Seaweed Invasions
Edited by
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
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 tak-en 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
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 trans-mitted 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.
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
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
6
Intentional introductions of commercially
harvested alien seaweeds
Timothy D. Pickering, Posa Skelton
and Reuben J. Sulu
18
Mechanisms of invasion: establishment,
spread and persistence of introduced
seaweed populations
Joseph P. Valentine, Regina H.
Magierowski and Craig R. Johnson
31
Mechanisms of invasion: can the recipient
community influence invasion rates?
Piers K. Dunstan and Craig R. Johnson
41
Methods for identifying and tracking
seaweed invasions
Alexandre Meinesz
53
Molecular approaches to the study of
invasive seaweeds
David Booth, Jim Provan and Christine
A. Maggs
65
Impacts of introduced seaweeds
Britta Schaffelke and Chad L. Hewitt
77
Control of invasive seaweeds
Lars W.J. Anderson
98
Invasive seaweeds: global and regional
law and policy responses
Meinhard Doelle, Moira L. McConnell
and David L. VanderZwaag
118
Conclusion
Seaweed invasions: conclusions and
future directions
Craig R. Johnson
131
Author information
139
Subject index
141
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
1School of Zoology and Tasmanian Aquaculture and
Fisheries Institute, University of Tasmania, GPO Box 252-05, Hobart, Tasmania 7001, Australia,
e-mail: [email protected]
2Department of Biology, Dalhousie University, Halifax,
NS, Canada, B3H 4J1 * 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 anthropogeni-cally-mediated translocation of species to regions out-side 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 bal-last water alone (Carlton and Geller 1993, Carlton 1999). Furthermore, there is good evidence to suggest that in some areas establishment of alien marine species origi-nating from hull fouling exceeds that attributable to trans-port in ballast water (Hewitt et al. 2004). Not surprisingly, the rate of establishment of alien marine species, includ-ing invasives, appears to be increasinclud-ing (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, fresh-water 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 sev-eral species have been invasive (e.g., Nyberg and Wal-lentius 2005, this special issue). However, as is the case in studying many other kinds of invasive marine species (Grosholz 2002), investigation of the seaweed compo-nent 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 refer-ence, for they comprise the knowledge base on the spe-cies 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 vari-ety 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 sur-vey, 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 Tur-kington 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 opport-unity 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 inva-sion ecology. We consider the means, both accidental and intentional, by which seaweeds are introduced to new biogeographic domains, mechanisms of their inva-sion 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 ele-ment of the interaction between society and invasive seaweeds. Inevitably, this work is also about identifying gaps, and, therefore, challenges and priorities for the future.
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 sea-weed invasions?
• Why do some communities appear to be more sus-ceptible 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
con-sequences of seaweed invasions?
• Can we expect that existing and, in particular, emerg-ing 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 deter-mined adequate?
We do not expect all of these questions to be answered with equal conviction, but we should be san-guine 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 empir-ical, although the interface of some of these issues with ecological theory is considered where appropriate. Dun-stan and Johnson (2007) in particular address the ques-tion 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 sea-weed species are intentional, Hewitt et al. (2007) focus
on reviewing modes of accidental introductions and iden-tify hull fouling as the most significant, but also the most poorly managed, transport mechanism for seaweeds. They emphasise that while eliminating risk is rarely pos-sible, 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 in-vasion 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 com-mon suite of traits of invasive seaweeds, in line with sug-gestions two decades ago (e.g., Crawley 1987).
But there is also pressure for further intentional intro-ductions. This is driven by ongoing demand for sea-weeds and their products, and perceptions that seaweed based industry offers an alternative and sustainable live-lihood 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 how-ever, Pickering et al. (op. cit.) find that intentionally intro-duced seaweeds are no more or less risk prone than unintentionally introduced seaweeds. Not surprisingly then, they too consider a careful risk assessment essen-tial when there are plans to introduce species for aqua-culture purposes. Clearly, species being considered for aquaculture should not be on watch lists of invasive spe-cies maintained by government agenspe-cies, NGOs or inter-governmental 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 anoth-er (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 con-sider 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 physio-logical attributes. Species tolerances to light, tempera-ture 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, Val-entine 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 Tas-mania, 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 larg-er spatial scales. Codium fragile ssp. tomentosoides does not reach nuisance proportions in all of the com-munities in which it has been introduced. This subspe-cies 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 (Chap-man et al. 2002). In fact several species that are invasive elsewhere in the world are not pests in their native com-munities (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 man-ifest 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 involve-ment 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 aware-ness campaign that distributed 300,000 brochures across eight Mediterranean countries. Cartographical data from an informed public, along with expert sam-pling, 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. tomen-tosoides) and Tuscany (on two invasive Caulerpa species co-existing with two introduced red turfing algae from Australia), there have been few comprehensive experi-mental works on seaweed invasion ecology. In most cas-es 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 differ-ent 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 Medi-terranean 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 com-plexity 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 consider-able genetic diversity within invader populations of Unda-ria pinnatifida, probably reflecting multiple introductions from different sources. This species is a vigorous invader, first found outside its native Japanese range in the Med-iterranean Sea in 1971 (Meinesz 2007), but with sub-sequent 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 inva-sive 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 hybridisa-tion 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. ser-ratus L.; interestingly, the hybrids occur in a restricted hybrid zone on the shore. Although hybridisation involv-ing 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 genotype-phenotype-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
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 devel-opment effort. Nonetheless, eradication in concert with early detection emerges as a cost effective response (Anderson op. cit.).
No consideration of human responses to invasive spe-cies would be complete without consideration of the reg-ulatory 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 sci-entists or environmental agencies, but coordination among governments at a global scale is poor. For exam-ple, only Australia, New Zealand, USA, Canada, Switzer-land and Germany have legislation controlling introductions for aquaculture. This legislation has devel-oped from both global and regional policies including the Law of Sea, the Convention on Biodiversity, the Inter-national 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 Fish-eries by the Food and Agriculture Organization. The Inter-national 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 Con-trol of Harmful Anti-Fouling systems on Ships) will pro-mote increased fouling by seaweeds as the use of toxic compounds in antifoulants such as tributyltin (TBT) are phased out. Importantly, none of the international con-ventions 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 seri-ously than other nation states but, in general, the devel-opment 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
envi-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.
References
Anderson, L.W.J. 2007. Control of invasive seaweeds. Bot. Mar. 50: 418–437.
Bax, N., A. Williamson, M. Aguero, E. Gonzalez and W. Geeves. 2003. Marine invasive alien species: a threat to global bio-diversity. Mar. Policy 27: 313–323.
Booth, D., J. Provan and C.A. Maggs. 2007. Molecular approaches to the study of invasive seaweeds. Bot. Mar. 50: 385–396.
Bruno, J.F., J.D. Fridley, K.D. Bromberg and M.D. Bertness. 2005. Insights into biotic interactions from studies of species invasions. In: (D.F. Sax, J.J. Stachowicz and S.D. Gaines, eds) Species invasions: insights into ecology, evolution and biogeography. Sinauer Associates, Sunderland. pp. 13–40. Cadotte, M.W. 2006. Darwin to Elton: early ecology and the
problem of invasive species. In: (M.W. Cadotte, S.M. Mc-Mahon and T. Fukami, eds) Conceptual ecology and invasion biology: reciprocal approaches to nature. Springer, Dor-drecht. pp. 15–33.
Carlton, J.T. 1996. Marine bioinvasions: the alteration of marine ecosystems by nonindigenous species. Oceanography 9: 36–43.
Carlton, J.T. 1999. The scale and ecological consequences of biological invasions in the world’s oceans. In: (O.T. Sandlund, P.J. Schei and A˚. Viken, eds) Invasive species and biodiversity management. Kluwer Academic Publishers, Dordrecht. pp. 195–212.
Carlton, J.T. and B.J. Geller. 1993. Ecological roulette: the global transport of nonindigenous marine organisms. Science 261: 79–82.
Chapman, A.R.O. 1984. Reproduction, recruitment and mortality in two species of Laminaria in southwest Nova Scotia. J. Exp. Mar. Biol. Ecol. 78: 99–109.
Chapman, A.R.O. 1986. Population and community ecology of seaweeds. Adv. Mar. Biol. 23: 1–161.
Chapman, A.S., R.E. Scheibling and A.R.O. Chapman. 2002. Species invasions and changes in marine vegetation of Atlantic Canada. In: (R. Claudi, ed.) Alien invasive species: threats to Canadian biodiversity. Natural Resources Canada, Ottawa. pp. 133–148.
Cohen, A.N. and J.T. Carlton. 1998. Accelerating invasion rate in a highly invaded estuary. Science 279: 555–558. Crawley, M.J. 1987. What makes a community invasible? In:
(A.J. Gray, M.J. Crawley and P.J. Edwards, eds) Colonization, succession and stability. Blackwell Scientific, Oxford. pp. 429–453.
Daskalov, G.M. 2002. Overfishing drives a trophic cascade in the Black Sea. Mar. Ecol. Prog. Ser. 225: 53–63.
Daskalov, G.M. 2003. Long-term changes in fish abundance and environmental indices in the Black Sea. Mar. Ecol. Prog. Ser. 255: 259–270.
Davis, M.A. 2006. Invasion biology 1958–2005: the pursuit of science and conservation. In: (M.W. Cadotte, S.M. McMahon and T. Fukami, eds) Conceptual ecology and invasion bio-logy: reciprocal approaches to nature. Springer, Dordrecht. pp. 35–64.
Davis, M.A., K. Thompson and J.P. Grime. 2001. Charles S. Elton and the dissociation of invasion ecology from the rest of ecology. Divers. Distrib. 7: 97–102.
C.R. Johnson and A.R.O. Chapman: Seaweed invasions: introduction and scope 325 Didham, R.K., J.M. Tylianakis, M.A. Hutchinson, R.E. Ewers and
N.J. Gemmel. 2005. Are invasive species the drivers of eco-logical change? Trends Ecol. Evol. 20: 470–474.
Doelle, M., M.L. McConnell and D.L. VanderZwaag. 2007. Inva-sive seaweeds: global and regional law and policy responses. Bot. Mar. 50: 438–450.
Dunstan, P.K. and C.R. Johnson. 2007. Mechanisms of invasion: can the recipient community influence invasion rates? Bot. Mar. 50: 361–372.
Fridley, J.D., J.J. Stachowicz, S. Naeem, D.F. Sax, E.W. Sea-bloom, M.D. Smith, T.J. Stohlgren, D. Tilman and B. Von Holle. 2007. The invasion paradox: reconciling pattern and process in species invasions. Ecology 88: 3–17.
Grosholz, E.D. 1996. Contrasting rates of spread for introduced species in terrestrial and marine systems. Ecology 77: 1680–1686.
Grosholz, E.D. 2002. Ecological and evolutionary consequences of coastal invasions. Trends Ecol. Evol. 17: 22–27.
Gurevitch, J. and D.K. Padilla. 2004. Are invasive species a major cause of extinctions? Trends Ecol. Evol. 19: 470–474. Hewitt, C.L. 2003. Marine biosecurity issues in the world oceans: global activities and Australian directions. Ocean Yearbook 17: 193–212.
Hewitt, C.L., M.L. Campbell, R.E. Thresher, R.B. Martin, S. Boyd, B.F. Cohen, et al. 2004. Introduced and cryptogenic species in Port Phillip Bay, Victoria, Australia. Mar. Biol. 144: 183–202.
Hewitt, C.L., M.L. Campbell and B. Schaffelke. 2007. Introduc-tions of seaweeds: accidental transfer pathways and mech-anisms. Bot. Mar. 50: 326–337.
Kinlan, B.P. and S.D. Gaines. 2003. Propagule dispersal in marine and terrestrial environments: a community perspec-tive. Ecology 84: 2007–2020.
Levine, J.M. and C.M. D’Antonio. 1999. Elton revisited: a review of evidence linking diversity and invasibility. Oikos 87: 15–26. Lodge, D.M. 2001. Lakes. In: (F.S. Chapin, O.E. Sala and E. Huber-Sannwald, eds) Future scenarios of global biodiversity. Springer-Verlag, New York. pp. 277–312.
Lu¨ning, K. 1990. Seaweeds. Their enivironment, biogeography and ecophysiology. Wiley, New York. pp. 527.
MacDougall, A.S. and R. Turkington. 2005. Are invasive species the drivers or passengers of change in degraded ecosys-tems? Ecology 86: 42–55.
Mack, R.N., D. Simberloff, W.M. Lonsdale, H. Evans, M. Clout and F.A. Bazzaz. 2000. Biotic invasions: causes, epidemio-logy, global consequences, and control. Ecol. Appl. 10: 689–710.
Meinesz, A. 2007. Methods for identifying and tracking seaweed invasions. Bot. Mar. 50: 373–384.
Nichols, F.H., J.K. Thompson and L.E. Schemel. 1990. Remark-able invasion of San Francisco Bay (California, USA) by the Asian clam Potamocorbula amurensis. II. Displacement of a former community. Mar. Ecol. Prog. Ser. 66: 95–101. Nyberg, C.D. and I. Wallentius. 2005. Can species traits be used
to predict marine macroalgal introductions? Biol. Invasions 7: 265–279.
Pickering, T.D., P. Skelton and R.J. Sulu. 2007. Intentional intro-ductions of commercially harvested alien seaweeds. Bot. Mar. 50: 338–350.
Pimentel, D., L. Lach, R. Zuniga and D. Morrison. 1999. Envi-ronmental 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 sea-weeds. Bot. Mar. 50: 397–417.
Schaffelke, B., M.L. Campbell and C.L. Hewitt. 2005. Reproduc-tive phenology of the introduced kelp Undaria pinnatifida (Phaeophyceae, Laminariales) in Tasmania, Australia. Phy-cologia 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 cteno-phore Mnemiopsis leidyi and recent changes in pelagic com-munity structure. Fish. Oceanogr. 7: 305–310.
Stachowicz, J.J. and D. Tilman. 2005. Species invasions and the relationships between species diversity, community satura-tion, and ecosystem functioning. In: (D.F. Sax, J.J. Stacho-wicz 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 sub-species. 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 persist-ence of introduced seaweed populations. Bot. Mar. 50: 351– 360.
Review
Introductions of seaweeds: accidental transfer pathways and
mechanisms
Chad L. Hewitt1,*, Marnie L. Campbell1and
Britta Schaffelke2,a
1National Centre for Marine and Coastal Conservation,
Australian Maritime College, PMB 10, Rosebud, Victoria 3939, Australia,
e-mail: [email protected]
2CRC 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 inten-tional introductions (translocations for aquaculture, aquarium or live seafood trade) or accidental introduc-tions (mainly as hull-fouling). A number of potential man-agement 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 mecha-nism for macroalgae, is of urgent need. Our current ability to identify which species are likely to invade next is lim-ited. 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-medi-ated 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, Carl-ton 2001, Ruiz and CarlCarl-ton 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
a
3, Townsville, QLD 4810, Australia.
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-indige-nous) 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, Ore-sanz 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 environmen-tal risks for which we have limited post-incursion control and management options (e.g., Ribera and Boudou-resque 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 Chlo-rophyta, 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 sig-nificant efforts, our understanding of the processes determining invasion success remains limited, specifi-cally 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 acci-dental 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 mecha-nisms; 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 in-depth 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).
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).
Transport mechanisms
Humans have undoubtedly transported species inten-tionally and accidentally for several thousand years (di Castri 1989). However, these movements are likely to have been spatially restricted and of relatively low fre-quency. 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 mech-anisms in the marine environment are largely associated with commerce and exploration. These include: wooden-hulled 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 aqua-rium 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 communi-ties of tens to hundreds of species being transported between disparate bioregions. These vectors of transport typically result in the unidirectional movements of spe-cies 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 trans-port 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 mech-anisms has a unique set of constraints that act as selec-tion 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 com-mon 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
Table 1 Specific constraints associated with identified transport mechanisms.
Transport mechanism Uptake Transport
Exposure Planktonic Association Shear Desiccation Darkness Crushing Exposure to
(settlement/ phase with target stress and changing
uptake) period species or physical environment
habitat stress
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 Deli-vopoulos 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 mech-anism 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 ship-worms (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 gal-leries. In one of the few modern studies of wooden hull fouling, Carlton and Hodder (1995) observed the inad-vertent 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, subsequent-ly, 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 mech-anism of transport for a variety of species including
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 hard-selection 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 angi-osperms 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 asso-ciated with hull fouling are small or have crustose or fil-amentous 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 voy-ages (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 commer-cial 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 spe-cies present in these habitats were available to this trans-port mechanism. Little is known about which species were transported by this vector. However, the cosmo-politan nature of many beach and shore fauna may be attributable to the large quantities of shoreline transport-ed 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 spe-cies that use vegetative fragments to establish. Dis-charge 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 Salz-mann ex A. Braun (Luther 1979, Wallentinus 2002) but see also the discussion of Lewis (1999) for Gymnogron-gus 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 trans-port 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 organ-isms, as well as demersal species (e.g., Carlton and Gel-ler 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 Platyhel-minthesx while Carlton and Geller (1993) reported 367 taxa, including taxa from three macrophytic phyla (includ-ing the angiosperms) – represent(includ-ing all of the major tro-phic 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 Boudour-esque 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 stag-es, 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 sea-chests can have an established and reproductive resi-dent benthic fauna (Coutts et al. 2003). While this mechanism is active for microalgal species, no macro-algal 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 intro-ductions (Elton 1958, Neushul et al. 1992, Wallentinus 2002, Ribera Siguan 2003). Numerous unequivocal examples exist: the introduction of Pacific oysters wCras-sostrea gigas (Thunberg 1793)x into the Northeast Pacific Ocean resulted in the establishment of Sargassum muti-cum (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, Sargas-sum muticum, Undaria pinnatifida, Porphyra yezoensis Ueda, and Grateloupia.
Species which are intentionally released for stocking or aquaculture purposes are pre-selected for likely estab-lishment in the new environment and are typically released at high inoculation densities, often with multiple releases through time to ensure establishment (see Pick-ering 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 com-mercial incentive to keep the target species alive during transport. Uptake of associated species would be predi-cated 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 toler-ances 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 com-mercial industry with well-prescribed quarantine practic-es. 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) includ-ing specific quarantine periods and containment require-ments. To be exempted, non-approved species must undergo an individual risk assessment process. However, these measures have been developed to prevent the
inadvertent introduction of fish pathogens and parasites and, consequently, do not deal effectively with macroal-gae, especially if these are imported as target species. Despite quarantine and IHS efforts, a number of high pro-file macroalgal introductions, many through accidental escapes, have been associated with this transport mech-anism. 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, conse-quently, numerous non-approved species bypass man-datory 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 import-ed 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 sig-nificant 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 ostensi-bly compassionate gesture will, at its best, see the spe-cies 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 trans-port 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, lob-sters, and oysters). Packing material for bait is the most likely to be discarded in the marine environment (Wallen-tinus 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. Typ-ically, packing material is not covered under IHS and therefore slips through biosecurity or quarantine detec-tion. Numerous packing material-associated invasions have been recognised in the Mediterranean Sea (Wallen-tinus 1999, 2002, Ribera Siguan 2002), the NE Pacific Ocean (Zostera japonica Aschers. et Graebner: Harrison and Bigley 1982, Ascophyllum nodosum (L.) Le Jolie:
Mil-ler et al. 2004), and the NW Atlantic Ocean wCodium frag-ile (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 Indi-an OceIndi-ans; 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 pack-ing material is likely to survive the transport process (Table 1).
Association with marine and maritime equipment Observations and anecdotal evidence indicate that intro-duced 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, Schaf-felke 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 fash-ion, especially from areas with high abundance of intro-duced macroalgae and during periods of high reproductive output. Rapid settlement of Undaria pinna-tifida zoospores was shown on glass slides suspended from ropes in an infested Tasmanian bay (B. Schaffelke, unpublished data). Constraints for this pathway are des-iccation and, possibly, freshwater stress, e.g., by wash-ing gear before re-immersion in salt water.
Scientific research
Several intentional introductions of macroalgae for sci-entific 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, Kiri-bati with a commercially cultivated species, Kappaphy-cus 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). Sev-eral transport mechanisms are either under current man-agement 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 Man-agement and Control of Ships’ Ballast Water and Sedi-ments (BWM) Convention, IMO 2004, see Doelle et al. 2007x. These efforts are largely focused on a few trans-port mechanisms, with the establishment of agreed quar-antine standards (through IHS) and treatment options. Risk mitigation addresses either the uptake of propa-gules at the source location (e.g., prevention of hull foul-ing by anti-foulfoul-ing 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 Trans-fers of Marine Organisms (ICES 2003), which advises on how to evaluate the risk of introducing a marine organ-ism, including determining the likelihood and conse-quences 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 crite-ria. It has been applied successfully in a variety of cir-cumstances 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 avail-ability 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 vari-ety 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 stage-gate 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 trans-port 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 sig-nificant numbers of associated species, despite signifi-cant efforts to ‘‘clean’’ the valves prior to transport (Wallentinus 1999).
Lastly, consideration should be given to the aquacul-ture (or mariculaquacul-ture) 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 dis-eases, pathogens or other associated species, such as macroalgae. Caution must be taken to ensure that quar-antine facilities are sited where the impact of natural dis-asters that possibly breach containment (e.g., cyclones, earthquakes, floods, tsunamis) is reduced (E. Gonzalez personal communication). Similarly, adequate filter sys-tems 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 spe-cies, management of species translocations associated with commercial shipping (including slow-moving oil plat-forms, ocean going barges and tugs) and recreational vessels is more problematic. At present, international opinion has identified ballast water as the transport path-way with the highest chance for successful management in the immediate future. This mechanism has been per-ceived 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 concert-ed 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 macroal-gae (e.g., Rainer 1995, Ribera and Boudouresque 1995, Cranfield et al. 1998, Hewitt et al. 1999, 2004, Wallenti-nus 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 orga-notin (e.g., tributyl tin; TBT) as active ingredients to pre-vent 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).