either natural stocks are abundant and accessible, or their agronomic traits lend them to large-scale aquaculture. It is a general rule that aquaculture (compared with wild fisheries) is based upon relatively few species. This scarcity of suitable species for aquaculture then results in trade and translocation. This occurs either when appropriate native species do not occur in a country, or when there is insufficient information about the culture requirements of native species to enable commercialization (Stickney 2002). The most likely target species for intentional introductions of macroalgae will, therefore, be either one of the select few high-value, aquacultured food species like Porphyra or Undaria pinnatifida, or easily propagated food-constituent species like Kappaphycus alvarezii. Such a small group of likely targets should greatly ease the research and risk-assessment demands posed by intentional introductions of macroalgae for aquaculture, compared with the host of candidate species that would need to be anticipated in the case of unintentional introductions. A second major driver for intentional introductions, particularly of the eucheumoids ‘‘cottonii’’ (Kappaphycus w19x

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species) and ‘‘spinosum’’ (Eucheuma denticulatum (N.L. Burman) F.S. Collins et Hervey), has been demand for new sustainable livelihoods for coastal communities in less-developed countries, as alternatives to unsustainable practices that cause degradation of tropical nearshore environments. Providing alternative livelihoods as a means to alleviate poverty and reduce pressure on wild marine resources is now a major strategy in global initiatives to protect coral reefs (e.g., the International Coral Reef Initiative). In many countries of Asia, Africa, the Caribbean Sea and the Pacific Ocean, cultivation of macroalgae is widely perceived as one of the most benign types of aquaculture activity in terms of environmental impacts, and one worth promoting to lessen exploitive pressure on fisheries/reefs and to lessen destructive practices like dynamite fishing (Zertuche-Gonzalez 1998, Ask 1999). For example, the environmental non-government organization (NGO) ‘‘The Nature Conservancy (TNC)’’ has worked with park authorities to establish a Kappaphycus seaweed farming industry in Komodo National Park in Indonesia (Howard 2003). The Swedish-funded CORDIO programme (Coral Reef Degradation in the Indian Ocean) has identified Kappaphycus aquaculture as one alternative livelihood suitable for some parts of the Indian Ocean region (Souter 2000, Howard 2003), as have Malaysian academics seeking to protect the coral reefs of Pulau Bangii in Sabah (Koh et al. 2002). ‘‘Cottonii’’ and ‘‘spinosum’’ cultivars have now been intentionally introduced to over 20 other countries both within and outside the native range of eucheumoids (Ask 2003, Ask et al. 2003a). A recent review of the history of Kappaphycus introductions in Pacific island countries (Sulu et al. 2004) lists eleven countries in the region where this species has been introduced at one time or another. There have been multiple introductions to some countries like Fiji, Kiribati and the Solomon Islands, either to introduce new cultivars like sacol, or as part of efforts to revive defunct Kappaphycus aquaculture projects in countries where self-sustaining populations did not establish. Kappaphycus farming has been promoted by governments and NGOs because it requires a low level of technology and investment, can be operated at the household level, has relatively little environmental impact, does not require refrigeration or high-tech post-harvest processing within the country, and is normally compatible with traditional fishing and other subsistence uses of the inshore environment. It is a potential source of income and employment in rural areas with few other incomegenerating opportunities, and in particular is an activity that can provide income for women (South and Pickering 2006). Alternative views about the claimed benefits and environmental impacts from Kappaphycus alvarezii introductions for aquaculture have been expressed, for example by Zemke-White (2004) and Zemke-White and Smith (2006). Nevertheless the prevailing general perception that eucheumoid aquaculture is a good alternative to other much-less sustainable coastal livelihoods in coral reef areas is likely to continue to be a driver for intentional w20x

introductions of the economically-important cultivars of eucheumoids, both within and outside their native range. Though strongest in tropical least-developed countries, similar interest in macroalgal aquaculture livelihoods also exists in the less-developed rural provinces of more-developed countries. In Ireland, two objectives were set for recent seaweed aquaculture development initiatives; one is to meet growing market demands for macroalgae, while the other is to create jobs in ‘‘peripheral communities in coastal areas’’ (Werner et al. 2004). Arguments for or against such development based upon ‘‘utility’’ in terms of contribution to national GDP by the new seaweed industry tend to under-value their potential economic impact. This is because, although benefits in dollar terms will certainly be small relative to GDP, even small increases in the income of marginalized or peripheral communities can make a big difference to peoples’ lives. For this reason, seaweed aquaculture projects in Asia, Africa and Pacific islands often enjoy political and community support out of proportion to the potential contribution to GDP.

The nature of commercial seaweed species: are they more or less risk-prone with respect to establishment and impact of feral populations in exotic locations? Species that readily establish and which threaten other species or habitats or ecosystems in exotic locations are termed ‘‘alien invasive species’’, however, the term can be subjective because it depends upon the investigator’s perspective of what constitutes ecological and/or economic harm. Not all alien species are ‘‘invasive’’, despite the recent proliferation of studies using these terms synonymously. Invasive species form high-density populations in which the species are structurally and/or functionally major components of the invaded communities (Trowbridge 2006). Trowbridge’s (2006) review of non-native macroalgae lists approximately 250 species reported to have been introduced to new geographic regions worldwide, but notes marked regional asymmetries including the fact that some regions are much better studied in this regard than others. Of these species, relatively few have become serious ecological pests, although several have become very widely distributed and abundant in their new habitats. A species of the red seaweed Asparagopsis introduced to the North Sea, probably from Australia (Elton 1958), and the introduction of Asian Undaria pinnatifida to France, New Zealand, Australia, Mexico and Argentina (Hay 1990) are well known examples. At least three macroalgal species are now regarded internationally as serious pests, but none of them are of interest for aquaculture or commercial harvesting: Asian Sargassum muticum (Yendo) Fensholt introduced to the Pacific seaboard of North America and to northern Europe; the introduction of a variety of Caulerpa taxifolia (M. Vahl) C. Agardh, possibly from northern Australia, to the Mediterranean Sea (Meinesz and Boudouresque 1996) and most recently to southern California; and the introduction of Codium fragile ssp. tomentosoides from

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Europe to the Atlantic seaboard of North America (Carlton and Scanlon 1985) and to Australia (http:// www.epa.vic.gov.au, Publication 67, April 1999). Characteristics and impacts of these species that render them ‘‘invasive pests’’ are reviewed by Zemke-White and Smith (2006). How many alien seaweed species proliferate and become aquatic nuisances? Trowbridge (2006) cites Boudouresque and Verlaque’s (2002) report that 8 of 84 macroalgal species introduced to the Mediterranean Sea are now ‘‘invasive’’; that of 113 species introduced to Europe as a whole, 27 are ‘‘invasive’’ and 9 are ‘‘very invasive’’ (Wallentinus 2002); and that of 19 species introduced to Hawaii, 5 have become ‘‘successful’’ (Smith et al. 2002). This is a relatively small yet significant proportion of the macroalgal species introduced. In making such comparisons, the potential for differences of interpretation about ‘‘invasive’’ must be borne in mind. So should the fact that a species can be ‘‘invasive’’ in some places but not in others (Trowbridge 2006). It does mean that the issue of macroalgal ‘‘invasiveness’’ must be taken seriously, along with other risks (including the unpredictability of risk), whenever intentional macroalgal introductions are contemplated. Two indicators of risk for commercially-cultivated macroalgae is whether or not any species important for aquaculture is prominent either in the literature on invasive macroalgae, or in any of the ‘‘watch lists’’ of invasive species maintained by government agencies, environmental NGOs, or inter-governmental organizations. Currently, Undaria pinnatifida, and Caulerpa taxifolia appear on the ICUN Global Invasive Species Database’s ‘‘100of-the-Worst’’ list (ICUN Invasive Species Specialist Group, http://www.issg.org). These species and Sargassum muticum and Codium fragile ssp. tomentosoides make up the four high-profile macroalgal ‘‘invaders’’ prominent in the scientific literature. According to Trowbridge (2006), public-awareness ‘‘watch lists’’ or publicity materials about marine invasives currently feature U. pinnatifida (in Europe, USA, Australia, New Zealand, Mexico), S. muticum (in Europe), Caulerpa taxifolia (in the Mediterranean Sea, California), and (in Hawaii) exotic Rhodophyta like Kappaphycus alvarezii, Acanthophora spicifera (M. Vahl) Børgesen, Eucheuma denticulatum, Gracilaria salicornia (C. Agardh) E.Y. Dawson, and Hypnea musciformis (Wulfen) J.V. Lamouroux. Of the above ten species there are two, U. pinnatifida and K. alvarezii that are important commercially for aquaculture. Environmental risks in culturing Undaria pinnatifida Undaria pinnatifida is one of the main commercially harvested and cultivated species in Asia, and has had its range extended by intentional introductions and translocations for aquaculture in China (Wu and Pang 2006) and to Atlantic France from Mediterranean France (Floc’h et al. 1991), but particularly by unintentional introductions to Europe, USA, Australia, New Zealand, Mexico and Argentina (see literature reviewed by Casas et al. 2004). The main vector for spread of U. pinnatifida to all of the new locations outside of Asia has been via shipping (as biofouling on hulls), with one exception, viz. the controversial case of an intentional translocation to establish a

new seaweed aquaculture industry in Brittany (Floc’h et al. 1991). U. pinnatifida has microscopic gametophyte stages, and is not known to propagate vegetatively. As described by Floc’h et al. (1991), translocation to Brittany first occurred in 1983 prior to consideration of risks by the ICES Working Group on Introductions and Transfers of Marine Organisms during 1984–1989. The decision was based upon the hypothesis that, while suitable for sporophyte growth, water temperatures of the North Atlantic Ocean would be too cold for the life cycle to be completed in nature, so that cultivation in coastal waters of laboratory-produced propagules would be ‘‘controllable’’. This hypothesis subsequently proved to be incorrect. In recent reviews, Schaffelke et al. (2006) and ZemkeWhite and Smith (2006) point out that Undaria pinnatifida is not considered invasive in some places (such as the Mediterranean Sea), nor is it always associated with declines in native species richness and abundance. However, it is considered invasive on the Atlantic coast of Europe and has been linked with reduction of native seaweed diversity in places like Argentina and New Zealand. Characteristics of U. pinnatifida that contribute to invasiveness are its rapid colonization of new or recentlydisturbed hard substrata, wide seasonal windows for propagule dispersal owing to the potential for multiple sporophyte generations within a population, and a propensity for colonizing floating objects (Hay 1990, Valentine and Johnson 2003, Valentine and Johnson 2004). These characteristics mean that U. pinnatifida is easily spread over long distances as biofouling on vessels or marine equipment. It has a microscopic gametophyte stage which, once established in the wild, is virtually impossible to eradicate. Its sporophytes can form large thalli (1–3 m) of a type that can be considered ‘‘invasive’’ in the sense of forming high-density populations that are structurally and/or functionally major components of invaded communities. Note, however, that in Tasmania U. pinnatifida is unable to displace native algae in the absence of disturbance and that formation of high density stands depends on disturbance to limit native canopy-forming algae (Valentine and Johnson 2003, 2004). Furthermore, ongoing persistence of U. pinnatifida at high densities on natural substrata requires persistent mechanisms to limit native canopy forming algae (Valentine and Johnson 2005a,b). There are no reports to suggest that U. pinnatifida is regarded as invasive within Asia in locations outside its natural range. One possible explanation is that the species here has commercial value that places it in high demand. Environmental risks in culturing Kappaphycus The eucheumoid Kappaphycus alvarezii, introduced intentionally to over 20 countries and territories including 11 in the oceanic Pacific (Eldredge 1994, Ask et al. 2003a, Sulu et al. 2004, Zemke-White 2004), has been publicized as ‘‘invasive’’ in one of those places. K. alvarezii was intentionally introduced to the Kaneohe Bay area of Oahu in Hawaii for aquaculture experiments in the 1970s, and Smith et al. (2002) report that, unlike the other non-indigenous macroalgae they surveyed, K. alvarezii is still restricted to this one locality; from the original w21x

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point of introduction it has spread at the rate of approximately 260 m year-1 (Rodgers and Cox 1999). However, K. alvarezii is considered ‘‘invasive’’ in Kaneohe Bay given its high abundance (up to 80% cover) on some patch reefs in this nutrient-enriched bay where it overgrows corals. For this reason it became the subject of a research program and public-awareness campaign to eradicate it and prevent further spread. The publicity generated about K. alvarezii at Kaneohe Bay has led to bans or restrictions upon introduction of this species for aquaculture in countries of the Caribbean Sea and Central America (Oliveira and Paula 2003, Alan T. Critchley personal communication). The situation at Kaneohe Bay in Hawaii contrasts with the mainly anecdotal evidence (some of which is reviewed by Ask et al. 2003a and Sulu et al. 2004) from other places where Kappaphycus alvarezii has been introduced for aquaculture. Many Pacific island countries attempted to cultivate Kappaphycus aquaculture during the 1980s and 1990s but did not move on to commercialization of this species (Luxton 2001) except in Kiribati, Fiji and, most recently, in the Solomon Islands (Pickering 2006). In some of these places, the abandonment of K. alvarezii aquaculture trials resulted in the thalli dying out, while in other places there are small residual populations such as in some parts of Tonga (Luxton 2001), Fiji Islands (Ask et al. 2003b) and Solomon Islands (Sulu et al. 2004). In these oligotrophic reef environments, K. alvarezii may be said to ‘‘persist’’ but it is not ‘‘proliferating’’ (Ask et al. 2003b, Sulu et al. 2004). For example, during a survey of commercial seaweeds in the Kaba and Kiuva area of Fiji in 1995 (where K. alvarezii had been farmed continuously from 1986 to 1993) only a single K. alvarezii thallus was found (Pickering and Mario 1999); it was a large (3 kg), unattached callus-like mass with few apical growing points and exhibiting damage consistent with grazing by siganid-fish and sea-urchins (e.g., Tripneustes gratilla Linnaeus). A farm still operates at Kiuva; this enterprise is destroyed occasionally by severe storm events; the farmer does not know of any residual K. alvarezii population in the area, so must move quickly to recover any drifting farmed plants for re-planting before they are washed away completely, or else must bring in new cuttings from farms in other parts of Fiji (Timothy D. Pickering personal observations). Currently, Fiji has two emerging problems with nuisance algae (May 2005, http://www.etravelblackboard. com/index.asp?ids38688&navs21). However, the species implicated are native to Fiji and the problems are most likely linked to anthropogenically driven nutrient loading of near-shore waters by sugar cane agriculture, tourism and piggeries (Lovell et al. 2004). In recent years Sargassum species have greatly increased in abundance on many coral reefs, for example along the tourismdeveloped Coral Coast of Viti Levu. Gracilaria edulis (S.G. Gmelin) P.C. Silva has also become abundant among sub-tidal seagrass beds adjacent to hotels like the Fijian Resort on the Coral Coast, and resorts like Beachcomber and Treasure Island in the Mamanuca Group (Timothy D. Pickering, personal observations). Beach-cast material is creating stench and requires removal and burial at the rate of 4 t (wet wt) per day at the two latter resorts alone. w22x

K. alvarezii has not been implicated in any of these problems with seaweeds in Fiji. The information presented by Smith et al. (2002) indicates that, of the five non-indigenous invasive species surveyed in their Hawaiian study, the entity they refer to as Kappaphycus alvarezii has the least invasive characteristics. Reproductive K. alvarezii plants are rare in nature, so the species propagates almost entirely by vegetative fragments. Smith et al. (2002) states that the ability to fragment readily, disperse widely before recruitment, and re-attach successfully, are all hallmarks of a highly invasive vegetatively propagating species. K. alvarezii, on the other hand, is a robust species, negatively buoyant and so unable to disperse long distances, and fragments do not have the capability of re-attaching to the substratum. K. alvarezii plants rely on either entanglement or their sheer weight to stay in place, though the fate of many plants is to become washed away into unfavorable habitats and perish. While K. alvarezii can slowly spread laterally from a point of introduction by fragmentation, unlike Undaria pinnatifida it has a very low probability of spreading as a component of biofouling communities on vessels. This suite of characteristics can account for its apparent inability to spread over long distances or between islands (Smith et al. 2002). There are additional factors in the Kaneohe Bay case that must be taken into account in any discussion of the risk of Kappaphycus alvarezii introductions for aquaculture. First, Kaneohe Bay is an urbanized and nutrientenriched environment, in response to which several macroalgal species including natives, for example Dictyosphaeria cavernosa (Forsska˚l) Børgesen (Smith et al. 2002), have become ecologically dominant. Second, the level of herbivory that might otherwise act as a biological control is low in Kaneohe Bay, owing to low numbers of sea urchins like Tripneustes gratilla (Zemke-White and Smith 2006) as a result of heavy fishing pressure, and to siganid fishes being completely absent from the Hawaiian islands (Woodland 1990). K. alvarezii is a palatable macroalga and siganid fishes are known to exert considerable grazing pressure on both farmed and free-living plants elsewhere (see, for example, Ask 1999, Luxton 2001). Thirdly, recent molecular (Zuccarello et al. 2006) and chemical evidence (the plants yielded iota, not kappa carrageenan; E. Ask, personal communication) now indicates that the K. alvarezii epithet has been misapplied in much of the literature about the Kaneohe Bay problem, because of the misidentification of Eucheuma denticulatum (another eucheumoid also introduced in the 1970s). According to Zemke-White and Smith (2006), the dominant invader in Kaneohe Bay is in fact E. denticulatum, and K. alvarezii is not nearly as abundant as previously thought. Further, Zuccarello et al. (2006) have found that K. alvarezii from Kaneohe Bay is a lineage genetically distinct from the K. alvarezii currently cultivated around the world. While the problems caused at Kaneohe Bay by proliferations of algae such as K. alvarezii are certainly not trivial, in terms of environmental risk, these took almost three decades to develop and are still at a local, rather than an international or even national scale, when compared with other high-profile invasive macroalgae like Sargassum muticum, Caulerpa taxifolia, Codium fragile or Undaria pinnatifida.

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Do commercial species of macroalgae pose enhanced environmental risks? Regarding the question ‘‘are commercial species more or less risk-prone’’ than other macrolagae, the answer is that two of the main commercially-cultivated species have gained a high profile for invasiveness, one well-justified and one less-justified. While Undaria pinnatifida is certainly used for aquaculture in Asia and in France, its dispersal and rapid establishment globally has been mainly unintentional due to shipping or to aquaculture of other organisms, owing to its propensity for spore settlement onto floating objects. Laminaria and Porphyra species have similar reproductive (and hence potentially bio-fouling) characteristics, and have been unintentionally introduced to places by the same vectors, but are not reported to be ‘‘invasive pests’’ in the sense of Trowbridge (2006). The characteristics of Kappaphycus alvarezii render it much less likely to become invasive, but this species is not completely without risk as is shown by the Kaneohe Bay experience. From the information currently available, it is possible to construct cases either in favor (for example, Zertuche-Gonzalez 1998, Oliveira and Paula 2003) or against (for example, Zemke-White and Smith 2006) introduction of K. alvarezii for aquaculture. This will continue to be controversial, and is destined to become a classic example of the differential application of the Precautionary Principle available to states under international law when competent authorities weigh up the benefits and risks of intentional exotic-species introductions. Points raised both for and against are based on arguments ranging from established empirical observation to hypothetical scenarios. Verification of the various claims made about socio-economic benefits (for example, Ask 1999) or environmental risks (for example, Zemke-White and Smith 2006) regarding K. alvarezii introductions for aquaculture will be fertile ground for further research. Advocates for and opponents against intentional Kappaphycus alvarezii introductions will agree, however, that anytime a species is introduced intentionally, it must have clear socio-economic benefits. There are now many cases of K. alvarezii having been introduced, often without quarantine protocols in place, for aquaculture projects that were under-financed, or commercially, socio-economically or institutionally unsound, or without guaranteed markets, and so these ventures were set up to fail from the outset (Ask 2003). By so doing, these states expose themselves to the full extent of potential environmental risk, without capturing any of the anticipated socio-economic benefits against which they had balanced this risk. The guidelines and advice provided by McHugh (2002) and Ask (2003) for anyone contemplating establishment of new eucheumoid cultivation projects are worthy of careful consideration. The record so far shows that intentional introductions of commercial macroalgae for aquaculture have caused few serious effects (Eucheuma denticulatum, not Kappaphycus alvarezii, at Kaneohe Bay has recently emerged as the main example). The bulk of problems caused by invasive macroalgae have arisen from unintentional intro-

ductions, either by shipping, or by ‘‘hitchhiking’’ with other organisms introduced for aquaculture (Oliveira and Paula 2003). But what of the future? What new species may be targeted for intentional introductions, and what species already introduced may result in unforeseen problems emerging over longer time-frames? Environmental risk associated with the introduction of any commercial macroalgae is, as for any exotic-species introduction, always uncertain. Schaffelke et al. (2006), Trowbridge (2006) and Valentine et al. (2007) review various factors that might assist to better predict which macroalgal species may become invasive and under what circumstances. These include possible connections between invasiveness and the latitude of introduction, species diversity of the receiving community (but see also Dunstan and Johnson 2007), nutrient status of receiving waters, degree of anthropogenic modification of environments, propagule pressure (for example, at nodes for inter-regional vectors such as ports), grazing pressure by herbivores, and the invading species’ environmental tolerances, growth and reproductive characteristics. No useful generalizations have emerged from consideration of these factors, apart from propagule pressure, suitability of the habitat, and previous success in other invasions (www.issg.org). These three factors appear useful to predict whether a macroalga introduced unintentionally will establish and survive. However, these factors are all ‘‘givens’’ in the case of an intentional introduction for aquaculture (after all, no one will want to introduce a species for aquaculture that is not going to survive) so they are not necessarily predictors of whether a commercial macroalga will become an ‘‘invasive pest’’ in the sense of causing negative environmental or socioeconomic impacts.

Risk management and risk mitigation of intentional introductions of commercially harvested exotic macroalgae Environmental, political, regulatory and moral imperatives Intentional introduction of species, whether for aquaculture, the aquarium trade, scientific research or other purposes, will always have an element of associated risk. Any introduction poses a potential threat to the ecosystem, and two important unknowns are the impact of the alien species on the recipient community, and its performance in the new ecosystem. Decisions about introduction of species rest with the governments of sovereign states and not with scientists. Competent authorities will take into account not only scientific advice but also social, economic and political considerations in assessing whether to accept or reject proposals for introduction. States are not unfettered in the exercise of their sovereign right to introduce species however, and must have regard to international norms and to possible responses from the international community or neighboring states. Policies and norms concerning introduction of species are found in a number of the international conventions, w23x

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guidelines and codes of practice that have stemmed from the Law of the Sea Convention and Chapter 17 of Agenda 21. While these are described more fully by Doelle et al. (2007), here we highlight the FAO Technical Guidelines for Responsible Fisheries 5: Aquaculture Development (FAO 1997) because these explicitly contemplate intentional introductions of commercial species. While conventions are binding upon those states that are signatories, the FAO guidelines are not legally binding but do carry moral force as an embodiment of international norms. The FAO Technical Guidelines principles on species introductions include: • Article 9.2.3 ‘‘States should consult with their neighboring states, as appropriate, before introducing nonindigenous species into transboundary aquatic ecosystems’’; and • Article 9.3.2 ‘‘States should cooperate in the elaboration, adoption and implementation of international codes of practice and procedures for introductions and transfers of aquatic organisms’’. It is implicit in these guidelines that states do not regard intentional introductions of species as completely prohibited. It is expected, however, that decisions whether or not to introduce should be made in accordance with codes of practice regarding risk assessments, safeguards and procedures to avoid environmental threats. This leaves states to grapple with application of a precautionary approach, given that any introduction will have risks against which prevention is the best defense. Doelle et al. (2007) note that prevention of unintentional introductions by placing a prohibition on some activities like shipping is simply not possible owing to its utility to society, while for other less-compelling activities like intentional introductions for aquaculture, decision makers would have to engage in the difficult task of weighing the risk against the utility of the proposed activity. In so doing, they are likely to also take into account the opportunity costs of not allowing the activity. Demand to introduce species for aquaculture can therefore be expected to continue, and the main issue for resource managers is to have in place mechanisms to assess risks and benefits and, if the decision is affirmative, strategies to manage the introduction in a way that maximizes the benefits and avoids, remedies or mitigates the risks. While international instruments like the FAO Guidelines can be criticized for being non-binding and for pitting ‘‘utility’’ against ‘‘prevention’’, the fact remains that the burden of proof and duty of care regarding environmental threats is now much higher than it was until a decade or so ago. States should be urged, persuaded, pressured and cajoled into adopting and implementing these guidelines. The FAO Guidelines hold up the ICES Code of Practice as a model for those states or regions that have not yet adopted any guidelines for species introductions (FAO 1997). This Code of Practice seeks to address (1) inadvertent coincident movement of harmful organisms (that is, non-target species), (2) ecological and environmental impacts of the target species, and (3) the genetic impact of alien species (ICES 1995). w24x

The annotations to Article 9.3.2 in FAO (1997) encapsulate the international norms on introduction of species as follows: ‘‘Basic elements of codes of practice such as ICES include: • A proposal to introduce a particular species in a particular area for a particular purpose, • An independent review of the proposal by competent authority, the review should include ecological and socio-economic risk assessments, and • Rejection, revision, or acceptance of the proposal. Once an introduction has been approved, governments should request aquaculturists to: • Create a fish health management program including quarantine and disease diagnosis, • Monitor and evaluate ecosystem and socio-economic effects, • Notify international organizations and neighboring States. A country’s ability to carry out the elements of the code will depend upon the state of knowledge on its human and aquatic communities and on the financial and human resources available. Faunistic and floristic surveys of local aquatic ecosystems can help determine what local species may be affected by aquaculture development and what local species may be utilized instead of importing an alien species. Socio-economic information on the fishing sector and on the fish-consumers will also help identify those people benefiting or at risk from aquaculture development. In addition, marketing surveys can help determine the cost-effectiveness and target consumer for a proposed introduction.’’ Managing intentional introductions in practice These guidelines provide a useful framework for states to adopt when contemplating intentional introductions for aquaculture, and many states should, as a first step, base their policies upon them. For marine macroalgae, however, risk mitigation and management of intentional introductions has often been poorly implemented (Zemke-White and Smith 2006), and this can be attributed to a variety of reasons. First, most of the intentional introductions of macroalgae predated the development of current norms on exotic species introductions. Second, financial and human resources in many countries (particularly less-developed countries) are insufficient to fully implement all the measures now deemed appropriate. These countries need assistance from their more developed counterparts or regional/international organizations to increase their capacity for compliance with international norms. For the same reason, there is a dearth of floristic surveys for some regions, and so it can be problematic to ascertain the status of a species as alien or otherwise and to assess the impact of an intentional introduction once it establishes as a feral population. A significant problem is that information on the biology of most marine taxa is relatively limited, and this increases scientific and public uncertainty about the outcomes of control actions (Thresher and Kuris 2004).

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Ribera and Boudouresque (1995) noted that it is relatively difficult for many regions of the world to recognize which species have been introduced, mainly because current species inventories are incomplete or recent, and because many authors are reluctant to specify whether certain species are introductions. For those regions that are relatively well-known, introductions comprise 4–5% (Mediterranean Sea) to 2–3% of the known flora (Ribera and Boudouresque 1995). Third, states in several regions lack the capacity for baseline ecological studies prior to introductions, or for follow-up monitoring (Schaffelke et al. 2006). According to Zemke-White and Smith (2006) there are no published reports that any baseline floristic studies have been taken prior to the introduction of macroalgae for aquaculture purposes, and there is little literature on effects that these introductions may have had on the recipient community’s structure or function. In the case of Kappaphycus alvarezii, the reasons given for this were usually either that funds were not available for the research, or funds were not allocated because K. alvarezii was not seen as a major environmental threat, so was not given high ranking among research priorities (Zemke-White and Smith 2006). Fourth, while impacts of the high-profile invasive species have engaged many researchers’ interest after they have become established, and numerous studies on these species are now completed, there has been very little research to support development of appropriate quarantine protocols for intentional or unintentional macroalgal introductions associated with aquaculture (Ask et al. 2003a, Sulu et al. 2004). This is an issue that the Aquaculture Programme of the Secretariat for the Pacific Community has begun to address on behalf of its member states (Sulu et al. 2004), but there is scope to improve recommended protocols further, and extend them to other target species. Fifth, most cases of post-introduction monitoring have been done over relatively short time-scales following an unintentional or intentional introduction (Trowbridge 2006), whereas surveys decades after introduction may yield different outcomes because the impact of an alien species may change over time (Schaffelke et al. 2006). This whole issue is difficult because, in many cases, the capacity to predict environmental effects is poor, reflecting the fact that invasion biology is a complex and relatively new area of study (Trowbridge 2006). Understanding one introduction does not necessarily enable prediction of other introductions of the same species, because of site- or time-specific factors and lack of ecological principles with general application (Schaffelke et al. 2006). Proposers of introductions relying upon containment measures must do so on the assumption that there will be escape and establishment in the wild (Stickney 2002). Further, eradications of macroalgae once established are very difficult and expensive, and have been successful in only very few cases where the response was rapid and the timeframe since introduction very short (for example, of Undaria pinnatifida from the hull of a sunken vessel at Chatham Is., New Zealand; Schaffelke et al. 2006). Thresher and Kuris (2004) have noted that eradication of marine invasive species in gen-

eral has a low probability of success once the invaders are well established. The only certain way to ensure there is absolutely no risk from a macroalgal introduction is not to introduce it, and instead develop suitable native species for aquaculture. For many countries, however, suitable native species will either be unavailable or have agronomic or marketing constraints. States balance ecological risk against socio-economic benefits, and no state has ruled out the possibility that at some time, in some place, there will be some species appropriate for introduction. While introductions are governed by international norms, a further problem, as pointed out by Ribera and Boudouresque (1995), is that texts in various conventions and agreements are generally couched as recommendations, and no authority is designated to assess risks ‘‘w«x which leaves the door wide open to whatever interpretation the country concerned considers (rightly or wrongly) to best suit its interests.’’ While states are now clearly expected to move in the direction of giving teeth to recommended practices and protocols, the fact remains that only a few countries (Australia, New Zealand, USA, Canada, Germany, Great Britain and Switzerland) have adopted implementing legislation relating to deliberate introductions of alien species. It is no coincidence that 51% of marine macrophytes introduced into Europe were first introduced in France where legislation controlling introductions of alien species is relatively lax. Progress is being made, however. In Europe alone, a total of 124 alien species (some of which are macroalgae) have been introduced in association with aquaculture activities (Zenetos and Strefaris 2004). These authors (op. cit.) note that the rate of introductions has been declining recently, suggesting that European Union measures are becoming more effective. The Berne Convention in 1979 provides that ‘‘w«x each contracting party undertakes w«x to strictly control the introduction of nonnative species’’. EU directives legislate for the protection of ecosystems against the adverse effects of aquaculture-related introduced organisms (for detail see Zenetos and Strefaris 2004). The Convention on Biological Diversity recommends a hierarchical approach, based primarily on the prevention of unwanted introductions. States in other regions need to be encouraged to adopt similar measures, and if necessary should be provided with assistance to develop capacity for their implementation. Identifying low- and high-risk species Schaffelke et al. (2006), and Trowbridge (2006) have reviewed the risks associated with invasive macroalgal introductions, and noted that there is no correlation between particular groups (Orders) and the levels of associated risk. There are few well documented examples of risk assessment for intentionally introduced marine macroalgae: well documented case histories, such as those for Kappaphycus alvarezii, Porphyra, Saccharina japonica and Macrocystis pyrifera (L.) C. Agardh illustrate the fact that the issue of risk came after, rather than before, the introductions. While environmental threats from intentional introductions of commercial macroalgae have been few compared with those arising from unintentional introduction by other vectors such as w25x

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shipping, this should be attributed more to the much greater propagule pressure posed by, and much greater difficulty in regulating, those other vectors. The historically lower threat posed by intentionally introduced species is not a result of their being intrinsically any more or less risky than unintentionally introduced species. The determination of low- vs. high-risk species remains largely arbitrary for marine macroalgae. In Table 1 we suggest some of the criteria potentially useful in the determination of low- vs. high-risk species. A number of intentionally introduced algae, however, do not match all criteria in either category, possessing intermediate characteristics of both low- and high-risk species. The introduction of Eucheuma/Kappaphycus from the Philippines into Hawaii, and later into the Pacific Islands, demonstrates the complexity of this issue (summarized in Sulu et al. 2004). While K. alvarezii (but now known to be Eucheuma) is regarded as a serious pest at one location in Hawaii (Zemke-White and Smith 2006), K. alvarezii has failed to establish in many localities in the Pacific Ocean where it was introduced (Eldredge 1994, Ask et al. 2003a, Sulu et al. 2004). Thus, site-specific information is needed since the same species may behave quite differently in different environments. Risk assessment procedures have resulted in some proposals for introduction of commercially valuable species to be declined. For example, plans to introduce the giant kelp Macrocystis pyrifera into Europe in the 1950s and 1970s were dropped because of public protests and a recommendation from ICES (Boalch 1981). It is possible that this species could otherwise have colonized the European Atlantic coast from Spain to Norway, with unpredictable consequences. While it is possible to identify macroalgal species that pose higher or lower levels of risk in terms of undesirable environmental or socio-economic impacts, for example by utilizing criteria like those in Table 1 or like those reviewed by Schaffelke et al. (2006) and Trowbridge (2006), no introduction of commercial macroalgae will be completely without risk and it may be difficult to predict what the actual consequences will be. Quarantine and farm procedures to minimize risk Quarantine procedures address the issue of inadvertent coincident movement of harmful organisms, that is, unintentional introduction of non-target species (sometimes termed ‘‘hitchhikers’’). For example, Sulu et al. (2004) showed that Kappaphycus alvarezii shipped to new loca-

tions may contain fauna such as copepods, amphipods, isopods or polychaete worms, smaller seaweeds attached as epiphytes, a microscopic epibiota possibly including harmful dinoflagellates, and potentially any disease organisms within the seaweed tissues. They also showed, however, that relatively simple propagule selection, treatment (e.g., washing with seawater) and quarantine procedures can greatly reduce this risk. Quarantine can thus play a vital role in reducing the risk of introductions of unwanted species and pathogens associated with host aquaculture species. States are urged to develop quarantine protocols for the species of macroalgae most likely to be proposed for introduction. Unfortunately, very little research has been directed at development of quarantine protocols for macroalgae, and hardly any protocols or guides to best practice are available for macroalgae. Those states that have paid attention to quarantine matters have usually resorted to ad hoc approaches. A detailed case-history of quarantine for Kappaphycus alvarezii aquaculture and transportation to various countries in the Pacific Islands region is provided by Sulu et al. (2004). While there are a small number of examples where quarantine measures have been applied to K. alvarezii (Ask et al. 2003a, Sulu et al. 2004), it is clear that in the majority of cases no quarantine protocols have been applied. However, a modicum of forethought to the application of simple and low-technology protocols can go a long way toward reducing ecological risk. For seaweeds like Kappaphycs alvarezii with a relatively simple growth habit, quarantine procedures such as washing, physical surface cleaning, mild chemical treatment or washing in freshwater may be effective in the removal of larger fauna and flora, although the complete removal of micro-organisms or endophytes is virtually impossible (Sulu et al. 2004). An alternative is to translocate just apical tips of the plant and then grow these at the new location using tissue culture methods. This was the procedure followed by Brazilian authorities when they introduced K. alvarezii. They imported axenic cultures, cultured apical tips in agar, grew plants, then grew a second generation of cultures from those plants before releasing K. alvarezii into the sea. The process took about four years and required fairly sophisticated laboratory facilities (Oliveira and Paula 2003). For quarantine of whole plants of Kappaphycs alvarezii, placement in holding tanks using closed water systems for an observation period can allow detection of ‘‘hitch-

Table 1 Criteria to determine the potential risk factors of deliberately introduced macroalgae. Low-risk

High-risk

• • • •

• • • •

Does not cause serious impacts to the environment Reproduction and life-cycle non aggressive Life-cycle stages can be controlled or manipulated Not known to out-compete local species when released

• Relatively easy to remove/discard if no longer desired • Lacks mechanisms or strategies (e.g., vegetative propagation) aiding its dispersal • Causes minimal economic and health costs • Has natural predators to help keep it in check

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• • • •

Causes serious impacts to the environment Aggressive reproduction and life-cycle Complex life-cycle that is difficult to control or manipulate Grows rapidly in a new environment and displaces local species in the wild Out-competes local species for space and habitat Possesses mechanisms aiding dispersal such as vegetative propagation Potential high cost of eradication Lack of natural predators in its new environment

T.D. Pickering et al.: Intentional introductions of commercially harvested alien seaweeds 347

hikers’’ or disease on the imported plants, and this can be done in relatively simple land-based tank systems with daily water exchange provided by pumping seawater from an adjacent shore, as was done in Fiji in 1984 (Sam Mario, personal communication). Water exchanged from this kind of quarantine system should be chlorinated and discarded at least 500 m from the coastline to ensure that no aquatic organisms escape into the local ecosystem. Zemke-White (2004) recommends that plants should be maintained for at least two weeks in a contained system, and that during that time they should be checked for the growth of microalgae and animals on the thalli. The health of K. alvarezii plants can greatly deteriorate in enclosed systems over two weeks, however. This observation time could be safely shortened to 7–10 days if washing and/or chemical-dip treatments precede the observation period (Sulu et al. 2004). Ribera and Boudouresque (1995) stress the importance of raising public awareness of the risks involved in species introductions. Good examples of public awareness initiatives are the brochures distributed to people arriving in Australia explaining the main Australian quarantine regulations, published in 49 languages, and the massive public awareness campaigns carried out in Spain, Italy and France on the invasion of Caulerpa taxifolia (see Meinesz 2007). The Aquaculture Programme of the Secretariat for the Pacific Community is now raising awareness among member countries about the need for quarantine protocols in aquaculture (www.spc.int/aquaculture), and is taking steps to develop guidelines for best practice. Once a species has been introduced and the quarantine period has ended, particular farm procedures may sometimes be applied to further minimize environmental risk. These are usually directed at containment of the introduced macroalga, to confine it to the aquaculture site and prevent it from becoming established in the wild as self-sustaining populations or spreading to other areas. One practice has been to propagate the target macroalga under controlled conditions onshore, then carefully select grow-out sites where environmental conditions are suitable for growth but not for reproduction (usually by virtue of unsuitable temperature and/or daylength regimes). This approach was not successful in containing Undaria pinnatifida on the French Atlantic coast, because estimates of the temperature limits for reproduction of this species were incorrect. Proposers and resource managers would be prudent to assume that any marine species, once introduced, will not be successfully contained, and ought to address all issues of environmental risk at the pre-introduction risk assessment stage of any proposal. Risk of spreading the introduced macroalga to other areas away from the farm can be reduced by taking care that seaweed farm equipment (ropes, floats etc.) or vessels do not harbor the macroalga through entanglement or biofouling before they are taken to other places. There is a growing literature on ways to reduce the risk of unintentional translocation of Undaria pinnatifida, Caulerpa taxifolia and Sargassum muticum. For example, Gunthorpe et al. (2001) report on tests of protocols to avoid spreading U. pinnatifida via mussel-farming equipment in

Australia, and the resultant guidelines are being built in to publicity materials and awareness campaigns. One component of environmental risk from introduction of a macroalgal species for aquaculture is the risk of adverse effects from the farm activity itself. Usually these can be avoided, remedied or mitigated by appropriate farm management practices. For example, Kappaphycus alvarezii plants in off-bottom culture can abrade and damage living corals that they contact (Zemke-White and Smith 2006). This can be avoided through selection of farm sites on clean-bottomed sandy areas away from either corals or seagrasses (Ask 1999), which brings the twin benefits of avoiding damage to corals, and reducing herbivory of the cultivated macroalga by fishes which are often much more abundant around corals and in seagrass beds. Conflicts between risk mitigation and commercial production The intentional introduction of organisms for the purpose of aquaculture requires a balance between minimizing ecological impact (i.e., risk), and maximizing economic gain (Watson et al. 2000). The rapid depletion of natural fish stocks through over-fishing in most parts of the world is placing increasing pressure on governments to seek alternative means of production, and as a consequence aquaculture is increasing (FAO 2004). Depletion of natural stocks of seaweeds is a much less significant issue because many stocks are inaccessible to fishing (e.g., because they are located on exposed rocky coasts with high wave action). However, this very inaccessibility can be one driver for increased production through aquaculture in order to meet demand. The majority of marine plant aquaculture (and thus the movement of organisms from country to country) is centered in SE Asia. Seed stock from this region has been used to establish aquaculture operations elsewhere, with the majority of the recipients being developing countries. For most of these developing countries, economic potential has almost always taken precedence over ecological concerns. The majority of these countries lack legislation regarding introductions or it is sketchy, they lack the capacity to conduct meaningful baseline surveys or monitoring, or they perceive aquaculture of an alien macroalga as preferable to alternatives such as continued unsustainable livelihoods or destructive practices associated with coral reefs. In these countries, where there are real or potential conflicts between risk mitigation and commercial production, the ‘‘utility’’ of commercial production often takes precedence. In other countries with greater capacity for coastal-management and weaker economic imperatives (for example, Australia, New Zealand, USA), further intentional introductions of non-indigenous macroalgae are now almost impossible. As mentioned earlier, in the example of the proposed introduction of Macrocystis pyrifera to Europe, deliberate introductions can be prevented in the face of public pressure and scientific advice, and potentially serious ecological impacts can be averted. These will always be political rather than purely scientific decisions. However, expectations are increasing that states should exercise a much greater w27x

348 T.D. Pickering et al.: Intentional introductions of commercially harvested alien seaweeds

duty of care and impose a greater burden of proof upon proposers of introductions than was the case twenty or even ten years ago. The aquaculture sector is a relatively easy target for regulators compared with the shipping sector (Hewitt et al. 2004) which continues to be the main vector for macroalgal introductions owing to technical constraints and powerful lobbyists. For the majority of deliberate seaweed introductions, the risks have proved to be relatively small and conflicts between risk mitigation and commercial production have not been an issue. Monitoring should be an important part of risk mitigation however, such as that being carried out with introductions of Porphyra yezoensis Ueda from Japan to the Atlantic coast of the USA (Watson et al. 2000). In cases where commercial production does not take precedence, the conflict becomes one where communities may have an opportunity cost imposed upon them either by an appropriate introduction being declined, or where permission for the introduction is granted but with onerous conditions attached. A hypothetical example might be if a jurisdiction were to require four years of high-technology tissue culture before Kappaphycus alvarezii propagules can be released from quarantine, as occurred in Brazil (Oliveira and Paula 2003), in order to completely eliminate the risk of introducing associated microbiota, compared with a two-week low-technology quarantine protocol (Sulu et al. 2004) that might perhaps reduce the risk by 99%. Conflicts about macroalgal introductions, where they exist, can best be solved through transparent and accountable decision-making processes by competent authorities in which there is wide and meaningful consultation and participation by relevant stakeholders. Decision support tools and other formal risk management frameworks Governments are cautious with aquaculture enterprises involving the introduction of exotic species into a new environment, due to potential adverse impacts on native flora and fauna (see Doelle et al. 2007). In addition, members of the World Trade Organization (WTO) are required to use formal risk analysis to justify any restrictions on international trade based on risks to human, animal or plant health (WTO 1994). The risk analysis process takes into account risk management and appropriate actions necessary to mitigate the risks’ impacts. Orr (2003) describes a detailed risk analysis review process for non-indigenous aquatic organisms, developed by the US Aquatic Nuisance Species Task Force in 1996. The review process provides a standardized means for evaluating the risk of introducing non-indigenous organisms into a new environment and, if necessary, determining the correct risk management steps needed to mitigate that risk. However, the process has some uncertainties about methodology, human error or the biology and environmental requirements of the organisms, and it should be seen as a constantly evolving system that takes into account new information derived from research. Orr (2003) provides a copy of the Organism Risk Assessment Form (including uncertainty reference w28x

codes) which is the main decision support tool of that assessment. A risk management framework should result from a risk analysis process. The US system provides the necessary guidelines and tools for the development of a risk management framework for marine macroalgae, although the complex infrastructure and taxonomic expertise required for the task is lacking in most countries where seaweeds are brought in for aquaculture purposes. Even (or perhaps, especially) under circumstances where there is lack of knowledge, limited capacity and low funding, the benefits in developing a risk management framework include identification of the core values that need to be protected, and prioritization in the delivery of control mechanisms and in contracting of scientific research (Hewitt et al. 2004). A risk management framework is a useful tool to identify gaps in knowledge or shortcomings in capacity, and to highlight the most pressing issues.

Acknowledgements The authors are indebted to Alan T. Critchley, Erik I. Ask, G. Robin South, Cameron Hay, A.R.O. Chapman and Craig Johnson for providing comments on this manuscript, and to Shital Swarup for assistance with literature searches and interloans.

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Valentine, J.P., R.H. Magierowski and C.R. Johnson. 2007. Mechanisms of invasion: establishment, spread and persistence of introduced seaweed populations. Bot. Mar. 50: 351–360. Wallentinus, I. 2002. Introduced marine algae and vascular plants in European aquatic environments. In: (E. Leppakoski, S. Gollasch and S. Olenin, eds) Invasive aquatic species of Europe. Distribution, impacts, and management. Kluwer Academic Publishers, Amsterdam. pp. 27–52. Watson, K., D. Cheney and I. Levine. 2000. Biomonitoring of an aquacultured introduced seaweed, Porphyra yezoensis (Rhodophyta, Bangiophycidae) in Cobscook Bay, Maine, USA. In: (J. Pederson) Marine bioinvasions. Proceedings of a Conference, January 24–27, 1999. MIT, Cambridge, MA. pp. 260–264. Werner, A., D. Clarke and S. Kraan, eds. 2004. Strategic review of the feasibility of seaweed aquaculture in Ireland. National Development Plan Marine RTDI Desk Study Series DK/01/ 008. Marine Institute, Galway. pp. 120. Woodland, D.J. 1990. Revision of the fish family Siganidae with descriptions of two new species and comments on distribution and biology. Indo-Pacific Fishes series No. 19. Bishop Museum, Honolulu. pp.136. WTO. 1994. Agreement on the Application of Sanitary and Phytosanitary Measures. In: The results of the Uruguay Round of multilateral trade negotiations: the legal texts. General Agreement on Tariffs and Trade (GATT), World Trade Organization, Geneva. pp. 69–84. Wu, C.Y. and S.J. Pang. 2006. The seaweed resources of China. In: (A.T. Critchley, M. Ohno and D. Largo, eds) Seaweed resources. Expert Centre for Taxonomic Identification (ETI), University of Amsterdam (CD-ROM series).

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Zemke-White, W.L. (2004). Assessment of the current knowledge on the environmental impacts of seaweed farming in the tropics. Marine science into the new millenium: new perspectives and challenges. Proceedings of the Asia-Pacific Marine Science and Technology Conference, 12–16 May 2002, Kuala Lumpur, Malaysia: 465–476. Zemke-White, W.L. and J.E. Smith. 2006. Environmental impacts of seaweed farming in the tropics. In: (A.T. Critchley, M. Ohno and D. Largo, eds) Seaweed resources. Expert Centre for Taxonomic Identification (ETI), Univ. Amsterdam (CD-ROM series). Zenetos, A. and N. Strefaris. 2004. Introduced species in marine and coastal waters. Aquaculture as a mode of introduction. Indicator Fact Sheet. pp. 1–9. http://themes.eea.europa.eu/ Specific_media/water/indicators/WEC07b,2004.05/ WEC7b_IntroducedSpecies_140504.pdf. Zertuche-Gonzalez, J.A. 1998. Macroalgal cultures as a sustainable coastal livelihood in coral reef areas. In: (M.E. Hatziolos, A.J. Jooten and M. Feodar, eds) Coral reefs: challenges and opportunities for sustainable management. Proceedings of an associated event of the fifth annual World Bank Conference on Environmentally and Socially Sustainable Development. October 9–11, 1997. The World Bank, Washington DC. pp. 53–54. Zuccarello, G.C., A.T. Critchley, J.E. Smith, V. Sieber, G.B. Lhonneur and J.A. West. 2006. Systematics and genetic variation in commercial Kappaphycus and Eucheuma (Solieriaceae, Rhodophyta). J. Appl. Phyc. 18: 643–651.

Received 31 January, 2006; accepted 17 April, 2007

Botanica Marina 50 (2007): 351–360

2007 by Walter de Gruyter • Berlin • New York. DOI 10.1515/BOT.2007.040

Review

Mechanisms of invasion: establishment, spread and persistence of introduced seaweed populations Joseph P. Valentine1,*, Regina H. Magierowski2 and Craig R. Johnson2 Marine Research Laboratories, Tasmanian Aquaculture and Fisheries Institute, University of Tasmania, GPO Box 252-49, Hobart, Tasmania 7001, Australia, e-mail: [email protected] 2 School of Zoology and Tasmanian Aquaculture and Fisheries Institute, University of Tasmania, GPO Box 252-05, Hobart, Tasmania, 7001 Australia

Keywords: establishment; invasion process; life history characteristics; management; persistence; spread.

1

* Corresponding author

Abstract Understanding the mechanisms that facilitate or inhibit invasion of exotic seaweeds is crucial in assessing the threat posed by their incursion and to define control options. In this paper, we consider how life history characteristics of the invading species and properties of the recipient environment influence the likelihood of invasion, giving particular emphasis to how disturbance influences the establishment, spread and persistence of introduced seaweed populations. Very few commonalities in key life history traits emerge since each species possesses a unique set of traits that confers a high capacity for invasiveness. Consequently, for seaweeds at least, predictions of invasibility based on life history characters alone are unlikely to be useful. In contrast, it is clear that disturbance is an important process in the establishment of these invasive species. With the possible exception of Caulerpa taxifolia, disturbance appears to be a critical factor that is either a key requirement (e.g., Codium fragile ssp. tomentosoides, Sargassum muticum and Undaria pinnatifida), or which accelerates (e.g., Fucus serratus) establishment and spread. The role of disturbance in the persistence of the invaders is more complex and depends on the species concerned. In several cases there is substantial evidence for positive feedback mechanisms that enable introduced species to persist in the absence of the disturbance factor that facilitated establishment in the first place. These circumstances define examples of ecological hystereses that pose particular challenges for management and control. The evidence suggests that, in several cases, preventing anthropogenically mediated disturbance to canopies of native seaweeds should be considered as a potential control option to minimise the risk of establishment of exotic species at high densities. However, for these kinds of introduced species, once they are established, control options that primarily target the disturbance are unlikely to represent viable management options.

Introduction Seaweed invasions are increasingly recognised as a major problem worldwide with often dramatic effects on ecosystem structure and function (Walker and Kendrick 1998, Piazzi and Cinelli 2000). The rapid acceleration in spread of introduced seaweeds poses a major challenge for management of marine ecosystems. Informed and responsible management should attempt to identify potential ‘‘next pests’’ and mitigate against their arrival and establishment at new locations (Hayes and Sliwa 2003, Nyberg and Wallentinus 2005, Hewitt et al. 2007). Once incursions do manifest, if presented with a number of introduced seaweeds, managers must decide which species have immediate priority for control, which to attempt to control if time and finances are available, and which to leave alone (Hiebert 1997). In this context, knowledge of the threat posed by an introduced species is essential to effectively prioritise species for management purposes (Byers et al. 2002). Understanding the mechanisms of invasion is an important element defining both threat and control options for introduced seaweeds. If a species is capable of displacing native algae in the absence of any primary mechanism of facilitation such as disturbance, then it represents a major threat to the integrity of native algal communities. For these species that are able to ‘‘drive’’ the community dynamic de novo, management may need to target the alga directly. Alternatively, if a species ‘‘tracks’’ community dynamics and requires disturbance or another facilitating mechanism to establish and persist, then there exists a greater range of management options that include targeting the cause of the disturbance rather than the alga itself. In addition to defining threat and options for management, understanding mechanisms of invasion may also be important in identifying both potential future invaders and potential recipient communities vulnerable to invasion (Mack et al. 2000, Dunstan and Johnson 2007). Most studies of seaweed invasions have been ‘‘case studies’’ that simply document the occurrence and spread of the introduced species. Relatively few studies have examined the mechanisms that underpin successful seaweed invasion, which is a complex multi-stage process that can be arbitrarily, but usefully broken down into the separate phases of arrival, establishment, spread and persistence (Mollison 1986). Defining these key phases precisely is vital to adequately describe and understand the invasion process and to enable meaningful comparisons among different species. We define establishment w31x

352 J.P. Valentine et al.: Mechanisms of seaweed invasion

of an introduced seaweed species as the development of a population of macroscopic thalli in an area that did not previously support the species. Spread involves dispersal (natural or human-assisted), establishment and subsequent expansion of the species range, while persistence (sensu Johnson and Mann 1988) refers to turnover of more than one generation of macroscopic thalli. For seaweed species with an alternation of generations between microscopic and macroscopic forms, the presence of microscopic stages, in the absence of the macroscopic form, is not considered to constitute either establishment or persistence. We attempt to provide a synthesis of current levels of understanding of the mechanisms facilitating successful seaweed invasions, focusing on life history characters and the factors influencing the processes of establishment, spread and persistence (factors influencing arrival are addressed elsewhere in this issue; see Hewitt et al. 2007). Where possible, particular emphasis has been placed on understanding mechanisms of invasion using evidence from experimental studies; however, correlative studies have also been included in the absence of experimental data. This article focuses on five species that have been recognised as successful seaweed invaders

which have been the subject of experimental studies examining invasion mechanisms, viz. Caulerpa taxifolia (Vahl) C. Agardh, Codium fragile ssp. tomentosoides (van Goor) P.C. Silva, Fucus serratus Linnaeus, Sargassum muticum (Yendo) Fensholt and Undaria pinnatifida (Harvey) Suringar. The significant research efforts that have been directed at understanding the invasions of these particular species provides sufficient basis for discussion of mechanisms that promote successful seaweed invasion.

The invasion process How important are life history characteristics in determining invasibility? Are there commonalities in the life history traits of invasive seaweed species that provide important insights into mechanisms of invasion? In the case of terrestrial plants, the attributes of invading species that render them ‘‘invasive’’ have frequently been investigated as a means of identifying future invaders (Mack et al. 2000). Is a similar approach likely to be fruitful in the case of seaweeds? Examining life history traits of the few well studied inva-

Table 1 Life history traits of selected invasive seaweeds in relation to the three primary strategies of plant evolution (adapted from Clayton 1990). Algal type*

Algal species

Opportunistic Stresstolerant

Biotically Caulerpa competent taxifolia

Fucus serratus

Codium fragile Sargassum Undaria ssp. tomentosoides muticum pinnatifida

Life span

Short

Long

Long

Long1 (pseudoperennial)

Long8

Long11,12,13 (pseudoperennial)

Long20 (pseudoperennial)

Herbivore resistance

Low

High

High

High2

High9,c

Moderate14

Moderate20 Low27

Growth rate

High

Low

High

High3

Low8

High8,15

High21

Reproductive episodes

Single

Repeated Repeated

Repeated4 Repeated10 Repeated11,12,16 (vegetative (zygotes) (parthenogenesis reproduction only) only)

Capacity for vegetative reproduction

Varies

Varies

Varies

Yes5,6,7

No8

Propagule number

High

Varies

Varies

n/aa

Moderate9,c High17

Moderate24 High26

Propagule size Small

Varies

Varies

n/aa

Large10

Small18,b

Large25

Small9

Dispersal shadow

Varies

Varies

Small and large3,4

Small10

Small and large16,19

Small and large22,24

Small and large29

Large

Yes8,12,16

Short26 (annual)

High28

Repeated22 Single26 (zygotes) (spores) No23

No28

While recognizing that qualitative assessment of life history strategies in this way is subjective, the tabulation nonetheless clearly shows that invasive seaweeds manifest a wide range of life history attributes, and that classification by life history type, qualitative or otherwise, does not reveal consistent patterns. *Definitions of ‘‘algal type’’ in accordance with Grime (1977). a Not applicable, since invasive Caulerpa taxifolia reproduces by vegetative fragmentation. b Although larger than other Codium species. c Inferred from studies of other species within the order Fucales. References: 1Meinesz et al. 1995, 2Boudouresque et al. 1996, 3Meinesz et al. 1993, 4Smith and Walters 1999, 5Sant et al. 1996, 6ˇ Zuljevic and Antolic 2000, 7NIMPIS 2002a, 8Chapman et al. 2002, 9Clayton 1990, 10Arrontes 2002, 11NIMPIS 2002b, 12Fralick and Mathieson 1972, 13Trowbridge 1999, 14Trowbridge 1998, 15Dromgoole 1975, 16Carlton and Scanlon 1985, 17Chapman 1999, 18Prince and Trowbridge 2004, 19Campbell 1999, 20Britton-Simmons 2004, 21Norton 1977, 22Fletcher and Fletcher 1975, 23Stæhr et al. 2000, 24 Deysher and Norton 1982, 25Norton 1992, 26Schaffelke et al. 2005, 27Valentine and Johnson 2005a, 28Saito 1975, 29Forrest et al. 2000.

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J.P. Valentine et al.: Mechanisms of seaweed invasion 353

sive seaweeds in relation to the three primary strategies associated with plant evolution, namely opportunistic, stress tolerant, and competitive or biologically competent strategies (Grime 1977, Vermeij 1978), provides a basis for detailed examination of traits that may contribute to invasion success (Table 1). While we acknowledge the limitations of comparing life history characteristics on the basis of subjective qualitative criteria, it is nonetheless apparent that few consistent patterns emerge. Invasive seaweed species clearly do not emerge with a common suite of life history characteristics, and most exhibit a range of characteristics that include components of what is usually recognised as opportunistic, stress-tolerant and biotically competent strategies. Undaria pinnatifida is arguably the exception, possessing life history traits entirely consistent with an opportunistic life history strategy. The most consistent trait shared by most species, with the notable exception of Fucus serratus, is that these invasive species have rapid growth rates and capacity for both short- and longdistance dispersal. Our overall conclusion is that, for the most part, there is limited value in using the life history features of these seaweeds to predict traits that confer invasiveness, highlighting the fact that each species possesses a unique suite of characters (Table 1; Clayton 1990). This is consistent with a view emerging from detailed examination of invasive terrestrial plants, which is that while invaders have some traits in common, exceptions abound and generalisations are only applicable to a relatively select group of invasive species (Mack et al. 2000). This notion is contrary to the more conventional viewpoint that most invasive plants, at least in terrestrial environments, are opportunistic species (Kolar and Lodge 2001, Sakai et al. 2001). It is necessary to reconcile this view with the results of a recent approach to predict macroalgal introductions by quantitative ranking of 13 seaweed species’ traits under the broad categories of dispersal, establishment and ecological impact (Nyberg and Wallentinus 2005). Using this approach, introduced species consistently ranked higher than native species, and invasive species ranked higher than non-invasive species. However, an important factor contributing to high rankings for invasive species was their ‘‘ecological impact’’, i.e., their overall impact on the recipient community through development of dense cover. When the life history characteristics that contribute to this ecological impact were considered alone, there were no consistent patterns. These results are, therefore, consistent with the tenet that invasiveness of seaweed species is poorly predicted from their life history characteristics alone. Community vulnerability to invasion The properties of the recipient community can be of major importance in determining invasion success (Dunstan and Johnson 2007). The presence of vacant, under-, or un-utilised niches has been proposed as a reason why some communities are particularly vulnerable to invasion (Mack et al. 2000). However, it is difficult to define a vacant niche until it is occupied (Trowbridge 1998), and so in this sense arguments centred on the concept of niche invoke a circular logic. It is more useful

to relate invasibility to the availability of resources. Emerging evidence suggests that increasing invasibility may be universally associated with increases in resource availability and/or variance in resource availability (reviewed by Dunstan and Johnson 2007). Certainly, there is evidence of this for some invasive seaweeds. For example, in New Zealand, Codium fragile ssp. tomentosoides is predominately found in association with Corallina turfs where bare space for attachment of thalli is abundant (Trowbridge 1995). Similarly, Undaria pinnatifida sporophytes can grow at high densities on otherwise unstable cobble and shell substrata (Valentine and Johnson 2003) on which foliose native macroalgae rarely develop. U. pinnatifida is able to utilise these unstable substrata because of the high growth rates of the sporophytes which rapidly attain sufficient size to stabilise the loose material to which they are attached. Native species coexisting with U. pinnatifida in Tasmania are unable to colonise unstable substrata because they grow too slowly and, therefore, are subject to a higher risk of damage by movement of the substrata during heavy surge or wave action. Similarly, for Sargassum muticum, there is evidence to indicate that it establishes in areas that usually support low cover of native algae (Critchley 1983, Fernandez et al. 1990). More recent work with Sargassum muticum examined the mechanisms underlying invasion resistance in greater detail (Britton-Simmons 2006). Experimental manipulations showed that increased availability of space with removal of crustose and turf algae is required for strong recruitment of S. muticum, but that recruits did not survive unless the foliose understorey and canopy species were reduced to increase light levels. This work suggests that resistance to invasion by S. muticum is realised by sequential pre-emption of key resources (space and light) by different functional groups at different stages of development of the invading alga. Other attributes of potential recipient communities that may influence vulnerability to invasion include properties that create a high likelihood of escape from biotic constraints (Mack et al. 2000). Thus, it is sometimes suggested that communities with low species richness tend to be invaded more readily than areas with high species richness (Trowbridge 1998, Levine and D’Antonia 1999, Shea and Chesson 2002). It is usually implied that reduced levels of interspecific competition are associated with lower diversity systems, which makes invasion by other species more likely. However, work with marine invertebrate systems show that in some cases invasion of patches significantly increases with diversity (Dunstan and Johnson 2004). These authors argue that resistance to invasion will be determined by the particular properties of component species and their community dynamic, and not by an aggregate community property such as richness. This dynamic is readily interpreted in terms of the availability of resources (in this case the resource is space) and variability in resource availability (see Dunstan and Johnson 2007). An ecosystem level attribute frequently cited as an important factor determining invasion success in terrestrial systems is disturbance (Elton 1958, Cavers and Harper 1967, Crawley 1986, 1987, Hobbs and Atkins w33x

354 J.P. Valentine et al.: Mechanisms of seaweed invasion

1988), which acts to free resources. While there is ample evidence that disturbance often plays a critical role in facilitating invasion of introduced plants in terrestrial systems, its importance in facilitating the invasion of exotic seaweeds has not been assessed in detail. Mechanisms of establishment: the importance of disturbance If the few studies that have critically examined the role of disturbance in the invasion process for introduced seaweeds are broadly indicative, then it is clear that disturbance can also play an important role in facilitating establishment and spread of marine plants (Table 2). Disruption of the native algal canopy has been demonstrated to be a key factor facilitating establishment of introduced seaweeds, particularly for Sargassum muticum, Undaria pinnatifida and Codium fragile ssp. tomentosoides. Experiments involving manipulation of native seaweed canopies (Deysher and Norton 1982, De Wreede 1983, Andrew and Viejo 1998, Britton-Simmons 2006) and observations following natural canopy disappearance (Ambrose and Nelson 1982) indicate that S. muticum requires the provision of free space to establish successfully. Stable native algal canopies inhibit invasion, and there have been several mechanisms proposed. Manipulations by Britton-Simmons (2006) suggest that an intact canopy prevents establishment of S. muticum by maintaining low light levels on the benthos, while Deysher and Norton (1982) suggested that canopy species can prevent S. muticum germlings from reaching the substratum.

Evidence from experimental manipulations, combined with observations following natural disturbances to the integrity of the canopy, indicate that Undaria pinnatifida sporophytes also require disturbance to establish at high densities, while the presence of a stable native canopy inhibits sporophyte development (Floc’h et al. 1996, Valentine and Johnson 2003, Edgar et al. 2004, Valentine and Johnson 2004). Development of U. pinnatifida sporophytes appears to be inhibited mainly by competition for light, since the presence of a native canopy does not prevent development of U. pinnatifida gametophytes on the rocky reef substratum (Valentine and Johnson 2003). The most important disturbance, enabling seasonal development of dense virtually monospecific stands of U. pinnatifida, is overgrazing of native algae by a native sea urchin (Heliocidaris erythrogramma Valenciennes) to form so-called ‘‘barrens’’ habitat (Valentine and Johnson 2003, Johnson et al. 2004, Valentine and Johnson 2005a). Clearly, this phenomenon is contrary to the pattern in other (non-marine) systems where activities of native herbivores provide resistance to invasions by reducing the survival and abundance of invasive exotic species (Parker et al. 2006). Observational studies also indicate that disturbance to the native algal canopy is a critical precursor to the establishment phase in the invasion of Codium fragile ssp. tomentosoides in the Northwest Atlantic Ocean. C. fragile ssp. tomentosoides does not directly displace native kelp species wSaccharina longicruris (Pyl.) Kuntzex, but establishes following decline of the kelp, which has occurred in recent decades after infestation with the

Table 2 Summary of the role of disturbance in the establishment process for selected invasive seaweeds. Species

Role of disturbance

Reference

Establishment – required?

Nature of disturbance?

No, but dense seagrass beds less vulnerable to invasion No, but dense seagrass beds less vulnerable to invasion No, but removal of erect algal species enhanced establishment intensity Yes

Specific disturbance not tested (different seagrass densities observed naturally) Experimental thinning of seagrass canopy

deVille´le and Verlaque 1995

Experimental removal of erect algal species

Ceccherelli et al. 2002

Stormwater discharge implicated as disturbance

Jaubert et al. 2003

Codium fragile ssp. tomentosoides

Yes

Kelp canopy loss (due to epiphytic bryozoan Membranipora) Sea urchins eliminated by disease

Levin et al. 2002 Chapman et al. 2002

Fucus serratus

No, but disturbance accelerates establishment

Experimental canopy removal

Arrontes 2002

Sargassum muticum

Yes Yes Yes Yes

Natural canopy disappearance Experimental canopy removal Experimental canopy removal Experimental canopy removal

Ambrose and Nelson 1982 Deysher and Norton 1982 De Wreede 1983 Andrew and Viejo 1998

Undaria pinnatifida

Yes Yes Yes Yes

Experimental canopy removal Experimental canopy removal Canopy dieback Experimental canopy removal

Floc’h et al. 1996 Valentine and Johnson 2003 Valentine and Johnson 2004 Edgar et al. 2004

Caulerpa taxifolia

w34x

Yes

Ceccherelli and Cinelli 1999

J.P. Valentine et al.: Mechanisms of seaweed invasion 355

invasive epiphytic bryozoan Membranipora membranacea (Linnaeus 1758) (Levin et al. 2002). C. fragile ssp. tomentosoides has also been observed to establish on former sea urchin barrens following epidemics of Paramoeba invadens (Jones 1985) that eliminate sea urchins (Scheibling and Hennigar 1997), and in estuarine areas following dieback of eelgrass (Trowbridge 1998). There is conflicting evidence regarding the importance of disturbance in the establishment of Caulerpa taxifolia. Manipulative experiments and observations of seagrass beds indicate that C. taxifolia can invade native beds in the absence of disturbance, however, it appears that dense seagrass meadows are more resistant to invasion than are sparse meadows (deVille´le and Verlaque 1995, Ceccherelli and Cinelli 1999). Similarly, manipulation by removal of native macroalgal canopies on rocky reefs has shown that communities with a native canopy are more resistant to C. taxifolia invasion than those without a developed canopy (Ceccherelli et al. 2002). Some authors hypothesise that C. taxifolia drives community dynamics by accelerating the decline of native seagrass wPosidonia oceanica (Linnaeus) Delilex, although the decline may be instigated by multiple disturbance agents (deVille´le and Verlaque 1995). Others provide correlative evidence suggesting that although C. taxifolia occurs in many habitat types, in high density stands it occurs principally in stressed environments where resources (e.g., light and space) are available because other species are rare, as in sites subject to stormwater and sewage outfalls (Jaubert et al. 2003). Clarifying the role of disturbance in the process of establishment of C. taxifolia, and, therefore, a more complete assessment of the risk that this alga poses to native communities, requires further research. Observations of Fucus serratus in the western Atlantic Ocean indicate the likely importance of disturbance in the invasion process, since it occurs most abundantly on friable sandstone substratum (Chapman et al. 2002) and in areas with frequent ice scour (A.R.O. Chapman personal communication). An experimental study examining the invasion of F. serratus on eastern Atlantic shores of Spain has further demonstrated the importance of disturbance in the invasion process (Arrontes 2002). While disturbance to native canopies is not required for establishment of F. serratus, disturbance affects the rate at which establishment occurs, and can significantly accelerate the establishment process. For this species, disturbance increases the probability of successful establishment. Mechanisms of spread Mechanisms of spread are often tightly coupled with the processes facilitating establishment. Clearly, if disturbance is required for establishment then it follows that spread will also be dependent on the frequency and intensity of disturbance. The dispersal characteristics of introduced seaweeds also have a major bearing on their capacity to spread. While the exact dispersal mechanisms among invasive species differ, each of the remaining species that have demonstrated large scale incursions possess efficient strategies for dispersal, with the exception of Fucus serratus, which has poor dispersal characteristics (Table 1).

Some species, including Codium fragile ssp. tomentosoides, Sargassum muticum and Undaria pinnatifida, manifest adaptations for both short- and long-distance dispersal via microscopic propagules and drift thalli. Many of these invasive seaweeds also possess characteristics that make them particularly susceptible to dispersal aided by human activity. For example, Caulerpa taxifolia and C. fragile ssp. tomentosoides have the capacity to propagate vegetatively via fragmentation, a trait that not only aids natural spread, but also enhances the likelihood of human-assisted transport, in association with anchors, nets, or ships’ sea chests. Anthropogenic transport of fragments can account for ‘‘leaping’’ of introduced species to geographically distant sites without an intermediate station (Ribera and Boudouresque 1995). Mechanisms of persistence: is continued disturbance important? While disturbance may be an essential process facilitating establishment and spread of introduced seaweeds, its role in the persistence of invaders is complex, and varies substantially among species. Codium fragile ssp. tomentosoides stands that have established are selfmaintaining and able to persist in the absence of the disturbance mechanism that enabled initial establishment. Experimental manipulations of established C. fragile ssp. tomentosoides canopies indicate that they inhibit recruitment of native kelps (Levin et al. 2002). This mechanism establishes a positive feedback to maintain dominance by C. fragile ssp. tomentosoides. A similar mechanism has been reported for Sargassum muticum which, once established, largely prevents recruitment and development of other algae (Ambrose and Nelson 1982). Recent experiments indicate that the negative effects of S. muticum on other algae are mainly the result of shading, rather than changes in water flow, sedimentation rates, or nutrient availability (Britton-Simmons 2004). A positive feedback mechanism facilitating persistence has also been demonstrated for Undaria pinnatifida. However, the mechanism is more complex than that described for Codium fragile ssp. tomentosoides and Sargassum muticum. Manipulative experiments in Tasmania indicate that the primary disturbance factor, sea urchin grazing, facilitates establishment of U. pinnatifida after overgrazing and subsequent destruction of native canopy-forming algae (Valentine and Johnson 2003, 2005a). However, removal of the sea urchins is not sufficient to promote recovery of native species because following native canopy removal, a matrix of sediment and filamentous algae accumulates on the reef surface (Valentine and Johnson 2005a,b). Development of the sediment matrix appears to inhibit settlement and development of native macroalgal propagules and, therefore, the recruitment and recovery of native canopy-forming species. U. pinnatifida is thought to persist on sedimentaffected reefs due to either a greater tolerance of sediment stress, or by exploiting periods of low sediment following events of elevated surge (Valentine and Johnson 2005b). It is able to exploit these windows of opportunity by virtue of the rapid growth of sporophytes. Thus, U. pinnatifida and the sediment matrix coexist and persist indefinitely. This complex mechanism is a stronger posw35x

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itive feedback than arises in the case of C. fragile ssp. tomentosoides and S. muticum since removal of the alien species does not realise recovery of native canopy-forming seaweeds in the case of U. pinnatifida, but may affect recovery of canopy forming natives in the cases of C. fragile ssp. tomentosoides and S. muticum. Indeed, in the Tasmanian system, removal of both the sea urchins and U. pinnatifida, and an enhanced inoculum of spores of native species, was insufficient to regenerate native canopy forming species (Valentine and Johnson 2005a). Factors influencing persistence have not been examined in detail for Caulerpa taxifolia. However, there is correlative evidence that positive feedback mechanisms also exist for this species. Invasion by C. taxifolia results in significant alteration to the environmental characteristics of seagrass beds, altering rates of sediment accumulation and inhibiting the growth of native seagrass shoots by shading (deVille´le and Verlaque 1995). In addition, C. taxifolia is known to produce toxic secondary metabolites that assist in competition for space (deVille´le and Verlaque 1995). The phase-shift from native algae to dominance by introduced species as a result of positive feedback (particularly for Caulerpa taxifolia, Codium fragile ssp. tomentosoides and Undaria pinnatifida) is cause for concern. Phase-shifts in community structure providing for alternative stable states have been demonstrated previously for marine systems, usually associated with humanmediated disturbance (e.g., Jackson et al. 2001, Bellwood et al. 2004), with potentially dramatic and longterm implications for ecosystem structure and function. The likelihood that a system affected by an invasive seaweed can return to its original state is a matter of critical interest. When a system manifests little capacity to recover after cessation of a mechanism responsible for a phase-shift, or when the recovery trajectory differs from that observed during the decline, the system is said to demonstrate an ecological hysteresis (Hughes et al. 2005, Schro¨der et al. 2005). In all cases, the advent of hystereses through positive feedback provides particular challenges for management. Phase-shifts to dominance by introduced seaweeds also have potentially serious implications for secondary consumers. For example, research has shown that C. fragile ssp. tomentosoides, C. taxifolia and S. muticum are not the preferred food of native grazers in their introduced range (Prince and LeBlanc 1992, deVille´le and Verlaque 1995, Britton-Simmons 2004). Moreover, diets based on consumption of exotics can result in lowered reproductive output of grazers (Scheibling and Anthony 2001).

Mechanisms of invasion: potential ‘‘system management’’ control options Where disturbance to the native algal canopy is a necessary prerequisite for the establishment and persistence of invasive seaweeds at high densities, managers are provided with alternative options to targeting seaweed thalli directly. Where disturbance can be linked to human activity, indirect control options may exist by w36x

focusing efforts to minimise anthropogenically-mediated disturbance. A range of disturbances or physiological stresses can lead to a reduction in cover of native algal canopies, including physical damage by storms (Kennelly 1987, Dayton et al. 1992), high water temperatures (Tegner and Dayton 1987), burial or abrasion by sediments (Airoldi et al. 1996, Airoldi and Virgilio 1998), sea urchin grazing (Lawrence 1975, Ayling 1981, Harrold and Reed 1985, Johnson and Mann 1988, Keats et al. 1990, Andrew 1993, Bulleri et al. 1999, Scheibling et al. 1999, Villouta et al. 2001), pollution (Hardy et al. 1993), and the presence of introduced species (Levin et al. 2002). Many of these disturbances are influenced by human activities and there is significant evidence to suggest that there has been a world-wide decline of canopy-forming species over the last 30 years (Benedetti-Cecchi et al. 2001). For example, sedimentation is increasing on rocky coasts around the world as a direct result of industrial and domestic discharges and as an indirect result of modifying coastlines and river catchments (reviewed by Airoldi 2003). Similarly, the predicted effects of global warming include not only increased water temperatures, but also increased frequency and intensity of storms, both of which can lead to reductions in cover of canopy-forming species (Coelho et al. 2000). There are significant opportunities for researchers to undertake studies designed to recognise and minimise the impact of human activities on canopy-forming species. Insight from this work has the potential to provide management with a range of options to prevent and/or control seaweed invasions. The available evidence suggests that focusing management efforts to prevent disturbance to the native canopy may represent a feasible control option for Codium fragile ssp. tomentosoides, Sargassum muticum and Undaria pinnatifida, all of which require disturbance for successful establishment. Recent work in Tasmania represents a case-study demonstrating the potential for system-level management of U. pinnatifida. In Tasmania, the spiny lobster Jasus edwardsii (Hutton 1875) is more important than reef fishes as a predator of sea urchins (Heliocidaris erythrogramma Valenciennes 1846) and, moreover, reduced abundances of lobsters as a result of fishing activity are sufficient to account for formation of ‘‘urchin barrens’’ (Pederson 2003, Johnson et al. 2004, Pederson and Johnson 2006). It is likely that overfishing of sea urchin predators is ultimately the single most important factor causing reduced native algal cover on the east coast of Tasmania, which has facilitated establishment of dense U. pinnatifida stands. Managing populations of J. edwardsii to maintain sea urchin populations at low levels, therefore, provides a potential option for control of U. pinnatifida. For Codium fragile ssp. tomentosoides in the north western Atlantic Ocean, there is strong evidence that the organisms demonstrated to facilitate the invasion (Membranipora membranacea and Paramoeba invadens) are themselves introduced species (Chapman et al. 2002). Consequently, efforts to prevent introduction of these exotic species represents a potential control option to avoid further C. fragile ssp. tomentosoides outbreaks. For Caulerpa taxifolia and Fucus serratus, efforts to minimise canopy disturbance may slow down the pro-

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cess of establishment and spread, and may reduce the ultimate density at which they establish, but would not be sufficient to prevent establishment from occurring. Clearly not all disturbances can be prevented or controlled. For example, where the facilitating canopy disturbance is natural (e.g., storm scour), practical control options are not available. In such instances the magnitude of invasion will depend on the frequency and intensity of natural disturbance events, and control options will be restricted to targeting the seaweed directly. While the current synthesis demonstrates a degree of predictability in the response of invasive seaweeds to disturbance, it needs to be acknowledged that responses to manipulations (in experiments) or external forcings (in nature) may differ from one location to another. In relation to introduced seaweeds, this kind of spatial variability may be a result of introduced species displaying characteristics not evident in their native range (Ribera and Boudouresque 1995). For example, Sargassum muticum occupies a wider ecological niche and grows much larger in European shores compared with Japanese populations (Norton 1977, Rueness 1989). Spatial variability in invasion dynamics is clearly an important consideration and can provide particular challenges to formulating management responses to invasions. Consequently, it should be stressed that wherever possible, experiments should be replicated in space and time to establish generality in the response of invasive seaweeds to disturbance, and so ensure greater confidence in predicting invasion dynamics. It should be emphasised that once an exotic species is established, management to control it by targeting the disturbance may be effective if the disturbance that facilitates establishment is also required for persistence. In contrast, the existence of positive feedbacks and ecological hystereses, which reinforce dominance of invasive seaweeds (such as those described for Undaria pinnatifida, Codium fragile ssp. tomentosoides and Sargassum muticum), present a significant challenge for management. In these examples, ‘‘system management’’ is problematic, since targeting the enabling disturbance after establishment is unlikely to represent an effective control strategy. Where the existence of different ‘‘stable states’’ (reflected as dominance by native canopy-forming seaweeds or introduced seaweeds) can be demonstrated, it is particularly important that future studies examine the factors that promote recovery of native canopy-forming species. From a management perspective, it is clearly preferable to sustain a resilient ecosystem than to attempt to rehabilitate it after a phase-shift has occurred (Hughes et al. 2005).

Mechanisms of invasion: the importance of spatial and temporal scale Scale is well recognised as a fundamental aspect of ecology (Levin 1992) and must be considered when examining mechanisms of invasion and potential management options. Another layer of complexity in the dynamics involving interactions between Undaria pinnatifida, disturbance, native seaweeds and the sediment matrix

relates to effects of spatial scale. The sediment matrix and high cover of U. pinnatifida develop regardless of whether disturbance to native canopy species occurs at small scales (e.g., in manipulation experiments; Valentine and Johnson 2003) or at large scales as a result of overgrazing of native algae by sea urchins (Johnson et al. 2004, Valentine and Johnson 2005a). However, the response of U. pinnatifida to cessation of disturbance to native canopy species is scale-dependent. Where there has been large scale destruction of the native canopy (spatial extent approximately 1.2 ha), there is no recovery of native canopy-forming algae following removal of sea urchins, U. pinnatifida, or both (Valentine and Johnson 2005a). However, when holes in the canopy are relatively small (16 m2), and there is no sea urchin grazing, the native species displace any U. pinnatifida thalli that develop in response to the initial disturbance within two years (Valentine and Johnson 2003). Thus, the strong positive feedback mechanism seems only to operate for canopy removals at larger scales. The mechanism shaping this non-linear scale-dependent response is unknown. A possible explanation is that sediment accumulation following canopy disturbance is itself scale-dependent. Greater development of the sediment matrix associated with the larger-scale disturbance may inhibit recovery of native algae, while the accumulation of sediment in the smaller experimental plots is insufficient to prevent recovery of native canopy-forming species. Propagule supply may also be influenced by the scale of the disturbance, potentially affecting recruitment processes for native canopy-forming algae, particularly those with low dispersal abilities. These kinds of mechanisms have implications for assessing the threat of an introduced seaweed species. In this particular example, it appears that the results from the small-scale experimental plots would have resulted in an underestimate of the threat posed by Undaria pinnatifida. Where possible, experiments examining mechanisms of invasion should be conducted on a range of spatial scales to infer generality of results. The temporal scale of observation may also have a large bearing on the ability to critically examine mechanisms of invasion, particularly in relation to persistence. Unfortunately, studies are rarely conducted on a time scale sufficient to examine more than one turnover of a generation of macroscopic thalli, particularly for species with longer life spans. The lack of long-term studies may represent a potential bias when examining persistence mechanisms, potentially overestimating the strength of positive feedback loops. Long-term studies are clearly required to examine whether native perennial species eventually exert dominance over pseudo-perennial (i.e., aspect annual) or annual invasive species.

Conclusion Our broad findings are that knowledge of life history traits provides only limited insight into the potential for invasion of exotic seaweeds, but that disturbance can play an important role in the invasion process, particularly in relation to establishment and spread. This can provide w37x

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important options for management to control seaweed invasions. Where a facilitating disturbance is linked with human activity, it may be more practical and effective to target the disturbance rather than the alga directly. However, this approach is not likely to be effective once an invader is established, particularly when positive feedbacks arise to maintain the dominance of invasive seaweeds. The clear message is that ‘‘prevention’’ is easier and less costly than ‘‘cure’’. There are significant opportunities for further research in this area, particularly in relation to understanding the mechanisms that enable invasive species to outcompete native species following disturbance. This research would fruitfully examine nutrient dynamics and macroalgal physiology, an aspect of seaweed invasions that has received scant attention to date. Work in this area has been conducted only with Codium fragile ssp. tomentosoides, and has shown that invasion in this species is due in part to its ability to compete for and acquire nitrogen (Trowbridge 1998). There are significant opportunities to extend this line of enquiry to other species. Given the importance of disturbance and the properties of the recipient community (Dunstan and Johnson 2007) in the invasion process, there is also considerable scope for research to examine the factors conferring resistance to invasion in native seaweed communities. What aspects of the native community provide this resistance? It is not simply a function of species richness (Dunstan and Johnson 2007), because other elements of community complexity may influence resistance (BrittonSimmons 2006). This question will only be answered with carefully designed experimentation and analysis (Ceccherelli et al. 2002, Britton-Simmons 2006, Dunstan and Johnson 2007).

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Pederson, H.G. and C.R. Johnson. 2006. Predation of the sea urchin Heliocidaris erythrogramma by rock lobsters (Jasus edwardsii) in no-take marine reserves. J. Exp. Mar. Biol. Ecol. 336: 120–134. Piazzi, L. and F. Cinelli. 2000. Effects of the spread of the introduced Rhodophyceae Acrothamnion preissii and Womersleyella setacea on the macroalgal community of Posidonia oceanica rhizomes in the Western Mediterranean sea. Cryptogam. Algol. 21: 291–300. Prince, J.S. and W.G. LeBlanc. 1992. Comparative feeding preference of Strongylocentrotus droebachiensis (Echinoidea) for the invasive seaweed Codium fragile ssp. tomentosoides (Chlorophyceae) and four other seaweeds. Mar. Biol. 113: 159–163. Prince, J.S. and C.D. Trowbridge. 2004. Reproduction in the green macroalga Codium (Chlorophyta): characterisation of gametes. Bot. Mar. 47: 461–470. Ribera, M.A. and C.F. Boudouresque. 1995. Introduced marine plants, with special reference to macroalgae: mechanisms and impact. In: (F.E. Round and D.J. Chapman, eds) Progress in phycological research. Vol. 11. Biopress Ltd, Bristol. pp. 217–268. Rueness, J. 1989. Sargassum muticum and other introduced Japanese macroalgae: biological pollution of European coasts. Mar. Poll. Bull. 20: 173–176. Saito, Y. 1975. Undaria. In: (J. Tohida and H. Hirose, eds) Advance of phycology in Japan. Junk, The Hague. pp. 304–320. Sakai, A.N., 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. Ann. Rev. Ecol. Syst. 32: 305–332. Sant, N., O. Delgado, C. Rodrı´guez-Prieto and E. Ballesteros. 1996. The spreading of the introduced seaweed Caulerpa taxifolia (Vahl) C. Agardh in the Mediterranean Sea: testing the boat transportation hypothesis. Bot. Mar. 39: 427–430. Schaffelke, B., M.L. Campbell and C.L. Hewitt. 2005. Reproductive phenology of the introduced kelp Undaria pinnatifida (Phaeophyceae, Laminariales) in Tasmania, Australia. Phycologia 44: 84–94. Scheibling, R.E. and S.X. Anthony. 2001. Feeding, growth and reproduction of sea urchins (Strongylocentrotus droebachiensis) on single and mixed diets of kelp (Laminaria spp.) and the invasive alga Codium fragile ssp. tomentosoides. Mar. Biol. 139: 139–146. Scheibling, R.E. and A.W. Hennigar. 1997. Recurrent outbreaks of disease in sea urchins (Strongylocentrotus droebachiensis) in Novia Scotia: evidence for a link with large-scale meteorologic and oceanographic events. Mar. Ecol. Prog. Ser. 152: 155–165. Scheibling, R.E., A.W. Hennigar and T. Balch. 1999. Destructive grazing, epiphytism, and disease: the dynamics of sea urchin–kelp interactions in Nova Scotia. Can. J. Fish. Aquat. Sci. 56: 2300–2314. Schro¨der, A., L. Persson and A.M. De Roos. 2005. Direct experimental evidence for alternative stable states: a review. Oikos 110: 3–19.

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Shea, K. and P. Chesson. 2002. Community ecology theory as a framework for biological invasions. Trends Ecol. Evol. 17: 170–176. Smith, C.M. and L.J. Walters. 1999. Fragmentation as a strategy for Caulerpa species: fates of fragments and implications for management of an invasive weed. Mar. Ecol. 20: 307–319. Stæhr, P.A., M.F. Pederson, M.S. Thomsen, T. Wernberg and D. Krause-Jensen. 2000. Invasion of Sargassum muticum in Limfjorden (Denmark) and its possible impact on the indigenous macroalgal community. Mar. Ecol. Prog. Ser. 207: 79–88. Tegner, M.J. and P.K. Dayton. 1987. El Nino effects on Southern California kelp forest communities. Adv. Ecol. Res. 17: 243–279. Trowbridge, C.D. 1995. Establishment of the green alga Codium fragile ssp. tomentosoides on New Zealand rocky shores: current distribution and invertebrate grazers. J. Ecol. 83: 949–965. Trowbridge, C.D. 1998. Ecology of the green macroalga Codium fragile (Suringar) Hariot 1889: invasive and non-invasive subspecies. Oceanogr. Mar. Biol. Ann. Rev. 36: 1–64. Trowbridge, C.D. 1999. An assessment of the potential spread and options for control of the introduced green macroalga Codium fragile ssp. tomentosoides on Australian shores. CSIRO Marine Research, Hobart, Tasmania. pp. 43. Valentine, J.P. and C.R. Johnson. 2003. Establishment of the introduced kelp Undaria pinnatifida in Tasmania depends on disturbance to native algal assemblages. J. Exp. Mar. Biol. Ecol. 295: 63–90. Valentine, J.P. and C.R. Johnson. 2004. Establishment of the introduced kelp Undaria pinnatifida following dieback of the native macroalga Phyllospora comosa in Tasmania, Australia. Mar. Freshw. Res. 55: 223–230. Valentine, J.P. and C.R. Johnson. 2005a. Persistence of the exotic kelp Undaria pinnatifida does not depend on sea urchin grazing. Mar. Ecol. Prog. Ser. 285: 43–55. Valentine, J.P. and C.R. Johnson. 2005b. Persistence of sea urchin (Heliocidaris erythrogramma) barrens on the east coast of Tasmania: inhibition of macroalgal recovery in the absence of high densities of sea urchins. Bot. Mar. 48: 106–115. Vermeij, G.J. 1978. Biogeography and adaptation, patterns of marine life. Harvard University Press, Cambridge. pp. 332. Villouta, E., W.L. Chadderton, C.W. Pugsley and C.H. Hay. 2001. Effects of sea urchin (Evechinus chloroticus) grazing in Dusky Sound, Fiordland, New Zealand. New Zeal. J. Mar. Freshw. 35: 1007–1024. Walker, D.I. and G.A. Kendrick. 1998. Threats to macroalgal diversity: marine habitat destruction and fragmentation, pollution and introduced species. Bot. Mar. 41: 105–112. Zˇuljevic, A. and B. Antolic. 2000. Synchronus release of male gametes of Caulerpa taxifolia (Caulerpales, Chlorophyta) in the Mediterranean Sea. Phycologia 39: 157–159.

Received 19 January, 2006; accepted 12 June, 2006

Botanica Marina 50 (2007): 361–372

2007 by Walter de Gruyter • Berlin • New York. DOI 10.1515/BOT.2007.041

Review

Mechanisms of invasions: can the recipient community influence invasion rates? Piers K. Dunstan1,* and Craig R. Johnson2 CSIRO Marine and Atmospheric Research, GPO Box 1538, Hobart, Tasmania 7001, Australia, e-mail: [email protected] 2 School of Zoology and Tasmanian Aquaculture and Fisheries Institute, University of Tasmania, GPO Box 252-05, Hobart, Tasmania 7001, Australia 1

* Corresponding author

Abstract At least since Elton’s work in the 1950s it has been argued that properties of recipient communities influence invasion rates. The widespread view, initiated by Elton and later supported by both empirical and modelling studies, is that species-rich communities are more resistant to invasion than species-poor communities in otherwise identical habitats. However, the dynamics of some empirical systems do not support this idea, but reveal instead a positive relationship between species richness and invasion, reflecting the particular attributes of component species and their interactions. These findings cannot be explained solely by Shea and Chesson’s (2002) attempt to reconcile divergent observations of both positive and negative relationships between invasion rates and richness of the recipient community. Here, based on the behaviours of several models, we develop an alternative view that invasion success is related positively to variability in resource availability. This variability can arise in many ways, dependent in part on the life history features of component species, the nature of interactions among species, and the spatial arrangement of individuals on a landscape. Our overall conclusion is that the ability of a community to resist the establishment of new populations comes not from species richness, diversity or the average connectivity of the system per se, but from the ability of established species to utilise the maximum amount of resources in the long term and realise the smallest amount of variability in resource availability. We argue that seaweed communities adhere to this principle. Keywords: invasion resistance; invasions; resource limitation; spatial interactions.

Introduction Invasion resistance and richness: the conventional view Two elements critical to the establishment and spread of invasive populations are the arrival of new propagules of

the potential invader and the ability of the recipient community to resist their establishment. The ability of a community to resist invasions has been a subject of ongoing research since Elton’s (1958) seminal work. Invasions into marine ecosystems continue to occur, arguably with increasing frequency (Cohen and Carlton 1998, Ruiz et al. 2000), but they are not as frequent as might be expected from a virtually constant supply of propagules from anthropogenic vectors (Carlton and Geller 1993). This observation keeps sharply in focus the question of why it is that alien species often fail to establish or invade at new sites despite inoculation. The impetus to more fully understand the mechanism(s) of resistance of communities to invasion is as urgent as at any time. Elton’s (1958) principal conclusion was that the ‘‘balance of nature’’ in simple communities was more easily ‘‘upset’’ (i.e., more prone to fluctuations) than in more complex communities. He argued that the simple low diversity communities that could be more easily ‘‘upset’’ were more susceptible to invasion by alien species. Based on qualitative observations across a diversity of habitats and ecosystems, his argument was compelling and provoked substantial debate. However, it needs to be emphasised that Elton stated his hypothesis as two separate ideas. First, he suggested that more complex communities, that is, communities with relatively more species and niches, were more stable than simpler, species-poor communities. MacArthur (1955) suggested a similar line of reasoning. Secondly, Elton (op. cit.) suggested that communities with varying species abundances (i.e., with greater oscillations in species abundances), should be more prone to invasions than communities that are more stable and less prone to oscillations. It follows that, as abundances of species vary, so too will patterns of resource utilisation. Consequently, total levels of resource availability will vary, even if the total amount of resource does not change. These remarkably prescient ideas have remained part of the framework of ecological theory over the past 50 years. While the lines of evidence forwarded by Elton do not stand up to exacting scrutiny, other more quantitative evidence has emerged to support his ideas. Most of the debate about invasions has focused on the idea that species diversity confers resistance to invasions. Considerable empirical evidence has revealed a negative relationship between the species richness of particular areas of interest (e.g., patches) and the invasibility of those areas (e.g., McGrady-Steed et al. 1997, Tilman 1997, Stachowicz et al. 1999, Levine 2000, Symstad 2000, Hector et al. 2001, Kennedy et al. 2002, Kolb et al. 2002). However, there is also a growing body of empirical evidence showing a positive relationship between richness and invasibility (e.g., Robinson et al. 1995, Wiser et al. 1998, Stohlgren et al. 1999, Cleland et al. 2004, w41x

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Dunstan and Johnson 2004, Meiners et al. 2004, Jiang and Morin 2004). These conflicting results, all based on analyses of real ecological systems, have provoked intense debate (Levine and D’Antonio 1999). Shea and Chesson (2002) suggested that the divergence can be reconciled by considering the scale of observation. They suggest that since environmental heterogeneity increases with scale, observations at large scales will encompass several habitat types, including poor habitats unable to support many species (native or alien), and favourable habitats able support a large number of species (native or alien). Accordingly, observations at large scales across multiple habitat types will show a positive relationship between species richness and the number of aliens. These patterns may be explained by the spatial variation in resource use and supply (Davies et al. 2005), but not by the average conditions across large scales. This hypothesis has been supported by recent empirical work (i.e., Stohlgren et al. 1999, Levine 2000, Brown and Peet 2003). Conversely, Shea and Chesson (2002) argue that at smaller scales, within the same habitat type, species-rich patches will be more resistant to invasion than species-poor patches, so that the relationship between richness and invasibility is negative, in line with Elton’s initial ideas. Invasion resistance: an alternative viewpoint based on resource variability Shea and Chesson’s (2002) synthesis does not account for the mounting number of studies that find positive or varying relationships between richness and invasibility (as a function of the number of alien species present) at small scales within the same habitat type (Davis and Pelsor 2001, Meiners et al. 2004, Smith et al. 2004, Cleland et al. 2004, Dunstan and Johnson 2004). As the broad thesis of this paper, we suggest that the common unifying feature across all of these studies, which cover a diversity of ecosystem types, is that the positive relationship between richness and the likelihood of invasion stems from particular properties of component species and their interactions to generate particular patterns of resource utilisation and community dynamics. Since the results of this suite of studies are diametric to the conventional view (after Elton 1958) and its modern synthesis (e.g., Levine 2000, Shea and Chesson 2002), it is useful to examine the underlying mechanisms in more detail. Meiners et al. (2004) found that invading species did not always respond in a monotonic way to increasing richness. Rather, the relationship between richness and probability of invasion depended on the particular composition of the recipient community. Communities with particular species were more resistant to invasion than others without these species, independently of total richness. Cleland et al. (2004) found that patterns of native and alien species richness were positively correlated within four sites located in grassland and desert biomes. However, they also found that the relative abundance of alien species was negatively correlated with the richness of native species. This implies that a community with one or two dominant species will be more resistant to invasion than those where the abundances of species are more even. They suggested that these patterns were w42x

driven by the higher resource utilisation of the resident communities with one or two dominant species, which reduced the ability of invasive (and other native) species to establish and spread in the community. Dunstan and Johnson (2004) similarly found that the frequency of recruitment into an established marine epibenthic community increased with increasing species richness of patches. They identified two different mechanisms that underpin this pattern. The primary mechanism arose as a result of size-specific mortality. Larger colonies were less likely to die (and so free up space for new recruits) than smaller colonies. Thus, opportunities for invasion of patches dominated by relatively few large long-lived colonies/individuals were notably fewer than in patches characterised by a large number of small colonies/individuals, which were inevitably more species-rich. The common element in these studies is the capacity of the established community to utilise resources and, thus, limit the establishment of the invader into the community. Davis et al. (2000) first suggested that the magnitude of fluctuation in resources could be the key to determining the likelihood of establishment of new species into a community. Because resident individuals will be unable to respond instantaneously to fully utilise a resource that becomes available (either through the resource increasing or the release of resources through the death of resident individuals), when there are pronounced fluctuations in resource availability there will be more opportunities for a new species to establish. Davis and Pelsor (2001) tested this hypothesis in a terrestrial plant community and found that resource supply was a critical determinant of invasion success. When resources were limiting (in their study, water supply was the varying resource), competition was intense and new plants had a reduced rate of successful establishment. When water was not limiting, competition was weak and establishment of new plants was significantly easier. Davis et al. (2000), who were principally concerned with invasions into plant communities, argued that there is no reason to suspect that richness per se should be correlated with invasibility. Here we suggest that this notion applies to ecological systems in general. In any community, species (whether alien or native) will tend to colonise until either the resources available are exceeded or environmental limits are reached. At this point, further colonisation will be limited by the availability of resources, including space. If the available resource fluctuates more rapidly than the established individuals are able to respond, then ‘‘peaks’’ in resource availability will facilitate establishment of new individuals, including invasive aliens. This mechanism need not be directly related to richness. Thus, within a habitat type, depending on the nature of the limiting resource base and the mechanism(s) underlying its variability, the relationship between richness of the recipient community and resistance to invasion may be either positive or negative. Invasion resistance: the importance of spatial organisation Closely related to the issue of variability in resource availability is that of spatial effects. Many communities,

P.K. Dunstan and C.R. Johnson: Mechanisms of invasion 363

including seaweed vegetation (e.g., Dayton et al. 1984, Sousa 1984), exist as dynamic mosaic of patches. A given mosaic will, at least in part, reflect a particular history of resource variability. For example, it is well established that seaweed community dynamics can be strongly influenced by spatial patterns in disturbance, the identity of species adjacent to disturbed areas (and their reproductive state at the time of the disturbance), and opportunties for dense patches of particular species to develop that limit colonisation by other, even dominant, species (e.g., Sousa 1979, 1984, Sousa et al. 1981, Dayton et al. 1984, Johnson and Mann 1988, Chapman and Johnson 1990, Valentine et al. 2007). These examples emphasise that the dynamics of a particular patch can depend in large part on spatial patterns in disturbance, and on the identity and phenology of local neighbours. Thus, since communities can often be viewed as collectives of local ‘‘patches’’, it follows that spatial pattern can potentially strongly influence broader scale dynamics. While these kinds of spatially explicit mechanisms are well recognised by experimentalists, they have not been embraced widely by modellers. Modelling is useful as a powerful way to generalise empirical results, and to more fully explore concepts. However, conventional models of invasion dynamics (e.g., Case 1990, 1991) are usually based on a mean field approach, which implicitly assumes a fully mixed and homogenous arrangement of individuals. This is, of course, a gross simplification of the structure of natural systems. Most communities, whether dominated by seaweeds or not, show a nonrandom spatial arrangement of species within and between habitats at most spatial scales. Invasion resistance: approaches using modelling Models are useful tools to help develop and clarify thinking about invasion dynamics and the underlying processes. Thus far, results of models have supported the tenet that invasion resistance increases with species richness, and models have not reproduced the empirical finding that, in some cases, invasion resistance declines with increased richness (e.g., Robinson and Valentine 1979, Shigesada et al. 1984, Case 1990, 1991, Drake 1990, Law and Morton 1996, Doak et al. 1998, Byers and Noonburg 2003, Fridley et al. 2004). Among the most frequently cited papers are those of Case (1990, 1991). His model simulates invasions into artificially constructed communities of differing sizes. The methods were initially established by MacArthur and Levins (1967), and considered further by Roughgarden (1974) and Abrams (1975), among others. Pivotal to the construction of these models is the assumption that species consume resources from a single resource dimension. Each species has a resource utilisation curve that defines the shape of resource utilisation over that resource dimension. The extent of overlap of the resource utilisation curves of any two species defines the intensity of competition between them, and it is assumed that two species with identical resource utilisation curves cannot coexist. MacArthur and Levins (1967) and Roughgarden (1974) found that the success of an invasion was severely constrained by the size and shape of the utilisation curves, and the invaders would be excluded under

many scenarios. The structure of these communities was limited by similarities in resource utilisation, so that multiple species with similar resource needs could not survive. Case (1990, 1991) extended the ideas of this early work and considered multispecies systems of more than three species. The model communities assembled by Case (1990, 1991) were constructed in such a way that they were always analytically stable, and they were allowed to run to stable equilibrium after the addition of an invasive species whose resource utilisation was the average of the existing community. The models of Case (1990, 1991) and later versions using a similar approach (Byers and Noonburg 2003) show that, for a fixed resource limit, invasion success declines as the number of species in the recipient community increases. Law and Morton (1996) also found that invasion rates decreased as they increased the number of species in their assembled systems. Again, their models relied on the stability of the Lotka-Voltera equations to define the success or otherwise of a new species entering a community. They found that as the number of species increased, the resistance of the community to new species increased. These results are in accord with the patterns found by May (1972) who showed that the potential parameter space for multispecies Lotka-Voltera equations decreases as species number increases. In other words, the more species there are, the harder it is to find a stable equilibrium. However, we question the two key assumptions that underpin these models, namely stable equilibria and limiting similarity. The outcomes of models are, of course, dependent on their inherent assumptions (e.g., Hewitt and Huxel 2002), and exploring the effects and validity of assumptions is an important exercise. Empirical observation suggests that marine communities in general, and seaweed communities in particular, are not in equilibrium. Importantly, by discarding the assumption of equilibrium it is possible to consider effects of fluctuating resource availability, which may generate behaviours different from those reported thus far. Secondly, we explore the possibility that niches and limiting similarity are not necessary to explain the success or failure of an invasion into a system. If there is a single resource, then by definition, the niches of all species will completely overlap and the fate of an invasion will depend on other properties of the system. Invasion resistance: this paper Against this background, we developed two models to examine effects of fluctuations in resource availability, spatial structure, patterns of resource utilisation, and patterns of interspecific interactions on invasion dynamics. The two models make quite different assumptions about the dynamics of species, but demonstrate some common behaviours. In the first model we focus on the importance of resource variability and the interactions this may have with species richness and invasion resistance. In the second model we consider the importance of spatial interaction between adults and the influence of network topology (i.e., richness and connectivity) on the outcomes of invasions. In both we examine whether invasibility is dependent more on particular properties of w43x

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communities, or on aggregate community measures such as richness. Finally, we relate the findings of these general models to the particular attributes of seaweeds in predicting broadscale patterns of their invasions. The design of the models was deliberately kept simple so that we could search for general patterns and dynamics. It was also a deliberate choice that the models are schematic because this simplicity enables clear explication of the underlying principles. We could make the models more complex and ‘‘realistic’’ in many ways, but over large areas of parameter space this would not change their qualitative behaviours. We are not attempting to develop accurate simulations of particular equilibrium or non-equilibrium scenarios for seaweeds, but are looking for our models to deliver clear signatures of nontrivial general behaviours.

Methods Model 1: non-spatial individual-based model Our first simple model allows fluctuations in resource availability. The model is a non-spatial individual-based system where individuals do not compete explicitly as adults. In other words, once an individual is established it is able to maintain its share of the resource base until it dies, and it cannot be displaced by another individual, either recruit or adult, of any species. When an individual dies, its resource base is freed and added to the pool of free resources. The model rules are: 1. Each species (i) has a resource consumption ri, that determines how much resource an individual requires to survive. Each individual consumes r resources from the total pool of resources K. The value of K is held constant for the duration of the simulation, but the total resources consumed varies as individuals reproduce and die. 2. Each species has a death rate, di (0, 1), which determines the probability of an individual dying. The probability of mortality di can be either independent of the species resource consumption or inversely related to the resource consumption of the species r

1q

so that dsd . In the latter case, if a species (i) i i consumed 100 resource units (rs100), the probability of mortality would be di2. Since di ranges between 0 and 1, increasing resource use will reduce the probability of mortality. 3. Each species has a birth rate, bi, which determines the likelihood of an individual producing a single offspring. Offspring from all species are randomly assigned to the community until all new offspring have been assigned to the community, or all resources are consumed and no new offspring can recruit to the community. In all simulations bis di /0.9. This ensures that the population will grow towards the carrying capacity, rather than following a random walk. For species where di is reduced, the magnitude of bi is reduced by a similar amount. This trade-off between births and deaths ensures that all species remain neutral in terms of their ability to sur100

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vive within a given habitat. If a species goes extinct locally then no more offspring are produced. 4. The community is initialised with 20 species and a total resource pool (K) of 1000. The number of individuals is varied to ensure that each species consumes the same total resource (50) when the model begins. Thus, if rx is 1, there will be 50 individuals of species x and if ry is 10 there will be 5 individuals of species y. 5. After 2000 iterations of the model, a new species invades the system. Depending on resource availability, up to 50 individuals of the new species are added to the system. If there are insufficient resources to support 50 individuals, then individuals are added until the system reaches K. The number of species at this point is uncontrolled and can vary between 2 and 20, depending on the random walk process that has occurred. The time when the new species is added is arbitrary, and adding species to the model at different times makes no difference to the final results (other than ts0). 6. The model is run for two different scenarios, first where mortality is independent of the resource consumption of individuals (i.e., mortalitysdi) and second, where mortality is dependent on resource r

1q

consumption (i.e., mortalitysdi ). For each scenario, each species i is randomly allocated a resource consumption rate ri, drawn uniformly between 1 and 20. For any given combination of parameters the simulations are repeated 10,000 times. 7. The variability of free resource is calculated as the variance in unconsumed resources between 1000 and 2000 iterations. The number of species present at 2000 iterations, and the number of individuals of the new species at 2500 iterations is recorded. Mean resource variability, species richness and number of successful invaders over the 10,000 independent simulations are calculated. 100

Model 2: spatially explicit individual-based model This model introduces two complexities absent in Model 1, namely interspecific competition between adults, and a spatially explicit environment. We introduce competition because any relationships between species richness and invasion evident from Model 1 might be altered by competition between adults. As the strength of interspecific competition increases, competitive interactions may play an increasingly important role in preventing the establishment of an invader. For example, Davis and Pelsor (2001) showed that as resources became limiting, competition increased and the establishment of invading species was limited. This model is made spatially explicit in part because it provides a mechanism to introduce resource variation, a critical element of Model 1. For example, depending on patchiness in the distribution of species, and the particular location and timing of disturbance, particular individuals or patches will have more opportunity to respond to free space than others. This logic suggests that spatially explicit models of invasion dynamics may yield dissimilar dynamics from their mean field counterparts, which

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assume a uniform distribution of component species. Indeed, it is well established that spatially explicit models may demonstrate qualitatively different behaviours from otherwise identical non-spatial models (e.g., Durrett and Levin 1994, Johnson 1997, Molofsky et al. 1999, Johnson and Seinen 2002). Note that in spatial models of the kind used here, species do not usually exist as individual cells except when they first recruit to the landscape. Instead, they develop as patches or colonies as a result of vegetative growth (e.g., Johnson 1997, Johnson and Seinen 2002, Habeeb et al. 2005, Dunstan and Johnson 2005). Model rules All spatially explicit models were undertaken using the Compete software (available at http:// www.zoo.utas.edu.au/CJPblist/PubListCJohnson2.htm), which enables individual-based models of competition among species on a 2-dimensional landscape. All models used a landscape of 400=400 cells with zero boundary conditions (i.e., the landscape surface was toroidal to avoid edge effects), rule structures for overgrowth defined a probabilistic cellular automaton based on the von Neumann neighbourhood (i.e., the four adjacent cells to the north, south, east and west), and updating of the landscape was synchronous. We ran the model under two different scenarios; first, with spatial organisation where ‘‘colonies’’ were allowed to form, and second, without spatial organisation where the positions of cells on the landscape were randomised between each time step, simulating the dynamics of a mean-field model. In comparing the dynamics of spatially explicit and equivalent mean-field models, all parameters relating to the biology of individual species and outcomes of interactions between species were identical in the two models. In all simulations, all species on the landscape had identical recruitment and growth rates, and all species could propagate vegetatively. Interactions among any two species X and Y were determined by three probabilities (sPr) that sum to unity, viz. PrwX displaces and overgrows Yx, PrwY displaces and overgrows Xx, Prwstandoff between X and Yx (Figure 1). Growth into an empty cell with at least one occupied neighbour in the von Neumann neighbourhood is determined by selecting the state of a neighbour chosen at random. When a cell Cx is occupied, a neighbour Cy is selected at random, and the likelihood of Cy displacing Cx is given by PrwCy)Cxx, which defines the probability of species Cy winning an encounter with Cx. In all simulations, the only instantaneous mortality that occurred was through disturbance. The likelihood of a disturbance event associated with any single cells10-4 per timestep, and each disturbance event cleared a square of 49 cells. Thus, on average, in any one time step 784 cells were cleared (;0.5% of the landscape). This does not mean that a given patch/colony was immortal, since patches were highly dynamic, continuously overgrowing and displacing one another. Recruitment was to empty space on the landscape made available by random disturbances which cleared clusters of cells on the landscape, and recruitment was open, i.e., recruits continued to arrive at the landscape

Figure 1 Possible outcomes of interactions between a pair of species X and Y. ): indicates overgrowth; s: indicates standoff.

at the same rate irrespective of the presence and abundance of adults on the landscape. Using this basic model architecture, we examined the effect of connectivity in the network topology and species richness on the capacity of ‘‘alien’’ species to invade self-organised landscapes of native species. The qualitative results do not depend on our simplifications to assign all species identical growth and recruitment rates and eliminate instantaneous mortality. Note, however, that if mortality and recruitment rates were made very large and competitive interactions reduced, then this model would behave similarly to Model 1. In this context, the two models as presented represent different ends of a spectrum. Effect of connectivity on invasion rates The prevailing paradigm is that communities characterised by high connectivity among species are more resistant to invasion than poorly connected communities (e.g., Case 1990). Connectivity is usually defined on the basis of the number and strength of pairwise interactions (e.g., Borrett and Patten 2003, Garlaschelli et al. 2003, Arii and Parrott 2004). In our models, the system is maximally connected in the sense that each species potentially competes with all others for space win an S species system there are S(S-1)/2 possible direct pairwise interactionsx, so connectivity is set by the strength of the interaction. Effectively, a pair of species does not interact if their interaction is defined as a standoff. We allowed an initial recruitment of 18 ‘‘native’’ species to an empty landscape, with each species recruiting with equal likelihood, and all recruits of all species initially covering a total of 2% of the landscape. These species were allowed to grow and the system spatially self-organise to occupy the entire landscape, such that an w45x

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approximately consistent community structure was attained. This landscape was then opened to potential invasion by two ‘‘alien’’ species, both with recruitment rates identical to the native species. For all 20 species in this system, the probability of recruitment to an empty cells0.01, thus, on average, ;0.4 new individuals of each species recruited to the landscape each time step. Because the dynamics of both alien species were virtually identical, the mean covers of the two invasive species were recorded after 200 time steps. We define connectivity as the proportion of standoffs in the 190 possible pairwise interactions in a 20-species system. A standoff between a species pair X and Y is defined as PrwXsYxs1.0. For each level of connectivity, the interaction matrix defining the network topology of the system was determined at random. For pairwise interactions that did not result in a standoff, we first ran sets of models where reversals in interaction outcomes were not possible, thus, where either PrwX)Yx or PrwY)Xxs1.0, determined at random. For each level of connectivity we conducted Monte Carlos of 1000 runs. Thus, in different runs within any given Monte Carlo series, the network topology defined by the interaction matrix was determined randomly, but in every case the number of standoffs in the interaction matrix was identical, with all other interactions defined as a binary outcome (win or lose), with the direction of win or loss determined at random. Thus, based on the number and strength of interactions (the usual way to define connectivity), each different run in any Monte Carlo series had identical connectivity. In nature, it is highly unlikely that the outcome of interactions between any two species is binary, i.e., where a species X displaces a species Y on all occasions in time and space. Depending on differences in size, physical environment, the identity of other neighbours, and other factors, it is more usual that reversals will arise, such that X will prevail in some circumstances and Y in others. Thus, we also conducted Monte Carlo sets where reversals were permitted, i.e., where PrwX)Yx, PrwY)Xx and PrwXsYx can take any value in the interval 0–1.0 such that sum of all three probabilitiess1.0. Again, each Monte Carlo series consisted of 1000 runs. Note that for these runs, while the number of 100% standoffs was identical in all runs, the strengths of interactions not involving 100% standoffs could vary at random. Effect of network topology on invasion rates If invasion rates are determined largely by the richness and connectivity of the recipient community as is suggested, then for a given rate of propagule arrival, the level of invasion of an alien species should be approximately constant for a fixed number of native species and connectivity in the recipient community. We examined this premise using the models described above for a simple recipient community of 10 native species, with a single alien as a potential invader, and where the number of standoffs in the interaction matrix was fixed (arbitrarily) at 45% and assigned randomly in each run. Where interactions were not 100% standoffs, we ran cases where outcomes were binary (either species X displaces Y, or Y displaces X, determined randomly, in every encounter), w46x

and where reversals were possible, as outlined above. Note that where the outcomes of non-standoff interactions are binary, the connectivity of each random configuration of the interaction network, defined in terms of the number and strength of interactions, is identical. We also compared the effect of differential recruitment rates, with the recruitment rate of all species (native and alien) held at either 0.01 or 0.001 per vacant cell per time step. We recorded the cover of each alien species 500 time steps after the first opportunity for introduction. Other aspects of the models were as described above.

Results Model 1: resource variability and the relationship between mortality and resource utilisation Although the model is intentionally simple, the dynamics demonstrate several important behaviours. When mortality is independent of individual resource consumption, the abundance of the invading species is negatively correlated with species richness (Figure 2A). Thus, species-rich communities resist potential invaders more effectively than species-poor communities, showing the same general relationships between species richness and invasion as other model communities (e.g., Case 1990, 1991, Law and Morton 1996). In marked contrast to this pattern, if mortality rates are dependent on levels of resource utilisation, then invasion success is positively correlated with species richness (Figure 2A). Note, however, that the correlation between invasion success and resource variability is positive, irrespective of whether mortality is dependent or independent of resource utilisation (Figure 2B). The mechanisms underpinning these patterns reflect patterns of resource availability. When mortality is independent of individual resource consumption, the more resources an individual utilises, the fewer the number of individuals that can be packed into the community, since the total utilisation is always limited to 1000. Thus, species-rich communities have more species that consume fewer resources. The loss of an individual from a species that requires few resources frees fewer resources than loss of an individual utilising a greater proportion of the total resources. Thus, resource variability associated with mortality of an individual in a species-rich system is less than in a species-poor system (Figure 3D), and invasion success is correspondingly reduced (Figure 3E,F). When a species-rich community is invaded, there are fewer resources for the newly arriving individuals and it is more difficult for the invading species to establish. This process is equivalent to the process proposed by Doak et al. (1998) where fluctuations in the abundance of a single species are averaged out as species richness increases. Stachowicz et al. (2002) demonstrated that this mechanism reduced invasions in some subtidal marine invertebrate communities (but see Dunstan and Johnson 2006). Where mortality is dependent on resource utilisation, invasions are also more successful in communities where resource consumption is variable between species (Figure 3A), as occurs in communities where mortality is independent of resource utilisation. However, unlike the

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Figure 2 The correlation (Kendall coefficient) between (A) species richness and invasion success, and (B) resource variability and invasion success, as a function of the maximum resource units consumed. Maximum resource units consumed defines the potential maximum number of resource units that can be consumed by an individual. Simulations with mortality dependent on resource utilisation are closed circles while simulations with mortality independent of resource utilisation are open triangles.

case where mortality is independent of resource utilisation, it is the species-poor communities that have the lowest resource variability and the least successful invasions (Figure 3B,C). This is because species in speciespoor communities are more likely to be large consumers, and species consuming a relatively large amount of available resources have lower mortality rates. These species retain a larger proportion of the available resource, and species-richness declines as those species that consume less, but die fastest, become locally extinct. Model 2: effects of connectivity and network topology on invasion In all spatial models, the likelihood of invasion at first increased with increasing connectivity (i.e., increasing

effects of competition) before levelling off (Figure 4). This qualitative pattern was evident irrespective of whether non-standoff interactions permitted reversals or not. In marked contrast, in the non-spatial, but otherwise equivalent models, the likelihood of invasion decreased with increasing connectivity of the community, consistent with the conventional view (e.g., Case 1990, 1991). Also, at higher levels of connectivity, the likelihood of invasion was notably higher in the spatial than in equivalent nonspatial systems (Figure 4). These patterns are also interpreted most readily in the context of resource availability. When the number of standoffs in a network is high, there is relatively little likelihood of a given patch of one species residing on the landscape adjacent to another species that it can overgrow. In contrast, if connectivity is high (i.e., the number of standoffs is low), there is a high likelihood of an individual being adjacent to another species that can be overgrown, or being overgrown by a neighbour. Thus, depending on the local arrangement of species, a given individual patch is more likely to overgrow other species or be overgrown by others than in communities where connectivity is lower. This tendency is exacerbated because the self-organised patch size in systems with high connectivity is, on average, larger than that in models with a high proportion of standoffs. If connectivity and richness of the recipient community is kept constant, but the topology of the interaction network is allowed to vary, the resultant level of invasion can be highly variable, irrespective of the level of recruitment of the alien, or whether reversals occur in pairwise interactions that are not standoffs (Figure 5). While in most cases the alien species occupies )0% but F10% of the landscape after 500 time steps, there are many occasions (ca. 10% of random topologies) in which the alien species fails to establish at all, despite repeated reintroductions of propagules. In ca. 25% of cases, the alien invaded to occupy )10% of the landscape at 500 time steps, and in about 2% of cases the invader spread to occupy )50% of the landscape over the same time period. This example shows clearly that, for a given fixed connectivity and richness of the recipient community, the tendency to resist invasion is highly variable depending on network topology. Note that repeated runs of the same network topology give virtually identical outcomes.

Discussion: invasions into marine algal communities General The key finding in the behaviour of both models presented here is that the likelihood of invasion increases with variability in resource availability in the recipient community, and there is no consistent relationship between invasion success and species richness. It is reasonable to expect that richness can be either positively or negatively correlated with invasions at a local scale depending on the nature of component species, the dynamics among species, and external forces (Davis et al. 2000, Dunstan and Johnson 2004, 2006). Both patterns arise in both models although they have very different underlying structures and assumptions. Variation w47x

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Figure 3 The relationships between invasion success, species richness and resource variability for simulations with dependent mortality (A)–(C) and independent mortality (D)–(F). Maximum resource utilisation (r) equals 15.

Figure 4 The effect of connectivity on levels of invasion in a model 20-species system. Connectivity is defined as the percentage of 100% standoffs in the matrix of possible pairwise interactions among species. The probability of potential invaders attaining )2% or )5% cover after 200 time steps is estimated from Monte Carlo sets of 1000 runs. Solid circles and squares show results for spatial models considering alien species that exceed 2% and 5% cover, respectively. Solid triangles and crosses are results for equivalent non-spatial, but otherwise identical models. Dotted lines are results of spatial models in which interactions that are not 100% standoffs allow reversals in outcomes (indicated by ‘‘with reversals’’). Non-spatial models show increased resistance to invasion with increasing connectivity, while the spatial equivalents demonstrate decrease in resistance to invasion with increasing connectivity, irrespective of whether non-standoff interactions allow reversals or not.

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in resource availability is influenced not only by external environmental forcings, but can be strongly influenced by the population dynamics of component species (i.e., characteristics of recruitment, growth and mortality), patterns of interspecific interactions among species, and structure in the spatial arrangement of individuals. Notably, it is not only system-wide patterns of interspecific relationships (e.g., connectivity) that are important but, for a given connectivity (and richness), the precise topology of the interaction network has a large influence on the susceptibility of the system to invasion (Figure 5). At least in part, differences in demography and outcomes of interspecific interactions reflect interspecific differences in the capacity of species to utilise resources. On average, there will be more opportunity for successful invasions when there are large fluctuations in resource availability, however, if resources are limiting, then even small fluctuations may release sufficient resource to enable a successful invasion. These results are consistent with the behaviours of the real ecological systems reported by Davis and Pelsor (2001) and Dunstan and Johnson (2004). If we exchanged variation in time (our models) with variation of resources in space, our results would be similar to those of Davies et al. (2005). They suggested (op. cit.) that spatial variability in resource availability was responsible for the positive correlation between richness of aliens and natives at large scales, but at small scales (within a patch) natives and aliens were negatively correlated. Clearly, our results are limited to a single patch but we have shown that both patterns of positive and negative correlations between native and alien species are possible at small scales, and that this pattern is generated as a function of temporal resource variability. By consid-

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Figure 5 Levels of invasion of a community of 10 native species defined by the frequency of cover attained by an alien species 500 time steps after the first arrival of propagules. Results are given for communities where (A) all non-standoff interactions have a unanimous winner and recruitment is high; (B) outcomes of non-standoff interactions are characterised by reversals and recruitment is high; and (C) all non-standoff interactions have a unanimous winner and recruitment is low. In all cases, connectivity is held constant, with 45% of interactions in the network as 100% standoffs. Frequencies are determined from Monte Carlo sets of 1000 runs. On the y-axis, ‘‘0 (DNE)’’s‘‘did not establish’’, indicating runs where the alien species was absent from the landscape after 500 time steps; while non-zero values indicate the upper limit of a class covering 10 percentage points, e.g., ‘‘80’’ indicates runs where the cover of the invader after 500 timesteps was )70% and F80%.

ering both our work and that of Davies et al. (2005) a more general theory of invasions should be possible, utilising information on species identities, resource supply and utilisation and propagule supply. It is worth noting that in the spatially explicit model (Model 2) we could devise run sets that showed, in general, that invasion success decreased with increasing richness of the recipient community, in line with the established view. In these runs mortality was independent of patterns of resource use, just as in the configuration of Model 1 that yielded a similar result. However, even in these circumstances, a single set of runs showed either significant positive or negative relationships between invasion success and richness depending on the magnitude of invasion that was considered (results not presented in detail here). For low-level invasions

where the invading alien species manages to cover at least 2% of the landscape 200 time steps after the first arrival of propagules, the relationship between invasion success and richness is positive. If a higher threshold of invasion is considered (the alien covers G5% of the landscape after 200 time steps), the relationship between invasion success and richness of the recipient community is negative. We highlight this point to emphasise that inconsistencies in the relationship between invasion success and richness of the recipient community arise at several scales of resolution of the system. In our model communities there is a single resource (i.e., space) that all species must utilise to survive. A single species, without any other species present, would utilise all the resource available until it ultimately utilises the maximum level of resource available. In terms of classic niche theory, all the species share the same niche as the overlap in resource requirements is 100%. However, multiple species can co-exist, despite utilising the same resource and ‘‘competing’’ as recruits for limited resource. The same principle has been demonstrated in spatial models of particular natural systems that proved to be good predictors of the global community dynamics of the real system (e.g., Dunstan and Johnson 2005). Both of the models presented here emphasise population stochasticity and resource utilisation either with (Model 2) or without (Model 1) competition between adults for space. Where competition for space is included, the network topology is not strictly hierarchical and, furthermore, some run sets allowed reversals in outcomes whereby at some points in space/time, a species X competitively displaces Y, while in others Y displaces X. Reversals of this kind arise commonly in benthic marine systems (e.g., see introduction of Johnson and Seinen 2002, Dunstan and Johnson 2005). In contrast, the models of Levins and Macarthur (and, by extension, Case 1990, 1991) emphasise a deterministic framework with a hierarchical competition structure. By relaxing assumptions about resource partitioning we have been able to demonstrate that both positive and negative correlations between invasion success and species richness are possible. Notably, the models presented here are generalised schematics designed to elucidate patterns that may occur in seaweed (and other) communities. We have not attempted to present a general theory of invasion dynamics, but have added to a developing list of considerations that is moving towards a more generalised and integrated viewpoint. Other factors not included in our models will almost certainly affect the outcomes of invasions (e.g., propagule pressure, predation, pathogens). It remains to thoroughly test those principles deemed as general, and to identify which particular combinations of factors act to promote or resist invasions. Seaweed communities Is there any reason to expect that seaweed communities may demonstrate any similarities with our models? In a broad sense, seaweeds share many ecological characteristics with, and are subject to a suite of evolutionary pressures similar to those of any other group of organisms which co-occur in a given habitat. Seaweeds, like other species, require a wide variety of resources (e.g., w49x

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trace and other elements, vitamins, nutrients, light, space) for survival, growth and reproduction (e.g., Lobban and Harrison 1997), and it is usual that only a relatively small number of these resource dimensions (typically nitrogen, light and space) are limiting. Like other organisms, seaweeds have evolved a plethora of strategies to optimise their survival, growth and reproduction in the face of variability in the availability of limiting resources. These strategies realise tradeoffs in competitive ability, recruitment potential and mortality, reflecting different underlying capacities to utilise resources and different strategies (in an evolutionary sense) in allocating them. We therefore expect that the characteristics of the established individuals at a site will influence the patterns of resource variability at that site, and that the patterns of resource variability will be an important determinant in the invasion success for a given density of alien propagules. In these generic properties, seaweed communities are unlikely to differ from many other kinds of communities. Seaweed communities will typically experience fluctuations in the availability of space, light are nitrogen as limiting factors. The availabilities of space and light are usually closely linked, and play a key role in shaping the mosaic of patches that so often arises in seaweed systems (e.g., Dayton et al. 1984, Sousa 1984, examples in Lobban and Harrison 1997), and also influences the growth of clonal species (Santelices 2004). The role of environmental stresses or disturbances, grazing and other ecological interactions in creating space and facilitating establishment of alien seaweeds is well documented (e.g., Valentine et al. 2007). However, depending on the species, ongoing persistence of an introduced seaweed may require mechanisms to prevent natives from eventually regaining domination of space (e.g., Valentine et al. 2007). Large-scale variation in nutrient availability often has a strong seasonal component, and so interactions in the availability of nitrogen and light may arise. Seaweeds manifest a variety of strategies in responding to variation in these resources, ranging from storage of nitrogen to be used for growth in times when nutrients are not available or light levels are more suitable (e.g., Gagne´ et al. 1982, Zimmermann and Kremer 1986), to a capacity for highly efficient uptake of nitrogen at low concentrations, and then using this nutrient directly for growth rather than storage (e.g., Probyn and Chapman 1983). For species such as the kelp Saccharina longicruris (de la Pylaie) Kuntze which realises a prodigious reproductive output (Chapman 1984) and uses stored nitrogen to achieve high growth rates under optimal light conditions when available nitrogen in the water column is low (Gagne´ et al. 1982), it is perhaps not surprising that individuals are able to rapidly capitalise on breaches in the canopy that locally release resources of light and space (Johnson and Mann 1988). It is worth noting that by this mechanism S. longicruris appeared able to maintain its dominance of the system and that opportunities for invasion were remote (Johnson and Mann 1988). However, this dynamic was changed markedly with invasion of an alien bryozoan w(Membranipora membranacea (L.)x which overgrows the kelp blades, shading the photosynthetic w50x

tissue and preventing high recruitment rates and rapid growth. This then enabled Codium fragile subsp. tomentosoides (van Goor) P.C. Silva to invade when light (and space) resources became available, and to form dense patches that effectively excluded re-establishment of the kelp (Chapman et al. 2002; see also Valentine et al. 2007). This example highlights a complex suite of mechanisms that realised sufficient variation in resource availability to enable a significant invasion event. Small-scale variation in nutrient availability can also arise. Species able to efficiently and rapidly uptake available nutrients may locally limit the availability of the resource to newly arriving propagules (e.g., Korb and Gerard 2000, Rees 2003, Phillips and Hurd 2003). This mechanism may be particularly prevalent at lower latitudes where nutrients may be limiting (Steneck et al. 2002). Spatial and temporal patchiness in the distribution of species able to commandeer nutrients in this way may generate small-scale patchiness in resource availability, influencing the survival of both juveniles (e.g., Hernandez-Carmona et al. 2001) and adults (Gerard 1982, 1997), and consequently increase the risk of invasion. Finally, we suggest that seaweed communities are also likely to demonstrate properties that our models suggest will realise positive relationships between invasion rates and species richness of the recipient community. The most obvious is size-specific mortality, particularly as applied to patches. The weight of evidence suggests that larger and denser patches of a particular species are better able to locally monopolise resources, with less vulnerability to ‘‘mortality’’, than smaller patches (e.g., Dayton et al. 1984, Sousa 1984, Chapman 1984, Chapman and Johnson 1990).

Conclusion The success or failure of an invasion will depend on at least three aspects of the recipient community: (1) the amount of resource available across the resource spectrum, (2) the pattern of utilisation of resources by the recipient community, and (3) the strength and pattern of interspecific competition and spatial organisation of the community. The resource spectrum is defined as the magnitude of the resource dimensions (e.g., space, light and nutrients) and will vary in space and time. Species will respond to this variation and during periods of excess resource availability, demonstrate differential capacity to utilise these resources in optimising their growth and reproduction. The particular dynamics of any community will release resources in space and time, and it is through these windows of opportunity that alien species can initially recruit to the community. Whether the invaders survive will be determined by continuity in resource supply and competition for those resources with other species in the recipient community. The critical point is that the ability of a community to resist the establishment of new populations comes not from species richness or diversity but from the ability of the suite of species that is present to utilise the maximum amount of resources over the long term, with the smallest

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amount of variability. We see no reason to suppose that seaweed communities are an exception to this principle.

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MacArthur, R.H. 1955. Fluctuations of animal populations and a measure of community stability. Ecology 36: 533–536. MacArthur, R. and R. Levins. 1967. The limiting similarity, convergence, and divergence of coexisting species. Am. Nat. 101: 377–385. McGrady-Steed, J., P.M. Harris and P.J. Morin. 1997. Biodiversity regulates ecosystem predicability. Nature 390: 162–165. May, R.M. 1972. Will a large complex system be stable? Nature 238: 413–414. Meiners, S.J., M.L. Cadenasso and S.T.A. Picket. 2004. Beyond biodiversity: individualistic controls of invasion in a selfassembled community. Ecol. Lett. 7: 121–126. Molofsky, J., R. Durrett, J. Dushoff, D. Griffeath and S. Levin. 1999. Local frequency dependence and global coexistence. Theor. Popul. Biol. 55: 270–282. Phillips, J.C. and C.L. Hurd. 2003. Nitrogen ecophysiology of intertidal seaweeds from New Zealand: N uptake, storage and utilisation in relation to shore position and season. Mar. Ecol. Prog. Ser. 264: 31–48. Probyn, T.A. and A.R.O. Chapman. 1983. Summer growth of Chordaria flagelliformis (O.F. Muell.) C. Ag.: physiological strategies in a nutrient stressed environment. J. Exp. Mar. Biol. Ecol. 73: 243–271. Rees, T.A.V. 2003. Safety factors and nutrient uptake by seaweeds. Mar. Ecol. Prog. Ser. 263: 29–42. Robinson, J.V. and W.D. Valentine. 1979. The concepts of elasticity, invulnerability and invadability. J. Theor. Biol. 81: 91–104. Robinson, G.R., J.F. Quinn and M.L. Stanton. 1995 Invasibility of experimental habitat islands in a California winter annual grassland. Ecology 76: 786–794. Roughgarden, J. 1974. Species packing and the competition function with illustrations from coral reef fish. Theor. Popul. Biol. 5: 163–186. Ruiz, G.M., P. Fofonoff, J.T. Carlton, M.J. Wonham and A.H. Hines. 2000. Invasions of coastal marine communities in North America: apparent patterns, processes and biases. Annu. Rev. Ecol. Syst. 31: 481–531. Santelices, B., 2004. A comparison of ecological responses among aclonal (unitary), clonal and coalescing macroalgae. J. Exp. Mar. Biol. Ecol. 300: 31–64. Shea, K. and P. Chesson. 2002. Community theory as a framework for biological invasions. Trends. Ecol. Evol. 17: 170–176. Shigesada, N., K. Kawasaki and E. Termoto. 1984. The effects of interference competition on stability, structure and invasion of a multi-species system. J. Math. Biol. 21: 97–133.

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Botanica Marina 50 (2007): 373–384

2007 by Walter de Gruyter • Berlin • New York. DOI 10.1515/BOT.2007.042

Review

Methods for identifying and tracking seaweed invasions

Alexandre Meinesz Laboratoire Environnement Marin littoral, Universite´ de Nice-Sophia Antipolis, Parc Valrose, 06100 Nice cedex 2, France, e-mail:[email protected]

Abstract Tracking macroalgal invasions relies on a variety of approaches and techniques, including random sampling, active tracking near sources of introduction, and identification of alien species and their vectors of introduction using both classical means (based on taxonomical and biogeographical knowledge) and molecular tools. To identify the invaded ecosystems, and to describe the rate of spread and invasion dynamics, other approaches must be used, including mapping techniques and public awareness campaigns. In these endeavours, standardisation of geographical and density data is important. The tracking of alien species before eradication or control measures are instigated requires special cartographic techniques. This paper provides a general overview of these different approaches and the specific strategies adapted to the biological and ecological characteristics of particular species. Keywords: introduced species; invasive seaweeds; mapping; monitoring; tracking.

Introduction In recent decades, introductions of alien benthic algal species have increased dramatically in the oceans and in landlocked seas (Ribera and Boudouresque 1995, Ribera Siguan 2003). Rapid detection of a species new to a geographic region can be crucial for eradication or control operations. Rapid detection is also important for identifying sources of introduction so that transport vectors can be restricted. Furthermore, details on the invasive capability of an introduced seaweed indicated by rates of spread are needed to assess potential impacts on native ecosystems. Based on numerous case studies, clear directions are now emerging for optimal tracking strategies to (a) detect alien species as early as possible, (b) monitor their spread, and (c) assess their invasive potential.

Detecting alien invaders Passive and random detection The discovery of an alien species is often a matter of chance collection. Alien species occur perchance in

sample collections, and are identified through the perspicacity of a phycologist. Detection is, thus, more frequent in areas that are most often visited by specialists, for instance in the vicinity of marine research stations or in marine protected areas. This passive detection is aided by laypersons using the oceans for a living or for recreation (e.g., sailors, fishers, and divers), who often report any anomaly detected in the marine environment. When an introduced species has a size or shape very different from those of native species, they may be detected more quickly by amateurs than by professional biologists. This was the case for the green alga, Caulerpa taxifolia (Vahl) C. Agardh, initially identified in the Mediterranean Sea by SCUBA divers when it covered barely 1 m2 of sea floor (Meinesz and Hesse 1991). However, early reporting is not always the case, even for large introduced and invasive species, where the first sighting is often reported only several years after introduction. Farnham (Farnham et al. 1973, Farnham 1974) first detected the brown seaweed Sargassum muticum (Yendo) Fensholt in 1973 on the Isle of Wight (south coast of Britain), two years before its discovery on the south side of the English Channel at SaintVaast-La-Hougue (France) (Cosson et al. 1977). The alga had probably been imported into Brittany or Normandy in the late 1960s from Japan on spat of the oyster Crassostrea gigas (Thunberg) (Gruet et al. 1976), but was not detected for many years in northern France. Similarly, in 2000, a SCUBA diving scientist working on marine phanerogams in California (San Diego) discovered Caulerpa taxifolia by chance. This was immediately identified as the well known invasive Mediterranean strain (Jousson et al. 2000, Woodfield and Merkel 2004). At the time of discovery in California, more than 1000 m2 of the lagoon were already covered by this species. An invasion of Caulerpa racemosa (Forsska˚l) J. Agardh var. cylindracea (Sonder) Verlaque, Huisman et Boudouresque originating from Australia was discovered in Libya in 1990, thanks to the discerning eye of a phycologist sampling off Tripoli harbour (Nizamuddin 1991). The real date and site of introduction to the Mediterranean Sea are still not known. Small algae whose invasive character is unspectacular, or algae that are difficult to distinguish from native species (cryptic invasions) often go unnoticed for long periods before detection and reporting. This is the case for a small member of the Ceramiales originating in Australia including Acrothamnion preissii (Sonder) E.M. Wollaston, which has invaded the Mediterranean Sea. Since this species is generally identifiable only with light microscopy, observations in the Mediterranean Sea have been spread widely in time and space, dependent on the collection and meticulous analysis of samples by specialists. Even though the site of first sighting for A. preissii was Leghorn in Italy (Cinelli and Sartoni 1969), the real w53x

374 A. Meinesz: Tracking invasive seaweeds

site of first introduction cannot be known with certainty. Similarly, for Codium fragile (Suringar) Hariot ssp. tomentosoides (van Goor) P.C. Silva, a large green alga, long confused with Codium tomentosum Stackhouse, the site and date of first introduction into the North and Mediterranean Seas are unknown and debatable. With so much uncertainty, it is clearly important to effectively monitor the arrival of introduced species, and this will depend on maintaining facilities with specialists in the taxonomy of algae and regional biodiversity. Investing in ‘‘outreach’’ activity, such as training members of the interested lay public about the shape and colour of native algae, and the potential impact of alien species, is also a means to ensure early detection of new alien species.

ranean coast, where there is a major oyster farming facility. Oyster spat has long been imported to Thau from Japan (from the Seto Inland Sea). After the discovery of several introduced species of Japanese origin in the early 1970s, phycologists organised tracking campaigns for new introduced species, and to date have discovered 45 introduced macroalgae, the majority of them (43 taxa) most likely originating from the Pacific region and including some large phaeophytes we.g., Saccharina japonica (J.E. Areschoug) C. Lane, C. Mayes, Druehl et G.W. Saunders, Sargassum muticum and Undaria pinnatifida (Harvey) Suringarx (Verlaque 1994, 2001).

Active detection

Various criteria to recognise and designate a species as alien have been proposed, including:

Active detection is a dynamic process. When an invasive introduced algal species is reported, it is then sought in neighbouring regions or countries. The newly discovered species, once reported or given media exposure, usually then arouses interest in the local and regional scientific community. Once the introduced alga has been properly described and recognised, it then becomes more easily identifiable, and an active search can be organised. A dynamic of detection is, thus, established, especially if the species is known to be potentially invasive with potentially deleterious impacts on the invaded environment. Numerous examples of this kind of ‘‘detection dynamics’’ may be cited, e.g., the introductions of Sargassum muticum (see the compilation of 300 publications concerning this invader by Critchley et al. 1990) and of the two invasive Caulerpa species (C. taxifolia and C. racemosa var. cylindracea). A revealing example is the tracking of Caulerpa taxifolia in some areas of the Mediterranean Sea. In Tunisia, this species had not yet been reported in the 1990s. A public awareness campaign about its possible arrival in Tunisia was organised in 1998 by local phycologists based on the distribution of well-illustrated brochures (Langar et al. 1998). Thanks to this information campaign, a fisherman reported the presence of the alga to an oceanographic centre in 2000 (Langar et al. 2000). Detection effort should also focus on the mode of introduction. For example, a site where several introduced species have been reported can become a focus of sustained interest for phycologists. In this case, it is the mode of introduction that justifies the active detection of introduced species. Certain sites are particularly favourable for the introduction of species, e.g., shipping harbours where deballasting takes place, or aquaculture facilities where juveniles of shellfish species are imported. Areas within the vicinity of public aquaria built on the coast may also be actively investigated. In fact, two cases of the introduction of Caulerpa taxifolia, in the Mediterranean Sea and in Japan, were reported in the vicinity of aquaria (Meinesz and Hesse 1991, Komatsu et al. 2003). Around potential sites of elevated risk of introduction, more intensive observation means that introduced species may be discovered more easily. A spectacular case is that of the Thau lagoon on the French Mediterw54x

Determining the alien character of a species, its origin and vectors of introduction

• it should be new to the area studied (the original range should be clearly distinct); • there should be identifiable invasion dynamics (temporal and geographical); and • sources of introduction in the area of first sighting should be identifiable. In the vast majority of cases of algal introduction, the designation as alien (either through range extension or anthropogenically mediated introduction) is not contested. Classical techniques of identification and biogeographical knowledge of the range of the natural populations of the introduced alga are usually sufficient to establish status as an alien. Similarly, in the vast majority of cases, the suspected vector has not been contested. However, in several cases outlined below, genetic tools have proven very useful in establishing alien invader status. The molecular tools to track and identify alien invasive species are described elsewhere in this issue (Booth et al. 2007). Codium fragile ssp. tomentosoides This sub-species is a worldwide invader. In the management and control of such alien species, it is important to determine the frequency with which it was introduced into the different invaded areas and the subsequent pattern of spread. For C. fragile ssp. tomentosoides, several authors have demonstrated that the invasive strain is monophyletic and genetically homogeneous. This suggests that there are only a few (at least two) separate introductions from the native populations localised in the North Pacific Ocean (Goff et al. 1992, Coleman 1996, 1997, Provan et al. 2005). Caulerpa taxifolia This green alga was first discovered as an invader in 1984 beneath a public aquarium in Monaco where it was cultivated. It had never been observed previously in the Mediterranean Sea (Meinesz and Hesse 1991). Its status as an introduced species and mode of introduction seemed obvious at the time, but two other hypotheses were formulated by scientists based in Monaco (Chisholm et al. 1995). These authors raised the possibility that the alga may have been carried from the Red Sea by currents (more than 2500 km from Monaco), or was native to the Mediterranean Sea as a

A. Meinesz: Tracking invasive seaweeds 375

‘‘metamorphosis’’ of Caulerpa mexicana Sonder ex Ku¨tzing, known on the eastern shores of the Mediterranean Sea (nearly 3000 km from Monaco) since 1945. To settle the issue, Jousson et al. (1998) and Olsen et al. (1998) confirmed by genetic analysis that C. taxifolia and C. mexicana are two distinct species, as is readily observable from their morphological differences (Meinesz et al. 1994, Meinesz and Boudouresque 1996). Genetic analysis of samples of C. taxifolia from the Red Sea, other neighbouring oceans (Atlantic Ocean, Indian Ocean) and from aquaria, including those of Monaco and Stuttgart (which was the source of the strain cultivated in the Monaco aquaria since the 1980s), demonstrated that the invasive strain in the Mediterranean Sea is identical to strains cultivated in the Monaco Aquarium, but different from those originating in the Red Sea and the oceans neighbouring the Mediterranean Sea (Jousson et al. 1998). Given the history of Caulerpa taxifolia as a (controversial) invasive species, its arrival in California (San Diego and Los Angeles) immediately prompted genetic analysis (Jousson et al. 2000). The genus Caulerpa was unknown on the Pacific coasts of the USA and C. taxifolia had never been reported off temperate and tropical coasts of the Pacific Ocean. Genetic knowledge accumulated during the debate about the introduction of C. taxifolia in the Mediterranean Sea made it possible to quickly identify the invasive strain in California as the one that had developed in the Monaco Aquarium and in the Mediterranean Sea (Jousson et al. 2000). This confirmation led the local authorities to attempt eradication of C. taxifolia in San Diego and Los Angeles and to ban trade in Caulerpa species for aquaria. Knowing the temperature tolerances of the invasive strain of Caulerpa taxifolia in the Mediterranean Sea (Komatsu et al. 1997) helped to identify a population of this species in subtropical Australian waters (Meinesz 2001, p. 307). It was then established quickly, on the basis of various molecular analyses, that the invasive (Mediterranean, Japan and California) and aquarium strain did indeed originate in Moreton Bay, adjacent to the city of Brisbane, Australia (Jousson et al. 2000, Meusnier et al. 2001, 2002, 2004, Wiedenmann et al. 2001, Fama` et al. 2002a,b, Schaffelke et al. 2002, Komatsu et al. 2003). Finally, the appearance of new range extensions of Caulerpa taxifolia into southern Australia (Sydney, Adelaide) (Millar 2001) led to the realisation that transport of vegetative fragments by boat traffic from Moreton Bay was an important vector of introduction (Schaffelke et al. 2002). Caulerpa racemosa var. cylindracea This species was first recognised as an invader in the Mediterranean Sea off Tripoli harbour, Libya (Nizamuddin 1991). However, there is no certainty regarding the date and location of first introduction, and the vector is unknown. The problem posed by this species, which rapidly invaded the coasts of most Mediterranean countries, and even the Canary Islands in the Atlantic Ocean (Verlaque et al. 2003a,b, Piazzi et al. 2005), was to determine whether it was derived by hybridisation between several different

varieties of C. racemosa, known since 1925 on the southern coasts of the Mediterranean Sea (Hamel 1926), or whether it came from the Red Sea or from other oceans. Several contradictory genetic analyses were published, but it was finally determined that the invasive strains were genetically close to a population developing in a temperate region of Western Australia near Perth (Fama` et al. 2000, Verlaque et al. 2000, 2003a, Durand et al. 2002). Grateloupia species Another example of the utilisation of genetic tools to elucidate the origin of an introduced and invasive alga is that of a rhodophyte identified initially as Grateloupia doryphora (Montagne) Howe, originally described from Pacific South America. This species was reported as introduced and invasive since 1973 in western Europe, since 1997 in eastern North America and since 1982 in the Mediterranean Sea. However, genetic analysis demonstrated clearly that this Grateloupia was misidentified, and that it corresponds to G. turuturu Yamada, originally described from Japan (Gavio and Fredericq 2002, Verlaque et al. 2005). Undaria pinnatifida Debate on this species is often focussed on the vector. It has invaded coastlines worldwide, being first observed outside its native Asian range in the Mediterranean Sea in 1971 and later in the eastern North Atlantic Ocean in Brittany, with subsequent establishment in the United Kingdom, Spain and the Netherlands. This alga has also established in New Zealand, Australia, Argentina and the American north Pacific coast (Silva et al. 2002, Valentine and Johnson 2003). It is often queried whether it arrived in ship ballast water, from hull fouling, or from aquaculture facilities where this edible alga is cultivated (under the name of wakame). Genealogical analyses point to aquaculture as a major vector of introduction and spread in Europe, but implicate maritime traffic in promoting recurrent migration events from the native range to Australasia (Uwai et al. 2005, Voisin et al. 2005). These analyses also make it possible to estimate the number of sources of introduction or distinct origins within an affected area. Acanthophora spicifera (Vahl) Børgesen Genetic tools were used to investigate the geographic origin of Hawaiian populations of this alga believed to have been introduced 50 years ago by way of a biofouled barge from Guam. Subsequently it reached most of the islands of the archipelago and became the most invasive alien macroalga on coral reefs throughout the main Hawaiian Islands. DNA sequencing revealed no variation for the two markers, even when collections from other areas of the Pacific Ocean and Australia were included. This genetical analysis can neither confirm the supposed origin nor secondary introductions in Hawaii. In contrast, ISSR analyses revealed highly structured Hawaiian populations of A. spicifera with a substantial range of both within- and among-population variation, with individual plants forming discrete clusters corresponding to geographic locality (Sherwood 2005, O’Doherty 2007). w55x

376 A. Meinesz: Tracking invasive seaweeds

Asparagopsis armata Harvey and A. taxiformis (Delile) Trevisan Genetic tools were used to differentiate these introduced species in the Mediterranean Sea wthe tetrasporophyte stages (‘‘Falkenbergia’’) are difficult to distinguishx and to describe genetic variability within native and introduced strains (Andreakis et al. 2004, 2005, Nı´ Chuala´in et al. 2004, Procaccini et al. 2005). For some introduced species, genetic analysis has not been able to resolve disagreements regarding sites and modes of introduction. A certain number of species introduced into the Mediteranean Sea and referred to as Lessepsian (coming from the Red Sea via the Suez canal, built by Ferdinand de Lesseps in 1855–69) by Por (1978), call for genetic investigation since they may represent relics of an algal flora originating in the tropical Atlantic Ocean. This is the case for Caulerpa mexicana Sonder ex Ku¨tzing, C. scalpelliformis (Brown ex Turner) C. Agardh and certain varieties of C. racemosa developing in the southern Mediterranean Sea. Similarly, genetic tools would provide a basis for testing the opinions of Cormaci et al. (2004) who contested the status of 14 taxa considered by other authors as introductions into the Mediterranean Sea. In summary, these different cases demonstrate that, where there is debate, genetic analyses usually support initial observations that use classical morphological and biogeographical criteria indicating that the species is indeed introduced. Genetic analysis may inform us on the status of a species of interest (its identity and origin), may confirm or identify the mode of introduction, assess the genetic diversity of introduced populations, and thus, enable estimates of the number of secondary independent introductions. Finally, genetic studies of alien algae can elucidate some evolutionary and genetic consequences for native and invading species (see Booth et al. 2007).

duction occurs was raised. The reproductive cycle of the genus Caulerpa involves production of gametes. All the nuclei are associated with chloroplasts and when the gametes are released the thallus, emptied of its content, disintegrates (this is a process of holocarpy). Species are dioecious or monoecious. C. taxifolia was known to be monoecious (Goldstein and Morall 1970). However, observations of the reproduction pattern of the strain introduced into the Mediterranean Sea showed that it is exclusively dioecious, producing only male gametes (without eyespots) (Meinesz et al. 1994, Zuljevic and Antolic 2000). This characteristic had already been observed in Caulerpa prolifera (Forsska˚l) J.V. Lamouroux on the French Mediterranean coast (Meinesz 1979). Since then, numerous phases of gamete production have been observed, without any female gametes being detected. Thus, all the signs are that sexual reproduction does not occur. Reproduction appears to be solely vegetative. If we consider that the introduced strain has a single origin, then all the populations of C. taxifolia that have developed in the Mediterranean Sea are clonal. Genetic analyses would appear to support this notion, since there is very little genetic differentiation among populations. In the Mediterranean Sea, the only one that appears to differ from the others is the population of C. taxifolia that has developed in Tunisia, but the only sample examined was in poor condition (Jousson et al. 1998). On this basis we may predict a slow rate of dispersal attributable solely to ‘‘secondary’’ anthropogenic dispersal, e.g., by boats with fragments attached to their anchors. In this case, tracking C. taxifolia can be undertaken in the immediate vicinity of already identified sites to assess natural rates of vegetative spread, and searches for the alga can be undertaken within a wider radius where human activities are likely to have transported fragments, such as to harbours and fishing or mooring areas.

Tracking alien algae to determine their spread and invasiveness

Caulerpa racemosa For this species, two modes of reproduction that might explain its very rapid spread have been described. The alga reproduces sexually (Panayotidis and Zuljevic 2001), and there is a highly efficient system of vegetative reproduction in which the spherical branchlets of the thallus can break off and form vegetative propagules. This production of specialised propagules more easily disseminated than fragments of thallus increases the colonisation potential of an affected site (Renoncourt and Meinesz 2002). These mechanisms can account for the very rapid spread of this alga at some sites. Tracking of C. racemosa should reveal a more rapid progression of cover over a more extensive area than has been demonstrated for C. taxifolia.

As soon as an alien species is detected and reported, its spread and invasion dynamics must be determined. In the first instance, this requires a good knowledge of patterns of reproduction and habitat use of the introduced alga, and then the application of a variety of monitoring techniques, depending on the biological and ecological characteristics of the species. Importance of knowledge of the biology of the alga Basic biological knowledge is important to estimate the potential for natural or anthropogenic dissemination. For example, knowledge of the production pattern of spores or zygotes or of vegetative fragments or propagules that may be spread by currents or by man, and a clear understanding of the phenology, is essential to define the potential for spread. A variety of case studies underline the importance of this kind of knowledge. Caulerpa taxifolia At the time of its initial discovery at Monaco and the first studies of this species (Meinesz and Hesse 1991), the question of whether sexual reprow56x

Other invasive introduced algae Other invasive introduced algae have a variety of reproductive modes that influence invasion dynamics. Codium fragile var. tomentosum reproduces parthenogenetically (Feldmann 1956, Chapman 1999), which must have favoured its dissemination. In contrast, some Rhodophyta introduced into the Mediterranean Sea wWomersleyella setacea (Hollenberg) R.E. Norris (sPolysiphonia setacea Hollenberg) and Antithamnionella elegans (Berthold) Price et John)x

A. Meinesz: Tracking invasive seaweeds 377

appear to be sterile (Cormaci et al. 1994, Athanasiadis 1996, Rindi et al. 1999), so their dissemination occurs only through natural (currents) or anthropogenic (mainly fishing nets in the case of these two algae) dispersal of thallus fragments. Knowledge of the ecology of introduced species Understanding a species’ autecology is also essential for efficient tracking of range extension. Knowing whether a species is able to colonise a large or small range of habitats determines the choice of monitoring techniques for the alga. Similarly, knowledge of the bathymetric limits attainable by an introduced species is useful to define the ‘‘search space’’ for tracking. For the two invasive species most extensively mapped in the Mediterranean Sea, Caulerpa taxifolia and C. racemosa, it was soon established that they are able to invade most benthic habitats from the surface to 50 m depth. Conversely, Sargassum muticum and Undaria pinnatifida develop just at the level of the lowest tides with a vertical distribution range of -5 m. These algae are, thus, more easily detected from aerial photographs or teledetection images. In some cases, the isothermic limits characteristic of certain species offer a basis for predicting the limits of spread. However, this kind of prediction can also fail. This was the case for Undaria pinnatifida, the cultivation of which in Brittany (France) was permitted on the basis that low water temperatures would not allow reproduction by spores. However, estimates of the thermal threshold for this species proved inaccurate, since the alga spread rapidly in the natural environment, invading further to the North (on the southern shores of the United Kingdom) than had been anticipated. Responses to requests to cultivate other large phaeophytes have been more conservative. Cultivation of the giant kelp Macrocystis pyrifera (L.) C. Ag. was banned in Normandy (France) on the basis of evidence that the alga could complete its reproductive cycle in the temperatures of the English Channel and the North Sea. If this were the case, then its development may have become uncontrolled (Boalch 1981). Collection of cartographic data through community observation Although discovery and monitoring of small or cryptic algae usually depends on observations by professional phycologists, tracking large algae has been greatly assisted by the community at large (e.g., SCUBA divers, fishers, bathers, sailors and other lay persons). Brochures or posters to inform the public helped monitor the spread of Sargassum muticum in various countries on the coasts of the north-eastern Atlantic Ocean, the English Channel and the North Sea. Important public awareness programmes were intitiated in New Zealand to track Undaria pinnatifida. Several government agencies have distributed brochures to the public including the Department of Conservation, Ministry of Fisheries and Biosecurity of New Zealand. Brochures have generally been made available both in hard-copy and more recently on the Internet (M. Stuart personal communication).

Properly informing the public made it possible to identify most of the Mediterranean sites affected by Caulerpa taxifolia and Caulerpa racemosa, and to monitor their development. More than 300,000 brochures have been distributed since 1991 in six Mediterranean countries (Spain, France, Italy, Croatia, Tunisia, Turkey), with similar brochures written in several languages in some countries (Cottalorda et al. 2001). This effort to collect cartographical data with the assistance of an informed public was also supported by the mass media (we estimate that 2000 press articles relating to the invasion by C. taxifolia and C. racemosa have been published by the media in Mediterranean countries). Finally, numerous Internet websites have been developed to raise public awareness and collect information (almost 1500 press articles and 36 websites are listed on the website http:// www.caulerpa.org). Keeping the public informed has proven to be an effective means of collecting the information needed to estimate the overall extent of invasion in a region as vast as the Mediterranean Sea. Collection of cartographic data by systematic monitoring Monitoring marine species involves two distinct types of observation (Meinesz et al. 1981). The first allows the localisation of populations (or individuals) from information obtained through a distributed (but usually haphazard) collection of point samples (by divers or by using a grab), or linear samples (i.e., using transects), e.g., by the passage of SCUBA divers or towed or autonomous video cameras (ROV). The second involves collecting and analysing two-dimensional images of the seabed, usually at large scales, using aerial photographs, spectral responses obtained from satellite teledetection, by compact airbone spectographic imaging (CASI), or sonograms of the seafloor obtained by side scan sonar. These monitoring techniques always require confirmation by divers since the spectral responses of different communities or species are often very similar and vary with depth (in the case of spectrographic images) or type of bottom (in the case of side scan sonar sonograms). Mapping using point samples For most introduced species, plotting the dynamics of population expansion has been limited to drawing up a list of scattered sightings on relatively small scale maps (more than 1/5000). This kind of monitoring effort can produce maps of the distribution of invasive algae over large areas (e.g., Figure 1). This is the case for Sargassum muticum and Undaria pinnatifida, which have been relatively easy to map, reflecting the fact that they are readily located and identified because of their large size, tendency to form dense aggregations in a narrow depth range and, in the case of S. muticum, floating thalli that often reach the surface. For these algae that are easy to see and identify, most sightings have been recorded as occurrences (plotted as points on distribution maps) without any estimation of the areas covered or concerned (e.g., Farnham 1997, Stuart 2004). Considering the timing of reports of species occurrences at different points in space can enable production of useful maps of the rate of spread, i.e., the temporal component of the dynamics (Figure 2). w57x

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Figure 2 Example of a map of sightings of invasive species on small scales. Isocontours indicate the rate of spread of Sargassum muticum in the northeastern Atlantic Ocean, the English Channel and the North Sea (from Boudouresque 1994). Figure 1 Example of a map of sightings of invasive species on small scales. Scattered sighting (dots) of Sargassum muticum with no data on density or colony sizes (area). From Farnham (1997).

In a minority of cases, where there is high confidence that detection is soon after initial establishment, efforts to describe the density or coverage of invading algae are limited to relatively localised areas. The abundance and spread of the invasive red alga Kappaphycus sp. in Hawaii was followed by monitoring 15 sites within a single bay (using quadrats of 0.25 m2) (Conklin and Smith 2005). Mapping using linear observations For invasive species with large populations, other types of assessment enable spread to be estimated and reported on larger scale maps (e.g., scales between 1/500 and 1/5000). This has been the case for Caulerpa taxifolia, where major mapping operations have been undertaken on the coasts of all the countries affected in the Mediterranean Sea (26 cartographic publications were cited by Meinesz et al. 2001). This species occurs in large, often very dense and extensive perennial populations, albeit with marked seasonal fluctuations. Several methods have been used to describe the extent of its colonisation. The simplest and least costly strategy is to involve those members of the public who routinely observe the seabed in shallow waters, including fishers, divers, swimmers and sailors. Most of the 130 independent infested areas known to date along the French Mediterranean coasts were discovered by amateurs, but were subsequently checked and verified by scientists. To find reported colonies, or to locate the alga in the vicinity of colonised areas, free divers can be towed on a surface buoy with a GPS device to continually record position. With this method, transects 200–500 m in length (according to depth) can w58x

be readily explored to search for or localise invasive Caulerpa. For more severely affected areas (mostly in the form of multiple scattered colonies), divers seek to delimit the area over which the colonies are dispersed. For this, using a small boat towing a diver at the surface or underwater provides a suitable means. This technique is also used to actively search for algae in areas where there have been no previous sightings. In the Mediterranean Sea, visibility is often more than 10 m, which means the diver can detect colonies in a 20 m swath bisected by the transect. A GPS can be used to localise the area covered by the diver, and each population observed can be localised with precision. A diver can cover 2–5 km in 2 h while observing from the surface in this way, often in water depths to 20 m. These techniques have been used extensively in monitoring the spread of Caulerpa racemosa var. cylindracea in the Mediterranean Sea (see the 23 cartographic publications cited by Piazzi et al. 2005). For Sargassum muticum, estimates of invaded areas at a regional scale together with assessments of density have been recorded on large scale maps (1/500–1/5000), based on density data for a series of transects (Thomsen et al. 1998). Another method of linear observation consists of towing a camera mounted on a benthic sled or hovering device that maintains an approximately fixed distance off the seafloor. The coverage depends on the lens used, but given the variations in the height of the camera above the bottom, a band of only 2–4 m width can usually be monitored effectively. This method has been utilised successfully in monitoring Caulerpa taxifolia in large bays at depths of 10–40 m. Belt transects several tens of km long can be filmed in one day using this methodology (Belsher et al. 1994, 2001, 2003).

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Mapping based on two-dimensional images of the seabed When invasive species densely cover large areas in shallow waters (particularly between 0 and -10 m), they can be detected by aerial photography or by using a compact airbone spectographic imager (CASI). Since the responses of several species of the same genus are similar, interpretation usually requires diving to ‘‘ground truth’’ the remotely sensed data. CASI is rarely used in detecting underwater alliens. Its efficacy was tested on only one occasion in the case of Caulerpa in the Mediterranean Sea, in some densely covered areas. No new records have been indicated by this costly method, which was unable to delineate the total area of infestation (Jaubert et al. 2003). Comparison of the densely covered areas detected by CASI with larger areas that have been invaded and observed previously by divers in the same region is not possible. Satellite teledetection has been used to estimate the spread of Sargassum muticum (Belsher and Pommellec 1988) but, again, this technique is likely to prove useful for detecting and estimating the area occupied by thalli at relatively high densities, and it is also very costly. Standardisation of cartographic data Once monitoring techniques have been implemented, the information collected must be recorded and visualised appropriately, usually as maps or tables. To compare data from one year to another and between neighbouring countries that have invested in monitoring the invasion, standardised criteria for the geographical assessment of the invasion are required. In the vast majority of cases, sites of observation are recorded in point fashion (one point per area where occurrence has been observed) on maps, as was the case, for example, in monitoring the invasion of Sargassum muticum in Europe (Critchley et al. 1983, Farnham 1997; see Figure 1), and in monitoring the invasion of Acrothamnion preissii and Codium fragile in the Mediterranean Sea (Ferrer et al. 1994, Ribera 1994). If the boundaries of the invasion can be described with reference to time, it is possible to estimate the invasion dynamics of an alga (e.g., Sargassum muticum, Figure 2). Numerous examples of invasions by introduced species in the terrestrial environment have been described by dividing the landscape into a large number of discrete cells of similar or identical size, and indicating those cells that are colonised. The use of isocontours for invasive species based on densities is frequently utilised in conjunction with temporal wave fronts of the invasion (Hengeveld 1989). Other representations are based on presence/absence within square grids of highly variable size, each containing information on the incidence of the invasive species. Each sighting of the alien species within a square, irrespective of density, is seen as a basis for considering the entire square invaded (and coloured uniformly on maps or recored as ‘‘present’’ in tables). Other spatial conventions have been proposed such as presence/absence within broader polygons, such as states, regions or other administrative subunits (county, city, country). For the sea, this form of demarcation is less relevant, but it has been used in the Mediterranean Sea to indicate the number of countries affected by the inva-

sion of the two species of Caulerpa. For example, as regards C. taxifolia, at least six Mediterranean countries are affected (in 2001) while C. racemosa is established in 12 countries (in 2005). Since the beginning of the invasion in the Mediterranean Sea, several international teams have been involved in monitoring the spread of Caulerpa taxifolia in several countries. In order to standardise the mapping, an international pool of experts (from France, Spain, Monaco, Italy, Tunisia and Croatia) established a standardised method (Vaugelas et al. 1999, Meinesz et al. 2001) (Figure 3). These guidelines for mapping C. taxifolia define three levels of invasion and associated descriptive terms for assessing an impacted area (Figure 4): (1) Level I is the first stage of colonisation, in which one or more colonies occurring less than 100 m apart occupy a total surface area of less than 1000 m2. Under these conditions, it is relatively easy to estimate the surface area covered by the alga (referred to as the covered area) and to delimit the exact perimeter of the area of dissemination by SCUBA diving (the concerned area). (2) Level II refers to the next step of spreading. It is characterised by several colonies that occur within 250 m of each other, that cover a total area of more than 1000 m2, and that are dispersed over an area -10 ha. During this stage of the invasion, it is both time-consuming and futile to measure the covered area, as the alga grows and spreads very quickly. The best method to estimate the extent of coverage is to delimit the perimeter enclosing all colonies (i.e., the concerned area). (3) Level III is attained when dozens or hundreds of colonies of various sizes are dispersed over a surface area )10 ha, with a total covered area of more than 1000 m2. At this stage, it becomes impossible (and in any case not very useful) to map the location of each colony with any precision, or to measure the covered areas. Rather, the outer boundary, or concerned area, is estimated by identifying the positions of the peripheral colonies. By summing the colonised areas (Levels I, II and III), an evaluation of the overall status of a given region in terms of the area affected by the colonisation is obtained. For Caulerpa racemosa var. cylindracea in the Mediterranean Sea, invasions proceed so rapidly that only the linear coastal extent of colonies is a valid criterion for description. Thus, maps summarising the invasion of this alga typically show only the lineal extent of the invaded area along the shoreline (Piazzi et al. 2005, Ruitton et al. 2005) (Figure 5). Spread modelling Numerous theoretical and empirical approaches for assessing and predicting the spread of invasive terrestrial species have been proposed based on the characteristics of the invasion dynamics observed during the first years of colonisation of an environment (Hengeveld 1989, Hastings et al. 2005). These studies suggest that long range dispersal events from initial points of establishment have a strong influence on the rate of the overall range w59x

380 A. Meinesz: Tracking invasive seaweeds

Figure 3 Example of a map of sightings of invasive species on small scales. Report of occurrences with spatial extension for each spot. Invasion of Caulerpa taxifolia in the Mediterranean Sea (from Meinesz et al. 2001).

expansion. For invasive algae, a predictive spread model has been developed only for Caulerpa taxifolia (Vaugelas et al. 1997, Hill et al. 1998, 2001, 2002) at a particular site. In the marine domain, predictions are hampered by poor knowledge of both coastal currents and the spatial distribution of the subtidal habitats that an introduced species is likely to colonise.

Tracking for eradication purposes In a few rare cases, monitoring invasive algae must be carried out with a high level of precision. Very precise monitoring is required if it is necessary to completely eradicate either the whole of an introduced population or all of a particular sub-population of developing colonies within a restricted area. The eradication of Caulerpa taxifolia in southern California, near San Diego (Agua Hedionda lagoon) and near Los Angeles (Huntington Harbor), where more than 1000 m2 of C. taxifolia were discovered in 2000 (Jousson et al. 2000, Williams and Grosholz 2002, Woodfield and Merkel 2004), is a useful example. The success of eradication operations usually depends more on the precision of monitoring than on the eradication technique itself, since if a few colonies have not been detected, the whole eradication investment may be rendered worthless within a few years. Commencing in summer 2001, different methods of tracking the algae for eradication purposes were tested in the turbid waters of these areas (visibility w60x

is typically -2 m in Agua Hedionda lagoon), and included towed divers, towed cameras and laser line scan, and divers using a guide-line deployed by a small boat using differential GPS. The latter technique appeared to be the most effective in locating very small fronds of C. taxifolia (Woodfield and Merkel 2004). This method seeks to ensure 100% coverage of the afflicted area. Agua Hedionda lagoon near San Diego was covered several times by SCUBA divers along hundreds of transects 1 m apart. Permanent grids of ropes were deployed in the colonised areas in the third year of quarterly surveys. This technique made it possible to detect all colonies, which were then successfully eradicated. Off the French Mediterranean coast, the Parc National de Port-Cros authorities decided in 1994 to control the spread of Caulerpa taxifolia by eradicating the colonies as soon as they became established. Each year since 1994, between 40 and 60 SCUBA divers or skin divers combed the waters of the national park for three days to detect developing colonies. Each year, colonies were discovered and eradicated. Even if the detection of the colonies is less than exhaustive, the occurrence of the alga at Port Cros has since been highly restricted, whereas in an uncontrolled neighbouring island (Porquerolles) the invasion spread to cover several tens of hectares by 2004. At Port-Cros, the extensive area of seabed favourable for the development of C. taxifolia cannot be monitored in its entirety during each campaign. Therefore, those particular sites most favourable for the introduction (mooring areas in sheltered bays or areas where fishing

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Figure 4 Configuration of the three levels of colonisation of Caulerpa taxifolia on large scale maps. In Level III the position of the dozen or hundred main colonies involving more than 100 ha and covering much more than 1000 m2 is not presented. Example: the rade of Villefranche (France, Mediterranean Sea; 438419580N, 781895899 E), (after Vaugelas et al. 1999 and Meinesz et al. 2001).

nets are placed underwater) have been most closely inspected. Monitoring methods based on groups of 6 to 10 divers swimming in line on a pre-established course holding a rope have been used. Since the waters at PortCros are very clear (visibility often more than 10 m), the distance between divers is 5 m. A buoy attached to a weight by a rope is placed at each of the colonies detected. Their position can, thus, be easily determined with a GPS, with subsequent eradication (Cottalorda et al. 1996, Robert 1996, Robert and Gravez 1998).

Conclusion The manner of discovery of new invasions and their subsequent monitoring provides a basis for assessing the effectiveness of several different tracking protocols that have been attempted. In the majority of cases, the activity of phycologists with a good knowledge of the algal flora of their region, plus informing the lay public about potential invaders, led to relatively rapid detection of alien algal species.

Figure 5 Representation of the invasion of Caulerpa racemosa var. cylindracea on large scale maps in terms of the linear extent of coast line affected (from Ruitton et al. 2005). Colonisation Levels I–III are as outlined in Figure 4.

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To assess spread, it is necessary to match appropriate marine environmental monitoring techniques to the biology and ecology of the alga. Comparing standardised cartographic data over time and between different areas has proven extremely useful in this context. Although in a majority of cases classical identification and biogeographical knowledge provides an adequate basis for determining that particular individuals represent an alien algal species, and often to ascertain their mode of introduction, genetic tools can be particularly helpful in illuminating the likely history of a given introduction. Above all, cartographic methods make it possible to measure the extent of the spread of an invasive alga and can assist in helping to predict potential impacts. Early discovery of major invaders followed immediately by exhaustive monitoring is the exception rather than the rule, and so successful prevention of spread as a result of early detection and tracking, such as that of the eradication of Caulerpa taxifolia in California, are relatively rare.

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Botanica Marina 50 (2007): 385–396

2007 by Walter de Gruyter • Berlin • New York. DOI 10.1515/BOT.2007.043

Review

Molecular approaches to the study of invasive seaweeds

David Booth, Jim Provan and Christine A. Maggs* School of Biological Sciences, Queen’s University, Belfast BT9 7BL, Northern Ireland, e-mail: [email protected] * Corresponding author

Abstract A wide range of vectors is currently introducing a plethora of alien marine species into indigenous marine species assemblages. Over the past two decades, molecular studies of non-native seaweeds, including cryptic invaders, have successfully identified the species involved and their sources; we briefly review these studies. As yet, however, little research has been directed towards examining the genetic consequences of seaweed invasions. Here we provide an overview of seaweed invasions from a genetic perspective, focusing on invader species for which the greatest amount of information is available. We review invasion processes, and rationalize evolutionary and genetic consequences for the indigenous and invader species into two main groups: (1) changes in genepool composition, in population structure and allele frequencies; and (2) changes in genome organization at the species level through hybridization, and in individual gene expression profiles at the levels of expressed messenger RNA and the proteome (i.e., all proteins synthesized) and thus the phenotype. We draw on studies of better-known aquatic and terrestrial organisms to point the way forward in revealing the genetic consequences of seaweed invasions. We also highlight potential applications of more recent methodological and statistical approaches, such as microarray technology, assignment tests and mixed stock analysis. Keywords: alien species; cryptic species; DNA; hybridization; invasion.

Introduction Non-native species that become established into existing ecosystems and subsequently threaten biodiversity and/ or result in economic damage are referred to as invasive alien species (Shine et al. 2000). Non-natives without such effects are known as alien or non-indigenous species (Shine et al. 2000). Shipping (Carlton and Hodder 1995, Ribera and Boudouresque 1995, Minchin and Gollasch 2003), aquaculture (Farnham 1980, 1994) and the aquarium trade (Jousson et al. 1998, Olsen et al. 1998)

are vectors that introduce a multitude of alien marine species into indigenous species assemblages (Hewitt et al. 2007), resulting in an uncontrollable biological experiment of immense proportions (Mooney and Cleland 2001). Invasive alien species are now widely accepted as one of the leading threats to biodiversity, after habitat destruction (Vitousek et al. 1997, Wilcove et al. 1998, Gurevitch and Padilla 2004), yet invasions also provide many opportunities for ecological and evolutionary research (Sax et al. 2005). Due to the convoluted nature of modes of invasion, establishment and subsequent outcomes (Cronk and Fuller 1995, Mooney and Cleland 2001, Mu¨ller 2001, Cox 2004, Valentine et al. 2007), it seems pertinent to outline invasion processes (Figure 1). Whilst this is necessarily a simplified model that does not take into account the possibility of multiple, sequential or coinciding metainvasions (Davies et al. 1999, Bohonak et al. 2001), it illustrates the potential changes that can occur. The consequences for the indigenous and invader species can be rationalized into two main groups: (1) changes in gene-pool composition, in population structure and allele frequencies; and (2) changes in genome organization at the species level through hybridization, and in individual gene expression profiles at the levels of expressed messenger RNA (the transcriptome) and the proteins synthesized (the proteome), and thus the phenotype. Current studies of bioinvasions are idiosyncratic and narrowly focused (reviewed in Grosholz 2002) and future studies must draw on a wider variety of techniques in the field of ecological and evolutionary genetics, including the exploitation of developments in bioinformatics (Lesk 2002). Marine ecosystems are potentially in a high-risk category, with estimates of approximately 10,000 species being transported daily (Ribera and Boudouresque 1995), yet marine research has lagged behind investigations of terrestrial and freshwater ecosystems (Grosholz 1996, Grosholz 2002). Nevertheless, some recent marine studies (e.g., de Rivera et al. 2005) are currently in the forefront of invasion biology. To our knowledge, there has been no previous general review of the genetic or evolutionary aspects of invading alien seaweeds. However, a number of reviews, briefly described here, include marine species and are relevant to the genetics of seaweed invasions. Holland (2000), reviewing the genetics of marine invasions, eloquently summarizes the key features of theoretical invasion genetics, and points towards the utility of molecular genetic techniques in an integrated pest management approach. Kolar and Lodge (2001) describe evidence for characteristics that determine the success of an invasive species, and assess the likelihood of predicting invasion w65x

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Figure 1 Key invasion processes, showing possible outcomes following introductions of an invasive species into a new region. There are insufficient data to give algal examples demonstrating all the processes on this diagram (cf. Schaffelke and Hewitt 2007), but examples are provided where possible based on the following studies: Fucus evanescens vs. F. serratus (Coyer et al. 2003); Sargassum muticum (Yendo) Fensholt wStrong et al. (2006) showed that it occupies a previously vacant niche on European coasts, where it can grow on soft sediments unlike native macroalgae; Sanchez et al. (2005), found that in Spain S. muticum is competitively excluding red algae, perhaps due to competition for lightx; Codium fragile subsp. tomentosoides (this subspecies is parthenogenetic: Trowbridge 1998, Prince and Trowbridge 2004); Sphaerotrichia divaricata (C. Agardh) Kylin (introduced populations of which are noninterfertile with native North Atlantic populations: Peters et al. 1993). An extinction vortex is defined as a situation where recruitment is close to the replacement level, so a small reduction in fitness changes a population from slightly positive growth to negative growth (McGinnity et al. 2003).

success. Mooney and Cleland (2001) and Lee (2002) examine specific aspects of the impact of invasions on evolutionary genetics. They highlight the rapidity of the invasion process, and note that successful invasions may be related to both the provenance and genetic responses of the invader. Grosholz (2002) points out that at larger geographical scales and longer timescales, our knowledge of introduced species dynamics breaks down; few studies deal with periods of more than ten years. Allendorf and Lundquist (2003) examined the population biology, evolution and control of invasive species, highlighting the point that, whilst genetic variation can identify invaders, control of invasives should perhaps involve a ‘‘shoot first, ask questions later’’ approach to eradication. Gurevitch and Padilla (2004) reviewed evidence for invasive species as general or primary agents of extinction, concluding that there are insufficient data available to support claims to this effect. Novak and Mack (2005) discuss genetic bottlenecks and how mating systems and genetic diversity influence invasions. In seaweeds, the majority of molecular studies to date have sought to identify the source of aliens, with particw66x

ular focus on cryptic invasions, in Caulerpa (Jousson et al. 1998, Meusnier et al. 2001, Fama` et al. 2002, Schaffelke et al. 2002, Verlaque et al. 2003), Codium fragile ssp. tomentosoides (Van Goor) P.C. Silva (Goff et al. 1992, Provan et al. 2005a, Pikea (Maggs and Ward 1996), Asparagopsis (Andreakis et al. 2004) and Polysiphonia harveyi J. Bailey (McIvor et al. 2001). Comparatively little work has been directed towards examining the consequences of seaweed invasions, for example the inheritance of adaptively significant traits and the population genetics of native and non-native species. Because of the vast range of life history patterns and mating systems in seaweeds (which include representatives of at least three phyla), many pertinent evolutionary and population genetic questions can be addressed by molecular approaches. Do populations of alien species show greater genetic diversity or lower diversity than conspecific native populations? Have alien species usually arrived as multiple introductions? What are the genetic threats in the form of introgression and loss of adaptive alleles? Is it possible to identify genetic traits of successful invaders? Do invading seaweeds exhibit predominantly clonal

D. Booth et al.: Genetics of seaweed invasions 387

reproduction, as is frequently observed in terrestrial and freshwater ecosystems (Elton 1958)? Do they show different mating systems (e.g., more selfing) compared to non-invasive congeners? Overall, are there any general trends in these parameters for invasive seaweeds or is each example case-specific? The aims of this paper are to provide an overview of seaweed invasions from a genetic perspective, focusing on invader species for which the greatest amount of molecular information is available, and to draw on relevant studies of better-known marine and terrestrial organisms to point the way forward in molecular approaches to the study of seaweed invasions. Where possible, we draw attention to evidence that may lead to answers to some of the questions posed above.

Identification of invasive species, sources and vectors Molecular markers A wide array of molecular markers is available to the phycologist today to discriminate invaders from native species. A variety of diagnostic markers for the mitochondrial, chloroplast and nuclear genomes can be used to classify organisms in a hierarchical fashion (Avise 1994, Ferraris and Palumbi 1996, Jarne and Lagoda 1996, Parker et al. 1998, Emerson et al. 2001, Provan et al. 2001). Depending on the level at which studies are aimed, from the species to the individual, markers that range from hypervariable microsatellite DNA markers (which vary in length due to different numbers of nucleotide repeats) to conserved gene sequences can be deployed. However, it is important to note for the assumptions of population genetic analyses that biological invasions are far from normal or equilibrium conditions (Davies et al. 1999, Bohonak et al. 2001). Identification of invasive species and their provenance Arguably this step is fundamental to any research on invasive species as without a working knowledge of these elements, little can be done to understand, predict, or mitigate the consequences. Due to the difficulty in identifying algae by morphology alone, introductions may be cryptic in nature or composed of multiple morphologically similar species; this phenomenon has been reported for many marine species groups (e.g., Carlton 1996, Geller et al. 1997, McIvor et al. 2001, Zuccarello and West 2003, Andreakis et al. 2004). The earliest attempts to identify sources of introduced seaweeds used plastid genome markers. The vector for an invasion of Codium fragile (Suringar) Hariot into San Francisco Bay in the late 1970s (Silva 1979) was claimed to be boxes of live fishing bait from New England packed in algae that had been discarded into the Bay (Dawson and Foster 1982). Goff et al. (1992) investigated the source of this invader using plastid DNA restriction fragment length polymorphism (RFLP) analysis, which cuts the circular plastid genome into fragments that form barcode-like patterns on gels. Restriction patterns con-

firmed the invader as the non-native subspecies C. fragile ssp. tomentosoides, but could not distinguish between an independent introduction from Japan and secondary transfer from introduced populations in the Atlantic Ocean (Goff et al. 1992). In a similar study by Maggs and Ward (1996), plastid DNA restriction fragment length polymorphism (RFLP) analysis was employed to address the origin of Pikea sp. in the Isles of Scilly, England, by comparison to Californian and Japanese samples. A high level of similarity indicated conspecificity between Californian and English Pikea californica Harvey and a possible vector was identified in the form of Catalina flying boats, over 500 of which were manufactured in San Diego and transported to England during the Second World War. Whilst plastid genome RFLP produces favorable results due to its conserved nature (Soltis et al. 1992), it does require whole DNA restrictions, and isolation steps to yield concentrated plastid DNA. Such analysis has now been largely superseded by polymerase chain reaction (PCR) methodologies to obtain informative haplotypes (Geller et al. 1994, Ferraris and Palumbi 1996, Hillis and Moritz 1996). The seaweed invasion that has received by far the most attention is that of Caulerpa taxifolia (M. Vahl) C. Agardh into the Mediterranean basin (Meinesz 2007). Initially it was shown that the invasive alga is not conspecific with the widespread tropical species C. mexicana (Olsen et al. 1998). Phylogenetic analyses of the noncoding internal transcribed spacer sequences (ITS) of nuclear ribosomal DNA (rDNA) demonstrated that Mediterranean C. taxifolia is closely related to strains cultivated in European aquaria prior to the first observations of the species in the Mediterranean Sea (Jousson et al. 1998). Several recent studies have focused on this highly invasive ‘‘aquarium strain’’ of C. taxifolia and similar results have been obtained with various markers. The origin of this strain is believed to be Australia (Meusnier et al. 2001, Fama` et al. 2002, Meusnier et al. 2002, Schaffelke et al. 2002, Meusnier et al. 2004). Meusnier et al. (2002) found that the Brisbane population of C. taxifolia was identical to the Mediterranean form, but two other Queensland populations (Kissing Point and Kelso Reef) showed high diversity of ITS2 types, none of which was the invasive genotype. Invasive C. taxifolia on the Californian coast was later shown to be the aquarium strain also (Jousson et al. 2000). Schaffelke et al. (2002) used ITS in an attempt to determine the origins of three newly discovered populations of C. taxifolia in New South Wales, Australia. They were able to adequately identify two of the three populations as being derived from tropical native stocks, but for the third they did not have enough resolution to exclude the invasive aquarium strain. Similarly, Kusakina et al. (2006) used morphological data and inter-simple sequence repeat (ISSR) markers to identifiy three independent invasions of Codium fragile ssp. tomentosoides into Atlantic Canada in New Brunswick, Malpeque Bay and Caribou Harbour respectively. Interestingly, these studies exemplify the point that invasions can occur as multiple events or from multiple sources, as noted by Geller et al. (1997) and discussed below. Recently, microsatellite markers have been applied to identify the provenance of an introduction. Coyer et al. w67x

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(2006) investigated Fucus serratus L. in Iceland and the Faroe Islands. By incorporating historical information on the shipping trade with their population structure data they identified the sources of the introductions and dated them to the 19th and late 20th centuries. Cryptic introductions Geller (1997) predicted that multiple conspecific introductions would be detected by molecular investigations of aliens. His prediction was borne out in a study of the red seaweed Polysiphonia harveyi Bailey, originally described from North America and considered an alien in European waters, with observations in locations over the past 170 years. McIvor et al. (2001) employed the gene coding for the large subunit of rubisco, rbcL, which had already proven useful in distinguishing species and populations of red algae. From a minimum spanning network of P. harveyi, two invasive haplotypes were identified that indicated two separate introductions into the North Atlantic Ocean from Japan. One came from Hokkaido, or nearby, into the northern North Atlantic Ocean (Nova Scotia and northwestern Europe). A second haplotype invaded from Honshu into the southern North Atlantic Ocean (North Carolina), and is also found in New Zealand and California. The presence of P. harveyi in New Zealand represents a cryptic invasion due to its morphological similarity with the native P. strictissima J.D. Hooker et Harvey, although the 4–5% sequence divergence between the two allowed unambiguous discrimination of the two species (McIvor et al. 2001). Freshwater et al. (2006) reported the detection of a semi-cryptic invasive species of Gracilaria in North America. By sequencing the rubisco spacer they were able to discriminate vegetative Gracilaria vermiculophylla (Ohmi) Papenfuss from native taxa with ‘‘nearly identical’’ morphology. Increase in both abundance and distribution of this species was found at 138 sites surveyed along the coast of North Carolina. Independent introductions into Europe of the green seaweed Codium fragile ssp. tomentosoides were reported by Provan et al. (2005a) after examination of eight indigenous and 15 introduced populations. Analysis of haplotypes combined from chloroplast microsatellite loci and chloroplast sequences indicated a minimum of two separate introductions from the native populations in Japan, one into the Mediterranean and the other into the North Atlantic Ocean and Chile. Tracking invasions in space and time Herbarium and museum samples potentially hold a unique record when considering the introduction of an invasive species, as they allow the exploration of past events, including cryptic invasions. The ability to isolate and amplify DNA from extinct ‘‘ancient’’ or archival specimens (Pa¨a¨bo 1989) is now commonplace, although precautions must be taken to avoid contamination by DNA from contemporary samples (Willerslev and Cooper 2005). For seaweeds, these problems can be minimized by using species-specific or lineage-specific primers. Whilst isolated ancient DNA molecules are replete with oxidative and hydrolytic damage (Pa¨a¨bo 1989, Hoss and w68x

Pa¨a¨bo 1993), short PCR primers can be developed that amplify small sections of genes studied. The ability to identify such temporal samples allows for increased accuracy in determining the first introduction of a species, particularly cryptic species such as Mytilus galloprovincialis (Lamarck) (Geller 1999). The application of molecular markers to herbarium material opens up new avenues of research, such as modelling using Bayesian and coalescent methodologies to estimate mutation parameters, population growth rates and effective population size from gene genealogies in seaweeds, as has been achieved previously for exploited fish species (Hauser et al. 2002, Hutchinson et al. 2003). Sufficient numbers of individuals (often more than 100) to allow for population genetics are available for some seaweeds: a wide range of species was preserved during the 19th and 20th centuries as multiple replicate herbarium samples known as ‘‘exsiccatae’’.

Invasions and population genetic structure Population genetics A key goal to be addressed by the molecular ecologist today is the identification and characterization of both indigenous species and invaders. Resolution of putative populations into management or evolutionary significant units (Moritz 1994) through an understanding of their biogeography or genetic structure, has the potential to identify the routes of transfer in order to recommend mitigation steps. In addition to this, population genetics studies can be used to estimate demographic parameters that can then be related to the stage and magnitude of invasion and the response of the indigenous species. Population parameters At each stage in the invasion process from introduction, succession or displacement to post-invasion, both the invader and the indigenous populations can be examined for a range of parameters. Given the nature of invasive algae, it is highly probable that such populations will have recently undergone a bottleneck (i.e., a reduction in genetic diversity that occurs in populations experiencing rapid, severe reductions in the number of individuals for one or more generations). The significance of bottlenecks for alien plant species has recently been reviewed by Novak and Mack (2005), with particular reference to the role of mating systems. The processes involved at each stage of invasion can be elucidated through estimation of population statistical parameters, such as effective population size Ne, most recent common ancestor estimates, migration rates, and evidence of genetic impoverishment. While these estimates are all intrinsically interlinked, a variety of statistical approaches have been suggested for most genetic markers. Tests for population bottlenecks involve assessment of the relationships between gene diversity and the observed number of alleles in bottlenecked versus nonbottlenecked populations (Luikart and Cornuet 1998, Luikart et al. 1998). Using the framework of coalescent

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theory (Kingman 1982, Hudson 1991), bottlenecked populations can be discriminated as they tend to exhibit star-like haplotype networks. It is also possible to determine whether such a loss of diversity came about due to population bottlenecks or selective sweeps (Galtier et al. 2000). Bottlenecks reduce variation throughout the whole genome due to drift, whereas selective sweeps, through directional selection, tend towards fixation of beneficial alleles and associated linked regions of the genome (Maynard Smith and Haigh 1974). Earlier studies provided single point estimates of coalescent time (Beaumont 1999, Beerli and Felsenstein 2001) and measures of genetic disequilibrium (Hill 1981, Vitalis and Couvet 2001). These have now been succeeded by more powerful temporal methods that employ samples distributed over several generations (Nei and Tajima 1981, Wang 2001, Wang and Whitlock 2003) and by Bayesian temporal methods (Beaumont et al. 2002, Berthier et al. 2002). Genetic depauperacy Changes in population structure and reduced genetic diversity associated with the genetic bottlenecks of introduction and colonization have been widely observed, mirroring the classical view from theoretical population genetics (e.g., Avise 1994, Williamson 1996, reviewed in Petit et al. 2004). Yet to fully understand the genetic consequences of an invasion, it is necessary to embrace the paradox that the invader has gone through a population bottleneck (Allendorf and Lundquist 2003, Tsutsui et al. 2003) and yet is still successful. Observations of reduced genetic variation within invasive species and populations clearly do not equate with lower evolutionary potential, reduced fitness or increased likelihood of extinction (Frankham and Ralls 1998, Reed and Frankham 2003), at least in the short term. In the invasive aquatic plant Phragmites australis (Cav.) Trin. ex Steud., Pellegrin and Hauber (1999) noted a low percentage of polymorphic isozyme loci, few alleles per locus and low genetic diversity. Likewise, Hofstra et al. (2000) found a lack of variation at the DNA level for both isozymes and randomly amplified polymorphic (RAPD) DNA in New Zealand Hydrilla verticillata (L.f.) Royle, an invasive submerged freshwater weed. Both studies attributed this pattern to the inconsequential amount of sexual reproduction. Alien populations of seaweeds likewise tend to exhibit reduced genetic diversity. The aquarium strain of Caulerpa taxifolia found in the Mediterranean Sea (and off San Diego, California) has severely reduced nuclear diversity within and between individuals compared with expectations for normal populations (Jousson et al. 1998, 2000). McIvor et al. (2001) also found higher levels of diversity in the indigenous Japanese populations of P. harveyi than those in the Atlantic, a trend parallel to that noted by Provan et al. (2005a) and Kusakina et al. (2006) in Codium fragile subsp. tomentosoides. Likewise, low genetic differentiation in Asparagopsis armata Harvey in the Mediterranean Sea for nuclear, mitochondrial and plastid markers (Andreakis et al. 2004) is thought to be the result of a recent invasion from Australia. Clonal reproduction is characteristic of all these species.

By contrast, increased genetic diversity has been observed in a number of invasive species such as mollusc aliens (reviewed in Holland 2000). The potential explanation for this apparent paradox lies in the research of Voisin et al. (2005) on the invasive kelp Undaria pinnatifida (Harvey) Suringar. High-resolution markers in non-coding regions of the mitochondrial genome showed increased genetic diversity in introduced populations relative to native populations, which was attributed to multiple introductions from highly differentiated sources. Admixtures from multiple sources result in higher diversity within introduced than in natural populations in many plants (Novak and Mack 2005). This mechanism can result in ‘‘patchwork quilt’’ patterns in population structure, and may lead to introgression of adapted genomes (Cox 2004). Indeed Novak and Mack (2005) suggest that this process may foster invasions, with some introductions being the product of admixed genotypes from the native range. These naturalized admixtures can then contain within-population diversity that exceeds the native range, with the potential for subsequent recombination and the generation of novel genotypes that can adaptively radiate. Postglacial colonization Postglacial colonization (reviewed in Hewitt 1999) reflects, on a grander scale, the dynamics of the invader. Refugial populations are equivalent to native populations in having high genetic diversity and the genetically impoverished colonies derived from the refugium have similar features to introduced populations. In this context, therefore, the postglacial colonization process may represent a model with which to examine the nature of changes in gene-pools over space and time due to invasion. There have been a few recent studies of postglacial recolonization of seaweed species in Europe, but they can provide illumination for some invasion scenarios. The genetic patterns characteristic of postglacial recolonization that have developed on millennial timescales may be detected with more rapidly evolving markers on the much shorter timescales of anthropogenic invasions. Studying the postglacial distribution of Fucus serratus in Europe, Coyer et al. (2003) collected data for seven microsatellite loci. They found high levels of genetic diversity throughout the species range, though the patterns in population structure were attributable to sequential colonization events and ‘‘edge’’ populations similar to those inferred by Comps et al. (2001). Heterozygote deficiencies found in 50% of the populations in the Atlantic Ocean and North Sea were attributed to sweeps of temporal colonization, possibly from a refugium in Brittany, causing population admixture. Conversely, Provan et al. (2005b), using a combined approach of nuclear and mitochondrial analysis in a phylogeographic analysis of the red seaweed Palmaria palmata (Linnaeus) Kuntze, found star-like topologies in minimum spanning trees of haplotypes that are normally indicative of sudden population expansions. The distribution of pairwise nucleotide mismatches, however, indicated a much deeper and slower colonization. A coalescent estimate of changes in effective population size revealed an expansion starting w69x

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around the penultimate glacial maximum, highlighting the time scale over which such colonization events can occur under natural circumstances. Hybridization Hybridization between species, and resulting introgression of genes from one species into another, are thought to be important components of the speciation process (see Hewitt 2001). Both processes have also been identified as elements of biological invasions. Hybrid zones have been well studied in natural populations of various organisms including the bee Apis mellifera (Linnaeus) (Clarke et al. 2002), the oak tree Quercus sp. (Craft et al. 2002), the aquatic plant Rorippa austriaca (Crantz) Bess. (Bleeker 2004) and the fishes Gadus morhua (Linnaeus) (Nielsen et al. 2003) and Oncorhynchus sp. (Rubidge and Taylor 2004). In the context of invasiveness, hybridization can work at the species and population levels by delivering additive genetic variation to depauperate gene pools (Soltis and Soltis 2000), by transferring and generating novel adaptations (Ellstrand and Schierenbeck 2000, Mooney and Cleland 2001, Arnold 2004), and by rearranging coevolved gene complexes into novel combinations through recombination in the introgressed F2 generation (reviewed in Soltis and Soltis 2000, McGinnity et al. 2003). There is little evidence of natural interspecific hybridization in seaweeds, with the exception of Fucus, which is a relatively recent radiation. Fucus vesiculosus Linnaeus, F. spiralis Linnaeus and F. ceranoides Areschoug are all closely related, with F. vesiculosus and F. spiralis hybridizing naturally, as evidenced both by intra-individual ITS sequence variation (Serra˜o et al. 1999) and microsatellite allele frequencies (Wallace et al. 2004, Billard et al. 2005, Engel et al. 2005). A recent study of the invasion of F. evanescens C. Agardh into the native range of F. serratus provides interesting details of some key genetic processes in invasions, including the formation of a hybrid zone (Figure 1). Fucus serratus and F. evanescens are naturally sympatric only in Iceland and NW Norway (Coyer et al. 2002). However, F. evanescens was introduced into Oslofjord (S Norway) in the late 19th century and spread to Sweden, reaching the western Baltic by 1992 (Wikstro¨m et al. 2002). Fertile hybrids have been identified by morphology in several areas where the two species now occur together. Using microsatellites, nuclear ITS1 sequences, plastid and mitochondrial markers, Coyer et al. (2002) demonstrated that individuals with hybrid morphology were indeed genetic hybrids, and investigated the population dynamics of the hybrid zone. F. serratus is dioecious whereas F. evanescens is hermaphroditic, and hybrids were found to be dioecious. Coyer et al. (2002) found only hybrids when both parents were above a threshold abundance level. Hybridization was asymmetric. Natural hybrids were all the result of mating between female F. evanescens and male F. serratus. This corresponded well with culture results in which F. evanescens male = F. serratus female crosses were only 20% as successful as the reciprocal cross. A subsequent study by Coyer et al. (2006) using Bayesian microsatellite analysis confirmed hybridization between F. serratus and F. evaw70x

nescens, with eight of 18 morphologically identified hybrids being F1. Bayesian admixture analysis of closely related F. spiralis and F. vesiculosus (Engel et al. 2005), and F. serratus and F. evanescens (Coyer et al. 2006) provides additional molecular evidence for this hybridization. The invasive strain of Caulerpa taxifolia has sometimes been interpreted as a hybrid (e.g., Makowka 2000), but these suggestions have not been substantiated. In Australia, individuals of C. taxifolia with the invasive genotype are up to six times larger and much more robust than the other genotypes, but there is no evidence that the invasive strain is the result of hybridization (Meusnier et al. 2004). The invasive strain is cold water tolerant compared to the native tropical strains, and sexual reproduction occurs rarely, as a stochastic event. However, in native Australian populations, hybridization and sexual recombination do occur. Meusnier et al. (2002) exploited single strand conformation polymorphism w(SSCP), which can detect single base mutations (Wattier and Maggs 2001)x of the ITS and a non-coding intron in the plastid 16S rDNA gene to evaluate the extent of sexual reproduction in Mediterranean and Australian C. taxifolia populations. SSCP polymorphism revealed several shared bands that constituted 10 different profiles. One individual from Kissing Point, Townsville, Australia showed a combination of two ITS2 profiles that were otherwise found in different individuals collected at the same site. Mediterranean populations showed no polymorphism with these markers. A second invasive taxon in the Mediterranean Sea (Verlaque et al. 2000), belonging to the Caulerpa racemosa (Forsska˚l) J. Agardh species complex, was also interpreted initially as the result of recent hybridization between the native C. racemosa var. turbinata (J. Agardh) Eubank -uvifera (C. Agardh) J. Agardh and an unknown tropical strain, based on 18S rDNA sequences (Durand et al. 2002). Later work by Verlaque et al. (2003) utilizing ITS1, 5.8S rDNA and ITS2 sequences clarified the taxonomic status of the invasive strain as C. racemosa var. cylindracea (Sonder) Verlaque, Huisman et Boudouresque and indicated a likely source population in Australia.

The future for molecular studies of seaweed invasions Genomics Several genome sequencing projects are now underway for seaweed species, including that of the first florideophyte, Chondrus crispus Stackhouse (J. Colle´n personal communication). Complete plastid genomes are already available for red, green and brown seaweeds. Further availability of genomic data will undoubtedly open up new fields for exploring seaweed invasions. Lee (2002) highlighted the role of genome architecture studies in investigating additive genetic variation, noting that hybridization processes generate ‘‘genetic substrate’’ from which traits can be selected in the invasion process. Adaptive traits include dispersal characteristics, growth rate, life history variation, mating systems and seed or propagule variation (Rejma´nek and Richardson

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1996, Clements et al. 2004, Petit et al. 2004). Lee (2002) also noted the role of gene-environment interactions (reviewed in Pigliucci 1996) and commented that invasive species are able to adapt to environments that are significantly different from that of their endemic range. Although adaptive genetic traits have not yet been sought in any invasive seaweeds, it is clear that a better knowledge of genomic arrangements, quantitative trait loci (QTL), phenotypic plasticity and associated adaptive traits could be fundamental to the future study of invasions, as these features may be associated with the underlying causes of success or failure of particular algal introductions. Knowledge of genomic architecture has gone hand in hand with the development of whole genome sequences and linkage studies in the form of chromosomal (Montgomery et al. 2001), microsatellite (Buetow et al. 2003) and single nucleotide polymorphism (SNP) maps (International SNP Map Working Group 2001). These maps are used in the study of commercially important or model organisms and are associated with large institutes (or, more commonly, groups of institutes) and research grants. With the advent of amplified fragment length polymorphism (AFLP) analyses, affordable high-throughput genome scans and QTL mapping have become available to molecular ecologists (see Liu and Cordes 2004). AFLP maps are now commonplace and are recognized as a powerful tool in predicting heterosis, investigating genome evolution and determining the proportion of the genome involved in adaptive functions (Klein et al. 2000, Wu et al. 2000). Although not yet developed in seaweeds, there are notable examples of such studies in marine invertebrates and freshwater fishes. Wilding et al. (2001) utilized 306 AFLP loci to examine speciation in parapatric morphs of the winkle Littorina saxatilis (Olivi). Wilding et al. (2001) found that 5% of the loci studied exhibited unusually high levels of differentiation. Cluster analysis of all loci grouped organisms by morph, whereas when these variable loci were excluded from the analysis, samples were grouped by site instead. These findings were attributed to the role of either divergent selection on a small proportion of the genome or differential introgression following secondary hybridization. Campbell and Bernatchez (2004) employed this approach also in assessing the role of selection and differential adaptation in sympatric pairs of Coregonus clupeaformis (Mitchill) (whitefish) ecotypes. Approximately 3% of some 440 AFLP loci screened deviated from expectations, and thus may be linked to adaptively significant genes. However, both studies come with strong caveats, namely that the tests for neutrality assume that loci studied are unlinked. This approach could be applied to the study of invading seaweeds as AFLP studies have already been pioneered in red algae (Donaldson et al. 1998, 2000) and in kelps (Erting et al. 2004). The combined approach of examining QTLs with AFLP genome scanning represents an ideal methodology for examining adaptively significant traits and phenotypic plasticity in invasive species (Rogers and Bernatchez 2005). Another approach in the identification of quantitative traits or adaptively significant genes is the use of microarray technology.

Microarrays Microarrays represent a novel way to determine the nature of gene-environment-phenotype relationships that could provide a tool to better understand how some seaweeds become invasive. Microarrays are essentially arrays of known probes fixed to a matrix that can allow subsequent hybridization of the expressed messenger RNA in the form of cDNA, and screening of gene expression in tissues and types of organisms (Schena et al. 1995, reviewed in Gibson 2002, Ranz and Machado 2006). Post-genomic research is now commonplace in model species and human gene expression studies, and is increasingly used to study fundamental questions in ecology, phenotypic plasticity and gene-environment interactions (reviewed in Morin et al. 2004) that have been highlighted as key traits in successful invaders (Grosholz 2002, Lee 2002). An affordable approach that is already being employed by phycologists uses the technique of subtractive hybridzation (Su et al. 1997, Pearson et al. 2001) to generate libraries of cDNA that are enriched for unique sequences and eliminate those common to both transcriptomes (the suite of transcribed cDNA sequences). This library of expressed sequences can be plotted onto a microarray, and sequenced to generate expressed sequence tags (EST). Gene expression in response to environmental parameters can then be examined by reverse-transcribing expressed RNA, and hybridizing the microarray. Upor down-regulation of genes in response to the chosen parameter can be visualized by fluorescence levels for each EST. Finally, these ESTs can be used to search protein databases using programs such as BLAST and subsequently isolate the gene expressed to a cellular location or biochemical pathway (Gene Ontology Consortium 2004). Whilst there is little possibility that there is an ‘‘invasiveness’’ gene and although microarray studies have limitations in terms of reduced sensitivity to low levels of gene expression, they may offer a valuable insight into the nature of gene-environment interactions, and may identify the gene pathways involved in producing invasive phenotypes. If we consider that a key trait in the invasion success of Codium fragile subsp. tomentosoides may lie in desiccation tolerance (Schaffelke and Deane 2005), and that in Caulerpa taxifolia the invasive strain is cold-water tolerant compared to the native tropical strains (Meusnier et al. 2004), relatively simple microarray studies could reveal the genes involved, affording detection of invasive and potentially invasive types. Assignment tests can reveal gene flow from invasive species Assignment tests have also recently been applied as an index to estimate population differentiation and admixture (for details see Paetkau et al. 2004). These tests allow the determination of an invader’s origin on the basis of the most likely population in which its multilocus genotype originated (Rannala and Mountain 1997). Tests involve either frequency (Paetkau et al. 1995) using a Bayesian method (Rannala and Mountain 1997, Pritchard w71x

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et al. 2000) or the likelihood-based Markov Chain Monte Carlo approach (Chikhi et al. 2001). Whilst assignment tests use population allele frequencies to ‘‘assign’’ individuals, they can be modified to implement both genetic distance and allele-sharing based measures. In these cases, for each locus, the individual is classified as a population sample of only two genes. Although potentially useful, however, assignment tests have a number of drawbacks. The most fundamental of these is that every individual is assigned, even with a low likelihood, to a given sample included in the baseline (even if this is not the true population of origin). Secondly, these tests tend to be most powerful when a large number of informative markers (i.e., 40 or more loci) are used (Bowcock et al. 1994), meaning that large amount of data must be collected. Assignment tests are best suited to detect potential immigrants and gene flow rather than to determine proportional contribution of stocks in a group of ‘‘unknowns’’. Mixed stock analysis Through admixture analysis it is possible to determine the genetic constitution of the mixed groups of populations or species (Utter and Ryman 1993). Mixed stock analysis has become powerful enough to allow the determination of the relative contribution to a mixture of native source populations (Galvin et al. 1995). The application of genetic stock identification procedures has allowed the evaluation of mixed stock in a variety of species, predominantly fishes, for several decades and may prove invaluable in studying invasive conspecifics. First, there must be a reasonable understanding of the individual stocks (i.e., populations) that are potentially contributing to the system. This must be linked with the existence of detectable temporally stable genetic differences among those groups and methodologies to detect them. Finally, there must be reliable methods for estimating the proportions of the various groups in the population. GSI uses complex algorithms to estimate the relative contributions of the putative populations in order to make up the multilocus genotypic frequencies observed in the mixed stock (Fournier et al. 1984, Pella and Masuda 2001). Recent advances in computational science and molecular tools have prompted the development of two analytical procedures. The conditional maximum likelihood estimate, which is readily implemented in the statistical package SPAM (Statistical Program for Analyzing Mixtures; Debevec et al. 2000), works by sequentially improving a computed ‘‘guess’’ until convergence at a maximum likelihood perceived to be the best estimate. This methodology is known to have a number of limitations, the main one being the assumption of Hardy-Weinberg equilibrium for the baseline samples (Pella and Masuda 2001). The pseudo-Bayesian analytical procedure, recently implemented by Pella and Masuda (2001) in the computer program BAYES, uses Bayesian likelihood functions to generate a prior probability density based on the relative frequencies of the alleles present in both the baseline samples, and in the stock mixture. Among the advantages of the pseudo-Bayesian procedure is that w72x

baseline groups are not necessarily required to be in Hardy-Weinberg equilibrium. Furthermore, this analytical approach seems to be much more robust with regard to incomplete sampling or missing stocks, which is likely given the cosmopolitan or wide distribution of many seaweeds.

Conclusion In conclusion, whilst it is critical to identify the source of the invading alga to fully characterize the biological niche it occupies, it is also necessary to determine the potential genetic changes that will occur following invasion. It is clear that future studies of hybridization, phenotypic plasticity, gene-environment interactions and differential gene expression will play a role in determining what makes an organism invasive. Since there has been relatively little work carried out on the genetics of invasive algae relative to other invasive groups, we have provided here a summary of valuable technical approaches. Examining the adaptive genetic traits that underlie the success of a particular invasive species may yield valuable information that can then be used to produce predictive models.

Acknowledgements Our genetic studies of invading seaweeds have been funded by ALIENS (EVK3-CT-2001-00062) and the Esme´e Fairbairn Foundation (Marine Aliens project).

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lation size and migration rates from genetic samples over space and time. Genetics 163: 429–446. Wattier R. and C.A. Maggs. 2001. Intraspecific variation in seaweeds: the application of new tools and approaches. Adv. Bot. Res. 35: 171–212. Wikstro¨m, S.A., T. von Wachenfeldt and L. Kautsky. 2002. Establishment of the exotic species Fucus evanescens C. Ag. (Phaeophyceae) in Øresund, Southern Sweden. Bot. Mar. 45: 50–517. Wilcove D.S., D. Rothstein, J. Dubow, A. Phillips and E. Losos. 1998. Quantifying threats to imperiled species in the United States. Bioscience 48: 607–615. Wilding, C.S., R.K. Butlin and J. Grahame. 2001. Differential gene exchange between parapatric morphs of Littorina saxatilis detected using AFLP markers. J. Evol. Biol. 14: 611–619. Willerslev, E. and A. Cooper. 2005. Ancient DNA. Proc. Roy. Soc. London Ser. B 272: 3–16. Williamson, M. 1996. Biological invasions. Chapman and Hall, London. pp. 244. Wu, R.L., Y.F. Han, J.J. Hu, J.J. Fang, L. Li, M.L. Li and Z.B. Zeng. 2000. An integrated genetic map of Populus deltoides based on amplified fragment length polymorphisms. Theor. Appl. Genet. 100: 1249–1256. Zuccarello, G.C. and J.A. West. 2003. Multiple cryptic species: molecular diversity and reproductive isolation in the Bostrychia radicans/B. moritziana complex (Rhodomelaceae, Rhodophyta) with focus on North American isolates. J. Phycol. 39: 948–959. Received 23 December, 2005; accepted 23 November, 2006

Botanica Marina 50 (2007): 397–417

2007 by Walter de Gruyter • Berlin • New York. DOI 10.1515/BOT.2007.044

Review

Impacts of introduced seaweeds

Britta Schaffelke1,* and Chad L. Hewitt2 Australian Institute of Marine Science, PMB 3, Townsville MC QLD 4810, Australia, e-mail: [email protected] 2 National Center for Marine and Coastal Conservation, Australian Maritime College, PMB 10, Rosebud, Victoria 3939, Australia 1

* Corresponding author

Abstract We analyzed 69 publications on the impacts of introduced seaweeds. The predominant impacts were changed competitive relationships in the recipient habitat, indicated by high abundances of invaders, resultant space monopolization, and reduced abundances/biomass of native macrophytes. Changes in biodiversity, effects on fish and invertebrate fauna, toxic effects on other biota, and habitat change were also identified. The mechanisms underlying the manifestation of impacts are uncertain and inferences about common patterns were hampered because impact studies were available for only a few introduced seaweeds, covered only a fraction of their introduced distribution and generally were conducted over short time scales. There was no information about evolutionary effects or changes of ecosystem processes. Knowledge of socio-economic impacts of invasive seaweeds is poor. We collated costs associated with control/eradication activities and for national spending on marine biosecurity in Australia, New Zealand and the United States. Prevention of impacts is the driving force for costly surveillance, eradication and control programs. Until we are able to understand, predict and measure impacts of introduced seaweeds, the management of species incursions needs to remain focused on early detection, rapid response and control to reduce the likelihood of negative impact effects. Keywords: control; ecological impacts; economic impacts; eradication; introduced macroalgae.

Introduction It is now widely accepted that global marine biodiversity and resource values of the world’s oceans are threatened by anthropogenic influences. In particular, overfishing, habitat alteration and destruction, global climate change and the introduction of alien marine species are identified stressors, especially in coastal regions (Norse 1993, Vitousek et al. 1997, Carlton 2000). The rate of marine introductions, including introductions of seaweeds, has

increased over the last 20 years, reflecting increased global trade and changes in economic activities; however, more awareness of the problem and increased survey effort are likely to have increased the detection of introductions (Ruiz et al. 2000, Perrings et al. 2002, Ribera Siguan 2002, Hewitt 2003a; see also Costello and Solow 2003). The assessment of ecological impacts of alien marine species has been recognized as a research priority in recent years. However, there are still very few rigorous studies of the impacts of aliens (Ruiz et al. 1999, Grosholz et al. 2000, Grosholz 2002). The threats posed are often inferred from estimates of introduction status and observations of negative impacts in other invaded areas, especially when aliens attain high abundances in a particular ecosystem. An alternative view is that most marine alien species have negligible impacts on their recipient environment, or are merely an addition to the ecosystem (Farnham 1980, Reise et al. 1999). However, there are some well-known examples of catastrophic effects of marine alien invaders on recipient ecosystems, e.g., the Asian clam wPotamocorbula amurensis (Schrenck)x in San Francisco Bay (Nichols et al. 1994), the comb jelly wMnemiopsis leidyi (A. Agassiz)x in the Black Sea (Kideys 2002), and the predatory sea star Asterias amurensis (Lu¨tken) in Tasmania and Victoria, Australia (Ross et al. 2003). Evidence is now also mounting that synergistic effects with other stressors play an important role for the establishment and spread of marine aliens species, and, hence, for any negative impacts (Ruiz et al. 1999, Occhipinti-Ambrogi and Savini 2003). Ecosystem alterations due to global change coupled with species introductions are thought to result in ‘‘biotic homogenization’’ (e.g., Olden et al. 2004, Olden and Poff 2004, Wilkinson 2004), a process by which ecosystems will become dominated by generalists and opportunistic species. This pattern has already been observed in some locations affected by environmental degradation and species’ invasions (McKinney and Lockwood 1999). Formal assessment frameworks for impacts of marine aliens, or introduced species in general, are scarce, both for ecological effects and for associated economic costs (Parker et al. 1999, Ruiz et al. 1999, Pimentel et al. 2000, Perrings et al. 2002, Hewitt 2003b). Marine macroalgae (seaweeds) are a significant component of marine alien taxa (Schaffelke et al. 2006) with current global estimates of introduced macroalgae ranging from 163 (Ribera Siguan 2002) to 260 species (J.E. Smith unpublished data). The current knowledge of impacts of alien macroalgae is even sparser than for other taxonomic groups of aliens. This is in contrast to the perception that invading macroalgae have potentially serious impacts, because they may alter ecosystem structure and function by monopolizing space, developw77x

398 B. Schaffelke and C.L. Hewitt: Impacts of introduced seaweeds

ing into ecosystem engineers, and altering foodwebs. Of particular concern is their potentially rapid spread beyond initial points of introduction through efficient dispersal, coupled with significant environmental and economic consequences (Thresher 2000). Documented impacts of seaweed invaders are known mostly from a few, well-studied, high profile species we.g., Caulerpa taxifolia (Vahl) C. Agardh, Codium fragile (Suringar) Hariot ssp. tomentosoides (Van Goor) Silva, Sargassum muticum (Yendo) Fensholt and Undaria pinnatifida (Harvey) Suringar; e.g., Ribera and Boudouresque 1995, Trowbridge 1998, Walker and Kendrick 1998, Boudouresque and Verlaque 2002, Levin et al. 2002, Ribera Siguan 2002, Wallentinus 2002, Occhipinti-Ambrogi and Savini 2003, Schaffelke et al. 2006x. In this review we synthesize and analyze current knowledge of impacts of alien seaweeds using published sources. We categorize reported impacts and classify the quality of the information (e.g., observational information, data from manipulative field experiments). Our aim is to find patterns of impacts, to examine whether certain species are more likely to cause significant impacts than others, and to identify mechanisms contributing to the observed impacts.

• Genetic effects – Within a species (e.g., introgression) – Between species (e.g., hybridization). Economic and societal impacts: • Direct – Costs of loss of ecosystem functions or values – Impacts on environmental amenity – Impacts on human health • Indirect – Management costs (government/non government) – Costs of research into introduced species – Costs for eradication and control measures – Costs for education/extension campaigns. Information about economic impacts was collated from quantitative assessments of cost or effort for control and eradication measures. While there are potentially other kinds of economic impacts (see list above), this was the only type of socio-economic impact for which there were sufficient quantitative data.

Ecological impacts Methods We examined 69 publications (;1980s to 2005) that present data on impacts of alien seaweeds; reviews and publications offering only distributional or observational data were omitted. Some original publications cited elsewhere proved difficult to obtain (e.g., reports to government agencies, unpublished proceedings). These are cited as the original author(s’) based on the secondary source (e.g., Wear and Gardner 1999, cited in Sinner et al. 2000). Studies that reported results from several alien species were listed as separate case studies for each species, unless they explicitly addressed interactions between the species. Results reported in multiple publications but leading to the same conclusion with regard to impact were listed as one case study, with all relevant references. Impacts or risks of impacts have been variously categorized (e.g., Gollasch and Leppa¨koski 1999, Parker et al. 1999, Ruiz et al. 1999, Grosholz 2002, Hewitt 2003b). For the purpose of this review we consider potential impacts into the following categories: Ecological and evolutionary impacts: • Direct and indirect competition with native biota (e.g., for light or substratum) – Space monopolization – Change in community composition • Effects on higher trophic levels (e.g., herbivores, associated fauna, toxicity) • Habitat change (e.g., changed structure, sediment accumulation) • Change of ecosystem processes (e.g., alteration of trophic structure). w78x

The 60 collated case studies report ecological impacts for only 17 species of introduced seaweeds (Table 1). These existing studies of impact address only a small fraction (;6.5%) of the current estimate of the total number of globally introduced macroalgal species (circa 260). The predominant effect of alien macroalgae reported in the case studies were changed competitive relationships in the recipient habitat (43 case studies, Table 1). These were generally indicated by high abundances of the alien species, resultant space monopolization and reduced abundances/biomass of native macroalgae or seagrasses. Changes in biodiversity, generally a decrease in richness of native macroalgal species in invaded areas compared to non-invaded areas, were reported in nine case studies. An additional 13 case studies reported effects of alien macroalgae on fish and invertebrate fauna in the recipient environment, with most cases reporting decreases in the number and abundance of species (Table 1). Six examples of toxic effects on other biota were reported for Caulerpa species. Less clear is the occurrence of habitat change, such as changes in productivity or habitat complexity (e.g., addition or loss of canopy species); seven case studies report such changes, mainly, however, based on observations or assumptions. One case study reported a genetic effect, the occurrence of fertile hybrids between an alien and a native congener (Table 1). Three case studies found no significant impacts of the alien seaweed studied. We were unable to find quantitative information about evolutionary effects or about changes in ecosystem processes caused by seaweed introductions. The majority of case studies focused on Caulerpa taxifolia, followed by Undaria pinnatifida, Sargassum muticum and Codium fragile ssp. tomentosoides. These species are the geographically most widely distributed alien macroalgae, and they are also able to attain high

Competitively superior to native Laurencia nidifica J. Agardh. High abundance at some sites, increasing total algal biomass

R

R

R

R

G

G

G

G

G

Acanthophora spicifera (M. Vahl) Børgesen

Acanthophora spicifera

Avrainvillea amadelpha (Montagne) A. Gepp et E.S. Gepp

Bonnemaisonia hamifera Hariot

Caulerpa filifomis (Suhr) K. Hering

Caulerpa racemosa (Forsska˚l) J. Agardh

Caulerpa racemosa

Caulerpa taxifolia (M. Vahl) C. Agardh

Caulerpa taxifolia

Negative effect on shoot density of Cymodocea nodosa, amplified by nutrient enrichment. In long-term, species co-exist, no influence of nutrients

Invaded areas: decrease in number, width, and longevity of leaves; health of Posidonia oceanica (L.) Delile. After 3 years of competition, mortality of sparse seagrass beds

Reduced species number, diversity and abundance of native macroalgae

Overgrowth increased shoot density of Cymodocea nodosa (Ucria) Ascherson and decreased shoot density of Zostera noltii Hornem. in mixed meadows

Increase in abundance since first record in 1920s to become dominant species in several locations in New South Wales, Australia

Increase in abundance, now common component of community (24% maximum cover)

Co-occurrence in previously monospecific Halophila hawaiiana Doty et Stone meadows

Most common introduced seaweed in Hawaii, common in intertidal and tide pools, displaces native macroalgae (but no data given)

Summary

Species

Table 1 Summary information of case studies of impacts of alien macroalgae.

E (enrichment, removal)

Sur (comp)

Sur (comp)

E (removal)

O (Serial collections)

O, Sur

O, Sur

O, Sur

O, Sur, E

Method

CC

CC

CC

CC

SM

SM

CC

SM

SM HC

Effect

Italy (Med, Elba)

Italy (Med, Elba)

Italy (Med, Tuscany coast)

Italy (Med, Tuscany coast)

Australia (Pacific coast)

USA (Atlantic coast) Gulf of Maine

USA (Hawaii)

USA (Hawaii)

USA (Hawaii)

Location

Ceccherelli and Cinelli 1997 Cecherelli and Sechi 2002

De Ville`le and Verlaque 1995

Piazzi et al. 2001a

Ceccherelli and Campo 2002

May 1976

Harris and Tyrell 2001

Smith et al. 2002

Smith et al. 2002

Russell 1992

Reference

B. Schaffelke and C.L. Hewitt: Impacts of introduced seaweeds 399

w79x

w80x

Higher density and diversity of invertebrate epifauna and fish in C. taxifolia compared to Cymodocea nodosa meadows species composition changed, mainly Polychaeta on C. taxifolia. Important economic fish species absent in C. taxifolia meadows

G

G

Caulerpa taxifolia

Caulerpa taxifolia

G

G

G

G

G

G

Caulerpa taxifolia

Caulerpa taxifolia

Caulerpa taxifolia

Caulerpa taxifolia

Caulerpa taxifolia

Caulerpa taxifolia

Caulerpa taxifolia

No evidence of decrease in Posidonia oceanica abundance, C. taxifolia and seagrass patches well isolated, indication of no significant competition

G

Caulerpa taxifolia

Caulerpenyne and C. taxifolia extracts inhibit development of sea urchin eggs

Consumption by sea urchins results in impaired gonadal development and loss of spines

Caulerpenyne and C. taxifolia extracts inhibit or delay the proliferation of phytoplankton strains

Reduced abundance of invertebrates compared to Zostera marina L.

Biomass of Ruppia maritima L. 20x lower in invaded patches

Colour change of a number of fish species inhabiting C. taxifolia meadows

No clear effect on composition and species richness of ichthyofauna, no change in fish feeding habits, reproduction, recruitment. Fish density and biomass slightly lower in invaded sites

Diversity higher in non-invaded area (note: Chisholm et al. 1997 also found large difference in water quality, sediment organics and pollution between the two areas)

Caulerpa taxifolia

Less biomass and diversity of native algal and invertebrate species

Summary

Species

(Table 1 continued)

Lab

Lab

Lab

Sur (comp)

Sur (comp)

Sur (comp)

Sur (comp)

Sur

Sur (comp)

Sur (comp)

Sur (comp)

Method

TO

TO

TO

HT

CC

HT

HT

–

CC

CC HT

HT HC

Effect

USA (Pacific coast, San Diego)

USA (Pacific coast, San Diego)

France (Med, Ligurian Sea)

France (Med, Ligurian Sea)

France (Med, Ligurian Sea)

France (Med, Ligurian Sea)

France (Med)

Italy (Med, Ligurian Sea)

Location

Leme´e et al. 1993, Pedrotti et al. 1996, Pesando et al. 1996, Amade and Leme´e 1998, Pedrotti and Leme´e 1999

Boudouresque et al. 1996

Merino et al. 1994 (in Boudouresque et al. 1995), Leme´e et al. 1997

Tippets 2002

Williams and Grosholz 2002

Arigoni et al. 2002

Francour et al. 1995

Jaubert et al. 1999

Verlaque and Fritayre 1994

Boudouresque et al. 1992

Relini et al. 1998a–c, Relini et al. 2000

Reference

400 B. Schaffelke and C.L. Hewitt: Impacts of introduced seaweeds

G

G

G

Codium fragile ssp. tomentosoides

Codium fragile ssp. tomentosoides

R

R

G

G

G

G

G

G

Caulerpa taxifolia

Caulerpa taxifolia Caulerpa racemosa Womersleyella setacea (Hollenberg) R.E. Norris Acrothamnion preissii (Sonder) E.M. Wollaston

Caulerpa taxifolia Caulerpa racemosa

Caulerpa taxifolia Caulerpa racemosa

G

Caulerpa taxifolia Caulerpa racemosa

G

Caulerpenyne changes behavior of marine ciliate possibly causing mortality

G

Caulerpa taxifolia

Codium now dominant, establishment only after previous disturbance of canopy-forming kelps, then preventing recruitment of native kelps. Native algal abundance decreased. Less abundance of juvenile fish in invaded patches

Codium now dominant, establishment only after previous disturbance of canopy-forming kelps, then preventing recruitment of native kelps. Native algal abundance decreased

Richness and species number of invertebrates (mainly amphipods) slightly reduced in invaded areas

50–100% cover of benthic macroalgae is introduced; cover and diversity of native species lower in invaded areas. No turf in C. racemosa patches, some W. setacea in C tax patches

C. taxifolia extracts and caulerpin from C. racemosa inhibit membrane extrusion pump (protective mechanism against toxins) in mussel and sponge, making sponges less resistant to tributyl tin

Laboratory testing of Caulerpenyne: antibacterial and antifungal properties, sea urchin larval and fish toxicity. Avoidance of treated feed pellets by fish, mortality in molluscs fed with treated algae

Reduced cover and number of native species colonized by one or both Caulerpa species, more pronounced in C. racemosa patches, compared to invaded controls. More Womersleyella in C. taxifolia areas

Summary

Species

(Table 1 continued)

O, E (removal, herbivory assays), Sur

O, Sur

Sur (comp)

Sur (comp)

Lab

Lab

Sur (comp)

Lab

Method

CC HC HT

CC HC

HT

CC

TO

TO

CC

TO

Effect

USA (Atlantic coast, Gulf of Maine)

Canada (Atlantic coast, Nova Scotia)

France (Med, Ligurian Sea)

Italy (Med, Tuscany coast, harbor area)

Italy (Med, Tuscany coast)

Location

Levin et al. 2002

Chapman et al. 2002

Bellan-Santini et al. 1996

Piazzi and Cinelli 2003

Schro¨der et al. 1998

Paul and Fenical 1986

Piazzi et al. 2003, Balata et al. 2004

Dini et al. 1994 (in Boudouresque et al. 1995), Ricci et al. 1999

Reference

B. Schaffelke and C.L. Hewitt: Impacts of introduced seaweeds 401

w81x

w82x

Shift in abundance over 25 years from kelpdominated to Codium and introduced red algae-dominated plus native opportunistic species wDesmarestia aculeata (Linnaeus) J.V. Lamourouxx, but high annual variation

G

G

B

B

R

R

R

R

R

R

Codium fragile ssp. tomentosoides

Codium fragile ssp. tomentosoides

Fucus evanescens C. Agardh

Fucus evanescens

Gracilaria salicornia (C. Agardh) E.Y. Dawson

Heterosiphonia japonica Yendo

Hypnea musciformis (Wulfen) J.V. Lamouroux

Hypnea musciformis

Kappaphycus spp. (as Eucheuma striatum F. Schmitz)

Kappaphycus species

Dominant at some sites, dense mats attached to reef substrata

Higher invertebrate diversity compared to noninvaded reef site, observation of coral mortality after smothering. No competition with native seaweeds, higher total algal biomass at some sites

Epiphyte on other macroalgal species, forms large monospecific blooms at some sites and/or grows intermingled with dense Ulva fasciata Delile

Often epiphytic on introduced Acanthophora spicifera, dominant species at some sites, together with A. spicifera leading to higher total algal biomass on some reefs

Most common species in sheltered and semiexposed subtidal (6 to )12 m), overgrowing benthos such as rhodoliths

Increase in distribution, dominant species at some sites

Lower biomass and fewer species of epiphytes and grazers, different epifauna composition and lower abundance of amphipods compared to native species

Fertile hybrids with F. serratus L.

Not preferred by sea urchins, no gonadal development on Codium diet

Summary

Species

(Table 1 continued)

O, Sur

O, Sur

O, Sur

O, Sur

O, Sur

O, Sur, Lab

Sur (comp)

Lab

Lab

O, Sur

Method

SM

SM HT HC

SM

SM HC

SM

SM

CC HT

G

HT

CC HC

Effect

USA (Hawaii)

USA (Hawaii)

USA (Hawaii)

USA (Hawaii)

Norway (North Sea coast)

USA (Hawaii)

Sweden (Baltic Sea)

Sweden (Baltic Sea)

USA (Atlantic coast, Gulf of Maine)

Location

Smith et al. 2002, Conklin and Smith 2005

Russell 1983

Smith et al. 2002

Russell 1992

Husa et al. 2004

Smith et al. 2002, Smith et al. 2004, Conklin and Smith 2005

Wikstro¨m and Kautsky 2004, Wikstro¨m et al. 2006

Coyer et al. 2002

Scheibling and Anthony 2001

Harris and Tyrell 2001

Reference

402 B. Schaffelke and C.L. Hewitt: Impacts of introduced seaweeds

Overgrowth of reef-building corals, leading to partial mortality

R

R

B

B

B

B

B

B

B

B

Kappaphycus species

Neosiphonia harveyi J. Bailey) M.-S. Kim, H.-G. Choi, Guiry et G.W. Saunders (as Polysiphonia harveyi J. Bailey)

Sargassum muticum (Yendo) Fensholt

Sargassum muticum

Sargassum muticum

Sargassum muticum

Sargassum muticum

Sargassum muticum

Sargassum muticum

Sargassum muticum

Recruitment after Macrocystis pyrifera (L.) C. Agardh dieback; seasonal S. muticum canopy at high density during peak of M. pyrifera recruitment, inhibiting re-colonization

Colonization of areas previously colonized by Zostera marina, no re-colonization by Z. marina

Most abundant species in lower indertidal and subtidal, decrease of Laminaria digitata (Hudson) J.V. Lamouroux abundance

Sargassum rapidly colonized exp. cleared areas, canopy then decreased recruitment of Rhodomela larix (Turner) C. Agardh

High abundance in tidepools caused decreased abundance of leathery and foliose macroalgae

Increased abundance of native kelp and understorey species after experimental removal of canopy. Fewer sea urchins at invaded sites

Canopy reduced cover of native species, esp. Laminaria saccharina (L.) J.V. Lamouroux and Fucus vesiculosus L.

Decreased cover and native algal species number under S. muticum stands

100% cover in one location

Summary

Species

(Table 1 continued)

E (removal)

O, Sur

O, Sur

E (removal)

E

E (removal), Lab

Sur

O, Sur (comp)

O, Sur

O, Sur

Method

SM

CC

CC

CC

CC

CC HT

CC

CC

SM

HT

Effect

USA (Pacific coast)

France (Atlantic coast)

France (Atlantic coast)

Canada (Pacific coast)

Spain (Atlantic coast)

USA (Pacific coast)

Denmark (Limfjord)

Italy (Med, Adriatic Sea, Venice)

USA (Atlantic coast, Gulf of Maine)

USA (Hawaii)

Location

Ambrose and Nelson 1982

den Hartog 1997

Cosson 1999

De Wreede 1983

Viejo 1997

Britton-Simmons 2004

Stæhr et al. 2000

Curiel et al. 1998

Harris and Tyrell 2001

Woo 2000 cited in Conklin and Smith 2005

Reference

B. Schaffelke and C.L. Hewitt: Impacts of introduced seaweeds 403

w83x

w84x

B

B

B

B

B

B

B

Undaria pinnatifida

Undaria pinnatifida

Undaria pinnatifida

Undaria pinnatifida

Undaria pinnatifida

Undaria pinnatifida

Undaria pinnatifida Sargassum muticum

Womersleyella setacea Acrothamnion preissii

B

Undaria pinnatifida

R

R

B

No detectable effect on native algal assemblage over 3 years

B

Undaria pinnatifida (Harvey) Suringar

Reduced functional diversity of seagrass rhizome epiphytes at sites invaded by turf species compared to unaffected sites

Dominant component in Venice lagoon, competition with Sargassum muticum assumed

High density of U. pinnatifida after experimental removal of native canopy, recovery to near control-level after 1 year

Establishment of U. pinnatifida at high abundance after dieback of canopy-forming Phyllospora comosa (Labillardie`re) C. Agardh

High density of U. pinnatifida after experimental reduction of native canopy, recovery after 2 years, but with changed native community composition

Decreased species richness and diversity of native seaweeds

Decreased cover of understorey species under 100% U. pinnatifida cover

Wellington Harbour: U. pinnatifida subcanopy invertebrate assemblages different, many ascidians, polychaetes and hydroids; accumulation of fine sediment; low cover of Corallina and turfs. Queen Charlotte Sound: increase in subcanopy species diversity (algae, molluscs, echinoderms), possibly because of increased habitat complexity

No detectable effect on native algal assemblage

Summary

Species

(Table 1 continued)

Sur

O, Sur

E (removal)

Sur (comp)

E (removal)

E (removal)

O, Sur (comp)

O, Sur (comp)

Sur (BACI)

Sur (BACI)

Method

CC HC

SM

SM

SM

CC

CC

CC

CC HT

–

–

Effect

France, Italy, Spain (several sites in W-Med)

Italy (Med, Adriatic Sea, Venice)

Australia (Tasmanian east coast)

Australia (Tasmanian east coast)

Australia (Tasmanian east coast)

Argentina (Patagonia)

Italy (Med, Adriatic Sea, Venice)

New Zealand

New Zealand

New Zealand

Location

Piazzi and Cinelli 2000, Piazzi et al. 2002

Curiel et al. 2001

Edgar et al. 2004

Valentine and Johnson 2004

Valentine and Johnson 2003

Casas et al. 2004

Curiel et al. 1998

Battershill et al. 1998

Wear and Gardner 1999, (cited in Sinner et al. 2000)

Forrest and Taylor 2002

Reference

404 B. Schaffelke and C.L. Hewitt: Impacts of introduced seaweeds

R

Methods: field survey (Sur); field survey comparing invaded and non-invaded sites wSur (comp)x; field survey with temporal comparisons before/after invasion wSur (BACI)x; field experiment (E); laboratory experiment or assay (Lab); observational study (O). Impact categories: competition with native biota, subcategories: space monopolization (SM); change in community composition (CC); Gsgenetic effects; HTseffects on higher trophic levels (HT), subcategory toxicity (TO), habitat change (HC), no significant impact shown (–). Bsbrown algae (Phaeophyceae), Gsgreen algae (Chlorophyta), Rsred algae (Rhodophyta). MedsMediterranean Sea.

Piazzi and Cinelli 2001 Italy (Med, Tuscany coast) Sur (comp) Species co-occur and compete with one another; overall space monopolization is independent of the respective dominant species R

Womersleyella setacea Acrothamnion preissii

SM CC

Airoldi et al. 1995, Airoldi 1998 Italy (Med, Ligurian Sea) Sur Dominant species R Womersleyella setacea

SM

Method Summary Species

(Table 1 continued)

Effect

Location

Reference

B. Schaffelke and C.L. Hewitt: Impacts of introduced seaweeds 405

abundances, or become the dominant benthic species in some locations. It is important to note that case studies of impacts of the above species are available only for small portions of the geographical ranges that have been invaded (see countries or regions marked by an asterisk in the following list): C. taxifolia, introduced to the: Mediterranean Sea (Croatia; France*; Italy*, Monaco; Spain; Tunisia), NW Pacific Ocean (Japan – failed introduction), NE Pacific Ocean (USA: California*), SW Pacific Ocean (Australia: South Australia, New South Wales). U. pinnatifida introduced to the: Mediterranean Sea (France; Italy*), NE Atlantic Ocean (Belgium; Netherlands; England; France; Spain), NE Pacific Ocean (USA: California; Mexico: Baja California), Australasia (Australia: Tasmania*, Victoria; New Zealand*), S Atlantic (Argentina*). S. muticum, introduced to the: Mediterranean Sea (France; Italy*), NE Atlantic Ocean (Belgium; Denmark*, Great Britain: England, N Ireland, Scotland, Wales; France*; Germany; Ireland; The Netherlands; Norway; Portugal; Spain*; Sweden), NE Pacific Ocean (Canada: British Columbia*; USA: Washington and Oregon*: Mexico: Baja California. C. fragile ssp. tomentosoides, introduced to the: Mediterranean Sea (France), NE Atlantic Ocean (Belgium; Denmark; Great Britain: England, N Ireland, Scotland; France; Germany; Ireland; The Netherlands; Norway; Portugal: Azores; Spain; Sweden), NE Pacific Ocean (USA: California and Oregon), NW Atlantic Ocean (Canada: Nova Scotia* and Prince Edward Island, USA: Connecticut, Maine*, Maryland, Massachusetts, New York, North Carolina, Rhode Island), SW Pacific Ocean (Australia: Tasmania, Victoria, New South Wales; New Zealand). The nature and, where known, the underlying mechanisms of the ecological impacts of these four species are discussed in detail below (also refer to Table 1). We briefly discuss case studies of red algal introductions to the Hawaiian Islands, a small but well-studied area with a relatively high number of abundant alien seaweeds (N.B.: While the number of alien seaweed species is higher in other areas, e.g., in the Mediterranean Sea, the proportion subjected to impact-related studies is relatively higher for Hawaii). Caulerpa taxifolia A large research effort has addressed the ecological impacts of Caulerpa taxifolia in the Mediterranean Sea. The presence of C. taxifolia had a negative effect on shoot density of the seagrass Cymodocea nodosa in short-term studies, especially under nutrient enrichment (Ceccherelli and Cinelli 1997), whereas long-term experiments suggested that the two species are likely to coexist and that high nutrient availability will not change competitive relations (Ceccherelli and Sechi 2002). In contrast, the dominant seagrass in the Mediterranean Sea, Posidonia oceanica, is negatively affected by competition with C. taxifolia, leading to decreased productivity and shoot mortality, especially in sparse meadows (De Villele and Verlaque 1995). P. oceanica facilitates C. taxifolia colonization and growth by providing physical w85x

406 B. Schaffelke and C.L. Hewitt: Impacts of introduced seaweeds

protection, rather than shade (Ceccherelli and Cinelli 1998, 1999). The distribution and abundance of P. oceanica has, however, not changed in the Bay of Menton (French Mediterranean Sea) over 7 years since C. taxifolia was introduced; C. taxifolia- and P. oceanica-dominated areas seem well separated, implying minimal competition at larger geographic scales (Jaubert et al. 1999). Sites on the French Mediterranean coast colonized by C. taxifolia typically show reduced biomass and diversity of native macroalgae and invertebrates and low fish abundance (Boudouresque et al. 1992, Verlaque and Fritayre 1994, Francour et al. 1995, Bellan-Santini et al. 1996). In contrast, Italian studies (only about 50 km from the French study sites) report higher biomass and diversity of invertebrates and fish in C. taxifolia meadows (presumably as a result of increased habitat complexity), but a significant lack of some important economic species that require open sand habitats (Relini et al. 1998a–c; 2000). Toxic secondary metabolites of C. taxifolia had negative effects on sea urchin larvae and protists in the laboratory (Table 1), but whether similar effects manifest in the field is unknown. Since the early 1990s, a second Caulerpa species has been spreading in the Mediterranean Sea, recently proposed as C. racemosa var. cylindracea (Verlaque et al. 2003). This species was also recently recorded as introduced in a water body in South Australia (Collings et al. 2004). In Italy, overgrowth by C. racemosa var. cylindracea reduced diversity and abundance of native macroalgae, especially turf and encrusting species (Piazzi et al. 2001a), and in mixed meadows of C. nodosa and Zostera noltii decreased shoot density of the former species but increased density of the latter (Ceccherelli and Campo 2002). Where they co-occur, C. racemosa var. cylindracea has higher growth rates and is competitively superior to C. taxifolia (Piazzi et al. 2001b, Piazzi and Ceccherelli 2002). Colonization by either species reduced diversity and abundance of native macroalgae compared with uninvaded areas, with C. racemosa var. cylindracea having the most pronounced effect (Balata et al. 2004). On the Tuscan coast, Italy, the two introduced Caulerpa species also interact with two introduced turf-forming red algae, Womersleyella setacea and Acrothamnion preissii (Piazzi and Cinelli 2003). The four species form a mosaic of largely introduced assemblages, with different species dominating, depending on habitat. Native species abundance and diversity are lower than in uninvaded areas (op. cit.). The introduced turf assemblages also promote growth and spread of introduced Caulerpa, whereas areas with a higher complexity and species diversity were less conducive (Ceccherelli et al. 2002). In summary, introduced Caulerpa species have monopolized benthos in some areas of the Mediterranean Sea, and through increased competition have caused significant changes to community composition, usually evident as reduced cover and richness of native seaweeds and marine plants. Impacts of introduced Caulerpa taxifolia in other parts of the world are scarcely known. In California, biomass of the seagrass Ruppia maritima was 20 times lower in meadow patches colonized by C. taxifolia (Williams and Grosholz 2002), and abundance of invertebrates was w86x

lower in C. taxifolia patches than in Zostera marina meadows (Tippets 2002). Information about impacts in the southern states of Australia is at present primarily anecdotal (Glasby et al. 2005). Undaria pinnatifida Undaria pinnatifida populations dominate space in many regions where the species has been introduced (e.g., Sinner et al. 2000, Forrest and Taylor 2002, Hewitt et al. 2005). Manipulative field experiments demonstrate that the establishment of U. pinnatifida is facilitated by increased substratum availability created by disturbance (Valentine and Johnson 2003, Edgar et al. 2004, Valentine and Johnson 2004). Once established, it forms predominantly seasonal canopies that act to decrease cover, and sometimes the diversity of understorey species (Battershill et al. 1998, Curiel et al. 1998, Casas et al. 2004). However, other studies have detected either no significant differences in diversity or cover of native macroalgal assemblages in invaded versus non-invaded areas (Wear and Gardner 1999, cited in Sinner et al. 2000, Forrest and Taylor 2002) or, more rarely, an increase in subcanopy species diversity (Battershill et al. 1998). Re-establishment of native assemblages after 1 to 2 years has been observed where high abundances of U. pinnatifida have been removed by manual clearing (Valentine and Johnson 2003, Edgar et al. 2004), albeit with changed species composition (Valentine and Johnson 2003). Sea urchin grazing can significantly reduce U. pinnatifida abundance but not enough to prevent canopy establishment (Valentine and Johnson 2005a). However, U. pinnatifida seems not to inhibit recruitment of native understorey species (Valentine and Johnson 2005a,b; see also Valentine et al. 2007). At low grazing pressure, U. pinnatifida persists while native canopy-forming seaweeds recover poorly, due to build-up of sediment in areas where native canopy-forming species are lost. Sargassum muticum Shortly after the discovery of Sargassum muticum on the south coast of England, the species was reported to have profoundly altered the coastal ecology, albeit without supporting data (Fletcher and Fletcher 1975). Recruitment and establishment of this species is often facilitated by disturbance creating available substratum (Ambrose and Nelson 1982, Deysher and Norton 1982, Critchley et al. 1987). The seasonal canopy of S. muticum then prevents re-establishment of native macroalgae (Ambrose and Nelson 1982, De Wreede 1983) and eelgrass (den Hartog 1997). Reduced abundances and sometimes reduced richness of native seaweeds have been found in invaded areas (Viejo 1997, Curiel et al. 1998, Stæhr et al. 2000, Britton-Simmons 2004). Underwater light measurements support the notion that shading by S. muticum is the most likely factor preventing re-growth of native species (Britton-Simmons 2004). The reduced abundance of native seaweeds has lead to decreased abundance of the sea urchin Strongylocentrotus droebachiensis (op. cit.), which avoids consumption of S. muticum, indirectly supporting the persistence of the introduced seaweed.

B. Schaffelke and C.L. Hewitt: Impacts of introduced seaweeds 407

Codium fragile ssp. tomentosoides Establishment and impacts of Codium fragile ssp. tomentosoides in the NW Atlantic Ocean have been facilitated by interactions with other introduced species. Periodic overgrazing by sea urchins (Johnson and Mann 1988) provided a disturbance to native seaweeds enabling establishment of C. fragile ssp. tomentosoides (Harris and Tyrell 2001, Chapman et al. 2002, Levin et al. 2002). Sea urchins (Strongylocentrotus droebachiensis) prefer kelp as a food source, only consume C. fragile ssp. tomentosoides when no other seaweeds are available (Sumi and Scheibling 2005) and have impaired gonad development on a diet of only this species (Scheibling and Anthony 2001). However, more importantly, natural sea urchin/kelp dynamics are disrupted by the spread of the introduced bryozoan Membranipora membranacea (Linnaeus), which overgrows kelp blades and leads to reduced growth, defoliation and gap formation in New England and Nova Scotian Saccharina latissima (L.) Lane, Mayes, Druehl et Saunders was Laminaria saccharina (L.) Lam.x beds (Harris and Tyrell 2001, Chapman et al. 2002, Levin et al. 2002). C. fragile ssp. tomentosoides recruits into these gaps and persists by inhibiting recruitment of kelp zoospores, the number of which is possibly further reduced by decreased kelp abundance (op. cit.). Space monopolization by C. fragile ssp. tomentosoides in this manner has resulted in reduced abundance of native macroalgae and of juvenile fish (Harris and Tyrell 2001, Levin et al. 2002). Ecological impacts of established C. fragile ssp. tomentosoides in other parts of the introduced range have not been studied. Space monopolization by C. fragile ssp. tomentosoides does not occur in the NE Atlantic Ocean, and Chapman (1999) suggested that high native floral diversity and grazing pressure prevent high abundances of C. fragile ssp. tomentosoides in this region. The introductions to the NE Atlantic Ocean occurred more than 30 years earlier than those in the NW Atlantic Ocean (reviewed in Chapman 1999). While changes in abundance are likely to occur over decades, there is unfortunately no information as to whether C. fragile ssp. tomentosoides in the NE Atlantic Ocean ever attained higher abundances in the past. Hawaiian macroalgal invasions At least 21 seaweed species have been introduced to the Hawaiian Islands, both accidentally and intentionally for seaweed aquaculture (Godwin 2001, Smith et al. 2002). Several red algal species (Acanthophora spicifera, Avrainvillea amadelpha, Gracilaria salicornia, Hypnea musciformis, Kappaphycus spp. and Eucheuma spp.) have established at high abundances and are spreading on Hawaiian coral reefs (Smith et al. 2002, Conklin and Smith 2005, G. Zucarello personal communication). These species monopolize space and increase overall macroalgal productivity and biomass on coral reefs (Table 1). Overgrowth of reef-building corals has been observed (Woo 2000, cited in Conklin and Smith 2005). Quantitative assessments of their ecological impacts and competitive relationships between each other and with native benthos are, however, not available. The introduced seaweeds exacerbate the problem of persistent

macroalgal blooms in some locations, e.g., Kaneohe Bay, which began in the 1960s with the establishment of high abundances of the native Dictyosphaeria cavernosa (Forsska˚l) Børgesen after disturbance and chronic nutrient enrichment (Smith et al. 1981). Alien and native bloom-forming macroalgal species now form a mosaic with overall high total algal cover sustained by low and spatially variable grazing rates (Stimson et al. 2001) and supported by sediment nutrient levels that remain elevated (Stimson and Larned 2000).

Economic impacts Information about economic impacts of alien seaweeds is generally rare, indeed the paucity of estimates of economic values in the marine sector in general has been identified as a significant gap (Colgan 2004). Direct impacts of marine macroalgae are largely unquantified, unlike impacts of macrophytes in freshwater systems. Cases of observed or anecdotal reports of economic impacts, summarizing effects on fisheries and aquaculture due to fouling of nets, ropes, floats and other maritime equipment, are collated in Ribera and Boudouresque (1995), Trowbridge (1998; for Codium fragile ssp. tomentosoides) and Sinner et al. (2000; for Undaria pinnatifida), but there are no quantitative data. One component of the economic impacts of invasive seaweeds is the cost of rapid response, control and eradication efforts (Table 2). Costs differ widely between reports (Table 2), but in most instances it is not obvious how estimates were calculated, so direct comparisons are potentially problematic. However, detailed breakdowns of costs are reported in three recent studies: Anderson (2005) for Caulerpa taxifolia in California, Wotton et al. (2004) for Undaria pinnatifida in the Chatham Islands, New Zealand, and Miller et al. (2004) for Ascophyllum nodosum (L.) Le Jolis in California (Table 2). The total sum of )US$ 7.5 million for the containment of C. taxifolia in California included immediately available emergency funds (the incursion was considered an environmental emergency similar to an oil spill) to commence the eradication and substantial funds for ongoing monitoring, research and public awareness (Anderson 2005; see also Anderson 2007 for further details on the eradication process). The costs of the successful eradication of U. pinnatifida from a sunken trawler in New Zealand were for (failed) salvage attempts (85% of total costs), in situ treatment of gametophytes and small-sized sporophytes on the ship’s hull (13%) and regular monitoring of the ship’s hull and adjacent shoreline (2%), all paid by the vessel’s insurer (Wotton et al. 2004). In both cases, there were unquantified costs for involvement of government agencies, local authorities, scientists and other stakeholders. Impacts on amenity and recreational value can be expected in situations where high abundances of introduced seaweed occur. Removal of beach wrack derived from Hypnea musciformis blooms in a coastal town in Hawaii costs ;US$ 55,000 year-1 (Van Beukering and Cesar 2004). The authors also predict a significant longterm economic benefit to the local economy via improved real estate values were the algal blooms controlled, e.g., w87x

408 B. Schaffelke and C.L. Hewitt: Impacts of introduced seaweeds

Table 2 Economic costs associated with eradication and control efforts for invasive seaweeds. Where no monetary value was available, an estimate of effort is given. Species

Summary

Cost/effort

Reference

Ascophyllum nodosum

Eradication by manual removal from small incursion area (total of 174 thalli)

US$ 4680

Miller et al. 2004

Caulerpa taxifolia

Rapid response, containment and ongoing monitoring of incursion in California, USA (2000–2005)

US$ 7.6 million over 5 year

Anderson 2005

Caulerpa taxifolia*

New South Wales, Australia, application of sea salt

US$ 5–23 m-2

Glasby et al. 2005

Estimated cost to treat all colonized areas (;8 km-2) in the State

US$ 46 million

Caulerpa taxifolia*

South Australia, freshwater treatment

US$ 4 million over 3 years

Neverauskas pers. comm.

Hypnea musciformis

Kihei coast, Maui, Hawaii. Removal of biomass from beaches

US$ 55,000 year-1

Van Beukering and Cesar 2004

Kappaphycus spp.

Removal from coral reefs in Hawaii

;2 person h m-2

Conklin and Smith 2005

Sargassum muticum

Manual removal by volunteers (group size unknown)

10–70 kg wet weight trip-1

Critchley et al. 1986

Sargassum muticum**

Estimated costs for mechanized removal (only applied at experimental scale)

;38 US$ t-1 (wet weight)

Hurley 1981 cited in Critchley et al. 1986

Undaria pinnatifida***

Successful eradication from a sunken vessel at the Chatham Islands, New Zealand (heat treatment and monitoring)

)US$ 1.9 million

Wotton et al. 2004

Undaria pinnatifida*

Manual removal at experimental scale

)US$ 23,000 over 3 years (5 person day 800 m-2 month-1)

Hewitt et al. 2005

Original figures were converted to US$ using exchange rates on 10/09/2006. Conversion factors used: *1 AU$s0.76 US$; **1 GB£s1.88 US$; ***1 NZ$s0.66 US$.

by reduction of nutrient inputs. We were unable to find other estimates of revenue loss caused by incursions of invasive seaweeds, as may arise, e.g., at dive sites that were previously attractions because of their high benthic diversity, or by impacting recreational boating or fishing activities (Critchley 1983). Estimates are generally unavailable for the indirect costs of invasive or potentially invasive seaweeds. These include associated costs of research and education/ extension activities. The New Zealand public good science funding agency, the Foundation for Research, Science and Technology (FRST), has explicitly allocated NZ$ 1.2 million year-1 (;US$ 0.8 million year-1) towards marine biosecurity research (C.L. Hewitt personal communication). However, this underestimates total marine biosecurity expenditure of FRST, given significant overlaps in research focused on biodiversity and biosecurity. In the United States, National Sea Grant allocates an estimated US$ 2.4 million year-1 towards research and outreach associated with marine invasions (C.L. Hewitt personal communication). w88x

Costs of management activities are usually not separated by taxonomic group, except in cases where there is a direct response to a particular invasion. In Australia, the establishment of a National System for the Management of Marine Pests is estimated to cost AU$ 7 million over the three-year period 2004–2007 (;AU$ 2.3 million year-1 s ;US$ 1.9 million year-1; N. Parker personal communication). This total is derived from a combination of appropriation funds within the Commonwealth Government (Department of Agriculture, Forestry and Fisheries, DAFF) and Natural Heritage Trust funding (shared between DAFF and the Department of Environmental Heritage). New Zealand has also recently adopted a Biosecurity Strategy (Biosecurity Council 2003) in which Marine Biosecurity was identified as a priority. As a consequence, the government agreed to a significant investment in enhanced marine biosecurity delivery in the 2004/05 budget, leading to an increase in marine biosecurity expenditure of almost 300% to ;NZ$ 6.9 million year-1 (;US$ 4.8 million year-1), representing ;4% of total biosecurity expenditure (Hewitt and Bauckham

B. Schaffelke and C.L. Hewitt: Impacts of introduced seaweeds 409

2004, Hewitt et al. 2004b). While this is proportionally much less than the economic contribution of primary marine industries to New Zealand’s GDP, it is a large improvement over previous investment. The management of marine introduced species in the United States is vested within a large number of Federal and State agencies including the US Coast Guard, US Geological Survey, US Fish and Wildlife Service, and individual State natural resource management agencies. Identifying all expenditure on managing marine introductions is beyond the scope of this review. We were unable to find any quantitative information about societal impacts of seaweed invasions.

Discussion Our review of available published literature showed that quantitative assessments of ecological and economic impacts of invasive seaweeds are still scarce. The lack of these data, for both marine and terrestrial ecosystems, is generally bemoaned in the invasion biology literature (e.g., Parker et al. 1999, Ruiz et al. 1999, Gurevitch and Padilla 2004). The data are urgently required to adequately inform and guide the management of invasive or potentially invasive species. The mechanisms underlying impacts of alien seaweeds are uncertain (see Levine et al. 2003 for discussion of this issue for better-studied higher plant introductions). In the majority of reported cases, impacts are typically expressed as community dominance of the invader through monopolization of space, and changing competitive relationships in the native assemblage. However, the mechanisms causing these community changes are mostly unknown (but see Valentine et al. 2007). Impacts of alien species cannot be viewed in isolation from the preceding stages in the invasion process, namely successful establishment and spread (for further discussion see Valentine et al. 2007 and Dunstan and Johnson 2007). These preceding stages and the manifestation of impacts through high abundances and space monopolization reflect characteristics of i) the recipient environment (e.g., disturbance, resource availability, competition and community composition) and ii) the invader (e.g., high growth rates, high fecundity). Closer examination of these two factors and of interactions between invaders may suggest the underlying mechanisms for the observed community changes. Influence of the recipient environment The analysis of the invasion history of a species is often used to predict whether that species would become invasive elsewhere, and hence likely to cause negative impacts (Lodge 1993, Williamson 1999, Hayes and Sliwa 2003, Branch and Steffani 2004). However, impacts observed in one location often do not predict the effects in another location, because the factors determining success of establishment and further spread may be site- or time-specific (Grosholz 1996). A good example of this is the significant impact of introduced Codium fragile ssp. tomentosoides on western Atlantic coasts, compared to

the relatively benign effect of this species on benthic communities in the east Atlantic Ocean (see above, Table 1). There are indications for a relationship between disturbance, which may lead to resource variability in the recipient habitat, and the establishment of invasive species and their proliferation to high abundances with associated impacts (Davis et al. 2000, Mack et al. 2000, MacDougall and Turkington 2005, Dunstan and Johnson 2007, Valentine et al. 2007). Once established, positive feedback mechanisms can enable invasive seaweeds to persist and flourish, even in the absence of the original disturbance (Valentine et al. 2007). Anthropogenic disturbance leading to changes in resource availability (e.g., high nutrient availability, water and sediment pollution, structures providing artificial substrata and altered temperature regime due to effluents) often leads to higher incidence and abundance of invaders (reviewed in Carlton 1996, Gollasch and Leppa¨koski 1999). For example, Undaria pinnatifida often forms dense stands on artificial substrata (e.g., Floc’h et al. 1996) and abundant populations of C. fragile ssp. tomentosoides in Australia are generally found in engineered environments, e.g., marinas, wharfs, jetties, bund walls and riprap (B. Schaffelke personal observations). Highly abundant Caulerpa taxifolia has been found on sediments enriched with nutrients and organic matter from urban wastewater, resources which C. taxifolia can utilize, whereas uninvaded sites or sites with low invader abundance are less polluted (Chisholm et al. 1997). The less polluted sites also have higher cover of native macrophytes, which may be due to impacts of the invader or the pollution, or both. Extensive blooms of other introduced Caulerpa species have also been linked to local nutrient enrichment by sewage inputs, C. brachypus var. parvifolia Harvey, recently discovered in Florida, and C. ollivierii Dosta´l in the Bahamas (Lapointe et al. 2005a,b). Facilitation of introductions by climate change (Stachowicz et al. 2002) has not yet been demonstrated for introduced seaweeds, although biogeographic limits of many macroalgae are known to be temperature-controlled (Breeman 1988). Biological interactions also play a major role in controlling high abundances and space monopolization. Recipient habitats with low cover and diversity of native species (either chronically or after acute disturbance) often have a higher incidences and abundances of alien species (e.g., Gollasch and Leppa¨koski 1999), although this view has been challenged recently (discussed in Dunstan and Johnson 2007). Nevertheless, low diversity algal turf assemblages and seagrass meadows promoted the establishment of high abundances of introduced Caulerpa species in the Mediterranean Sea (Cecherelli and Cinelli 1998, Ceccherelli et al. 2002) and proliferation of several invading seaweed species is facilitated by reduced native macroalgal cover (see Table 1; and, e.g., Valentine et al. 2007). Avoidance by herbivores may be an important mechanism that causes shifts in community composition. Control of introduced macroalgal biomass by herbivory is often ineffective, either because invasive seaweeds are not preferred by native grazers (examples in Table 1; Caulerpa taxifolia, Boudouresque et al. 1996, Leme´e w89x

410 B. Schaffelke and C.L. Hewitt: Impacts of introduced seaweeds

et al. 1996; Codium fragile ssp. tomentosoides, Prince and LeBlanc 1992; Fucus evanescens, Schaffelke et al. 1995) or preferred by only a few grazers (C. fragile ssp. tomentosoides, Trowbridge 1995, 1998; Undaria pinnatifida, Thornber et al. 2004). However, in some instances no change of herbivores’ feeding habits was observed (C. taxifolia, Francour et al. 1995). Role of species’ functional traits Functional traits may influence whether some species are more likely to cause significant ecological or economic impacts. Nyberg and Wallentinus (2005) compared species traits (relating to dispersal, establishment and impact) between European alien and native species. Traits relevant to the manifestation of impacts were size (most invasive green and brown macroalgae were larger than their native counterparts) and growth strategies (invasive species more often form dense covers and inhabit a larger depth range than native species). In that analysis (op. cit.), species most likely to be successful invaders, and hence, likely to have significant negative impacts were Codium fragile ssp. tomentosoides, Caulerpa taxifolia, Undaria pinnatifida, Asparagopsis armata Harvey and Grateloupia doryphora (Montagne) M.A. Howe (currently accepted synonym: G. turuturu Yamada, Gavio and Fredericq 2002). A number of species’ traits known from well-studied invader seaweeds are likely to facilitate establishment of high abundances, ultimately leading to impacts. For example, shading by the canopy-forming Sargassum muticum was an important mechanism that reduced native biodiversity in invaded areas (Levin et al. 2002). Asexual reproduction and fast growth also have the potential to enable alien seaweeds to quickly colonize available space. However, traits observed in an invasive species are often also found in conspecifics or congenerics that are not known to be invasive (Paula and Eston 1987, Trowbridge 1996, Vroom and Smith 2001). Indeed, species’ traits alone are unlikely to help predict the likelihood and impacts of invasions (Valentine et al. 2007). Establishment of high abundances more likely depends on characteristics of the recipient environment that result in traits of aliens being advantageous for recruitment and growth, and on sufficient inoculation pressure (Davis et al. 2000, Davis and Pelsor 2001, Dunstan and Johnson 2007, Valentine et al. 2007). Interactions between invaders Multiple invasions into one location can synergistically disturb an ecosystem and facilitate the establishment of further alien species – a process that has been termed ‘‘invasional meltdown’’ (Simberloff and von Holle 1999; but see Lohrer and Whitlatch 2002). Examples for facilitation of establishment and growth of one alien seaweed by another is the promotion of Caulerpa taxifolia in the Mediterranean Sea by invasive red turf algae (Ceccherelli et al. 2002), and in California by the disturbance of native eelgrass beds by the mussel Musculista senhousia (Benson) and the anemone Bunodeopsis sp., both of which are alien (Reusch and Williams 1998, 1999, Williams 2002). In the Mediterranean Sea, where two introduced w90x

Caulerpa species co-occur, it is as yet unknown whether C. racemosa, which is competitively superior (Piazzi et al. 2001a, Piazzi and Ceccherelli 2002), will eventually replace C. taxifolia, or whether the two species will facilitate one another. The outcome of either scenario could be more serious ecological impacts than presently observed. Another example is the facilitation of space monopolization by Codium fragile ssp. tomentosoides by the invasive bryozoan Membranipora membranacea, as discussed above. Limited inference space Evidence of impacts of alien marine species is often hampered by the lack of suitable baseline data prior to invasion. Ross et al. (2003) suggest a weight-of-evidence approach to overcome the lack of pre-invasion data, and assessed impacts of a predatory seastar using information from small-scale experimental manipulations, detailed field observations and field surveys at various spatial scales in invaded and uninvaded areas. Such an approach has not yet been applied to assess impacts of seaweed incursions. Typically, studies are only initiated after an incursion has already occurred and use comparisons of sites colonized and not colonized by alien species (see Table 1 for examples). For example, Forrest and Taylor (2002) found no differences in native species richness and abundance due to the establishment of Undaria pinnatifida using a control-impact design. However, they suggest that the lack of benthic community data before establishment of U. pinnatifida significantly limited their ability to draw inferences. Uncolonized sites may be inherently different from colonized sites, and these differences may have resulted in the lack of the alien species establishment in uncolonized sites, and significant differences in community composition could thus be the result of confounding artifacts. In situ experimental introduction of species for impact studies is, typically, deemed unethical and in New Zealand it is illegal. In New Zealand, U. pinnatifida is classified as an ‘‘unwanted organism’’ under the Biosecurity Act of 1993, and so it is illegal to disseminate or transport this species. Scientists have tried to circumvent this dilemma through the experimental removal of established invading kelp for comparison with unmanipulated invaded control sites. The manipulated sites are used to simulate species composition in communities that have not been invaded. Results, however, may be difficult to interpret because the experiment may reset the assemblage to an earlier successional stage, which is different from the initial, undisturbed, community (Valentine and Johnson 2003). Re-establishment of native species is possible but full recovery may take several years (Valentine and Johnson 2003, Edgar et al. 2004) and may be impaired by the lack of native species in the immediate vicinity to provide for sufficient recruitment of spores. Known impacts of other species are limited mostly to single studies at small geographic scales, making comparisons difficult, and inferences about common patterns impossible. The impact studies assessed here cover only a small part of the introduced distributional range for even the best-studied introduced seaweeds (see above).

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Caulerpa taxifolia is the one introduced seaweed for which ecological impacts are well documented (Table 1). The majority of studies, however, are from two highly urbanized coastal regions in France and Italy, where C. taxifolia reaches very high abundances (see Table 1), and where impacts are most likely. Interestingly, contrasting results were found (see above and Table 1). Most alien marine species are found in the coastal zone (Carlton 1996), and urbanized embayments, estuaries and ports are considered to be ‘‘hot spots’’ of introductions (Hewitt and Martin 2001, Ruiz and Hewitt 2002, Hewitt 2003a). While environmental disturbance facilitating establishment of aliens may be greater in these environments, they also have a high inoculation pressure, i.e., one or more significant introduction vectors are generally present (Ruiz et al. 2000, Hewitt 2002, Ruiz and Hewitt 2002, Hewitt et al. 2004a). It is debatable whether reported impacts are inherent, species-specific consequences or whether they would be manifested only in disturbed environments. There is also some dispute about how much area of the Mediterranean Sea is colonized by C. taxifolia, and hence possibly impacted. Meinesz et al. (2001) estimated a colonized area in the Mediterranean of 131 km2, whereas remote sensing results suggest that C. taxifolia cover along the south coast of France may have been overestimated by a factor of ten (Jaubert et al. 2003). Impacts of invaders may also change through time. However, most impact studies are conducted over periods ranging from only weeks to at most a few years, and there is currently no quantitative information about invasive seaweed abundances or impacts on decadal or longer time scales. Long-term monitoring of Caulerpa taxifolia in the Mediterranean Sea (Meinesz et al. 2001, Meinesz 2007) is focused on tracking the distribution and spread of this invader, but does have limited abundance or impact information for specific sites. However, this monitoring indicates that areas of highest cover and colonized area are close to the initial incursion point (along the Ligurian coast) and that the spread of C. taxifolia is not slowing down. Observations of Codium fragile ssp. tomentosoides in the Mediterranean Sea and off the coast of Maine (USA) indicate that this species peaked about a decade after first discovery and then declined (reviewed in Trowbridge 1998). However, the reasons for this are unknown. For example, herbivore preferences may change over time from native to alien seaweeds, reducing invader abundance, and alleviating impacts (Stimson et al. 2001). In contrast, invading marine species often persist at low levels and later start to increase in abundance and spread, which Stockwell et al. (2003) attribute to either an initial period of adaptation or a change to previously functional environmental controls such as competition or herbivory. Other reasons may be density-dependent thresholds in survivorship or exponential growth after a lag phase. Even though rhodophytes are the most prevalent group of alien macroalgae (Ribera Siguan 2003), ecological impacts of this group are known from only a handful of species, mainly those introduced to the Hawaiian islands (Table 1), and are possibly underestimated. Rhodophytes are often inconspicuous and difficult to identify

to species level, there may have been separate introductions of morphologically dissimilar generations (e.g., gametophytes vs. tetrasporophytes of Asparagopsis armata, Maggs and Stegenga 1999), or cryptic invasions of sibling species that are morphologically indistinguishable from previously introduced species or native species (e.g., McIvor et al. 2001, Booth et al. 2007). The wider ecological consequences of genetic effects of seaweed invasions (the only example we found is the formation of fertile hybrids between the native Fucus serratus and the alien F. evanescens; Coyer et al. 2002) are currently unknown. Economic impacts The data are too limited to even roughly assess the economic impacts of invader seaweeds. An economic assessment of the impacts of seaweed invasions should cover all potentially affected values including use and non-use values (Perrings et al. 2002, Born et al. 2004, also see Nunes and van den Berg 2001 for a review of economic valuation of biodiversity). The economic costs of species invasions must also include other societal costs such as management and research. We have presented figures for a limited set of countries. However, it is impossible to identify the proportion of these expenses that apply to seaweed invasions only. We have indications of some costs involved with seaweed invasions, e.g., costs for eradication and control (Table 2). Other costs (e.g., for de-fouling of maritime structures) are perceived to be ongoing costs regardless of the presence of introduced species (Sinner et al. 2000). Even though costs for vessel maintenance (i.e., hull antifouling) are significant for commercial and recreational shipping sectors, they are unusually not considered to be specific to alien marine species (Hassall & Associates Pty Ltd. 2002). The management regimes currently under consideration for hull fouling in Australia and New Zealand may lead to specific, and alien marine species-associated additional costs of maintenance. However, the use of tributyltin in antifouling paints will be phased out globally by 2008 and costs of hull maintenance may increase. Aquaculture imports and transfers are the main vectors for invading seaweeds in Europe (Ribera Siguan 2002, Wallentinus 2002, Hewitt et al. 2007, Pickering et al. 2007). The ICES Code of Practice for the Introductions and Transfers of Marine Organisms (updated 2003, available at http://www.ices.dk) prescribes quarantine and disinfection procedures to alleviate this pathway; however, the costs of compliance with the Code are unknown. The potential for harvest of commercially valuable seaweeds, either accidentally or intentionally introduced, is generally viewed as a positive impact (see Pickering et al. 2007, for detailed information about intentional seaweed introductions). Invasive Undaria pinnatifida is harvested commercially in Australia (Tasmania) and, at least briefly, in Spain (Cremades 1993, cited in Wallentinus 2002). A commercial harvest policy is in place in New Zealand. The species has been cultured in France since 1983, albeit with limited success (Ribera and Boudourw91x

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esque 1995, Fletcher and Farrell 1999). In Argentina this species was first considered to be a new resource (Casas and Piriz 1996), but is now rather viewed as an ecological and economic threat to native seaweed resources (Casas et al. 2004). Introductions of seaweeds for aquaculture are common practice, especially of tropical carrageenophytes (Zemke-White 2004). Impacts of these introductions are poorly understood and are inferred from knowledge about impacts from red algae introduced to Hawaii for aquaculture trials (see Table 1 and Smith et al. 2002). A quarantine protocol for introductions of tropical seaweed has been established, targeting epiphytes and epifauna (Sulu et al. 2003); however, costs for these quarantine measures are unknown. The risks of intentional seaweed introduction have not yet been evaluated with cost-benefit analyses, and such analyses would be difficult to perform currently due to lack of data about impacts. The potential of future introductions of genetically modified seaweeds for aquaculture may add another dimension of uncertainty to this issue. Management of impacts Prevention of impacts is the driving force for costly surveillance, eradication and control programs. Managing the impacts of introduced seaweeds goes hand in hand with management strategies aimed at preventing new introductions in the first place and at controlling or eradicating established invading species (Hewitt 2003a, Hewitt et al. 2004b). Clearly, impacts will be avoided if species are prevented from arriving through a range of pre-border management options (op. cit.). Similarly, impacts are likely to be minimized if eradication/control measures are in place to limit the establishment and/or prevent high abundances of established invasive species wfor a description control measures for invading seaweeds see Wotton and Hewitt (2004) and Anderson (2007)x. Research needs Biological invasions have human causes and consequences (Perrings et al. 2002, Hewitt 2005). Future research on impacts of alien seaweeds (and other alien marine species) should focus on multidisciplinary research with biological, social and economic approaches. As impacts are intricately linked to the transport and establishment of alien marine species, much more knowledge is needed about the mechanisms involved in these preceding two stages of the invasion process. Frameworks need to be developed to better predict which species are likely to invade which habitats. The mechanisms that lead to high abundances of introduced seaweeds need to be identified, including the role of anthropogenic disturbance and climate change as confounding factors. The spatial and temporal variability of impacts need to be explored, which will improve the understanding of ecosystem vulnerability and adaptation. This knowledge will support implementation of Article 8h of the Convention of Biological Diversity (prevention, control and eradication of introduced species which threaten ecosystems or species). Without the capacity to measure and predict impacts of alien marine species, w92x

scarce funds for research and management are unlikely to be allocated where they are most needed.

Conclusion We were unable to find sufficient substantial quantitative information about the impacts of alien seaweeds to detect common patterns. Even though a number of studies have shown significant ecological impacts, the underlying mechanisms are largely unidentified and impacts may be specific to the invaded system or the period of time since establishment and/or past disturbance. In addition, knowledge about socio-economic impacts is extremely scarce. Currently, decisions about management of alien invasive seaweeds are mostly unsupported by best science. Until we are able to understand, predict and measure impacts of alien seaweeds on various spatial and temporal scales, the management of species incursions needs to remain focused on early detection, rapid response and control to reduce the likelihood of impact manifestation.

Acknowledgements We thank Craig Johnson and Anthony Chapman for the invitation to contribute to this special issue and for their helpful comments on the manuscript. We also thank Marnie Campbell, Jennifer Smith, Sven Uthicke and two anonymous reviewers for constructive comments and discussions.

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2007 by Walter de Gruyter • Berlin • New York. DOI 10.1515/BOT.2007.045

Review

Control of invasive seaweeds

Lars W.J. Anderson USDA-Agricultural Research Service, Exotic and Invasive Weed Research, One Shields Ave., Davis, CA 95616, USA, e-mail: [email protected]

Abstract Seaweeds have invaded ecosystems along the coasts of many countries where they can displace native algae and seagrasses, reduce biodiversity and impair habitat of fish and invertebrates. The most widespread and notorious cases have been introductions of Caulerpa taxifolia, which now infests over 20,000 ha of Mediterranean subtidal zones. Few attempts to control seaweed invasions have been successful, due to often harsh and highly variable physical conditions in marine environments, and the lack of efficacious methods. Use of heat, copper, chlorine, salt, freshwater and various mechanical (removal) approaches have been successful in reducing or eradicating some infestations. Biological control by herbivorous mollusks and sea urchins has been investigated, but has yet to result in any operational programs. Nutrient inputs from near-shore sources have exacerbated the spread of some species (e.g., off the Florida coast). To counter the increase in seaweed introductions and the spread of these species, it will be useful to adapt, where feasible, methods that have proven successful in controlling freshwater weeds. New methods will need to be developed as well. This will require better communication among researchers and managers working to reduce introductions and negative impacts of these seaweeds. Keywords: biocontrol; Caulerpa; Codium; eradication; herbicide; management; Undaria.

Introduction This article focuses on control of alien marine macroalgae as invasive species of tidal ecosystems that may be natural or altered habitats such as harbors, marinas, bays, estuaries and lagoons. Invaded areas are usually gradients from naturally occurring littoral zones toward highly modified and artificial habitats formed by construction of physical barriers and substrata to accommodate maritime activities. Though there is a long history of attempts to prevent or reduce the attachment of ‘‘fouling’’ marine species (including algae), these actions are directed primarily to protect the integrity of pilings, wharfs, and vessel hulls and to reduce drag on ships. Control of fouling species (e.g., through use of bottom paints, coatings, films, etc.) will not be discussed here. w98x

In contrast to typical fouling species, efforts to control alien invasive seaweeds have been extremely limited, and pertinent research is sparse (Schaffelke et al. 2006). On the other hand, basic and applied research on management of freshwater macrophytes has a rich and long history spanning over 100 years (Sculthorpe 1967, Gallagher and Haller 1990, Pieterse and Murphy 1990). Though there are obviously important physiological and ecological differences between freshwater angiosperms and seaweeds, there are many physical similarities in the habitats, particularly related to logistics of prevention, rapid response actions, containment, safe and effective chemical treatments, non-chemical methods, and removal. Furthermore, while there are several international, national, regional and even local societies organized to deal expressly with nuisance freshwater plants, there is no such extensive network of researchers and practitioners whose purpose is the improvement of management methods for invasive seaweeds. Similarly indicative is the lack of any examples of invasive marine algae in the relatively recent summary of invasive species as a global issue (Mooney and Hobbs 2000). It is important, therefore, to glean what is applicable from the freshwater environment. There are guides, manuals and practical management plans and strategies for freshwater ‘‘weeds’’, and an associated industry that encompasses aquatic herbicide development and use, biological control, mechanical control and prevention (e.g., AERF 2005). Given the paucity of verifiable, efficacious methods for controlling seaweed infestations, I have highlighted a few examples where management has been achieved at some measurable level, and have utilized parallels between the marine and freshwater physical environments to propose practical approaches for control of seaweed invasions.

Goals of seaweed control: management or eradication I have chosen to use the terms ‘‘management’’ and ‘‘eradication’’ from the well established vernacular of ‘‘pest control’’ in general. Actions aimed at management usually presume an unending infestation (i.e., self-sustaining), but one that may be reduced to some acceptable level, through effective management actions. Thus, ‘‘management’’ and ‘‘control’’ refers to methods used to mitigate or reduce the negative impacts resulting from an invasive species, short of complete removal. The goal of management is to reduce impacts to an acceptable level, either based on economic criteria or criteria that result in tolerable effects on the environment. Usually, management approaches are taken when complete extirpation of

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the target species is deemed not feasible, or when costs of eradication cannot be borne. This approach encompasses nearly all current attempts to deal with invasions of aquatic organisms, because their detection has usually occurred far too late for complete removal (eradication). The rapid increase in costs, and the potential non-target impacts of these typically tardy responses have been noted for invasive species in general (Rejmanek and Pitcairn 2002). The simple analogy of responses to a fire in a home, business or a ‘‘wildfire’’ provides apt clarity. ‘‘Management’’ of fire alone is clearly not a tenable approach and it explains the tremendous efforts directed at both prevention and detection of fires, and why action analogous to ‘‘eradication’’ of the fire is the expected goal. Eradication of an invasive alga, likewise, presupposes very early detection, coupled with rapid, costeffective responses whose goal is to completely remove, or kill all of a species’ vegetative parts, and any other viable part or propagule. I have emphasized propagule removal because those unfamiliar with the practical aspects of eradication often use ‘‘management’’, ‘‘control’’ and ‘‘eradication’’ interchangeably; they are quite different. For example, some level of propagule ‘‘bank’’ may be tolerated in management projects (e.g., via biological control), but cannot be permitted if eradication is to be achieved. Thus, the thoroughness of removal and requisite, concomitant intensity of surveillance required for successful eradication usually far exceed (at least initially) that needed for management alone. In addition, the importance of rapid, effective responses to algal introductions is supported by dispersal rates and capacities of marine macroalgae that can exceed by orders of magnitude those modeled for invertebrate larvae (see figure 3 in Kinlan et al. 2005, Wright 2005). For example, Schaffelke et al. (2005) report that Undaria pinnatifida can release zoospores at rates up to nearly 150 million per hour per sporophyte, and that zoospore releases occurred over a three-month period. These authors also outline release rates from other studies on laminarian species ranging from 2000 to 800,000 h-1 cm-2 surface area of sorus tissue. Notwithstanding the importance of rapid response (i.e., eradication) to ‘‘new infestations’’, the scale or scope of an existing infestation should not necessarily preclude eradication. For example, if some populations can be contained and eradicated locally, then an overall strategy for full eradication may be feasible. This is dependent on site characteristics, target species biology, and available containment/eradication methods. If new, highly efficacious methods become available even after significant spread, then the potential for eradication of large infestations may be greatly improved. An example of management would be the efforts to contain further spread of Caulerpa taxifolia (Vahl) C. Agardh in the Mediterranean Sea by constantly removing new phalanxes of the population spreading at the broad edges of the infestation. While this would be a Herculean task given some 20,000q ha infested, it might be possible to curtail further spread. Likewise, cutting or physical removal of Codium fragile spp. tomentosoides (Van Goor) P.C. Silva might be a feasible management approach for keeping well-defined intertidal areas free of further intrusion (Trowbridge 1999).

In contrast, the successful eradication of Caulerpa taxifolia in California exemplifies the highly intense and initially costly type of program required for complete extirpation of an invasive marine alga (Anderson 2005). Thus, the ultimate expense of over $US 7 million for eradication must be weighed against potential impact on the environment and economic activity had this species been allowed to spread to susceptible sites along the Pacific coast. One might argue that there is a continuum of goals (and attendant effort) between eradication at one end and ‘‘management’’ the other. However, experience suggests this is not the case, and that successful eradication programs necessitate a zero-tolerance goal. A firm commitment to that goal also unequivocally defines for all stakeholders the expectations, scope of resources, and time frame needed. A decision to eradicate as opposed to ‘‘manage’’ must also provide a scientifically driven endpoint that once attained, causes a significant reduction in further costs. The clear identification of a successful endpoint also justifies effective surveillance costs and provides incentives to preclude any further introductions (e.g., enforcement of regulatory constraints). In long-term ‘‘management’’ programs, it is much more difficult to argue as forcefully to block all possible pathways of introduction. There are also equally important social ramifications associated with management versus eradication goals because the latter will almost always require significant changes in behavior, reduced ‘‘public access’’ to the infested site, and may even necessitate access to private property for extensive periods. Finally, the rationale used to arrive at a decision to embark on a management or an eradication program must reflect the overall (generally long-term) goal, as well as the feasibility of achieving it. Thus, if sufficient opportunity for native organisms to sustain pre-invasion populations can be provided by simply suppressing populations of an invasive seaweed, then eradication may not be warranted. However, the ability to sustain a management program for many years needs to be assessed, as well as the competitive interactions between native species and an invasive alga, and the potential impacts on non-target species. These interactions are usually poorly understood unless the invasion is a reoccurrence in a well-understood ecosystem, or has occurred elsewhere in very similar ecosystems. Similarly, long-term impacts of not controlling or eradicating invasive seaweeds must be considered. For example, York et al. (2006) recently reported that infestations of Caulerpa taxifolia in Botany Bay (New South Wales, Australia) were associated with lower species richness of fishes, and at one site, reduced fish abundance when compared with sites containing native seagrasses.

General principles of management and eradication In the past few years, plans and strategies for effective responses to invasive species have been developed by various countries (e.g., Australia National Task Force w99x

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2000, Biosecurity Council 2003, Hewitt et al. 2004, Wotton and Hewitt 2004) and US state and federal agencies (e.g., Keppner 2002, FICMNEW 2003, 2004, CDFA 2005, US Fish and Wildlife Service 2005). Given the limited tools available to manage or eradicate invasive seaweeds, what reasonable set of objectives can be established to address seaweed infestations? How will success be measured? Who is responsible for coordinating control actions and where do the resources come from? Marine communities typically manifest complicated interspecific interactions, and the invasive algal species that pose a threat to them have multiple pathways of introductions. A general model for ‘‘invasive species response’’ can be adapted from other existing pestprevention and control plans cited above, and from successful control and eradication projects (e.g., Kuris and Culver 1999, Bax et al. 2002, Thresher and Kuris 2004, Anderson 2005). These response plans and successful projects share several essential components, although specific methodologies must be tailored to fit both the target seaweed species as well as the invaded habitat. I have summarized these actions in Figure 1, which presumes that efforts directed at ‘‘prevention’’ have failed, thus, resulting in a new infestation. In addition to adequate resources (funding, equipment, expertise), a consistent hallmark of successful programs has been an underpinning from, and reliance on, sound science. Only through that foundation can adequate, objective assess-

ments of all actions be made, while at the same time, sustaining public (stakeholder) confidence. Finally, it is essential to implement an adaptive management approach because, inevitably, unforeseen problems arise and new knowledge gained from monitoring and assessments of treatment actions will lead to adjustments and, hopefully, improvements in methods. In short, one cannot simply initiate or direct a given set of actions and presume the outcome. Instead, the process requires a culture of flexibility among collaborating individuals, agencies and public stakeholders, with the common understanding that multiple routes and changes may be necessary to attain the goals of the project. The ascending, arrowed lines in Figure 1 represent this principle. Establishing criteria for success Undertaking complex, technically challenging, and usually expensive, pest-management or eradication projects tends to create a high level of public and political interest. These efforts also often require commitment of resources for several years. Therefore, developing credible, defensible and quantifiable measures of progress is not only essential for evaluating actions, it is also crucial to attaining a recognizable endpoint. Under the best circumstances, these criteria can be defined very early in the project. The clear choice between ‘‘management’’ and ‘‘eradication’’ needs to be made very early since that

Figure 1 Generic scheme for developing response to invasive seaweed infestations.

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decision will drive different efforts with very different expected endpoints. For example, in the early discussions of options for dealing with the first US infestation of Caulerpa taxifolia (in Agua Hedionda Lagoon, California), the decision to eradicate the population was made within two weeks of discovery (Anderson 2005). The clarity this brought to subsequent actions facilitated acquisition of resources, drove public outreach and education, and eliminated discussions such as ‘‘What is an acceptable level of growth and occurrence of Caulerpa taxifolia?’’. It also became clear that trying to answer that question by field experiments or ‘‘monitoring growth’’ was not compatible with the process of eradication, given the track record of this species in temperate waters of the Mediterranean where it had spread to over 20,000 ha since 1985 (Meinesz 1999, 2001). Once a goal is established, two critical activities must be implemented: treatment actions and reliable evaluation of their field efficacy. Though this may seem obvious, the specific evaluative, or monitoring methods, the consistency of their validity and assured objectivity are not trivial matters. The process is analogous to establishing quality assurance/quality control (QA/QC) requirements for monitoring contaminants (e.g., chemicals) in the environment. The approach taken with the USA Caulerpa taxifolia project provides a good example. Field efficacy was evaluated by removing sediment cores from treated areas and providing adequate growth conditions for any viable propagules (Anderson et al. 2005). Continued QA/ QC for post-treatment monitoring was accomplished through SCUBA diver training, use of GPS-established transects, and multiple offset survey grids (i.e., search grids were rotated slightly between diver passes to ensure better surveillance). Equally important were routine evaluations of divers’ ability to detect Caulerpa,

which was determined by assessing their ability to locate various sizes of plastic ‘‘Caulerpa’’ placed randomly within the search areas, but without divers’ knowledge. This approach also answered a question that was later crucial to determining the duration of monitoring required, viz. what is the minimum size of a ‘‘colony’’ that would be detected 100% of the time in a single multi-diver grid survey even with poor visibility? The answer was between 0.5–1.0 m in diameter, though when visibility was especially good, success was nearly as good even for smaller patches. When the growth rate of C. taxifolia was considered, the likelihood of a missed colony going undetected for two years was virtually nil. By adding an additional year of surveillance for conservative assurance, the projected criteria for successful eradication was three full growing seasons following the last known real ‘‘find’’. The history of surveillance and occurrence of C. taxifolia in Agua Hedionda is shown in Figure 2. A similar pattern has been observed in the other C. taxifolia infested site in Huntington Harbor, USA. Based upon the surveillance and criteria established, declaration of eradication was made in July 2006 (California Department of Fish and Game 2006). The importance of post-treatment monitoring cannot be overemphasized, nor can it be ignored when estimating the costs of such projects. In the California project described, of the total cost (ca. $US 7 million), 80% was associated with post treatment monitoring.

Approaches to management and eradication of invasive seaweeds Invasive seaweeds are able to establish in a variety of marine environments ranging from relatively placid

Figure 2 Surveillance history and presence of Caulerpa taxifolia at Agua Hedionda Lagoon, California, USA. Note: lack of plants was the result of containment under PVC tarpaulins followed by injections of sodium hypochlorite, or placement of solid chlorine-releasing pellets (see Anderson 2005). Sprsspring; Sumssummer.

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lagoons, bays and estuaries to highly energetic and variable open coastal systems. These species also have a variety of growth forms, and reproductive and dispersal capacities (Zuljevic and Antolic 2000, Ceccherelli and Piazzi 2001, Ribera Siguan 2002, Wallentinus 2002, Miller and Chang 2004, Ruitton et al. 2005, Schaffelke et al. 2006). Clearly, the most problematic invaded sites are open coastal areas having high-energy wave action, rocky substrata with uneven surfaces, and subjected to strong currents. Similarly, seaweeds having pelagic occurrences, such as Sargassum muticum (Yendo) Fensholt, poses very difficult problems for containment or treatments. Conducting physical (or chemical) manipulations and surveillance in these types of environment pose tremendous logistic problems. Therefore, the physical and biological characteristics of an invaded site will very often determine the methods that may be feasible to use. This is also the case for freshwater ecosystems. Several strategies that have been successfully used for freshwater weed management or eradication could be adapted to marine systems. These potential seaweed ‘‘control’’ actions are separated into containment and treatment approaches. The importance or urgency of ‘‘containing’’ an infestation, rather than quickly attempting a particular management or eradication method depends upon the lag-time from introduction to ‘‘discovery’’, the likely rate of spread, likely impacts (if known), and the resources at hand. However, in most cases, and certainly in eradication programs, containment is a crucial and integral part of the entire strategy. Containment of invasive seaweed infestations There are no ready-made products or protocols for invasive seaweed management or eradication. Each invasion presents unique constraints and opportunities and, therefore, each will require creative problem solving tailored to the prevailing physical, biological, socioeconomic and political conditions. However, the first and most prudent action will usually be effective containment of the site to allow time for notification of appropriate stakeholders and for initial, rational, science-based evaluation of the potential impacts or threats from the infestation. The objective is simple: to prevent movement of viable parts of the invasive alga outside a delineated perimeter, and to reduce opportunities for further spread within the site. For example, Wright (2005) has shown that fragmented fronds of Caulerpa taxifolia may reach densities of 6,000 m-2 and that 30 to 45% of newly established stolons in several infested sites were derived from these types of fragments. The importance of containing (or removing) C. taxifolia fragments to prevent spread of colonies was recently demonstrated through modeling of their growth, coupled with various ‘‘removal’’ scenarios (Ruesink and Collado-Vides 2006, Wright and Davies 2006). The modeling showed that multi-seasonal removal of propagules could dramatically reduce the size of infestations. Thus, potential dispersal capacity must be curtailed immediately as a first response, whether through physical containment, or by judicious removal of propagules (Smith and Walters 1999). Due to the patchiness of many infestations, even when discovered early, a number of containment devices and w102x

methods may be needed, rather than a single massive one. These may range from simple enclosures (e.g., anchored plastic sheeting) to the complicated installation of large robust structures (e.g., cement pipes or metal dikes) to surround the infested site. In addition, any activities that may compromise containment must be curtailed, including, as appropriate, boating, fishing, anchoring, swimming, SCUBA or snorkel diving, or dredging. In some circumstances, complete closure or quarantine of an infested site may be necessary. These types of action almost always raise serious social, economic, legal and even political issues that must be dealt with quickly, effectively and constructively. A number of questions invariably arise with any discussion of containing the affected area: Who has jurisdiction of the site? Who ‘‘owns’’ the site? What legal and administrative processes are required to restrict access to the site? Who will be affected by such restrictions and how will communications be handled with these stakeholders? Who has authority to enforce restrictions, such as physical access and quarantine? Answers to these questions will not only affect containment approaches, they will ultimately also circumscribe proposed management or eradication efforts. Table 1 summarizes possible strategies and methods for containing seaweed infestations. Some have been successfully used for relatively small areas (e.g., less than 0.2 ha covered with PVC tarpaulins). Large containments would most likely rely on use of natural or existing constructed boundaries such as coves, groins, jetties, retaining walls in marinas, or movable locks, possibly in conjunction with additional solid barriers. The cost and probable effectiveness of a given method must be weighed against the risks of not containing the infestation. Treatment and control action for management or eradication The following sections describe specific methods for management or eradication of invasive seaweeds. Some of these approaches have been successful on varying scales of infestations in marine systems, and nearly all have been used for control or eradication of freshwater invasive plants. As with all pest control actions, those undertaken most quickly and with least delay from the onset of the infestation generally have been most effective and, in the long run, should be least costly (Rejmanek and Pitcairn 2002). At present, however, our ability to detect introductions of invasive seaweeds (and most other aquatic invasive species as well) is poor. There are numerous and varied pathways for such incursions, yet monitoring efforts are minimal whether at points of origin or at recipient habitats. Therefore, having management and eradication tools readily available is essential for adequate responses to these invasions when they are finally detected. The methods described, or at least suggested here, may have utility whether the goal of the response is long-term management or eradication. The difference in the outcome is as much a function of how the methods are used as in the choice of a specific method.

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Table 1 Physical containment methods for invasive seaweed infestations. Strategy

Methods and constraints

References/Comments

Cover and isolate infested area(s) with solid or flexible materials

Physically cover infestation with either weighted, flexible, solid or fine-mesh material (e.g., nylon or stainless steel, polyvinyl chloride, polyethylene sheeting). Practical limitations include size of infestations and the type of bottom contours (smooth sandy or mud bottom vs. jagged rocks). Adequate, secure anchorage may be achieved using either weighted materials (sand/rock/gravel bags/chains) or metal or fiberglass rods driven into the substrate. Cover infested area with adequate depth (thickness) of gravel, coarse sand or other dense material. Ideally, material can be gravity-fed through flexible hosing or pipes that allow SCUBA divers to place the material accurately

Merkel and Woodfield 2003 (tarpaulins). Sand, dense material cover: proposed, no reference found

Isolate by solid perimeter barriers

Insert linked steel plates (sheet piling), or concrete dikes, or other dense, rigid material. Alternatively, large, steel or aluminum cylinders could be used for small colonies. Lighter material can be anchored using weights placed near the bottom of the cylinder. The size and mass of these materials would necessitate shipboard hoists and coordinated diver-assisted placement. In deep areas of high surge or surface swells, containment cylinders could be off-loaded adjacent to the infested areas and later placed where needed

Proposed, no reference found for marine systems

Isolate by flexible barrier

Install weighted fine mesh or solid flexible curtains to surround the infestation. These are similar to oil spill containment skirts, but extending to the bottom where the lower edge is sealed with sand or gravel bags, or other dense materials. This method is probably only practical when little or no tidal or currents affect the site, or to cover infested vessel hulls

Madsen 2000 (freshwater systems), no references for marine systems

Drain or ‘‘De-water’’ site

After isolating and sealing the area, either allow tidal flux to drain water or pump out. This may only be practical in upper tidal zones, or where a narrow access channel can be closed off (e.g., small harbor or marina)

Madsen 2000, CDFA 1990 (freshwater sites), no reference found for marine systems

Prevent movement into/outside and within sites

Quarantine site. This decreases potential for further disturbance dispersal and isolates site for human activities

Merkel and Woodfield 2003, Anderson 2005, CDFA 2000

Mechanical These methods consist of various physical manipulations in or around infestations, including the substratum, with the objective of either removing, burying, or killing established seaweed. There is a long history of mechanical uses in freshwater plant and algal infestations (e.g., table 6 in Madsen 2000). Successes of these freshwater mechanical systems can be ascribed to many factors such as the many decades of development, trial and error, and redesign. Often the systems have to be tailored to specific bathymetric, sediment-types and logistic constraints on mobility of large pieces of equipment. In addition, these systems are typically deployed in relatively placid, accessible, lentic environments in contrast to tidally influenced and potentially high-energy wave conditions where invasive seaweeds have become established. The most effective methods incorporate some means for collection and removal of plant fragments after cutting or dislodging from the substratum, or simply complete burial (Glasby et al. 2005b). Conveyor belt transport sys-

tems can remove the bulk of cut plants, but there are nearly always large numbers of viable propagules released. Dispersal of propagules can be reduced somewhat by using floating curtains. Since seaweeds often produce microscopic gametes and zoospores, as well as very small multicellular stages, containment of these reproductive structures during mechanical operations may be impossible under most circumstances. Conklin and Smith (2005) point out how difficult it is to manually remove all propagules, as evidenced by the regrowth of Kappaphycus species in hand-cleared areas of Kaneohe Bay, Hawaii. The difficulty in physical removal is also illustrated in the recurrence of a population of Caulerpa taxifolia in the Croatian coastal harbor at Malinska within two years after it was initially ‘‘removed’’ by pumping (dredging) (Zuljevic 2001, Ivesa et al. 2006). Notwithstanding these difficulties, when conditions are favorable, even hand-removal can be effective as with the successful extirpation of Ascophyllum nodosum (L.) Le Jolis from a relatively localized infestation in San Francisco w103x

424 L.W.J. Anderson: Control of invasive seaweeds

Bay, California (Miller et al. 2004). Over 170 ‘‘free living’’ A. nodosum ecad mackayi (Turner) Cotton thalli and fragments were physically removed, bagged and disposed at a landfill offsite. The effort took 3 days, with monitoring following for a few months, and further surveillance over a longer period. The only discouraging aspect of this effort is that over two months passed between the discovery of the infestation and actions taken to remove it. Another interesting example of using physical (hand) removal for seaweed control are responses to some infestations of the Japanese-native alga Undaria pinnatifida (Harvey) Suringar. This kelp (Laminariales) has expanded dramatically from its native, temperate Asian range over the past 30 years to diverse habits in the USA (Silva et al. 2002), Mexico (Aguilar-Rosas et al. 2004), Australia (Sanderson and Barrett 1989, Sanderson 1990, Campbell and Burridge 1998, Valentine and Johnson 2003), New Zealand (Hay and Luckens 1987), Argentina (Piriz and Casas 1994, Martin and Cuevas 2006) and several European Atlantic and Mediterranean coastal areas as well (Boudouresque et al. 1985, Fletcher and Manfredi 1995, Floc’h et al. 1996, Salinas et al. 1996, Wallentinus 2002). Management of this species is particularly difficult since it can establish on a range of substrata and temperature conditions, and because it can disperse via fragments as well as numerous microscopic zoospores. Furthermore, disturbances of native agal assemblages appear to facilitate its dispersal, so relatively indiscriminate disruption (e.g., mechanical removal) could spread an infestation (Valentine and Johnson 2003, 2004). However, in a recent and particularly well-documented southern Tasmanian infestation (Tinderbox Marine Reserve), SCUBA divers performed systematic removal of nearly 5000 sporophytes in monthly efforts spanning 1997–1999 (Hewitt et al. 2005). As with the US Caulerpa taxifolia project, divers’ search and surveillance transects were carefully planned. This project also exemplifies the importance of understanding the reproductive biology of target seaweeds. Thus, removal of sporophytes before releases of zoospores is a strategic action to reduce recruitment. This knowledge was critical in the successful eradication of U. pinnatifida on the metal hull of the trawler Seafresh I, which sank in Hanson Bay, New Zealand in 2000 (Wotton et al. 2005). To contain the infestation, divers removed the two existing sporophytes within days of the sinking, and continued to remove sporophytes (produced from attached, microscopic gametophytes) to prevent further release of spores. Localized heating (708C for 10 min) was eventually used (over one year later) to destroy attached gametophytes. There are other ongoing attempts to understand and curtail the spread of U. pinnatifida in the United States (Thornber et al. 2004), New Zealand (Ministry of Fisheries, Marine Biosecurity New Zealand 2001), and Australia (Sanderson 1990, 1996), but without fully dedicated funding and long-term wellexecuted programs (probably 5 or more years), success is not likely. Notwithstanding limitations of mechanical and hand removal methods, I have listed (Table 2) approaches that are probably most amenable to management or eradication of seaweeds. However, as with any management or eradication strategy, the practicality of each method w104x

will be dictated by the characteristics of the infested environment (see Dunstan and Johnson 2007) and as well as the morphology and phenology of the invasive seaweed. It is also important to note that, due to physical disruption of the benthic habitat by any of these methods, adverse effects on non-target organisms are virtually unavoidable. Therefore, evaluation of these collateral impacts must be thoroughly evaluated in light of any advantages accruing from these ‘‘non-chemical’’ approaches. Biological control Although there are several examples of biological control programs for exotic freshwater macrophytes (e.g., Confrancesco 1998), currently, no successful programs have been reported for marine algae. The most likely candidates for control agents are herbivorous sacoglossan mollusks, or possibly sea urchins such as Strongylocentrotus droebachiensis (OF Muller) (Secord 2003, Sumi and Scheibling 2005). The Mediterranean infestations of Caulerpa taxifolia have generated the greatest interest in this approach due to widespread and dense populations (Meinesz 1999, Coquillard et al. 2000, Thibaut and Meinesz 2000, Thibaut et al. 2001, Zuljevic et al. 2001, Meinesz 2001). Four species of sacoglassans that have received most attention as potential agents for control of C. taxifolia are listed in Table 3. As in all approaches using classical biological control, introduced agents must be sufficiently host-specific, efficacious, and well-suited to the conditions of their new habitat. For example, though Elysia subornata (Verrill), a shell-less species feeds well on two invasive Caulerpa species, its tolerance of cool waters (e.g., Mediterranean Sea) is poor. Similarly, the feeding behavior of Lobiger serradifalci Calcara produces viable fragments of C. taxifolia, thus, making it ill suited for biological control. Davis et al. (2005) examined feeding preferences in some generalist herbivores, including mollusks, sea urchins, and fish assemblages (i.e., Parma microleptis Gunther, P. uifasciata Steindacher and Chromis hypsilepis Gunther). Some feeding trials employed artificial diets such as agar discs containing extracts of Caulerpa filiformis (Suhr) Hering, or Corallina officinalis L. In these tests, Turbo torquatus Gmelin favored the C. officinalislaced discs. When offered mixed diets of live algae, Turbo undulatus Lightfoot (gastropoda) showed a significant preference for Ulva species and Sargassum species and least preference for C. filiformis. The presence of C. filiformis extracts had marginal effects, yet did appear to deter consumption in some cases (e.g., with the sea urchin Helopcidaris tuberculata Lamarck). These authors emphasized the ‘‘low probability that consumers will control these invasive algae’’ due to the low preference for C. filiformis. These were laboratory studies, however, and clearly field conditions may influence grazing behavior (Alcoverro and Mariani 2004). Compared with research on freshwater exotic weed biological control agents, studies of biological control of invasive seaweeds have barely begun. Expanded exploration coupled with adequate host preference testing could yield more candidate agents. Generalist herbivores such as fish may reduce biomass to some degree, but

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Table 2 Potential mechanical methods for control or eradication of invasive seaweeds. Methods

Conditions and constraints

References/Comments

Small-scale diverassisted dredging (e.g., 8–15 cm diameter dredge opening)

Dredged material is either pumped onshore, or into holding/settling tanks, or passed through sufficiently fine mesh to retain algal fragments and other propagules. This is suitable for small to moderate scale infestations (e.g., 0.1–2 ha) that are fairly shallow (less than 5 m depth). Poorly consolidated substrata may preclude this due to high turbidity generated during dredging

Zuljevic 2001, Van Way 2006 (freshwater system), California Department of Agriculture 1990 (freshwater system)

Large-scale hydraulic dredging (e.g., 30–60 cm diameter dredge opening)

Due to scale of equipment, power and disposal needed, this approach is limited to near-shore areas having adequate storage volume (e.g., tanks), or where floating barges can receive dredged spoils. This has potential for clearing large areas in relatively shallow water. However, deployment and logistics make this costly and complicated, and not suited for rough water conditions

Proposed only, no references found for marine systems, though largescale dredging is commonly done in marine harbors and marinas to remove sediment

Rotovation, hydroraking within contained area

These systems usually include an articulated, hydraulically powered rotating head with rigid or flexible tines that are directed along the substrate. Following rotovation, loosened algae would need to have been removed via suction dredges or diverassisted dredges. Unless the site can be contained (e.g., floating curtains or solid walls), this approach will result in spread of propagules and usually creates very poor visibility, which interferes with monitoring and surveillance

Madsen 2000, Newroth 1979, Newroth and Soar 1986. These methods have been used in freshwater systems to remove Myriophyllum spicatum in Canadian lakes, Bugbee and White 2004 (Hydroraking was used in conjunction with herbicides.)

Complete drying of site (plus removal of dried seaweed)

Efficacy requires extensive air-exposure for drying and probably removal of sediment containing the seaweed. This would probably be done as part of a large-scale containment installation that is sufficient to permit pumping out water behind retaining walls. This approach could also be timed to coincide with extremely low tides if the infestation is limited to the upper tidal zones

Proposed for marine systems, Madsen 2000, California Department of Agriculture 1990 (freshwater systems)

Heated or superheated water/steam

The high heat capacity of water makes this method problematic, but if it were coupled with transient de-watering or focused heating of infested surfaces, then high, lethal temperatures could be achieved quickly in localized areas

Wotton et al. 2005

may not provide the requisite selectivity needed to differentially reduce problematic invaders. However, as Secord (2003) correctly points out, there are inherent and largely unknown risks with the introduction of yet another exotic marine species (i.e., a control agent), and nontarget assessments are often limited temporally and spatially. In order to better select and ‘‘match’’ a potential biological control agent, more genetic analyses of invasive seaweeds are needed such as those reported by Provan et al. (2005). These authors used chloroplast microsatellite analysis to determine silmilarities and origins of several invasive populations of Codium fragile spp. tomentosoides Van Goor and found that there have

been at least two independent introductions from native sources. How this might affect host specificity of herbivores is unclear, but certainly needs to be considered. At this juncture, it is prudent to move with extreme care and thoroughly investigate not only host-range, and herbivore/host interactions, but also secondary and tertiary effects resulting from new species introductions. Use of native hebivores (e.g., sea urchins) via augmentative approaches coupled with protective exclosures may offer the most conservative and acceptable opportunities for biological control (Conklin and Smith 2005). In addition, the ability of algal species to produce ‘‘repellent’’ compounds has long been recognized and certainly restricts w105x

w106x

Provenance

Mediterranean

Tropical

Tropical

Mediterranean

Tropical (Hawaii)

Northerntemperate (Nova Scotia)

Temperate (SE Australia)

Tropical

Control agent: (species)

Lobiger serradifalci Calcara (w/shell)

Elysia subornata Verill (w/o shell)

Oxynoe azuropunctata Jensen (w/shell)

Oxynoe olivacea Rafinesque (w/shell)

Tripneustes gratilla (L.) (sea urchin)

Strongylocentrotus droebachiensis (sea urchin)

Turbo undulatus Lightfoot

Various herbivorous fish: Acanthurus blochii Valenciennes, A. triostegus L., Zebrasoma flavescens Bennett, Z. veliferum Bloch, Parma microlepis Gunther, Chromis hypsilepis Gunther

Sargassum species, Caulerpa species Gracilaria salicornia (C. Agardh) E.Y. Dawson, C. taxifolia

Codium fragile ssp. tomentosoides

Kappaphycus species

C. taxifolia

C. taxifolia

C. taxifolia

Caulerpa taxifolia

Target species

Benthic

Pelagic

Benthic

Pelagic

Pelagic

Direct, benthic

Direct, benthic

Pelagic

Larval development

Tropical, 18–26

14–22

Tropical/subtropics, 18–26 8–12

12–25

17–25

17–25, -15 is lethal

17–25

Optimal temperature range for feeding (8C)

Generalists, preferred native: Graciliaria coronopifolia J. Agardh

Preferred Laminaria species, poor feeding in warm waters Generalist Sargassum species

Kappaphycus species

C. prolifera, C. taxifolia

C. taxifolia, C. racemosa

C. prolifera (Forsska˚l) Lamouroux, C. taxifolia C. taxifolia, C. racemosa (Forsska˚l) J. Agardh

Host-specificity (feeding ‘‘preferences’’)

Thallus consumption

Thallus consumption, hold-fasts, Codium as only source did not support gonadal development Thallus consumption

Incisions on all parts, kleptoplasty, stores caulerpenin, high feeding rate (10= others) Incisions on all parts, stores caulerpenin, medium feeding rate Feeds on individual pinnules, low feeding rate Thallus consumption

Incisions lead to holes and fragments, low feeding rate

Feeding characteristics

Paul and Hay 1986, Smith et al. 2004, Davis et al. 2005

Davis et al. 2005

Scheibling and Anthony 2001, Sumi and Scheibling 2005

Conklin and Smith 2005

Thibaut and Meinesz 2000

Thibaut and Meinesz 2000

Coquillard et al. 2000, Thibaut et al. 2001

Thibaut and Meinesz 2000

References

Table 3 Characteristics of some potential biological control agents for invasive seaweeds (modified from Anderson 2002b, T. Thibaut personal communication 2002).

426 L.W.J. Anderson: Control of invasive seaweeds

L.W.J. Anderson: Control of invasive seaweeds 427

the pool of candidate herbivores (e.g., McConnel et al. 1982). Further complicating this approach is the overall uncertainty of future environmental conditions, such as changes in oceanic temperature regimes and related population shifts, both in affecting invasive species and, presumably, their potential control agents (Stachowicz et al. 2002). Chemical control Chemicals used for control of plants and algae are called ‘‘herbicides’’, which are a sub-set of all chemicals used to control or mitigate ‘‘pests’’, more broadly termed ‘‘pesticides’’ (FIFRA, as amended 1974, 1996, 2003). Algicides are generally considered a sub-set of herbicides that are used to kill or reduce the growth of algae. Due to a long history of economic, agricultural, aesthetic and human-health impacts of freshwater algae and freshwater plants (‘‘aquatic weeds’’), there are several aquatic herbicides, including algicides, registered for use by the US Environmental Protection Agency (and similar regulatory entities in other countries). However, the use of chemicals to control seaweeds has been extremely limited both in the USA and in most other countries. In fact, except for products used as marine antifouling boat bottom paints, or surface coatings, there are no algicides registered expressly for use in marine ecosystems in the USA (Don Stubbs, US Environmental Protection Agency personal communication 2005). Even for freshwater aquatic weed control, the available active ingredients are quite limited (Madsen 2000, AERF 2005). Furthermore, among those products cleared for use in freshwater systems, many active ingredients are targeted primarily for control of angiosperms (flowering plants), not algae (e.g., fluridone, 2,4-D, triclopyr, carfentrazone). Freshwater algicides currently registered for use are ‘‘contact’’-type herbicides; that is, they only affect parts of the algae that are directly exposed to the active ingredient. These active ingredients include copper (both inorganic and organo-chelated forms), chlorine (in various forms), diquat, endothal, and acrolein, and ozone (the latter is usually generated on-site). There are also some long chain akyl-type algicides, but these are primarily registered for use as anti-fouling agents in cooling tower systems, or other systems where water is recirculated. Notwithstanding the lack of commercially available marine algicides, recent introduction, establishment and detrimental impacts of invasive seaweeds have prompted attempts to use chemicals to manage, and in some cases, eradicate these infestations. Use of freshwater is included since, like high concentrations of NaCl, its action is to chemically overwhelm physiological mechanisms that maintain normal osmotic balance in marine algae. A summary of these chemicals is provided in Table 4. I have included some laboratory scale studies as well, though there are very few published reports available. Symptoms in Caulerpa taxifolia from exposures to these chemicals vary from complete chlorosis and necrosis and loss of integrity (e.g., chlorine), loss of integrity (freshwater; chelated copper) to conditions where necrosis and chlorosis occur, but fronds may appear to remain intact (e.g., acetic acid). While the physical appearances vary,

these symptoms appear fairly quickly after a few hours (chlorine) to a few days (copper, acetic acid, freshwater). As with most algicides, the onset of symptoms is generally earlier, and effects are more pronounced at higher doses (e.g., Figure 3B,C,D). Field-level efficacy using chemicals has ranged from poor (copper sulfate) to fairly successful (freshwater) to highly effective (chlorine). The variability is due to specific physical conditions within the infested sites, ability to achieve adequate contact time, and availability of resources to continue treatments. Although contact herbicides would be expected to provide the most rapid control of seaweeds, anchored seaweeds may survive treatments if holdfasts, stolons or rhizoids are protected from chemical exposures by bottom sediments or encrusting epiphytes. Thus, a kind of ‘‘chemical mowing’’ might be achieved, but regrowth from viable, anchored parts would necessitate continued re-treatments. However, since some of these anchoring structures such as rhizoids of Caulerpa species can absorb and translocate nutrients, they might also be capable of absorbing and moving systemically active compounds to susceptible points such as active meristematic regions (Williams 1984, Chisholm et al. 1996, Ceccherelli and Sechi 2002). This strategy is used in both freshwater and terrestrial weed control with soil/sediment-active herbicides. The only systemic herbicide that appears to have been examined for control of invasive seaweeds is fluridone (Anderson unpublished data). SonarTM (SePRO Corporation, Carmel, USA), which contains fluridone, has been marketed since the mid-1980s for control of susceptible, weedy freshwater angiosperms and is normally used for 6 to 8 weeks at very low doses (6 to 50 ppb). It is generally not considered an algicide because it has little effect on most problematic freshwater algae. Fluridone inhibits a key enzyme (phytoene desaturase) in the higher plant carotenoid biosynthesis pathway. The pellet formulation sinks to the substratum where it releases the active ingredient over a few days to several weeks, depending upon the formulation. This approach might be quite useful for localized treatments of invasive seaweeds, if they were suceptible. Partial chlorosis was observed at the elongating (distal) fronds in small scale tests using explants of C. taxifolia that were exposed to 50 ppb fluridone for 12 d (Figure 3G). Whether longer exposures could kill this or other macrophytic marine algae is questionable, but is worthy of further investigation, particularly with the pelleted (solid) formulation. However, the need for a very long contact time (several weeks) might preclude the use of fluridone in high-energy zones, unless the infestations could be covered first to help retain effective concentrations of the compound. Probably the three most successful examples to date of field-scale chemical treatments (all to control or eradicate Caulerpa taxifolia) relied on different methods: (1) chlorine (sodium hypochlorite) in the USA (Anderson 2001, 2002a, 2005, Merkel and Woodfield 2003, Southern California Caulerpa Action Team 2006), (2) large applications of granular sodium chloride in New South Wales, Australia (Creese et al. 2004, Glasby 2004, Glasby et al. 2005a), and (3) use of freshwater pumped from the Torrens River into saline ‘‘West Lakes’’ in South Australia (Di w107x

428 L.W.J. Anderson: Control of invasive seaweeds

Table 4 Examples of chemicals used to control marine macroalgae. Chemical

Site

Method of application

Effect

References

Acetic acid

Laboratory

Liquid/direct exposure

Kills fronds

Glasby 2004, Forrest et al. 2004, Anderson (unpublished)

Aluminum

Laboratory

Liquid/direct exposure

Inhibition of photosynthesis

Thake et al. 2003

Chlorine

Field

Liquid (sodium hypochlorite solution), solid tablets (Trichloro-STriazinetrione) Liquid (sodium hypochlorite)

Killed fronds/rhizoids

Anderson 2002a, 2005, Merkel and Woodfield 2003, Williams and Schroeder 2004, Perlich 2004

Laboratory

Copper

Laboratory/field

Liquid/CuSO4 saturated textile mats

Laboratory

Liquid (copper sulfate solution) Liquid chelate (copper alkanolamine complex)

Laboratory

Killed fronds/some rhizoids Inhibition of photosynthesis, killed fronds Partially effective

Uchimura et al. 2000

Killed fronds

Anderson (unpublished)

Creese et al. 2004

Glyphosateq Imazapyr

Laboratory

Direct exposure as liquid formulation to Gracilaria salicornia (C. Agardh) E. Y. Dawson

Negative growth rates

Smith et al. 2004

Glyphosate

Laboratory

Ecklonia radiata (C. Agardh) J. Agardh

Little effect on zoospores

Burridge and Gorski 1998

Diuron Simazine

Laboraory

E. radiata (zoospores)

Little effect

Burridge and Gorski 1998

Furanone

Laboratory

Undaria pinnatifida

Killed gametophyes at )10 mg l-1

Burridge and Gorski 1998

Sea Nine 211 (Rhom Hass)

Laboratory

U. pinnatifida, E. radiata

Killed gametophytes at )1.5 mg l-1

Burridge and Gorski 1998

Salts (NaCl)

Laboratory/field

Granular (50–200 kg/m-2)

Osmotic stress, localized kill

Creese et al. 2004, Smith et al. 2004

Freshwater

Laboratory/field

Direct contact

Osmotic stress, killed fronds/rhizoids

Di Fava 2003, Collings et al. 2004

Fava 2003, Collings et al. 2004). In the USA effort, several large and small colonies that were discovered in 2000 at two separate locations were first contained under vinyl tarpaulins (sealed along the bottom with gravel-filled bags). The Agua Hedionda Lagoon site in Carlsbad, California was approximately 57 ha, but only a small proportion was actually infested (cumulatively less than 0.2 ha). The other USA site, Huntington Harbor, California, contained a patchy distribution of colonies over approximately 2.4 ha. The colonies of C. taxifolia were first contained under PVC (polyvinyl chloride) tarpaulins, then liquid sodium hypochlorite (approximately 12% solution) was injected through ports near the top (Figure 3H). Smaller colonies found later were treated with solid chlorine-generating tablets and immediately covered with vinyl tarpaulins. The tarpaulins were left in place for the duration of the project, but were removed in the seventh year in 2007. To assess the efficacy of these treatments, sediment core samples were removed from beneath w108x

some of the tarpaulins and placed in conditions in a growth chamber that would facilitate growth of viable algal propagules. No regrowth was observed from any of these cores, even though the uninfested sediments removed from the site supported growth of inoculated fronds (Anderson et al. 2005). Following the assessments of the sediment cores, parts of some tarpaulins were removed in situ and recolonization of benthic organisms was noted. As with the removed sediment cores, no growth of C. taxifolia was observed up to five years after initial containment and treatments, though other benthic fauna and flora (e.g., Zostera marina L.) eventually colonized exposed areas. It is important to note that both sites in California were in relatively shallow water (2–4 m) and in highly protected areas where natural, high-energy waves and surge were negligible. These applications, coupled with intensive post treatment surveillance, have eliminated live C. taxifolia for six years and resulted in successful eradication at the two USA sites (Anderson et

L.W.J. Anderson: Control of invasive seaweeds 429

Figure 3 Responses of Caulerpa taxifolia to chemical treatments. Exposed fronds were maintained at 208C under 300 mmol m-2 s-1 cool-white flurosescent light (L:D 14:10). Frond lengths range from 5 to 15 cm. (A) Example of of C. taxifolia exposed to ‘‘household’’ concentration (5% vol/vol) of sodium hypochlorite for 24 h in sea water; (B) unexposed (control) fronds; (C) fronds exposed to 1.0 ppm copper (Cutrine Plus) for 8 days; (D) fronds exposed to 10.0 ppm copper (Cutrine Plus) for 8 days; (E) fronds (on left) exposed to freshwater continuously for 7 days (controls on right); (F) fronds 7 days after initial 90 min exposure to 2% (vol/vol) acetic acid; (G) fronds (left) exposed to 50 ppb fluridone for 12 days (controls on right), arrow indicates typical chlorosis symptoms; (H) example of frame and tarpaulin covering used to contain and treat C. taxifolia with sodium hypochlorite solution in the California infestations (Agua Hedionda) during 2000 and 2001.

al. 2006, Southern California Action Team 2006, California Department of Fish and Game 2006). The efficacy of salt applications in New South Wales, Australia was variable, but in several sites complete kill was achieved after 50 to 200 kg m-2 sodium chloride was applied on top of the colonies (Creese et al. 2004). These authors noted, however, that rates above 50 kg m-2 also resulted in long-term (one to six months at least) negative impacts on some native flora such as Zostera capricorni Aschers. Other benthic fauna (on sediment surfaces and infauna) were also adversely affected by the higher rates of salt application. Transporting and applying large masses of salt required some specialized equipment and logistic support. This probably restricts the utility of this method to relatively shallow waters where the salt can be directed properly onto the target colonies without too much dilution (Glasby 2004).

The replacement of seawater with freshwater in West Lakes, South Australia resulted in excellent kill rates of Caulerpa taxifolia where flushing was complete (Collings et al. 2004). This infestation was discovered in 2002 and consisted of a range of sparse to dense colonies in a fairly narrow tidal-lake system that was amenable to hydraulic control. Mesocosm tests with some herbicides (e.g., copper sulphate, hydrogen peroxide) and altered salinity (e.g., 10, 17, 25, 65, 80 psu) showed that either low (10 to 17 psu) or high salinity should be lethal. The decision was made to dilute the water in West Lakes, since a source of freshwater was relatively close (Torrens River). Over 2000 megaliters of water were pumped into the infested site to reduce the salinity from ;35 psu to less than 10 psu. Because seawater is denser than the freshwater pumped from the Torrens River, it took nearly four months to dilute and displace the deep layers and w109x

430 L.W.J. Anderson: Control of invasive seaweeds

to achieve sufficiently low, lethal salinity near the bottom. During the pumping and freshwater exposure period (ca. five months: July 2003 to December 2003) some efficacy sampling was conducted which showed that none of the C. taxifolia exposed to the low salinity was capable of regrowth (Collings et al. 2004). As expected, other stenohaline, non-target organisms were killed as well. However, recovery of these species over time was anticipated, since adjacent areas where freshwater intrusion was limited should have provided sources for recolonization. Figure 3E shows typical symptoms of C. taxifolia resulting from exposure to freshwater. These examples illustrate that site-specific physical conditions (hydraulic, bathymetric etc.) may allow for effective containment and control actions that might not be practical elsewhere. There are also limitations to the effective use of available algicidal chemicals to manage or eradicate seaweeds. The high levels of salts, coupled with high pH of seawater tend to lower the efficacy of inorganic copper products. Such conditions shift equilibria away from the most active, ionic (dissolved and dissociated) forms of Cu toward complexes such as carbonates, which then become less available to enter target algal tissues. This may be overcome partially by using higher application rates, but costs then become restrictive and there may be potential non-target impacts on fish and invertebrates. Recently Arnold (2005) also pointed out that even moderate levels of dissolved organic carbon (DOC, in ranges from 0.5 to 10 mg C l-1) can increase EC50 values for copper for sensitive species and/or stages such as mussel (Mytilus galloprovincialis Lamarck) embryos and larvae. It is likely that under some conditions of moderate to high DOC (or fronds that are laden with detritus or epiphytes), applications of copper will be less effective. Thake et al. (2003) showed the low pH of cytosol in Caulerpa taxifolia tends to retain metals such as aluminum (and presumably copper) in their more reactive ionic form (e.g., Al3q), thus, leading to disruption of photosynthetic electron transport and probably other physiological reactions. These authors reported total inhibition of photosynthesis following exposures of C. taxifolia fronds to 1 mM aluminum chloride. Thus, it is not surprising that in laboratory exposures, chelated copper was lethal to C. taxifolia fronds with exposures of a few days to 1 or 10 ppm copper (e.g., Figure 3C,D). This response suggests that proper formulations of metals may provide a tool for management and perhaps eradication under conditions that permit adequate duration of exposure (Anderson et al. 1982). Interestingly, some higher plants such as the marine angiosperm Poisidonia oceanica (L.) Delile and freshwater angiosperms such as Potamogeton species are far less sensitive to acute exposures to some metals (Al and Cu, respectively) than are target invasive weeds. As Thake et al. (2003) point out, this selectivity might provide a benefit in releasing populations of less susceptible native plants (or algae?) following successful removal or suppression of the target seaweeds. However, studies are needed to ascertain more precise dose/responses of both target seaweeds and desirable plants, other algae, and non-target animals. Table 5 lists several potential aquatic herbicides that may be efficacious against some target invasive seaw110x

weeds. Though there are certainly many other compounds that may affect seaweed growth, only those with prior US EPA registration for freshwater uses were included since these at least have been reviewed for environmental fate and some non-target effects. It is important to note, however, that additional studies on efficacy must be coupled with thorough assessments of environmental impacts, including non-target organisms as well the dissipation and fate of these chemicals in marine ecosystems. Lack of these data is a major impediment to the development of effective chemical controls for invasive seaweeds. Management by resource limitation One could argue that if critical resources for growth were lacking at a given point of introduction (spatially and temporally), then successful invasions would not occur, or at least would be greatly confined (see Dunstan and Johnson 2007). Indeed, Blumenthal (2005) has argued that both responses to increased resources as well as ‘‘release from enemies’’ should be examined when trying to assess the invasiveness of alien plants. He points out that high-risk species include those freed from coevolved host-specific herbivores, pathogens or competitors, and especially risky are those also able to capitalize on newly available resources. Conversely, insufficient light or nutrients, substratum instability (spatial opportunity), intolerable temperature and salinity ranges are obvious conditions that could preclude establishment of seaweeds. Soluble dyes (e.g., Aquashade ) that attenuate photosynthetically active radiation (PAR) have been used to reduce growth of freshwater macrophytes and algae. Unfortunately, light and most other resources are not easily controlled in marine, near-shore environments, or even within lagoons or small estuaries. Tidal fluctuations, shifting sediments and the large scale of marine systems make these approaches impractical under most circumstances. In fact, just the opposite often occurs as a result of development along coastal zones. Construction of marinas and harbors usually provide low-energy, protected areas as well as increased nutrients, while at the same time, serve as entry points from various pathways of introduction (Bax et al. 2002). Some marinas, aquaculture facilities and other engineered embayments could be manipulated to some extent to reduce light availability and restrict nutrient inputs. These strategies require thoughtful design since they must mitigate impacts from runoff, percolation, sedimentation and other nutrient loading sources and pathways. Thus, land use practices must be integrated into an overall management objective. It is clear, however, that for invasive species such as Caulerpa taxifolia, which have low light requirements (Komatsu et al. 1997), deepening most marinas would be impractical, particularly where annual sediment inputs gradually elevate the bottom. Furthermore, since most marinas and aquaculture facilities have large floating surfaces (docks, nets or screen structures) and artificial depth-gradients that span surfaces to the bottom (e.g., pilings, bulkheads, retaining walls), the opportunity for settlement of algal propagules is high, regardless of depth.

L.W.J. Anderson: Control of invasive seaweeds 431

Table 5 Herbicides with potential for controlling invasive seaweeds (WSSA 2002). Herbicide (trade name)

Active ingredient

Type

Mode of action

Concentration use range (ppm)

Approximate half-life (freshwater)

Aquathol-K/ Hydrothol 191

Endothall

C

Inhibits protein/lipid synthesis, membrane disruption, electron transport uncoupler

0.5–5

3–7 days

Diurex

Diuron

S

Photosynthesis inhibitor, lipid peroxidation

0.1–0.5

90 daysq

Velpar

Hexazinone

S

Photosynthesis inhibitor

0.25

4–5 months

Shark

Carfentrazoneethyl

C

Cell membrane disruption

0.1–0.5

1–5 days

Reward

Diquat

C

Free radical generation, lipid oxidation

0.1–0.37

Few h (water), months (soil); binds to clay (biologically inactive)

Aquazine

Simazine

S

Photosynthesis inhibitor, lipid peroxidation

0.1–0.5

30 daysq

Norosac

Diclobenil

S

Inhibits cell division, inhibits cell wall synthesis

0.5–2.0 (also soil use at 2 to 6– 10 kg ha-1 active ingredient)

30–80 days

Magnacide-H

Acrolein (2-propenal)

C

Disrupts cell membranes (general biocide)

1–15

1–4 h

Chlorine (various forms)

Chlorine gas, sodium hypochloride

C

Oxidant, general bioside

1–15

1–6 h

C: contact type herbicide; S: systemic type herbicide.

Spatial limits to establishment: Limiting or altering ‘‘suitable’’ space for colonization might be considered as a resource management approach. However, what drives spatial competition among native and invasive seaweeds is poorly understood, and even less known are the physical conditions of benthic environments (shapes, textures, elevations, composition) that would predictably favor a native species over an introduced seaweed. At the extremes, such as loose, unstable sand versus wellanchored rock, exclusion of some species could be achieved, but the problem is, of course, the likelihood of excluding desirable native flora as well. Nutrient limitation: Increases in available nutrients from sediments can certainly support or increase algal population biomass to problematic levels. Lapointe (1997) demonstrated that in some cases, sources of increased nutrients include groundwater discharges, which suggests that improved land-based nutrient management might reduce growth of invasive seaweeds. In studies on Kaneohe Bay, Hawaii, Larned (1998) concluded that nutrient sources in low energy protected sites were primarily benthic, whereas growth at higher flow sites relies on continuous water-column inputs. Stimson and Larned (2000) also reported that efflux of dissolved inorganic nitrogen (DIN) in Kaneohe Bay, which supports abundant Dictyosphaeria cavernosa (Forsska˚l) Børgesen, was near-

ly 500 mM m-2 d-1. Pore water levels there were also higher than in less alga-infested sites. In an attempt to assess the role of benthic nutrients and herbivory on species biomass and distribution, Lapointe et al. (2004a) used cages (to exclude or include grazers) placed at increasing distances from high DIN inputs in the Bahamas. They found, using C:N ratios, that biomass was positively correlated with proximity to near shore, high DIN input, but that the composition of species seemed to result from herbivore preferences. In another study in the Florida Keys, Lapointe and coworkers (Lapointe et al. 2004b) showed through analysis of 15N/14N that populations of the red alga Laurencia intricata Lamouroux and the green alga Cladophora catenata (L.) Kuetzing utilized landbased sources of nitrogen. These sources included both large-scale movement from agricultural inputs via the Everglades as well as sewage discharges. They also found a correlation between rainfall events and increased abundance of epiphytic algae on Thalassia testudinum Banks ex Koenig. Taken together, these studies and others suggest that reduction in land-based inputs (e.g., sewage/waste disposal, runoff) could reduce growth of some seaweeds. These are complex problems since they usually encompass effects of large-scale commercial and industrial development, their associated waste-management infraw111x

432 L.W.J. Anderson: Control of invasive seaweeds

structures, as well as increased near-shore private home development. Establishing causal relationships between nutrient sources and observed increases in intertidal and subtidal algal infestations, or increases in algal biomass, can take many years. Reversing these effects can take even longer due to time lags between detection of the nutrient source(s), and implementation of effective alternative treatments. These systems are further complicated if shifts in herbivores occur, which may in turn impede the recovery of native algal populations even after nutrient inputs have been curtailed. Climate change: Further confounding and overriding effects of specific resource availability on the establishment and competitive success of invasive seaweeds, there is now a well-established multi-decade trend in increasing temperature of marine waters (Stachowicz et al. 2002). At least for some invertebrates that have been studied (e.g., acidians), it is clear that the present warming conditions have led to more favorable environments for non-native species compared to native species, and that the critical influences appear to be increased temperature maxima and minima now being experienced at the colonizing sites. Further evidence for the potential expansion of species to formerly ill-suited habitats was provided by Root et al. (2003). These authors showed in a meta-analysis of over 140 studies covering diverse taxa winvertebrates, amphibians, birds, trees, other plants (but not algae)x, that phenological indicators (e.g., onset of reproductive stages) as well as species densities were consistent responses to global warming, particularly in the higher latitudes. Adding more evidence of changing oceanic conditions, Bryden et al. (2005) have recently detected altered circulatory velocities in the northern Atlantic region (Atlantic overturning) that may reduce transport of warm waters to the upper latitudes along European coasts. The net result of all these perturbations seems uncertain, except that benthic organisms would seem to be in line for significantly different future environments. Perhaps there are parallel changes occurring in seaweeds, such as changes in morphology and reproductive modes or timing that foreshadow population shifts. The architectural plasticity exhibited by seaweeds suggests that this is certainly possible (Carruthers et al. 1993, Collado-Vides 2002). How this malleability is manifested in marine communities, including competitive interactions with introduced species, will be interesting to see. Just as alarming, and certainly pertinent to growth of invasive seaweeds, is the increase in atmospheric CO2, which will probably double before the end of this century. How this, and a concomitant reduction in pH and calcium carbonates in marine waters may affect seaweed growth is unsure. It is likely to lead to stresses and reduction in coral cover on tropical reefs (Tanhua et al. 2007, Chem. Eng. News March 16, 2005). Will these changes increase potential habitat for algal species? There is no a priori reason to believe that responses to these temperature and CO2 trends are not occurring in seaweeds, and that tropical species may find suitable conditions in higher latitudes over the next 50 to 100 years. Thus, increases in temperature, coupled with more available carbon (dissolved CO2), regardless of direct land-based nutrient w112x

inputs, could exacerbate the invasiveness of some algal species. Although control and management of invasive seaweeds may be somewhat ameliorated by reductions in nitrogen and phosphorous loads that end up in coastal waters, these approaches will likely be limited to very specific locales, and perhaps, ultimately, be far overshadowed by more widespread global climate and oceanic changes.

Research needs and solutions Worldwide, pest control is a multibillion dollar industry, with a diverse range of markets that include agricultural pests, domestic and commercial structural pests, aquatic (freshwater) and riparian weeds, rangeland weeds, health related pests, and forest pests. It is not difficult to understand how these markets drive research in private industry as well as in various national, state and local agency programs: there is a tremendous profit motive, and successful production of food and fiber depends on efficient, cost-effective pest control methods. For example, in the USA alone, there are over 300 active ingredients registered as herbicides, but only 9 registered as aquatic herbicides, and about 40 active ingredients registered as algicides, although there are no active ingredients registered for control of marine algae (Don Stubbs, US Environmental Protection Agency personal communication). In addition, there are hundreds of non-chemical devices marketed for weed control in non-marine environments. It is clear that, at least for the last 50 years, the economic incentive for development of seaweed control methods (chemical or non-chemical) has been absent. As noted previously, only products aimed at preventing ship hullfouling organisms seem to have generated private support for research. Simberloff et al. (2005) recently pointed out this disparity between the federal funding levels of research for agriculturally related pests compared to pests that primarily impact natural ecosystems. These authors also point out the lack of consistent funding for control actions that could lead to eradication. The reponses to introductions of Undaria pinnatifida noted earlier illustrate the gaps in both strategic responses to, and adequate resources for mitigating invasive seaweeds. The current situation parallels closely circumstances during the late 1940s and 1950s when freshwater weeds began to cause serious impacts and pose threats in various aquatic sites in the USA and other countries. Recognizing the limits of private industry and the lack of market incentive, USA federal and state agencies, and like entities in the UK, Australia, and Canada developed public-supported research programs to develop new, effective control methods for aquatic weeds (Sculthorpe 1967, Gallagher and Haller 1990, Anderson 2003). Today, public agencies still undertake most of the research and provide technology transfer related to aquatic and riparian weed management. Aside from some significant economic losses to the aquaculture industry such as Codium-impaired production and harvests in coastal Canada (Colautti et al. 2006), invasive seaweeds primarily impact large, natural, open

L.W.J. Anderson: Control of invasive seaweeds 433

habitats that are usually under governmental stewardship, and which are not directly associated with a highvalue income-generating activity. Thus, there is almost no support for private, government or academically based research on control methodologies, or even on basic studies on the interactions of invasive seaweeds and native communities. However, compared to conditions in the 1950s, there is now considerably more opportunity to adapt current knowledge, practices and products used for various freshwater angiosperm and algae control to the management and even eradication of invasive seaweed infestations. First, there are many coastal research facilities throughout the world that already have most of the infrastructure and support systems needed to conduct seaweed control research. Research could be supported through in-country (governmental) funds as well as sponsorship through multinational industries (e.g., aquaculture, mariculture, aquarium). In the USA, federal

(e.g., NOAA-Fisheries), EPA, private university and stateuniversity affiliated marine laboratories could house programs aimed at a wide range of control research. Second, there are hundreds of scientists and aquatic plant management specialists who have useful experience with non-chemical (including biological control), herbicidal and fully integrated approaches for controlling freshwater weeds. Third, a network of international, national and regional scientific societies already exists whose focus is on management of freshwater aquatic plants and algae in a wide range of habitats. And, lastly, there already is a foundation of specialized equipment, dozens of candidate herbicides and newly developed surveillance methods (e.g., hydroacoustics), all of which could be evaluated for efficacy, environmental safety and utility in control projects. With a multitude of national and state priorities for marine research, it is reasonable to ask how these

Table 6 Research needed for development of seaweed management and eradication. Research area

Control strategy

Potential method(s)

Non-target effects

Chemical efficacy, translocation ability, fate and metabolism of herbicides in seawater and marine sediments

Block photosynthesis, inhibit transport, uncouple respiration, disrupt electrolyte regulation

Sediment injection, localized pelleted formulations, containment and treatment

Benthic organisms, bottom-dwelling fish (e.g., sculpins), native seagrasses and native algae

Biotic and abiotic reproductive controls

Block dispersal and reproduction capacity

Alter nutrients, chemical exposure (e.g., plant growth regulators), alter photoperiod

Native seagrasses and native algae

Viability of vegetative structures

Prevent dispersal, use of best disposal methods

Cutting, rotovating, scraping, dredging coupled with removal off site

Ability to reestablish natives, revegetation methods

Susceptibility to osmotic stress

Overload osmoregulatory systems

Contain and expose to freshwater or hypersaline conditions

Susceptibility to osmotic shock, recovery conditions

Host-specific herbivores (biological control)

Increase herbivory pressure and impacts

Augmentative releases (culture and release)

Ensure specificity

Natural product effects (allelopathy)

Interfere with growth/reproduction

Augment biomass of allelopathic (native) species

Susceptibility, replanting options

Temperature stresses

Produce lethal conditions, disrupt reproduction

Localized hot water or steam, radio frequency heating

Susceptibilty, re-planting options, selectivity by isolating heated area

Light requirements (quality/irradiance), carbon sources and carbon metabolism

Block photosynthesis, alter carbon availability

Burial with various materials, differentially block carbon sources

Susceptibility of photosynthetic pathways in native algae or seagrasses

Nutrient responses

Deprive essential nutrient(s)

Curtail inputs, sequester nutrients

Differential effects on native algae

Interspecies interactions (competition)

Out-compete invader (space, light, nutrient, canopy)

Facilitate space for native growth, provide conditions/resource for growth

Identify conditions favoring native algae or seagrasses

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434 L.W.J. Anderson: Control of invasive seaweeds

Figure 4 Diagrammatic summary of the key components required for successful response to invasive seaweed infestations. Note: arrows indicate the interaction of all critical inputs and knowledge base.

resources could be directed toward developing seaweed control. In the USA, the importance of, and threat from invasive species was recognized in a Presidential Executive Order No. 13112, signed February 3, 1999, during the Clinton administration. Over the past eight years, implementation of that directive has begun to result in ‘‘cross-cutting’’ federal agency budgets that can address particular invasive species needs. What has been lacking is a strategic plan that truly focuses federal (and state) agency resources and expertise on invasive species, like seaweeds, that have not received adequate attention to date. Any such plan needs to provide for a rational risk and benefit assessment as well (Maguire 2004.) Specifically, I have compiled a list of research needs or gaps, based upon the kind of information that has been useful in developing technologies for the management of freshwater weeds (Table 6). Ideally, a fully vetted review of research priorities could be generated through one or more workshops with participants from several disciplines, including phycology, marine ecology, marine fisheries, and scientists working on freshwater plant/algae management and marine pest management. Figure 4 provides a simple graphic representation of how these disciplines must eventually interact to provide integrated management and eradication approaches.

Summary and conclusion The ability to control invasive seaweeds will become more important as both the pathways of introduction expand along with increased ship movements and other near-shore activities, and because habitat conditions are changing in ways that will often favor success of new introductions (Thresher et al. 1999, Carlton 2001, Buschbaum et al. 2006, Walters et al. 2006). Coastal human populations worldwide have been increasing tremendously over the past 100 years and this trend will continue. With these increases, inadvertent introductions by aquarium hobbyists will probably increase as well. Population growth will also bring with it higher land-based w114x

nutrient loading in some coastal areas, and this will surely accelerate growth and spread of not only exotic algae, but some native species as well. Further compounding these events, global changes in seawater temperature maxima and minima appear to be allowing range extensions of some sessile species, and these will likely include macrophytic algae, or strains that are better suited to the new conditions. The fact that some early responses to these types of invasions have been successful is particularly encouraging in that there has not yet been any concerted, systematic effort to develop technologies expressly for control and eradication of invasive seaweeds. To meet the inevitable increase in invasion episodes, it will be essential to develop well-coordinated national and statesponsored research projects that focus on identifying likely invasive species and habitat characteristics that facilitate invasions, as well as a range of effective and safe control and eradication technologies.

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Silva, P.C., R.A. Woodfield, A.N. Cohen, L.H. Harris and J.H.R. Goddard. 2002. First report of the Asian kelp Undaria pinnatifida in the northeastern Pacific Ocean. Biol. Invasions 4: 333–338. Simberloff, D., I.M. Parker and P.N. Windle. 2005. Introduced species policy, management, and future research needs. Front. Ecol. Environ. 3: 12–20. Smith, C. and L. Walters. 1999. Fragmentation as a strategy for Caulerpa species: fates of fragments and implications for management of an invasive weed. PZSN 1. Mar. Ecol. 20: 307–319. Smith, J.E., C.L. Hunter, E.F. Conklin, R. Most, T. Sauvage, C. Squair and C.M. Smith. 2004. Ecology of the invasive red alga Gracilaria salicornia (Rhodophyta) on O’ahu, Hawi’i. Pacific Sci. 58: 325–343. Southern California Caulerpa Action Team (SCCAT). 2006. Final report on eradication of the invasive seaweed Caulerpa taxifolia from Agua Hedionda Lagoon and Huntington Harbour, California. pp. 88 (access via http://sccat.net). Stachowicz, J.J., J.R. Terwin, R.B. Whitlatch and R.W. Osman. 2002. Linking climate change and biological invasions: Ocean warming facilitates nonindigenous species invasions. Proc. Nat. Acad. Sci. 99: 15497–15500. Stimson, J. and S.T. Larned. 2000. Nitrogen efflux from the sediments of a subtropical bay and potential contributions to macroalgal nutrient requirements. J. Exp. Mar. Biol. Ecol. 252: 159–180. Sumi, C.B.T. and R.E. Scheibling. 2005. Role of grazing by sea urchins Strongylocentrotus droebachiensis in regulating the invasive alga Codium fragile spp. tomentosoides in Nova Scotia. Mar. Ecol Prog. Ser. 292: 203–212. Tanhua, T., A. Kortzinger, K. Friis, D.W. Waugh and D.W.R. Wallace. 2007. An estimate of anthropogenic CO2 inventory from decadal changes in oceanic carbon content. Proc. Nat. Acad. Sci. 104: 3037–3042. Thake, B.L. Herfort, M. Randone and G. Hill. 2003. Susceptibility of the invasive seaweed Caulerpa taxifolia to ionic aluminium. Bot. Mar. 46: 17–23. Thibaut, T. and A. Meinesz. 2000. Are the Mediterranean ascoglossan molluscs Oxynoe ulivacen and Lobiger serradifalci suitable agents for a biological control against the invading tropical alga Caulerpa taxifolia? C.R. Acad. Sci. Paris. 323: 477–488. Thibaut, T., A. Meinesz, P. Amade, S. Charrier, K. De Agnelis, S. Ierardi, L. Mangialazo, J. Melnick and Vidal. 2001. Elysia subornata (Molusca) a potential control agent of the alga Caulerpa taxifolia (Chlorophyta) in the Mediterranean Sea. J. Mar. Biol. Ass. U.K. 81: 497–504. Thornber, C.S., B.P. Kinlan, M.H. Graham and J.J. Stachowicz. 2004. Population ecology of the invasive kelp Undaria pinnatifida in California: environmental and biological controls on demography. Mar. Ecol. Prog. Ser. 268: 69–80. Thresher, R.E. and A.M. Kuris. 2004. Options for managing invasive marine species. Biol. Invasions 6: 295–300. Thresher, R.E., C.H. Hewitt and M.L. Campbell. 1999. Synthesis: introduced and cryptogenic species in Port Phillip Bay. In: (C.L. Hewitt, M. Campbell, R.E, Thresher and R.B. Martin, eds) Marine biological invasions of Port Phillip Bay, Victoria. Centre for Research on Introduced Marine Pests Tech. Report No. 20. Hobart, CSIRO Marine Research. pp. 283–295.

Trowbridge, C.D. 1999. An assessment of the potential spread and options for control of the introduced green macroalga Codium fragile ssp. tomentosoides on Australian shores. CSIRO Marine Research Report. pp. 43. Uchimura, M., A. Ribal, A. Mato, R. San deaux, J. Sandeaux and J.C. Baccou. 2000. Potential use of Cu2q, Kq and Naq for the destruction of Caulerpa taxifolia: Differential effects of photosynthetic parameters. J. Appl. Phycol. 12: 15–23. U.S. Fish and Wildlife Service. 2005. National management plan for the genus Caulerpa. pp. 135. Valentine, J. and C.R. Johnson. 2003. Establishment of the introduced kelp Undaria pinnatifida in Tasmania depends on disturbance to native algal assemblages. J. Exp. Mar Biol. Ecol. 295: 63–90. Valentine, J.P. and C.R. Johnson. 2004. Establishment of the introduced kelp Undaria pinnatifida following dieback of the native macroalga Phyllospora comosa in Tasmania, Australia. Mar. Freshw. Res. 55: 223–230. Wallentinus, I. 2002. Introduced marine algae and vascular plants in European aquatic environments. In: (E. Leppa¨koski, S. Gollasch and S. Olenin, eds), Invasive aquatic species of Europe: distribution, impact and management. Kluwer Academic Publishers, Dordrecht. pp. 27–52. Williams, S.L. 1984. Uptake of sediment ammonium and translocation in a marine macroalga Caulerpa cupressoides. Limnol. Oceangr. 29: 374–379. Williams, S.L. and S.L. Schroeder. 2004. Eradication of the invasive seaweed Caulerpa taxifolia by chlorine bleach. Mar. Ecol. Prog. Ser. 272: 69–76. Wotton, D.M. and C.L. Hewitt. 2004. Marine biosecurity postborder management: developing incursion reponse systems for New Zealand. N. Z. J. Mar. Freshw. Res. 38: 553–559. Wotton, D.M., C. O’Brien, M.D. Stuart and D.J. Fergus. 2005. Eradication success down under: heat treatment of a sunken trawler to kill the invasive seaweed Undaria pinnatifida. Mar. Poll. Bull. 49: 844–849. Wright, J.T. 2005. Differences between native and invasive Caulerpa taxifolia: a link between asexual fragmentation and abundance in invasive populations. Mar. Biol. 147: 559–569. Wright, J.T. and A.R. Davis. 2006. Demographic feedback between clonal growth and fragmentation in an invasive seaweed. Ecology 87: 1744–1754. WSSA (Weed Science Society of America). 2002. Herbicide handbook, 8th edition. WSSA Publications, Lawrence, KS. pp. 493. York, P.H., D.J. Booth, T.M. Glasby and B.C. Pease. 2006. Fish assemblages in habitats dominated by Caulerpa taxifolia and native seagrasses in south-eastern Australia. Mar. Ecol. Progr. Ser. 312: 223–234. Zuljevic, A. 2001. Pojava i sirenje vrste Caulerpa taxifolia (Vahl) C. Agardh (Caulerpales, Chlorophyta) us Jadranu. M. Sc. Thesis, University of Zagreb, Zagreb. pp. 134. Zuljevic, A. and B. Antolic. 2000. Synchronous release of male gametes of Caulerpa taxifolia (Caulerpales, Chlorophyta) in the Mediterranean Sea. Phycologia 39: 157–159. Zuljevic, A., T. Thibaut, H. Elloukal and A. Meinesz. 2001. Sea slug disperses the invasive Caulerpa taxifolia. J. Mar. Biol. Ass. U.K. 81: 343–344. Received 29 December, 2005; accepted 2 August, 2007

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Botanica Marina 50 (2007): 438–450

2007 by Walter de Gruyter • Berlin • New York. DOI 10.1515/BOT.2007.046

Review

Invasive seaweeds: global and regional law and policy responses

Meinhard Doelle*, Moira L. McConnell and David L. VanderZwaag Marine & Environmental Law Institute and Dalhousie Law School, Dalhousie University, Halifax, NS, Canada, B3H 4H9, e-mail: [email protected] * Corresponding author

Abstract We consider law and policy responses to invasive seaweeds at global and regional levels. Key global regimes considered include the 1982 United Nations Convention on the Law of the Sea, the Convention on Biological Diversity, the Ramsar Convention and the Bonn Convention on Migratory Species. Contributions from the Food and Agriculture Organization and the International Maritime Organization are also considered in the global context. At a regional level, examples of efforts in North America and Europe are offered to illustrate challenges and opportunities for regional responses to invasive seaweeds. We conclude with law and policy recommendations, most notably the need to approach the issue of invasive seaweeds in a manner consistent with the precautionary principle. Keywords: global; law and policy; precaution; regional.

1. Introduction As demonstrated in a number of the articles in this Special Issue, the introduction of alien invasive species poses one of the most serious threats to both terrestrial and marine biodiversity. In fact, habitat loss, climate change, and alien invasive species are generally considered to top the list of biodiversity threats. Concern about invasions is not limited to biodiversity per se but extends to its broader socio-economic impacts on agriculture, forests, fisheries, aquaculture, and other human activities dependent on the stability of living resources in a particular ecosystem. As a result, invasive species pose almost incalculable economic, socio-cultural and human health security risks. Estimates of the cost of responding to this problem around the globe vary widely. One estimate of the cost to the US economy is US$137 billion per year (Murray et al. 2004). Although concern about the issue of introduction of alien species was evident in the late 1970s, the scope of the problem only gained widespread attention of law and w118x

policy makers in the 1990s. Most of the effort in policy development to date has been on terrestrial invasive species, particularly in relation to intentional introductions, and has taken the form of border control or quarantine measures. On the aquatic side, attention appears to have been focused more on intentional introductions of fish species and on specific sectoral pathways or vectors for the unintentional transfer of these and other species including pathogens. To date, however, law and policy efforts have tended to be generic and focused on managing pathways rather than on the problems posed by particular species or organisms, such as seaweeds. Although these efforts in part impact on the problem, invasive seaweeds nevertheless pose a serious threat in the context of unintentional transfer and introduction through shipping, aquaculture, the aquarium trade, fishing activities, and the opening of new canals and waterways. Some of these activities also involve intentional transfer of invasive seaweeds. Here we provide an overview of law and policy responses to aquatic alien invasive species generally and invasive seaweeds more specifically. We are primarily concerned with the, arguably, more difficult regulatory problems posed by unintentional pathways for species transfer and introduction. While a comprehensive analysis of law and policy responses at the global and regional level is not possible here, we describe key global and selected regional efforts to deal with invasive species to demonstrate the challenge of developing a comprehensive and coordinated response to the threat of invasive species. Sectoral and selected regional and sub-regional efforts are also highlighted. In the process, the fragmented nature of the response at the global and regional level is exposed. Finally, national responses are not considered. However, it should be noted that states such as Australia and New Zealand are recognized as leaders in terms of domestic law and policy responses (Doelle 2003, Hewitt et al. 2004). As explained in the other articles in this Special Issue, terminology varies greatly in both law and academic literature on this topic, with the terms alien species, alien invasive species, non-native species, non-indigenous species, or harmful aquatic organisms used, often interchangeably. Here we use the terms ‘‘invasive species’’ or ‘‘invasive seaweeds’’ unless the context requires otherwise. Similarly, we do not attempt to distinguish between various species but simply use the term ‘‘seaweed’’. In section 2 we point out that at least two global regimes provide an international framework for States to take action preventing the introduction and spread of aquatic invasive species, viz. the 1992 Convention on

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Biological Diversity (CBD), and the 1982 United Nations Convention on the Law of the Sea (UNCLOS). The two main sectoral responses at an international level to the problem of transfer and introduction of invasives as an unintentional consequence of ships’ operations (ballasting and hull fouling) and fisheries and aquaculture are also outlined. An overview of regional responses set out in section 3 is followed with a generic discussion of what could be done at the national level to implement the principles developed internationally as the basis of an effective response to the threat. We argue that even though the global and regional law and policy responses are, as yet, neither comprehensive nor coordinated, this problem needs not be replicated at a national level. It proposes the adoption of a precautionary and integrated approach to regulatory design and implementation at a national level to address the problem of invasive species, including invasive seaweeds. Precaution has emerged as a key principle in the response to the problem of invasive species at both the global and regional level. Consistent with precaution, there has been a general recognition internationally that prevention is the best and, in many cases, the only effective defense against invasions: eradication, control and containment are risky at best, in most cases very costly and, more often than not, ineffective. The implications of building a national response on these principles are explored in the last part of this article. This article is, therefore, based on the view that the harm arising from this problem, including both socioeconomic and biodiversity impacts and costs, are such that prevention-oriented strategies are the best regulatory approach. In cases of doubt, precaution and prevention rather than eradication or containment or control is the way to deal with activities that are likely to be a vector or pathway for the introduction of invasive species. Note: As the law in this area is constantly changing, we would like to emphasize that the materials are current up to 10 January, 2006, the date of submission of this article.

2. Global responses to invasive organisms such as seaweeds While numerous international instruments and institutions are concerned with invasive species (McNeely et al. 2001), this chapter briefly discusses four of the main global agreements targeting invasives and highlights sectoral attempts by the Food and Agriculture Organization (FAO) and the International Maritime Organization (IMO) to address the problem of introductions of alien species in the contexts of fisheries and aquaculture, and shipping, respectively. 2.1. Four main global agreements The four global agreements of special relevance to controlling invasive marine species fall into three sub-categories. UNCLOS, which, inter alia, addresses the issue of State obligations to protect and preserve the marine environment from pollution from both land and ocean-

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based activities, provides the general legal framework for addressing the problem of marine invasive species. The 1992 CBD casts a wide net to address the impacts of introduction of ‘‘alien species’’ warticle 8(h)x on broad biodiversity protection. The 1971 International Convention on Wetlands of International Importance especially as Waterfowl Habitat (Ramsar Convention) and the 1979 Convention on Migratory Species of Wild Animals (Bonn Convention) tangentially aim to curb the introduction of alien species in the context of specific area and species conservation. 2.1.1. General legal framework Considered the ‘‘Constitution of the Oceans’’, UNCLOS, with its 320 articles and nine annexes, devotes just one article specifically to the problem of ‘‘introduction of species, alien or new’’. Article 196(1) states: ‘‘States shall take all measures necessary to prevent, reduce and control pollution of the marine environment resulting from the use of technologies under their jurisdiction or control, or the intentional or accidental introduction of species, alien or new, to a particular part of the marine environment, which may cause significant and harmful changes thereto.’’(emphasis added)

There has been some debate as to the precise meaning of this provision, which is found in Part XII of UNCLOS, the part that deals with preservation of the marine environment. The question is whether on a strict reading it can be interpreted as meaning that the introduction of potentially harmful alien species is pollution of the marine environment, or whether this constitutes some other category of environmental harm (McConnell 2002, Firestone and Corbett 2005). Article 1 of UNCLOS defines ‘‘pollution of the marine environment’’ as: ‘‘(4) w«x the introduction by man, directly or indirectly, of substances or energy into the marine environment, including estuaries, which results or is likely to result in such deleterious effects as harm to living resources and marine life, hazards to human health, hindrance to marine activities, including fishing and other legitimate uses of the sea, impairment of quality for use of sea water and reduction of amenities.’’

Although this distinction may have some implications for national level regulatory responses, particularly in connection with international shipping, the obligation, under Article 196(1), on Parties to UNCLOS to take steps to address the problem of transfer and introduction of alien or new species into the marine environment is clear. In addition to this specific provision on species introduction, UNCLOS imposes a general duty on all States to protect and preserve the marine environment (Article 192). This obligation includes a duty to prevent pollution of the marine environment and to protect and preserve rare or fragile ecosystems as well as the habitat of depleted, threatened or endangered species and other forms of marine life from all sources of pollution wArticle 194(1) (5)x (Firestone and Corbett 2005). An environmental impact assessment requirement is also set out in Article 206 for activities, which may cause pollution or significant and harmful changes to the marine w119x

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environment. This language may capture intentional introductions of seaweed for cultivation: ‘‘When States have reasonable grounds for believing that planned activities under their jurisdiction or control may cause substantial pollution of or significant and harmful changes to the marine environment, they shall, as far as practicable, assess the potential effects of such activities on the marine environment w«x.’’

It may also have implications for assessing the effect of any approaches adopted to address unintentional introductions. The use of the term ‘‘may’’ emphasizes the need to take action where a material risk is indicated. The Convention also establishes a continuing obligation for States to protect and preserve the marine environment through global and regional cooperation, which might be the foundation for filling regulatory gaps in addressing invasive seaweeds. Article 197 provides: ‘‘States shall cooperate on a global basis and, as appropriate, on a regional basis, directly or through competent international organizations, in formulating and elaborating international rules, standards and recommended practices and procedures consistent with this Convention, for the protection and preservation of the marine environment, taking into account characteristic regional features.’’

If indeed introduction of alien species (under Article 196) can be considered a form of pollution, then numerous other UNCLOS provisions would apply (McConnell 2002, Firestone and Corbett 2005). For example, Article 194(2) includes an obligation regarding protection of the environment of other States. ‘‘Article 194 Measures to prevent, reduce and control pollution of the marine environment w«x 2. States shall take all measures necessary to ensure that activities under their jurisdiction or control are so conducted as not to cause damage by pollution to other States and their environment, and that pollution arising from incidents or activities under their jurisdiction or control does not spread beyond the areas where they exercise sovereign rights in accordance with this Convention.’’11

States might also be obliged to notify other States of the potential for invasive seaweeds to cause transboundary damage (Article 198). States would be urged to develop joint contingency plans for responding to seaweed invasions (Article 199). States would have an obligation to ensure that alien species that may/can harm do not spread beyond areas of national jurisdiction wArticle 194(2)x. States might also be liable for transboundary invasions of invasive species (Article 235). 2.1.2. Biodiversity protection The CBD, calling for preventative and precautionary approaches to addressing the causes of biodiversity losses, has one provision specifically aimed at addressing the issue of introduced invasive species or alien species, as they are described Similar wording as to state responsibility is found in Article 3 of the CBD. An Expert Group has linked this obligation to invasive species (Subsidiary Body on Scientific, Technical and Technological Advice 2005). 1

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in the CBD (and in later discussion, ‘‘marine alien species’’). Article 8(h) calls on Parties to ‘‘prevent the introduction of, control or eradicate those alien species which threaten ecosystems, habitats or species.’’ To the extent that it deals specifically with marine biodiversity, the CBD can be seen as building upon and elaborating the State obligations set out in UNCLOS concerning conservation and preservation of the marine environment. Article 22 (2) of the CBD specifically notes this relationship: ‘‘Contracting Parties shall implement this Convention with respect to the marine environment consistently with the rights and obligations of States under the Law of the Sea.’’

The Conference of the Parties (COP), having agreed in 1995 to a program of action called the Jakarta Mandate on Marine and Coastal Biological Diversity, in 1998 adopted a program of work with one of five thematic areas being devoted to invasive alien species (CBD Secretariat, undated). The program of work has become a ‘‘living tree’’ spawning numerous initiatives to address the threat of introduction of alien species to ecosystems. For example, in Decision VI/23 (2002) on alien species that threaten ecosystems, habitats or species, the COP urged Parties to develop national invasive alien species strategies and action plans and to develop regional strategies where appropriate. The Decision also adopted, through an Annex (Guiding Principles for the Prevention, Introduction and Mitigation of Impacts of Alien Species That Threaten Ecosystems, Habitats or Species), setting out 15 guiding principles for the prevention and mitigation of impacts of alien species. Besides embracing the precautionary approach and ecosystem approach for dealing with invasive species, the Guiding Principles urge a three-stage hierarchical approach involving prevention as a first priority, followed by eradication and containment. The Ad Hoc Technical Expert Group on Gaps and Inconsistencies in the International Regulatory Framework in Relation to Invasive Alien Species (AHTEG), established at the request of the Conference of the Parties in Decision VII/13 (2004), met in New Zealand in May 2005 and issued an informative report (Subsidiary Body on Scientific, Technical and Technological Advice 2005). Inadequate national implementation of international obligations and limited national capacity were identified as key impediments for addressing the introduction and spread of invasive alien species (paragraphs 17, 35). The AHTEG emphasized the need for capacity building efforts including technology transfer and training in relation to invasive species. The Expert Group noted that liability regimes for damages caused by invasive species may be an important issue and recommended that the issue be raised at the Experts Meeting on Liability and Redress under the Convention scheduled for October 2005.2 The Group of Legal and Technical Experts on Liability and Redress in the Context of Paragraph 2 of Article 14 of the Convention, met from 12 to 14 October 2005 in Montreal and while finding it premature to recommend development of a liability protocol, the Group initiated discussions about the numerous legal issues including how to define damage to biological diversity and how to value damage to biodiversity (CBD 2005). 2

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The AHTEG urged Parties to take seriously their responsibilities under Article 3 of the Convention, that is, not allowing activities within their jurisdiction or control to cause damage to the environment of other States or areas beyond national jurisdiction. Various export controls were listed as examples, including notifying potential importing countries about particular species that may be invasive and prohibiting the export of some species. The AHTEG highlighted the many specific gaps and inconsistencies in the international regulatory framework. While the COP, through Decision VII/5 (2004) on marine biological diversity, recommended that Parties use native species and subspecies in aquaculture, the Expert Group noted there are no specific binding international requirements addressing impacts, including transboundary impacts, regarding the use of alien species in aquaculture or in relation to the problem of transfer of invasive species through ships, fouling of ships and on other equipment and vessels. The exclusion of certain ships from IMO regulatory treaties was also viewed as a gap, as were other potential pathways for the introduction of invasive species including scientific research, tourism and the aquarium trade. 2.1.3. Specific area and species conservation The Ramsar Convention was signed in Ramsar, Iran in 1971 and has been amended by two Protocols (in 1982 and in 1987) (Ramsar Convention 1971). It committed Parties to establish nature reserves on wetlands and to include at least one wetland on the List of Wetlands of International Importance. It has largely addressed the issue of invasive species, through hortatory resolutions. Resolution VII.14 on invasive species and wetlands, adopted at the 7th Conference of Contracting Parties in 1999, urged Parties to: prepare inventories of alien species in wetlands; establish control and eradication programs; review existing legal and institutional measures relating to invasive species control; and, where necessary, adopt legislation and programs to prevent the introduction of ‘‘new or environmentally dangerous alien species’’ (Ramsar Conference 1999). Resolution VIII.18, adopted at the 8th Conference of the Parties in 2002, urged Parties to: undertake risk assessments of alien species which may pose a threat to the ecological character of wetlands; identify the presence of invasive alien species in wetlands and for listed sites to report to the Ramsar Bureau the nature of the invasion; and cooperate in preventing and controlling transboundary invasions including those in shared coastal/marine zones (Ramsar Conference 2002). The 1979 Bonn Convention is aimed at protecting endangered migratory species (Bonn Convention, Appendix I) and listed migratory species having an unfavorable conservation status or which would significantly benefit from international cooperation (Bonn Convention, Appendix II). The Convention establishes quite general obligations on Parties in relation to invasive species. Pursuant to Article III (4)(c) of the Convention, Range States (of a migratory species listed in Appendix I as endangered) are required to endeavor: ‘‘To the extent feasible and appropriate to prevent, reduce or control factors that are endangering or are likely to further

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endanger the species, including strictly controlling introduction of, or controlling or eliminating, already introduced exotic species.’’

‘‘Range States’’ are defined in the Convention to include states that exercise jurisdiction over any part of the range of the migratory species and states whose flag vessels are engaged outside national jurisdictional limits in taking the migratory species. For migratory species listed in Appendix II, Range State Parties are encouraged to conclude Agreements to restore or maintain migratory species at a favorable conservation status. In particular agreements are encouraged to include provision for protection of habitats from disturbances, ‘‘including strict control of the introduction of, or control of already introduced, exotic species detrimental to the migratory species’’ wArticle V(5)(e)x. The potential impact of invasive seaweeds on the ecology of habitats of protected species would raise concerns relevant to this obligation. 2.2. Global sectoral initiatives 2.2.1. The FAO and invasives in the fisheries and aquaculture sectors The FAO has accepted a Code of Conduct for Responsible Fisheries (FAO 1995), a non-binding document setting out principles and standards for fisheries and aquaculture practices. It directly addresses the introduction of non-native species only in the context of aquaculture and intentional introductions. Article 9.2.3 proposes that States adhering to the Code consult with the neighboring States before introducing non-indigenous species into transboundary aquatic ecosystems. Article 9.3.1 urges States to minimize the harmful effects of introducing non-native species or genetically altered stocks used for aquaculture. The Code may, however, be relevant to the issue of intentional and unintentional introductions of invasive seaweeds through its general habitat protection and research exhortations. Article 6.8 urges States to protect and rehabilitate all critical fisheries habitats in marine and freshwater ecosystems. Article 12 calls for strengthening national research capacities and for assessing the impacts of environmental changes on fish stocks and aquatic ecosystems. Perhaps of greater relevance in addressing invasive species than the Code of Conduct itself is the development of the non-binding Technical Guidelines on the Precautionary Approach to Capture Fisheries and Species Introductions (FAO 1996). The Guidelines equivocate over what a precautionary approach should mean to species introductions. After noting in paragraph 104 that a strictly precautionary approach would not permit deliberate species introductions and would require strong measures to prevent unintentional introductions, the Guidelines proceed to suggest moderated versions of the precautionary approach, especially for deliberate introductions of species. Rather than prohibiting intentional new species introductions, the Guidelines suggest ‘‘controlled introductions’’ where proponents of introductions shall be required to demonstrate caution through following the non-binding ICES Code of Practice on the Introduction and Transfer of Marine Organisms, or a similar code, where a risk assessment approach is followed and w121x

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pilot/experimental introductions supported. For unintended introductions, for example, by ballast water discharge, aquarium trade, or biologically contaminated fishing gear, the Guidelines recommend that authorities should ‘‘establish regulations to reduce these risks, commensurate with the severity of potential adverse impacts’’ (paragraph 129). The Guidelines also suggest the development of an accessible database on ballast or fouling organisms that have an impact upon fisheries and the creation of a network of experts charged with identifying introduction problems and areas of impact (paragraph 131). The Guidelines encourage the development of effective non-biocidal antifouling paints or treatments to reduce the risk of introduction from ship fouling (paragraph 137). This approach can, in part, be seen as complementary to the international obligations in the Anti-fouling Convention, discussed in the next section. 2.2.2. The IMO and the regulation of shipping as a vector for invasive species Shipping has been recognized as one of the primary vectors for the national and international transfer of ‘‘harmful aquatic organisms’’, which are defined in Article 1, paragraph 8, of the International Convention on the Control and Management of Ship’s Ballast Water and Sediment 2004 (BWM Convention 2004), as: ‘‘w«x aquatic organisms or pathogens which, if introduced into the sea including estuaries, or into fresh water courses, may create hazards to the environment, human health, property or resources, impairment of biological diversity or interfere with other legitimate uses of such areas.’’

This vector will be partly regulated at an international level by the BWM Convention 2004, when it comes into force3, and, when finalized, the associated Guidelines for implementation (IMO, Guidelines 2004). This Convention, which took many years to negotiate, is under the auspices of the IMO. The BWM Convention 2004 is based, in part, on a 1997 Resolution of the IMO General Assembly, A. 868(20), Guidelines for the Control and Management of Ships’ Ballast Water to Minimize the Transfer of Harmful Aquatic Organisms and Pathogens. It is part of a web of regulatory Conventions, such as the International Convention for the Prevention of Pollution from Ships 73/78, more commonly known as MARPOL 73/78, developed by the IMO member States, to meet its mandate of securing ‘‘safer shipping and cleaner seas’’ by regulation of the shipping industry through the implementation of standards for many of the flag State obligations found under UNCLOS (Articles 91, 94). Under international law, and as codified by UNCLOS, a ‘‘flag State’’4 has the primary responsibility for regulating the operation of ships flying its flag, irrespective of where they operate in the world. On the ratification by 30 States whose combined merchant fleets constitute not less than thirty-five percent of the gross tonnage of the world’s merchant shipping fleet. It is implemented through a multifaceted phase in system for the standards linked to the date of construction (before or 2009 or 2012) and the size of the ship’s tanks (Annex, Section B, Regulation B-3). 4 A ‘‘flag State’’ is generally the State in which a ship is registered (although a ship can be on more than one registry, it can only fly one flag at a time). 3

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The BWM Convention 2004 is one of the newer generation of IMO regulatory Conventions, in that it includes more coastal State responsibilities and rights in the regulation of this sector.5 The regulatory system set out in the BWM Convention 2004 adopts the formula that is now well established in many Conventions developed by the IMO to help prevent many different ship-based sources of marine pollution. This involves a combination of initial and on-going compliance inspections, certification, coastal zoning and alternative discharge options. It is an important step forward in addressing concerns about the introduction of potentially harmful organisms (including invasive species) by transport in ballast water and in sediment in the bottom of ships’ ballast water tanks. It can be seen as a specific delineation of State obligations under both UNCLOS and the CBD, discussed above, to adopt and implement national laws to address this problem. However, it is suggested that it may not be an effective regulatory response to either prevention of or reducing the risk of spread of invasive seaweeds by ships. A recent study of invasive seaweeds on the Pacific Coast of North America (Murray et al. 2004, pp. 1, 2) noted that: ‘‘w«x the major pathways or vectors of introducing marine NIS wnon-indigenous speciesx into non-native waters are: 1) shipping transport, either in ballast water or as hull fouling organisms; 2) aquaculture enterprises, either as targeted species or as unintentional hitch-hiker associates w«x Besides canals and waterways, the most significant vectors for seaweed NIS appear to be: 1) aquaculture w«x; 2) shipping, mostly as fouling organisms attached to hulls and other ships’ parts; w«x Seaweeds appear to be much less likely than other marine NIS to be introduced through the discharge of ballast water but are very likely to be moved along the coast as fouling organisms on ships’ hulls or other marine gear.’’ wemphasis addedx

This comment highlights a regulatory gap, largely related to the dynamics of international institutions and lawmaking. It means that not all aspects of shipping that can serve as vectors for the unintentional introduction of harmful organisms are regulated under the BWM Convention 2004. As the foregoing quote indicates, potentially harmful organisms, such as invasive seaweeds, are transferred between countries in other ways related to ships’ operations. These include attaching to the ship’s hull (a process called fouling), the ship’s sea chest (Cawthron 2004), attaching to the anchor and other parts of a vessel as well as cargo, cargo packaging and loading equipment. These may be considered, along with ballast water, as unintentional transfers in the sense that they are by-products of the operation of shipping, as distinct from intentional transfers of alien species, albeit on board a ship. Concerns have been expressed about these other unintentional or operational vectors in various fora, but so far there is no specific international regulatory develThe term ‘‘coastal State’’ is a shorthand term usually used to refer to the interest of a State in protecting its coastline and resources in the waters under its jurisdiction. States often have a ship owning/regulation interest (flag State) and a coastal State interest (coastal and marine environment and resource management) and a port State role (a specific regulatory/enforcement role vis a vis the other two interests). 5

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opment.6 For example, in 2000 concern was expressed about this piecemeal, gap-filling approach to dealing with related issues in the meetings relating to the CBD (e.g., Invasive Alien Species, Options for Future Work, SBSTTA VI/8, 20 December 2000, and SBSTTA/6/ paragraphs 20–22; McConnell 2002). More recently, concern has been expressed in the Report of the AHTEG meeting, discussed elsewhere in this article (Subsidiary Body on Scientific, Technical and Technological Advice 2005; at paragraphs 64–70). In particular the AHTEG noted the need to encourage the IMO to address the issue and also, to address the issue in connection with the Antarctic Treaty area. The need to raise the issue with respect to the United Nations openended Informal Consultative Process on Oceans and the Law of the Sea (INICPOLOS) was also noted at paragraph 70(g). A further, perhaps ironic, complication in connection with the transport of invasives, such as seaweeds that attach to the exterior of the ship and its equipment (hull fouling), arises as a result of the fact that the problem may be exacerbated by the adoption and implementation of another recent IMO marine environmental protection Convention, the International Convention on the Control of Harmful Anti-fouling Systems on Ships 2001 (Anti-fouling Convention). The Anti-fouling Convention is aimed at eliminating harmful biocides such as tributyl tin (TBT) used in the coating, paint, surface treatment, surface, or devices on a ship to control or prevent attachment of unwanted organisms. This may then result in use of less effective surface treatments and increased hull fouling and risk of transport of organisms such as invasive seaweeds. The BWM Convention 2004, once it comes into force and is implemented at a national level, may, over time, be an effective response, to the extent that seaweeds are transported and introduced through ballast water and sediment discharges. However, to avoid the spread of species along the coastline within a country, and to adjoining countries, regulation of ships on international voyages should be coupled with regulation of domestic and any regional trade fleets. In addition, once adopted, the Convention’s Guidelines (IMO, Subcommittee on Bulk Liquids Cases 2004 2.2.2) will also include coastal/ port responsibilities, which could help prevent the risk of ships encountering seaweeds that may attach to hulls and equipment. Port area authorities are expected to identify and warn ships of areas where ballast should not be taken up or discharged. These include areas of phytoplankton blooms, outbreaks, infestations or known populations of harmful aquatic organisms, including

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pathogens, as well as any particularly sensitive areas or activities. The process of identifying the ‘‘harmful aquatic organisms’’ in or near coastal and port waters can be taken into account in evaluating ‘‘risky ships’’ by the next port. Although a few countries have adopted regulations to address transport of organisms such as invasive seaweeds by hull fouling, it is not yet entrenched as an acceptable port or coastal State practice under a specific international agreement, although it could perhaps be supported on the basis of more general UNCLOS or CBD obligations to protect the marine ecosystems.7

3. Comparative regional coordination and cooperation In the absence of a comprehensive global response to invasive seaweeds, and given the regional nature of the problem in many respects, it is worth considering the effectiveness of regional efforts to deal with aquatic invasive species generally and invasive seaweeds more specifically. To this end, this section considers regional efforts to respond to the threat of invasive species. Regions recognized by the United Nations Environment Program (UNEP) include Europe, North America, Latin America and the Caribbean, Asia and the Pacific, and West Asia. It is important to recognize that there are regional efforts to address invasive species in most regions of the world (McNeely et al. 2001). A complete overview of regional efforts on this issue is, however, not possible here. Instead, regional efforts in North America and Europe are used to illustrate what has taken place to date, and to consider the challenges and opportunities associated with a regional approach to invasive species such as seaweed. By way of introduction, a regional approach can, in theory, have advantages over both global and national responses to invasive species. In comparison to global efforts, the regional approach has the advantage of allowing law and policy responses to be tailored to the unique circumstances of each region. It also allows States within a region to cooperate in the absence of global consensus. Even where this may not be a viable regulatory response for all vectors, e.g., international shipping, it can be an important component to ensure the effectiveness of international regimes. Regional approaches have potential advantages over national efforts in that they may be better able to tailor responses according to ecological boundaries as opposed to political ones. Furthermore, regional approaches may have The Northern Territory of Australia has recently implemented mandatory hull fouling checks on recreational and fishing vessels seeking entry into any of the enclosed marinas. These guidelines are being evaluated by the National Introduced Marine Pest Coordination Group for implementation in either a voluntary or regulatory framework. New Zealand has developed a Voluntary Code of Practice for vessels departing the two main Islands for the sub-Antarctic Islands, Chatham Islands and transiting to Fiordland (South Island). In addition, a comprehensive research program examining hull fouling across all international vessel sectors (recreational, commercial merchant, commercial fishing, passenger cruise, petroleum and slow moving barges) to underpin a risk evaluation has commenced. 7

An electronic list serve posted a notice in early July 2001 of a proposed ‘‘Planning Meeting: Workshop on Ship Fouling and Biological Invasions in Aquatic Ecosystems’’. The Workshop was proposed by a member of the US Navy, Naval Surface Warfare Center and a member of the USCG Environmental Standards Division. The proponents note that: ‘‘Historically, hull fouling has been the most important means by which shipping has transported non-indigenous species w«x impending limitations on the use of the most effective antifouling paint worganotin basedx and on the conduct of hull cleanings, may result in increased fouling of ships and the subsequent transport of non-indigenous species.’’ 6

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the advantage of minimizing competitiveness otherwise associated with law and policy responses that either prohibit or increase the cost of certain economic activities. The following sections consider whether there are indications that regional efforts in Europe and North America have been able to materialize on these theoretical advantages. 3.1. North American cooperation Cooperation in the North American context means cooperation among Canada, the United States of America, and Mexico. Any regional cooperation involving all three countries is likely to involve the Commission for Environmental Cooperation (CEC) established under the (1992) North American Free Trade Agreement (NAFTA) to encourage cooperation on environmental issues of interest to the three countries. Beyond this, there are opportunities for bilateral or sub-regional cooperation on the Pacific, the Atlantic, and to some extent the Arctic Oceans. In addition, there are opportunities for cooperation on aquatic invasive species in freshwater ecosystems, most notably the Great Lakes system. An interesting issue with respect to regional and sub-regional cooperation in North America is the role of First Nations peoples, and the role of states and provinces in the US and Canada respectively, particularly given the confusing jurisdictional picture with respect to environmental and aboriginal issues in Canada and the US. 3.1.1. Regional cooperation in North America: the CEC The CEC has been actively involved in coordinating action of aquatic invasive species since it held a workshop on the topic in March 2001 (CEC Proceedings 2001). At this workshop, representatives of interested departments and agencies from the three member states identified a number of objectives as well as some specific steps to be taken to improve regional cooperation on this issue. Topics discussed at the workshop included cooperation with respect to science and information sharing, including prediction and modeling work, as well as public awareness issues. The other main area for discussion was the response to the threat of aquatic invasive species through regulatory and voluntary measures to prevent and respond to invasions. Possible areas of cooperation include joint processes for prediction, identification and response to invasions. On the regulatory side, consistent rules to discourage intentional and unintentional behavior that may lead to invasions was seen as important to help reduce the risk of invasions in a manner that equally distributed potential costs involved in eliminating high risk activities. With respect to actions that require no outside motivation, voluntary measures were seen as serving to bring about the desired behavior to reduce or eliminate the risk of invasions without the need for regulations. Within the CEC, discussions on how to achieve regional cooperation through information sharing, regulation, voluntary measures, and awareness raising, are ongoing. At the March, 2001 workshop, participants identified five priority areas for regional cooperation: w124x

(1) A North American Invasive Species Information Network; (2) A directory of legal and institutional frameworks for the prevention and control of invasive species; (3) Identification of invasive species and pathways, particularly those of potential concern to more than one country; (4) Tools for raising awareness; (5) Tools to provide economic incentives to take voluntary action to reduce the risk of invasions (CEC Proceedings 2001, p. 49, 50). Efforts to implement these priorities are ongoing; however, there are no concrete results to report. With respect to invasive seaweeds, the CEC recently commissioned a report on the status of seaweed invasions on the Pacific coast of North America (Murray et al. 2004). The Report, referred to above in section 2 on shipping, which is currently in draft form, considers the environmental threats of identified invasions, and makes some general science and policy recommendations. 3.1.2. Sub-regional cooperation in North America Sub-regional cooperation is generally ecosystem driven and often involves various levels of government with responsibility for a threatened ecosystem. In North America, there are at least three examples of sub-regional cooperation, one involving the Pacific Ocean, one involving the Great Lakes system, and one involving the Gulf of Maine. The Western Regional Panel is considered in more detail below as an example of sub-regional cooperation in the North American context. In the Gulf of Maine, institutions for inter-jurisdictional cooperation are well established in the form of the Gulf of Maine Council. As discussed below, however, there has been limited action on invasive species to date. With respect to the Great Lakes, the International Joint Commission has been active in promoting cooperation on the response to aquatic invasive species in the Great Lakes Region. The focus of these efforts has been on invasions resulting from shipping, specifically ballast water and hull fouling. 3.1.2.1. Western regional panel on aquatic nuisance species The Western Regional Panel on Aquatic Nuisance Species was initially called for in Section 1203 of the US National Invasive Species Act of 1996 (NISA). The Panel was set up following a meeting in autumn 1996. It was made up of representatives from four existing advisory groups on invasive species, including one with representation from British Columbia, Canada (Annual Report, 2000–2001, at 1). Membership of the Panel includes representatives from federal, state and provincial governments in Canada and the United States. In addition, there are members from aboriginal organizations, industry, conservation groups, academia, and other related interests. An interesting aspect of the Panel is that it includes jurisdictions from two countries, however, its statutory base is federal legislation in the United States. The US National Invasive Species Act of 1996 goes beyond requiring the establishment of the Panel. It also dictates its goals. They include the following:

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• Identify priorities for the Western Region with respect to aquatic nuisance species; • Make recommendations to the US Federal Task Force on Invasive Species regarding an education, monitoring (including inspection), prevention, and control program to prevent the spread of the zebra mussel west of the 100th Meridian; • Coordinate, where possible, other aquatic nuisance species program activities in the western region that are not conducted pursuant to the NISA; • Provide advice to public and private individuals and entities concerning methods of preventing and controlling nuisance species infestations; • Submit annual reports to the Task Force. The work of the Panel to implement these goals is ongoing. Based on the 2000/2001 Annual Report (the most recent report available), the work of the Panel is in its early stages. Concrete law and policy recommendations on how to prevent, eradicate or control aquatic invasions have so far not been brought forward by the Panel. The focus so far has been on information sharing, voluntary cooperation and coordination. 3.1.2.2. Gulf of Maine Council On the Atlantic coast of North America, the Gulf of Maine Council on the Marine Environment has been the main vehicle for regional cooperation on marine environmental protection. With respect to aquatic invasive species, the focus to date has been on education and awareness raising. In fact, the current action plan (Gulf of Maine Action Plan 2001) provides for awareness raising and improved management as the two pillars of its aquatic invasive species strategy. On the management side, the plan focuses on information sharing as an initial step toward more effective responses within the Gulf of Maine. The bottom line in North America is that it is too early to tell whether opportunities associated with regional approaches will materialize. The CEC is still in the early stages of trying to engage the member states in taking a coordinated approach to the problem. Similarly, subregional efforts are in the early stages, too early to assess guiding principles let alone evaluate law and policy responses to the threat of invasions. 3.2. European initiatives on aquatic invasive species The multitude of jurisdictions in Europe generally, and in various aquatic ecosystems at risk of invasions more specifically, has resulted in considerable pressure for regional and sub-regional cooperation in Europe, perhaps more than anywhere else. As pointed out in other papers in this issue, aquatic invasions have, for some time, been documented in marine ecosystems throughout Europe, including the Mediterranean Sea, the Baltic Sea, and the Caspian Sea. It is to be expected, therefore, that efforts at regional cooperation may have advanced further in Europe than in other parts of the world. 3.2.1. Regional cooperation in Europe under the Bern Convention Efforts at regional cooperation on aquatic invasive species in Europe generally have been based on global efforts under the CBD, but have been coordinated

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through the 1979 Bern Convention on the Conservation of European Wildlife and Natural Habitats, which came into force in 1982. The Bern Convention has certainly been the main legal instrument to guide regional cooperation on this issue. The Convention initially did not deal in any detail with invasive species, but it does include provisions in Article 11 that bring the issue within the mandate of the Convention. Specifically, Article 11(2)(b) requires each party to ‘‘strictly control the introduction of non-native species’’. This provision, in combination with numerous invasions documented in various ecosystems in Europe, has caused the parties to the Convention to consider the need for regional action on the issue. This has led to the preparation of a European Strategy on Invasive Alien Species prepared on behalf of the Standing Committee under the Bern Convention (Bern Convention 1979). The Strategy was presented to the Standing Committee at its December 2003 meeting in Strasbourg. The Strategy was endorsed by the Standing Committee in Recommendation 99 at the December meeting (Standing Committee Proceedings 2003). Recommendation 99 recommends that all parties to the Bern Convention implement national strategies in light of the European Strategy endorsed by the Standing Committee, and report back to the Standing Committee on progress. Observers (nonparties) are also invited to implement the Strategy. The Strategy is largely based on work carried out on invasive species under the CBD. It adopts the terminology from the CBD, and generally cross references work done under the CBD throughout the document. Areas covered include awareness raising, research and information sharing, and the importance of legal, policy and institutional frameworks. The Strategy furthermore highlights the role of regional cooperation, and endorses the hierarchy of prevention, eradication and control. The Strategy accepts as its basis the precautionary and ecosystem approaches. 3.2.2. Sub-regional cooperation in Europe Due to the significant number of invasions and the multitude of jurisdictions with a stake in responding to those invasions as well as taking measures to prevent future ones, it is not surprising that there are a number of sub-regional efforts to deal with aquatic invasions. There are efforts underway in the Mediterranean Sea, the Baltic Sea, and the Caspian Sea. Such efforts at multi-jurisdictional cooperation are generally carried out under regional agreements, such as the 2003 Framework Convention for the Protection of the Marine Environment of the Caspian Sea (Article 12), the 1995 Protocol concerning Specially Protected Areas and Biological Diversity in the Mediterranean Sea, and the 1992 Helsinki Convention on the Protection of the Marine Environment of the Baltic Sea Area.

4. National level law and policy options One clear message from the foregoing overview of international efforts on invasive species is that regional and global efforts are fragmented in terms of both regulatory w125x

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responses and at an institutional level. Furthermore, it is left to national governments to implement specific law and policy measures to implement the general principles that have been developed at a global and regional level. It is at the national level, therefore, that we can expect to see the results of the global and regional efforts to develop effective responses to the threat of invasions. It will not come as a surprise that the problem of fragmentation does not stop at national borders. Few countries have developed integrated strategies on invasive species that consider long-term local and global harm and benefits of the activities involved and the invasions linked to those activities. A country that has taken this issue more seriously than most is New Zealand (Wotton and Hewitt 2004). In many States, the responsibility is assigned to agencies with a focus on a particular aspect of the problem, usually either associated with the utility of the activity involved, or with a mandate to protect ecosystems or components of ecosystems threatened by invasions. 4.1. General considerations This section will consider in very general terms the issues facing domestic law and policy measures. It is assumed for purposes of this overview that the starting point for national measures will be those commitments, obligations and principles developed under the various international instruments described elsewhere in the article. Given the limited success in addressing the problem globally and regionally, implementation of international agreements is only the starting point at a national level. An effective national response clearly has to go beyond the implementation of international commitments. A national level precautionary approach to regulatory system design is needed. This means that a clear identification and understanding of the problems and gaps found in the international system and the implications for national implementation is needed. The overall message from international efforts on this issue seems clear. Prevention is the most effective way of slowing the increasing rate of invasions. In many cases, prevention has proven to be the only effective way of avoiding the invasion from becoming permanent. This means any effective strategy on invasive species will consider available preventative measures first. Such measures can fall into two categories; they can involve changes to the activity that leads to the invasion to make it safe, or it can involve prohibition of the activity. Whether either of these options can reasonably be implemented with respect to a particular pathway will depend on a number of factors, including the utility of the activity, the availability of an alternative method of achieving the utility of the activity, and the cost of implementing alternatives. In the context of both unintentional and intentional transfer of invasive species, the international character of many of the activities inherently places some limits on the ability to regulate, other than through border control measures that are acceptable under international law. In case of pathways where current conditions do not reasonably allow for measures to prevent the risk of invaw126x

sions, a combination of measures to reduce the risk and motivate further efforts to reduce and even, perhaps, eliminate the risk in the future would constitute a second level of response. An example might be invasions from hull fouling and ballast water. In using this example as a starting point for the discussion of possible national level responses, it is important to take into account the unique characteristics of the activity involved. The example of international shipping is one involving an unintentional introduction from a long-established activity which most would consider as providing an essential service to society. This means that the prohibition of shipping is generally not considered a reasonable response to this threat. As was discussed in section 2, implementation of international agreements, such as the BWM Convention 2004, can be effective in reducing the risk of invasions that may result from ballast water and sediment discharge. This leaves the question of what can or should be done about other risks related to shipping, and about other vectors such as those associated with the aquarium trade and fishing. One part of the answer would be to motivate research and development on ways to prevent invasions resulting from shipping. Another would be the allocation of responsibility for compensation for damages arising from invasions that do occur in spite of measures to those who benefit from the activity. However, this can be problematic because a causal connection can be difficult to establish in that an invasion may not be apparent for many years and, if baseline data on the relevant ecosystem are not available, difficult to prove. Responses to risks that are deemed unavoidable as a result of the utility of the activity and the absence of alternatives have the potential to serve a dual function. They can serve to motivate those who benefit from the activity responsible for the risk to look for ways to reduce or eliminate that risk. They can also serve to fund efforts for responding to invasions that do take place by funding efforts to identify responses as early as possible, and by funding eradication, control or containment of invasions to reduce the long-term damage. An effective national level response to the risk of invasions from vessel traffic might therefore consist of the following components: • Identification of reasonable measures to reduce the risk of invasions from ship traffic, including those set out in international agreements; • Regulation and enforcement to ensure all reasonable measure to reduce or eliminate the risk are being taken; • Identification of the residual risk; • Internalization of the cost of the residual risk (erring on the side of overestimating the risk and the resulting cost); • Use of the funds generated from the process of internalization for early identification, eradication, control, and containment. Assuming this approach works for well established activities with clear utility resulting in unintentional introductions, how would the response have to be adjusted

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to deal with differences in these basic characteristics? Clearly, the less the utility of an existing activity, and the higher the risk associated with it, the greater the pressure to respond to the threat by prohibiting the activity. With respect to new activities, the main difference is that there is an opportunity to implement the precautionary approach in a more direct manner, by delaying or preventing the activity until or unless it can be proven to be sufficiently safe by its proponent. While interpretative debates have surrounded the precautionary approach (VanderZwaag 1998, Ellis and FitzGerald 2004, Scott 2005), a strong version of precaution would place the burden of proof on the proponent of an activity to meet some requisite standard of proof (Hildreth et al. 2005), such as ‘‘no significant’’ threat to ecosystem health. Principle 10 in the CBD Guiding Principles for the Prevention, Introduction and Mitigation of Imports of Alien Species, discussed above, suggests that the burden of proof should be with the proposer of intentional introductions of potentially invasive species to show the introduction will be unlikely to threaten biological diversity. In principle, whether the activity is a new or existing activity, this decision would be made based on the same precautionary approach. The reason for considering them separately is not to diminish the importance of applying a precautionary approach in case of existing activities, but to highlight the opportunity to do so for new activities in a manner that avoids the difficult question of how to turn back the clock and eliminate an activity that has become accepted and relied upon within a society. This leaves the issue of intentional versus unintentional introductions. It is important to first define the boundary between intentional and unintentional introductions. Introductions from ships clearly fall into the unintentional category. On the other hand, fish stocking is the clearest example of an intentional introduction. In between is a range of activities whose classification really depends on the level of effort applied to prevent the escape and/or establishment of an invasive or alien species. In the mariculture context for example, one could consider all non-native mariculture to be intentional. Alternatively, depending on the level of effort made in a particular operation to prevent the spread of a species used in mariculture, it is possible to consider the introduction to be unintentional. Examples might be mariculture in closed systems on land, or in cages, as opposed to in the open sea without barriers to prevent the spread of the species. For purposes of this discussion of appropriate law and policy responses, it is suggested that nothing turns on where the line between intentional and unintentional introductions is drawn. Rather, the issue is one of whether a strategy has been developed whereby the risk of an invasion can be identified and how well it can be evaluated. Having said this, a precautionary approach to invasive species would generally result in the prohibition of intentional introductions in the sense of an introduction of a non-native species without any barriers to prevent its spread, unless the species by its nature is known not to be invasive, something that is generally accepted to be difficult to establish.

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4.2. Institutional issues Having considered the range of decisions that may have to be made at a national level to implement an effective response to invasive species, how should these decisions be made? Inevitably, the decision-making responsibility will have to fit within an existing institutional framework. Even so, existing frameworks may need to be modified to facilitate appropriate decisions on issues ranging from whether to allow an activity, what conditions to impose on an activity to minimize the risk of invasion, to how to ensure motivation to reduce the risk and to ensure that the cost of any invasion is born by those who benefit from the activity that is creating the risk. The fundamental choice is between the creation of a new decision-making agency and the use of one or a combination of existing decision makers, ideally operating within an integrated framework. Some countries, such as New Zealand, Australia, and the United States, have created new agencies to either oversee the national implementation by existing agencies, implement directly, or a combination of both. Other countries have focused on imposing decision-making responsibilities on existing departments and agencies. Existing decision makers typically fall into one of two categories. The first consists of regulators of the sectoral activities that pose a risk of invasions. These generally include departments of fisheries, aquaculture, transportation, and agriculture. The other includes decision makers with an environmental protection and conservation focus. These generally include departments responsible for environmental protection, biodiversity, resource conservation, and agencies responsible for processes such as environmental approval and environmental assessment processes. It is not surprising that one of the main challenges with the reliance on existing decision-makers is how to ensure that the objective of preventing, reducing and controlling invasive species receives sufficient weight when compared to the pre-existing mandate of the decision maker. This problem also arises in the context of environmental decision-making superimposed on government decision makers through environmental assessment processes that are based on the self assessment model. A precautionary approach here would suggest that decision-making responsibility should not rest with regulators of the activity involved, but with decision makers whose primary mandate is as closely as possible connected to the prevention of invasions. 4.3. Regulatory tools A domestic regulatory response to prevent the transfer of harmful aquatic organisms and pathogens should be based on a principled approach with sustainability as the ultimate objective. The following principles initially developed in the shipping context provide a sound basis for a national response to invasive seaweeds more generally: • Responses to an ecological problem in the context of an international activity, such as shipping, should be based on an approach that seeks to fulfill international responsibilities to protect the global environment, w127x

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•

•

•

• •

•

•

•

•

•

•

integrates economic and ecological protection concerns and is based on international cooperation to develop rules and technological or other solutions to environmental problems arising out of the globalization of the economic system. A precautionary approach should be adopted for both regulatory design and implementation. For example, all regulatory determinations must, as much as possible, be based on scientific research and an analysis of both local and global ecological implications of any action, with preference given to measures designed to ensure either no, or the least possible, long-term negative impact on the environment. Risks to the ecosystem8 should be minimized by designing and adopting measures that are commercially and practically viable and that encourage compliance rather than avoidance and conflicts. Responses should allow for and explicitly encourage continuous technological and operational improvement to better protect the marine ecosystem. It will be important to ensure transparency, sustainability and integration of agency responses. Responses should encourage the involvement of all parties affected by the issue (and any decisions about regulating the issue), including the regulated sectors and other sectors, in helping to develop a solution. Compliance should be encouraged by making use of a range of modern regulatory devices such as economic incentives and voluntary compliance agreements. There should be a focus on measures to prevent the uptake of harmful organisms and pathogens at the source as well as preventing their introduction. It will be important to develop local and regional contingency responses and compensation plans for all those negatively affected by the activity, based on a ‘‘polluter-pay’’ model. Requirements should be put in place that are environmentally safe, practicable, designed to minimize cost and delays to the shipping industry and, as much as possible, based on the internationally accepted standards such as the IMO Guidelines and the BWM Convention 2004 and any guidance developed under it. Requirements operating at a national level also need to take into account ecosystem differences within each country, and must be applied in a fair, uniform and consistent manner in each port. It will be important to ensure that there is ongoing review and monitoring to evaluate the impact of any action that is taken.

Goals to be achieved in designing a regulatory system, including legislation, might include: • Preventing the problem at the earliest possible point and with the highest level of effectiveness possible; • Maximizing opportunities for risk assessment and prevention; This includes the environment or ecosystem of the enacting State and other States and common areas as noted in Guiding Principle 4. 8

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• Maximizing administrative efficiency and cooperation through holistic approaches and integrated management; • Reducing unnecessary costs to the public and the regulated industry; • Avoiding unnecessary conflicts between shipping and other coastal zone users and amongst regulatory agencies; • Minimizing uncertainty for all affected parties; • Ensuring transparency; • Maximizing accountability – internationally, regionally and nationally; • Ensuring flexibility to respond to and incorporate developments in scientific information, technology or the development of new related concerns, and to accommodate local ecosystem conditions and requirements in a harmonized manner (McConnell 2002). Having now considered the substance of the decisions that would have to be made at the national level as well as some basic choices about the decision maker, this leaves the question of the regulatory tools that may be used to implement the decisions made. The traditional tool for environmental protection is generally known as command and control. This refers to the general process of identifying what should or should not be done, setting out those requirements in law, imposing penalties for not following those requirements, and then designing and implementing effective enforcement mechanisms to ensure that legal obligations are observed. A common regulatory tool is the permitting process, which requires individuals who wish to engage in an activity to seek approval, and allows regulators to specify conditions for approval on a case by case basis. Another common form of regulating activities is through standards imposed through regulations. Here those engaged in activities are required to meet the standard set in the regulation rather than to be forced to individually apply for permission. Anyone who complies with the standards set in the regulation is entitled to engage in the activity in question. There are a number of alternatives to the traditional command and control approach that have evolved over time. Of most interest in this context is what has become known as economic instruments. They are of interest here in particular with respect to activities that are permitted to continue in spite of an ongoing risk of invasions, presumably due to the utility of the activity and the absence of an alternative that would eliminate the risk without eliminating the utility. Economic instruments provide the opportunity to ensure that the cost of the risk of invasion is borne by those who engage in or otherwise benefit from the activity. This can be done through user fees, requirements for insurance, security requirements, or the establishment of a liability fund. The common thread is the objective of quantifying the cost of the risk of invasion, and requiring that the cost be paid by those who benefit from it. This process is often referred to as ‘‘internalizing the cost’’. An alternative use of the same range of economic instruments is to motivate those engaged in the activity to find ways to reduce or eliminate the risk. While this can be achieved by internalizing the cost, it only does so

M. Doelle et al.: Law and policy responses to invasive seaweeds

if the cost cannot be passed on easily and if the cost is sufficiently high to create a motivation to invest in finding ways to reduce or eliminate the risk. If the primary objective of an economic instrument is therefore the ultimate elimination of the risk, rather than the internalization of the cost, the price will be set based on what is required to motivate research and development, not based on the cost of responding to invasions.

5. Conclusion: the need for a precautionary integrated approach to regulatory design and implementation decisions The development of an effective law and policy response to the threat of invasive seaweeds is still in its early stages. Efforts to move towards an effective global response are ongoing. While these efforts do show some promise, the current state is one of fragmentation. Similarly at the regional level, much remains to be done. Some jurisdictions have made significant progress, while others are just starting to take the problem of invasive species seriously. There is little indication to date of the specific issue of invasive seaweeds as a distinct challenge that has been identified in many jurisdictions or regions, let alone at the global level. At the same time, law and policy makers do have a solid basis of principles and potentially powerful regulatory tools to draw upon.

References International agreements and documents Bern Convention on the Conservation of European Wildlife and Natural Habitats, September 19, 1979, Council of Europe – ETS no. 104. Online: -http://conventions.coe.int/Treaty/EN/ Treaties/Html/104.htm). Convention on Biological Diversity of the United Nations Conference on the Environment and Development, June 5, 1992, 31 I.L.M. 818 (1992). Convention on Migratory Species of Wild Animals, June 23, 1979, 19 I.L.M. 15 (1980). Convention on the Protection of the Marine Environment of the Baltic Sea Area, April 9, 1992. Online:-http://www.helcom.fi/ Convention/en_GB/text/). Convention on Wetlands of International Importance especially as Waterfowl Habitat, February 2, 1971, 11 I.L.M. 963 (1972). International Convention on the Control of Harmful Anti-fouling Systems on Ships (Anti-fouling Convention), February 1, 2002, IMO AFS/CONF/26 (2001). Online: -http://www. imo.org/). International Convention for the Prevention of Pollution from Ships (MARPOL 1973), November 2, 1973, 1340 U.N.T.S. 184 (1973), 1978 Protocol, 17 February 1978, 1340 U.N.T.S. 6. International Convention on the Control and Management of Ship’s Ballast Water and Sediment, June 1, 2004, IMO BWM/ CONF/36 (2004). Online: -http://www.imo.org/). North American Free Trade Agreement, December 17, 1992, 32 I.L.M. 289 (1993), 32 I.L.M. 605 (1993). North American Agreement on Environmental Cooperation, September 14, 1993, 32 I.L.M. 1480 (1993). Protocol concerning Specially Protected Areas and Biological Diversity in the Mediterranean, June 10, 1995. Online: -http: //www.oceanlaw.net/texts/unepmap2.htm).

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Rio Declaration on Environment and Development, June 13, 1992, 31 I.L.M. 874 (1992). The Framework Convention for the Protection of the Marine Environment of the Caspian Sea, November 4, 2003. Online: -http://www.caspianenvironment.org/NewSite/ Convention-FrameworkConventionText.htm). United Nations Convention on the Law of the Sea, December 10, 1982, 21 I.L.M. 1261 (1982).

Other sources Cawthron Institute, ‘‘Cawthron reveals unwelcome ‘Octopus Garden’’’. Press release. December 14, 2004. Online: -http: //www.cawthron.org.nz/news/news_sea_chests_PR.htm). CBD Secretariat, ‘‘Jakarta Mandate Marine and Coastal Biodiversity – Introduction’’. Online: -http://www.biodiv.org/ programmes/areas/marine/default.asp). CBD Secretariat, Report of the Group of Legal and Technical Experts on Liability and Redress in the Context of Paragraph 2 of Article 14 of the Convention on Biological Diversity. 2005. UNEP/CBD/COP/8/27/Add.3. Online: -http://www. biodiv.org/doc/metings/cop/cop-08/official/cop-08-27-add3en). Doelle, M. 2003. The quiet invasion. Law and policy responses to invasive species in North America. 2003. Int. J. Mar. Coast. Law 18: 261–294. Ellis, J. and A. FitzGerald. 2004. The precautionary principle in International Law: lessons from Fuller’s internal morality. McGill Law J. 49: 779–800. European Strategy on Invasive Alien Species, Bern Convention on the Conservation of European Wildlife and Natural Habitats Standing Committee. December 2003. Online: -http:// www.coe.int). FAO, Code of Conduct for Responsible Fisheries. 1995. Online: -http://www.fao.org/fi/,default.asp). FAO, Precautionary approach to capture fisheries and species introductions. FAO Technical Guidelines on the Precautionary Approach to Capture Fisheries and Species Introductions No. 2. 1996. Online: -http://www.FAO.org/documents/ show_cdr.asp). Firestone, J. and J.J. Corbett. 2005. Coastal and port environments: international legal and policy responses to reduce ballast water introductions of potentially invasive species. Ocean Dev. Int. Law 36: 291–316. Globallast, Global Ballast Water Management Programme website. Online: -http://globallast.imo.org). Gulf of Maine Council on the Marine Environment, Action Plan, 2001–2006. Online: -www.gulfofmaine.org). Hildreth, R.G., M.C. Jarman and M. Langlas. 2005. Roles for a precautionary approach in marine resources management. Ocean Yearb. 19: 33–61. Hewitt, C.L., J. Willing, A. Bauckham, A.M. Cassidy, C.M.S. Cox, L. Jones and D.M. Wotton. 2004. New Zealand marine biosecurity: delivering outcomes in a fluid environment. NZ J. Mar. Freshw. Res. 38: 429–438. ICES, Code of Practice on the Introduction and Transfer of Marine Organisms, International Council for the Exploration of the Sea, 2005, available online at -http://www.ices.dk/). IMO, Resolution A.868(20) Guidelines for the Control and Management of Ships’ Ballast Water to Minimize the Transfer of Harmful Aquatic Organisms and Pathogens, 1997. Online: -http://globallast.imo.org/resolution.htm). IMO, Sub-Committee on Bulk Liquids and Gases. Development of Guidelines for Uniform Implementation of the 2004 BWM Convention. Note by the Secretariat, IMO Doc. BLG 9/11, November 12, 2004. Invasive Alien Species, Options for Future Work, SBSTTA VI/8 and SBSTTA/6/ paras 20–22, 20 December 2000. Online: -http://www.biodiv.org). March 2001 CEC Workshop Proceedings, Workshop to Identify Cooperative Opportunities for Preventing the Introduction and Spread of Aquatic Invasive Species, North American

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Commission for Environmental Cooperation. Online: -http:// www.cec.org/programs_projects). McConnell, M.L., GloBallast Legislative Review, 2002 Final Report, Globallast Monograph Series, International Maritime Organization No. 1. London. 121 pp. q app. 2002. Online: -http://globallast.imo.org). McNeely, J.A., H.A. Mooney, L.E. Neville, P. Johan Schei and J.K. Waage. Global Strategy on Invasive Species. IUCN Gland, Switzerland and Cambridge, 2001. U.K. 50. Notes of Standing Committee to the Bern Convention, 23rd Meeting, December, 2003, Strasbourg. Online: -http:// www.coe.int/T/E/Cultural_Co-operation/Environment). Murray, S., L. Fernandez, J. Zertuche-Gonzalez. Status, Environmental Threats, and Policy Considerations for Invasive Seaweed for the Pacific Coast of North America, Commission on Environmental Cooperation (presentation) 2004. Online at -http://www.icais.org/pdf/21Tuesday/C/tues_c_ e_am/Hans_Herrmann.pdf). Ramsar Conference, 1999. Resolution VII.14 on Invasive Species and Wetlands. Online: -http://www.ramsar.org/res/key_ res_vii.14e.htm). Ramsar Conference, 2002. Resolution VIII.18 on Invasive Species and Wetlands. Online: -http://www.ramsar.org/res/ key_res_viii_18_e.htm). Resolution VII.14 1999, Ramsar Convention COP7. Online: -http://www.ramsar.org/res/key_res_vii.14e.pdf). Resolution VIII.18, 2002, Ramsar Convention COP8. Online: -http://www.ramsar.org/res/key_res_viii_18_e.pdf). SBSTTA 2004, SBSTTA 2000 and other SBSSTA docs and COP decisions, Convention on Biological Diversity. Online: -http: //www.biodiv.org/). Scott, D.N. Shifting the burden of proof: the precautionary principle and its potential for the ‘‘democratization’’ of risk. In: (Law Commission of Canada, ed.) 2005. Law and risk. UBC Press, Vancouver. pp. 50–85. Subsidiary Body on Scientific, Technical and Technological Advice, Convention on Biological Diversity. Report of the Ad Hoc Technical Expert Group on Gaps and Inconsistencies in the International Regulatory Framework in Relation to Invasive Alien Species. 2005. UNEP/CBD/SBSTTA/11/INF/4.

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US National Invasive Species Act of 1996, Pub. L. No. 104–332, 110 Stat 4073, available online at -http://thomas.loc.gov/ cgi-bin/query/D?c104:13:./temp/;mdbsexmOl0::). VanderZwaag, D. 1998. The precautionary principle in environmental law and policy: elusive rhetoric and first embraces. J. Environ. Law Practic. 8: 355–375. Western Panel on Aquatic Nuisance Species, 2000–2001 Annual Report. Online: http://www.fws.gov/answest/annualreport. htm). Wotton, D.M. and C.L. Hewitt. 2004. Marine biosecurity postborder management: developing incursion response systems for New Zealand. N. Z. J. Mar. Freshw. Res. 38: 553–559.

Internet resources Baltic Marine Biologists (BMB). http://www.smf.su.se/bmb. Baltic Sea Regional Project (BSRP). http://www.helcom.fi/ helcom/projectsmeetings/GEF-BSRP.html. Baltic Sea Transportation System Projects. http://www. oce.pg.gda.pl/oce2/eureka/index.htm. Caspian Environment Programme (CEP). http://www. caspianenvironment.org. CIESM Portal Programme. http://www.ciesm.org/people/task. html. Global Ballast Water Management Programme. -http:// globallast.imo.org). Nordic Network on Introduced Species (NNIS). -http://www. sns.dk/natur/nnis). The EU Northern Dimension. http://europa.eu.int/comm/ external_relations/north_dim. The Nordic Council. http://www.norden.org/search/sk/search. asp. The Nordic Environmental Action Plan 2001–2004. http://www. norden.org/miljoe/sk/eng.pdf. Working Group 30 on Non-indigenous Estuarine and Marine Organisms (BMB WG 30 NEMO). http://www.ku.lt/nemo/ mainnemo.htm. 2003 HELCOM Bremen Declaration. http://www.helcom.fi/ helcom/declarations.html. Received 10 January, 2006; accepted 23 November, 2006

Botanica Marina 50 (2007): 451–457

2007 by Walter de Gruyter • Berlin • New York. DOI 10.1515/BOT.2007.047

Conclusion

Seaweed invasions: conclusions and future directions

Craig R. Johnson School of Zoology and TAFI, University of Tasmania, GPO Box 252-05, Hobart, Tasmania 7001, Australia, e-mail: [email protected]

In the introduction of this special issue of Botanica Marina, we established the scope of this topic around a series of questions, which also provided a framework for integrating the different contributions and a means of highlighting deficiencies in knowledge and understanding (Johnson and Chapman 2007). In defining challenges for the future and suggesting how those challenges might best be tackled to optimize returns on research investment, it is useful to revisit those questions under four main headings dealing with species introductions per se, the invasion process, consequences of invasions and human responses to the threat and occurrence of invasion.

Introductions of alien seaweeds The questions posed • What are the major modes of introduction of invasive seaweeds? • Is there tangible pressure for ongoing intentional introductions? Accidental introductions Introduced seaweeds have been detected in most marine bioregions of the world, the exceptions being some tropical areas and the polar latitudes (Hewitt et al. 2007). They can account for a significant proportion of the total flora (e.g., Ribera and Boudouresque 1995) and up to ;40% of all alien species (Schaffelke et al. 2006) in a given area. As is the case with most establishments of alien species, the great majority (;97%) of seaweed introductions are accidental, and hull fouling of ships is by far the most important vector (Hewitt et al. 2007). These observations pose two challenges, namely ascertaining the veracity of current knowledge about the extent of alien and invasive seaweeds, and whether there are meaningful responses to reduce the risk of introductions through hull fouling. Particularly in the last decade, there has been an accelerated effort to detect introduced seaweeds to the extent that up to ;260 species world wide have now been identified as alien to their native range (Hewitt et al. 2007). However, there is at least a qualitative correlation

between the number of alien species recorded in IUCN bioregions and the effort given to looking for them. Regions where interest among phycologists, authorities and the public alike has driven a large effort to detect alien species (Mediterranean), or where systematic surveys of susceptible areas (ports) have been implemented (Australia and New Zealand), reveal the greatest number of alien seaweed species. This is hardly surprising, but it begs the question about the real number of established alien seaweeds, and whether apparent patterns of larger numbers of alien seaweeds in Australia, New Zealand and the Mediterranean, but far fewer alien species in tropical waters, particularly in Africa and Asia, are real. Just as is the case for invasions of terrestrial systems (Mack et al. 2000), preventing invasions of seaweeds in the first place is much less costly than attempting control of post-establishment. However, given that hull fouling is the principal culprit, this poses large challenges, particularly in light of the intent to phase out use of tributlytin (TBT) and related organotins in antifouling paints by 1st January 2008 because of their deleterious effects on the marine environment. The International Maritime Organization’s (IMO’s) International Convention on the Control of Harmful Anti-fouling Systems on Ships prohibits the use of TBT and will enter into force 12 months after 25 States (representing 25% of the world’s merchant shipping tonnage) have ratified it. The 1999 IMO Assembly resolution called for complete prohibition on the application of TBT and other organotin compounds by 1st January 2008. Despite redoubled efforts to develop new antifouling paints based on chemical biocides less toxic than organotins, or with physical properties that inhibit settlement of marine organisms, hull fouling remains a significant problem. Engineering solutions (e.g., mechanical hull scrubbers) have presumably not been cost effective. Indeed, most engineering efforts to control translocation of marine species, including algae, by shipping have focused on treatment of ballast water (e.g., Cangelosi et al. 2007, Gollasch et al. 2007). Intentional introductions A minority of species (-3%) has been introduced intentionally, usually for aquaculture, but there are likely to be ongoing appeals for future introductions. This pressure is most acute in developing nations in tropical and subtropical regions where aquaculture of seaweeds is often seen as an alternative (and sustainable) economic activity to either growing species that require large inputs of artificial feeds or extractive harvesting in wild fisheries (Pickering et al. 2007), both of which have typically had a large impact on the environment and which may be ecologically and economically unsustainable in the long term. Given large ecological pressures on coral reefs as a result of, for example, overfishing and pollution associated with w131x

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aquaculture of species that require addition of nutrients (McManus 1997, Jackson et al. 2001, Hughes et al. 2003, Feng et al. 2004, Azanza et al. 2005), there is considerable ecological as well as socio-economic pressure to identify alternative and more sustainable livelihoods for human coastal populations. Aquaculture species that do not require exogenous inputs of nutrients, such as shellfish (Bell and Gervis 1999, Feng et al. 2004) and seaweeds (Feng et al. 2004, Pickering et al. 2007) are attractive on environmental grounds. Nonetheless, in many places in the Pacific where they occur, these relatively low-impact activities are still largely in an experimental phase (Bell and Gervis 1999, Pickering et al. 2007), although there are notable exceptions (Doty 1977, Trono 1999, Feng et al. 2004, Pickering et al. 2007). The reasons behind failed attempts to develop seaweed culture usually have a strong cultural and/or socio-economic basis (McManus et al. 2002, Pickering 2006, Pickering et al. 2007).

The invasion process The questions posed • Are there common life-history or genetic traits of successful invaders? • Why do some species become invasive while others do not? • Is it possible to predict the next pest seaweed? • Are there common mechanisms underpinning seaweed invasions? • Why do some communities appear to be more susceptible to incursions than others? • Do the traits of the recipient community influence invasion rates? As is the case with terrestrial plants (e.g., Crawley 1987, Mack et al. 2000), there is no consistent set of lifehistory or morphological traits, or particular taxonomic affinities, that define invasive seaweeds (Valentine et al. 2007). Thus, it remains unclear why some species are highly invasive while other, even closely related, species (e.g., Trowbridge 1998) are not. By corollary, the most reliable means of identifying a ‘‘next pest’’ seaweed is through reference to existing patterns of invasion rather than consideration of species’ traits in isolation. If there are any broad traits common to most (but not all) invasive seaweeds, it is the capacity for rapid growth and means of effecting both short and long-distance dispersal (Valentine et al. 2007), but many seaweeds manifest these properties without showing any sign of being invasive. Moreover, even the same species can be invasive in one area but not in another (e.g., Trowbridge 1998, Chapman 1999) or, in the same area, at one time but not another we.g., Undaria pinnatifida (Harvey) Suringar, CR Johnson, pers. obs.x. Collectively, these observations suggest that invasion might have as much to do with the properties of the recipient community as of the invader itself and, of course, this has been a topic of much debate since Elton’s (1958) first discussion of the idea (e.g., Levine and D’Antonio 1999, Levine 2000, Stachowicz and Tilman 2005, Fridley et al. 2007). w132x

The results of Dunstan and Johnson’s (2007) models substantiate the suggestions of Davis et al. (2000) and Davies et al. (2005) that resource availability is a critical determinant of invasion success and, in particular, that the likelihood of invasion increases with variability in resource availability. For nearly all seaweed species where there is empirical evidence of the mechanism of invasion, disturbance to native species in the recipient community is a key feature (Valentine et al. 2007). Caulerpa taxifolia (M. Vahl) C. Ag. in the Mediterranean may be the only seaweed where invasion does not depend on disturbance breaking a native species’ monopoly on resources (see Valentine et al. 2007), but even here resistance to invasion is greater in areas where cover of native seagrass or macroalgae is high (deVille´le and Verlaque 1995, Ceccherelli and Cinelli 1999, Ceccherelli et al. 2002), and there is evidence that high density stands do indeed depend on disturbance to limit the growth of native species (deVille´le and Verlaque 1995, Jaubert et al. 2003). Thus, it appears that invasive seaweeds are not especially competitive; competitive displacement of native seaweeds de novo by alien invasive seaweeds occurs rarely, if at all. At least in some quarters, a similar view is emerging for terrestrial plants (Bruno et al. 2005). To date, theoretical considerations and empirical observations are consistent with the idea that the risk of seaweed invasion is related to variability in resource availability (Dunstan and Johnson 2007, Valentine et al. 2007), but further carefully designed and executed experiments are required to more rigorously test these ideas. If variability in resource availability through disturbance, or any other mechanism that causes morbidity or mortality of native seaweeds, is ultimately shown to be the key predictor of invasion success, then this may explain apparent anomalies such as why a species is invasive in one area but not another, and why some communities appear more easily invaded than others. It would also explain why risk of invasion might not be predicted by the richness of the recipient community (Dunstan and Johnson 2007). The pattern of fluctuation in resource availability at any given locale will depend at least on the properties of the occupying native species and interactions among them, patterns of disturbance, and variability in physical properties such as temperature, wave action, and nutrient loading. Complex responses to several experiments suggest that considerations of invasion dynamics need to address separately the different stages of invasion since the mechanisms facilitating ongoing persistence of an invader may be different from those underpinning its initial establishment and spread (Valentine et al. 2007). Disturbance to native canopy-forming algae to provide space is necessary for Sargassum muticum (Yendo) Fensholt and Codium fragile ssp. tomentosoides (van Goor) P.C. Silva to establish at high densities, and presumably to spread beyond initial establishment. Having established a closed canopy, both species can effectively inhibit re-establishment of native seaweeds, at least in the short term (reviewed by Valentine et al. 2007). In Tasmania, Undaria pinnatifida also requires disturbance to native seaweeds to establish at high densities but, once established, it does not require ongoing disturbance to

C.R. Johnson: Seaweed invasions: conclusions and future directions 453

maintain its dominance. This is because a matrix of sediment and filamentous algae builds up on the substratum when native canopy-forming species are removed. This matrix inhibits development of native canopy-forming seaweeds, even in the absence of both disturbance (sea urchins) and Undaria pinnatifida and when the inoculum of spores from native seaweeds is enhanced (Valentine and Johnson 2005a,b). However, this effect is scale dependent. If native canopy forming seaweeds are removed in small patches (16 m2) then U. pinnatifida establishes and the sediment matrix develops, but within two years the native seaweeds displace the invader to recover dominance (Valentine and Johnson 2003). However, when native algae are removed on a much larger scale, U. pinnatifida establishes, the sediment matrix develops, and native canopy forming seaweeds do not recover (Valentine and Johnson 2005a). The explanation for this non-linear response is unclear. Moreover, it is difficult to estimate the prevalence of nonlinearities of this kind because too few experiments are undertaken to detect them. They certainly add complexity to the nature of invasion dynamics. Similarly, shortterm responses to manipulations at small scales may not be evident over longer periods or at larger scales. If all patches where Codium fragile ssp. tomentosoides (Chapman et al. 2002, Levin et al. 2002) or Sargassum muticum (Britton-Simmons 2004) established were able to inhibit recruitment of native seaweeds indefinitely, then patches of the invader might be expected to persist and maintain their dominance of ‘‘captured’’ sites. This mechanism would result in the gradual but inexorable displacement of native seaweeds over large scales, but this has not been observed for either species. The complete dynamic is more complex than is indicated by short-term experiments. Clearly, detailed and enlightened understanding of invasion processes requires careful experimentation and observations over the long term. There is yet much to be done to fully comprehend the nuances of the invasion dynamics of seaweeds. But even with a more complete knowledge, it is unlikely that particular seaweed invasions will ever be predictable. In this, invasions by seaweeds are no different from those of other kinds of organisms (Gilpin 1990).

Consequences of invasions The questions posed • What are the ecological, genetic and economic consequences of seaweed invasions? • Can we expect existing and, in particular, emerging techniques in genetics and genomics to provide a much deeper understanding of seaweed invasions? If there is issue about the real nature of worldwide biogeographic patterns of alien seaweeds, the uncertainty about the impacts of alien species is even more profound. It could be argued that, since the direct impacts of very few alien species have been considered at all (;17 of the ;260 alien species identified) while even fewer (;4 species) have been studied in any real depth with appropriate effort given to both experimental work

and surveys (Schaffelke and Hewitt 2007), it is premature if not wholly invalid to attempt any general synthesis of the impact of alien seaweeds. However, it seems clear that, like most other groups of alien species including terrestrial plants (e.g., Williamson and Fitter 1996, Gurevitch and Padilla 2004, Bruno et al. 2005), the great majority of alien seaweeds are not highly invasive and are unlikely to become pests. Indeed, the highly skewed attention to a minority of alien seaweed species arises because effort has focussed on those few that appear most likely to be problematic. While there are undoubtedly as yet undetected ecological impacts of alien seaweeds, it is just as certain that a greater number of species would have been investigated more fully to date if there were clear signs of them becoming invasive as opposed to establishing as another ‘‘background’’ alien species. Of those that have been investigated, effects have varied from significant reductions in both abundance and diversity of native species at a local scale, to no detectable effect, to local enhancement of invertebrate and fish biomass and/or diversity (see Schaffelke and Hewitt 2007, Table 1). In several cases, the introduced seaweed seems to simply increase the standing biomass of total algae. No invasive seaweed has had what could be described as massive effects on biodiversity on a large scale. While it is clear that some seaweeds have significant ecological impacts, at least locally, current evidence suggests that their impacts might not be as great as several marine animals, particularly some filter-feeding marine molluscs we.g., the Asian clam Potamocorbula amurensis (Schrenck), Nichols et al. 1990, and New Zealand screwshell Maoricolpus roseus (J.R.C. Quoy et J.P. Gaimard, A. Reid and C.R. Johnson, unpublished datax, and generalist predators in pelagic (e.g., the ctenophore Mnemiopsis leidyi, Shiganova 1998, Daskalov 2002, 2003) and benthic (e.g., Northern Pacific seastar, Asterias amurensis A. Agassiz; Ross et al. 2003) habitats. All of these animals have the capacity to effect deleterious impact on richness, diversity and ecosystem functioning over large spatial scales. Of course, comparing the impact of seaweeds in one area with that of animals in another is problematic because it is not possible to distinguish impacts as a property of the invader from impacts related to the nature of the recipient community. However, the pattern of greater impacts by animals than plants is sometimes evident in the same system. In Tasmania for example, despite the relatively rapid spread of Undaria pinnatifida and its capacity to form a closed canopy in dense stands (Sanderson 1990, Valentine and Johnson 2003, Hewitt et al. 2005), its overall effect has arguably been small compared to that of introduced benthic marine animals such as A. amurensis (Ross et al. 2003), M. roseus and the European green crab Carcinus maenus (Linn.) (Thresher et al. 2003). C. maenus has a voracious appetite for small bivalves including mussels, oysters and cockles, and appears to displace a similarly sized native crab species, Paragapsus gaimardii (Milne Edwards) (Walton et al. 2002). If these patterns are found to hold more generally, then it will parallel the pattern from terrestrial systems where invasive animals, particularly generalist predators, w133x

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have greater direct impacts than plants. In other words, invasive top-down effects may be more important than invasive bottom-up effects in changing the configurations of recipient benthic communities. There remains much to do to properly understand the impacts of invasive seaweed species. Their socio-economic impacts have hardly been addressed at all (Schaffelke et al. 2007). Not only have the impacts of few species been examined thoroughly, but most studies have focussed on the direct impacts on the abundance and diversity of natives. In considering effects on diversity, most work has examined local alpha diversity, while beta and gamma diversity have been ignored. In many cases, local alpha diversity might decrease while gamma diversity increases. For example, in Nova Scotia Codium fragile spp. tomentosoides has been spectacular in its displacement of the previously dominant kelps (largely Saccharina longicruris (Pyl.) Kuntze; see Johnson and Mann 1988 was: Laminaria longicruris Pyl.x) in patches. This green alga does so by capitalizing on disturbances from grazers, or in response to dieback of kelp as a result of smothering of their photosynthetic tissue by the introduced bryozoan Membranipora membranacea Linn. (see Chapman et al. 2002). Native seaweeds, particularly kelps, are less abundant in areas taken over by Codium, and so at a local scale alpha diversity is reduced. A similar phenomenon has been observed in the Gulf of Maine further to the south (e.g., Levin et al. 2000). While Codium can form meadows of 100s–1000s of square meters in extent, in Nova Scotia replacement of kelps at this scale has occurred mostly in two bays (Mahone and St Margarets Bay), but only in wave sheltered areas in shallow water (F8 m; R. Scheibling, pers. comm.). Thus, gamma diversity may well have increased. Descriptions of impacts on diversity and abundances are a useful starting point, but at least as important are the effects on ecosystem functioning and dynamics. However, these aspects are rarely examined, no doubt because of the effort and, in some cases, technology required. How are primary and secondary production, and resistance and resilience stability (sensu Dayton et al. 1984) affected by invasives? What is the nature of indirect effects, and of synergistic effects of co-occurrence of several alien seaweeds? The only measurable effect of the establishment of some alien species is apparently to increase total algal cover and local richness (see Schaffelke and Hewitt 2007). But do relatively benign alien species like this simply accumulate indefinitely in a linear manner, or is some threshold attained where effects become non-linear? Can otherwise noninvasive seaweeds become highly invasive in the face of significant interactions with anthropogenic modifications to the marine environment, e.g., disturbance and eutrophication? Is there any evidence of ‘‘invasional meltdown’’ (Simberloff and Von Holle 1999; see also Simberloff 2006) whereby establishment of some alien species increases the likelihood of subsequent incursions? Are there significant evolutionary consequences for the invader in establishing in new areas, and for native species as seaweed floras become increasingly homogenized with ongoing accumulation of aliens? These are all important questions that are poorly addressed, not only for seaweeds but for invaded sysw134x

tems in general. But they are also difficult questions to tackle, and typically require complex experimental fieldbased studies, often over several years combined with empirical surveys at several scales. We are hopeful that researchers will have the courage to address these questions in the near future. With the rapid emergence and uptake of new techniques in molecular genetics, there is every chance that some of the evolutionary questions, particularly those relating to interactions between genotype, phenotype and environment, might be resolved before the ecological ones (Hofmann et al. 2005, Booth et al. 2007). This would be a welcome development given that the genetic consequences of seaweed invasions, including effects on gene pool composition, genome organization and mating systems, are currently so poorly appreciated (Booth et al. 2007).

Human responses to the threat and occurrence of invasions The questions posed • How have seaweed invasions been tracked, and can existing approaches be improved? • Is it possible to predict the course of an invasion? • What are sensible approaches to reducing risk of further introductions? Responses to the threat of alien invasive seaweeds have included (1) attempts to mitigate the risk of alien species becoming established in the first place, (2) the development and application of science and technology to better predict ‘‘next pests’’, more rapidly detect alien species before they have opportunity to spread, and better understand the space-time dynamics of invasions, and (3) attempts to eliminate or control invasive seaweeds when they do establish in a new locale. The earlier intervention occurs in the chain of events described by uptake at a donor site ™ translocation by a vector ™ release at a donor site ™ establishment at donor site, the easier and less expensive it usually is to thwart potential introductions. Ships have long been recognized as vectors for translocation of marine species (e.g., Ostenfeld 1908) but, as we emphasized earlier, practical responses to minimize translocation of species by ships have focused largely on ballast water (e.g., Cangelosi et al. 2007, Gollasch et al. 2007), despite the fact that hull fouling is a major source of introduction of seaweeds and many other kinds of marine organisms (Hewitt et al. 2007). Particularly as TBT-based antifouling is phased out, and given the cost of engineering solutions to hull fouling, translocation of alien seaweeds by shipping is likely to be with us for a long time to come unless hull scrubbing or other hull treatment is forced through regulation, which seems unlikely, at least in the medium term (see Doelle et al. 2007). This is despite the economic incentive to reduce drag, and thus fuel costs, by minimising hull fouling. Notwithstanding the problem of hull fouling, international and most national policy and law makers are acutely aware of the risks associated with translocation of marine species. They also recognise the fragmented nature and lack of integration of existing reg-

C.R. Johnson: Seaweed invasions: conclusions and future directions 455

ulations aimed at controlling translocation of marine species across national borders (Doelle et al. 2007). Researchers must take all opportunities to urge their governments to establish regulatory frameworks that minimize the risk of translocating species among bioregions, and to cooperate with other nations in effecting this goal. The lead taken in establishing pertinent legislation by the governments of Australia, New Zealand, Canada and Germany needs to be replicated globally. Rapid response to new introductions requires being able to identify an alien species as soon as it establishes. This requires not only implementation of some kind of screening procedure, which might sensibly be based on risk management principles, but also the existence of appropriate taxonomic and, increasingly, molecular genetics skills. With declining funding for taxonomy and fewer taxonomists being trained worldwide (Godfray 2002, Hopkins and Freckleton 2002), a potential problem in early detection is looming rapidly. There have been successful, if expensive, campaigns to completely eradicate alien invasive seaweeds, but in every case this has relied on early detection (Anderson 2007). There are limited options to control invasive seaweeds once the opportunity for eradication has passed (Anderson 2007). In this context it is interesting that modelling seaweed invasions and various options for their control has not featured in responding to establishment of alien seaweeds. This omission is puzzling given that modelling could be highly beneficial in providing understanding of the epidemiology of invasions, predicting the course of seaweed invasions and thus improving strategic responses to invasions and informing optimum allocation of effort for surveillance, and in helping assess various options for control. Models would be of particular value if they could be calibrated against the known dynamics of initial range expansion. In this context we note that a plethora of techniques has been brought to tracking seaweed invasions (Meinesz 2007), but little attempt has been made to use this information in models. Given problems with remotely sensing invasive seaweeds, there is necessarily a large ‘‘on ground’’ effort if an accurate estimate of the pattern of range expansion of an invasive seaweed is to be ascertained (Meinesz 2007). The high cost of obtaining these kinds of data from field-based monitoring begs optimum use of the information, and this surely must include modelling. There exist a plethora of both well-established and emerging techniques to predict species’ distributions (Guisan and Zimmerman 2000, Elith et al. 2006), model the epidemiology of spread, including the effects of unpredictable long-distance dispersal events (e.g., Russell et al. 2005) which might be a hallmark of many marine invasions (Kinlan and Gaines 2003), and account for the spatial structure of populations in considering control options (e.g., Travis and Park 2004).

Coda We have identified a raft of questions and challenges that will, hopefully, help shape research efforts into invasive seaweeds into the future. No task is more pressing how-

ever than the need for better integration of research on invasive seaweed species with ‘‘mainstream’’ theory on ecological invasions. Seaweed ecologists are occupied with many of the questions that are more broadly considered in invasion biology; how the likelihood, nature and impact of invasions are influenced by propagule pressure, the nature of the recipient community, the role of disturbance, and biology of the invader. The critical gains in understanding, not only of seaweed invasions but of invasion dynamics in the broad, will arise through a heady mix and purposeful integration of in situ fieldbased experiments, empirical observation at a range of scales, and ecological theory.

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Author information Anderson, Lars W.J. USDA-Agricultural Research Service Exotic and Invasive Weed Research One Shields Ave. Davis, CA 95616 USA E-mail: [email protected] Booth, David School of Biological Sciences Queen’s University Belfast BT9 7BL Northern Ireland E-mail: [email protected] Campbell, Marnie L. National Centre for Marine and Coastal Conservation Australian Maritime College Private Mail Bag 10 Rosebud, Victoria 3939 Australia E-mail: [email protected] Chapman, Anthony R.O. Department of Biology Dalhousie University Halifax, NS Canada B3H 4H9 E-mail: [email protected] Doelle, Meinhard Marine & Environmental Law Institute and Dalhousie Law School Dalhousie University Halifax, NS Canada B3H 4H9 E-mail: [email protected] Dunstan, Piers K. CSIRO Marine and Atmospheric Research GPO Box 1538 Hobart, Tasmania 7001 Australia E-mail: [email protected] Hewitt, Chad L. National Centre for Marine and Coastal Conservation Australian Maritime College Private Mail Bag 10 Rosebud, Victoria 3939 Australia E-mail: [email protected]

Johnson, Craig R. School of Zoology and Tasmanian Aquaculture and Fisheries Institute University of Tasmania GPO Box 252-05 Hobart, Tasmania 7001 Australia E-mail: [email protected] Maggs, Christine A. School of Biological Sciences Queen’s University Belfast BT9 7BL Northern Ireland E-mail: [email protected] Magierowski, Regina H. School of Zoology and Tasmanian Aquaculture and Fisheries Institute University of Tasmania GPO Box 252-05 Hobart, Tasmania 7001 Australia E-mail: [email protected] McConnell, Moira L. Marine & Environmental Law Institute and Dalhousie Law School Dalhousie University Halifax, NS Canada B3H 4H9 E-mail: [email protected] Meinesz, Alexandre Laboratoire Environnement Marin Littoral Université de Nice-Sophia Antipolis Parc Valrose 06100 Nice cedex 2 France E-mail: [email protected] Pickering, Timothy D. School of Marine Studies The University of the South Pacific Private Mail Bag Suva Fiji Islands E-mail: [email protected] Provan, Jim School of Biological Sciences Queen’s University Belfast BT9 7BL Northern Ireland E-mail: [email protected]

Schaffelke, Britta CRC Reef Research PO Box 772 Townsville, Queensland 4810 Australia E-mail: [email protected] Skelton, Posa International Ocean Institute Regional Center for Australia and the Western Pacific P.O. Box 1539 Townsville, Queensland 4810 Australia E-mail: [email protected] Sulu, Reuben J. Solomon Islands Center The University of the South Pacific P.O. Box 460 Honiara Solomon Islands E-mail: [email protected] Valentine, Joseph P. Marine Research Laboratories Tasmanian Aquaculture and Fisheries Institute University of Tasmania GPO Box 252-49 Hobart, Tasmania 7001 Australia E-mail: [email protected] VanderZwaag, David L. Marine & Environmental Law Institute and Dalhousie Law School Dalhousie University Halifax, NS Canada B3H 4H9 E-mail: [email protected]

Subject index Special Issue Seaweed Invasions (Note: Grouped entries refer to expansive treatment of a topic; they are marked in bold face)

aboriginal issues 124 Acanthophora spicifera 21, 55, 79, 82, 87 accidental/unintentional introductions 1–3, 6–17, 18, 21, 23, 24, 25, 26, 27, 87, 91, 118, 119, 120, 121, 122, 131 acetic acid to control introductions 107, 108, 109 acrolein, herbicide 107, 111 Acrothamnion preissii 53, 59, 81, 84, 87 adaptive radiation 66 aerial photography 57, 59 AFLP analysis 71 AHTEG 120, 121, 123 algal types (strategic resource allocation) 2, 23, 32, 33, 35, 71, 77, 82 algicide 107, 112 alien species, definition 54 allele frequency 65, 70, 72 allelopathy 113 alpha diversity 134 alternative stable states 36, 37 aluminum to control introductions 108, 110 amphipods 26, 81, 82 ancient DNA (herbarium) 68 anemones 90 Antarctic Treaty area 123 anti-fouling compounds 4, 8, 11, 12, 91, 107, 122, 123, 131, 134 Antithamnionella elegans 56 Antithamnionella spirographidis 13 Antithamnionella ternifolia 13 aquaculture (mariculture) 2, 4, 6, 7, 8, 9, 10, 11, 12, 13, 18–30, 54, 55, 65, 87, 91, 92, 110, 112, 118, 119, 121, 122, 127, 128, 131, 132 aquarium trade 6, 7, 8, 9–10, 11, 13, 23, 54, 55, 65, 67, 69, 113, 114, 118, 121, 122, 126 ascidians 9, 84 Ascophyllum nodosum 4, 10, 87, 88, 103, 104 asian clam 77, 133 Asparagopsis 20, 56, 66, 69, 90, 91 Asparagopsis armata 56, 69, 90, 91 Asparagopsis taxiformis 56 assignment tests (genetic) 71–72 Avrainvillea amadelpha 79, 87 BACI sampling 84, 85 bait 8, 10, 67 balance of nature 41 ballast dry/semi-dry 7, 8, 9, 12 ballast water 1, 4, 6, 7, 9, 11, 12, 13, 55, 122, 123, 124, 126, 131, 134 barren grounds 34, 35, 36 Bayesian methodology 68, 70, 71, 72 benthic sled 58 Bern Convention 25, 125, 129 biodiversity protection 119, 120 biodiversity → see species richness

biogeographic region (bioregion) 1, 7, 10, 13, 131, 135 biological control of introductions 22, 104–107, 113 biosecurity (→ see also quarantine) 6, 10, 58, 77, 88, 90, 100, 104 Bonn Convention 118, 119, 121 Bonnemaisonia hamifera 13, 79 boring organisms 7, 8, 9 bottleneck, genetic 3, 66, 68, 69 burden of proof 18, 24, 28, 127 BWM Convention 2004 11, 122, 123, 126, 128 canopies of algae 21, 31, 33, 34, 35, 36, 37, 50, 78, 81, 83, 84, 86, 90, 114, 132, 133 carfentrozone, herbicide 107, 111 carrying capacity 44 cartography 53–64 CASI spectrographic imaging 57, 59 Caulerpa brachypus var. parvifolia 89 Caulerpa filiformis 79, 104 Caulerpa mexicana 55, 56, 67 Caulerpa ollivierii 89 Caulerpa racemosa var. cylindracea 53, 54, 55, 56, 57, 58, 59, 62, 70, 79, 80, 86, 90, 106 Caulerpa racemosa var. turbinata 70 Caulerpa scalpelliformis 56 Caulerpa taxifolia 3, 4, 10, 13, 20, 21, 22, 27, 32, 33, 34, 35, 36, 53, 54– 55, 56, 57, 58, 59, 60, 61, 62, 67, 69, 70, 71, 78, 79, 80, 81, 85–86, 87, 88, 89, 90, 91, 98, 99, 101, 102, 103, 104, 106, 107, 108, 109, 110, 132 caulerpenyne 80, 81 caulerpin 81 CEC 124, 125 Ceramium 10 Chara connivens 9 chlorine to control introductions 98, 101, 107, 108, 111 ciliates 81 climate change 4, 77, 89, 92, 112, 118 clonal reproduction, species 50, 56, 66, 69 Code of Conduct for Responsible Fisheries 4, 24, 121 Code of Practice for Introductions and Transfers of Marine Organisms 11, 24, 91, 121 Codium fragile ssp. tomentosoides 2, 3, 10, 12, 13, 19, 20, 21, 22, 31, 32, 33, 34, 35, 36, 37, 38, 50, 54, 56, 59, 66, 67, 68, 69, 71, 78, 81, 82, 85, 87, 89, 90, 99, 105, 106, 119, 132, 133, 134 comb jelly 77, 133 command and control protection 128 commercial seaweeds 19, 10, 18–30 Commission for Environmental Cooperation → see CEC

community composition change 78, 84, 85, 86, 89 community dynamics, ecological 1, 3, 31, 33, 35, 38, 41, 42, 43, 44, 45, 46, 47, 49, 50, 66, 70, 87, 132, 133, 134 compensation plans 128 competition 3, 13, 33, 34, 36, 42, 43, 44, 45, 47, 49, 50, 66, 77, 78, 79, 80, 82, 84, 85, 86, 87, 89, 90, 91, 99, 110, 111, 112, 113, 124, 132 competitive reversals 46, 47, 48, 49 complexity of communities 31, 35, 36, 37, 38, 41, 46, 50, 78, 86, 132, 133 Conference of the Parties (COP) 120, 121 connectivity 41, 43, 45, 46, 47, 49 conserved gene sequences 67 Constitution of the Oceans 119 containment of introductions 9, 10, 25, 27, 87, 88, 98, 99, 100, 101, 102, 103, 104, 105, 108, 109, 110, 113, 119, 120, 126 contingency response 120, 128 control of introductions 1, 4, 8, 10, 11, 21, 22, 24, 25, 28, 31, 36–38, 53, 54, 60, 66, 77, 78, 87–89, 91, 92, 98–117, 118, 119, 120–128, 131, 134, 135 Convention on Biological Diversity (CBD) 4, 25, 92, 118, 119, 120, 122, 123, 125, 127 Convention on Harmful Anti- Fouling Systems on Ships 4, 12, 123, 131 COP → see Conference of the Parties copper to deter introductions 7, 8, 12, 98, 107, 108, 109, 110 coral reef degradation 18, 20, 22, 27, 55, 82, 83, 87, 88, 131 Corallina 33, 84, 104 “cottonii” cultivar 19, 20 cryptic invasions 3, 53, 57, 65, 66, 67, 68, 91 ctenophore → see comb jelly Cymodocea nodosa 79, 80, 85 decline in canopy algae (global) 36 Desmarestia aculeata 82 detection of aliens 3, 4, 8, 9, 10, 26, 53, 54, 56, 57, 58, 59, 60, 61, 62, 68, 71, 72, 77, 99, 101, 102, 131, 133, 134, 135 developing (less-developed) countries 2, 8, 18, 20, 24, 27, 131 diclobenil, herbicide 111 diquat, herbicide 107, 111 disease 11, 19, 24, 26, 27, 34, 35 dispersal 4, 21, 23, 26, 32, 33, 35, 37, 56, 57, 59, 70, 78, 90, 99, 102, 105, 113, 132, 135 disturbance (ecological) 3, 13, 21, 31, 33, 34, 35, 36, 37, 38, 43, 44, 45, 50,

Subject index

81, 86, 87, 89, 90, 91, 92, 103, 104, 121, 132, 133, 134, 135 diuron, herbicide 108, 111 drainage to eradicate introductions 103 dredging for control/eradication 102, 103, 105, 113 dry dock hull cleaning 11, 12 drying for control/eradication 105 Ecklonia radiata 108 ecological consequences of invasion 1, 2, 3, 26, 91, 133 ecological theory 1, 2, 41, 135 economic incentives 128 economic instruments 128 economics 1, 2, 3, 6, 11, 18–30, 65, 77, 78, 80, 86, 87–89, 90, 91–92, 98, 99, 102, 107, 112, 118, 120, 124, 128, 129, 131, 132, 133, 134 ecosystem 20, 77, 23, 24, 25, 27, 31, 33, 36, 37, 41, 42, 53, 65, 67, 77, 78, 89, 90, 92, 98, 99, 102, 107, 110, 112, 118, 119, 120, 121, 123, 124, 125, 126, 127, 128, 133, 134 effective population size 68, 69 Elton 41, 42, 132 endothall, herbicide 111 environmental forcing 37, 47, 48 equilibrium view of communities 43, 44, 67 eradication of introductions 3, 4, 25, 26, 53, 55, 60, 61, 62, 67, 77, 78, 87, 88, 91, 92, 98–117, 119, 120, 121, 125, 126, 135 establishment of introductions 2, 9, 10, 12, 20, 23, 25, 31, 32, 33, 34–35, 36, 37, 41, 44, 50, 55, 58, 59, 65, 78, 81, 84, 86, 87, 89, 90, 91, 92, 107, 110, 111, 112, 127, 131, 132, 134, 135 Eucheuma denticulatum 18, 20, 22, 23 Eucheuma striatum 82 eucheumoids 19, 20 European initiatives on invasives 125 evolutionary consequences of invasion 56, 65, 66, 77, 78, 134 experimental approaches 1, 21, 34, 35, 37, 78, 85, 86, 101, 132, 133, 135 expressed sequence tags 70 extinction 44, 48, 66, 67, 69 FAO 19, 24, 27, 118, 119, 121 FAO Guidelines for Responsible Fisheries 24, 91 FAO Technical Guidelines on the Precautionary Approach to Capture Fisheries and Species Introductions 121 fecundity of algae 10, 32, 50, 89 feedback mechanisms 11, 31, 35, 36, 37, 38, 89 fish 8, 9, 10, 22, 27, 36, 68, 70, 71, 72, 77, 78, 80, 81, 86, 87, 88, 98, 99, 102, 106, 110, 113, 114, 118, 119, 121, 127, 133 fishing 4, 10, 21, 22, 24, 27, 36, 56, 60, 67, 77, 88, 102, 118, 119, 122, 123, 126, 131 fitness 66, 69 fluridone, herbicide 107, 109 freshwater to control introductions 10, 12, 26, 98, 107, 108, 109, 110, 113 142

freshwater weeds 98, 104, 110, 112, 113, 114 Fucus ceranoides 70 Fucus evanescens 3, 66, 70, 82, 90, 91 Fucus serratus 3, 31, 32, 33, 34, 35, 36, 66, 68, 69, 70, 83, 91 Fucus spiralis 70 Fucus vesiculosus 70, 83 functional traits of seaweeds 6, 12, 33, 90 furanone, herbicide 108 gamma diversity 134 gastropods 104 gene pool composition 65, 134 genetic consequences of invasion 2, 3, 24, 56, 65–76, 78, 121, 133, 134 genetic impoverishment/depauperacy 68 genetic integrity 66 genetics 2, 3, 22, 24, 54, 55, 56, 62, 65–76, 78, 85, 91, 92, 105, 121, 132, 134, 135 genomics 2, 3, 70–71, 133 Gigartina 10, 19 Gigartina skottsbergii 19 global sectoral initiatives 121–123 global warming 36, 112 glyphosate to control introductions 108 Gracilaria coronopifolia 106 Gracilaria edulis 22 Gracilaria salicornia 21, 82, 87, 106, 108 Gracilaria vermiculophylla 68 Grateloupia 9, 13, 55, 90 Grateloupia doryphora 55, 90 Grateloupia subpectinata 13 Grateloupia turuturu 13, 55, 90 grazing/herbivory 3, 13, 22, 23, 27, 32, 34, 35, 36, 37, 50, 78, 81, 82, 86, 87, 89, 90, 91, 98, 104, 105, 106, 107, 110, 111, 112, 113, 134 green crab 133 growth rates of introduced species 32, 33, 45, 50, 68, 70, 86, 89, 101, 108 Gulf of Maine Action Plan 125 Gulf of Maine Council 124, 125 Gymnogongrus crenulatus 9 habitat change 4, 20, 34, 65, 77, 78, 85, 98, 118, 120, 121 Halophila hawaiiana 79 hand removal of introductions 103, 104 haplotype network 69 Hawaii 10, 18, 21, 22, 26, 55, 58, 79, 82, 83, 85, 87, 88, 91, 92, 103, 106, 111 health, human 78, 107, 112, 118, 119, 122, 127 heat to kill invaders 8, 88, 98, 104, 105, 113 herbicides 98, 105, 107, 109, 110, 111, 112, 113, Heterosiphonia japonica 13, 82 heterozygote deficiency 69 hexazinone, herbicide 111 “hitch-hikers” 2, 9, 18, 26 Hizikia fusiformis 19 homogenization, biotic 77 hull cleaning, in water 12 hull fouling 1, 2, 6, 8, 11–13, 55, 122–126, 131, 134 hybridisation 3, 55, 65, 70, 71, 72, 78 hydrogen peroxide as control agent 109 hydroids 84

Hypnea 19, 20, 82, 87, 88 Hypnea musciformis 21, 82, 87, 88 hystereses, ecological 31, 36, 37 ICES Code of Practice on the Introduction and Transfers of Marine Organisms 24, 91, 121 IMO (International Maritime Organization) 118, 119, 121, 122, 123, 127, 131 impact assessment 119 impacts of introductions 1–3, 10, 11, 13, 18, 20, 21, 23, 24, 25–28, 33, 53, 54, 59, 62, 66, 77–97, 98, 99, 102, 104, 107, 109, 110, 112, 113, 118–120, 121, 122, 128, 131–133, 134, 135 industry, seaweed 2, 9, 18–30, 112 insurance 128 intentional introductions 1, 2, 6, 7, 9–11, 18–30, 87, 91, 92, 118–122, 124, 126, 127, 131 International Convention for Management and Control of Ships’ Ballast Water and Sediments → see BWM Convention 2004 International Convention for the Prevention of Pollution from Ships 73/74 → see MARPOL 73/74 introgression 66, 69, 70, 71, 78 invasibility of communities 13, 15, 31, 32, 33, 40, 41, 42, 43 invasion process 1, 2, 6, 7, 11, 12, 13, 14, 31, 32, 34, 36, 37, 38, 65, 67, 68, 70, 89, 93, 131, 132, 133 invasional meltdown 90, 134 invasive species response model 100, 114 invertebrates 8, 11, 33, 46, 71, 77, 78, 81, 82, 84, 86, 98, 99, 110, 112, 133 ISSR markers 55, 67 ITS sequences 67, 70 IUCN “100-of-the-worst list” 21 IUCN 7, 13, 21, 131 Jakarta Mandate 120 Kappaphycus alvarezii 10, 18, 19, 20, 21, 22, 23, 25, 26, 27, 28 kelp forest 3, 34, 35, 50, 81, 82, 83, 87, 90, 105, 134 Laminaria japonica → see Saccharina japonica land-based aquaculture 11, 27 Law of the Sea (UNCLOS) 4, 24, 118, 119, 120, 123 legal issues 3, 4, 10, 11, 23, 24, 90, 102, 118–130, 135 legislation 4, 25, 27, 121, 125, 128, 135 levels of invasion 48, 49, 59 liability (legal/financial) 120, 128 life history 2, 3, 6, 8, 11, 12, 14, 19, 21, 25, 31, 32, 33, 37, 41, 56, 57, 66, 69, 132 limiting similarity-model assumption 43 “live” rock 10 live seafood trade 6, 8, 10, 13 Lomentaria hakodatensis 13 Macrocystis pyrifera 25, 26, 27, 57, 83 management by resource limitation 110

Subject index

management of introductions 28, 31, 36, 37, 38, 54, 65, 77, 78, 88, 89, 91, 92, 98, 99, 100, 101, 102, 104, 110, 111, 112, 113, 114, 122, 125, 128, 135 mapping (→ see also cartography) 53–64 maritime equipment 8, 10, 87 MARPOL 73/74 122 mating system 66, 67, 68, 70, 134 mechanical methods to control introductions 98, 103, 104, 105, 131 mechanisms of invasion 1, 2, 21, 26, 31–40, 41–52, 57, 90, 132, 133 Membranipora membranacea 34, 35, 36, 50, 87, 90, 134 messenger RNA 65, 71 microarrays 71 microsatellite DNA 67, 68, 69, 70, 71, 105 microscopic stages 8, 9, 21, 32, 35, 103, 104 mixed stock analysis 72 mode of introduction 2, 54, 56, 62, 65, 131 modelling invasion 41, 43, 44–49, 59, 60, 65, 68, 69, 72, 100, 102, 124, 132, 135 molecular approaches 3, 22, 53, 54, 55, 65–76, 134, 135 molecular markers 67, 68 molluscs 77, 81, 84, 98, 104, 133 monitoring 4, 18, 24, 26, 28, 53–64, 87, 88, 91, 100, 101, 104, 105, 125, 127, 135 mortality 42, 44, 45, 46, 47, 48, 49, 50, 79, 80, 82, 83, 85, 132 mosaic, community 42, 43, 50, 87 multiple introductions 3, 20, 65, 67, 69, 90 national level law and policy 125 Neosiphonia harveyi 83 network topology 43, 45, 46, 47, 49 next pest 2, 6, 31, 132, 134 niche 13, 33, 37, 41, 43, 49, 66, 72 non-spatial model 44, 45, 47, 48 North American cooperation 124–125 North American Invasive Species Network 124 nutraceuticals 19 nutrients (resource for algae) 13, 22, 23, 35, 38, 50, 79, 85, 87, 88, 89, 98, 107, 110, 111, 112, 113, 114, 132 overfishing 4, 36, 77, 131 oyster culture associates 9, 53, 54 ozone treatment 107 Pacific islands 20, 26, packing material as a vector 6, 7, 8, 10, 12, 13 Palmaria palmata 69 parapatric speciation 66 patch structure/dynamics 3, 33, 41–52, 69, 81, 86, 101, 102, 108, 133, 134 PCR 67, 68 permitting process 128 persistence of introduced species 21, 31, 32, 35–36, 37, 50, 86, 132 pharmaceuticals 19 phase shift (community) 36, 37, 91 Phloiocaulon 8 Phyllospora comosa 84 physiology, algal 2, 9, 12, 36, 38, 107, 110 Pikea californica 67

plastid genome 67, 70 policy responses to alien species 1, 4, 118–130 pollution 36, 80, 89, 119, 120, 122, 128, 131 polychaetes 26, 80, 84, Polysiphonia brodiaei 13 Polysiphonia harveyi (=Neosiphonia harveyi) 13, 66, 68, 69, 83 Polysiphonia strictissima 68 population dynamics 47, 48, 70 population structure 65, 68, 69 Porphyra yezoensis 9, 28 Posidonia oceanica 35, 79, 80, 85, 110 Posidonia oceanica 80, 85 postglacial colonization 69 precautionary approach/principle 23, 24, 118, 120, 121, 125, 126, 127, 128, 129 predation 2, 26, 36, 49, 77, 90, 133 predator impacts 133 predicting introductions 2, 3, 6, 12, 13, 18, 23, 31, 33, 37, 44, 57, 59, 60, 65, 68, 72, 89, 91, 92, 124, 132, 133–135 prevention of introductions 11, 24, 25, 38, 92, 98, 100, 119, 120, 124, 125, 126, 127, 128, 131 propagules 7, 9, 11, 21, 23, 26, 28, 32, 35, 37, 41, 46, 47, 49, 50, 56, 70, 99, 101, 102, 104, 105, 108, 110, 135 proportion of aliens in flora 1, 21, 131 proteome 65 public awareness 21, 22, 27, 53, 54, 57, 87, 101, 124 public education (outreach) 3, 54, 78, 88, 101, 125 Punctaria 8 QTL mapping 71 quarantine 2, 9, 10, 11, 23, 24, 25, 26, 27, 28, 91, 92, 102, 103, 118 Ramsar Convention 118, 119, 121 random walk 44 rate of invasion 1, 126 rbcL 68 recipient community, traits of 2, 3, 6, 12, 14, 25, 31–40, 78, 89, 132–135 recruitment 8, 22, 33, 35, 37, 42, 45, 46, 47, 48, 49, 50, 66, 80, 81, 83, 86, 87, 90, 104, 133 refugial population 69 regional coordination and cooperation 123–125 regulations, legal, political 1, 2, 4, 10, 11, 23, 27, 28, 99, 107, 118–130, 134, 135 regulatory tools 127 remote sensing 3, 91 reproduction, algal 4, 10, 12, 22, 23, 26, 27, 32, 43, 50, 56, 57, 67, 69, 70, 90, 102, 103, 104, 112, 113 resilience stability 134 resistance stability 134 resistance to invasion → see also invasibility 12, 33,34, 35, 38, 41, 42, 43, 45, 46, 47,48, 49, 50, 132, 134 resource (commercial wild stock) 6, 19, 20, 77, 89, 92, 118, 119, 122, 127

resource (ecological) availability 3, 13, 31–40, 41–52, 89, 110, 111, 112, 113, 132 resource (ecological) use 13, 41, 42, 43, 44, 46, 47, 48, 49 resource supply 42, 49, 50 resource variability 3, 42, 43, 44, 46, 47, 48, 50, 89 response to introductions 1, 2, 3, 4, 22, 23, 25, 37, 57, 59, 66, 68, 71, 77, 87, 88, 92, 98–100, 102, 104, 110, 112, 114, 118–130, 131–135 restriction fragment length polymorphism 67 rhodoliths 82 Rhodomela larix 83 risk assessment 2, 9, 11, 18, 19, 24, 25, 26, 27, 28, 121, 128 risk mitigation 2, 6, 9, 10, 11, 23, 24, 27, 28 risk of introduction 2, 6, 9, 10, 11, 12, 13, 18–28, 31, 35, 50, 54, 65, 78, 92, 102, 105, 110, 114, 118, 120, 121, 122, 123, 124, 126, 127, 128, 131, 132, 134, 135 rotovation for control/eradication 105 ROV (remotely operated vehicle) 57 Ruppia maritima 80, 86 Saccharina (Laminaria) japonica 9, 18, 19, 25, 54 sacoglossans as control agents 104, 106 “sacol” cultivar 20 salt to control introductions 88, 98, 108, 109, 110 sampling design 3, 4, 53, 57, 69, 71, 72, 108 Sargassum muticum 9, 12, 20, 21, 22, 27, 31, 32, 33, 34, 35, 36, 37, 53, 54, 57, 58, 60, 66, 78, 83, 84, 85, 86, 88, 90, 102, 132, 133 satellite teledetection 57, 59 scale (spatio-temporal) 1, 3, 6, 9, 11, 13, 19, 25, 35, 37, 42, 43, 44, 47, 48, 49, 50, 57, 58, 60, 61, 66, 69, 70, 77, 86, 88, 90, 91, 92, 99, 102, 105, 107, 110, 133, 134, 135 scale dependence 1, 3, 37, 42, 47, 48, 49, 57, 58, 66, 68, 77, 86, 90, 91, 92, 99, 105, 110, 133, 135 scientific research, alien escapes from 7, 8, 10, 11, 13, 23, 121 screw shell mollusc 133 SCUBA survey 53, 57, 59, 60, 101, 102 sea star 77 sea urchins (→ see also barren grounds) 22, 34, 35, 36, 37, 80, 81, 82, 83, 86, 87, 98, 104, 105, 106, 133 seagrass 10, 22, 27, 34, 35, 36, 78, 79, 80, 84, 85, 86, 89, 98, 99, 113, 132 sediments 4, 8, 9, 11, 35, 36, 37, 66, 78, 80, 84, 86, 87, 89, 101, 103, 105, 107, 108, 109, 110, 111, 113, 122, 123, 126, 133 selection pressure (evolutionary) 8, 69, 71 shipping 4, 7, 11, 12, 19, 21, 23, 24, 26, 28, 36, 54, 65, 68, 87, 92, 99, 112, 114, 118, 119, 121, 123, 124, 126, 127, 128, 131, 135 143

Subject index

side scan sonar 57 simazine, herbicide 108, 111 SNP maps 71 socio-economics 19, 23, 24, 25, 26, 77, 78, 92, 102, 118, 119, 132, 134 sources of introductions 3, 11, 53, 54, 55, 65, 66, 67, 68, 69, 70, 72, 105, 128 space monopolization 77, 78, 85, 87, 89, 90 spatial organization (individuals) 41, 42, 43, 45, 48, 50 spatially explicit model 43, 44, 45, 49 species co-existence 49, 79 species richness/diversity 3, 4, 6, 13, 18, 19, 21, 23, 25, 33, 38, 41–52, 54, 65, 77, 78, 79, 80, 81, 82, 84, 86, 87, 89, 90, 91, 92, 98, 99, 118, 119, 120, 121, 122, 125, 127, 132, 133, 134 “spinosum” cultivar 20 spread modelling 59 spread of introduced species 2, 10, 12, 18, 21, 22, 27, 31, 32, 34, 35, 37, 41, 42, 47, 53, 54, 55, 56, 57, 58, 59–60, 62, 70, 77, 78, 86, 87, 89, 91, 98, 99, 101, 102, 114, 118, 120, 122, 123, 125, 127, 132, 133, 134, 135 stability of communities 34, 36, 37, 41, 43, 134 strategies and action plans (legal/ political) 120 stress tolerance 33 Striaria attenuata 13

144

substratum 9, 12, 13, 21, 22, 33, 34, 35, 78, 82, 86, 89, 98, 100, 102, 103, 104, 105, 107, 110, 133 subtractive hybridization 71 surveillance 92, 99, 100, 101, 102, 104, 105, 108, 113, 135 survey 1, 21, 22, 24, 25, 27, 60, 68, 77, 85, 89, 90, 101, 131, 133, 134 synergisms 77, 90, 134 system management 37 Tasmania 3, 8, 10, 21, 33, 35, 36, 77, 85, 86, 91, 104, 132, 133 taxonomic training 135 taxonomy 28, 53, 54, 70, 88, 100, 132, 135 TBT → see tributyl tin Thau lagoon 9, 54 top-down vs. bottom-up invader effects 134 toxicity of invaders 9, 36, 77, 78, 81, 85, 86 tracking introductions 1, 2, 3, 53–64, 68, 91, 134, 135 transcriptome 65, 71 transport → see vectors treatment options for control/eradication 6, 11, 100 tributyl tin 4, 11, 12, 81, 91, 123, 134 trophic structure 13, 78 turfing algae 3, 33, 81, 84, 86, 89, 90 Ulva fasciata 82 unavoidable risk 126

Undaria pinnatifida 2, 3, 8, 9, 10, 13, 18, 19, 20, 21, 22, 23, 25, 27, 31, 32, 33, 34, 35, 36, 37, 55, 56, 78, 84, 85, 86, 87, 88, 89, 90, 91, 99, 104, 108, 112, 132, 133 understorey (algal) 33, 83, 84, 86 United Nations Environment Program (UNEP) 123 United States Federal Task Force on Invasive Species 125 United States National Invasive Species Act 124 vacant niche 33, 66, vectors of introduction 4, 6, 7, 8, 9, 10, 12, 13, 18, 21, 23, 25, 26, 28, 41, 53, 54, 55, 65, 67, 91, 118, 119, 122, 123, 126, 131, 134 vegetative propagation 3, 9, 12, 19, 21, 22, 26, 32, 35, 45, 55, 56 Venice lagoon 83, 84 voluntary compliance agreements 128 von Neumann neighbourhood 45 watch lists 2, 21 Western Regional Panel 124, Womersleyella setacea 56, 81, 84, 85, 86 Zostera capricorni 109 Zostera marina 10, 80, 83, 86, 108, 109