Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

David W. Langor · Jon Sweeney Editors

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

Reprinted from Biological Invasions, volume 11, number 1, January 2009

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Editors David W. Langor Natural Resources Canada Canadian Forest Service Northern Forestry Centre Edmonton, AB Canada T6H 3S5 [email protected]

Jon Sweeney Natural Resources Canada Canadian Forest Service Atlantic Forestry Centre Fredericton, NB Canada E3B 5P7 [email protected]

ISBN: 978-1-4020-9679-2

e-ISBN: 978-1-4020-9680-8

DOI: 10.1007/978-1-4020-9680-8 Library of Congress Control Number: 2008943244 © Springer Science+Business Media B.V. 2009 No part of this work may be reproduced, stored in a retrieval system, or transmitted in any form or by any means, electronic, mechanical, photocopying, microfilming, recording or otherwise, without written permission from the Publisher, with the exception of any material supplied specifically for the purpose of being entered and executed on a computer system, for exclusive use by the purchaser of the work. Cover illustration: Gypsy moth larva, Lymantria dispar (L.). Photo by Steve Marshall, Guelph University, Canada Printed on acid-free paper springer.com

Contents

Foreword Ecological impacts of non-native invertebrates and fungi on terrestrial ecosystems D.W. Langor & J. Sweeney . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1–3

Papers Diversity of non-native terrestrial arthropods on woody plants in Canada D.W. Langor, L.J. DeHaas & R.G. Foottit . . . . . . . . . . . . . . . . . . . . . . . .

5–19

Ecological effects of invasive alien insects M. Kenis, M.-A. Auger-Rozenberg, A. Roques, L. Timms, C. Péré, M.J.W. Cock, J. Settele, S. Augustin & C. Lopez-Vaamonde . . . . . . . . . . . . . . . .

21–45

Beyond Pandora’s Box: quantitatively evaluating non-target effects of parasitoids in classical biological control D. Parry . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

47–58

Distribution and impacts of invasive earthworms in Canadian forest ecosystems J.A. Addison . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

59–79

Ecological impacts of non-indigenous invasive fungi as forest pathogens J.A. Loo . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

81–96

The impact of invasive fungi on agricultural ecosystems in the United States A.Y. Rossman . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

97–107

Responses of Chalcidoidea (Hymenoptera) parasitoids to invasion of the cabbage seedpod weevil (Coleoptera: Curculionidae) in western Canada L.M. Dosdall, G.A.P. Gibson, O.O. Olfert & P.G. Mason . . . . . . . . . . . . 109–125 Non-native insects in agriculture: strategies to manage the economic and environmental impact of wheat midge, Sitodiplosis mosellana, in Saskatchewan O. Olfert, R.H. Elliott & S. Hartley . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 127–133 Climate suitability and management of the gypsy moth invasion into Canada J. Régnière, V. Nealis & K. Porter . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 135–148 We can eliminate invasions or live with them. Successful management projects D. Simberloff . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 149–157

Ecological impacts of non-native invertebrates and fungi on terrestrial ecosystems David W. Langor Æ Jon Sweeney

Originally published in the Journal Biological Invasions, Volume 11, No 1, 1–3. DOI: 10.1007/s10530-008-9326-y Ó Her Majesty the Queen in Right of Canada 2008

Since the arrival of Europeans about 500 years ago, an estimated 50,000 non-native species have been introduced to North America (including Hawaii) (Pimentel et al. 2000). This averages two species every week; however, the rate of entry is generally thought to have been much higher in the last century as the amount of international trade rapidly increased. Non-native or exotic species figure prominently in our lives. Many of the species that we consume are not native. In urban environments we are inundated by exotic species, especially plants; however, a large proportion of exotic invertebrates are also anthropogenic (Langor, unpublished data). In forestry and agriculture many serious insect and fungal pests are non-native. Rivers, lakes and ponds are increasingly becoming breeding grounds for a wide variety of aquatic invaders. Marine environments, especially inland waters, have been colonized by a large number of exotic species. Since Charles Elton’s seminal book, The Invasion Ecology of Animals and Plants, published in 1958, scientific interest in invasions has rapidly increased.

D. W. Langor (&) Natural Resources Canada, Canadian Forest Service, 5320-122 Street, Edmonton, AB, Canada T6H 3S5 e-mail: [email protected] J. Sweeney Natural Resources Canada, Canadian Forest Service, P.O. Box 4000, 1350 Regent Street, Fredericton, NB, Canada E3B 5P7 e-mail: [email protected]

However, in Canada it is only over the last two decades that invasive alien species (IAS) received significant public attention. This attention came about largely because of the incursion of a few highly visible species that had significant impacts on the economy, security and human health. Species such as zebra mussel, Asian longhorned beetle, sudden oak death, Russian wheat aphid, SARS, West Nile Virus caught the media and public attention. The enormous costs of IAS in the USA, estimated at $137 billion per year caught the attention of the federal administration and resulted in dedicated funding for IAS research (Pimentel et al. 2000). Soon after, IAS also gained federal attention in Canada, again largely because of the economic and health impacts of these species, and eventually resulted in An Invasive Alien Species Strategy for Canada (Environment Canada 2004). Although economic, security and health impacts are deemed newsworthy, the ecological impacts of IAS has generally received much less public and scientific attention, despite the fact that IAS ranks second to habitat destruction as a cause of species loss. Although the ecological impacts of marine, freshwater and plant invaders have received some attention, there is much less available information about the ecological impacts of hyper-diverse groups such as fungi and terrestrial invertebrates. To bring attention to this paucity of information and to encourage a review of the current state of knowledge, the Canadian Forest Service, Biological Survey of Canada and Canadian Food Inspection Agency convened a symposium, Ecological

D.W. Langor & J. Sweeney (eds.), Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems. DOI: 10.1007/978-1-4020-9680-8_1

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impacts of non-native invertebrates and fungi on terrestrial ecosystems, held on November 18, 2006 during the Joint Annual Meeting of the Entomological Society of Canada and Entomological Society of Quebec in Montreal, Quebec. This symposium brought together 13 scientists from the USA and Canada to review the state of knowledge in this field of work. Most of these papers are contained herein. As well, we have added two papers that were not part of the original symposium. This set of 10 papers represents a crosssection of work on ecological impacts of IAS. Most are review/synthesis papers, but we have balanced this with three case-studies. Although there is a strong focus on Canadian work, there is also significant focus on work in the northern USA and Europe. David Langor and colleagues provide an overview of the diversity of non-native arthropods feeding on woody plants in Canada. Almost every genus of woody plants hosts non-native species, and in many cases exotic species often exist at high population levels. Some plant genera host many exotic species and the potential for significant impacts on native phytophagous insects and mites is high, although this has rarely been documented. Langor argues strongly for better investment in taxonomy and improved information management/synthesis to allow ready identification of invasion by exotic species and to learn from the wealth of knowledge about IAS that is largely buried in gray reports and restricted databases. Marc Kenis and colleagues review and synthesize the ecological impacts of invasive alien insects based on 401 primary research publications, mainly from North America, concerning 72 insect species. They show that most research focused on effects of IAS on native biodiversity at population or community level. Genetic effects and, to a lesser extent, effects on ecosystem services and processes were rarely explored. Ecological impacts may occur through simple trophic interactions such as herbivory, predation or parasitism, but also through more complex mechanisms such as competition for resources, disease transmission, apparent competition, or pollination disruption, among others. Dylan Parry examines the potential for non-target effects among insect parasitoids introduced for biological control in North America. He highlights three techniques, quantitative food webs, life table analysis and experimental populations, to assess non-target effects in different systems. He also explores three methods to ascertain the strength of competitive

D.W. Langor & J. Sweeney (eds.)

interactions between native and introduced parasitoids, a potential non-target effect that has received little attention in the literature. Switching our focus to another group of invertebrates, Jan Addison points out that although exotic earthworms are generally considered to be beneficial in agricultural soils, their effects can be less benign in forested ecosystems, where they can significantly alter the forest floor, affecting the distribution of carbon, nitrogen and other chemicals, roots, microbes and other elements of the soil fauna, and even understory vegetation. She summarizes the current distribution of exotic and native earthworm species in Canadian forests and draws on the results of studies of invasion patterns and environmental impacts in northern forests in North America and Europe to discuss potential outcomes for forests in Canada. Two papers examine the ecological impacts of non-native fungi. First, Judy Loo focuses on impacts in forested ecosystems, and reviews several treefungus systems in northern North America. Impacts range from functional elimination of abundant tree species to loss of populations or all trees in the larger size classes. She shows that impacts are most significant when highly successful invading pathogens attack foundation species, setting in motion a long-lasting cascade of effects on the host and associated species. Such impacts have generally not been well documented at the ecosystem level. Amy Rossman summarizes the impacts of some exotic fungi on agricultural ecosystems. Although agricultural systems are unnatural, usually feature nonnative plant species and impacts by pests are largely economical, the persistence of non-native fungal species in these systems can result in spill-over to native ecosystems, resulting in ecological impacts. Rossman also asserts that taxonomic challenges in this group inhibit early detection and identification of non-native species. In addition to review papers, three case studies of non-native insects are included in these proceedings, two from agricultural systems and one forest pest. Lloyd Dosdall and colleagues report on response of native parasitoid and hyperparasitoid assemblages to invasion by the cabbage seedpod weevil. Owen Olfert and colleagues describe the impacts of the wheat midge and a successful management program to minimize economic and environmental impacts. Using the gypsy moth as a model, Jacques Re´gnie`re and colleagues

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

demonstrate how use of historical data and modeling based on climatic suitability can be used to predict future distribution of this species, providing forest managers with an opportunity to apply early intervention to reduce economic and ecological impacts. To wrap up the proceedings, Dan Simberloff encourages us to take heart from the many success stories concerning management of IAS. Using many examples (often unpublished and not widely known) from around the world, he shows that it is possible to eradicate IAS or at least to apply management to reduce them to low population levels. We need not be eternally pessimistic about managing IAS but take inspiration from the many successes and push forward with optimism to add to these successes. This symposium would not have been possible without support from the Canadian Forest Service and Biological Survey of Canada (Terrestrial Arthropods).

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We also thank Dr. Hugh Danks for his encouragement and practical suggestions that helped us through the planning stages. Finally, we wish to thank all of the participants in this symposium for contributing to the success of the event and ensuing proceedings. We hope that this set of papers will help inspire others to pursue work on ecological impacts of non-native species.

References Environment Canada (2004) An invasive alien species strategy for Canada. www.ec.gc.ca/eee-ias/Default.asp?lang=En&n =98DB3ACF-1. Accessed 12 March 2008 Pimentel D, Lach L, Zuniga R, Morrison D (2000) Environmental and economic costs of nonindigenous species in the United States. Bioscience 50:53–65

Diversity of non-native terrestrial arthropods on woody plants in Canada David W. Langor Æ Laura J. DeHaas Æ Robert G. Foottit

Originally published in the Journal Biological Invasions, Volume 11, No 1, 5–19. DOI: 10.1007/s10530-008-9327-x Ó Her Majesty the Queen in Right of Canada 2008

Abstract A list of non-native phytophagous insects and mites on woody plants (trees, shrubs, vines) in Canada was compiled using information from literature and input from taxonomists. The 419 recorded species include Hemiptera (53% of species), Lepidoptera (22%), Coleoptera (13%) and Hymenoptera (9%). Almost all species originate from the Palearctic, especially Europe, reflecting historical trade patterns. About 41% of species were directly introduced to Canada from countries of origin, and the remainder spread from the United States of America (USA) after initial establishment there. Major ports on the east and west coasts, on Lake Erie and Lake Ontario are the main points of entry for exotic species directly introduced, and southern British Columbia (BC), Ontario (ON) and Quebec (QC) are the major points of entry for species spreading from the USA. Consequently, BC, ON, QC and Nova Scotia have the highest diversity of non-native species, and the prairie provinces and northern territories have the

lowest. The extent of the distribution of individual species is related to length of time in Canada, number of introductions and dispersal abilities. Almost all native woody plant genera in Canada have been invaded by exotic phytophages. The large majority of phytophages occur on angiosperms. Woody plant genera with the largest distribution, highest species diversity and highest local abundances tend to host the greatest number of non-native species, including Picea, Pinus, Malus, Prunus, Salix, Betula, Quercus, Pyrus and Populus. The arrival rate of species in Canada increased from the late nineteenth century until about 1960, and declined rapidly thereafter. Quarantine legislation enacted in the USA in 1912 and in Canada in 1976 seems to have reduced the rate of insect invasion.

Electronic supplementary material The online version of this article (doi:10.1007/978-1-4020-9680-8_2) contains supplementary material, which is available to authorized users.

Introduction

D. W. Langor (&)
L. J. DeHaas Natural Resources Canada, Canadian Forest Service, 5320-122 Street, Edmonton, AB, Canada T6H 3S5 e-mail: [email protected] R. G. Foottit Agriculture and Agri-Foods, Ottawa, ON, Canada

Keywords Exotic species
Invasive arthropods
Insects
Mites
Woody plants
Invasion pathways
Exotic species database

Since the discovery and colonization of North America by Europeans over 500 years ago, there has been an increasing invasion of the continent by plants, animals and fungi from other regions of the globe (Sailer 1983; Stuckey and Barkley 1993; Hendrickson 2002). However, it is only relatively recently that the scale, extent and impacts of invasion

D.W. Langor & J. Sweeney (eds.), Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems. DOI: 10.1007/978-1-4020-9680-8_2

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by alien species have been given full appreciation. Despite the increasing concern about the adverse effects of invasive alien species (Krcmar-Nozic et al. 2000; Pimentel et al. 2000; Colautti et al. 2006), there are large information gaps concerning this significant component of North American biota. For example, the species introduced by modern humans into North America have not been adequately catalogued. The best known groups include the vertebrates and vascular plants. In Canada, about 20–27% of the ca. 5,800 species of vascular plants are alien, and the distribution of most species is well known (Haber 2002). However, there is a relative dearth of information about species-rich, taxonomically challenging and poorly surveyed groups such as terrestrial arthropods, fungi (Yarwood 1983), nematodes and micro-organisms, and among these groups, the arthropods are certainly best known. Of the estimated 50,000 non-native species that have been introduced to the United States (including Hawaii), only about 4,500 (9%) are arthropods (Pimentel et al. 2000). Of the arthropods, only about 2,000 are estimated to occur in the continental United States (Pimentel et al. 2000). This is likely a large underestimate. Turnbull (1979) stated that he had produced a list of 301 alien insect species in Canada, but this list was never published. Many species lists and taxonomic monographs produced in Canada over the last 25–30 years have provided indications of which species were not native to Canada. For example, Bousquet (1982) tallied 469 non-native Coleoptera species in a total Canadian fauna of 7,447 species (i.e., 6.3% of species introduced), and Maw et al. (2000) tallied 337 non-native species of Hemiptera in a total fauna of ca. 3,900 species (i.e., 8.6% of species introduced). Recent efforts have tallied about 1,900 non-native terrestrial arthropods in Canada (Langor, unpublished data), and this number will likely exceed 2,000 by the time ongoing assessments are completed. Thus, non-native species account for approximately 3.2% of the insect fauna and 4.7% of the spider fauna of Canada (Biological Survey of Canada 2007a; Langor, unpublished data). Other groups such as mites are too poorly known to make a reasonable estimate of the percent of the fauna represented by non-native species. Among the non-native terrestrial arthropods in Canada, a large proportion feed on woody plants. Some of these are among the most notorious invasive

D.W. Langor & J. Sweeney (eds.)

alien species in Canada, as indicated by the amount of recent media coverage, and include the gypsy moth, Lymantria dispar (L.), Asian longhorned beetle, Anoplophora glabripennis (Motschulsky), brown spruce longhorned beetle, Tetropium fuscum (Fabricius) and the emerald ash borer, Agrilus planipennis Fairmaire. Hendrickson (2002) listed 180 species of alien insects established on woody plants in Canada, and this list was based largely on the data provided by Mattson et al. (1994). However, many species were overlooked by Hendrickson (2002) and several species have been detected in Canada since that time. To provide a foundation for future syntheses of data concerning distribution, spread and impact of alien species in Canada, herein we provide a synopsis of known non-native terrestrial arthropods feeding on woody plants in Canada.

Methods An initial list of alien species feeding on woody plants (trees, shrubs, vines) in Canada was developed by exhaustive literature searches (covering literature published up to December 2007), and especially by perusal of species lists, catalogues and taxonomic treatments of the Canadian fauna. This list was then vetted widely among arthropod taxonomists, especially those in Canada and the USA, and was modified based on feedback from experts. The species list on which the analyses are based is included as Appendix A—see Electronic supplementary material. We included only species that feed on living plants and excluded species associated exclusively with dead wood. For each species we sought out the following data through consultation of literature and taxonomic experts: common names; synonyms; current distribution in Canada, and current North American and global distribution; region of origin (Africa, Palearctic-Europe, Palearctic-Asia, Palearctic-Eurasia, South America); location and date of first report of the species in North America and in Canada; pathway of introduction (directly to Canada, indirectly via the USA, or unknown); known and preferred hosts in Canada, in the USA and in its native range; and preliminary notes on pest status, control measures, natural enemies, and life history. Alien arthropod species were classified as ‘directly introduced to Canada’ if reported in the literature as

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

such, reported in Canada before being reported in the USA (i.e., dates and locations of first reports are known for both countries), or reported as having been directly introduced into Canada as well as spreading from the USA. Arthropods were classified as ‘indirectly introduced’ if reported as spreading into Canada from the USA or reported as introduced into the USA (location and date of first report known) and also found in Canada, but the method (and usually date) of introduction into Canada is unknown. Many of these species probably spread into Canada from the USA. The current provincial and territorial distribution of non-native species was determined based on literature records and by examination of specimens in the six regional collections of the Canadian Forest Service and the Canadian National Collection.

Results Diversity of non-native species A total of 419 non-native species (416 insects and three mites) are recorded from woody plants in Canada (Appendix A—see Electronic supplementary material), consisting mainly of Hemiptera (53% of species), Lepidoptera (22%), Coleoptera (13%) and Hymenoptera (9%). Other insect orders with representatives are Diptera and Thysanoptera. Seven of those species present in Canada are not yet recorded from woody plants there; however, they commonly infest woody plants in their native range and are highly likely to occur on woody plants in Canada. These species are included in analyses. As well, 44 non-native species are recorded from woody plants in adjacent parts of the United States but have not yet been recorded from Canada, although some of these likely have already spread into Canada without our knowledge. These species were not considered herein. Of the 223 species of introduced Hemiptera, most species (65%) are in the suborder Sternorryhncha, dominated by aphids (Aphidoidea) and scales (Coccoidea), but also including Psyllidae and Phylloxeroidea (Fig. 1a). Clypeorrhyncha represent 21% of species, mainly leafhoppers (Cicadellidae), but also spittlebugs (Cercopidae). The 55 species of non-native Coleoptera on woody plants are

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dominated by weevils (69%), especially bark beetles in the subfamily Scolytinae (27% of total beetle species) (Fig. 1b). All of the 38 species of Hymenoptera feeding on woody plants are Symphyta, dominated by sawflies in the Tenthredinidae (79% of taxa) and Diprionidae (11%) (Fig. 1c). Among the 91 species of Lepidoptera on woody plants, the vast majority of species are micro-Lepidoptera, especially Tortricoidea (34%) and Gelechoidea (19%), and the dominant macro-Lepidoptera are Noctuoidea (12%) (Fig. 1d). The non-native species are divided into three feeding guilds. The majority (54%) of the non-native fauna are sap feeders, reflecting the dominance of Hemiptera; however, this guild also contains two species of thrips (Thysanoptera) and three species of mites (Acarina). Bark-and wood-feeders (including external feeders on roots and gall makers on twigs) represent 12% of the Canadian fauna, and 84% of the species in this guild are Coleoptera, dominated by weevils and bark beetles (Curculionidae) and longhorned beetles (Cerambycidae). However, this guild also includes two species of Hymenoptera and six species of Lepidoptera. Foliovores (including gallmakers on leaves and bud-feeders) represent 34% of the overall fauna, including Lepidoptera, Hymenoptera, Coleoptera and Diptera. Sources and geographic patterns of establishment The vast majority of non-native species on woody plants originated from the Palearctic, especially from Europe (82% of all species). Of the 380 species on woody plants whose origins are known only two originate from a region other than the Palearctic: Ischnaspis longirostris (Signoret) (Hemiptera: Diaspididae) from Africa, and Pinnaspis aspidistrae Signoret (Hemiptera: Diaspididae) from the Oriental Region. There are two pathways of entry for non-native species into Canada: direct introduction from the country of origin, and indirect introduction via spread from the USA after being directly introduced there. Of the 372 species of non-native species whose pathway of entry is known or suspected. about 41% (152 species) were directly introduced into Canada. It is thought that the largest proportion of non-native species (220 species) were first introduced to the USA and subsequently spread to Canada. For many

8 Fig. 1 Taxonomic composition of major orders of non-native insects on woody plants in Canada. (a) Hemiptera, (b) Coleoptera, (c) Hymenoptera, (d) Lepidoptera

D.W. Langor & J. Sweeney (eds.)

(a) Hemiptera (223 spp.)

(b) Coleoptera (55 spp.)

Psyllidae Aleyrodidae Coccoidea

Others

Heteroptera

Buprestidae Cerambycidae

Phylloxeroidea

Chrysomelidae

Cicadellidae

Scolytinae

Cercopidae Aphidoidea Curculionidae (excl. Scolytinae)

(d) Lepidoptera (91 spp.)

(c) Hymenoptera (38 spp.) Argidae Siricidae Pamphiliidae

Diprionidae Others

Gelechioidea

Cimbicidae Yponomeutoidea Geometroidea

Gracillarioidea

Tortricoidea

Noctuoidea

Tenthredinidae

species it is very difficult to retrace pathways of entry, especially for those introduced a long time ago and that are now present in both countries. Analysis of the geographic points of entry for nonnative species provides an indication of which parts of the country are most susceptible to invasion. Overall, most species were first detected in Ontario (ON, 28%), British Columbia (BC, 25%), the maritime provinces (12%) and Quebec (QC, 9%) (Fig. 2). Very few species were first detected in the prairie provinces and north (1.2%) and Newfoundland and Labrador (NL, 2.9%). Of the 152 species believed to have been directly introduced to Canada, 40% were first recorded in BC, 21% in the maritimes, 19% in ON, 12% in QC, 6% in NL and 2% in the prairie provinces and north (Fig. 2). It is clear from examination of specific locality records that the areas where most direct introductions occur are: the lower Fraser Valley of BC (at least 30 species introduced); the southern portion of Vancouver Island, BC (at least 11 spp.); southern ON, especially the western area of Lake Ontario, north shore of Lake Erie, and Niagara Peninsula (at least 16 spp.); St. John’s, NL (at least 7 spp.); and Halifax, Nova Scotia (NS, at least 6 spp.). A substantially different pattern of introduction is discernable for species that were

likely indirectly introduced to Canada via spread from the USA; the largest proportion were first detected in ON, especially southern ON (33%), followed by BC, especially the lower Fraser Valley (16%), southern QC (7%), the maritimes (7%), the prairie provinces and north (1%) and NL (1%) (Fig. 2). Distribution of non-native species in Canada BC has the highest diversity of non-native insects and mites on woody plants in Canada (270 species), followed by ON (263), QC (217) and NS (181) (Fig. 3). The Canadian north (Yukon [YK], Northwest Territories [NT], Nunavut [NU] and Labrador [LB]) have been colonized by very few species. The three prairie provinces (Alberta [AB], Saskatchewan [SK] and Manitoba [MB]) have the lowest non-native diversity of all the provinces, even lower than that of the tiny province of Prince Edward Island (PE). A plot of the range (number of occupied provinces and territories) by the number of non-native species best fits a log normal distribution (Fig. 4). About 29% of species have a limited distribution (one jurisdiction), and only 57 species are found in more than half of the political jurisdictions of Canada. The

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems Fig. 2 Points of first entry for non-native insects and mites on woody plants in Canada. Species were either directly introduced from the country of origin, or indirectly introduced via spread from the USA after being directly introduced there. Province abbreviations: BC, British Columbia; ON, Ontario; QC, Quebec; NF + LB, Newfoundland and Labrador

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140 All species (419 spp.) Directly introduced (152 spp.) Indirectly introduced (220 spp.)

No. of species

120 100 80 60 40 20

U nk no w n

LB + F N

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250 200 150 100 20 0 YK

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Provinces and territories Fig. 3 Number of non-native insects and mites on woody plants in each province and territory of Canada. Province and Territory abbreviations: YK, Yukon; NT, Northwest Territories; NU, Nunavut; BC, British Columbia; AB, Alberta; SK,

Saskatchewan; MB, Manitoba; ON, Ontario; QC, Quebec; NB, New Brunswick; NS, Nova Scotia; PE, Prince Edward Island; NF, Newfoundland (insular); LB, Labrador

regression between the length of time that a species was resident in Canada and its geographic range (no. of provinces and territories in which it is present) was significant (r2 = 0.246; P \ 0.0001) indicating that geographic range is positively influenced by length of time that a species has been in Canada (Fig. 5). The overall number of non-native species in each major insect order, as well as the proportional representation of each order in provincial and territorial faunas, varies across the country (Fig. 6). Of the 223 species of non-native Hemiptera on woody plants in Canada, 70% are found in BC and 58% in Ontario. In all provinces, Hemiptera represented the highest number of non-native species, and proportional representation of the non-native fauna ranged from 39% in NL to 57% in BC. Hemiptera were

uncommon in the north. Of the 55 species of nonnative Coleoptera, QC and ON each had 76% of the species followed by NS (60%) and BC (47%). Coleoptera comprised a higher proportion of the faunas in ON, QC and the maritime provinces (16–21%) than elsewhere in Canada. Although there are only 38 species of non-native Hymenoptera feeding on woody plants in Canada, 89% occur in ON and 82% in QC. Proportionate representation of non-native fauna by Hymenoptera is highest in the north (37–44%) and in SK (21%), and lowest in the maritimes (8–9%). Of the 91 species of Lepidoptera, 72% are found in BC and 60% in ON. As a proportion of provincial faunas, Lepidoptera exhibit much less variation across the country (16–30%) than other major orders.

10 140 120

No.of species

Fig. 4 Number of provinces and territories inhabited by non-native insect and mite species on woody plants in Canada. Records for Nunavut and the Northwest Territories are combined

D.W. Langor & J. Sweeney (eds.)

100 80 60 40 20 0 1

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Fig. 5 Mean (SE) number of years of residency for non-native species on woody plants in relation to geographic range in Canada. Records for Nunavut and the Northwest Territories are combined

Average no. years in Canada

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10

No. of provinces and territories inhabited

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Coleoptera (55 spp.) Hemiptera (223 spp.) Hymenoptera (38 spp.) Lepidoptera (91 spp.)

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Fig. 6 Distribution of four major orders of non-native insects on woody plants in Canada. For province and territory abbreviations, see caption for Fig. 3

120 100 80 60 40 20 0 YK

NT NU

The number of species in each feeding guild also varied among provinces and territories (Fig. 7). Of the 50 species of bark- and wood-feeders, the greatest number of species were in the faunas of ON (38), QC (34), NS (27) and BC (26). Only one species, the strawberry root weevil, Otiorhynchus ovatus (L.), was found in the north. The prairie provinces had 4–8 species in their faunas. The proportional representation of provincial and territorial faunas by bark- and

BC

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LB

Provinces and territories

wood-feeders varied relatively little, ranging from 5% in the Northwest Territories (NT) and Nunavut (NU) to 16% in QC. Of the 227 species of sap feeders (almost all Hemiptera), BC has the most species (141), followed by ON (132). Only seven species are found in the northern territories, and, among the provinces, the sap-feeding faunas of the prairie provinces are most depauperate. In terms of proportionate representation within provincial and territorial

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems 180 160

Bark- and wood-feeders (n=50 spp.) Sap feeders (n=227 spp.) Foliovores (n=142 spp.)

140

No. of species

Fig. 7 Distribution of nonnative insects and mites on woody plants in Canada by type of feeding damage. For province and territory abbreviations, see caption for Fig. 3

11

120 100 80 60 40 20 0 YK

NT NU

faunas, sap feeders are most dominant in BC (59% of fauna); however, proportional representation is also high in other provinces, ranging from 38% in NL to 57% in MB. There are no non-native sap feeders recorded from LB. Of the 142 species of foliovores, most are found in ON (93), BC (87) and QC (85). Foliovores are the dominant feeding guild in the northern territories and LB, representing 57–80% of the non-native faunas there. Among the provinces, the foliovore faunas of the prairie provinces are most depauperate, and proportional representation ranges from 32% in NS to 49% in NL. Host range The majority of non-native species (326 species) are found only on angiosperms, 59 species are found exclusively on gymnosperms and 34 species feed on both gymnosperms and angiosperms (Fig. 8). In terms of breadth of diet, 195 are stenophagous (feed on a single woody plant genus), 85 are oligophagous (on three or fewer genera) and 139 are polyphagous 450 400

No. of species

Fig. 8 Host range of nonnative insects and mites on woody angiosperms and gymnosperms in Canada

BC

AB

SK

MB

ON

QC

NB

NS

PE

NF

LB

Provinces and territories

(on more than three genera) (Fig. 8). Stenophagous and oligophagous species are proportionately more abundant on gymnosperms (58 and 27% of species, respectively) than on angiosperms (49 and 20%). A high proportion of angiosperm-feeders are polyphagous (31%), but polyphagy is present in only 15% of gymnosperm-feeding species. Of the 93 species that feed on gymnosperms, the greatest number of species are found on Pinus (54 spp.) and Picea (47 spp.) (Appendix B—see Electronic supplementary material). Of the 360 species that feed on angiosperms, the greatest number are found on genera of fruit trees, Prunus (85 spp.), Malus (79 spp.) and Pyrus (63 spp.). Other angiosperms with rich faunas of non-native species are Salix (73 spp.), Betula (72 spp.), Quercus (62 spp.) and Populus (59 spp.). Temporal pattern of colonization Of the 419 species of non-native insects and mites feeding on woody plants in Canada, approximate

Polyphagous (139 spp.) Oligophagous (85 spp.) Stenophagous (195 spp.)

350 300 250 200 150 100 50 0 All

Gymnosperms

Hosts

Angiosperms

Gymosperms & Angiosperms

Fig. 9 Year of first introduction into North America and into Canada for non-native insects and mites on woody plants

D.W. Langor & J. Sweeney (eds.)

No. of species

12 100 90 80 70 60 50 40 30 20 10 0

North America Canada

9

9

69

1 0-

0

16

0

17

9

74

1 0-

5

17

9

79

1 0-

0

18

2

1 0-

4

18

87

1 0-

6

18

9

9

85

83

1 0-

18

9

9

81

1 0-

8

18

9

89

1 0-

0

19

93

1 0-

2

19

95

1 0-

4

19

97

1 0-

6

19

No. of species

45 40

s re

p

0-

0 20

Year of introduction

Fig. 10 Year of introduction into Canada for directly and indirectly introduced species of insects and mites on woody plants

en

99

1 0-

8

19

t

9

9

9

9

91

1 0-

Direct to Canada Indirect to Canada

35 30 25 20 15 10 5 0

99

6 -1

00

16

00

17

49

7 -1

99

7 -1

50

17

00

18

19

8 -1

20

18

59

39

8 -1

8 -1

40

18

60

18

79

8 -1

80

18

99

8 -1

Year of entry

times of arrival in North America have been estimated for 389 species. Only 11 of those species are recorded from North America before 1840 (Fig. 9). The earliest record is the codling moth, Cydia pomonella (Lepidoptera: Tortricidae), which was recorded from apple trees in Ontario in 1635. After 1840, the rate of introduction of species into North America increased rapidly, peaking between 1900 and 1919. Thereafter, the rate of introduction declined steadily. A similar temporal pattern of colonization is evident for Canada, although the period of rapid colonization did not commence until later (1880–1899) and peaked later (1940–1959) than for North America as a whole (Fig. 9). The temporal pattern of introduction also differs for those species directly introduced to Canada from countries of origin compared to those species that were indirectly introduced to Canada via spread from the USA (Fig. 10). The number of species that

00

19

19

9 -1

20

19

39

9 -1

40

19

59

9 -1

60

19

79

9 -1

80

19

t

99

9 -1

en

s re

-p

00

20

indirectly reached Canada peaked between 1920 and 1939, and declined rapidly thereafter. The number of species that were directly introduced to Canada peaked in 1940–1979, followed by a rapid decline.

Discussion Diversity Clearly the number of established non-native insect and mite species feeding on woody plants in Canada and the United States is much higher than previously recorded. The most recent list of non-native insect species feeding on woody plants in Canada is that of Hendrickson (2002), who tallied 180 species, largely extracted from the list provided by Mattson et al. (1994) for North America. Thirteen species listed in

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

the current database were introduced into North America (including Canada) since the publication of Mattson et al. (1994), but only one of those was included by Hendrickson (2002). As well, about 40 species listed by Mattson et al. (1994) as present in the USA but not in Canada are, in fact, known from Canada. About 225 species of non-native species in our database were not recorded by Mattson et al. (1994) even though they were present in North America before 1990. About 150 of the new records are Heteroptera and the large majority of those records have recently come to light due to detailed faunistic analyses (Maw et al. 2000; Miller et al. 2002, 2005; Miller and Miller 2003; Foottit et al. 2006; Scudder and Foottit 2006). Finally several species listed by Mattson et al. (1994) and Hendrickson (2002) as phytophagous on woody plants in Canada were not included in our list as they feed on only seasoned wood (five species) or are native (three species). Combining current work with that of Mattson et al. 1994, there are at least 585 species of non-native insects feeding on woody plants in the USA and Canada. As well, we catalogued three species of phytophagous mites on woody plants, and these have not yet been tallied for the USA. Cataloguing the non-native species in Canada is perpetually a work-in-progress because: species invasion is a dynamic process with new species arriving on a regular basis; there is a dearth of knowledge about the composition of major components of the Canadian fauna (native and introduced species); and our knowledge of the native distribution of species is still evolving (e.g., some species thought to be introduced may be Holarctic). As knowledge of the Canadian and Nearctic fauna improves, there is little doubt that the list of non-native arthropod species on woody plants will grow. The overwhelming dominance of Hemiptera (53% of species) among the non-native species on woody plants is striking. Overall, Hemiptera is estimated to represent about 6% of the native insect fauna of Canada (Biological Survey of Canada 2007a), but approximately 23.5% of the total non-native insect and arachnid fauna of Canada (Langor, unpublished data) and 24.4% of the non-native insects and arachnid fauna of the 48 contiguous states of the USA (Sailer 1983). More notably, Hemiptera represents 45% of the non-native phytophagous insect species in Canada (Langor, unpublished). There are several possible reasons why the Hemiptera, and

13

especially the Sternorrhyncha (aphids and scales), have been the most successful phytophagous insect invaders. A high percent of aphid and scale species are parthenogenetic (Niemela¨ and Mattson 1996). Parthenogenetic capacity is a valuable trait for colonizing species and those inhabiting patchy or ephemeral habitats because it allows continual propagation even at low population levels. Parthenogenesis also promotes, more rapidly than sexual species, adaptation through random mutations (Ledig 1992). No wonder that about 40% of the non-native insects on woody plants in North America exhibit parthenogenesis, whereas only about 11% of the native species are parthenogenetic (Niemela¨ and Mattson 1996). The success of aphids and scales may also be related to their small size (especially their eggs) which may allow them to more easily escape detection than larger, more conspicuous insects. The success of cicadellid leafhoppers and mirid bugs cannot be explained by parthenogenesis; however, these groups generally have endophytic oviposition which may more easily allow eggs to escape detection during inspection. As well, many mirids are both phytophagous and predaceous (Wheeler and Henry 1992), which increases the niche/food options in new environments. The high diversity of Curculionidae among the non-native Coleoptera on woody plants may also be attributed to the fact that a large number (at least 100 species) of European weevils are parthenogenetic (Lanteri and Normark 1995; Niemela¨ and Mattson 1996). Also, Curculionidae is the most species-rich family of beetles, so it is not unexpected to see this family well represented in the non-native fauna. Some weevils, especially those, such as Otiorhynchus spp., that feed on plant roots and spend much time in the soil, were likely introduced in ships ballast (Lindroth 1957). Root inhabiting insects also easily escape detection because of their cryptic habits. It is notable that a high percentage of the non-native Coleoptera are bark- and wood-feeders, and many of these species were inadvertently brought into North America on solid wood packing material (Allen and Humble 2002; Haack 2006; Humble and Allen 2006). Among the non-native Hymenoptera (all of which are Symphyta, and can reproduce with only females), tenthredinid sawflies are the dominant family, and this likely just reflects the fact that this is the most diverse family of Symphyta in the northern

14

hemisphere. Although the diversity of non-native Hymenoptera on woody plants in Canada and the USA is much lower than that of other orders such as Hemiptera, Coleoptera and Lepidoptera, symphytans currently and historically represent many of the most notorious, widespread and destructive invasive species in Canadian forests. The remarkable success of sawflies in North American forests may be related to the fact that there are relatively few other sawfly competitors on Nearctic woody plants. In Europe there is a much higher diversity of sawflies per woody plant species (ca. five) than in North America (ca. one) (Neuvonen and Niemela¨ 1983; Haack and Mattson 1993). The combination of parthenogenetic abilities, low interspecific competition with native sawflies, the apparent abundance of unexploited or under-exploited niches and the general absence of specialist parasitoids common in their native range, have allowed populations of many sawfly species to quickly explode on reaching North American shores (Haack and Mattson 1993). Species such as the European pine sawfly, Neodiprion sertifer (Geoffroy), introduced pine sawfly, Diprion similis (Hartig), European spruce sawfly, Gilpinia hercyniae (Hartig), pine false webworm, Acantholyda erythrocephala (L.), larch sawfly, Pristiphora erichsonii (Hartig), European apple sawfly, Hoplocampa testudinea (Klug), and five leaf-mining sawflies on birch have been among the most damaging non-native insects on woody plants in Canada (Armstrong and Ives 1995). Fortunately, some of these have been brought under successful control by implementation of classical biological control programs (Mason and Huber 2002). Lepidoptera is the second largest order among the non-native fauna. The micro-Lepidoptera represents the large majority of introduced species in both Canada and the USA (Mattson et al. 1994), likely because they easily escape detection due to their small size and cryptic habits (many are miners of leaves, buds or twigs). Among the non-native macro-Lepidoptera, the Noctuidae and Geometridae are best represented in the Canadian fauna, likely simply reflecting that fact that these are the most diverse families of macro-moths and the odds of their introduction are simply proportional to their relative composition in the native fauna of the Palearctic (the source of all non-native Lepidoptera in Canada).

D.W. Langor & J. Sweeney (eds.)

The apparent poor representation of Diptera and Acarina among the non-native fauna of Canada is likely somewhat artificial. Of all the major groups of phytophagous arthropods, these two are the most poorly surveyed and studied in North America. There are likely many more non-native species in the Canadian fauna than we realize at this time. As many of these species are small and cryptic in habits (e.g., leaf, stem and twig miners), they are easily able to escape detection during inspection. Sources of invaders and points of entry The geographic origins and points of entry of nonnative arthropods on woody plants in Canada tend to largely reflect human colonization and trade patterns. All but two species of the 419 species known from Canada originate from Palearctic, 82% of them from Europe. This is not surprising as, historically, the overwhelming source of colonists and trade for Canada, excepting the USA, was Europe. Over the last 20 years, as trade with Asia (particularly China) increased dramatically (Asia Pacific Gateway 2007), the proportion of non-native species originating from Asia increased to at least 19% of newly introduced species compared to 6.7% before this period. As the USA has had a much greater volume of trade and immigrants with the Europe and northern Asia than Canada, it is not surprising that most of the Palearctic arthropods introduced to North America first arrived there. Consequently, about 60% of the non-native insect and mite species in Canada are thought to have arrived (naturally or aided by humans) via the United States. As well, some species directly introduced to Canada have also spread to the USA (Sailer 1983). This inadvertent exchange of non-native species between Canada and the USA clearly argues that both countries need to continue to work together in developing policies concerning invasive alien species. As a large portion of the historical and current trade between Canada and the Palearctic was via ship, it is the major shipping ports that served as the point of entry for the majority of species: Halifax and St. John’s on the east coast; Vancouver and Victoria on the west coast; and the Great Lakes, especially ports on Lake Ontario. The interior areas of Canada, such as the prairie provinces, have not been important entry points for exotic species. However, with the

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

large increase in source-to-destination container shipping in recent years, it may be expected that increasingly areas of Canada far from ports may be sites of new introductions (Allen and Humble 2002). Current distribution Herein we only coarsely describe the distribution of non-native species, i.e., at the provincial and territorial level. Ongoing work to database specimen records in the collections of the Canadian Forest Service, the Canadian National Collection of Insects, and selected university collections (Langor, unpublished) will soon allow more detailed distribution maps to be generated, and permit spatial and temporal analyses of invasion patterns and species spread. Clearly BC, ON, QC, and NS are the areas with the highest diversity of non-native species. To a large extent this pattern is a reflection of the main points of entry (i.e., ports) for many of the species directly introduced to Canada. As well, southern parts of these provinces are where species spreading from the USA most commonly enter Canada. Southern BC, ON and QC also have the highest diversity of woody plants in Canada, which maximizes habitat options for newly introduced species. Unsurprisingly, the most depauperate areas (the prairie provinces and the territories) are those furthest from major points of introduction. It is expected that the longer the period of time a species has been present in Canada, the wider its distribution should be, and indeed this relationship is evident and significant. As well, current distribution is undoubtedly influenced by the number of independent introduction events and the geographic spread of distribution points. Rate of spread of a species will also be affected by natural dispersal abilities, capacity to establish populations from a small number of immigrants (i.e., parthenogenetic versus sexual reproduction) and presence of suitable hosts. Many introduced species are also inadvertently spread by humans. For example, commercial tree nurseries have been very important in aiding the spread of birch leafmining sawflies (Langor et al. 2000). Knowledge of the current distribution of nonnative species is not equivalent for all areas of the country, nor for all taxonomic groups. Hemiptera are best surveyed in ON and BC where most recent and current taxonomists working on this order were located. Historically, most coleopterists working in

15

Canada were located in the eastern half of the country, and this area is likely best known, especially ON and QC. There are ongoing projects to inventory the beetles of the maritime provinces (Majka 2007) and NL (Biological Survey of Canada 2007b). The Symphyta are best known for ON, QC and the prairie provinces, as these are areas where taxonomists have been most recently active with collecting efforts. However, for the five species of exotic birch leafmining sawflies there has been a recent extensive survey of much of Canada, resulting in many new provincial records (Digweed and Langor, unpublished data), and this effort showed that there were large gaps in our knowledge about distribution even for species that we putatively know well. The macroLepidoptera has received much taxonomic and inventory focus throughout most of Canada, and this is one of the best known groups of insects in Canada. The distributions of non-native micro-moths are not nearly as well known. Similarly, there is a dearth of information about distribution of non-native Diptera and mites. Host range The high degree of taxonomic affinity (at the generic level) between European and North American woody plants (Huntley 1993) figures largely in the remarkably successful invasion of Canada by European phytophagous insects. In general, non-native insects and mites have colonized the same genera of woody plants in Canada that are hosts in their native range. However, in some cases, species have colonized host genera that have been extinct from Europe since the Pleistocene, e.g., Taxodium, Thuja, Tsuga, Carya, Liriodendron, Magnolia, Robinia and Sassafrass (Niemela¨ and Mattson 1996). Niemela¨ and Mattson (1996) suggest that some of the phytophagous insect fauna of tree genera that went extinct in Europe were able to adapt and persist on con-familial relatives. Some of these host-switching survivors may maintain the genetic capacity to recolonize their pre-Pleistocene relatives. Woody plant genera with the widest distribution and highest local abundance within its range tend to have the greatest diversity of phytophages (Strong et al. 1984). While this has certainly been shown to be a common pattern for native phytophages, it also appears to hold up for non-native faunas. Among

16

gymnosperms, Picea and Pinus are the most widespread, diverse and locally abundant genera in Canada (Farrar 1995), and these also have the most diverse faunas of non-native phytophages. For angiosperms, Malus, Prunus, Salix, Betula, Quercus and Populus had the most diverse non-native fauna, and these are among the most widespread and abundant genera in Canada. It was notable that Pyrus also had a highly diverse fauna, even though this genus is large localized and abundant in only in a few fruit-growing regions of Canada. It seems that commercial fruit trees (Pyrus, Malus and Prunus) have been especially susceptible to invasive insects. Many European woody plant species have been planted throughout Canada, especially in urban areas (Farrar 1995). The presence of their native hosts likely aids the establishment of some non-native phytophages. For example, the successful establishment of the pine shoot beetle, Tomicus piniperda (L.) in the Great Lakes region appears to have been aided by the presence of its native host, Scots pine, Pinus sylvestris L. (Niemela¨ and Mattson 1996). Almost all native woody plant genera in Canada have been invaded by non-native phytophages. The only genera that appear to be spared thus far are Cephalanthus, Gymnocladus, Hamamelus, Nyssa, Toxicodendron and Zanthoxylum. In fact, it would not be surprising if these genera have also been attacked by non-native insects or mites but have thus far escaped attention. Angiosperms host 3.5 times more non-native phytophagous insects than gymnosperms. This is to be expected as angiosperms are much more diverse and have more diverse faunas in both Europe and North America (Strong et al. 1984). Phytophages on angiosperms also tend to be much more polyphagous than those on gymnosperms (Strong et al. 1984). Temporal pattern of colonization Estimation of the date of arrival of a non-native species is difficult even for species introduced relatively recently, let alone for those introduced centuries ago, even before the advent of modern taxonomy. Nonetheless, historical literature (e.g., diaries and almanacs) at times draws attention to pestiferous species on crops (Sailer 1983). The earliest record is the codling moth, Cydia pomonella (Lepidoptera: Tortricidae), which was recorded from apple trees in Ontario in 1635.

D.W. Langor & J. Sweeney (eds.)

As well, dating of insects found in old latrines can give insights into approximate arrival times of some species (Bain and LeSage 1998). Examination of the dates of arrival in North America for the 419 non-native species on woody plants in Canada yielded a pattern virtually identical to that reported by Sailer (1983) for the almost 1,500 non-native species of insects then known from North America. Very few species were present before about 1840, and those were largely imported on crops, in ballast or on animal dung (Sailer 1983). After 1840, the rate of introduction of species into North America increased rapidly, largely through the accidental introduction of phytophagous insects and mites as the importation of foreign plants came into vogue in the mid-nineteenth century. The rate of introduction of non-native arthropods into the USA peaked between 1900 and 1919. The marked reduction in the rate of species introduction after 1920 likely reflects the deterrent effect of the Plant Quarantine Act which was enacted in 1912 (Sailer 1983). A similar temporal pattern of colonization is evident for Canada, although the period of rapid colonization did not commence until later (1860–1879) and reached a peak later (1940–1959) than for North America as a whole. Clearly the temporal pattern of colonization of Canada by non-native phytophages was influenced by the policies of the USA. Most of the earliest species introduced to Canada spread from the USA, and the enactment of the 1912 Plant Quarantine Act resulted in a decrease in species arriving in Canada via spread from the USA. However, the species introduced directly to Canada did not decline noticeably until the 1980s. In 1976 Canada enacted the Plant Quarantine Act, which more strictly regulated importation of exotic species. It may be that this legislation has been effective in reducing the rate of direct species introductions to Canada, and possibly the spread of exotic species from Canada to the USA. The future Canada and the USA will continue to receive new non-native species through direct accidental importation from other countries (Work et al. 2005; McCullough et al. 2006), albeit hopefully at an ever-decreasing rate. Legislation in the USA (1912 Plant Quarantine Act) and Canada (Plant Quarantine

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

Act of 1976) has apparently aided the fight against invasive species in North America, and it is clear that legislation and policies in each country will have implications for both countries. The North American Forestry Commission should continue to play a central role in fostering communication within North America concerning non-native species on trees. Furthermore, in Canada, federal regulatory controls have been adopted to prevent the introduction of new populations and spread of existing populations of non-native tree-feeding species such as the Asian longhorned beetle, the brown spruce longhorned beetle, the emerald ash borer, and the pine shoot beetle as well as inadvertent introduction of species associated with high-risk commodities, such as those requiring solid wood packaging (Canadian Food Inspection Agency 2007). The new international standards for treating wood packaging material (FAO 2002) should significantly reduce phytosanitary risks associated with such packaging materials, a known mode of entry for many non-native bark-and wood-boring insects in Canada (Humble and Allen 2001). However, as only a small proportion of shipping containers arriving from international destinations are ever inspected, and as containers can be moved practically anywhere in Canada before they are opened (Allen and Humble 2002), there are still many opportunities for non-native species to continue to invade Canada. There is a large number of non-native species on woody plants in the northern USA that have not yet been collected in Canada. Undoubtedly, many or all of these will eventually spread northward. Some of the most important recent introductions into Canada have come via the USA, including the emerald ash borer and the pine shoot beetle. The most recent known introduced insect on trees, the banded elm bark beetle, Scolytus schevyrewi Semenov, was just detected in southern Alberta in 2006 and likely came across the border (possibly aided by humans) from the USA where it has been established for several years (Haack 2006). It may be that this species will vector Dutch Elm Disease and will be an important complicating factor in the fight against DED in Canada and the USA (Negro´n et al. 2005). An under-appreciated and underestimated threat associated with non-native phytophagous insects and mites is the potential danger posed by organisms vectored by these arthropods. For example, many

17

introduced bark and wood-boring beetles have fungi associated with them (Hausner et al. 2005), and some of those fungi are potentially phytopathogenic. Some fungi introduced in association with non-native insects may also be transferred to native insect species that also serve as vectors (Jacobs et al. 2004). As well, aphids are well-known plant virus vectors (Chan et al. 1991). Humble and Allen (2006) argue persuasively that effective management of invasive insect species also requires a detailed understanding of their role as vectors of other organisms. Detection and identification of non-native arthropod species (and associated species of fungi, viruses, etc.) is simply not possible without access to extensive taxonomic expertise (Huber and Langor 2004). Many introduced species, especially those from the Palearctic, have closely related species in Canada. Attempting to identify non-native species using existing keys to Nearctic species will often not allow non-specialists to realize that they have a non-native species in hand, and will lead to a misidentification and a failure to recognize a new introduction. For example, Tetropium fuscum was consistently misidentified as a native species for 10 years before it was recognized by an expert as a Palearctic species (Huber and Langor 2004). Continued investment in taxonomic research, regional and national faunistic inventories and maintenance of reference collections of native and non-native species is essential to maintaining domestic capacity to identify new biotic threats to Canada. The sources of information about non-native species on woody plants in Canada are numerous. Specimens residing in Canadian collections provide information about presence (place and time) of nonnative species, which is useful for assessing patterns and rates of spread. As well, specimen labels often contain host information. Since the early 1930s until the mid-1990s, the Canadian Forest Service (CFS)Forest Insect and Disease Survey (FIDS) has routinely collected information about species affecting woody plants, and this information was published annually in regional and national reports. These reports contain a lot of information about notable invasive insects and fungi. In the near future the CFS will make those reports available in a searchable electronic format via the Internet. Many specimen and damage collections made by FIDS were accompanied by enclosure slips that contained an

18

abundance of ancillary information about damage symptoms, habitat and population levels. This information has been databased and is available for searches and analyses. A large number of web sites currently have information about non-native biota of North America, and most contain information relevant to species in Canada (Ricciardi et al. 2000; Crall et al. 2006). However, many of these are superficial and often do not do justice to the non-native biota of Canada, especially as regards distribution, phenology and hosts. Numerous journal papers and gray reports have been produced in Canada and the USA concerning non-native species, and many of those have extensive information about the biology, impacts, host, distribution and management of invasive species. As well, the taxonomic and faunistics literature contains a wealth of information about distribution, hosts and identity of non-native species. The Internet provides better access to published information than has ever previously been experienced, but still fails to capture (and certainly does not summarize or synthesize) much information concerning non-native species. A current initiative lead by the CFS aims at creating a one-stop-shop web site for information about invasive phytophagous arthropods and fungi on trees in Canada (Langor, unpublished). Clearly, the combined goals of protection of Canada against future invasions and minimization of the effects of current invasions, along with the development of science and policy to aid achievement of those goals, requires ready access to and synthesis of the most current information about non-native species. The catalogue of non-native insects and mites on woody plants will provide a key part of the foundation for development of this integrated information resource. Acknowledgements We thank R. Anderson, P. Bouchard, H. Goulet, A. Hamilton, B. Hemming, D. Lafontaine, J.-F. Landry, S. Laplante, L. LeSage, E. Maw, G. Pohl, G. Scudder and D. Smith for providing information about nonnative species of arthropods on woody plants in Canada. This work was funded by the Canadian Forest Service and Agriculture and Agri-Food Canada.

References Allen EA, Humble LM (2002) Nonindigenous species introductions: a threat to Canada’s forests and forest economy. Can J Plant Pathol 24:103–110

D.W. Langor & J. Sweeney (eds.) Armstrong JA, Ives WGH (eds) (1995) Forest insect pests in Canada. Natural Resources Canada, Canadian Forest Service, Ottawa, ON, 732 pp Asia Pacific Gateway (2007) Canada’s bilateral trade with Asia Pacific. http://www.asiapacificgateway.net/statistics/with_ asia.cfm. Accessed 19 June 2007 Bain A, LeSage L (1998) A late seventeenth century occurrence of Phyllotreta striolata (Coleoptera: Chrysomelidae) from North America. Can Entomol 130:715–719 Biological Survey of Canada (2007a) Canada’s insect fauna. http://www.biology.ualberta.ca/bsc/english/ insectfauna.htm. Accessed 20 June 2007 Biological Survey of Canada (2007b) Terrestrial arthropods of Newfoundland and Labrador. http://www.biology. ualberta.ca/bsc/english/nfld.htm. Accessed 20 June 2007 Bousquet Y (ed) (1982) Checklist of beetles of Canada and Alaska. Agriculture Canada, Ottawa, ON, Publication 1861/E, 430 pp Canadian Food Inspection Agency (2007) Plant protection policy directives. Canadian Food Inspection Agency. http://www.inspection.gc.ca/english/plaveg/protect/dir/ directe.shtml. Accessed 19 June 2007 Chan CK, Forbes AR, Raworth DA (1991) Aphid-transmitted viruses and their vectors of the world. Agriculture Canada, Ottawa, ON, Technical Bulletin 1991-3E, 216 pp Colautti RI, Bailey SA, van Overdijk CDA, Amundsen K, MacIsaac HJ (2006) Characterised and projected costs on nonindigenous species in Canada. Biol Invasions 8:45–59 Crall AW, Meyerson LA, Stohlgren TJ, Jarnevich CS, Newman GJ, Graham J (2006) Show me the numbers: what data currently exist for non-native species in the USA? Front Ecol 4:414–418 FAO (2002) Guidelines for regulating wood packaging material in international trade. Secretariat of the International Plant Protection Convention, Food and Agriculture Organization of the United Nations, International Standards for Phytosanitary Measures, Rome. Publ. 15 Farrar JL (1995) Trees in Canada. Fitzhenry and Whiteside Ltd and Canadian Forest Service, Ottawa, ON, 502 pp Foottit RG, Halbert SE, Miller GL, Maw E, Russell LM (2006) Adventive aphids (Hemiptera: Aphididae) of America north of Mexico. Proc Entomol Soc Wash 108:583–610 Haack RA (2006) Exotic bark- and wood-boring Coleoptera in the United States: recent establishments and interceptions. Can J For Res 36:269–288 Haack RA, Mattson W (1993) Life history patterns among North American tree-feeding sawflies. In: Wagner MR, Raffa KF (eds) Sawfly life history adaptations to woody plants. Academic Press, Boca Raton, LA, pp 503–545 Haber E (2002) Spread and impact of alien plants across Canadian landscapes. In: Claudi R, Nantel P, MuckleJeffs E (eds) Alien invaders in Canada’s waters, wetlands and forests. Canadian Forest Service, Ottawa, ON, pp 43–57 Hausner G, Iranpour M, Kim JJ, Breuil C, David CN, Gibb EA, Reid J, Loewen PC, Hopkin AA (2005) Fungi vectored by the introduced bark beetle Tomicus piniperda in Ontario, Canada, and comments on the taxonomy of Leptographium lundbergii, Leptographium terebrantis, Leptographium truncatum, and Leptographium wingfieldii. Can J Bot 83:1222–1237

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems Hendrickson O (2002) Invasive alien species in Canadian forests. In: Claudi R, Nantel P, Muckle-Jeffs E (eds) Alien invaders in Canada’s waters, wetlands and forests. Canadian Forest Service, Ottawa, ON, pp 59–71 Huber JT, Langor DW (2004) Systematics: it’s role in supporting sustainable forest management. For Chron 80:451–457 Humble LM, Allen EA (2001) Implications of non-indigenous insect introductions in forest ecosystems. In: Liebhold AM, McManus ML, Otvos IS, Fosbroke SLC (eds) Proceedings: integrated management and dynamics of forest defoliating insects, 1519 August 1999, Victoria, British Columbia. USDA Forest Service, Northeastern Research Station, Newtown Square, PA, General Technical Report NE-277, pp 45–55 Humble LM, Allen EA (2006) Forest biosecurity: alien invasive species and vectored organisms. Can J Plant Pathol 28:S256–S269 Huntley B (1993) Species richness in north temperate zone forests. J Biogeogr 20:163–180 Jacobs K, Bergdahl DR, Wingfield MJ, Halik S, Seifert KA, Bright BD, Wingfield BD (2004) Leptographium wingfieldii introduced into North America and found associated with exotic Tomicus piniperda and native bark beetles. Mycol Res 108:411–418 Krcmar-Nozic E, Wilson B, Arthur L (2000) The potential impacts of exotic forest pests in North America: a synthesis of research. Natural Resources Canada, Canadian Forest Service, Pacific Forestry Centre, Victoria, BC, Information Report BC-X-387, 33 pp Langor DW, Digweed SC, Williams DJM, Spence JR, Saunders C (2000) Establishment and spread of two introduced parasitoids (Ichneumonidae) of the birch leafminer, Fenusa pusilla (Tenthredinidae). Biocontrol 45:415–423 Lanteri AA, Normark BB (1995) Parthenogenesis in the tribe Naupactini (Coleoptera: Curculionidae). Ann Entomol Soc Am 88:722–731 Ledig FT (1992) Human impacts on genetic diversity in forest ecosystems. Oikos 63:87–108 Lindroth CH (1957) The faunal connections between Europe and North America. Wiley, New York, NY, 344 pp Majka C (2007) Atlantic Canada Coleoptera. http://www.che bucto.ns.ca/Environment/NHR/atlantic_coleoptera.html. Accessed 20 June 2007 Mason PG, Huber JT (eds) (2002) Biological control programmes in Canada, 1981-2000. CABI Publishing, Wallingford, UK, 583 pp Mattson WJ, Niemela¨ P, Millers I, Inguanzo Y (1994) Immigrant phytophagous insects on woody plants in the United States and Canada: an annotated list. USDA Forest Service, North Central Forest Experiment Station, St. Paul, MN, General Technical Report NC-169, 27 pp Maw HEL, Foottit RG, Hamilton KGA, Scudder GGE (2000) Checklist of the Hemiptera of Canada and Alaska. NRC Press, Ottawa, ON, 220 pp McCullough DG, Work TT, Cavey JF, Liebhold AM, Marshall D (2006) Interceptions of nonindigenous plant pests at US

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ports of entry and border crossings over a 17 year period. Biol Invasions 8:611–630 Miller GL, Miller DR (2003) Invasive soft scales (Hemiptera: Coccidae) and their threat to U.S. agriculture. Proc Entomol Soc Wash 105:832–846 Miller DR, Miller GL, Watson GW (2002) Invasive species of mealybugs (Hemiptera: Pseudococcidae) and their threat to U.S. agriculture. Proc Entomol Soc Wash 104:825–836 Miller DR, Miller GL, Hodges GS, Davidson JA (2005) Introduced scale insects (Hemiptera: Coccoidea) of the United States and their their impact on U.S. agriculture. Proc Entomol Soc Wash 107:123–158 Negro´n JF, Witcosky JJ, Cain RJ, LaBonte JR, Duerr DA, McElwey SJ, Lee JC, Seybold SJ (2005) The banded elm bark beetle: a new threat to elms in North America. Am Entomol 51:84–94 Neuvonen S, Niemela¨ P (1983) Species richness and faunal similarity of arboreal insect herbivores. Oikos 40:452–459 Niemela¨ P, Mattson WJ (1996) Invasion of North American forests by European phytophagous insects: legacy of the European crucible? Bioscience 46:741–753 Pimentel D, Lach L, Zuniga R, Morrison D (2000) Environmental and economic costs of nonindigenous species in the United States. Bioscience 50:53–65 Ricciardi A, Steiner WWM, Mack RN, Simberloff D (2000) Toward a global information system for invasive species. Bioscience 50:239–244 Sailer RI (1983) History of insect introductions. In: Wilson CL, Graham CL (eds) Exotic plant pests and North American agriculture. Academic Press, New York, NY, pp 15–38 Scudder GGE, Foottit (2006) Alien true bugs (Hemiptera: Heteroptera) in Canada: composition and adaptations. Can Entomol 138:24–51 Strong DR, Lawton JH, Southwood R (1984) Insects on plants: community patterns and mechanisms. Harvard University Press, Cambridge, MA, 313 pp Stuckey RL, Barkley TN (1993) Weeds. Flora of North America, vol 1. Oxford University Press, New York, NY, pp 193–198 Turnbull AL (1979) Recent changes to the insect fauna of Canada. In: Danks HV (ed) Canada and its insect fauna. Mem Entomol Soc Can No. 108, pp 180–194 Wheeler AG, Henry TJ (1992) A synthesis of the Holarctic Miridae (Heteroptera): distribution, biology and origin with emphasis on North America. Thomas Say Foundation Monograph No. 15, Entomological Society of America, Lanham, MD, 282 pp Work TT, McCullough DG, Cavey JF, Komsa R (2005) Arrival rate of nonindigenous insects species into the United States through foreign trade. Biol Invasions 7:323–332 Yarwood CE (1983) History of plant pathogen introductions. In: Wilson CL, Graham CL (eds) Exotic plant pests and North American agriculture. Academic Press, New York, NY, pp 40–63

Ecological effects of invasive alien insects Marc Kenis Æ Marie-Anne Auger-Rozenberg Æ Alain Roques Æ Laura Timms Æ Christelle Pe´re´ Æ Matthew J. W. Cock Æ Josef Settele Æ Sylvie Augustin Æ Carlos Lopez-Vaamonde

Originally published in the Journal Biological Invasions, Volume 11, No 1, 21–45. DOI: 10.1007/s10530-008-9318-y Ó Springer Science+Business Media B.V. 2008

Abstract A literature survey identified 403 primary research publications that investigated the ecological effects of invasive alien insects and/or the mechanisms underlying these effects. The majority of these studies were published in the last 8 years and nearly two-thirds were carried out in North America. These publications concerned 72 invasive insect species, of which two ant species, Solenopsis invicta and

Electronic supplementary material The online version of this article (doi:10.1007/978-1-4020-9680-8_3) contains supplementary material, which is available to authorized users. M. Kenis (&)
C. Pe´re´
M. J. W. Cock CABI Europe-Switzerland, 1 Rue des Grillons, 2800 Delemont, Switzerland e-mail: [email protected] M.-A. Auger-Rozenberg
A. Roques
S. Augustin
C. Lopez-Vaamonde INRA, Station de Zoologie Forestie`re, 45166 Olivet, France L. Timms Faculty of Forestry, University of Toronto, Toronto, Canada M5S3B3 C. Pe´re´ University of Neuchaˆtel, 2009 Neuchatel, Switzerland J. Settele Department of Community Ecology, UFZ—Helmholtz Centre for Environmental Research, 06120 Halle, Germany

Linepithema humile, accounted for 18% and 14% of the studies, respectively. Most publications investigated effects on native biodiversity at population or community level. Genetic effects and, to a lesser extent, effects on ecosystem services and processes were rarely explored. We review the effects caused by different insect invaders according to: their ecosystem roles, i.e. herbivores, predators, parasites, parasitoids and pollinators; the level of biological organisation at which they occur; and the direct and indirect mechanisms underlying these effects. The best documented effects occur in invasive ants, Eurasian forest herbivores invasive in North America, and honeybees. Impacts may occur through simple trophic interactions such as herbivory, predation or parasitism. Alien species may also affect native species and communities through more complex mechanisms such as competition for resources, disease transmission, apparent competition, or pollination disruption, among others. Finally, some invasive insects, particularly forest herbivores and ants, are known to affect ecosystem processes through cascading effects. We identify biases and gaps in our knowledge of ecological effects of invasive insects and suggest further opportunities for research. Keywords Invasive alien species
Insect invasions
Ecological effect
Ecological impact
Native species
Natural communities
Predators
Parasitoids
Pollinators
Review

D.W. Langor & J. Sweeney (eds.), Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems. DOI: 10.1007/978-1-4020-9680-8_3

21

22

Introduction The threat posed by invasive alien species on biodiversity is widely recognized (Williamson 1996; Wittenberg and Cock 2001; Pimentel 2002). Although insects form a large part of the alien fauna worldwide, invasive alien insects appear to have received disproportionately less attention regarding their effects on the environment compared to plants, vertebrates, or aquatic organisms (Parker et al. 1999; Levine et al. 2003; Long 2003). Alien insects can affect native biodiversity through direct interactions, e.g. a herbivore feeding on a native plant (Jenkins 2003), a predator or a parasitoid attacking a native prey or host (Boettner et al. 2000; Snyder and Evans 2006), an alien species hybridizing with a native species (Jensen et al. 2005), etc. They can also affect native species and ecosystems indirectly, through cascading effects, or through various mechanisms, such as carrying diseases, competing for food or space or sharing natural enemies with native species (NRC 2002). Ecological impact by invasive species can occur at different levels of biological organisation: genetic effects; effects on individuals, populations or communities of species; and effects on ecosystem processes (Parker et al. 1999). It can also occur at different spatial scales, from microhabitat to landscape (Williamson 1996). Parker et al. (1999) surveyed for published reports of quantitative data on impacts by various categories of invasive organisms. The majority focused on population effects and most studies were carried out in a correlative manner—e.g. comparing sites before and after invasion, or sites inside and outside the invasion range, but only a few of these studies used designed experiments to assess the mechanisms or pathways through which these impacts occur. For terrestrial invertebrates, they identified two dozen publications, more than half of them illustrating population-level effects. To date, reviews on the ecological effect of invasive alien insects have been published on particular taxa, such as ants (Holway et al. 2002), bees (Goulson 2003; Moritz et al. 2005) and mosquitoes (Juliano and Lounibos 2005), for specific regions, such as the Galapagos Islands (Causton et al. 2006), or for particular impact mechanisms, such as the ecological impact of generalist predators (Snyder and Evans 2006). This paper provides the first comprehensive

D.W. Langor & J. Sweeney (eds.)

literature review of the ecological effects of invasive insects. We first make a general analysis of the literature presently available on the topic and review the impacts caused by different invaders according to their ecosystem roles—herbivores, predators, parasites, parasitoids and pollinators. Within these groups we analyse ecological effects at different levels of biological organisation—genetic, population/community and ecosystem—and through different ecological mechanisms, e.g. herbivory, predation, parasitism, resource competition, and various indirect mechanisms. We also try to identify gaps in knowledge of ecological hazards by invasive insects and to stimulate new approaches to fill these gaps. The study was carried out as part of the EU project ALARM (Settele et al. 2005).

Published studies on the ecological impact of invasive insects Relevant primary research publications on the ecological effects of invasive alien insects were first identified by electronic searches in CAB Abstracts covering the period 1900–2007. Since the terminology used in the context of invasive species has changed during this period, the widest possible variety of terms (e.g. invasive or alien or non-indigenous or exotic; impact or effect, displacement) were entered in the search engine, in various combinations. Then the references in these sources were examined for additional relevant publications. Only papers published until 2007 were included in the general analysis, but some relevant papers in press are cited in the text. In a couple of cases, general publications were included in the analysis when they described essential unpublished research. Studies describing the effect on single individuals without information on the effect at population level (e.g. an alien parasitoid emerged from a native species, or an alien herbivore found attacking an alien plant) were not taken into account, unless the mortality described in the paper affected a measurable and significant proportion of the regional or world population of a native species (e.g. Stiling and Moon 2001; Fowler 2004). Studies showing negative results (i.e. no ecological effect) were included. In contrast, papers on risk assessments or on alien biological control agents showing a positive effect

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

23

45 40

No. of publications

35 30 25 20 15 10 5

2006

2004

2002

2000

1998

1996

1994

1992

1990

1988

1986

1984

1982

1980

1978

1976

1974

1972

1970

0

Fig. 1 Number of publications found on primary research investigating the ecological impact of invasive insects and/or the mechanisms underlying these impacts, from 1970 to 2007.

Four publications from 1930, 1952, 1961 and 1963 are not shown in the figure

on the environment, either through the control of a pest of ecological importance or through pesticide reduction, were excluded. Papers describing a negative impact of phytosanitary methods implemented to control invasive insects were also excluded, as well as those mentioning interference with pests in purely agricultural systems. Finally, we did not consider in the analyses the effect of alien species on other alien species, unless this effect had an indirect consequence on native biodiversity. Papers were classified following the biological organisation level at which the investigated effects occur (genetic, population/community, and ecosystem) and based on how the effect was quantified: (1) studies based on field observations, usually comparative studies between invaded and non invaded sites, or comparing data before and after invasion; (2) field studies with a significant experimental component, e.g. exclusion experiments, exposure of sentinel animals or plants, etc.; (3) laboratory experiments or mathematical models used to investigate impact mechanisms. Studies of the third type were included in the database only when they were based on field data showing or suggesting that impact by the invasive insect already occurs. The identity of the invasive species, the year of publication and the continent/region in which the study was carried out were also included in the database.

A total of 403 primary research papers were identified that investigate the ecological effect of invasive insects and/or the mechanisms underlying these effects (See electronic Appendix A available on the Biological Invasions web site). Although these represent only a fraction of the publications available, we believe that they are a representative sample of the published literature on the topic. Nearly 60% of the papers were published in the last 8 years, even though many of the alien insects were introduced several decades earlier (Fig. 1). Few papers on invasive alien insects published before the 1990s were included in our analyses because, in most cases, earlier papers described damage on individuals, economic injuries or anecdotal observations on ecological effects, but did not provide reliable and measurable data on the effects on native populations, communities and ecosystem processes. This suggests that the ecological impact of insects is a relatively new area of research, or that the ecological effects of invasive alien insects were of little concern until recently, and we can expect that much more information on impacts and impact mechanisms will become available in the next few years. About 62% of the studies were carried out in North America, followed by Oceanic Islands (13%) and Australia/ New Zealand (8%) (Fig. 2). Only a few studies were conducted in the other continents Vila´ et al. (2006)

24

D.W. Langor & J. Sweeney (eds.)

Oceanic Islands South and Central America North America Europe Australia/New Zealand Asia Africa 0

50

100

150

200

250

300

No. of publications

Fig. 2 Number of publications found on primary research investigating the ecological impact of invasive insects and/or the mechanisms underlying these impacts, from different continents and regions. Oceanic Islands includes islands from all oceans and of less than 20,000 km2

found a similar pattern for plants. They showed that 59% and 21% of studies assessing the ecological effects of invasive plants were carried out in North America and Oceania, respectively. This discrepancy in the number of studies on the effect of invasive species between regions and continents probably reflects the higher incidence of invasive species usually observed in Oceanic islands, Oceania, and North America compared to other regions (Simberloff 1986; Niemela¨ and Mattson 1996; Pimentel 2002). However, literature searches being usually made in English and references in other languages being generally under-represented in peer-reviewed literature, a bias towards English-speaking countries cannot be ruled out. Only nine of these 403 publications describe investigations on genetic effects. Publications on the effect of invasive insects on ecosystem processes are more numerous (25 publications), but concern nearly exclusively ants (e.g. Solenopsis invicta Buren) and, especially, forest herbivores in North America (e.g. Adelges tsugae Annand and Lymantria dispar (L.)). Most studies analyse effects at the population or community level. Effects were most commonly assessed or studied through field comparisons of populations in invaded and non-invaded areas (224 publications), but field studies involving an experimental component were more frequent than previously expected (106 publications). Laboratory experiments (104 publications) concern mainly intraguild predation tests with the invasive predators. Parker et al. (1999) also observed that population

level effects are by far the most commonly documented ecological effects for invasive terrestrial invertebrates, as for all other taxonomic groups. Impact studies were found for 72 invasive insect species and evidence for ecological effects in the field was found for 54 of them (Table 1). Two ant species, S. invicta and Linepithema humile (Mayr), account for 18% and 14% of the studies, respectively. Other extensively studied species include the honey bee Apis mellifera L.(7%), three Eurasian forest pests introduced into North America, L. dispar (6%), Adelges piceae (Ratzeburg) (5%) and A. tsugae (5%), and a biological control agent, the Asian ladybird Harmonia axyridis (Pallas) (6%). All together, invasive ants were the target of 41% of the studies, other predators 19%, parasitoids and parasites 6%, herbivores 24% and pollinators 10% (Fig. 3).

Genetic effects Hybridization between invasive and native species may be of major concern because of the disturbances it can induce in native genetic resources (Huxel 1999; Mallet 2005). Hybridization has been well documented in vertebrates and plants and, in several cases, has been shown to have a strong negative impact on native species (Rhymer and Simberloff 1996; Vila` et al. 2000, 2006; Long 2003). In contrast, genetic impacts related to invasions of insects and other terrestrial invertebrates remain largely unexplored. Indeed, most studies focused on the genetic structure of insect invaders (Tsutsui and Case 2001; Lee 2002), especially with the aim of tracing their origin (Scheffer and Grissell 2003; Grapputo et al. 2005; Havill et al. 2006). No case of horizontal gene transfer is reported except in some laboratory tests (Labrador et al. 1999) and good examples of interspecific hybridization are scarce, and mainly concern laboratory experiments, e.g. with bumblebees (Mitsuhata and Ono 1996). More examples concern hybridization between native and introduced bees and bumblebee subspecies. The shipment of vast numbers of non-native honeybees and bumblebees throughout the world has already resulted in noticeable genetic effects. The massive introduction in north-western Europe of two subspecies of A. mellifera originating from southern Europe, A. m. ligustica S. and

1 1

Bemisia tabaci (Gennadius) (Hem.: Aleyrodidae)

Cactoblastis cactorum (Berg) (Lep.: Pyralidae)

Yes

Yes

1 1 2 5 1 1 2

Pieris rapae (L.) (Lep.: Pieridae)

Pineus boerneri Annand (Hem.: Adelgidae)

Rhinocyllus conicus (Froelich) (Col.: Curculionidae)

Urophora affinis (Frauenfeld) (Dipt.: Tephritidae)

Urophora quadrifasciata (Meigen) (Dipt.: Tephritidae)

Zizina labradus (Godart) (Lep.: Lycaenidae)

Yes

Yes

Yes

Yes

Yes

Yes

Yes Yes

1 1

Orthezia insignis Browne (Hem.: Ortheziidae) Penicillaria jocosatrix Gene´e (Lep.: Noctuidae)

Yes

Yes

Nezara viridula (L.) (Hem.: Pentatomidae)

Lipara sp. (Dipt.: Chloropidae) 25

1 1

Larinus planus (F.) (Col.: Curculionidae)

Yes

Yes

Yes Yes

No

Yes

No

Yes

Yes

Yes

No

No

No

No

Yes

Yes

Ecological effect found in the field

Lymantria dispar (L.) (Lep.: Lymantridae)

1 1

Elatobium abietinum (Walker) (Hem.: Aphididae) Erythroneura variabilis Beamer (Hem.: Ciccadellidae)

Icerya purchasi Maskell (Hem.: Margarotidae)

1 1

Drosophila subobscura Collin (Dipt.: Drosophilidae)

Homalodisca coagulata (Say) (Hem.: Ciccadellidae)

8 3

Cryptococcus fagisuga Lindinger (Hem.: Eriococcidae)

1

1

Andricus quercuscalicis (Burgsdorf) (Hym.: Cynipidae)

1

1

Andricus lignicola (Hartig) (Hym.: Cynipidae)

Coptotermes formosanus Shiraki (Isopt.: Rhinotermitidae)

1

Andricus kollari (Hartig) (Hym.: Cynipidae)

Chilo partellus (Swinhoe) (Lep.: Pyralidae)

1

18

Andricus corruptrix (Schlechtendal) (Hym.: Cynipidae)

19

Adelges tsugae (Annand) (Hem.: Adelgidae)

No. of publications

Adelges piceae (Ratzeburg) (Hem.: Adelgidae)

Herbivores

Species

USA New Zealand

P/Gd

USA

USA

USA

Madeira

Guam

St. Helena

Japan

USA

USA

USA

Galapagos

French Polynesia

USA USA

USA

South Africa

USA

China, Australia

USA

North America

Country or region where the effect occurs

P

P

P

P

P

P

P

P

P/E

P

P

P

P

P P

P/E

P

P

P

P/E

P/E

Organisational levela

Table 1 Non exhaustive list of invasive alien insects for which primary research publications were found that describe investigations on their ecological effects at various levels of biological organisation

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems 25

3 71

Solenopsis geminata (F.) (Hym.: Formicidae)

Solenopsis invicta Buren (Hym.: Formicidae)

2 4 1 1 3

Pterostichus melanarius (Illiger) (Col.: Carabidae)

Tenodera sinensis Saussure (Orth.: Mantidae)

Trechus obtusus Erichson (Col.: Carabidae)

Vespula germanica (F.) (Hym.: Vespidae) Vespula vulgaris (L.) (Hym.: Vespidae)

Yes Yes

Yes

Yes

No

P/E P/E

P

P

P P

Yes?b

P

E

P

P

P

P

P

Yes

2

Yes

8

24

Harmonia axyridis (Pallas) (Col.: Cocinellidae)

Yes

Propylea quatuordecimpunctata (L.) (Col.: Coccinellidae)

1

Forficula auricularia L. (Derm.: Forficulidae)

Yes

Polistes dominulus (Christ) (Hym.: Vespidae)

6

Culex quinquefasciatus Say (Dipt.: Culicidae)

Yes

Yes

9

Coccinella septempunctata L. (Col.: Coccinellidae)

P

Yes

1

4

Chrysomya rufifacies (Macquart) (Dipt: Calliphoridae)

Lab onlyc

Yes

1

Chrysomya putoria (Wied.) (Dipt: Calliphoridae)

P P

P

P

P

P/E

P

P

P

P

P/E

Organisational levela

Yes Yes

1

2 2

Cicindelidia trifasciata (F.) (Col.: Carabidae) Chrysomya albiceps (Wiedemann) (Dipt.: Calliphoridae)

No

Lab onlyc

No

Yes

Yes

Yes

Yes

Yes

Yes

Yes

Yes

Ecological effect found in the field

Limnophyes minimus (Meigen) (Dipt.: Chironomidae) Oopterus soledadinus (Gue´rin-Me´neville) (Col.: Carabidae)

1

Aedes albopictus (Skuse) (Dipt.: Culicidae)

Calliphora vicina Robineau-Desvoidy (Dipt.: Calliphoridae)

4 10

Adalia bipunctata (L.) (Col.: Coccinellidae)

Other predators and detritivores

9

12

Pheidole megacephala (F.) (Hym.: Formicidae)

Wasmannia auropunctata (Roger) (Hym.: Formicidae)

1

Paratrechina fulva (Mayr) (Hym.: Formicidae)

1

56

Solenopsis wagneri Santschi (Hym.: Formicidae)

16

Linepithema humile (Mayr) (Hym.: Formicidae)

No. of publications

Anoplolepis gracilipes (Fr. Smith) (Hym.: Formicidae)

Ants

Species

Table 1 continued

New Zealand New Zealand

Hawaii

USA

USA

USA

Kerguelen

Marion Island

North America

USA

Hawaii, New Zealand

USA

USA

Galapagos Brazil

New Caledonia, Gabon, Galapagos

USA

USA

Australia, Polynesia, Galapagos

Australia, Hawaii, Fiji, Mexico

Colombia

USA, Southern Europe, South Africa, Pacific Islands, Japan

Several tropical islands, India

Country or region where the effect occurs

26 D.W. Langor & J. Sweeney (eds.)

Lab only

4 5 2 4 1 1

Microctonus aethiopoides Loan (Hym.: Braconidae)

Philornis downsi Dodge & Aitken (Dipt.: Muscidae)

Pteromalus puparum (L.) (Hym.: Pteromalidae)

Torymus sinensis Kamijo (Hym.: Torymidae)

Trichopoda pilipes (F.) (Dipt.: Tachinidae)

Trissolcus basalis (Wollaston) (Hym.: Scelionidae)

1

Megachile rotundata (F.) (Hym.: Apidae)

No

No

Yes

Yes

No

No

Yes

No

Yes

c

d

c

b

a

P/G

P/G

G

P

P

P

P

P P

P

Organisational levela

Worldwide

Worldwide

Japan

Galapagos

Italy

USA

Fiji Italy

USA

Country or region where the effect occurs

Genetic effect (hybridization) speculated but not proved

Effects found in laboratory experiments or using mathematical models, but good evidences for effects in the field are still lacking

Only in association with other Coccinellidae

Level of biological organisation at which effects were found in the field. G = Genetic effects; P = Effect on Populations or communities; E = Effects on ecosystem processes

All references are cited in the electronic Appendix A. Criteria for being included in the list are described in the text

1

Bombus terrestris (L.) (Hym.: Apidae)

Megachile apicalis Spinola (Hym.: Apidae)

28 13

Apis mellifera L. (Hym.: Apidae)

Pollinators

Yes

2

Lysiphlebus testaceipes (Cresson) (Hym.: Braconidae)

No

Yes

2 1

Yes Yes

Yes

Cotesia glomerata (L.) (Hym.: Braconidae)

1 1

Ecological effect found in the field

Compsilura concinnata (Meigen) (Dipt.: Tachinidae)

1

Bessa remota (Aldrich) (Dipt.: Tachinidae) Cales noaki Howard (Hym.: Aphelinidae)

No. of publications

Aphidius ervi (Haliday) (Hym.: Braconidae)

Parasitoı¨ds/parasites

Species

Table 1 continued

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems 27

28

D.W. Langor & J. Sweeney (eds.)

Pollinators

Herbivores

Parasitoids

Other predators*

Invasive ants 0

50

100

150

200

No. of publications

Fig. 3 Number of publications on primary research investigating the ecological effects of invasive alien insects belonging to different functional groups. * ‘‘Other predators’’ also include detritivores

A. m. carnica Pollmann, has caused large-scale gene flow and introgression between these subspecies and the native black honeybee, A. m. mellifera (Jensen et al. 2005), whose native populations are now threatened in north-west Europe (Goulson 2003) and in the Canary islands (De La Ru`a et al. 2002). A similar problem exists with bumblebee subspecies, Bombus terrestris dalmatinus Dalle Torre and B. t. sassaricus Tournier, originating from the Middle East and Sardinia respectively, which have been introduced in vast numbers worldwide as managed pollinators of glasshouse crops. Indeed, there is a real risk that commercial and native subspecies will hybridize (Ings et al. 2005a, b). Even if growers are advised to prevent the escape of sexuals (queens and males) that could interbreed with native bumblebees, this measure might not be enough since workers can successfully produce males (by arrhenotokous parthenogenesis) by invading congener colonies (LopezVaamonde et al. 2004). However, suitable molecular markers still need to be identified to genetically characterize B. terrestris subspecies and to identify evidence for recent hybridization. Other cases of hybridization between native and invasive species or subspecies are rather speculative, since they have not been confirmed by proper genetic studies. A frequently cited example is the Australian lycaenid butterfly Zizina labradus (Godart), which has apparently displaced the endemic Z. oxleyi (Felder) in several regions in New Zealand (Barlow and Goldson 2002). Hybridization is the most likely mechanism responsible for the displacement because the two species—often regarded as sub-species—interbreed

freely at sites where they still occur sympatrically (Gibbs 1980, 1987). However, other impact mechanisms such as competition for resources or indirect competition through shared natural enemies cannot be ruled out. Interestingly, hybridization was also suspected to be the cause of the observed displacement of the native Torymus beneficus Yasumatsu and Kamijo by the introduced Torymus sinensis Kamijo, hymenopteran parasitoids of the chestnut gall wasp, Dryocosmus kuriphilus Yasumatsu, in Japan (Yara 2006). However, recent molecular studies revealed that hybridization between the two species in the field was marginal and, thus, was probably not playing a significant role in the displacement of the native species (Yara et al. 2007). Due to the limited number of studies, the question of whether genetic risks, especially hybridization, due to invasive populations are weak or just underestimated in insects remains open. Further research is badly needed, in particular the molecular quantification of gene flow between introduced and native species.

Ecological effects due to herbivores Direct effect on native plant populations Alien invertebrate herbivores can be particularly harmful to native plant populations, sometimes driving them to local extinction. However, most publications reporting ecological effects of alien herbivores do not properly quantify these effects. The best documented effects of invasive invertebrate herbivores are undoubtedly those caused by forest insects. North America has been particularly affected by invasive forest insects from Eurasia. For example the balsam woolly adelgid, A. piceae, and the hemlock woolly adelgid, A. tsugae, are threatening unique forest ecosystems in eastern North America by killing Fraser fir (Abies fraseri (Pursh) Poir) and Eastern and Carolina hemlock (Tsuga canadensis (L.) Carr. and T. caroliniana Engelm.) on a large scale, so that they are gradually replaced by other tree species (e.g. Smith and Nicholas 2000; Jenkins 2003; Small et al. 2005; Weckel et al. 2006). In particular, A. tsugae poses a major threat to the viability of Carolina hemlock, a rare endemic tree species in the Appalachian Mountains. Since its accidental

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

introduction from Europe to North America in the nineteenth century, the gypsy moth, L. dispar, has become the main pest of broadleaved trees in Eastern North America. Repeated defoliation may lead to severe tree mortality, particularly in oak (Quercus spp.) stands (Kegg 1971; Allen and Bowersox 1989; Liebhold et al. 1995; Fajvan and Wood 1996). Other examples of Eurasian insects causing serious concern for North American tree species include the spruce aphid, Elatobium abietinum (Walker), threatening Engelmann spruce, Picea engelmannii Parry ex. Engelm. (Lynch 2004), and the newly introduced emerald ash borer, Agrilus planipennis Fairmaire, which, in a few years, has already killed 15 million ash trees (Fraxinus spp.) (Poland and McCullough 2006). Endemic flora on islands are particularly vulnerable to herbivore invasions. In St Helena, the scale insect Orthezia insignis Browne was in the process of pushing the endemic gumwood, Commidendrum robustum (Roxb.), to extinction when a successful biological control programme was implemented (Fowler 2004). Similarly, in the Galapagos, another scale, Iceria purchasi Maskell, has severely affected populations of endangered plants (Roque-Albelo 2003). Here again the introduction of its natural enemy, Rhodolia cardinalis Mulsant, clearly mitigated its effect. Sometimes, biological control agents introduced against exotic weeds may have an adverse effect on native plants. The weevil Rhinocyllus conicus (Fro¨lich), introduced in North America to control the exotic weed nodding thistle, Carduus nutans L., now feeds on many native thistles including the endangered Pitcher’s thistle (Cirsium pitcheri (Torr. ex Eaton) Torr. & A. Gray), significantly reducing the seed production of native thistles (e.g. Louda et al. 1997, 2005; Russel and Louda 2005). Native thistles, in particular Tracy’s thistle, Cirsium undulatum var. tracyi (Rydb.) Welsh, are also threatened by another European weevil, Larinus planus (F.) accidentally introduced in North America but deliberately distributed within the continent as a biological control agent (Louda and O’Brien 2002). The cactus moth, Cactoblastis cactorum (Berg), was introduced to the Caribbean to successfully control prickly pear cacti, Opuntia spp., in 1956. In 1989, it was found in Florida, where it is now threatening the survival of already endangered indigenous Opuntia species (Stiling and Moon 2001; Stiling et al. 2004).

29

It is also spreading along the coast towards Mexico, an important centre of diversity and endemism for Opuntia spp., where some species are also of significant economic importance (Perez-Sandi 2001). Indirect effects on native plant communities By killing and reducing host plant populations, invasive herbivores also indirectly affect populations of other native plant species and native plant communities. Defoliation by L. dispar can cause a major shift in tree species in North America, either directly through tree mortality (Allen and Bowersox 1989; Fajvan and Wood 1996) or via seed failures and mortality of oak seedlings (Gottschalk 1990). The dramatic mortality observed in Fraser fir and Eastern hemlock due to Adelges spp. has totally altered forest plant communities in these forest ecosystems (Jenkins 2003; Eschtruth et al. 2006; Weckel et al. 2006). Busing and Pauley (1994) showed that the loss of Fraser fir due to A. piceae has increased wind exposure and, consequently, mortality of remaining canopy trees, in particular red spruce, Picea rubens Sarg. Indirect effects on native fauna Invasive herbivores may affect populations and communities of native herbivores by competing for the same resource, although mechanisms underlying competition are not always fully understood (Reitz and Trumble 2002). An Asian adelgid, Pineus boerneri Annand, has been shown to be competitively superior and to displace a native congener, P. coloradensis (Gilette) in red pine (Pinus resinosa Aiton) plantations in Eastern USA, probably by reducing host plant quality and forcing P. coloradensis to less suitable sites (McClure 1984, 1989). The European weevil R. conicus, feeding on flowerheads of native thistles in North America, significantly decreases the density of native tephritid flies, which also feed on the flowerheads, at high weevil density (Louda et al. 1997). The scale insect I. purchasi, by killing endangered plants in the Galapagos, has also caused local extinctions of host-specific Lepidoptera (Roque-Albelo 2003). Fabre et al. (2004) suspect resource competition between native and exotic seed chalcids of the genus Megastigmus spp. in Europe, and displacement of the native species. The African

30

stem borer Busseola fusca (Fuller) seems to be displaced from grain sorghum fields by the Asian stem borer Chilo partellus (Swinhoe) (Kfir 1997), perhaps because the native species is deterred by the invasive species, or because of differences in their phenology. Oak defoliation by gypsy moth, L. dispar, may negatively affect populations of the northern tiger swallowtail, Papilio canadensis Rothschild & Jordan. Adult female swallowtails are incapable of distinguishing between damaged and undamaged leaves and laboratory experiments showed that defoliation by gypsy moths depressed swallowtail growth rate and survival (Redman and Scriber 2000). An invasive herbivore can also displace other indigenous species via behavioural interference. A striking example is the rampant invasion of biotype B of the sweet potato whitefly, Bemisia tabaci (Gennadius). This biotype is one of the world’s most damaging agricultural pests and has displaced two indigenous biotypes of this species because invading males interfere with mating by native males and invading females produce more female offspring (Liu et al. 2007). Biological control may provide an opportunity to confirm competitive displacement a posteriori. For example, the displacement of native Lepidoptera by the exotic noctuid moth Penicillaria jocosatrix Guene´e in Guam was confirmed by a successful biological control program against P. jocosatrix, which allowed the native species to recover (Schreiner and Nafus 1993). Invasive herbivores do not only affect closelyrelated species. The disturbance of Fraser fir forests by the balsam woolly aphid has had a detrimental effect on local birds, 10 out of 11 species declining, and six species by more than 50% (Rabenold et al. 1998). Similarly, the decline of eastern hemlock due to A. tsugae in North America strongly affects bird species composition (Tingley et al. 2002) and also has an effect on salamander populations (Brooks 2001), and on deer survival through modifications in forest microclimates (Lishawa et al. 2007). The indirect consequences of L. dispar outbreaks on native birds, have been extensively studied (e.g. Bell and Whitmore 2000; Gale et al. 2001). However, in contrast to Adelges spp., defoliations by L. dispar induced only temporary changes to bird populations and communities, probably because the general impact on the dominant tree species is less dramatic

D.W. Langor & J. Sweeney (eds.)

for L. dispar than for the two adelgids. Interestingly, Thurber et al. (1994) observed that nests in sites defoliated by L. dispar suffered a higher predation rate than did those in non-defoliated sites. Finally, invasive species may also affect native predators through intoxication. The glassy-winged sharpshooter, Homalodisca coagulate (Say), has recently invaded islands of French Polynesia, where it represents a poisonous prey for native spiders (Suttle and Hoddle 2006). Laboratory experiments showed that H. coagulata can be lethal for two native spider species, and preliminary field surveys suggest that the invasive species may already have adversely affected an endemic spider population on at least one island.

Indirect impact as vectors of plant and insect diseases Invasive herbivores may affect native plants by transmitting or facilitating diseases. The European elm bark beetle, Scolytus multistriatus (Marsham) is the vector of the infamous Dutch elm disease, Ophiostoma ulmi (Buisman) Nannf., and O. noviulmi (Brasier) in North America (Brasier 2000). The European beech scale, Cryptococcus fagisuga Lindinger is associated with the fungus, Neonectria faginata (Lohman et al.) Castl. & Rossman, to cause beech bark disease, which devastates American beech (Fagus grandifolia Ehrh.) in North America (Houston 1994; Morin et al. 2007). The insects themselves are relatively minor pests, but the related diseases have a tremendous impact on North American forest species and ecosystems. Non native insects may also affect native insects by transmitting diseases. An interesting case is the extinction of the Madeiran large white, Pieris brassicae wollastoni Butler. This remarkable endemic disappeared a few years after the introduction in Madeira of the congeneric pest species, Pieris rapae (L.), which is now one of the most abundant butterflies in the island (Wakeham-Dawson et al. 2002; Aguiar-Franquinho and Karsholt 2006). Gardiner (2003) suggests that the introduction of P. rapae brought a different strain of the granulosis virus for which P. brassicae wollastoni had no resistance, which in turn lead to the sharp decline and ultimate extinction of this island endemic.

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

Indirect effects through apparent competition Apparent competition occurs when the presence of one species indirectly decreases the fitness of another through the increased presence of a shared enemy (Holt 1977). Very few studies have investigated such interactions in invertebrates, and fewer still in the context of invasive insects. The earliest of these studies investigated the correlation between invasion of the variegated leafhopper, Erythroneura variabilis Beamer, and decreases in populations of a congeneric native grape leafhopper, E. elegantula Osborn, in California vineyards (Settle and Wilson 1990). Field experiments and collections revealed that although neither species was superior in direct competition, declines in E. elegantula populations were correlated with increased levels of parasitism by a native mymarid wasp, Anagrus epos Girault, in the presence of E. variabilis. The invasion of L. dispar in North America has provided numerous possibilities for apparent competition with native species. Efforts to biologically control the gypsy moth have led to the introduction of over 60 species of parasitoids from Europe and Asia. Many of the natural enemies found attacking gypsy moth in North America are generalist parasitoids of Lepidoptera. These include the polyphagous tachinid fly Compsilura concinnata (Meigen), which has been implicated in the decline of several endangered saturniid moths (Boettner et al. 2000) (see section on parasitoids below). Redman and Scriber (2000) also investigated the effect of gypsy moth on native northern tiger swallowtails, P. canadensis, through a suite of indirect interactions. Although they did not examine the competitive interactions between the two species in the absence of shared enemies, they did find that parasitism rates of the native caterpillar more than doubled in the presence of the gypsy moth (Redman and Scriber 2000). Gypsy moth outbreaks also favour a generalist predator, the white-footed mouse, resulting in an increase in tick populations and in the incidence of Lyme disease (Jones et al. 1998). Some studies failed to observe apparent competition between invasive and native insects. Scho¨nrogge and Crawley (2000) used quantitative linkage webs to investigate the effect of alien cynipid gall wasps on native gall wasps in the UK through shared native parasitoids and inquilines. They concluded that the

31

recruitment of parasitoids and inquilines by the invading species was unlikely to have a strong effect on the native species because the native parasitoids did not exhibit strong responses to the invasive wasps. In other cases, shared natural enemies between an invasive and a native species may favour the latter. In North America, Hoogendoorn and Heimpel (2002) compared parasitism rates in an invasive ladybeetle, H. axyridis, and a native ladybeetle, Coleomegilla maculata (DeGeer), by a native braconid wasp, Dinocampus coccinellae (Schrank). They used parasitism rates from field collections and a calculated measure of host susceptibility from lab trials to parameterize a model of parasitoid-mediated interactions between the two species. Hoogendoorn and Heimpel (2002) concluded that invasion by H. axyridis may actually be beneficial to C. maculata because limited susceptibility of the invader to D. coccinellae can allow it to act as an ‘egg sink’, reducing the abundance of the parasitoid in the community. Effects on ecosystem processes Good studies on the effect of invasive insects on ecosystem processes are rare and most examples concern the effect of herbivores on forest ecosystems through tree defoliation or mortality. Effects on North American oak forests by L. dispar defoliation have been extensively investigated (see review in Lovett et al. 2002, 2006). Defoliation decreases transpiration, tree growth and seed production and increases tree mortality, light penetration to the forest floor and water drainage. It alters tree species composition and consequently, faunistic composition. It may also alter carbon allocation and nitrogen cycling, which may have consequences such as acidification of stream waters. Adelges tsugae provides another example of a forest insect for which the impact on ecosystem processes has been well studied. A major effect of Eastern hemlock mortality caused by the adelgid is a dramatic increase in inorganic N availability and nitrification rates, resulting in nitrate leaching in regions experiencing adelgid infestations (Jenkins et al. 1999; Kizlinski et al. 2002). Yorks et al. (2003) made similar observations when girdling trees to simulate an adelgid attack. Stadler et al. (2005) showed that infestations by adelgids increased the

32

presence of bacteria, yeast and filamentous fungi in the canopy, and strongly altered the chemistry, quantity and spatial pattern of throughfall. Tree mortality due to the adelgid may also modify forest floor microclimate (Cobb et al. 2006; Lishawa et al. 2007) and hydrologic processes (Ford and Vose 2007). Other invasive forest insects such as A. piceae, C. fagisuga (and its associated fungus N. coccinea var. faginata), E. abietinum and A. planipennis are probably responsible for serious changes in forest ecosystems because they kill important tree species on a large scale, but the precise impacts, as well as the processes underlying these impacts, are largely unknown.

Ecological effects due to predators and detritivores Effects on native animal populations Many studies show that invasive predatory insects displace native species, but most fail to identify the mechanism behind displacement (Reitz and Trumble 2002). This can be caused by extensive direct predation, exploitative competition for food or space, relative immunity from shared natural enemies or a disruptive mating system (see review in Snyder and Evans 2006). Many invasive predators are especially detrimental to related native predators. For example, H. axyridis, and the European seven-spotted ladybird, Coccinella septempunctata L., are both strongly suspected to displace other aphidophagous ladybirds in various agricultural environments in North America (e.g. Elliott et al. 1996; Brown and Miller 1998; Colunga-Garcia and Gage 1998; Michaud 2002; Evans 2004). But it is not clear whether the main mechanism of displacement is predation on native coccinellids or local depletion of aphids. Several laboratory experiments showed that larvae of H. axyridis and, to a lesser extent, C. septempunctata, are aggressive intraguild predators and will successfully prey on immature stages of most indigenous ladybirds (e.g. Burgio et al. 2002; Snyder et al. 2004; Yasuda et al. 2004; Ware and Majerus 2008). In contrast, there are few studies that investigate the more complex mechanism of displacement of native ladybirds by food depletion. Those who did include the factor of food availability in laboratory competition tests (e.g. Obrycki et al. 1998)

D.W. Langor & J. Sweeney (eds.)

failed to reach firm conclusions regarding the mechanisms involved in the displacement. Both invasive ladybirds have a particularly broad diet, allowing them to persist at sites where aphid density is low and purely aphidophagous species have left (Michaud 2002; Evans 2004). Invasive ladybirds are not only detrimental to other ladybirds. A recent study (Mizell 2007) showed that the invasion of H. axyridis also dramatically reduced other groups of aphid predators and parasitoids in pecan and crape myrtle. Among predatory insects, alien ants show the highest and best documented records of ecological damage on the native fauna. The most dramatic impacts of invasive ants occur when native ants are displaced through resource competition or direct predation. But other invertebrates or even vertebrates may also be displaced through the same mechanisms. Only some examples are given here. More can be found in Holway et al. (2002). The red imported fire ant, S. invicta, is probably the invasive insect which has received the most attention for its impact on native biodiversity. Originating from South America, it has invaded southern North America, where it threatens several arthropods, molluscs, reptiles, birds, amphibians, and mammals (e.g. Porter and Savignano 1990; Vinson 1997; Allen et al. 1997, 2000, 2001; Forys et al. 2001; Morrison 2002). It also attacks beneficial insects such as parasitoids and predators (Eubanks et al. 2002; Ness 2003). The argentine ant, L. humile, has invaded most continents and is known to displace native ants, other arthropods, birds, lizards and mammals through a variety of mechanisms such as predation, by competition for nesting sites or by tending arthropods and plants (e.g. Human and Gordon 1996, 1997; Laakkonen et al. 2001; Go´mez and Oliveras 2003; Carpintero et al. 2005; Suarez et al. 2005; Lach 2007, 2008). The crazy ant, Anoplolepis gracilipes (Jerdon), greatly reduces populations of red land crab on Christmas Island (O’Dowd et al. 2003), ants and other invertebrate species in Tokelau (Lester and Tavite 2004; Sarty et al. 2007), many invertebrates in the Seychelles (Hill et al. 2003; Gerlach 2004) and, in conjunction with the big-headed ant, Pheidole megacephala (F.), excludes native spiders in the genus Tetragnatha from native and disturbed forests in Hawaii (Gillespie and Reimer 1993). Pheidole megacephala is also introduced in northern Australia, where it displaces native ants and other invertebrates (Hoffmann et al.

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

1999; Hoffmann and Parr 2008) in Florida, where it may pose a threat to native fauna, including sea turtle and sea bird nestlings (Wetterer and O’Hara 2002), and in Mexico, where it has a negative effect on termite populations (Dejean et al. 2007). The little fire ant, Wasmannia auropunctata (Roger), and the tropical fire ant, Solenopsis geminata (F.), reduce the diversity and abundance of invertebrates, birds and reptiles in the Galapagos (Causton et al. 2006). Wasmannia auropunctata is also present in Central Africa, where it is known to displace native ants (Walker 2006) and in New Caledonia, where it has a negative effect on populations of native arthropods and lizards (Jourdan 1997; Jourdan et al. 2001). Finally, the crazy ant Paratrechina fulva (Mayr), introduced in Colombia for the control of leaf-cutting ants and poisonous snakes, is now threatening local biodiversity, in particular soil insects, snakes and lizards (de Zenner-Polania and Wilches 1992). In some cases, however, climatic requirements may limit the impact of invasive ants and other species. For example, after 150 or more years of residence in Madeira, P. megacephala and L. humile occupy only a small part of the island and appear to have little impact. Most of the island may be too cool for P. megacephala and too moist for L. humile, which are excluded by a dominant, better adapted native ant, Lasius grandis Forel (Wetterer et al. 2006). Although most of these impacts on the native fauna were investigated by comparing infested and uninfested areas, quite a few used experimental designs, for example by using poisonous baits, hot water or sticky barriers to exclude the invasive species (Allen et al. 2001; King and Tschinkel 2006; Lach 2007), food baits to assess the importance of resource competition (Sarty et al. 2007) or artificial nests to assess predation on birds (Suarez et al. 2005). Social wasps can also be particularly damaging for both native wasp species and other indigenous species. European wasps, Vespula germanica (F.) and V. vulgaris (L.), have invaded New Zealand beech forests where they prey on vulnerable native invertebrates and strongly compete with rare birds and invertebrates for food (Beggs 2001). Vespula germanica is also invasive in Australia, where it is suspected of outcompeting the native paper wasp Polistes humilis (F.) because of its broader diet (Kasper et al. 2004). Another paper wasp, Polistes versicolor (Olivier) is considered as highly invasive in the

33

Galapagos. It is estimated to prey on 17–154 g insects per ha per day, mainly Lepidoptera, therefore competing for food with finches and other arthropod predators (Parent 2000, in Causton et al. 2006). A congeneric species, the European Polistes dominulus (Christ) has invaded North America where it may be competing with the native Polistes fuscatus (F.). Although displacement has not yet been proven, several studies showed that the invasive species is competitively superior to the native species (Gamboa et al. 2002; Armstrong and Stamp 2003; Curtis et al. 2005). Invasive mosquitoes have the potential to affect populations of native mosquitoes by various mechanisms. However, the only well illustrated cases of displacement are between two invasive species, particularly Aedes albopictus (Skuse) displacing A. aegypti (L.) in various regions (see Juliano and Lounibos 2005, for review). For example Juliano (1998) investigated the competitive interactions between A. albopictus, recent invader in Florida, and A. aegypti, the previously resident invader, which had been shown to be displaced by A. albopictus in some environments and not others (Juliano et al. 2004). Juliano (1998) used field and lab experiments to test both the effects of shared parasites, Ascogregarina sp., and interspecific competition on mosquito communities. His results show that A. albopictus was a superior competitor in the absence of the parasite, and that its rate of parasite infection in the field was actually higher than that found in A. aegypti. The variation in the outcome of the interaction between the two mosquitoes was attributed to differences in habitat suitability. Juliano (1998) concluded that, when it occurred, direct competition was the mechanism driving the replacement of A. aegypti by A. albopictus. Displacements of native mosquito species have been less extensively studied. The invasive Culex quinquefasciatus Say may have displaced the native Culex tarsalis Coquillett in California through competition for resource and by degrading larval breeding sites (Smith et al. 1995). A laboratory experiment showed that C. quinquefasciatus displaces C. tarsalis in laboratory cages within a single generation (Smith et al. 1995). Carrieri et al. (2003) carried out laboratory experiments to investigate potential competitive interactions between A. albopictus and the native mosquito Culex pipiens L. in Italy. They found evidence that A. albopictus is superior in resource

34

competition with C. pipiens but, to date, the displacement of C. pipiens has not been demonstrated in the field. Similarly, although laboratory experiments consistently showed that the invasive A. albopictus was competitively superior to the native North American Ochlerotata triseriatus (Say), there is no evidence for decline of O. triseriatus in the field (Lounibos et al. 2001). In some cases, extensive research programmes on abundant invasive predators fail to show a significant effect on the native fauna, as for the European carabid beetle Pterostichus melanarius Illiger, invasive but apparently harmless for native carabid populations in North America (Niemela¨ and Spence 1991; Niemela¨ et al. 1997). Several calliphorid blow flies have been introduced from the Old World to the Americas. While some are strictly saprophagous species feeding mainly on carrion, at least two species, Chrysomya albiceps (Wiedemann) and C. rufifacies (Maquart), are facultative predators on other maggots. Both species are displacing native flies, in particular the American species Cochliomyia macellaria (F.), in both field and laboratory experiments (Wells and Greenberg 1992; Wells and Kurahashi 1997; Del Bianco Faria et al. 1999). Interestingly, laboratory experiments showed that Old World species having co-evolved with the predatory species are more resistant to predation than C. macellaria (Wells and Kurahashi 1997; Del Bianco Faria et al. 1999). Effects as vectors of animal diseases Several invasive mosquitoes are vectors of various animal and human diseases (Juliano and Lounibos 2005). Some of them have severe consequences for biodiversity. For example, invasive Culex spp. are responsible for the transmission of avian malaria that devastates endemic bird populations in Hawaii, particularly at low elevations (Van Riper 1991; Atkinson et al. 1995; Woodworth et al. 2005). In New Zealand, Tompkins and Gleeson (2006) found a correlation between the distribution of the invasive C. quinquefasciatus and the occurrence of avian malaria. Indirect effects on plant communities and ecosystem processes Predators having a significant effect on native species may also indirectly affect plant communities and

D.W. Langor & J. Sweeney (eds.)

ecosystem processes through cascading effects. Well described cascading effects by invasive predators are rare, except for some invasive ants. For example, the yellow crazy ant, A. gracilipes, has caused a rapid, catastrophic shift in the rain forest ecosystem of Christmas Island by greatly reducing populations of the red land crab, which is the main endemic consumer of the forest floor (O’Dowd et al. 2003). The displacement of crab populations results in slower litter breakdown, followed by a release of seedling recruitment and an increase in tree and shrub species richness. Furthermore, new associations between the alien ant and scale insects have led to tree dieback and changes in tree community composition. Similar observations were made in other parts of the world where A. gracilipes has been introduced, such as in the Seychelles, where it developed an association with scale insects resulting in tree mortality (Hill et al. 2003). The invasion of the argentine ant, L. humile, in many parts of the world has disturbed seed dispersal through the displacement of myrmecochorous ants (Christian 2001; Carney et al. 2003; Go´mez and Oliveras 2003; Go´mez et al. 2003; Witt et al. 2004). It is also known to reduce fruit-set and seed set of some native plants (Blancafort and Go´mez 2005). Lach (2007, 2008) observed that L. humile displaces floral arthropods, including pollinators, on various plants of the South African fynbos, but this decline had no detectable effect on seed sets. In addition, the argentine ant is strongly suspected to affect soil chemistry, turnover and erosion. Nest building and foraging activities of the red imported fire ant, S. invicta, affect physical and chemical soil properties and strongly enhances plant growth though the increase of NH4+ (Lafleur et al. 2005). In general, the importance of ecosystem process effects in invasive ants is largely unknown and deserves further studies (Folgarait 1998; Holway et al. 2002). The Chinese mantis, Tenodera sinensis (Sauss.), in North America provides another interesting case of a trophic cascade effect triggered by a generalist predator. The mantis preys on both herbivores and spiders, which feed on the same herbivores. The result is a net herbivore reduction increasing plant biomass (Moran et al. 1996; Moran and Hurd 1998). In other cases, effects on ecosystem processes are strongly suspected but not fully ascertained. For example, in New Zealand, European wasps are

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

thought to alter nutrient cycling in New Zealand beech forests by removing honeydew, which reduces the flow of carbon to microorganisms in the phyllosphere and the soil (Beggs 2001).

Ecological effects due to parasitoids and parasites About 2,000 arthropod species have been released in new regions for biological control purposes, the majority of them being parasitoids (van Lenteren et al. 2006). A small number of these parasitoids have been subsequently reared from non target species. In several cases an effect on native non-target hosts and native parasitoids has been either documented or suspected (see examples in Lynch and Thomas 2000; van Lenteren et al. 2006; Parry 2008). The earliest example is probably that of the tachinid, Bessa remota (Aldrich), released in Fiji in the 1920s, which is suspected to have caused the extinction of the questionably native target species, the coconut moth Levuana iridescens Bethune-Baker, but also of a non-target native moth, Heteropan dolens Druce (Tothill et al. 1930; Kuris 2003). However, assessing the effect of alien parasitoids on non-target hosts/preys and native parasitoids long after their introduction is a complicated task because, in most cases, the necessary quantitative data on native species populations before the introduction, or in non-invaded areas, are not available. A good example is the introduction of the tachinid fly Trigonospila brevifacies (Hardy) from Australia to New Zealand to control the tortricid moth Epiphyas postvittana (Walker). An extensive study on the parasitoid food web of Tortricidae in New Zealand showed that the tachinid has become the dominant parasitoid of many Tortricidae in broadleaf/podocarp forests in central North Island (Munro and Henderson 2002). Although the introduced parasitoid is suspected to affect both native tortricid populations and their parasitoids, the authors conclude that, ‘‘as no pre-release data on the composition of the parasitoid guild or the relative abundance of lepidopteran species were gathered before the release of T. brevifacies, it is difficult to determine the exact effect the tachinid has had on the native fauna in this system. Empirical studies of a simplified controlled host– parasitoid community would be required to determine

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if native parasitoid displacement were actually occurring’’. Another tachinid, C. concinnata, has been implicated in the decline of several species of native Lepidoptera since shortly after its release in North America in 1906 for control of the gypsy moth. Boettner et al. (2000) state that C. concinnata has significantly contributed to the decline of several native saturniid moths. They base their conclusion, firstly, on the fact that field exposures of saturniid caterpillars resulted in very high parasitism by C. concinnata and, secondly, by presenting arguments that alternative hypotheses for the decline were very unlikely. A similar study by Kellogg et al. (2003) demonstrated high rates of parasitism by C. concinnata on native luna moths, Actias luna (L.), in Virginia, but concluded that long-term studies would be needed to determine the actual effects of the parasitoid on luna moth populations. Life tables are sometimes used to assess the effect of alien natural enemies on native species. For example, Johnson et al. (2005) applied life table studies to show that the decline of the native Hawaiian koa bug, Coleotichus blackburniae White, was due to accidentally introduced egg predators rather than parasitoids introduced for biological control. Alternatively, models may be used to assess the role of invasive species in the decline of native fauna. Keeler et al. (2006) used a stochastic simulation model to assess the respective role of an alien parasitoid, an alien plant and the loss of native host plants in the decline of the native butterfly Pieris napi oleracea Harris in North America. The model showed that the role of the parasitoid was probably negligible compared to the two other factors. Similarly, Barlow et al. (2004) and Barron (2007) modelled the impact of two introduced parasitoids in New Zealand on non-target hosts, using the intrinsic rate of host increase, the average abundance of the host in the presence of parasitism and the estimated mortality caused by the parasitoid. Barlow et al. (2004) predicted that the introduction of the alien braconid Microctonus aethiopoides Loan would decrease populations of some native weevil species by 8–30%. The models developed by Barron (2007) showed that the introduced pteromalid parasitoid, Pteromalus puparum (L.), is probably not responsible for the decline of populations of the endemic red admiral butterfly Bassaris gonerilla (F.).

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Accidentally introduced parasitoids may also cause ecological damage. For example, the ichneumonid Echthromorpha intricatoria (F.) is suspected to be partly responsible for the decline of B. gonerilla in New Zealand (Barron et al. 2004). Kenis et al. (2007) suspect that many parasitoids that are thought of as occurring on several continents may have been more or less recently introduced accidentally with their host or host plant. Several of these may cause undetected hazards on new hosts. It has often been suggested that introduced parasitoids may also displace native parasitoids by competition (Bennett 1993), but reliable examples are rare. The Nearctic aphid parasitoid Lysiphlebus testaceipes (Cresson), introduced in the Mediterranean region to control Aphis spiraecola Patch, has become a dominant parasitoid of other aphid species, including Toxoptera aurantii (Boyer de Fonscolombe), in which it may have displaced two congeneric parasitoid species, L. fabarum (Marshall) and L. confuses Tremblay & Eady (Tremblay 1984). Similarly, Schellhorn et al. (2002) provide evidence that the exotic braconid parasitoid Aphidius ervi (Haliday), introduced into North America to control the pea aphid Acyrthosiphon pisum Harris, has caused the decline of the native Praon pequadorum Viereck. Another example is the probable displacement of Encarsia margaritiventris (Mercet) as dominant parasitoid of the viburnum whitefly, Aleurotuba jelineki (Frauen.), in Italy, following the introduction of the exotic parasitoid Cales noaki Howard (Viggiani 1994). A recent study by Parry (2008) suggests that Compsilura concinnata may be involved in the apparent disappearance from areas of New England of Lespesia frenchii (Williston), a native polyphagous tachinid competing for the same lepidopteran hosts. Other potential cases are listed and discussed in Lynch and Thomas (2000) and van Lenteren et al. (2006). The small hive beetle, Aethina tumida Murray, a nest parasite/scavenger native to Sub-Saharan Africa has invaded North America and Australia, where it parasitizes domesticated honey bees. It also attacks and develops on bumble bees, and there is growing concern that may affect populations of native pollinators (Hoffmann et al. 2008). Some exotic ectoparasites are considered as ecological pests because of the damage on vertebrates. In the Galapagos, a parasitic fly, Philornis downsi

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Dodge & Aitken, significantly decreases fledging success of finches by infesting and killing juvenile birds (Fessl et al. 2006). A chewing louse, Damalina (Cervicola), is suspected to cause hair-loss syndrome in black-tail deer in North America, although firm evidence is still lacking (Bildfell et al. 2004).

Ecological effects due to invasive pollinators Invasive pollinators, in addition to causing hazards through hybridization (see section above), may also compete with native pollinators for floral resources and nesting sites. Other undesirable effects include co-introduction of natural enemies, inadequate pollination of native flora or undesirable pollination of exotic flora (Goulson 2003; Goulson et al. 2008). The honeybee, A. mellifera, has been widely introduced in many regions for pollination and honey production (Moritz et al. 2005). Although its introduction is often considered positive, various detrimental effects have been investigated and reported. In particular, it has been often reported to cause a decline in native bee and bird species, particularly on islands (e.g. Roubik 1978; Kato et al. 1999; Hansen et al. 2002; Dupont et al. 2003). Another important invasive pollinator, the European bumblebee, Bombus terrestris (L.), displaces native megachilid bees in Tasmania, where it is also suspected to have a negative effect on plant pollination (Hingston and McQuillan 1999). In addition, it has been shown that B. terrestris has superior reproductive rate than native Japanese bumblebees and there is serious risk of outcompetition since there is overlap in forage use and time of foraging (Matsumura et al. 2004; Inari et al. 2005). Non native pollinators can be vectors of pathogens which can threat native pollinators. An interesting case is the spread of bumblebee-specific pathogens (Critihidia bombi Lipa and Triggiani and Nosema bombi Fantham & Porter) and tracheal mite (Locustacarus buchneri Stammer) due to trafficking of commercial bumblebee colonies. Pathogen spillover from commercial bumblebee colonies can potentially have a devastating effect on native bumblebee populations (Colla et al. 2006). Indeed, Colla et al. (2006) have shown that commercial colonies have greater parasite load than wild colonies and that pathogen loads in wild bumblebee

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populations near commercial greenhouses are significantly increased. Introduced bees are also known to reduce fitness of some native plant species (Roubik 1996; Gross and Mackay 1998). On the other hand, they may enhance pollination and, consequently, invasiveness of exotic weeds, as shown by Barthell et al. (2001) for Centaurea solstitialis L. in North America and Stout et al. (2002) for Lupinus arboreus Sims in Tasmania. Another interesting example is provided by fig wasps. Many fig species are dependent on highly specific fig wasps (Agaonidae) for pollination, and without them the fig tree will bear no seeds. Three exotic fig trees, Ficus microcarpa L., F. benghalensis L., and F. altissima Blume grown in Florida gardens for over a century only started spreading and became invasive in the 1980s, when their fig wasp pollinators arrived (Nadel et al. 1992). It must be noted that no single experiment has clearly demonstrated long-term reductions in populations of native organisms following the introduction of exotic pollinators (Goulson 2003; Moritz et al. 2005). As stated by Goulson (2003) this probably reflects more the difficulty of carrying out convincing competition studies rather than a true absence of competitive effects. Whereas most studies on the ecological effect of introduced pollinators rely on correlational data or other indirect measures, two recent investigations use experimental approaches. Kenta et al. (2007) tested, in a greenhouse experiment, the potential disturbance caused by the introduction of the European bumblebee on native plant–pollinator interactions. They concluded that the alien bumblebee can disturb pollination on a plant even when only representing a small fraction of the total pollinator community. Thomson (2004) provides the first experimental demonstration of the negative effects of non native honey bees on native bumblebees. She experimentally introduced honey bees and found that proximity to hives significantly reduced the foraging rates and reproductive success of Bombus occidentalis Greene colonies. In addition, Thomson (2006) found significant niche overlap between foraging preferences of native bumble bees and introduced honey bees, which is maximum at the end of the season when floral resources are more limited. This indicates that both native bumblebees and introduced honey bees largely rely on the same restricted suite of plant species.

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Conclusions and future research This review has shown that invasive alien insects can affect native species and ecosystems through a variety of mechanisms. A surprisingly high number of primary research publications (403) were found that describe or investigate the ecological effect of invasive alien insects. Nevertheless, these studies concern only 72 species, and a clear impact in field conditions has been ascertained for only 54 of them. This represents a very low proportion of the alien insects in the world. For example, 311 alien insect species are established in Switzerland (Kenis 2006), more than 2,000 alien arthropods are found in the Continental USA and more than 2,500 in Hawaii (Pimentel 2002). It is not clear whether the low proportion of alien insects known to have an effect on biodiversity reflects a lack of effect or a lack of investigations. The vast majority of studies found during our survey ([80%) reported a significant effect, suggesting that more investigations would reveal more impacts. However, it may also be partly due to a publication bias towards studies showing significant results. Other important biases are observed towards species and ecosystems that are also considered important for the economy or public health, as illustrated by the high number of studies investigating the impact of alien ants, honey bees, plant pests or mosquitoes. Many studies showing effects on native insect biodiversity focused on groups that are considered as more ‘‘attractive’’ for the public and the researchers, e.g. butterflies or ladybirds, although there is no scientific reason to believe that, for example, aphids or flies are less affected by invasive species. The vast majority of studies on the effect of alien herbivorous insects have focused on forest pests and their damage on trees and forest ecosystems, probably because their effect is more visible and concerns keystone species of forest ecosystems. However, many other invasive herbivores would deserve more attention for their potential effect on indigenous plants. For example, in eastern North America, dozens of studies have investigated the impact of alien forest pests such as L. dispar, A. piceae and A. tsugae, whereas none has focused on the ecological effects of the lily leaf beetle, Lilioceris lilii (Scopoli), and the viburnum leaf beetle, Pyrrhalta

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viburni (Paykull), two species that may seriously threaten the survival of wild lilies (Lilium spp.) and Viburnum spp. in the same region (Ernst et al. 2007; Weston et al. 2007). Similarly, many investigations have focused on the effect of insects released for biological control because of the follow-up studies carried out in biological control programmes and because of the particular interest of conservation ecologists for the non-target effect of alien biological control agents (Louda et al. 2005; van Lenteren et al. 2006). The Asian ladybird, H. axyridis, a biological control agent that invaded North America and Europe is presently the target of extensive studies on its potential impact on native ladybirds (Burgio et al. 2002; Michaud 2002; Snyder et al. 2004; Yasuda et al. 2004; Ware and Majerus 2008) whereas the impact of many alien predators accidentally introduced in the same continents remains totally unexplored. Furthermore, all examples of ecological impact cited here concern terrestrial ecosystems, with the exception of mosquitoes, but freshwater ecosystems are probably not immune. Insects as effective biological control agents of alien water weeds clearly can have a profound effect on freshwater plant populations and hence ecosystem functioning (Mbati and Neuenschwander 2005). Thus, there are clear and important gaps in our knowledge of the effect of alien insects on native biodiversity and ecosystems, particularly in species and habitats that are of lower importance for the economy and the general public. Nevertheless, these gaps also represent exciting opportunities for further research and the remarkable increase in the number of studies on the ecological impact of invasive insects shows that these opportunities are presently being taken. Examples of effects on species populations and communities are far more numerous than those on ecosystem processes, an observation also made by Parker et al. (1999) for several groups of invasive species. In contrast, Levine et al. (2003) found that roughly equal numbers of studies on invasive plants examined effects on species and communities and effects on ecosystem processes. In general, effects of invasive species on native species populations are more easily observed, i.e. through comparative studies between invaded and non-invaded areas, or between conditions before and after invasion.

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However, many of these observations are rather anecdotal or quantified at very local scales only, and the mechanisms by which the impacts arise are often not clearly understood. Furthermore, making assessments of invasion impact on the basis of temporal or spatial correlations may be misleading (Thomson 2006). To fully assess and understand the ecological effects of an invasive species and the mechanisms behind variations in populations observed, or not, in the field, experimental approaches are needed, preferably under field conditions. This is particularly true for effects occurring at the same trophic level, i.e., the displacement of an herbivore by another herbivore or a predator by another predator. In cases such as these, the mechanisms underlying the impact are often indirect and complex. For example, until now the numerous field observations and laboratory experiments carried out to assess the effect H. axyridis on native ladybirds have failed to understand whether displacement is due to direct intraguild predation or through resource competition. The question may only be answered by field experiments involving the manipulation of prey and ladybird densities and analysis of gut contents. Direct effects of invasive insects on lower trophic levels through herbivory, predation and parasitism are easier to assess, at least at a local scale. Evaluating the effect on native species at a regional scale is often more complicated because it has to take into account the geographic, ecological and temporal variability throughout the distribution range. The effect of an invasive insect is known to vary with time, space and system, and, understandably, studies tend to focus on sites and systems where the impact is most likely. However, it would be of utmost importance to conduct parallel studies in systems where impacts are suspected to be lower. This would improve assessments of the regional importance of the invasive species, as well as increasing our understanding of the mechanisms underlying impacts. Finally it could also allow us to understand how native ecosystems and communities resist the impact of invaders, which could be the key for restoring invasion-resistant ecosystems and for developing control strategies (Levine et al. 2003). Investigations on ecosystem effects, albeit less numerous than population and community effects, are often of good quality because they require longer studies based on a priori hypotheses on impact

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

processes. Nevertheless, as pointed out by Levine et al. (2003) for plant invasions, the consequences of alterations in ecosystem processes for species populations and community structure are poorly explored. For example, invasive forest herbivores such as L. dispar and A. tsugae are known to alter nitrogen cycling in the invaded forests, but how these changes affect plant and animal communities remains unclear. Acknowledgments We thank Ecki Brockerhoff, Wolfgang Rabitsch and an anonymous reviewer for their useful comments on the manuscript, and Jon Sweeney and Dave Langor for having invited us to participate in this special issue. We also thank all the colleagues who provided assistance in literature search and acknowledge support from the European Commission Framework Programme 6 via the Integrated Project ALARM (Assessing LArge scale environmental Risks for biodiversity with tested Methods; GOCE-CT-2003-506675; www.alarmproject.net; Settele et al. (2005)).

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Yara K, Sasawaki T, Kunimi Y (2007) Displacement of Torymus beneficus (Hymenoptera: Torymidae) by T. sinensis, an indigenous and introduced parasitoid of the chestnut gall wasp, Dryocosmus kuriphilus (Hymenoptera: Cynipidae), in Japanese chestnut fields: possible involvement in hybridization. Biol Control 42:148–154 Yasuda H, Evans EW, Kajita Y, Urakawa K, Takizawa T (2004) Asymmetric larval interactions between introduced and indigenous ladybirds in North America. Oecologia 141:722–731 Yorks TE, Leopold DJ, Raynal DJ (2003) Effects of Tsuga canadensis mortality on soil water chemistry and understory vegetation: possible consequences of an invasive insect herbivore. Can J For Res 33:1525–1537

Beyond Pandora’s Box: quantitatively evaluating non-target effects of parasitoids in classical biological control Dylan Parry

Originally published in the Journal Biological Invasions, Volume 11, No 1, 47–58. DOI: 10.1007/s10530-008-9319-x Ó Springer Science+Business Media B.V. 2008

Abstract A seminal paper by Howarth (Proc Hawaii Entomol Soc 24:239–244, 1983) entitled ‘‘Classical biological control: Panacea or Pandora’s Box’’ ignited a sometimes acrimonious debate over the relative safety of introductions for classical biological control. Extolled for years as environmentally benign, the litany of negative non-target effects profiled by Howarth heightened awareness of this issue. Several factors have muddied this debate including the conflation of frequency of effects with their strength, grouping the effects of disparate biological control agents together, and the lack of quantitative data on either side of the argument. Here, I examine the potential for non-target effects among insect parasitoids, the most common group used for biological control of arthropods. In response to calls for better quantitative studies, I highlight three different techniques, quantitative food webs, life table analysis, and experimental populations, respectively, to quantitatively assess or reassess non-target effects in different systems. I also explore three methodological approaches employed to ascertain the strength of competitive interactions between native and introduced parasitoids, a potential non-target effect that has received little attention in the

D. Parry (&) College of Environmental Science & Forestry, State University of New York, 246 Illick Hall, 1 Forestry Drive, Syracuse, NY 13082, USA e-mail: [email protected]

literature. These types of studies may greatly increase our understanding of the nature of non-target interactions with introduced parasitoids and bring more rigor to a debate often dominated by rhetoric. Keywords Parasitoid drift
Non-target
Competition
Biological control
Invasive
Native
Lepidoptera Abbreviation CBC Classic biological control

Introduction Classic biological control (hereafter CBC), the deliberate introduction of the natural enemies of a non-indigenous species, was long considered the most environmentally desirable method for controlling the proliferation of exotic organisms (McEvoy and Coombs 2000). Francis Howarth (1983) published Panacea or Pandora’s Box, a paper that provocatively challenged the long-standing paradigm of biological control as a safe alternative to chemical pesticides. His article documented a litany of nontarget effects and unintended consequences of organisms released in CBC programs. The paper launched a heated and sometimes acrimonious debate often dominated by rhetoric rather than science. Despite bold pronouncements extolling the safety of CBC (e.g., Funasaki et al. 1988; Lai 1988; DeLoach 1991;

D.W. Langor & J. Sweeney (eds.), Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems. DOI: 10.1007/978-1-4020-9680-8_4

47

48

Frank 1998; Headrick and Goeden 2001), or its catastrophic consequences (Lockwood 1993; Howarth 1991) quantitative studies validating either viewpoint are few (Thomas and Willis 1998). There is little debate that some species introduced for CBC have had or are having detrimental impacts on non-target native species (Simberloff and Stiling 1996; Follett and Duan 2000; Strong and Pemberton 2000). Certainly, some early CBC programs, especially those where vertebrate predators were introduced to islands, had devastating consequences for non-target organisms (Smith and Remington 1996; Simberloff and Stiling 1996). Modern practitioners of CBC point out that none of these introductions would be considered today, while critics rejoin that some organisms introduced within the last 30 years have also had undesirable effects on non-target species despite extensive screening (Strong and Pemberton 2000; Louda and Stiling 2004). While less polarized in recent years, the debate about the risk posed by CBC continues (see Hoddle 2004 and Louda and Stiling 2004). The arguments about the safety of CBC and its non-target effects may be confounded by sweeping generalizations across diverse groups of organisms. For example, while the introduction of mammalian predators and freshwater fish often results in severe non-target consequences (Simberloff and Stiling 1996), fewer non-target effects have been documented for the biological control of plants and even fewer for efforts aimed at arthropods. However, nontarget effects from CBC programs against arthropods may be greater than has been recognized because only a few systems have been extensively studied (e.g., Lynch and Thomas 2000). To move beyond polarization, we need finer scale analyses that scrutinize specific groups of biological control organisms. Secondly, we must evaluate the effects on nontarget species in a more quantitative manner using realistic in situ field studies.

Parasitoids in classical biological control Insect parasitoids, primarily from the orders Diptera and Hymenoptera, are the most common organisms used for the CBC of other arthropods. In the Hymenoptera, several large super-families are exclusively parasitic or nearly so (Pennacchio and Strand

D.W. Langor & J. Sweeney (eds.)

2006), whereas the majority of dipteran parasitoids are in the family Tachinidae (Stireman et al. 2006), although this life-history strategy does occur in numerous other dipteran families. Parasitoid biology Parasitoids can be broadly divided into idiobionts and koinobionts based on life-history (Hawkins 1994). Idiobiont parasitoids kill their host at or shortly after parasitism, then feed externally on the host, a life history strategy that precludes interaction with the intact immune system of the host. Koinobionts on the other hand, complete most or all of their larval development in a living host, necessitating evasion or manipulation of host defenses in order to survive and complete development. This dichotomy has led to speculation that idiobionts should have larger host ranges than koinobionts although the evidence is equivocal (Hawkins and Marino 1997). Parasitoid drift The acquisition of novel non-target hosts by introduced parasitoids or ‘drift’ incorporates several processes, including host shifts, host range expansion, and host switching (Follett et al. 2000). Studies suggest that parasitoid drift may be extensive. For example, of 313 exotic parasitoids released in North America for CBC of holometabolous insects prior to 1990, 51 were recorded from native non-target insects (Hawkins and Marino 1997). While the acquisition of non-target species by so many introduced parasitoids is alarming, it is not particularly surprising. Few parasitoids are truly monophagous (Shaw 1994; Sheehan and Hawkins 1991) and drift has long been recognized (e.g., Webber and Schaffner 1926; Nishida 1956; Davis 1964; Duan et al. 1996; Follett et al. 2000). In fact, the expansion of host range by introduced parasitoids through the acquisition of native species was historically viewed as beneficial (e.g., Culver 1919; Webber and Schaffner 1926). Until recently, parasitoids introduced for CBC were thought to have little or no effect on populations of most native species (Coulson et al. 1991; Godfray 1994; Onstad and McManus 1996). The debate over non-target utilization by parasitoids has been significantly hampered by the conflation of frequency of occurrence with the

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

strength of their effects. Most reviews have concentrated on the former, which is relatively easy to assess. However, estimating the frequency of nontarget utilization without assessing impact on their population density may lead to exaggerated claims of risk or, conversely, underestimate their importance. While dramatic effects on non-target species and even extinctions have been proposed (e.g., Gagne and Howarth 1985), these claims have often not withstood scrutiny (Lynch and Thomas 2000). Despite increased attention to this issue, the greatest impediment to understanding non-target issues is the absence of quantitative data on the actual impact or strength of the effect. With few exceptions, nontarget species were not routinely surveyed, and we often have no knowledge about their population dynamics prior to introductions of biological control agents. Thus, implicating an introduced parasitoid in the decline of a native insect will always be difficult and has been shown with confidence in only a few cases.

Quantitatively evaluating non-target effects As the debate over the relative safety of parasitoids in CBC has matured, some researchers have attempted to bring a more quantitative approach to assessing non-target effects. Lynch and Thomas (2000) developed a hierarchical ranking for non-target effects that would allow comparison among different systems. Using a severity index, they ranked different systems from 0 to 9 based on the known effects on non-target species. The most severe non-target effects attributable to a parasitoid were those of Bessa remota (Ald.), a tachinid introduced to control a native pest of coconut, Levuana iridescens Betheune-Baker (Zygaenidae), which it may have ultimately extirpated (Table 1) along with other native lepidopterans (Kuris 2003). While one may quibble with their criteria, and the likelihood that data are inherently too incomplete for accurate ranking, the authors should be commended for advancing a methodology for assessing non-target impacts. Several recent studies have experimentally investigated non-target effects using more quantitatively rigorous methods than in the past. While this is certainly not an exhaustive list of potential experimental approaches, I illustrate three different

49

methodologies used in these studies to assess the non-target effects of exotic parasitoids and resolve longstanding issues about the nature and consequences of parasitoid drift. Quantitative food webs As early as 1958, the loss of lepidopteran biodiversity in Hawaii had been attributed to parasitism by exotic biological control agents (Gagne and Howarth 1985; Howarth 1983). While some have argued that evidence for this outcome is scant (Lynch and Thomas 2000), there is no doubt that Hawaii has been the recipient of a large number of CBC introductions (Funasaki et al. 1988; Messing and Wright 2006) and that a significant number of these attack Lepidoptera. To quantitatively assess the effects of introduced parasitoids on native Hawaiian Lepidoptera, Henneman and Memmott (2001) focused on community interactions among plants, lepidopteran herbivores, and their parasitoids. The study was conducted in the remote Alakai swamp, a relatively pristine Hawaiian forest. In the study, 58 species of Lepidoptera were systematically collected and reared from 60 plant species. The parasitoids reared from the collection were used to construct quantitative food webs (e.g., Memmott 2000). Despite the remote forested location, 83% of the parasitoids recovered were species introduced for biological control of primarily agricultural pests. An additional 14% were exotics introduced accidentally and only 3% of the complex was represented by native species. This type of study does not lend itself to directly assessing the role of parasitism in the population dynamics of a particular species. The parasitism rates recorded in several species (*25%) could have significant impact at the population level, but determination would require more detailed study of individual species. Trophic web studies are an excellent starting point for investigating non-target effects of parasitoids in the context of community structure. Life table analysis The decline of the koa bug, Coleotichus blackburniae White (Scutelleriidae), an endemic Hawaiian heteropteran, has been used as the poster-child for parasitoid drift by numerous authors including

50 Table 1 The severity ranking of selected nontarget effects caused by parasitoids introduced for biological control (adapted from Lynch and Thomas 2000)

The ranking (0–9, with 9 being most severe) was based on a severity index that incorporates varying degrees of mortality, population reduction, and extinction

D.W. Langor & J. Sweeney (eds.)

Introduced parasitoid

Family

Region/Country

Number of non-target species affected

Severity

Bessa remota

Tachinidae

Fiji

1

7

Cotesia flavipes

Braconidae

Mexico & Brazil

1

6

Aphytis holoxanthus

Aphelinidae

Florida, Texas & Brazil

2

6

Copidosoma floridanum

Encyrtidae

Several

5

Tetrastichus dryi

Eulophidae

New Zealand La Re´union

1

5

Cales noacki

Aphelinidae

Italy

4

5

Trigonospila brevifacies

Tachinidae

New Zealand

1+

5

Compsilura concinnata

Tachinidae

New England, USA

Several

3

Microctonus aethiopoides

Braconidae

New Zealand

3

3

Pteromalus puparum

Pteromalidae

New Zealand

1

3

Trichopoda pilipes

Tachinidae

Hawaii

1

3

Trissolcus basalis

Scelionidae

Hawaii

1

3

Howarth himself (Howarth 1983, 1991). This insect is attacked by several species of introduced parasitoids, most notably Trichopoda pilipes (F.), a tachinid originally introduced to control the southern green stink bug, Nezara viridula (L.) (Pentatomidae), an exotic agricultural pest. Post-release museum collections indicated that more than 17% of koa bug specimens had T. pilipes eggs attached to their bodies (Follett et al. 2000), although sample sizes were small and possibly biased. Thus, speculation about non-target effects on koa bugs is based on limited qualitative data. To determine the actual effect of T. pilipes on koa bugs relative to other sources of mortality, Johnson et al. (2005) used life table analysis. They determined that, although T. pilipes responded to koa bugs in a density-dependent fashion and could extirpate localized populations, it appeared to play a minor role in koa bug mortality at many sites. In this case, quantitative assessment of parasitism suggests that the general decline of the koa bug is unlikely to have been caused by T. pilipes alone, and that general predators and habitat loss should also be considered. Experimental populations Populations of several species of native giant silk moth (Saturniidae) have declined or been extirpated in parts of Northeastern North America (Schweitzer 1988; Tuskes et al. 1996). Several hypotheses have been advanced to explain this decline, but Boettner et al. (2000) proposed that an exotic parasitoid

introduced for control of gypsy moth a century ago may be at least partially responsible. Borrowing a technique used in studies of forest pest insects (e.g., Gould et al. 1990; Parry et al. 1997), Boettner et al. (2000) established experimental populations of two native giant silk moths, Hyalophora cecropia (L.) and Callosamia promethea (Drury) (Saturniidae), in western Massachusetts. Laboratory reared sentinel larvae were deployed in the field from 6 to 8 days before being retrieved and reared in captivity. In turn, these larvae were replaced in the field by another cohort at the same stage of development as the ones collected. Mortality within each time interval is then combined to estimate cumulative parasitism (Boettner et al. 2000; Kellogg et al. 2003; Elkinton et al. 2006). This sequential rear, release, and retrieve technique is very labor intensive but can be a powerful tool in estimating the impact of parasitoids on populations. The dominant source of mortality in the experimental silk moth populations was parasitism from the exotic tachinid, Compsilura concinnata Meig., confirming the authors’ original hypothesis. First released in 1906, C. concinnata is abundant and widely distributed throughout eastern and mid-western North American forests. In Massachusetts, the staggeringly high rates of parasitism (69–81%) by this tachinid in cecropia (H. cecropia) and promethea (C. promethea provides a plausible explanation for the well-documented decline in saturniids in this region. Subsequent studies have validated these results for some species, but have also indicated that C. concinnata may not be

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

the only or even the most important reason for regionwide declines in other saturniid populations. For example, Kellogg et al. (2003) conducted similar research in Virginia using the saturniids Actias luna (L.), H. cecropia and C. promethea. Although mortality from C. concinnata was high (80%) in first generation A. luna, parasitism was much lower in the second generation (20%) as well as in the other species studied. However, Kellogg et al. (2003) found native hyperparasitoids in the majority of C. concinnata recovered in Virginia, which may be limiting the impact of this parasitoid in some regions. A comprehensive study by Selfridge et al. (2007) found much lower parasitism (\10%) of Hemileuca maia (Drury) on Cape Cod, MA, than described from a small sample of this species elsewhere on the Cape (36%, Boettner et al. 2000) and from the closely related H. lucina in mainland Massachusetts (26–53%, Stamp and Bowers 1990). Comparison among these studies is difficult because of differences in methodology, density, species used, and experimental habitats. To address this issue, I have conducted experiments in New York and Massachusetts using several species of saturniids and compared parasitism by C. concinnata (Fig. 1). The results suggest that for two of the species, C. promethia and Antheraea polyphemus (Cramer),

C. concinnata has significantly greater impact in Massachusetts than it does in New York. Two other saturniids, Anisota senatoria (J.E. Smith) and H. maia, appear to be little affected by C. concinnata in either region. These studies suggest that C. concinnata has had and continues to have significant impacts on non-target species, but reveals large differences over a regional scale that will not be easy to explain. Some saturniids, especially in the sub-family Ceratocampinae, have declined (e.g., Schweitzer 1988) in the apparent absence of high mortality from C. concinnata, although quantitative studies validating this observation are lacking.

Competitive interactions between introduced and native parasitoids The focus of research on non-target effects of parasitoids released for biological control has been on their potential to attack indigenous host species. Relatively little attention has been paid to possible detrimental effects to native parasitoids with which they may directly or indirectly compete. Many native insect herbivores are associated with a community of indigenous parasitoids made up of both specialist and generalist species (Schaffner and Griswold 1934;

100

Massachusetts New York

80

Parasitism (%)

Fig. 1 Parasitism (mean ± SE) of four species of giant silk moth (Saturniidae) in Massachusetts and New York by Compsilura concinnata. Data for A. polyphemus from D. Parry, G. Boettner, and G. Tuttle (unpublished manuscript), C. promethea from Boettner et al. (2000) and Parry (2007), H. maia from Selfridge et al. (2007) and B. Hoven (unpublished), and A. senatoria, D. Parry (unpublished data). The number of sentinel larvae deployed for each species within a region are provided on each bar

51

60

40

n= 560

n= 378

n= 286

n= 260

n= 409

20

n= 320

n= 2720

n= 2962

0 Ant. polyphemus

C. promethea

Ani. senatoria

Species

H. maia

52

Hochberg and Hawkins 1992; Hawkins and Mills 1996; Marino et al. 2006). The insertion of an exotic parasitoid into these communities will inevitably change host-parasitoid relationships. One outcome may be competitive displacement, as has been seen with introduced arthropod predators such as lady beetles (e.g., Evans 2004). Such effects may be confined to one or a few native species in many systems, but for highly polyphagous species such as C. concinnata, negative interactions could conceivably extend to hundreds of native parasitoids. The role of competition in structuring parasitoid communities has long been debated (Hawkins 2000). Competitive exclusion has been shown for hymenopteran parasitoids introduced for CBC, although the excluded species have been other introduced parasitoids rather than native species (Luck and Podoler 1985; Bennett 1993; Murdoch et al. 1996). To my knowledge, the extirpation of a native parasitoid by an introduced parasitoid has not been documented. Determining that competition has occurred is difficult because studies of native parasitoid fauna are rarely conducted prior to introduction of species for CBC, thus limiting the utility of any post-hoc analyses (Bennett 1993). Even where exotic species have become a significant component of a host’s parasitoid load, determining if they are simply exploiting an unoccupied niche or if they have out-competed and replaced native species is difficult. Despite these hurdles, there are cases where the preponderance of evidence suggests that exotic parasitoids have displaced native parasitoids. Competition among parasitoids can take several forms, including hyper- and klepto-parasitism (Mills 2003). For the purposes of this paper, I will restrict discussion to exploitative and interference competition among obligate parasitoids. In exploitative competition, an exotic parasitoid reduces the density of native host species required by indigenous parasitoids. Interference competition results when one species uses physical combat, physiological suppression, or more rapid development to usurp host resources from a competitor (Mills 2003). Priority effects can be important in governing the outcome of competition among parasitoids, with the first arriving parasitoid species often intrinsically superior to later arriving species (Reitz 1996). One of the first studies to examine competition between introduced and native parasitoids was

D.W. Langor & J. Sweeney (eds.)

conducted in New Zealand. Following the introduction of the tachinid Trigonospila brevifacies (Hardy) in the 1960s for control of tortricid pests of fruit trees, researchers were surprised to find that the species adopted multiple species of native tortricids as hosts even though they occurred in natural forests rather than orchard environments. Using connectance and quantitative food webs, Munro and Henderson (2002) identified the host species utilized by T. brevifacies, quantified their parasitoid load, and attempted to determine if this species has become dominant in native hosts at the expense of indigenous parasitoids. Remarkably, within 30 years of introduction, T. brevifacies had become the dominant parasitoid of all but one of the tortricid species, and contributed from 16% to 80% of the total parasitism, the most of any parasitoid in the system. The host range of T. brevifacies overlaps in part or completely with that of 13 native parasitoids. The numerical dominance and degree of overlap with native parasitoids strongly suggests that this species is displacing native parasitoids of tortricids, although refugia in contiguous forest may prevent their complete loss. Perhaps our most detailed understanding of competitive interactions between introduced and native parasitoids comes from research on the braconid Aphidius ervi (Haliday). This wasp was introduced to the continental United States in 1959 to control the pea aphid, Acyrthosiphon pisum Harris (Homoptera: Aphidae), an important pest of a variety of leguminous crops (Angalet and Fuester 1977). This parasitoid is a good competitor and had displaced another introduced braconid, A. smithi (Sharma and Subba Rao) by the 1970s (Reitz and Trumble 2002). Until the introduction of A. ervi, the dominant parasitoid of pea aphids was the native Praon pequadorum Viereck (Braconidae), which accounted for as much as 42% of the parasitism (Schellhorn et al. 2002). The percentage of parasitism attributable to A. ervi is now 90–100% and parasitism by P. pequadorum has fallen to very low levels (Danyk 1993). To determine if the decline of the native P. pequadorum was caused by the exotic A. ervi, Schellhorn et al. (2002) coupled experimental manipulation with modeling. Competition was first addressed using population level manipulations of host density and subsequently by quantifying foraging ability on individual plants. In an apparent paradox, P. pequadorum was superior to A. ervi in larval–larval

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

competition within a host, an effect most pronounced when aphid populations were manipulated to create high parasitism levels by both braconids. This conundrum was resolved when the foraging efficiency of each braconid species was examined. The introduced species, A. ervi, was significantly more efficient and parasitized more aphids per unit time than its native competitor. The model developed by Schellhorn et al. (2002) showed that the superior foraging ability of A. ervi could overcome the superiority of P. pequadorum as a larval competitor. Anthropogenic factors tilt the scales further toward A. ervi because the harvesting practices used in crops like alfalfa create frequent, catastrophic disturbance in this host-parasitoid system. In the low density aphid populations following harvest, the superior foraging abilities of A. ervi enable it to dominate its native competitor. Identifying a mechanistic basis for competitive effects between native and introduced parasitoids as Schellhorn et al. (2002) did in an agroecosystem will be more challenging in unmanipulated natural environments. In many systems, parasitoids were released without any knowledge of existing host–parasitoid relationships. There are, however, a few natural systems where sufficient data were recorded to evaluate the effects of an introduced parasitoid on native parasitoid complexes. In eastern North America, we are fortunate to have information quantifying the native parasitoid complexes of Lepidoptera shortly after the introduction of C. concinnata. Coupling this historical data with contemporary field studies has allowed insight into the effects of a dominant introduced parasitoid on native competitors. The prevalence of C. concinnata in forest lepidopteran communities in Northeastern North America suggests that this species has had a long and extensive interaction with native parasitoids. In an extraordinarily prescient study, Schaffner and Griswold (1934) undertook one of the most ambitious systematic collections of hosts and parasitoids ever conducted, with more than 300,000 individual Lepidoptera from 164 species reared from 1915 to 1929. Most importantly, the project encompassed a 15-year period following the release of C. concinnata in New England. This study serves as a benchmark that can be used to retroactively assess the effects of this introduction on native parasitoid communities. Given the wealth of data in Schaffner and Griswold’s (1934) landmark compendium, it is surprising

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that few ecologists have taken advantage of it. To illustrate the utility of this data set in retrospective studies, I examined changes in the abundance of a native generalist tachinid, Lespesia frenchii (Williston), one of the most polyphagous native tachinids attacking forest Lepidoptera in North America (Arnaud 1978). This species was historically prevalent in several saturniids that are now heavily parasitized by C. concinnata and was not recovered from 2000 saturniid sentinel larvae deployed in central Massachusetts (Boettner et al. 2000). To assess the abundance of L. frenchii and C. concinnata, I utilized experimental populations of two lepidopterans in Massachusetts and northern New York State. One of the host species, C. promethea, is heavily parasitized by C. concinnata and has precipitously declined in Massachusetts. In New York, C. concinnata parasitism is significantly lower in C. promethea, which is still common across central and western parts of the state. The other species, eastern tent caterpillar (Malacosoma americanum F.), serves as a control because it has remained abundant in both regions, utilizes a host tree (Prunus serotina Ehrh.) shared with Promethea, yet is rarely attacked by C. concinnata. In this case study, the potential confounding effect of historical change in land use was minimized by comparing regions (MA and NY) that underwent a similar conversion from forest to agriculture followed by succession back to forest. Data from Schaffner and Griswold (1934) indicate that Lespesia frenchii was the most common parasitoid recovered from C. promethea from 1915 to 1929 (Table 2). Promethea larvae deployed in two experimental populations in western Massachusetts did not yield L. frenchii in 2004 (Parry 2007) nor did Boettner et al. (2000) recover it from their sentinel populations (Table 2). In contrast, L. frenchii was reared from three experimental populations in New York where C. concinnata parasitism was much lower (see Fig. 1). Given that L. frenchii is highly polyphagous, it is possible that it has shifted to alternative host species less favored by C. concinnata. To investigate this possibility, I quantified parasitism in eastern tent caterpillar, a species favored by L. frenchii and seldom used by C. concinnata. When the parasitoid complex of M. americanum in present day Massachusetts was examined, oligophagous tachinid species (L. exul and C. malacosomae) occurred in the same rank order as Schaffner and Griswald had found 70 years earlier, but

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D.W. Langor & J. Sweeney (eds.)

Table 2 A comparison of historical and present-day tachinid complexes for two species of Lepidoptera Species

Location (dates)

Sample size

C. promethea

New England (1915–1929)a

n = 4644 (multiple sites)

Rank abundance of Tachinid species 1. Lespesia frenchii (p) 2. Compsilura concinnata (p)

Massachussets (1999–2004)

b,c,d

n = 1607 (4 sites)

3. Chetogena sp. (p) 1. C. concinnata No L. frenchii collected

New York (2003–2004)c,d

n = 239 (4 sites)

1. C. concinnata 2. L. frenchii

M. americanum

New England (1915–1929)a

n = 60031 (multiple sites)

1. Carcelia malacosomae (o) 2. Leschenaultia exul (o) 3. L. frenchii (p) 4. Chetogena sp. (p) 5. Exorista mella (p)

Massachussets (2001–2005)d

n = 1246 (12 sites)

1. C. malacosomae (o) 2. L. exul (o) 3. Chetogena sp. (p) No L. frenchii collected

Michigan (2003)d

n = 840 (8 sites)

1. L. exul (o) 2. C. malacosomae (o) 3. L. frenchii (p) 4. Chetogena sp.

Polyphagous species are denoted by (p) and oligophagous species by (o). Promethea (C. promethea) is heavily parasitized by C. concinnata (see Fig. 1) whereas eastern tent caterpillar, Malacosoma americanum, is seldom attacked. Schaffner and Griswold (1934) collected C. promethea and M. americanum in New England shortly after the release of C. concinnata. I collected promethea and eastern tent caterpillar from experimental populations in Massachusetts nearly 100 years later. Additional populations of Promethea and eastern tent caterpillar were collected in New York where parasitism by C. concinnata is much lower. In addition, a collection of eastern tent caterpillar was made in Michigan where C. concinnata is a relatively recent arrival. All tachinids are native to eastern North America except C. concinnata a

Schaffner and Griswold (1934)

b

Boettner et al. (2000)

c

Parry (2007)

d

Parry (unpublished)

the polyphagous L. frenchii was completely missing (Table 2). In northern New York and in Michigan, where non-target parasitism from C. concinnata is much lower, L. frenchii resumed its rank (3rd most abundant) in the M. americanum parasitoid complex. While these data are not conclusive evidence for competitive displacement, they are certainly suggestive of it. In the absence of further experimentation, we can only speculate as to the actual mechanism(s) underlying the disappearance of L. frenchii. One possibility is that the multivoltine L. frenchii is dependent on saturniid silk moths as overwintering hosts. Even though alternative hosts are readily available to L. frenchii at other points in the season,

it may be that the loss or severe reduction of hosts at a critical point in the season can limit its population density. The effects of high C. concinnata attack rates on specialist parasitoids may be even greater. Collections of silk moth larvae from both wild and experimental populations in Massachusetts have failed to yield a number of specialists such as tachinids in the genus Belvosia, recorded by Schaffner and Griswold (1934) and others (Arnaud 1978). Intriguingly, native parasitoids have been recovered from saturniids on Martha’s Vineyard, an island off the coast of Massachusetts where C. conncinata is thought to be absent (G.H. Boettner, personal communication). While admittedly an extreme example, the C. concinnata

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

system does illustrate that going beyond simple hostparasitoid interactions is important when evaluating the ecological impact of CBC introductions.

Future directions in research and regulation Practitioners of biological control have suggested that the introductions responsible for the most severe effects on non-target species were conducted in earlier, less enlightened times, and that modern criteria for release would have prevented the occurrence of many of these effects (e.g., Hoddle 2004; Messing and Wright 2006). Indeed, the studies highlighted in this manuscript illustrate this point well. In Henneman and Memmott (2001), exotic parasitoids found attacking native Lepidoptera were all released prior to 1945, the species’ attacking the koa bug in Hawaii were introduced in the early 1960s (Johnson et al. 2005), and the first releases of Compsilura were in 1906 (Culver 1919). One recent introduction does cast doubt on the frequent assertion that generalist parasitoids would not be used in CBC today. In the late 1970s, a hymenopteran parasitoid from Asia, Pimpla disparis Vierick, was released for gypsy moth control. This was done despite existing knowledge about the highly polyphagous nature of this ichneumonid. Although originally released in Delaware, the state of Minnesota (MDNR 1997, 1999) acquired and released large numbers and a separate introduction occurred in Illinois (Ellis et al. 2005). I have collected P. disparis from eastern tent caterpillar and the exotic lymantriid Euproctis chrysorrhoea (L.) in Massachusetts (Elkinton et al. 2006; D. Parry, unpublished), indicating substantive movement from release sites. In retrospect, it is puzzling why a highly polyphagous, multivoltine parasitoid (Schaefer et al. 1989) was considered a candidate for release, especially given its low rate of attack on the target species (Fuester et al. 1997). Not only was the original release illconceived, but a state agency continued to release this species recently despite increased attention given to non-target concerns in recent years. This example highlights some pressing concerns regarding parasitoid introduction. Clearly, significant regulatory changes coupled with the development of better host-range testing protocols are long overdue. A review by Messing and Wright (2006) highlights

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‘regulatory chaos’ in the United States. They contrast the tangled web of sometimes contradictory rules and regulations regarding the introduction of arthropod natural enemies with the system developed by Australia and New Zealand where well-defined legislative frameworks have been implemented. Host-range testing protocols specifically for parasitoids have lagged behind those developed for CBC of plants (Sands and Van Driesche 2004). The range of behaviors and interactions exhibited by parasitoids in natural environments may be impossible to replicate in captivity rendering interpretation of results difficult (Briese 2005). The development of better risk assessment protocols for evaluating candidate species would be a significant step in assuring ecologists, conservation biologists and other stakeholders that biological control is an effective tool against invasive species rather than a source of additional problems. Insect parasitoids will continue to be the most widely used group for the biological control of arthropods. In contrast to introductions for the biological control of plants, parasitoids were not subjected historically to host-range testing. The slew of recent review and synthesis articles by biological control practitioners, ecologists, and conservation biologists suggest an increasing awareness that a laissez-faire approach to non-target effects is no longer acceptable in arthropod control programs. Although classical biological control is not the Pandora’s Box premised by Howarth, it certainly has not been a Panacea either. While catastrophic non-target effects have been identified in relatively few systems and generally have been restricted to parasitoids that would not be considered candidates today, it is disconcerting that we know so little about the interaction with native species for most releases. Recognition that less dramatic non-target effects can still have fundamental effects on the integrity and function of ecosystems through more subtle mechanisms such as competition or interactions with anthropogenic stressors should provide researchers with alternative avenues to fully assess the non-target effects of parasitoids released for classical biological control. Acknowledgments I thank D. Langor and J. Sweeney for organizing the conference and coordinating the ensuing manuscripts. The saturniid research would not have been possible without the efforts and data of G. Boettner, J. Selfridge, B. Hoven, and G. Tuttle. Special thanks are due

56 G. Boettner for much critical discussion of non-target issues and saturniid silk moths as well as constructive comments that improved the manuscript. N. Finley (Cape Cod National Seashore) and N. Gifford (The Nature Conservancy and Albany Pine Bush Preserve) provided accommodation, labor, and necessary permission to conduct various studies in Massachusetts and New York, respectively. The projects were funded by National Science Foundation grant DEB089699, The Edna Bailey Sussman Foundation, SUNY— College of Environmental Science & Forestry, and the Biological Research Institute (New York State Museum) grant.

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58 Thomas MB, Willis AJ (1998) Biocontrol: risky but necessary? Trends Ecol Evol 13:325–329 Tuskes PM, Tuttle JP, Collins MM (1996) The wild silk moths of North America. Cornell University, Ithaca

D.W. Langor & J. Sweeney (eds.) Webber RT, Schaffner JV Jr (1926) Host relations of Compsilura concinnata Meigen, an important tachinid parasite of the gipsy moth and the brown-tail moth. Bull 1363, United States Department of Agriculture, 31 pp

Distribution and impacts of invasive earthworms in Canadian forest ecosystems J. A. Addison

Originally published in the Journal Biological Invasions, Volume 11, No 1, 59–79. DOI: 10.1007/s10530-008-9320-4
Springer Science+Business Media B.V. 2008

Abstract In Canada it is generally accepted that most indigenous earthworms did not survive glaciation, and that the majority of the earthworms now inhabiting Canadian soils are relatively recent introductions of European origin. Although these exotic earthworms are generally considered to be beneficial in agricultural soils, their effects can be less benign in forested ecosystems. Studies have shown that invading earthworms can significantly alter the forest floor, affecting the distribution of carbon, nitrogen and other chemicals, roots, microbes and other elements of the soil fauna, and even understory vegetation. This paper summarizes the current distribution of exotic and native earthworm species in Canadian forests and draws on the results of studies of invasion patterns and environmental impacts in northern forests in North America and Europe to discuss potential outcomes for forests in Canada. The potential for variables such as temperature, pH, litter palatability and dispersal, to limit or promote the invasion of exotic earthworms in Canadian forests is discussed, and areas for future research are proposed. The same earthworm species that are invading forests in northern Europe and the US are also invading Canadian forests. Several species of exotic earthworms are already established

J. A. Addison (&) School of Environment and Sustainability, Royal Roads University, 2005 Sooke Road, Victoria, BC, Canada V9B 5Y2 e-mail: [email protected]

in a wide range of deciduous and coniferous forest types, including the boreal. Evidence is presented to suggest that further expansion into Canadian forests is likely. Keywords Boreal
Distribution
Earthworms
Exotic
Invasion
Native
Northern forest Abbreviations BOR Boreal forest region COAST Coastal forest region CWH Coastal Western Hemlock Biogeoclimatic Zone GL-ST Great Lakes/St. Lawrence forest region MON Montane forest region

Introduction In North America it is generally accepted that most indigenous earthworms did not survive the late Wisconsonian glaciation and that the majority of the earthworms now inhabiting soils in Canada and northern regions of the US are relatively recent introductions (Gates 1982; Reynolds 1994). Consequently, except for relatively small areas where relict native earthworm populations have survived, Canadian soils, their vegetation, and carbon and nutrient pools, have developed in the absence of earthworms.

D.W. Langor & J. Sweeney (eds.), Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems. DOI: 10.1007/978-1-4020-9680-8_5

59

60

Until recently, most Canadian research on earthworms has focused on agricultural environments where they are thought to have a generally beneficial influence on soil structure and fertility (e.g., Tomlin and Fox 2003; Clapperton et al. 1997). Biogeographical publications, largely the work of Reynolds and collaborators, confirmed that exotic earthworms were becoming widely distributed across Canada. As early as 30 years ago, Reynolds (1976) documented the presence of exotic earthworms in forests in North America, including forest types that occurred in Canada. However, only recently have ecologists and even regulatory bodies started taking a closer interest in the distribution and activities of exotic earthworms, especially those invading natural ecosystems. Much of the concern about exotic earthworms in Canada has been fueled by the rather alarming reports about the environmental impacts these organisms are having in northern temperate hardwood forests in the northeastern US. Those studies have shown that invading earthworms have significantly altered (or even eliminated) the forest floor (Alban and Berry 1994; Burtelow et al. 1998; Bohlen et al. 2004a; Hale et al. 2005) affecting the distribution and availability of carbon, nitrogen and phosphorus (Sua´rez et al. 2003; Bohlen et al. 2004a, b) in the soil. The biological impacts of the invading earthworms in soil included altering the distribution of fine roots (Fisk et al. 2004) and the size and activity of microbial biomass (Li et al. 2002; Groffman et al. 2004) and negatively affecting mycorrhizal colonization rates (Lawrence et al. 2003). The alteration of the soil physical, chemical and biological properties has been reflected in changes to the above-ground vegetation, including negative effects on regenerating maple seedlings (Hale 2004; Hale et al. 2005) and a rare native fern (Gundale 2002). In Canada, results of studies on invasive earthworms in aspen (Populus tremuloides Michx.) and lodgepole pine (Pinus contorta Dougl.) forests in the Kananaskis Valley on the eastern slope of the Rocky Mountains have shown remarkable similarities to the studies carried out in the northeastern US. The invading earthworms greatly modified the structure of the forest floor (McLean et al. 1996; Dymond et al. 1997; McLean and Parkinson 1997) and affected carbon and nitrogen distribution and availability (Scheu and Parkinson 1994; Migge 2001; Eisenhauer et al. 2007). Microbial distribution and activity were

D.W. Langor & J. Sweeney (eds.)

altered (McLean and Parkinson 1997; McLean and Parkinson 2000a), as were the structure and diversity of microarthropod communities (McLean and Parkinson 2000b; Migge 2001) and ultimately the understory vegetation (Eisenhauer et al. 2007). The Kananaskis study is the only Canadian investigation in which researchers were able to follow the patterns of invasion and their associated impacts over a relatively long period of time ([12 years). The aspen site was the location of numerous soil biological studies over the years, and it was well-known that there were no earthworms at that site prior to 1985 (Parkinson et al. 2004). Thus when earthworms first appeared at the site, researchers were able to make direct comparisons of environmental and soil parameters before and after invasion, providing a very clear example of the impacts of earthworm invasion into a forest soil. Other Canadian studies have not had that luxury. Some have described soil parameters and the associated earthworm community at a particular point in time and then tried explain the observed patterns and relationships (e.g., Wironen and Moore 2006). In other cases, all we have is anecdotal accounts of earthworms being found in remote forests, usually by surprised researchers engaged in non-earthwormrelated studies. Given the relative paucity of experimental data from Canadian sources, and the increasing reports of earthworms in northern Canadian forests, it may be useful to examine the patterns of earthworm invasions and their ecological consequences in other environments, and look for commonalities that might provide insights into the future distribution and potential impacts of exotic earthworms in Canadian forests. The research that is being conducted in the northeastern US has obvious relevance to the Canadian situation as that area and parts of southern Ontario and Quebec share a common forest type (Great Lakes/St. Lawrence; Rowe 1972). However, most of Canada is covered with forests that have a significant coniferous content (Forest Regions of Canada NRCan). How well can earthworms handle these less than ideal habitats? Invasion of exotic species of earthworms into forests is occurring over a wide geographical range, and not just in North America (Terhivuo 1988; Ra¨ty and Huhta 2004; Pop and Pop 2006; Tiunov et al. 2006). The Pleistocene glaciation that eliminated the

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

•

endemic earthworms in much of North America performed a similar function in Northern Europe. Given the broad similarities between climatic and soil conditions on the two continents, the northern European experiences, especially those of the Taiga region in Russia and the coniferous forests of Scandinavia, may provide insights relevant to Canada. Forests are important in Canada. They cover 43% of Canada’s landmass, and make a significant contribution to the country’s economy and environmental and social well-being. On a global basis, Canadian forests contribute 10% of the world’s forest cover, 30% of the world’s boreal forest and 23% world’s temperate rainforest (NRCan 2006). Earthworm-free forest soils in Canada tend to have relatively thick organic forest floors that are repositories of significant amounts of carbon, nutrients and biodiversity. Earthworms have the potential to change fundamental soil properties, with profound consequences that cascade up to include effects on ecosystem functioning and biodiversity. The widespread invasion of exotic earthworms into Canada’s forests could therefore have global implications. This paper will: •

•

61

evaluate the relevance of observations in other forests types for application to Canadian forests, discuss factors that may limit or promote the invasion of exotic earthworms into Canadian forests and propose areas for future research.

•

•

Earthworm fauna of Canada Exotic European lumbricids are by far the most abundant and widespread of all the earthworms in Canada, even in forests (Reynolds 1976, 1977; Tomlin and Fox 2003). However, Canada does have a few native earthworm species, and because there is a potential for interaction between these indigenous species and the invasive lumbricids, a brief summary of native earthworm distribution will be presented. Native earthworms At present only eight species of native earthworm are known to occur in Canada (Table 1). Most of the native species (Bimastos parvus (Eisen), B. beddardi (Michaelsen), B. lawrenceae Fender and Aporrectodea bowcrowensis Reynolds and Clapperton) have extremely limited distributions and are unlikely to be of any great significance in forests. Sparganophilus eisini Smith, although more widely distributed in Canada, is a limnicolous mud-eating species (Reynolds 1977).

summarize the current distribution of exotic and native earthworm species in Canada and in Canadian forests in particular, review ecological life-strategies of earthworms and their implications for forests,

Table 1 Distribution of native species of earthworms in Canada, by Province and Territory Name

BC

AB

K

MB

ON

QU

NS

NB

PEI

NL

YK

NU

NWT

*e

Aporrectodea bowcrowensis Reynolds and Clapperton 1996

Total no. Prov/Terr 1

Arctiostrotus perrieri (Benham 1892)

*d

1

Arctiostrotus vancouverensis McKey-Fender 1994

*d

1

Bimastos lawrenceae Fender 1994

*d

1

Bimastos parvus (Eisen 1874) Bimastos beddardi (Michaelson 1894)

*a

Sparganophilus eiseni Smith 1895

*a

*b

2

2

2 1

*b

Toutellus oregonensis (Smith 1937)

*

Total Native species

4

a

*c *f

3

d

1 1

0

0

0

0

0

0

1

0

0

Reynolds (1977); b Reynolds and Reynolds (1992); c Berman and Marusik (1994); d McKey-Fender et al. (1994); e Reynolds and Clapperton (1996); f McAlpine et al. (2001)

62

In British Columbia, the coastal rainforests of the Coastal Western Hemlock Biogeoclimatic Zone (CWH; Pojar et al. 1991) are home to at least three species of native earthworms, Arctiostrotus perrieri (Benham), Ar. vancouverensis McKey-Fender and Toutellus oregonensis (Smith), all members of the Family Megascolecidae. The distribution of the latter two species in Canada is restricted to Vancouver Island (McKey-Fender et al. 1994), but Ar. perrieri has also been collected on the Queen Charlotte Islands and numerous small islets off Vancouver Island (McKey-Fender and Fender 1982). Until very recently none of the three species was known to occur on the BC Mainland. However, in 2000 a mature specimen of Ar. perrieri was captured in a stand of big-leaf maple (Acer macrophyllum Pursh) at a contaminated site near Squamish, BC, and additional specimens, collected during a study by Hocking and Reimchen (2002) at two forested sites along the Clatse and Neekas Rivers near Bella Bella on the BC coast, were also identified as Ar. perrieri (J. Addison and V. Marshall, personal observation). This raises the possibility that this species may be more widely distributed in coastal rainforests than was thought. The ecological importance of native Canadian earthworms is poorly understood. Fender and McKey-Fender (1990) reported that Ar. perrieri is a parthenogenic habitat generalist, found in the upper layers of coniferous soil, rotting wood, moss layers and in well-aerated muck of cold seeps and small streams. It is apparently very intolerant of heat, and will decompose quickly at room temperature, which at least to some degree probably explains the dearth of information about the functional role of this animal. Spiers et al. (1986) reported that Ar. vancouverensis (at that time still referred to as ‘‘an undescribed indigenous earthworm species in the Megascolecidae’’) is widely distributed and relatively abundant (up to 200 individuals m-2) in CWH forests on Vancouver Island, where it plays an important role in decomposition, humus formation and nutrient cycling. Exotic earthworms Nineteen species of exotic earthworms of European origin are known in the wild in Canada (Table 2), and many are widely distributed across the country. The data were compiled using Reynolds (1977) as a starting point, as this benchmark reference

D.W. Langor & J. Sweeney (eds.)

synthesized what was known about earthworm distribution prior to 1977. Subsequent references reporting the occurrence of each species in the different Provinces and Territories are arranged in chronological order, so the letters provide an estimate of how long the species are known to have been present in the area. Most of the exotic species in Ontario, Quebec and, to a lesser extent BC and Atlantic Canada were reported by Reynolds (1977), and have therefore been present in those regions for at least 30 years, but many of the records for the Prairie Provinces and the northern Territories are more recent. While this table provides some insight into the spread of earthworms in Canada, obviously each record depends not only on the presence of the earthworm, but also the cooperation of someone to collect and identify it. Several of these species are unlikely to be of importance in forests, as they appear to have restrictive habitat requirements. Eisenia fetida (Savigny) is a popular vermicomposting species that can be found all over Canada (including the NWT). It requires substrates high in organic matter (such as manure and compost), a relatively high pH (6.8–7.6) and is frost-intolerant (Reynolds 1977). It is most frequently collected close to human habitation, although there is a single record of that species occurring in a boreal forest (along a riparian area within the city boundary of Thunder Bay in Northern Ontario) (Ferguson 2004). Similarly, Eisenia hortensis (Savigny) is also cultured for use in vermicomposting; according to Reynolds and Reynolds (1992), it has a very limited distribution in North America, being most commonly found in greenhouses, arboretums and in material rich in organic matter. Fender (1985) provides a vivid description of its association with the ‘‘foulest of habitats’’ in the western US (including drains and putrid organic sludge), and concluded this species was restricted to moist, warm habitats with an abundance of fresh organic matter, a description that at least presently, does not fit Canadian forests. Eiseniella tetraedra (Savigny), although widely distributed in Canada, prefers very moist habitats such as ditches and seeps, along streams or lake margins (Fender 1985; Reynolds and Reynolds 1992; Moore and Reynolds 2003) and so is unlikely to play a substantive functional role in forests. The habitat requirements of Lumbricus festivus (Savigny) are poorly known;

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

63

Table 2 Distribution of exotic species of earthworms in Canada, by Province and Territory Name

BC AB SK MB ON QU NS NB PEI NL YK NU NWT Total No. Prov./Terr.

Allobophora chlorotica (Savigny 1826)

*a

*g

*g

Aporrectodea icteria (Savigny1826)

*a

*a

*a

*a

*a

*a

*a

*a

*a

*

*i

1

*b

*g

Aporrectodea rosea (Savigny 1826) *a Aporrectodea trapezoides (Duge`s 1828) *a

*a

*h

*j

*a

*a

*a

*a

*a

*a

*h

*a

*a

*a

*a

*a

*a

Aporrectodea tuberculata (Eisen 1874)

*a

*a

*a

*a

*a

*a

*a

*a

*a

*a

Aporrectodea turgida (Eisen 1873)

*a

*d

*g

*a

*a

*a

*a

*a

*a

*a

*j

*a

*a

*a

*a

*a

*a

Aporrectodea longa (Ude 1885)

7 *a *l

11 10

o

*

Dendrobaena octaedra (Savigny 1826)

*a

*a

Dendrodrilus rubidus (Savigny 1826)

*a

*a

*a

*a

*a

*a

*a

*a

*a

*a

a

e

h

a

a

*

a

a

k

*

*

g

*i

*

c

*

10 9

Dendrobaena attemsi Michaelsen 1902

Eisenia fetida (Savigny 1826)

8

a

1

*

*

*

*

Eisenia hortensis (Savigny 1826) Eiseniella tetraedra (Savigny 1826)

*a

Lumbricus castaneus (Savigny 1826)

*d

*j

*

*n

10 10 *f

11

*l

11

1

*a

*a

*a

*a

*a

*i

*b

*a

*a

*a

*a

*a

*a

a

a

a

*a

*a

*n

4 10

*a

*a

*n

11

Lumbricus festivus (Savigny 1826) Lumbricus rubellus Hoffmeister 1843

* *a

*d

Lumbricus terrestris Linnaeus 1758

*a

*d

*h

a

g

g

Octolasion cyaneum (Savigny 1826)

*

Octolasion tyrtaeum (Savigny 1826)

*a

*d

Total exotic species

17

14

Total native species

4

1

Total earthworm species

21

15

*

*

*a

* *a

*a

* *a

*j

*a

*a

*a

*a

a

a

a

g

*

7

a

* *a *

*m

*

*

*

g

*a

*a

*a

*a

8

12

17

17

15

15

12

12

2

2

8

12

19

19

15

16

12

12

8 6 3

2

2

2

2

1 4

a

Reynolds (1977); b Fender (1985); c Reynolds and Reynolds (1992); d Scheu and McLean (1993); e Reynolds and Clapperton (1996); f Reynolds and Moore (1996); g Lee et al. (1998); h Reynolds and Khan (1999); i Reynolds (2000a); j Reynolds (2000b); k Reynolds (2001); l Reynolds (2003a); m Reynolds (2003b); n Teale (2007); o Marshall and Fender (2007)

although this species has been recovered from a wide variety of habitats, it is never abundant. In North America its distribution is limited to Canada (Fender 1985), so it can be assumed that it requires relatively cold conditions for survival. According to Reynolds (1976, 1977), Allobophora chlorotica (Savigny), Aporrectodea longa (Ude) and the infrequently collected Aporrectodea icteria (Savigny) are more characteristic of agricultural and urban environments. Although Whalen (2004) reported that both Ap. longa and A. chlorotica were present at a mixed deciduous forest at the Morgan Arboretum in Ste Anne-de Bellevue in Quebec, they represented \1% of the mature earthworms numbers at the site. Dendrobaena attemsi Michaelsen was found at Pacific Rim National Park on the west coast of Vancouver Island, but has only been collected once in Canada. All the remaining species of exotic earthworms listed in Table 2 have already been collected from

forests, or, based on their habitat requirements, would be expected to occur in forests.

Ecological life-strategies of earthworms and their implications for forest soils Bouche´ (1977) provided a useful framework for classifying the activity and hence the impacts of different life strategies on soil. In his classification, epigeic earthworms inhabit and feed on accumulations of organic matter at the soil surface, and are typically small, pigmented and highly mobile. Many epigeics are parthenogenic and have high reproductive rates. Endogeic species live and feed in the mineral horizons below the organic (LFH) layer, and are usually lightly (or non-) pigmented, mediumlarge in size and slower-moving with lower reproductive rates. Anecic earthworms live deep in the soil

64

but come to the surface to feed on litter and deposit their casts. They tend to be large, often pigmented at the anterior end, with a low reproductive rate. Clearly these different strategies will have very different consequences for forest soils. Typically epigeics (e.g., Dendrobaena octaedra (Savigny)) physically disrupt the organic layer by consuming and mixing the F and H layers, producing a homogeneous, granular form of organic forest floor. The endogeic earthworms (often Aporrectodea spp. or Octolasion tyrtaeum (Savigny)) mix the organic material with the underlying mineral material, while the anecic earthworms (e.g. Lumbricus terrestris L.) remove litter from the surface, pulling it down into the mineral layer, and deposit casts of mixed organic and mineral material on the soil surface. While Bouche´’s classification provides a useful first step for predicting the potential impacts of earthworm activity, in the field the situation is often more complicated. The actual behaviour of earthworms may not always conform to their published classification. For example, the endogeic Aporrectodea turgida (Eisen) feeds below the soil surface, but casts on the surface, a behaviour more characteristic of anecic worms (Fender 1985). Lumbricus rubellus Hoffmeister operates in two categories, and is therefore usually considered to be epi-endogeic in its habits, feeding on organic matter in the forest floor, but also mixing the organic material into the upper layer of mineral soil (Terhivuo 1988). In addition, the same species of earthworm may have different effects on different ecosystems. In the aspen and lodgepole pine forests of Alberta, the epigeic earthworm D. octaedra significantly altered soil strata and populations of other soil organisms, whereas in Minnesota, where that same earthworm species invaded maple/basswood (Acer spp./Tilia spp.) forests, the main impacts of earthworm invasion were not evident until later, when L. rubellus arrived at the site (James and Hendrix 2004).

Exotic earthworms in Canadian forests Distribution and activity of exotic earthworms in Canadian forests Ecological studies that include direct evidence of the presence and activity of identified species of earthworms in specific forest types are presented in

D.W. Langor & J. Sweeney (eds.)

Table 3. A common classification system has been used for forests (Forest Regions of Canada, NRCan, based on Rowe (1972)) and soils (Canadian System of Soil Classification 1998) to allow comparison of data from different geographical areas. The relative lack of ecological studies from southern Ontario and Atlantic Canada does not mean that earthworms are not present and active in forests of these areas. Indeed, one of the first documented accounts of the impacts of exotic earthworms on forest soils came from Langmaid (1964) who reported that the activities of exotic earthworms had created mull-like soils in podzols in a New Brunswick cutover. In southern Ontario, it is more likely that the invasion of earthworms into hardwood forests just passed unnoticed (Parkinson et al. 2004). It was only when exotic earthworms started turning up in areas that were thought to be inhospitable to earthworms, or were known to have been earthworm-free, that they were deemed to be worthy of study (and funding). At present, exotic earthworms are found in a wide range of forest types in Canada, including those with significant coniferous content. They also inhabit a broad range of forest soils, including podzols, brunnisols and luvisols, many of which are acidic (Table 3). Coast and Columbian forests of British Columbia No studies have directly addressed the invasion of exotic earthworms into the coastal rainforest of BC, but there are several reports of their presence. In a study of nematodes in Coastal Western Hemlock (CWH) forests near Robert’s Creek (*40 km north of Vancouver), Panesar et al. (2000) discovered some of their sampling areas had been invaded by L. rubellus, which was converting the 5–6 cm forest floor (typically humimoders and mormoders) to a mull humus form. Similarly, during a study of microarthropods in regenerating CWH forests, Addison et al. (2003) reported that Aporrectodea tuberculata (Eisen) had invaded one of the study plots (a mature stand *60 years old) and that the depth of the LFH layer in the stand had decreased. Dendrodrilus rubidus (Savigny) was also collected in several remote locations along the west coast of BC (J. Addison, personal observation) during a study on the impact of the relationship between grizzly bears and salmon on forest nitrogen cycling (Hocking and Reimchen 2002).

COAST COL BOR COAST GL-SL GL-SL

Ap. trapezoides

Dd. rubidus

D. octaedra

Ap. turgida

GL-SL

MON BOR BOR

GL-SL COAST

BOR

BOR BOR

MON

MON

MON

GL-SL GL-SL GL-SL MON MON

GL-SL GL-SL GL-SL

Ap. rosea

Ap. tuberculata

Forest type

Species

Alnus Fagus grandifolia, Acer rubrum Acer saccharum, Betula alleghaniensis Populus tremuloides, Populus balsamifera Pinus contorta, Abies glauca, Populus tremuloides Pinus contorta, Picea glauca 9 engelmanii, Populus temuloides Pinus contorta, Abies lasiocarpa, Picea glauca 9 engelmanii Abies lasiocarpa, Pseudotsuga menzeii, Picea glauca 9 engelmanii Abies balsamea, Betula papyrifera Abies balsamea, Picea glauca, Populus tremuloides Populus tremuloides, Populus balsamifera, Picea glauca Acer saccharum, Betula alleghaniensis Tsuga heterophylla, Pseudotsuga menziesii, Thuja plicata Populus tremuloides, Populus balsamifera Abies balsamea, Betula papyrifera Populus tremuloides, Populus balsamifera, Picea glauca Acer saccharum, Betula alleghaniensis

Pseudotsuga menziesii, Tsuga heterophylla Pseudotsuga menziesii, Betula papyrifera Abies balsamea, Betula papyrifera Pseudotsuga menziesii, Tsuga heterophylla Fagus grandifolia, Acer rubrum Acer saccharum, Fagus grandifolia

Alnus Fagus grandifolia, Acer rubrum Acer saccharum, Fagus grandifolia

Dominant tree species

Orthic Ferro-Humic Podzol

Orthic Gray Luvisol Brunisol Gray Luvisols, Brunisols

Orthic Ferro-Humic Podzol Ferro-Humic Podzol

Gray Luvisols, Brunisols

Humo-Ferric Podzol, Dystric Brunisol, Orthic Gleysol Brunisol Humo-Ferric Podzol

Gray Luvisol

Gray Luvisol

Brunisol Humic Gleysol Orthic Dystric Brunisols or Orthic Sombric Brunisols Humo-Ferric Podzol Brunisol, Luvisol Brunisol Humo-Feric Podzol Humic Gleysol Orthic Dystric Brunisols or Orthic Sombric Brunisols Brunisol Humic Gleysol Orthic Ferro-Humic Podzol Orthic Gray Luvisol Orthic Eutric Brunisol

Soil type

Mor

NA Modera NA

Mor NA

3.4

5.5–6.7 *7 NA

2.8–3.6 NA

NA

*7 3.4–4.5c

Modera Mor NA

4.3

4.6

4.8

Hemimor

Hemimor

Moder

NA 6.2 2.8–3.6 5.5–6.7 4.8–5.4

4.2 4.8–5.48 *7 3.6b 6.2 3.0–5.2

Modera NA Modera Modera/mull NA NA Mull NA Mor NA NA

NA 6.2 3.0–5.2

pH

Mull NA NA

Humus

Moore and Reynolds (2003)

Dymond et al. (1997) Moore and Reynolds (2003) Cameron et al. (2007)

Moore and Reynolds (2003) Addison (unpublished)

Cameron et al. (2007)

Moore and Reynolds (2003) Addison (unpublished)

Battigelli (2000)

Battigelli (2000)

Battigelli (2000)

Moore and Reynolds (2003) Whalen (2004) Moore and Reynolds (2003) Dymond et al. (1997) McLean and Parkinson (1997)

Addison (unpublished) Welke and Parkinson (2003) Moore and Reynolds (2003) Addison et al. (2003) Whalen (2004) Wironen and Moore (2006)

Moore and Reynolds (2003) Whalen (2004) Wironen and Moore (2006)

Reference

Table 3 Exotic earthworm distribution in Canadian forests in relation to forest (NRCan, Forest regions) and soil types (Canadian System of Soil Classification 1998)

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems 65

Populus tremuloides, Populus balsamifera Abies balsamea, Betula papyrifera Acer saccharum, Fagus grandifolia Fagus grandifolia, Acer rubrum Populus tremuloides, Populus balsamifera Alnus Abies balsamea, Betula papyrifera

MON BOR GL-SL

GL-SL MON GL-SL BOR

BOR GL-SL GL-SL

BOR Brunisol Brunisol Orthic Dystric Brunisols or Orthic Sombric Brunisols Orthic Gray Luvisol Brunisol Orthic Dystric Brunisols, Orthic Sombric Brunisols Humic Gleysol Orthic Gray Luvisol Brunisol Brunisol

Orthic Dystric Brunisol

Gray Luvisol

Humo-Ferric Podzol

Orthic Ferro-Humic Podzol Ferro-Humic Podzol

Soil type

NA NA Mull Modera

6.2 5.5–6.7 NA *7

5.5–6.7 *7 3.0–5.2

*7 NA 3.0–5.2

Modera/mull Mull NA NA Modera NA

4.8–5.3

3.8c 4.3c 4.6

Humimora Mormodera Moder Moder

3.2 NA

pH

Mor Moder

Humus

Whalen (2004) Eisenhauer et al. (2007) Moore and Reynolds (2003) Moore and Reynolds (2003)

Eisenhauer et al. (2007) Moore and Reynolds (2003) Wironen and Moore (2006)

Moore and Reynolds (2003) Moore and Reynolds (2003) Wironen and Moore (2006)

Addison (unpublished)

Battigelli (2000)

Panesar et al. (2000)

Moore and Reynolds (2003) Addison (unpublished)

Reference

NA, information not available; a ‘‘Mull-like’’ in the presence of earthworms; b Trofymow and Porter (1998); c Scarratt (1996); d Value not measured directly. Based on values for that humus type given in Klinka et al. (1981)

Octolasion tyrtaeum

L. terrestris

Picea contorta, Picea glauca 9 engelmanii, Populus tremuloides Abies balsamea, Picea glauca, Populus tremuloides Abies balsamea, Betula papyrifera Alnus Acer saccharum, Fagus grandifolia

MON

COAST

Acer saccharum, Betula alleghaniensis Tsuga heterophylla, Pseudotsuga menzeii, Thuja plicata Tsuga heterophylla, Pseudotsuga menziesii Thuja plicata

GL-SL COAST

L. castaneus L. rubellus

Dominant tree species

Forest type

Species

Table 3 continued

66 D.W. Langor & J. Sweeney (eds.)

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

Kranabetter and Banner (2000) investigated chemical properties of forest floors on the north coast of BC near Prince Rupert, and reported exotic earthworms were abundant at several of their sites. One site (Hammer Island) had to be treated as an outlier in the statistical analyses because it had much higher concentrations of calcium and microelements than the other replicates of the same type. The authors suggested that a large population of exotic earthworms had changed the chemical properties of the soil by mixing mineral subsoil into the upper mineral soil and forest floor horizons. Unfortunately the earthworms responsible for these effects did not preserve well in alcohol, and could be identified only as ‘‘perhaps’’ L. rubellus. For several years the students of the Bamfield Temperate Forest Ecology Course have searched for earthworms in old-growth and harvested CWH forests near Bamfield on the west coast of Vancouver Island. Lumbricus rubellus is frequently encountered in clear-cuts, while the riparian areas have yielded a mixture of exotic and native species (L. rubellus, Dd. rubidus, Octolasion cyaneum (Savigny) and Ar. vancouverensis) (unpublished data). High numbers of the exotic earthworm species Ap. trapezoides (Duge´s) (370 individuals m-2 in one pedon), were recently found in a clear-cut in the CWH forest on Vancouver Island (J. Addison, unpublished data) as part of a study to determine the impacts of various harvesting and retention treatments on forest productivity and biodiversity. Earthworm casts within the soil profile and on the soil surface were found at the nearby unharvested control, together with an obvious homogenization and reduction in the LFH layer. A 5-min search found several individuals of Ap. trapezoides and L. rubellus at two of the four sampling locations (those nearest the road). Apporrectodea trapezoides also occurs in Columbian forests where, at least in laboratory experiments, it has the potential to enhance the growth of Douglasfir (Pseudotsuga menziesii (Mirbel) Franco) seedlings (Welke and Parkinson 2003). Boreal and montane forests The species most commonly encountered in boreal and montane forests in western Canada were D. octaedra (McLean and Parkinson 1997; Dymond

67

et al. 1997; Ca´rcamo et al. 1998; Cameron et al. 2007) and, to a lesser extent, Dd. rubidus (Dymond et al. 1997; Cameron et al. 2007) and L. rubellus (Battigelli 2000). Both L. rubellus and D. octaedra were also collected in boreal forests sites in northern Ontario, *120 km northeast of Thunder Bay, as part of the Black Sturgeon Lake Boreal Mixedwood Research Project (Scarratt 1996). At the Black Sturgeon Lake site, L. rubellus was locally very abundant (80–176 individuals m-2) at the control plots nearest Eskwanonwatin Lake (*1 km from the lake), where its activities homogenized the LFH layer. Dendrobaena octaedra was present in relatively low numbers in several samples taken from both cut and uncut areas of the research site, but was abundant (up to 1,500 m-2 in one sample) only in the uncut plot nearest the road (J. Addison, unpublished data). More recently O. trytaeum and L. terrestris (Eisenhauer et al. 2007) have invaded the aspen site at Kananaskis, and Parkinson et al. (2004) reported that Aporrectodea spp. were also present in pine forests in the area. The impacts of these species on soil physical, chemical and biological properties have been well documented (see ‘‘Introduction’’). Moore and Reynolds (2003) found high earthworm density and abundance in a small patch of soil situated *600 m from a recreational building in the boreal Montmorency Forest in Quebec. Although the typical pH of soil in the area was *4.0, the pH was *7.0 in this discrete area and the soil exhibited a mull structure rather than the mor humus form characteristic of the rest of the area. This small area supported Ap. trapezoides, D. octaedra, L. rubellus, E. tetraedra, L. terrestris and O. tyrtaeum. Great Lakes/St. Lawrence forests Exotic earthworms are widely distributed throughout the GL-SL forest type. The Turkey Lakes Watershed (maple/yellow birch (Betula alleghaniensis Britton),*60 km north of Sault Ste. Marie) has at least two species: D. octaeda and L. rubellus; these plus Ap. turgida and Ap. tuberculata were abundant in a maple/beech (Fagus spp.) site near Wharnciffe ON (*60 km east of Sault Ste. Marie (J. Addison, unpublished data). Moore and Reynolds (2003) found five species (Ap. turgida, D. octaedra, Dd. rubidus, E. tetraedra and Lumbricus castaneus (Savigny)) in a

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very acidic (pH = 2.8–3.6) maple/yellow birch forest in Southern Quebec while Wironen and Moore (2006) reported L. terrestris, L. rubellus, Aporrectodea rosea (Savigny), A. tuberculata and D. octaedra in an old-growth beech/maple forest in the Gault Nature Reserve in southern Quebec. The latter study also reported dramatic decreases in the thickness and mass of organic layers in areas with high earthworm numbers, as well as increases in C and N concentrations in the lower layers of mineral soil. In their study, the epigeic D. octaedra was recovered only from locations distant from the lake. In other sites, particularly those with a high oak (Quercus spp.) content or sites at the northern limit of the GL-SL forest type, D. octaedra may be the only species present (J. Addison, unpublished data).

Earthworm species in northern forests of the US and northeastern Russia Table 4 compares the earthworm species identified as having a significant presence in Canadian forests with those known to be invading forests in northern Russia and the northeastern US. For the most part, the same

species are involved in all three regions. The droughttolerant Ap. trapezoides appears to be an important species in the drier forest ecosystems in BC but does not figure prominently in current invasions in other areas. In view of the similarities of the species involved in forest invasions, it may be useful to look for commonalities that could then allow extrapolation to a wider set of forest environments.

Patterns of invasions and impacts Earthworm invasions typically occur in waves (Hendrix and Bohlen 2002; James and Hendrix 2004; Sua´rez et al. 2006; Eisenhauer et al. 2007). Epigeic species are usually the first to arrive because they are small, easily transported and are able to utilize the undisturbed forest floor (Dymond et al. 1997). The first noticeable impacts tend to be physical disruption of the stratified humus layers in the forest floor. Endogeics such as Aporrectodea generally only invade after the organic layer has been modified by epigeic or epiendogenic species. In the northeastern US (Minnesota), D. octaedra (epigeic) was the first species to arrive, followed by L. rubellus (epi-endogeic),

Table 4 Invasive earthworms in forests in Canada, NW Russia, and northeastern USA Exotic earthworm species known from Canadian forests

NW Russia boreal coniferousa

NE US hardwoodsb

Ap rosead

*

*

NE US boreal aspen/firc

Ap. trapezoides Ap. tuberculata

*

*

Ap. turgidae

caliginosa

caliginosa

caliginosa

D. octaedra

*

*

*

Dd. rubidus L. castaneus

* *

* *

*

L. rubellus

*

*

*

L. terrestris

*

*

*

O. tyrtaeumf

*

*

*

a

Tiunov et al. (2006)

b

Shakir and Dindal (1997); Bohlen et al. (2004a)

c

Hale and Host (2005)

d

Ap. rosea is also reported as Eisenia rosea or Allobophora rosea in other studies (Reynolds 1977, 1995)

e

According to Reynolds (1995) ‘‘Ap. caliginosa’’ may refer to several different species. However the ‘‘Ap. caliginosa’’ in northern Europe generally is the same species as ‘‘Ap. turgida’’ in North America (Fender 1985; Reynolds 1995). For a European perspective on the taxonomic controversy please see Sims and Gerard (1999) f

O tyrtaeum may also be reported in the European literature as O. lacteum (Terhivuo 1988)

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

Aporrectodea spp. (endogeic) and finally L. terrestris (anecic) (James and Hendrix 2004). In the Kananaskis aspen site, D. octaedra (and to a lesser extent Dd. rubidus) were the first to arrive, followed by O. tyrtaeum (endogeic) and L. terrestris (Eisenhauer et al. 2007). At the nearby lodgepole pine site, D. octaedra and Dd. rubidus were the first colonizers, but more recently pine forests have been invaded by endogeic species such as Aporrectodea (Parkinson et al. 2004). Earthworm densities in the initial invasion of a site can be extremely high, but are probably transient. Dymond et al. (1997) reported an average density of 2,621 individuals m-2 of D. octaedra in the early stage of the invasion; this dropped to 76 individuals m-2 within a few years, a figure typical of this species in northern European forests. As these later densities were negatively correlated with the densities of L. terrestris and O. tyrtaeum, Eisenhauer et al. (2007) suggested that this second wave of invasive earthworms had essentially consumed the habitat required by D. octaedra. Wironen and Moore (2006) reported a similar observation regarding the distribution of D. octaedra in a beech/maple forest in Quebec. Measurable impacts of earthworm invasion soon become apparent. Podzolic moder soils were converted to mixed, mull structured soils in cutovers in New Brunswick within 3–4 years (Langmaid 1964). Earthworms first observed in an Alberta aspen site in 1985 (Dymond et al. 1997) had reached very high densities (*3,000 m-2 30 m from the road) by 1992. Significant impacts on soil structure, microbes and microfauna were documented within a few years (Migge 2001; Parkinson et al. 2004), and by 2007 impacts on the understory could be demonstrated (Eisenhauer et al. 2007). The order of arrival of different species can be critical (Frelich et al. 2006). Lumbricus rubellus can consume 10 cm or more of intact forest floor in a single growing season whereas L. terrestris consumes mainly the litter layer and so gradually reduces the input of organic material to the lower layers in the forest floor. If L. rubellus is the first earthworm species to arrive, the rapid rate of habitat alteration makes it difficult for vegetation rooted in the forest floor to adapt, and plant mortality may be high. L. terrestris may allow a more gradual change in the thickness of the forest floor, allowing more opportunity for other organisms to adapt (Frelich et al. 2006).

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Factors affecting exotic earthworm distribution Temperature Most lumbricid earthworms are relatively intolerant of freezing (Holmstrup and Zacharissen 1996; Holmstrup 2003), suggesting that winter temperatures may limit their expansion into northern forests. Earthworms can use three strategies to survive freezing temperature: they can be frost tolerant and survive the effects of freezing of body fluids, they can supercool or they can avoid freezing conditions by overwintering in a frost-free refuge, usually deep in the soil (Holmstrup 2003). As the response to temperature depends on the species, developmental stage, history of acclimation, duration of the exposure and the geographical source of the test population (Holmstrup 1994; Holmstrup et al. 1999; Berman et al. 2001; Rasmussen and Holmstrup 2002), it is difficult to predict exactly how important temperature will be as a factor limiting the expansion of earthworms into northern forests. Tiunov et al. (2006) presented a ranking of earthworm species of north-eastern Europe based on laboratory data of their frost tolerance as: A. chlorotica \ L. castaneus \ L. terrestris \ Ap. rosea B Aporrectodea caliginosa (Savigny) \ L. rubellus \ O. tyrtaeum \ Dd. rubidus \ D. octaedra. The epigeic D. octaedra is extremely frost-tolerant and can withstand freezing in all stages of development (Holmstrup 1994; Berman et al. 2001; Parkinson et al. 2004). Berman et al. (2001) reported the 50% mortality threshold for D. octadrea in soil was -12C, but cocoons of this species survived temperatures of -18C to -20C (Holmstrup and Westh 1995; Berman et al. 2001). Cocoons of Lumbricus and Aporrectodea species survive temperatures of -5C to -8C, but the post-embyronic stages are not so tolerant and probably rely on avoiding extreme temperatures for survival (Tiunov et al. 2006). Berman et al. (2001) found that the geographic distribution of D. octaedra in Northern Europe followed the winter minimum -12C isotherm for soil, which correlates well with its measured tolerance to freezing. Tiunov et al. (2006) used isotherms of the mean air temperature in January to examine the macro-scale distribution of lumbricid earthworms in north-eastern Europe and the Russian plain (north of

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Fig. 1 Mean minimum January temperatures. Dotted line highlights the -15C (+5F) isotherm and solid line is the -20C (-4F) isotherm.* indicates the approximate location of the following studies; (1) Kananaskis aspen and lodgepole pine sites (e.g., Dymond et al. 1997; McLean and Parkinson 1997; Eisenhauer et al. 2007) (2) Cameron et al. (2007). Only the location of the southern-most of the 17 sampling sites is shown. (3) Falling Snow Ecosystem (J. Addison, unpublished) (4) Boreal Mixedwood study (J. Addison, unpublished), (5) Montmorency Forest (Moore and Reynolds 2003)

55N, east of 30E). This approach indicated that the distribution of D. octaedra was limited by the -20C air isotherm while the distributions of Ap. caliginosa, Ap. rosea, O. tyrtaeum, L. rubellus, L. terrestris and Dd. rubidus were all constrained by the -15C isotherm. Figure 1 highlights the -15C (+5F) and the 20C (-4F) Canadian mean minimum air temperature isotherms for January. If the distribution of earthworms (or at least exotic endogeic earthworms) is limited by the -15C January isotherm in Canada, we would only expect to find earthworms in coastal areas, southern BC and the very southern tip of Ontario, a distribution that is not coincident with the observed distribution of earthworms presently established and active in Canadian forests (Tables 3 and 4). If we accept the -20C (-4F) isotherm as a reasonable first step towards defining lower temperature limit for exotic earthworm distribution, the area potentially available for earthworm establishment would include most of the forests in BC, a large portion of Alberta, southern areas of Ontario and Quebec, coastal areas of Labrador and all of Atlantic Canada. However even the -20C mean minimum

January isotherm appears to be a rather conservative predictor of earthworm distribution in forests, as there are already several instances of a significant presence of exotic earthworm species in forests subject to much colder temperatures, particularly in northern Alberta (Cameron et al. 2007) (Fig. 1). Litter palatability and pH Deciduous forests tend to have larger and more diverse earthworm populations than coniferous forests (Edwards and Bohlen 1996; Terhivuo 1989), which is thought to be due to the higher soil pH and more palatable litters (Satchell 1967; Kreutzer 1995). In practice it is often difficult to separate the effects of the two variables because acid soils tend to support trees with unpalatable litters. Yet litters produced by different conifer species differ considerably in their ability to support earthworms. In a common garden test with 14 tree species, Reich et al. (2005) did not find a clear contrast between conifers and deciduous species; the litter of several species of conifers supported a higher earthworm biomass than did some of the deciduous species. In their experiment, calcium

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

content of the litter was a key determinant of soil chemistry. Interestingly, the litter of Douglas-fir, one of the most important species in BC forests, supported a significant earthworm population even though soil pH (4.01) was fairly low. The importance of pH in determining the abundance of earthworms is illustrated by the study of Haimi and Einbork (1992), who found that humus from a spruce stand was too acid for Ap. caliginosa tuberculata (Eisen), but that the earthworms thrived in the same humus after liming had increased the pH to 4.8. In the field it is often difficult to distinguish between direct and indirect effects of liming. Kreutzer (1995) reported a 10-fold increase in earthworm biomass in a stand of Norway spruce (Picea abies (L.) Karsten) when liming increased pH levels in the organic layer from pH *3 to pH *5; the treatment also dramatically increased the productivity of herbaceous vegetation, thus also changing litter characteristics at the site. Dendrobaena octaedra, L. rubellus and Dd. rubidus are acid-tolerant species common in coniferous forests in Europe and North America (Reynolds 1977; Terhivuo 1988; Tiunov et al. 2006). Dendrobaena octaedra feeds readily on coniferous litter (Heugens 1969; Mclean et al. 1996; McLean and Parkinson 1997) and also grows and reproduces well in litter with a high content of oak (Addison and Holmes 1996). In Canada these three species are already well established in forests dominated by lodgepole pine, balsam fir (Abies balsamea (L.) P. Mill.), several species of spruce, Douglas-fir, western hemlock (Tsuga heterophylla (Raf.) Sarg.) and western red cedar (Thuja plicata Donn ex D. Don) (Table 3). However, even D. octaedra has its limits. Ca´rcamo et al. (1998) found that juveniles of D. octaedra reared in soils taken from a lodgepole pine forest near a sour gas plant (pH 2.9) in the eastern Foothills of the Canadian Rocky Mountains exhibited lower growth and survivorship relative to those reared in control soil (pH 5.7). In their field sampling, only a single earthworm was found at the site nearest the sulphur block where the soil pH was 2.9, but sites further from the block (pH 5.7 and 6.1) contained healthy populations of D. octaedra (150–500 individuals m-2). Unpalatable litter and low soil pH pose more of a problem for endogeic and anecic species, but even amongst these groups there are species that can cope.

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Terhivuo (1989) found that although the earthworm fauna of coniferous forests in southern Finland (Norway spruce and Scotch pine (Pinus sylvestris L.), pH 3.3–5.6) was dominated by D. octaedra with lesser components of Dd. rubidus and L. rubellus, populations of Ap. caliginosa, Ap. rosea and L. terrestris were also present. In the northeastern US, Shakir and Dindal (1997) reported Ap. tuberculata, Ap. turgida. L. rubellus, L.castaneus and O. tyrtaeum in single-species plantations of European larch (Larix decidua P. Mill.) and Norway spruce; the earthworm communities in these coniferous stands were very similar to those of the nearby hardwood (maple/beech/oak) forest. In Canada, the endogeic species Ap. tuberculata, Ap. trapezoides and O. tyrtaeum occur in coniferous forests (Table 3). L. terrestris can tolerate soils with pH values as low as 3.5–3.7 (Vimmerstedt 1983; Edwards and Lofty 1975). Chan and Mead (2003) established a positive relationship between pH and the abundance of Ap. trapezoides, and then used regression analysis to suggest that the abundance of this earthworm species would be reduced to zero at pH 4. Some species of earthworms increase the pH of the medium in which they live, modifying the environment so that less acidophilic species can invade. In laboratory experiments using coniferous material, Haimi and Huhta (1990) reported that L. rubellus raised pH by 0.2–0.6 units and D. octaedra by 0.1– 0.4 units over a 15 week period. The mechanisms are not fully understood, but the authors suggested NH+ excretion (which decreases the H+ pool), and the action of calciferous glands. Moreover, the ability to raise soil pH is not limited to epigeic species. In similar experiments using soil from a deciduous forest, Ra¨ty (2004) found a combination of A. caliginosa and L. terrestris significantly raised the pH in the topsoil and humus over a period of 60 weeks. In these species, mucous secretion as well as translocation of mineral material are implicated as the agents for raising soil pH (Haimi and Huhta 1990; Tiunov and Scheu 1999). Dispersal The dispersal rate of earthworms through soil is relatively low, usually only *5–10 m year-1 (Ghilarov and Perel 1984; Marinissen and van den Bosch 1992),

72

so under their own power they could only advance a rather modest *1 km every 100 years. Thus passive dispersal is critical for allowing the earthworms to expand their ranges. Anthropochorous dispersal (e.g., in agricultural products, fishing bait, along roads) has played a large role in the dispersal of exotic earthworms in the past (Reynolds 1977) and undoubtedly continues to facilitate the movement of exotic earthworms into new areas (Cameron et al. 2007). Many of the exotic earthworm species now invading forests are important commercial fishing bait species (e.g., Aporrectodea spp., L. rubellus and L. terrestris), and are discarded by anglers in remote areas of forest. Earthworms are also widely used in composting, and the cultures may contain unexpected species in addition to those traditionally used (Proulx 2003). The association of many earthworm species with areas of human habitation, especially at the extreme limits of their distribution, such the Northwest Territories and Nunavut (Reynolds 2003a, b) and northern regions of Finland (Terhivuo 1988) and Russia (Tiunov et al. 2006) all point to the importance of human-mediated dispersal. However, anthropochorous dispersal may be more important for some species than for others, an important consideration in formulating strategies to limit or slow down the expansion of earthworms into new areas. This does not mean that human activities do not assist dispersal of these species at all, but simply that dispersal into remote areas may be possible even in the absence of human intervention. Terhivuo (1988) suggested that human-made habitats do not play any major role in the habitat spectra of the epigeic species D. octaedra and Dd. rubidus. Both species are found in remote rocky islands, wrack beds of marine areas and in soils bordering brackish and fresh water, suggesting they can be spread by water (Terhivuo 1988). The cocoons of D. octaedra are often embedded within a sphere of fecal matter that would seem to provide a measure of protection from desiccation, allowing transportation by wind or other animals such as ungulates or birds (e.g., Meijer 1972). In a study of human-facilitated dispersal into boreal forests in central and northern Alberta, Cameron et al. (2007) found that while the bait species (Ap. turgida, Ap. tuberculata, L. terrestris and L. rubellus) were almost exclusively found in close proximity to boat launches, D. octaedra and

D.W. Langor & J. Sweeney (eds.)

Dd. rubidus were much more widely distributed. Although D. octaedra was most abundant in transects closely associated with roads, this species was also found (in low numbers) in areas of interior forest well away from areas of human influence. A similar situation was observed in boreal forests in northern Ontario (J. Addison, unpublished data). Terhivuo (1988) suggested that the dispersal of O. tyrtaeum may also be facilitated by water as the high concentration of hemoglobin in the blood allows it to occupy limnic and water-logged habitats usually devoid of other lumbricids. Indeed, distribution patterns at the Kananaskis aspen site suggest that O. trytaeum and L. terrestris invaded along a small creek crossing the site, but that D. octaedra invaded from the road (Dymond et al. 1997; Eisenhauer et al. 2007). Hale and Host (2005) attempted to use various measures of the influence of human development (e.g., distance to roads, trails, streams and campsites) to explain the distribution and abundance of earthworms in hardwood and aspen-fir forests in the western Great Lakes Regions. They found that, except for L. terrestris, distances to human development were poor predictors of earthworm biomass and assemblages. A similar study in south-central New York (Sua´rez et al. 2006) reported that while forest type, and distance to agricultural clearings and wet refugia were significant predictors of earthworm presence, the distance to streams and roads were not. Thus it appears that anthropochorous dispersal is only one of the many factors that influence invasion patterns and dynamics. In Finland, where lumbricid earthworms are also invading following glaciation, exotic earthworms are still absent from several areas of deciduous forest that that would seem to be eminently suitable for colonization. In laboratory experiments using soil collected from an earthworm-free birch forest, Ra¨ty (2004) demonstrated that both L. terrestris and Ap. caliginosa were able to survive in the collected soil for the duration of the experiment (60 weeks), readily feeding on birch and alder (Alnus spp.) leaves, transforming the soil, and suppressing the native enchytraeid populations. In addition, L. terrestris reproduced. Ra¨ty therefore concluded that edaphic factors alone could not explain the absence of earthworms, and suggested that the absence of earthworms in these forests was due to a lack of opportunity to colonize.

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

Interactions with native earthworms and community resistance to invasion The invasion of exotic earthworms into temperate and tropical ecosystems was recently reviewed by Hendrix et al. (2006) who concluded that, even in the absence of obvious disturbance, exotic earthworms were able to invade ecosystems already inhabited by indigenous earthworm species. In some cases the invading exotic species have replaced indigenous species (e.g., Pop and Pop 2006), and according to Kalisz and Wood (1995), remnant undisturbed areas of forests (\30 ha) are particularly vulnerable to invasion by exotic earthworms. However, Hendrix et al. (2006) presented an impressive list of studies in which native and exotic earthworms appear to coexist, at least in the short-term. They also suggested biological interactions with native earthworms may be less important than physical and chemical characteristics of the habitat in determining the degree of community resistance to exotic earthworm invasion. This point of view was shared by Fender (1995) who suggested native earthworm species in the Pacific Northwest appear to have a much higher tolerance than the introduced Lumbricidae for soils with high clay content, low pH and resinous, low nitrogen plant material. Indeed, Spiers et al. (1986) reported that the highest population of the native Ar. vancouverensis in their study was recovered from LFH material that had a pH of 2.9. Very little is known about the interactions between native and exotic earthworms in Canada. The rather limited knowledge we have about the distribution of the native earthworms on Vancouver Island suggests that they are largely confined to the wetter variants of the CWH forests. The records for the exotic endogeic species Ap. tuberculata and Ap. trapezoides in CWH forests (Table 3) refer to drier variants of the forest types, areas that do not support populations of native earthworms. The wetter variants of the CWH near Bamfield on the west coast of Vancouver Island contain both native and exotic species (especially L. rubellus and Dd. rubidus), with the exotics having a significant presence in disturbed areas. On the other hand, disturbance does not necessarily result in the replacement of native earthworms by exotic species. The earthworms of two sites within the Very Wet Maritime Coastal Western Hemlock variant (CWHvm1) (Pojar et al. 1991) were sampled

73

40 years after the forests had been clear cut and then replanted with single species stands of four different conifers. Despite the logging, and subsequent human interventions for tending and monitoring, the native Ar. vancouverensis was the only earthworm at these sites (J. Addison, unpublished data).

What is in store for Canadian forests in the future? The same earthworm species that are invading forests in northern Europe and the US are also invading Canadian forests. Several species of exotic earthworms are already present in a wide range of forest and soil types, and source populations are distributed throughout the country, suggesting that further expansion into forests is entirely possible. The ecosystem-level impacts of invasive earthworms in the northeastern US are taking place in a forest type (GL-SL) that is shared with southern Ontario and Quebec. At least to some extent, many of the same impacts can be expected to occur in Canada. For example, in Quebec, Pare´ and Bernier (1989a, b) found that stands, where earthworms had mixed the forest floor into the mineral soil, had lower concentrations of available phosphorus than stands with an intact forest floor. Similarly, in the northeastern US, Hale et al. (2005) reported sugar maple (Acer saccharum Marsh.) forests invaded by earthworms had lower available phosphorus than forests without earthworms. Several studies in the US reported changes in the distribution of carbon and nitrogen following earthworm invasion (e.g., Bohlen et al. 2004a), and in Quebec, Wironen and Moore (2006) have documented earthworm-related changes in the distribution of organic matter, carbon and nitrogen in a beech/maple forest. There may also be conditions that will tend to mitigate the scale of impacts in the GL-SL forests in Canada. In the Minnesota example, the dramatic impacts on understory vegetation are thought to have been caused by a combination of the effects of earthworms and white-tailed deer (Odocoileus virginianus Zimmerman) (Frelich et al. 2006). Deer (or other herbivores) may, or may not, play a significant role in other locations. In addition, in the more northern areas of the GL-SL forest, less palatable species such as beech become more

74

abundant, pH values tend to decrease, and temperatures are cooler, all of which may slow down the rate of change, allowing a more gradual adjustment to occur. However the prevalence of source populations of endogeic and anecic earthworms, in addition to the epigeic species throughout the GL-SL forests in Canada, suggest that earthworms will have (and indeed already have had) an impact on the forest. Hale and Host (2005) hypothesized that the level of impact of earthworms in forests will be related to the biomass and diversity of invading earthworms. They found the total earthworm biomass in sugar maple stands with very little (unpalatable) beech litter was double that of maple/beech forests, and four times greater than that in aspen/fir forests. Earthworm populations in northern, largely coniferous forests will tend to be smaller that those of the largely deciduous GL-SL forests, or even the montane aspen site in Alberta. Hale and Host (2005) suggest that an ash-free dry weight of 0.01 g of earthworm per m-2 marks the threshold above which we will begin to see proportional changes in forest soil characteristics, a hypothesis that is certainly worth testing. It is unclear how exotic earthworms will impact forests with a more significant coniferous content, but it is certainly reasonable to expect that the epigeic D. octaedra and Dd. rubidus are capable of colonising vast areas of the boreal forest. Dendrobaena octaedra is acid-tolerant, feeds readily on coniferous litter, is extremely frost resistant, is easily dispersed and moreover, it is parthenogenic. In Russia its range includes tundra biomes as well as northern forests (Tiunov et al. 2006), so its expansion into northern areas of Canada seems extremely likely. Both D. octaedra and Dd. rubidus are already present in boreal and montane forests in several provinces in Canada. Lumbricus rubellus is slightly less tolerant of extreme winter temperatures, but thrives in soils with a low pH; its potential for dispersal is high because it is a popular bait species. In addition L. rubellus is one of the species Terhivuo (1988) suggests is not entirely reliant on anthropochorous dispersal. Along with Dd. rubidus, it is the exotic earthworm species most commonly encountered in coastal rainforests in BC. We know from numerous studies that epigeics such as D. octaedra and L. rubellus tend to be the first species to arrive at a site, but that in general, the large-scale impacts on soil physical, chemical and

D.W. Langor & J. Sweeney (eds.)

biological properties are most likely to occur when all ecological groups of earthworms are present (Hale et al. 2005). Many areas of boreal, montane and coastal forests now have this first wave of epigeic species, but we cannot predict when (or if) these forests will support a full range of earthworms. In northern Europe, coniferous forests support a wide range of earthworm species and ecological groups (Terhivuo 1988, 1989; Tiunov et al. 2006), but populations tend to be smaller than in deciduous forests. Observations from the drier variants of the CWH, Columbian forests and the montane Kananaskis aspen and pine forests tend to suggest that some penetration of endogeics (particularly Aporrectodea spp. and O. tyrtaeum) is definitely possible. Source populations of endogeic and anecic species are also present in boreal forests of central and northern Alberta (Cameron et al. 2007). If endogeic and anecic species are able to establish viable populations in northern forests, there is a potential for significant impacts on soil properties. However, the effects that epigeic species alone can have on forest soil should not be underestimated. The activities of D. octaedra had dramatic impacts on soil structure, chemistry and biodiversity at the aspen and lodgepole pine sites in the Kananaskis Valley (Dymond et al. 1997; McLean and Parkinson 2000a, b). Lumbricus rubellus has a voracious appetite for the organic material of the forest floor (Frelich et al. 2006), and is the species implicated as causing significant changes in plant communities in Minnesota (Hale 2004), including the extirpation of the rare fern Botrychium mormo Wagner in the Chippewa National Forest (Gundale 2002). Thus its widespread presence in coastal rainforests should be of concern. Although poor quality litter and low pH will undoubtedly affect both the rate of expansion of earthworms into Canadian forests, and the diversity and abundance of the ensuing earthworm communities, these environmental characteristics will not exclude earthworms from northern forests. While numerous studies have shown that most species of exotic earthworm prefer deciduous litter and neutral soils, many are also capable of dealing with coniferous litter and low pH, and some can even modify the pH of their environment. In addition, not all species of coniferous litter are equally unpalatable (Reich et al. 2005). Furthermore, even in the largely

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

coniferous boreal forests, early successional stages that follow harvesting or fire have a high component of deciduous trees and shrubs, and riparian areas also provide habitat with more favourable soil and litter characteristics. Thus boreal forests contain a mosaic of patches that are constantly changing over time, providing a range of conditions that differ in their suitability to support earthworms. Even using a conservative estimate of the ability of earthworms to withstand cold winter temperatures, it is clear that based on temperature alone, large areas of Canadian forests are susceptible to invasion. This does not even take into account the potential impact of climate change on earthworm distribution.

Research priorities

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Canada lies within the boreal zone, and boreal forests, with their high accumulation of dead organic matter, have a greater potential for change in soil C relative to other climatic zones (Vogt et al. 1995). The mineralization of even a small proportion of these carbon stores could have far-reaching consequences. However, although earthworms mobilise carbon in the short-term, there is also evidence to suggest that in the longer-term they contribute to stabilization of soil carbon by the formation of stable aggregates and by removing organic material from the soil surface, often to considerable depths in the soil (Bohlen et al. 2004c; Wironen and Moore 2006). This balance between earthworm activities promoting carbon flux, and those contributing to carbon sequestration is critical for understanding how the invasion of exotic earthworms into northern forests will impact global carbon stores.

Impacts of earthworms Studies on the long-term impacts of earthworms on forests, rather than simply their presence in forests at particular points in time, are urgently required. The US experience should give cause for concern and the Kananaskis studies are an excellent beginning, but the forests examined in those investigations represent only a small portion of forest diversity in Canada. In particular, we lack ecological information on distribution and impacts of exotic earthworms in the Atlantic Forest Region and in forests in general in BC, both areas with relatively mild climates, well within the tolerance range for many species of invasive lumbricids. BC has a tremendous diversity of forests types, some of which are already home to indigenous earthworms. Earthworms and carbon flux and sequestration Several studies have demonstrated that earthworm invasion has the potential to enhance carbon flux from northern forests. For example, Alban and Berry (1994) reported carbon losses of *600 kg ha-1 year-1 for a period of 14 years from a mixed hardwood forest in Minnesota, while in an undisturbed sugar maple forest in central New York, earthworms decreased soil carbon storage by 28% (Bohlen et al. 2004b) in the top 12 cm of soil. Canadian soils contain 17% of the world’s total soil carbon (Bhatti et al. 2002). In addition, a large part of

Earthworms and mycorrhizae There have been few studies on the interaction between earthworms and mycorrhizae, and most have focussed on arbuscular mycorrhizae (McLean et al. 2006). Given the overwhelming importance of ectomycorrhizae in coniferous forests and their distribution within the organic layers (e.g., Goodman and Trofymow 1998), there is an urgent need to determine if and how invasive earthworms will jeopardize this vital association between coniferous trees and their fungal partners. Temperature limitations on earthworm distribution Any estimation of the area potentially available for earthworm colonisation is extremely sensitive to our understanding of the relationship between earthworms and temperature. For example, if mean winter temperatures were to increase even slightly in the area, there would a large incremental change in ‘‘acceptable’’ temperatures. The relationship between air temperatures and soil temperatures also deserves attention. Earthworms are directly affected by soil temperatures (which are rarely measured directly), and the relationship between air and soil temperatures cannot be assumed to be the same in different geographical locations (Zheng et al. 1993).

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Taxonomic expertise This paper has only considered exotic lumbricid earthworms of European origin, which are generally well-described in the taxonomic literature. However, it is likely that earthworms from other parts of the world may also be imported into Canada, either intentionally (e.g., for soil remediation or commercial applications) or inadvertently (e.g., in soil associated with horticultural and agricultural products). Hendrix and Bohlen (2002) reported that at least 20 species from Asia, Africa and South America have already been introduced into North America. In order to make informed decisions about the risks posed by invasive earthworms, it is essential to know what species we have. At present the lack of trained earthworm specialists in Canada jeopardizes the work of both ecologists and regulators (Reynolds 2004). Linking the traditional morphologically based taxonomy (in which only adults can be identified with certainty) with the newer molecular techniques will greatly enhance the capacity for rapid, accurate identification of earthworms at all developmental stages. Acknowledgements The valuable comments and suggestions provided by Dr. John Reynolds and Dr. Valin Marshall are gratefully acknowledged. The original map data were provided by the Atlas of Canada http://atlas.gc.ca/
2007, produced under licence from Her Majesty the Queen in Right of Canada, with permission of Natural Resources Canada.

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Ecological impacts of non-indigenous invasive fungi as forest pathogens Judy A. Loo

Originally published in the Journal Biological Invasions, Volume 11, No 1, 81–96. DOI: 10.1007/s10530-008-9321-3 Ó Her Majesty the Queen in Right of Canada 2008

Abstract Non-indigenous pathogenic fungi increasingly threaten North American tree species. Ecosystems may be fundamentally changed when abundant tree species are functionally eliminated, as occurred with American chestnut (Castanea dentata (Marsh.) Borkh.). Conversely, changes may be more subtle but still significant and long lasting when populations are lost, or all trees in the larger size classes are killed. Proposed approaches for characterizing the magnitude of ecological impacts use characteristics of both the non-indigenous pathogen and the host species. Impacts are most significant when highly successful invading pathogens attack foundation species, setting in motion a long-lasting cascade of effects on the host and associated species. Such impacts have generally not been well documented at the ecosystem level. Several North American forest tree species have been functionally eliminated or severely reduced by non-indigenous pathogens. Historical invasions, such as that of chestnut blight (Cryphonectria parasitica (Murr.) Bar), caused very significant ecological impacts that will never be completely understood because of lack of quantitative data. Beech bark disease, caused by a combination of an introduced scale insect

(Cryptococcus fagisuga Lindinger) and a fungus (Neonectria faginata (Lohman et al.) Castl. & Rossman), is still advancing and provides opportunities for studying ecosystem-level impacts when a major tree species is removed or markedly reduced in abundance from the overstorey. Butternut canker, caused by the fungus, Sirococcus clavigignentijuglandacearum N.B. Nair, Kostichka & Kuntz, has spread throughout the host range, endangering species’ survival. Other non-indigenous invasive fungi such as Cronartium ribicola J.C. Fisch and Phytophthora lateralis Tucker & Milbrath continue to move into new populations, causing high mortality and associated restructuring of these ecosystems. Global trade and environmental change trends will ensure new challenges by non-indigenous fungal pathogens, presenting an urgent need for improved understanding of long-term impacts across ecological systems. Keywords Beech bark disease
Butternut canker
Chestnut blight
Dutch elm disease
Ecological impacts
Invasive fungi
Non-indigenous species
Pathogen
Port-Orford cedar root rot
White pine blister rust

Introduction J. A. Loo (&) Natural Resources Canada, Canadian Forest Service – Atlantic Forestry Centre, P.O. Box 4000, Fredericton, NB, Canada E3B 5P7 e-mail: [email protected]

Non-indigenous invasive species constitute one of the greatest global threats to native biodiversity in North America, second only to habitat loss (McNeely 2005;

D.W. Langor & J. Sweeney (eds.), Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems. DOI: 10.1007/978-1-4020-9680-8_6

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Novacek and Cleland 2001; Enserink 1999; Walker and Steffen 1997; Desprez-Loustau et al. 2007). Pimentel et al. (2000) noted that about 400 of the 958 species listed in the US as threatened or endangered are threatened, at least in part, because of nonindigenous invasive species. Vitousek et al. (1996) suggested that biological invasions constitute a significant component of global environmental change. Among the thousands of non-indigenous species that have invaded North America since European colonization, insect pests and pathogens pose the greatest threat to forest ecosystems (Lovett et al. 2006). Vitousek et al. (1996, 1997) claimed that the greatest perturbations to forest ecosystems in the twentieth century resulted from waves of introduced pests and pathogens, and they suggested that nonnative species will continue to be the greatest threat to diversity of eastern North American hardwood forests in the future. These introductions have greatly reduced the frequency and relative dominance of at least seven eastern North American native tree species (Orwig 2002; Liebhold et al. 1995). More than 20 non-indigenous species of pathogens attack forest trees in the US (Pimentel et al. 2000). Many of them also occur in Canada. Non-indigenous fungal pathogens are often highly destructive to species that are closely related to the pathogens’ host species, which are relatively unaffected in the native range of the fungus, presumably due to host–pathogen co-adaptation (Parker and Gilbert 2004). White pine blister rust (Cronartium ribicola J.C. Fisch.) is an example of a pathogen that has devastating impacts on five-needle pines (Pinus spp.) in North America, but in Asia where the fungus and pine species coevolved, the impact is much less and the native Asian pine species have developed a relatively high level of resistance (Kinloch and Dupper 2002). Similarly, chestnut blight (Cryphonectria parasitica (Murr.) Bar) is very destructive to American chestnut (Castanea dentata (Marsh.) Borkh.), but Japanese chestnut (C. crenata Sieb. et Zucc.) and some Chinese chestnut (C. mollissima Blume) trees are relatively resistant to the disease (Anagnostakis 2001). Although it has been proposed that selection should favor less aggressive strains of non-indigenous pathogens over time, there is little or no empirical evidence of this occurring (Jarosz and Davelos 1995; Parker and Gilbert 2004; Brasier and Buck 2001).

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Examples of non-indigenous invaders that have changed ecosystem function and the fundamental nature of ecosystems are numerous (Mack et al. 2000). However, the magnitude of impacts of particular historical invasions remains unclear, even when the influence of an invader led to the virtual loss of a native species (Parker et al. 1999). This is in part because the ecological importance of the native species prior to the invasion was generally not quantified. The extent and ecological impacts of introduced fungi are unclear because of the lack of baseline information on fungal communities (Desprez-Loustau et al. 2007). Although introduced pathogenic fungi have received much more attention than non-parasitic fungal species, knowledge of impacts at the community or ecosystem level remains incomplete. Loss of foundation or keystone species is an extreme example of a change in the fundamental nature of an ecosystem. Ellison et al. (2005) identified several foundation species, including trees, that have been or are currently in the process of being effectively removed from forest ecosystems as a result of non-indigenous fungal invasions. They defined foundation species as those that are locally abundant and regionally common, usually occupying low tropic levels, and stabilizing fundamental ecosystem processes such as productivity and water levels. They noted that, in many ecosystems, a single foundation species controls ecosystem dynamics and modulates processes. In forest ecosystems, trees are usually foundation species because they define the structure and microclimate (Ellison et al. 2005). Loss of such species has serious and long-term impacts on energy and nutrient fluxes, hydrology, food webs, and biodiversity. A successful non-indigenous invasive species becomes part of a pre-existing complex ecosystem and may influence native species in unpredictable ways, whereas native species have concurrent effects on them, often ultimately stabilizing them. Castello et al. (1995) asserted that pathogens, whether native or not, regulate and are regulated by patterns and processes in forest ecosystems. They noted that the magnitude of pathogenic species’ impacts on forest ecosystems is often not recognized. Many nonindigenous pests and pathogens do not become serious problems. For example, a scleroderris pathogen, Gremmeniella abietina (Lagerb.) Morelet, introduced from Europe caused alarm in northeastern

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North America, but outside of plantations, it never became a serious pest (Liebhold et al. 1995; Laflamme et al. 1998). Non-indigenous pathogens are affected by other drivers of global change, often in ways that increase their invasive potential, because of the destabilizing effect of changing climate and pollution, and the greater vulnerability of simplified ecosystems in general (Liebhold et al. 1995; Vitousek et al. 1996; Walker and Steffen 1997; Mack et al. 2000; Anderson et al. 2004; Harvell et al. 2002). In this article, I review approaches for understanding ecological effects of non-indigenous fungal pathogens and the documented impacts of such species on the different levels of ecological organization from genes to ecosystems. Examples are described throughout the article and several case studies are discussed in greater detail to illustrate types of ecological impacts.

Approaches for characterizing ecological impacts of non-indigenous species The ecological effects of non-indigenous pathogenic invaders may be limited in time and space, affecting only the host species. Alternatively, there may be additional indirect effects on a large number of nonhost species, which may extend over multiple generations. The direct and indirect impacts are greatest when one or more foundation or keystone species are lost from ecosystems, resulting in cascading effects (Ellison et al. 2005). Conversely, the impacts of a nonindigenous pathogen may be small in terms of overall ecosystem integrity, even when host mortality results and when the effect is long term (Parker et al. 1999). Such is the case when an invader influences directly only one species, which is a relatively minor constituent of the ecosystems it inhabits. Short-term impacts include defoliation, seed crop failures, and temporary shifts in energy flows within ecosystems; long-term impacts include widespread host mortality and reduction in frequency, with subsequent replacement by other species (Lovett et al. 2006). Non-indigenous invasive fungi provide examples for each of these types of ecological effects. Long-term impacts of non-indigenous pathogens involve a change in ecosystem state such as shifting species composition and altering successional pathways with consequent changes in ecosystem processes.

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Chestnut blight provides a dramatic example of longterm impacts, including direct and indirect effects on species. The direct effect is the effective removal of the tree species itself and the consequent shift in tree species composition. Smock and MacGregor (1988) provided an example of long-term indirect impacts. They examined the potential for lower leaf shredder activity by aquatic invertebrates in streams, as a result of a compositional shift from chestnut to oak (Quercus spp.), and concluded that such long-term indirect effects on non-host species were likely. Parker et al. (1999) discussed the need for a consistent and quantitative approach to assessing impacts of non-indigenous invasive species. They used three measures to describe the level of impact of such species: the range, abundance, and per capita or per biomass effect of the invader. They suggested a multiplicative relationship between the three to describe the impact severity of an introduced species on the host ecosystem. Range and abundance of both the pathogen and the host influence the degree or gravity of ecological impacts at the ecosystem level. Characteristics of both the invasive species and the host were considered in a more comprehensive approach, proposed by Lovett et al. (2006). They identified sets of characteristics of the invader and the host tree by which to quantify ecological impacts. Characteristics of the pathogen or pest include mode of action, host specificity, and virulence. Those of the host are importance (dominance), uniqueness, and phytosociology (species’ associates and successional status). Three modes of action of pathogens are exemplified by the prevalent symptoms: defoliation, root rot, and stem cankers. Ecological effects of introduced pathogens causing root rot, Phytophthora lateralis Tucker & Milbrath, for example, and stem cankers, such as Sirococcus clavigignenti-juglandacearum N.B. Nair, Kostichka & Kuntz, the causal agent of butternut canker, tend to be more severe than of those causing defoliation. Host specificity is often high for forest tree pathogens. For example, pathogens causing beech bark disease (Neonectria faginata (Lohman et al.) Castl. & Rossman [previously Nectria coccinea var. faginata]), butternut canker (Sirococcus clavigignenti-juglandacearum), Dutch elm disease (Ophiostoma ulmi (Buisman) Nannf. and O. novo-ulmi Brasier) and chestnut blight, each have devastating impacts, and are restricted in North America to a single tree species or a low number of native species (Ehrlich 1934; Ostry

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1995; Ostry and Kastovich 1997; Schlarbaum et al. 1997; McKeen 1995). In each of these cases, virulence is high with widespread mortality, but the range and abundance of the pathogen is limited by the range and abundance of the host species. American beech (Fagus grandifolia Ehrh.) and American chestnut were both very abundant within the species’ ranges, but both are limited to temperate hardwood forests of eastern North America. Butternut (Juglans cinerea L.) has a similar range but much lower natural abundance within its native range, whereas elm species (Ulmus spp.) are more broadly distributed but occurred with relatively lower frequency than the first two, even before the introduction of Dutch elm disease. In contrast, white pine blister rust results in widespread mortality of all of the Hapoxylon (five-needled) pines (Liebhold et al. 1995), and a relatively new, not yet widespread pathogen, Phytophthora ramorum S. Werres, A.W.A.W. de Cock & W.A. Man, is lethal for Fagaceae species (Quercus spp. in the red oak group and Lithocarpus densiflora (Hook. & Arn.) Rehd.). It also causes damage ranging from shoot dieback to leaf blights on up to 100 other North American species, including Ericaceous shrubs and commercially important conifers (Goheen et al. 2006). Both of these introduced pathogens have had or potentially will have devastating impacts throughout North America. Ellison et al. (2005) used the concept of foundation species to describe the consequences of non-indigenous invaders. Dominant and co-dominant trees are ‘‘foundation species,’’ as their architectural, functional, and physiological characteristics define forest structure and alter microclimates, and their biomass and chemical makeup contribute greatly to ecosystem processes (Jones et al. 1997). High mortality of a widespread, abundant foundation species has a much greater ecological effect than the same mortality rate for a less common species (Lovett et al. 2006). However, if the less common species has a unique role, providing ecosystem services that are not provided by other species in the ecosystem, the impact is disproportionately greater (Parker et al. 1999).

How do alien invasive pathogens affect forest ecosystems? Impacts of non-indigenous invaders are expected to be greatest when the invading species performs a new

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ecological function in the invaded ecosystem (Parker et al. 1999). A new ecological function, although apparently minor, may result in a cascade effect. For example, the introduced beech scale, Cryptococcus fagisuga Lindinger does not apparently cause tree mortality or even significant damage alone, but it predisposes American beech trees for attack by the fungus Neonectria faginata (Houston 2005), which in turn kills or seriously debilitates the tree, resulting in a cascade of ecological effects. Crooks (2002) discussed the concept of invaders as ‘‘exotic engineers,’’ and invaders with greatest impact are those that directly modify ecosystems and have cascading effects. These cascading effects may include altering system-level flows and availability or quality of nutrients, food, and physical resources (living space, water, heat, light). ‘‘Ecosystem engineers’’ create, modify, and maintain habitats, changing the availability of resources used by other taxa (Jones et al. 1994). Little attention has been paid to date to the control of physical resources through modification of habitats by non-indigenous invasive species, particularly fungi, but invasive pathogens, such as the causal agent of chestnut blight, are clearly ecosystem engineers by this definition. Non-indigenous species are often considered to reduce biodiversity (Vitousek et al. 1996; Altizer et al. 2003). For example, McNeely (2005) stated that increasing global domination by a relatively few invasive species threatens to create a more homogeneous world rather than one characterized by great biodiversity and local distinctiveness, thus affecting the capacity of ecosystems to adapt to changing conditions. Crooks (2002) pointed out, however, that exotic ecosystem engineers may not always simplify ecosystems. Those that increase habitat complexity increase species richness whereas those that reduce the complexity reduce the species richness. Engineering impacts are often greatest when the modified resource flows are used by many other species or when there is an effect on abiotic factors, such as soil or hydrology (Jones et al. 1994). Pathogen-induced changes in leaf litter composition or rooting depth have such effects. The impacts on host species of invaders, such as non-indigenous fungi, as ecosystem engineers may be measured at five levels (Parker et al. 1999): (1) individuals, including demographic factors such as growth and survival; (2) genetic effects; (3)

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

population dynamics effects; (4) community effects including species richness, diversity, and tropic structure; and (5) effects on ecosystem processes, such as nutrient cycling and primary productivity. In examining effects of non-indigenous fungi in forest ecosystems, research has been focused mainly on impacts on host species, and sometimes on the cascading effects on other species that depend on the host, but little attention has been paid to the effects on native fungal species (Desprez-Loustau et al. 2007). Effects of pathogens that cause serious damage have been studied and can be categorized in various ways. However, there are undoubtedly many invasive fungal species that are unknown because their impacts are small, at least with respect to their effects on trees. In some cases, native fungal species are adversely affected as a result of competition by a new aggressive pathogen that may effectively exclude the host species as a resource for more benign co-adapted fungal associates. Alternatively, native pathogens and saprophytes may benefit from mortality (of a tree or tree part) caused by nonindigenous species. Effects on host species, at the individual level, are obvious for many non-indigenous pathogenic fungi, for example individual tree mortality, or reduced growth or reproduction rates. Genetic effects are less obvious, little studied (Parker et al. 1999), but no less real. The selection pressure imposed by non-indigenous pathogens may result in changed gene frequencies and, in the case of high mortality, loss of genetic diversity in the host species (Altizer et al. 2003). This may have long-term evolutionary effects, reducing the capacity of the host species to respond to other environmental changes over time. When a pathogen eliminates populations of a host species, gene flow is disrupted and genetic diversity of remaining populations is gradually lost over subsequent generations. Genetic effects may be exacerbated by human activities, if healthy trees are selected over diseased ones for harvesting. The third level named by Parker et al. (1999), changes in population dynamics, is a common outcome of invasion by non-indigenous pathogens. Most commonly, this is observed as changes in local abundance of host species, which in turn affect population sizes. There may be more subtle changes at the population level as well. Population level changes occur, for example, when few trees survive

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to reach reproductive maturity and most regenerating individuals are of root sprout origin. Under such conditions, the oldest age classes are lost and population structure may tend toward groups of clones rather than mixtures of less closely related individuals. This is observed in beech stands in northeastern North America where beech bark disease has caused heavy damage (Ostrofsky and McCormack 1986). Vitousek (1990) described three ways by which an invasive species may alter communities or ecosystems: by differing substantially from native species in their resource use; by changing the trophic structure of the invaded ecosystem; and by changing disturbance frequency and/or intensity. Although his work focused on animal and plant invasions, pathogens may be categorized using the same principles. For example, resource use by a non-indigenous fungal species such as C. parasitica differs from that of the native fungi associated with chestnut, in its aggressiveness, effectively eliminating its host in a short period of time. Trophic structure is changed through indirect means, through the cascading effects of the loss of a major food source for other species. When high-quality hard-mast sources are removed from an area, or are substantially diminished, as has occurred in areas where the native ranges of chestnut, beech, and butternut overlap, top trophic level species such as black bear (Ursus americanus Pallas) and marten (Martes americana Turton) may be affected (Jakubas et al. 2005). Another community-level change, as a result of impacts of invasive pathogens, is associated with the loss of canopy species in forest ecosystems, which may in turn result in greater or lower species diversity. The loss of chestnut from forest canopies in large areas of eastern North America resulted in increased tree species diversity because several species filled the niche abandoned by chestnut (Castello et al. 1995). Alteration of disturbance frequency and intensity is the most obvious ecosystem-level impact of nonindigenous fungal invaders that cause widespread mortality. Widespread mortality of a dominant or codominant species constitutes a severe disturbance to forest ecosystems. Dieback and mortality of conifer species in fire-prone areas contribute to the fuel load, indirectly affecting a second disturbance agent by increasing and intensifying the fire cycle (Castello et al. 1995). Despite the availability of severity

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evaluation frameworks such as this one, even when the impacts can be deduced to be extreme, for example when an ecologically important tree species is eliminated from ecosystems over a very broad area, the actual impact to the ecosystem has rarely been quantified and long-term implications are still debated. Dukes and Mooney (2004) noted that ecological impacts of alien species include large-scale effects, such as altering geomorphology, fire regime, hydrology, microclimate, atmospheric composition, nutrient cycling, and productivity. Plant pathogens that attack dominant forest trees can, at least temporarily, reduce the system’s productivity and live biomass. Thus, it is useful to categorize the ecosystem-level effects of non-indigenous pathogens in terms of their impacts on foundation species. Introduction of invasive pathogens is resulting in decline of foundation tree species throughout the world.

Case studies: non-indigenous fungi-mediated loss or reduction of foundation tree species and associated effects During the past 120 years, at least seven significant new fungal diseases have had serious impacts on native tree species of importance in North America forests. Six of these are summarized in Table 1. Several non-indigenous species attack more than one tree species. In Canada, a country with only about 140 tree species, at least one-tenth of the native tree species are subject to serious damage and mortality caused by non-indigenous pathogens Chestnut blight American chestnut was once a dominant or codominant canopy tree over much of eastern North America (Russell 1987). It occurred on a variety of sites, usually on well-drained moist sandy soils (Waldron 2003), often in association with oak species (Ellison et al. 2005). It was a prolific nut producer and provided an important source of wildlife mast (Waldron 2003). Chestnut blight was introduced from Asia in the late 1800s and was first observed in the US in 1904 (Bramble 1934). The fungus was thought to have arrived on nursery stock (McKeen 1995;

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Anagnostakis 2001), and had a rapid and devastating impact on eastern hardwood forests. In 30 years it had killed most of the mature native chestnut in the middle Atlantic and New England States (Bramble 1934). Nelson (1955) found that the basal area of chestnut had declined from 41% to less than 1% in the previous 18 years in western North Carolina. When American chestnut was effectively extirpated by the fungal pathogen within a few decades of the fungus’ appearance (Mack et al. 2000), a variety of mostly hardwood species replaced it. McKeen (1995) called the loss of the chestnut ‘‘the greatest single catastrophe in the annals of forest history.’’ Two decades after its appearance in New York, the disease had spread into southern Ontario and had caused similar destruction there (McKeen 1995). Nutrient cycling has been altered by chestnut blight because chestnut wood is slower to decay than any of the other hardwoods, such as oak species, that were associated with it (Ellison et al. 2005). Chestnut leaves decay more rapidly, however, and have higher nutritional value than oak. Woods and Shanks (1959) reported that in the Great Smoky Mountains National Park, 41% of all replacements of chestnut were oak species. They found that pure stands of hemlock (Tsuga canadensis (L.) Carr.) replaced chestnut in some areas, thereby changing the ecology of such sites significantly. Nelson (1955) reported that tuliptree (Liriodendron tulipifera L.) was the most frequent replacement species in western North Carolina. A study of the effect of naturalized chestnut at a site 600 km from the species’ natural distribution revealed that on sandy loam soils, total soil carbon and nitrogen, and net mineralization and nitrification rates were 10–17% higher under chestnut canopies than under mixed hardwood canopies (Rhoades 2006). These results indicate that nutrient status of soils may have declined as a result of the loss of chestnut from eastern hardwood forests. Ecological impacts of chestnut blight are numerous, but mostly not quantified. The loss of the highquality food source associated with effective extirpation of chestnut, undoubtedly influenced wildlife populations. Diamond (2000) estimated a 34% reduction in hard-mast output after chestnut blight killed mature chestnut trees, seriously decreasing the carrying capacity for some wildlife species. Smock and MacGregor (1988) looked at the impact on water courses of the replacement of chestnut by other

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Table 1 Examples of non-indigenous fungal pathogens and their ecological impacts in forest ecosystems Pathogen name and mode of action

North American host(s)

Impact on host species

Ecological impact: dominance, successional status, uniqueness

Chestnut blight (Cryphonectria parasitica)

American chestnut (Castanea dentata)

Very high mortality

Very high: devastation and conversion of 25% of eastern hardwood forest

Girdles stem

White pine blister rust (Cronartium ribicola)

Loss of wildlife food source

Five-needle pine species (Pinus spp.)

High mortality

High: conversion of pine forest; very high in montane regions

Elm species (Ulmus spp.)

High mortality

Vascular wilt, kills branches

Butternut canker (Sirococcus clavigignentijuglandacearum)

Butternut (Juglans cinerea)

Girdles stem

Beech bark disease (Neonectria faginata)*

Loss of dominant mid-successional tree species with unique ecosystem services

Stem cankers kill trees of all ages

Girdles stem

Dutch elm disease (Ophiostoma ulmi and O. novo-ulmi)

Highly virulent and aggressive pathogen effectively eliminated species

Highly virulent, moderately to highly aggressive pathogen reduces or destroys populations Stem cankers (rust) kill trees of all ages Highly virulent, variably aggressive pathogens eliminates large trees Crown death kills trees after age of reproductive maturity Very high mortality Highly virulent and aggressive pathogen effectively eliminates species Stem cankers kill trees of all ages

American beech (Fagus grandifolia)

Girdles stem

Reduction or loss of dominant early to late-successional species; very broad geographic area Loss of crucial food source for Clark’s nutcracker, grizzly bear Medium: loss of large trees, localized ecological importance Reduction of scattered dominant mid-successional species over very large geographic area Loss of food source and habitat for some bird species Medium: effective loss of species; localized ecological importance Loss of early successional sub-dominant species with scattered occurrence in eastern hardwood forests Loss of food source for birds and mammals

High initial mortality; severe damage in aftermath forest

Medium: conversion to short-lived, short-stature, defective trees

Highly virulent, moderately aggressive pathogen kills trees after age of reproductive maturity

Loss of dominant, late-successional species in eastern hardwood forest Reduction of important food source for birds and mammals

Stem cankers Port-Orford cedar root rot (Phytophthora lateralis) Root rot

Port-Orford cedar (Chamaecyparis lawsoniana

High mortality Highly virulent and aggressive pathogen Root rot kills trees of all ages

Medium: severe reduction of species with unique ecological role; localized ecological importance Loss of dominant, late-successional tree species in specialized habitat Loss of nutrient source

*Until recently, known as Nectria coccinea var. faginata; thought to be introduced. Origin of the pathogen is now in question, but the disease develops only in the presence of an introduced insect

species, in terms of processing rates and nutritional quality of chestnut leaves, and impacts on shredding macro-invertebrates. The forest that replaced chestnut has a greater diversity of tree species, so local biodiversity may have increased as a result of the disease (Castello et al. 1995), but the functional

loss of a species with such ecological dominance undoubtedly influenced species diversity at a variety of trophic levels, including mast-dependent birds and mammals and their associated predators, parasites, and pathogens, as well as soil micro flora and fauna.

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A gradual resurgence of chestnut has been observed in Ontario, and is thought to be the result of very low ambient levels of the fungal pathogen (McKeen 1995). The fungal population crashed as the great majority of its host trees died, allowing more trees to escape infection now, but likely C. parasitica will experience resurgence as well if chestnut recovers to a significant degree. Beech bark disease American beech was also a dominant or co-dominant canopy tree in temperate hardwood forests of eastern North America, and still maintains dominance beyond the area infected by beech bark disease. Although it can form pure stands, it is usually found in mixtures with species such as sugar maple (Acer saccharum Marsh.), yellow birch (Betula alleghaniensis Britt.), and red spruce (Picea rubens Sarg.) and attains best growth on well-drained alluvial soils, but occurs over a wide range of site conditions (Tubbs and Houston 1990). In the northern part of its range, beech is the only common nut-producing tree and the species provides an important source of mast for a variety of birds and mammals. Beech bark disease is caused by the combined effects of a non-indigenous scale insect, Cryptococcus fagisuga and one of two bark-killing fungal species. Although the disease is often initiated by the native species, Neonectria galligena Bres., N. faginata usually becomes more prevalent soon after the disease has established. N. faginata may or may not have been introduced (Castlebury et al. 2006; Houston 2005). It behaves as a non-indigenous species, because it becomes a problem only in association with the non-indigenous beech scale. The insect was introduced to Halifax, Nova Scotia, more than 110 years ago, and there have probably been one or more subsequent introductions in northeastern US (Houston 2005). Three phases of the disease represent different ecosystem states (Shigo 1972). The ‘‘advancing front’’ is the term for the movement and rapid build up of insect populations in newly invaded beech stands. Ecosystem change is minimal in this phase, but the stage is set for a cascade of effects initiated by high-density infestation by the scale insect, predisposing the trees for the fungal infection that follows, often resulting in 50– 85% mortality of large trees within 10 years at the

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‘‘killing front’’ (Houston et al. 1979; Houston 1994; Witter et al. 2005). The third phase, called the ‘‘aftermath zone,’’ consists in large part of small trees mainly of root sprout origin, that over time are attacked repeatedly by endemic levels of beech scale and Neonectria fungus. Thickets of often distorted and defective beech stems of low vigor may result (Shigo 1972; Houston 1975). Ecological impacts of beech bark disease are sequential, beginning with the loss of a dominant species from the canopy at the killing front. The disease is most severe in nearly pure stands of mature beech. Immediate ecological impacts are associated with a dramatic shift in forest structure (Storer et al. 2005) resulting from the concurrent loss of canopy trees and a proliferation of root suckers from dead and dying trees (Fahey 1998), an increase in coarse woody debris as dead trees fall (Paipak et al. 2005), and the loss of hard-mast source for wildlife (Storer et al. 2005). Black bear dependence on hard mast and beech nuts, in particular, in areas where beech dominates among nut tree species, has been studied by several researchers (Faison and Houston 2004). Jakubas et al. (2005) reported that 80% of female bears at a study site in Maine reproduced following good nut crops whereas only 20% reproduced following years of poor nut crops. As the aftermath forest develops from root suckers and seedlings, where causal agents of the beech bark disease are present at endemic levels, a shift in species succession is observed in some cases. For example, DiGregorio et al. (1999) found that sugar maple is stimulated by as much as 30% in the understorey due to the thinned canopy. Beech leaves have higher lignin concentration and break down more slowly than sugar maple and yellow birch leaves (Lovett et al. 2006). Replacement of beech by maple results in reduced forest-floor mass, increased nitrification, increased leaching of nitrate into stream water, and decreased retention of atmospherically deposited nitrogen, because soil organic matter formed from maple has a higher propensity for nitrification than that of beech (Lovett and Rueth 1999; Lovett et al. 2006). In other cases, there is a gradual shift to a greater proportion of hemlock and yellow birch in the overstorey (Papaik et al. 2005). A shift to hemlock probably will result in the opposite effect from that of a shift to maple, with increased forest-floor mass, and less nitrogen loss (Lovett et al. 2006).

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

Beech thickets may persist, however, constituting a high proportion of stand composition with little change in frequency of tree species, and nut production may rebound even to previous levels (McNulty and Masters 2005), but the trees usually do not attain dominance in the crown canopy and are easily broken, rendering the nut crop less available for black bear than in healthy stands. The main ecological effect in such stands is related to a shift in forest structure and age-class distribution. Papaik et al. (2005) found that in NY, diseased beech was almost twice as likely to experience stem break, than uprooting in severe windstorms. Stem break frequency was a function of disease severity. In the southeastern US, beech in intermediate size classes have high resistance to stem break, until affected by beech bark disease. Overall coarse woody debris increased two- to four-fold in stands affected by the disease. Even without severe windstorms, results reported by Papaik et al. (2005) show that the effect of beech bark disease on beech stands is to make them more disturbance prone, with a higher rate of canopy tree turnover and gap formation. Krasny and Whitmore (1992) examined gap formation in northern hardwood forest and reported that beech had a disproportionate importance in relation to its abundance in affected stands, with almost all gap formation being gradual rather than sudden. They found a two-fold difference between proportion of gap-maker trees that were beech and the frequency of the species in the stands. The implications of gradual gap formation are not well studied, but probably understorey shifts are less dramatic in stands with gradual rather than sudden gap formation. Long-term effects of beech bark disease on northeastern North American forest ecosystems are likely to be substantial, but the full implications are yet to be understood as the disease is still in progress. For example, it is unknown whether aftermath-zone forests will be stable over time, or if subsequent outbreaks will occur, resulting in periodic elevated mortality. Some of the diseased but relatively vigorous trees in the aftermath-zone may attain large size and canopy dominance, or they may be destroyed in windstorms because of a greater tendency to stem breakage. Continued evaluation of ecosystem change and beech development in aftermath-zone forests is thus important. The

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modeling approach proposed by Le Guerrier et al. (2003) may prove useful in identifying and understanding long-term impacts. Genetic resistance of beech to beech bark disease has been reported, but this resistance is indirect through resistance to the beech scale (Houston 2005; Koch and Carey 2005; Ramirez et al. 2007). Research is underway on a number of fronts to understand the mechanism and mode of inheritance of the resistance, to develop a somatic embryogenesis protocol, and to identify molecular markers. Butternut canker Butternut is another eastern North American, mastproducing, temperate hardwood species that has been devastated by an invasive pathogen. The species has best growth in riparian areas, and rich upland forests, though it also grows on dry outwash sands and gravels (Farrar 1995; Waldron 2003). The loss of the species is less significant ecologically than the loss of chestnut or beech as it rarely grows as a pure stand or a dominant canopy species, and is usually associated with a variety of other temperate hardwoods including mast-producing species such as oaks (Ostry and Pijut 2000). Butternuts are highly nutritious and are eaten by mice, squirrels, deer, other wildlife species, and humans (Waldron 2003). In areas without other high-quality wildlife mast sources, the species provides an important ecological service (Ostry et al. 1994). The species has been in decline for most of the past century, and was first reported to be in trouble in the US in 1923 (USDA Forest Service 1979). Widespread mortality in Wisconsin was first attributed to the fungus, Melanconis juglandis (Ellis & Everh.) Graves var. caryae Wehmeyer. However, in 1967, a previously unknown fungus, Sirococcus clavigignenti-juglandacearum was found to be the causal agent for butternut canker (Renlund 1971). The role and taxonomy of S. clavigignenti-juglandacearum was not fully understood until after extensive research and determination that the fungus merited description as a new taxon in 1979 (Nair et al. 1979). The origin of the species is not known but it is thought likely to be non-indigenous in North America because of its sudden appearance, its aggressiveness, and the lack of genetic variability (Furnier et al. 1999).

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By the early 1990s, infected butternut had been found throughout most of the species range (Ostry 1995). In Canada, the first report of this species was in Quebec in 1990 (Innes and Rainville 1996), then in Ontario in 1991 (Davis et al. 1992), and finally in New Brunswick in 1997 (Harrison et al. 1998). New Brunswick may have the last uninfected populations, but recent work there indicated that the disease is continuing to spread (Hopkins et al. 2001; Harrison et al. 2005). Butternut was listed as endangered on Canada’s Schedule 1 (list of wildlife species at risk), in July 2005 under the federal Species at Risk Act. In the US it is listed as a ‘‘species of Federal Concern’’ by the Fish and Wildlife Service (Schlarbaum et al. 1997). Butternut canker is reminiscent of chestnut blight, in that it occurs on trees of all ages and on all sites. Infection can occur through buds, leaf scars, and wounds (Ostry 1995) and spores may be spread by rain splash, birds, insects, or on seeds (Innes 1998). The fungus usually starts on small branches and twigs in the crown, although it may also be observed on the trunk and near the root collar. The pathogen is very aggressive and has spread rapidly since it was first detected in 1967 (Ostry 1998). Although it has been found on black walnut (Juglans nigra L.), infection has been limited (Ostry 1997; Ostry and Kastovich 1997). Trees are killed when multiple cankers girdle the stem or when much of the crown has been killed by branch-girdling cankers. Sprouts may develop but are also infected and are usually killed within a few years. In Wisconsin, Ostry and Woeste (1994) reported that in a span of 15 years, butternut declined by 58%, and by 84% in Michigan. In the US, the Forest Service estimated that in 1995, 77% of the butternut had been killed in the southeast (Schlarbaum et al. 1997), whereas in Canada mortality has been estimated to be 80% in Ontario (Fleguel 1996). Many of the surviving trees are already infected, so mortality will likely rise with time. There are few sprouts from dying trees and those that occur succumb to the disease. Seed production is reduced by crown dieback after infection, and the seedlings that are produced, are commonly infected as well. Genetic resistance may be present at very low frequency in natural populations (Ostry and Woeste 2004), but populations are being lost rapidly with their repositories of unique genetic variability (Schlarbaum et al. 1997).

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White pine blister rust Five-needle (Hapoxylon) pines include economically important white pine species in eastern and western North America, as well as several species with smaller ranges and lower economic importance, but having high ecological value. The white pine group extends into Mexico with large populations of economically and ecologically important species such as Pinus ayacuahuite Shaw and P. strobiformis Engelm. Five-needle pines inhabit a variety of sites, from moist, relatively rich, well-drained soil in eastern and western river valleys, to high-elevation windswept sites with poor shallow soil (Farrar 1995). Their prolific seed crops provide an important source of food for many wildlife species. All five-needle pines are susceptible to an introduced fungal stem rust, known as white pine blister rust (Kinloch 2003), which is thought to have been first brought to North America from Europe in the 1890s (Spaulding 1922; Liebhold et al. 1995). It arrived in western North America on a shipment of eastern white pine seedlings, grown in France, in 1910 (Allen and Humble 2002). The fungus has a complex life cycle involving five spore stages and an obligate alternate host in the genera Ribes, Pedicularis and Castilleja (Liebhold et al. 1995; McDonald et al. 2006). Blister rust rapidly girdles and kills shoots or stems of young trees. In an infected area, seedlings and saplings are destroyed and many mature trees die. The disease spreads unevenly, becoming established where conditions are appropriate, and eliminating or seriously reducing populations, resulting in the creation of metapopulation patterns (Kinloch 2003). The degree to which the artificial imposition of metapopulation structure affects genetic and other ecological processes has not been evaluated. Whitebark pine (Pinus albicaulis Engelm.), a foundation species, is native to high elevations of the US and southern Canadian Rockies. It has an important role in hydrology and its loss results in increased flashiness of streams and altered dynamics of wildlife populations (Ellison et al. 2005). Whitebark pine populations are being destroyed by a combination of factors, most importantly, white pine blister rust. Less than one tree in 10,000 is rust resistant according to Vitousek et al. (1996) and mortality is occurring throughout the species’ range.

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

The demise of whitebark pine over large areas of its native range is likely to have catastrophic impacts on a number of dependent species. Grizzly bear (Ursus arctos L.) depends on the large, highly nutritious seed produced by whitebark pine, and the loss of the tree species hinders bear recovery in the western US (Kendall 1995). The highly coevolved relationship between whitebark pine and Clark’s nutcracker (Nucifraga columbiana Wilson) (Lanner 1999) has given the species a unique ecological role, and its occurrence as pure stands at or near the tree line ensures high ecological importance. Clark’s nutcracker feeds on the seed, tearing apart closed cones to do so. The cones do not open if left alone. Many of the seeds harvested by the bird are cached for winter, just under the surface of the soil and all regeneration of whitebark pine is thought to result from the cached seed (Hutchins and Lanner 1982). Because of the nutcracker’s high dependence on whitebark pine seed, as populations of the tree species drop when they succumb to rust infection, bird populations are also expected to decline. Limber pine (Pinus flexilis James) occupies similar but usually not overlapping environments as whitebark pine. It also has a mutual dependency on Clark’s nutcracker (Lanner 1999), and is also declining rapidly because of white pine blister rust. Limber pine cones open readily, dropping seed, but the only means of dispersion aside from gravity is by birds or animals. The species has a classic metapopulation distribution with populations periodically eliminated by fire and re-established by Clark’s nutcracker planting and forgetting seed caches, up to 22 km from the source tree (Vander Wall and Balda 1977; Tomback and Kramer 1980). Langor (2007) reported that mortality due to the introduced blister rust increased dramatically between 1996 and 2004 to approximately 50% in two areas of Alberta. He noted that the low regeneration rate in heavily infected stands means that mortality exceeds recruitment, leading to general decline of the species in its Canadian range. Loss of limber pine will have significant effects on their montane habitat as the species is adapted to harsh conditions where it is often the only tree species providing cover and food for wildlife as well as modulating hydrology. The seeds are large and nutritious, and are an important source of food for black and grizzly bears (Langor 2007).

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Eastern and western white pine (Pinus strobus L. and P. monticola Dougl.) are both foundation species in their respective ranges and populations are being lost to the introduced pathogen. According to LaFlamme and Hofacker (2001), white pine populations dropped dramatically in many areas; in Minnesota, old-growth white pine covered 3,500,000 acres in 1837, but dropped to less than 60,000 acres by 1990, in part due to blister rust. From 1917 to 1970, there was a 60% reduction in white pine in Quebec. In western US, up to 95% of original western white pine stands have been killed or damaged (Liebhold et al. 1995). Areas in both eastern and western North America that formerly produced prime white pine timber are no longer managed for white pine because of the pathogen. Genetically based resistance has been found with low frequency in both eastern and western white pine and breeding programs have met with some success (Liebhold et al. 1995; Sniezko 2006). Sugar pine (Pinus lambertiana Dougl.) is also very susceptible to the pathogen. Sugar pine trees attain the largest size among all pine species in the world. In addition to the effects on forest structure and the ongoing mortality of the species, Millar et al. (1996) predicted that the impact on the genetic diversity of sugar pine would be highly significant within the next 50–60 years. According to Liebhold et al. (1995), only approximately 5% of the original stands of sugar pine remain undamaged. Damage often involves top death and loss of branches resulting in a shorter and thinner canopy. Recently, the first report of a natural blister rust infection on Rocky Mountain bristlecone pine (Pinus aristata Engelm.) served as a reminder that the pathogen continues to advance and threaten additional species (Blodgett and Sullivan 2004). Dutch elm disease Six elm species are native to eastern North America, and are found on a variety of soils from well-drained to periodically inundated, usually mixed with other temperate hardwood species such as maple, basswood (Tilia americana L.), and butternut (Farrar 1995). The best known of these species is white or American elm (Ulmus americana L.), which has one of the largest natural ranges of any North American tree (Gibbs 1978). Even before the introduction of Dutch

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elm disease, it was seldom found in pure stands; common associates are red and silver maple (Acer rubrum L., A. saccharinum L.), black and green ash (Fraxinus nigra Marsh., F. pennsylvanica Marsh.), balsam poplar (Populus balsamifera L.), various oak species, birch species (Betula spp.) and other species of elm (Bey 1990). White elm has been planted extensively in rural as well as urban settings (Hubbes 1999). The prolific seed production by healthy elm trees gives the species an important ecological role because the seed is consumed by many species of birds and small mammals. The branches provide important nesting sites and deer and rabbits browse on twigs and buds (Waldron 2003). White elm leaf litter decomposes more rapidly than that of maple and oak associates. The leaves have relatively high contents of potassium and calcium. White elm is considered to be a ‘‘soilimproving’’ species because of the combination of rapid leaf decomposition and the release of desirable nutrients during leaf processing (Bey 1990). Dutch elm disease has eliminated large elm trees throughout much of North America (Farrar 1995). In many populations, the disease has continued at epidemic levels until almost all trees have died (Gibbs 1978). The disease is caused by two fungal strains, a non-aggressive strain, Ophiostoma ulmi, and an aggressive strain, O. novo-ulmi (Sinclair and Lyon 2005). The non-aggressive strain was first introduced from Europe to the US in the late 1920s. It had made a dramatic appearance in Europe a few years earlier (Gibbs 1978). The aggressive strain appeared in North America in the 1940s and was carried to Great Britain and western Europe in the late 1960s, where it had a devastating impact on mature elm trees (Gibbs 1978). In eastern North America, as in western Europe, the more aggressive species, O. novo-ulmi has out-competed O. ulmi and now dominates the pathosystem (Houston 1985). O. ulmi, no longer is found in Great Britain. When trees are infected, they may die in a single year or decline for several years before succumbing (Sinclair and Lyon 2005). The overt signs of wilting leaves are an indicator of internal symptoms which shut down the vascular system. The disease depends on one or more species of bark beetles for transmission of fungal spores from dead or diseased trees to healthy host trees (Gibbs and Wainhouse 1986; Gibbs 1978; Sinclair and Lyons 2005). As for beech bark

D.W. Langor & J. Sweeney (eds.)

disease which is caused by the interaction of a native species and one or more non-indigenous ones, Dutch elm disease provides an example of an interaction between native and non-indigenous species (Allen and Humble 2002); in this case, an indigenous insect and non-indigenous fungus. The native elm bark beetle, Hylurgopinus rufipes Eichhoff, is the usual vector in Canada and the northern US, and the introduced European elm beetle (Scolytus multistriatus Marsh.) is also a common vector. A new potential vector introduced from Asia, Scolytus schevyrewi Semenov, has been identified in Colorado, where it behaves very similarly to the European elm beetle. It was recently detected in Canada, in both Alberta and Ontario, but is not yet established (G. Pohl pers. comm.). Most large elm trees have been eliminated from the North American landscape but not before producing seed. In many areas, young elm trees survive long enough to reproduce before being infected and killed by the pathogen, thus the tree size and age structure in ecosystems where elm was and sometimes continues to be a major component, have been altered significantly (Castello et al. 1995), but the species is not in danger of being extirpated in any part of its range (Bey 1990). In addition to loss of a food source for wildlife as a result of reduced seed production, and increased levels of coarse woody debris in streams, ecological effects of elm mortality include increased shrub density in canopy gaps as well as local shifts in tree species composition (Castello et al. 1995), and decreased availability of nesting sites for birds (Osbourne 1985). The existence of large, healthy, reproductively mature elm trees in the wild indicates the possibility of genetic resistance (Hubbes 1999; Loo et al. 2007), and four cultivars have been developed that have moderate resistance to the disease (Sinclair and Lyon 2005). Port-Orford cedar root rot Port-Orford cedar (Chamaecyparis lawsoniana (A. Murr.) Parl.) is an example of a less well-known foundation species found in a relatively small area of southwestern Oregon and northwestern California. Its small native range means that, on a North American scale, the species is insignificant, but locally, the species has high ecological importance in ecosystems

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

where it dominates or co-dominates the forest canopy (Ellison et al. 2005). Litter under this species has higher calcium content than that of associated species, and it is known as a soil improver (Zobel 1990). The species is highly shade tolerant and frequently grows in riparian zones stabilizing soil and stream banks (Hansen et al. 2000). The species is threatened throughout its range by Phytophthora lateralis, a root pathogen, believed to have been introduced at least 80 years ago (Murray and Hansen 1997; Hanson et al. 2000). It spread through infected nursery seedlings and causes high mortality (Hansen et al. 2000). Loss of Port-Orford cedar will have locally serious impacts because it is often the only tree species in riparian zones with ultramafic soils (Hansen et al. 2000, Ellison et al. 2005). It recycles calcium to surface soils, provides shade, stabilizes streamside soil, and its rot-resistant wood provides habitat structure long after trees die. Hansen et al. (2000) noted that although the tree species is not in danger of extinction, its frequency especially along stream banks and in sensitive habitats has dramatically declined with potentially serious ecological consequences. As well, the size and age structure has changed with large old trees now replaced by smaller, younger ones. Sniezko (2006) offers hope for the species in his description of the feasibility of breeding for resistance. An active breeding program has resulted in substantial improvement in survival level.

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on native populations are strongly correlated with impacts on ecosystem functions. In order to develop predictive models for long-term impacts, studies are needed that will integrate effects at multiple scales and multiple levels of organization. Among the research themes identified by Mack et al. (2000) two are crucial to generate a better understanding of ecosystem-level impacts: the effects on food webs and associated cascading impacts, and effects of sequential losses of tree species on nutrientcycling dynamics. The current difficulty in defining impacts of past invasions points to the importance of collecting and maintaining broadly based ecosystem baseline information before invasions occur. As Ellison et al. (2005) and Novacek and Cleland (2001) advised, the best insurance to counter the loss of foundation species is to maintain very large natural reserves of intact forests and adopt practices that preserve ecosystem integrity in managed forests. Genetic resistance to invasive pathogens occurs at low frequencies in natural populations of several of the host species. Maintaining large, relatively natural populations of all native tree species will allow natural selection to operate with sufficient intensity to ensure different mechanisms and levels of resistance and tolerance can develop over time, without catastrophic losses of genetic diversity. Acknowledgments A number of people contributed to this manuscript. Ken Harrison provided helpful advice on species names and technical aspects of mycology. Dave Houston, Dave Langor, and two anonymous reviewers provided very helpful comments.

Conclusion Non-indigenous invasive pathogens have caused large-magnitude changes in forest structure, food webs, nutrient cycling, and tree species composition in many areas of North America. Global trade in combination with climate change is almost certain to present new challenges. Despite many excellent studies on effects of the pathogens on their host species, in most cases little is known about the impacts on associated species. Information required for predicting long-term ecological effects is generally lacking. Many questions remain to be addressed to gain an understanding of the full ecological impacts of nonindigenous invasive species and of non-indigenous pathogens in particular. Parker et al. (1999) pointed out that in most cases it is not known whether impacts

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Mack RN, Simberloff D, Lonsdale WM, Evans H, Clout M, Bazzaz FA (2000) Biotic invasions: causes, epidemiology, global consequences, and control. Ecol Appl 10:689–710 McDonald GI, Richardson BA, Zambino PJ, Klopfenstein NB, Kim M-S (2006) Pedicularis and Castilleja are natural hosts of Cronartium ribicola in North America: a first report. For Pathol 36:73–82 McKeen CD (1995) Chestnut blight in Ontario: past and present status. Can J Plant Pathol 17:295–304 McNeely JA (2005) The problems with invasive alien species, and implications for GMOs. Collect Biosaf Rev 2:10–35 McNulty SA, Masters RD (2005) Changes to the Adirondack forest: implications of the beech bark disease on forest structure and seed production. In: Evans CA, Lucas JA (eds) Beech bark disease. Proceedings of the beech bark disease symposium. Saranac Lake, NY, 16–18 June 2004. USDA Forest Service, NE Res Station, Gen Tech Rep NE-331, pp 52–57 Millar CI, Kinloch BB Jr, Westfall RD (1996) Conservation of biodiversity in sugar pine: Effects of the blister rust epidemic on genetic diversity. In: Kinloch BB Jr, Marosy M, Huddleston M (eds) Sugar pine: status, values and roles in ecosystems. Univ Calif Div Agric Nat Resour, Publ 3362, pp 190–199 Murray MS, Hansen EM (1997) Susceptibility of Pacific yew to Phytophthora lateralis. Plant Dis 81:1400–1404 Nair VMG, Kostichka CJ, Kuntz JE (1979) Sirococcus clavigignenti-juglandacearum: an undescribed species causing canker on butternut. Mycologia 71:641–646 Nelson TC (1955) Chestnut replacement in the southern highlands. Ecology 36(2):352–353 Novacek MJ, Cleland EE (2001) The current biodiversity extinction event: scenarios for mitigation and recovery. PNAS 98(10):78–82 Orwig DA (2002) Ecosystem to regional impacts of introduced pests and pathogens: historical context, questions and issues. J Biogeogr 29:1471–1474 Osbourne P (1985) Some effects of Dutch elm disease on the birds of a Dorset dairy farm. J Appl Ecol 22:681–691 Ostrofsky WD, McCormack ML Jr (1986) Silvicultural management of beech and the beech bark disease. N J Appl For 3:89–91 Ostry ME (1995) Conservation of butternut in the Eastern Forests. USDA Forest Service. In: Forest gene conservation principles to practice – workshop proceedings – 008, Ottawa, pp 47–49 Ostry ME (1997) Butternut canker: history, biology, impact and resistance. In: Knowledge for the future of black walnut. Proceedings of the fifth black walnut symposium, 1996. USDA North Central Forest Experimental Station, p 10 Ostry ME (1998) Butternut canker in North American 1967– 1997. In: Proceeding IUFRO WP foliage, shoot and stem diseases in forest trees. Quebec City, p 8 Ostry ME, Kastovich S (1997) First report of Sirococcus clavigignenti-juglandacearum on black walnut. Plant Dis 81:830 Ostry ME, Pijut PM (2000) Butternut: An underused resource in North America. HortTechnol 10:302–306 Ostry ME, Woeste K (2004) Spread of butternut canker in North America, host range, evidence of resistance within

96 butternut populations and conservation genetics. In: Michler CH, Pijut PM, vanSambeek J, Coggeshall M, Seifert J, Woeste K, Overton R (eds) Black walnut: a new century, Proceedings of 6th walnut council research Symposium. 25–28 July, Lafayette, IN. Gen tech Rep NC243. St. Paul MN USDA For Serv North Central research Station, 188 pp Ostry ME, Mielke ME, Skilling DD (1994) Butternut – strategies for managing a threatened tree. USDA For Serv North Central Forest Experiment Station, Gen Tech Rep NC-165, p 7 Papaik MJ, Canham CD, Latty EF, Woods KD (2005) Effects of an introduced pathogen on resistance to natural disturbance: beech bark disease and windthrow. Can J For Res 35:1832–1843 Parker IM, Gilbert GS (2004) The evolutionary ecology of novel plant-pathogen interactions. Annu Rev Ecol Syst 35:675–700 Parker IM, Simberloff D, Lonsdale WM, Goodell K, Wonham M, Kareiva PM, Williamson MH, Von Holle B, Moyle PB, Byers JE, Goldwasser L (1999) Impact: toward a framework for understanding the ecological effects of invaders. Biol Invaders 1:3–19 Pimentel D, Lach L, Zuniga R, Morrison D (2000) Environmental and economic costs of nonindigenous species in the United States. Bioscience 50:53–65 Ramirez M, Loo JA, Krasowski M (2007) Evaluation of genetic resistance to the beech scale insect (Cryptococcus fagisuga) and propagation of American beech (Fagus grandifolia) by grafting. Silvae Genet 56:163–169 Renlund DW (1971) Forest pest condition in Wisconsin. Wisconsin Dept Nat Res Ann Rep 1–53 Rhoades CC (2006) The influence of American Chestnut (Castanea dentata) on nitrogen availability, organic matter and chemistry of silty and sandy loam soils. Pedobiologia 50(6):553–562 Russell EWB (1987) Pre-blight distribution of Castanea dentata (Marsh.) Borkh. Bull Torrey Bot Club 114(2): 183–190 Schlarbaum SE, Hebard F, Spaine PC, Kamalay JC (1997) Three American tragedies: chestnut blight, butternut canker, and Dutch elm disease. In: Britton KO (ed) Exotic pests of eastern forests. USDA Forest Service, Southeastern Research Station, Asheville, NC. Tennessee Exotic Pest Council, 8–10 April 1997, pp 45–54 Shigo AL (1972) The beech bark disease in the northeastern United States. J For 70:263–269 Sinclair WA, Lyon HH (2005) Diseases of trees and shrubs, 2nd edn. Cornell University Press, Ithaca Smock LA, MacGregor CM (1988) Impact of the American chestnut blight on aquatic shredding macroinvertebrates. J N Am Benthol Soc 7:212–221 Sniezko RA (2006) Resistance breeding against nonnative pathogens in forest trees–current successes in North America. Can J Plant Pathol 28:S270–S279

D.W. Langor & J. Sweeney (eds.) Spaulding P (1922) Investigations of the white pine blister rust. USDA Bull No. 957 Storer AJ, Rosemier JN, Beachy BL, Flaspohler DJ (2005) Potential effects of beech bark disease and decline in beech abundance on birds and small mammals. In: Evans CA, Lucas JA (eds) Beech bark disease. Proceedings of the beech bark disease symposium. Saranac Lake, NY, 16–18 June 2004. USDA Forest Service, NE Res Station, Gen Tech Rep NE-331, pp 72–78 Tomback DF, Kramer KA (1980) Limber pine seed harvest by Clark’s nutcracker in the Sierra Nevada: timing and foraging behaviour. Condor 82:467–468 Tubbs CH, Houston DR (1990) American beech. In: Burns RM, Barbara H, Honkala BH (tech. coords.) Silvics of North America: 1. Conifers; 2. Hardwoods. Agriculture handbook 654. U.S. Department of Agriculture, Forest Service, Washington, D.C., vol 2, 877 pp USDA Forest Service (1979) Butternut dieback. In: Guide to common insects and diseases of forest trees. The Northeastern United States Forest Insect and Disease Management, Northeastern Area State and Private Forests, Broomall Vander Wall SB, Balda RP (1977) Co-adaptations of the Clark’s nutcracker and the pin˜on pine for efficient seed harvest and dispersal. Ecol Monogr 47:89–111 Vitousek PM (1990) Biological invasions and ecosystem processes: towards an integration of population biology and ecosystem processes. Oikos 57:7–13 Vitousek PM, D’Antonio CM, Loope LL, Westbrooks R (1996) Biological invasions as global environmental change. Am Sci 84:468–478 Vitousek PM, D’Antonio CM, Loope LL, Rejmanek M, Westbrooks R (1997) Introduced species: a significant component of human caused global change. N Z J Ecol 21:1–16 Waldron G (2003) Trees of the Carolinian Forest: a guide to species, their ecology and uses. The Boston Mills Press, Toronto Walker B, Steffen W (1997) An overview of the implications of global change for natural and managed terrestrial ecosystems. Conserv Ecol [online]1(2):2 http:// www.consecol.org/vol1/iss2/art2/. Accessed 11 Jan 2008 Witter JA, Stoyenoff JL, Petrillo HA, Yocum JL, Cohen JI (2005) Effects of beech bark disease on trees and ecosystems. In: Evans CA, Lucas JA (eds) Beech bark disease. Proceedings of the beech bark disease symposium, Saranac Lake, NY, 16–18 June 2004. USDA Forest Service, NE Res Station, Gen Tech Rep NE-331, pp 28–35 Woods FW, Shanks RE (1959) Natural replacement of chestnut by other species in the Great Smoky Mountains National Park. Ecology 40:349–361 Zobel DB (1990) Port-Orford-Cedar. In: Burns RM, Honkala BH (tech coords) Silvics of North America: 1. Conifers. USDA, For Serv, Agric Handbook 654

The impact of invasive fungi on agricultural ecosystems in the United States Amy Y. Rossman

Originally published in the Journal Biological Invasions, Volume 11, No 1, 97–107. DOI: 10.1007/s10530-008-9322-2 Ó US Government 2008

Abstract Invasive fungi and other non-indigenous plant pathogens have had a significant effect on American agriculture for hundreds of years. At present crop loss due to invasive plant pathogens, especially fungi, is estimated at $21 billion per year in the United States, greater than the loss caused by non-indigenous insects. Plant pathogenic fungi are difficult to detect and identify. Thus knowledge of which fungi pose a threat is essential to prevent their entry by means others than inspection. In this paper, examples are presented of invasive fungi on agricultural commodities introduced into the United States. In all cases two factors have been crucial: first, the pathway through which these fungi have entered, and second, systematic knowledge to prevent and respond to the new invasive species. Historically important plant pathogens such as black stem rust of wheat still cause considerable damage while others such as late blight of potato appear to be having a resurgence. Known previously in Australia, then moving to Africa and South America, the virulent species of soybean rust appeared in the U.S. in 2004 but has not been as devastating as anticipated. Plant pathogenic fungi on specialty crops such as daylily, gladiola and chrysanthemum are threatened by rust fungi recently found in the U.S.

A. Y. Rossman (&) Systematic Mycology & Microbiology Laboratory, USDA Agricultural Research Service, Rm. 304, B011A, 10300 Baltimore Ave., 20705 Beltsville, MD, USA e-mail: [email protected]

apparently brought in on infected germplasm. A crisis in the export of U.S. wheat occurred in the late 1990’s when the molecular diagnostic test for Karnal bunt gave a false positive response to a closely related but previously unknown species. Many potentially dangerous plant pathogens of crop plants have not yet been introduced into the U.S. It is critical that meticulous surveillance be conducted as plant material enters the country as well as where crops are grown prior to shipment. In addition, the scientific infrastructure is needed to be able to respond quickly to new invasive fungi. This requires sound systematic knowledge of plant pathogenic fungi both in the U.S. and around the world and a cadre of systematic experts who can characterize invasive fungi. Keywords Black stem rust
Karnal bunt
Non-indigenous pathogens
Plant pathogens

Invasive fungi and other non-indigenous plant pathogens have had a significant effect on American agriculture for hundreds of years. With the increase in international trade the rate of introduction of these destructive organisms is increasing. Some reports estimate that up to 65% or an estimated $21 billion per year of crops are lost in the United States (U.S.) due to introduced plant pathogens (Pimental et al. 2002, 2005. With control costs included, this translates into about $23.5 billion annually in the U.S. Surprisingly, the crop loss due to non-indigenous

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plant pathogens, most of which are fungi, is greater than that caused by non-indigenous insects. For most of their life histories, plant pathogenic fungi cannot be seen on the plant surface and thus it is difficult to determine their presence and prevent their movement on agricultural commodities. Effective measures to prevent the introduction of invasive fungi must take these factors into account. Many potentially dangerous plant pathogens of crop plants have not yet invaded the U.S. Thus, it is critical that meticulous surveillance be conducted as plant material enters the U.S. at ports of entry as well as where this material is grown prior to shipment of agriculture commodities.

Why are fungi such a big problem as invasive species? Fungi are organisms that obtain their nutrition via absorption. As a result of this mode of nutrition they are always associated with other organisms, often plants, but also insects and humans, as parasites, symbionts, or saprobes. One reason that fungi pose such a problem as invasive species is that most fungi are composed of hyphae, which are very narrow threadlike structures only 5–10 lm diameter, thus the body of the fungus itself occurs inside the substratum and may not be visible from the outside. Only when the fungus is sporulating, often after the plant starts dying, can a fungal pathogen be detected by those who are inspecting agricultural commodities. A second reason the fungi are such a big problem as invasive species is that there are so many different kinds of them, most of which are unknown. At present it is estimated that about 1.5 million species of fungi exist, second only to the insects in total estimated number of species. Of the total number of fungi thought to exist, only 7–20% or about 150,000 fungal species have been described and illustrated (Hawksworth 2001; Hawksworth and Rossman 1997). Even fewer are known in the modern literature with only about 22,000 species or 1.5% of the known fungi having been sequenced. Thus, it is not surprising that about half of the invasive fungi in North America were not described prior to their appearance as invasive species. Given this lack of knowledge, it is often difficult to determine if a new disease outbreak is caused by an invasive fungus or is simply a fungus encountering a stressed host plant

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or a new climatic situation. For example, a fungus that is known primarily as a weak parasite or harmless endophyte in a living plant may become aggressive once the plant is weakened or stressed such as when an agricultural plant is grown in a monoculture or climatic conditions change (Manter et al. 2005; Ortiz-Garcia et al. 2003). Because fungi are relatively unknown and difficult to detect, it is also difficult to anticipate and prevent the arrival of fungal pathogens that have not been adequately described and characterized. One would think that molecular diagnostic tools might be the answer to the detection of plant pathogenic fungi in agricultural commodities and, indeed, progress has been made in that regard. However, such molecular diagnostic tools are only as good as the systematic underpinnings upon which these tools are based. Molecular tests are developed by finding a unique sequence for the organism to be detected, thus it is necessary to know what is the most closely related species in order to determine if a sequence is truly unique to the target species. The consequences of an inaccurate molecular diagnostic tool are exemplified by the test for the Karnal bunt fungus of wheat that gave false positive responses and resulted in a serious agricultural trade crisis as discussed below. Although rapid progress is being made with the use of molecular tools, the identification of fungi has only recently moved from being based on microscopic characteristics to molecular diagnostics that are the result of phylogenetic studies. Within the major groups of fungi, species are now being reevaluated based on the use of multigene phylogenies with some unexpected results. In some groups of fungi, sexual state characters that have traditionally been used for species definition are found to be misleading, masking a number of cryptic or similar looking species. For some groups the morphological variability exhibited by the asexual state, rather than the sexual state, may reflect the phylogeny. For example, in the powdery mildews or Erysiphales the sexual states have morphologically variable appendages that have served as the basis for their identification. Recent phylogenetic studies have confirmed that the morphological characteristics that correlate with phylogeny are those of the conidial or asexual states (Takamatsu 2004). Such is also the case with the leaf parasites in the relatively non-descript sexual genus Mycosphaerella, many of which have highly variable asexual states in such

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Historically important diseases caused by introduced fungi: most are still a problem Black stem rust of wheat

Fig. 1 Phomopsis sp. on agar plate. Species of Phomopsis had been considered host specific but molecular data have shown that one species can occur on several hosts or one host can harbor more than one fungal species. Photo courtesy of USDAARS

genera as Cercospora, Phloeospora, Septoria and others (Hunter et al. 2006). Conversely, species defined based on differences in their plant hosts such as in the asexual genus Phomopsis with over 1,000+ described taxa (Fig. 1) may or may not be hostspecific (Castlebury et al. 2001). Many species of Phomopsis defined based on host differences have been discovered actually to be one species that may attack numerous hosts (Black et al. 1996; Li et al. 2001). Conversely, a single plant host may harbor numerous species of Phomopsis (Farr et al. 2002). Lack of accurate systematic knowledge hinders the ability to develop the accurate molecular diagnostic tests needed to detect invasive fungi on agricultural commodities. In this paper I will focus on examples of invasive fungi of agricultural commodities introduced into the U.S. that have been historically important as pathogens, fungi that have recently been introduced, and those that have influenced the exchange of agricultural commodities. In all cases two factors have been crucial to each situation. One factor is the pathway through which these fungal pathogens have invaded. The second factor is the critical role that knowledge of systematics has played in responding to invasive fungi and the even greater role that systematics could play in preventing the entry of new invasive fungal species. Finally, a discussion is presented of the actions that could be taken to prevent the continuing entry of invasive fungi into the U.S. on agricultural commodities.

Historically, many fungi have negatively affected agricultural ecosystems in the U.S. Wheat rusts are mentioned in the literature from antiquity with a festival, the Robigalia, practiced for over 1,700 years, dedicated to pleasing the forces that control rust diseases (Schumann 1991). Given the close association of rust fungi with their host plants, it is not surprising that the early European settlers who brought wheat seed to the North American continent also introduced the fungus that causes black stem rust. The disease is caused by the rust fungus, Puccinia graminis Pers. (Uredinales), which has a complex life history with five different spore types that infect hosts in two diverse plant families. The uredinal and telial states occur on 77 genera of grasses (Poaceae, primarily subfamily Pooideae), while the aecial state occurs on more than 70 species of barberry and related plants (Berberis and Mahonia, family Berberidaceae). In the early 1900s black stem rust was an especially damaging disease of wheat in the U.S. Steps were taken to eliminate the alternate barberry host, which continued until 1990 (Peterson 2001). A concomitant drop in wheat rust epidemics has been documented (Roelfs 1982). At the same time plant breeders were discovering genes in wheat for resistance to black stem rust; this program has been successful although the research continues to this day. The combination of barberry eradication in the major wheat-growing states and the constant inclusion of black stem rust resistant genes has resulted in limited damage from this disease in recent years. Although the barberry eradication program was successful in reducing the amount of wheat disease, one wonders what impact eliminating these plant species from the Mid-West has had on the existence of native species such as Berberis canadensis P. Mill. used as a cure for diarrhea by the Cherokee Indians as well as the other organisms, some possibly beneficial, associated with these plants. Despite the worldwide occurrence of P. graminis, the systematics of this species is still relatively unknown. Over the decades many segregate taxa have been described, particularly segregate species and biologically specialized formae specialis, which are taxa defined by their occurrence on a specific host.

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Subdivisions of P. graminis into subspecies, varieties and formae speciales have been proposed based on the morphology of the urediniospores and the host range. Crossing studies support the separation of at least two subspecies, but not the proposed separation based on spore size. Recent molecular sequence data have been used to determine that P. graminis is truly one species with broad genetic diversity that attacks many genera of plant hosts. Its high genetic variability suggests that its center of origin is in Asia. This confirms previous studies demonstrating that this disease was introduced into the U.S. at least twice (Abbasi et al. 2005). These data suggest that strains from diverse hosts may be undergoing genetic exchange at a low level. Recently a new strain (Ug99) of this fungus was detected in eastern Africa, has spread to the Middle East, and is projected to move eastward to India and beyond (Stockstad 2007). Thus, black stem rust is still a problem with new races evolving and new introductions occurring that result in a continuous need to develop resistant wheat cultivars (Leonard 2001). Late blight of potatoes Late blight of potatoes caused by Phytophthora infestans (Mont.) de Bary was introduced into Europe in the 1840’s. In 1845 it moved into Ireland where 25% of the potato crop was lost. The next year the disease was even more severe when 80% of the potato crop was destroyed. With their food staple gone about a quarter of the Irish population, or two million people, died of starvation and one million Irish immigrated to the U.S. Although copper fungicides are now used to protect potatoes from this organism and resistant potato cultivars have been developed, losses due to late blight of potatoes still occur when the weather is unusual or new strains of the organism are introduced. In the last decade a new mating type of P. infestans was found to occur in the U.S. and this disease has re-emerged as a serious problem for potato growers (Fry et al. 1993; Fry and Goodwin 1997). In addition to increased genetic diversity due to recombination, changes in weather patterns have also been implicated in the increased disease severity. As a result there is rekindled interest in this historically important disease and the threats posed by genetic variants of this organism. Using specimens deposited in herbaria over the past one hundred years, Jean Ristaino and her

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associates have been able to determine the routes and approximate times in which P. infestans was transported around the globe. May and Ristaino (2004) determined that this organism was moved with potatoes from South America to Europe in about 1830 and then to North America in 1910. Recent data suggest that P. infestans originated in the Andes (South America) and was transported to Europe, Mexico, and North America (Go´mez-Alpizar et al. 2007). Several genetic variants reside outside the U.S., the most threatening in Mexico; thus, actions to prevent the introduction of invasive isolates of P. infestans are still needed. These recent molecular analyses, especially of historical specimens, have contributed to the ability to track the spread of this pathogenic organism. Although not known for certain, it seems likely that the pathogen was spread with diseased germplasm that was transported for planting into new geographic areas. Phytophthora infestans belongs to a group of organisms that has traditionally been considered fungi but recently has been determined to be phylogenetically distinct. Phytophthora, Pythium and the downy mildews (Peronosporales) are now known to belong to the Chromista, more closely related to the yellowbrown algae than to the true Fungi (Rossman and Palm 2006). Many species of Phytophthora, Pythium and downy mildews are obligate parasites of living plants and cause diseases throughout the world. New invasive species of Phytophthora are being discovered regularly, such as P. ramorum S. Werres, A.W.A.W. de Cock & W.A. Man in’t Veld which causes sudden oak death, a disease that has recently killed hundreds of trees in the western U.S., and ramorum blight that affects numerous woody ornamental hosts and has had a significant impact on trade. Southern corn leaf blight Southern corn leaf blight is of considerable significance in the history of American agriculture. This disease reached epidemic proportions in 1970 when it devastated much of the corn crop at an estimated monetary loss of $1 billion (Schumann 1991). Southern corn leaf blight is caused by an ascomycetous fungus with the asexual state Bipolaris maydis (Nisikado) Shoemaker and the sexual state or teleomorph Cochliobolus heterostrophus Drechsler. The fungus causing southern corn leaf blight occurs on several hosts including corn, sorghum and teosinte.

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The fungus overwinters on crop debris, primarily on the soil surface, as mycelium, conidia and chlamydospores. Conidia are windblown or splashed by water to fresh plant tissue in the spring. The spores germinate on the leaf surface and infect the host directly and through stomata. Disease development is favored by warm moist weather and the presence of free moisture on the host tissue surface. The fungus is very prolific and is able to complete an entire life cycle in 60–72 h under favorable weather conditions. The fungus causing southern corn leaf blight exists in two distinct races each of which produces phytotoxins: Race O (old race) and Race T (virulent on corn containing Texas male sterile cytoplasm). The phytotoxin produced by Race O is nonspecific to cytoplasm types and is produced in small amounts. The more aggressive Race T produces a phytotoxin that specifically affects corn containing Texas male sterile cytoplasm. Although present in the U.S. prior to that time, Race T became prevalent in 1970 in the Midwest where almost all of the corn was genetically uniform and contained Texas male sterile cytoplasm, a desirable trait that was, unfortunately, correlated with susceptibility to this pathogen. Following the 1970 disaster corn was bred with single gene and polygenic resistant to Race T. Today corn is resistant to both races of the pathogen, the primary means of controlling the plant’s health. In this example the new race of the pathogen was noticed but the impact was not appreciated. The pathogen appears to have been introduced through natural means. However, the genetic uniformity of the crop resulted in extreme susceptibility to this disease (Ullstrup 1972).

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spores of P. pachyrhizi carried by wind currents over long distances may have led to its introduction into South America from Africa and then into the U.S. from South America. Schneider et al. (2005) suggest that spores of soybean rust were blown northward from northern South America into the U.S. by Hurricane Ivan. Spores of soybean rust have been found high in the atmosphere (Barnes et al. 2006). Both species of Phakopsora infect a broad range of leguminous hosts; thus, control of the disease may not be easy. The arrival of this invasive fungus was long anticipated following an outbreak of a Phakopsora disease on legumes in Puerto Rico in the mid-1970’s (Bromfield 1984). In this situation the fungus on soybean was the less harmful P. meibomiae, but the government was alerted to the more serious potential problem posed by P. pachyrhizi. At that time a research program was begun anticipating the introduction of the more virulent species. In 2001 P. pachyrhizi was found in South America, causing serious damage to the soybean crop in Brazil. It was considered inevitable that the disease would arrive in North America and plans were made for actions to be taken when/if the fungus was found in the U.S. A molecular test was developed to provide rapid diagnosis. State diagnosticians were alerted and trained in the morphological identification of the disease. Finally in the fall of 2004 following Hurricane Ivan, Asian soybean rust caused by P. pachyrhizi was detected in Louisiana (Levy et al. 2005) (Fig. 2). The response plan was put into place. Soon thereafter Asian soybean rust was also detected in a number of southeastern states (Krupa et al. 2006).

New invasive fungi arriving daily Soybean rust One of the most recent invasive fungi to arrive in the U.S. is soybean rust caused by two fungi, Phakopsora pachyrhizi Syd. & P. Syd. and P. meibomiae (Arthur) Arthur. Phakopsora pachyrhizi is by far the more aggressive of these two species and the disease it causes is referred to as Asian soybean rust. Native to Asia and present in Australia for some time where it has had significant economic impact, the fungus has recently been found in Africa, South America, Hawaii, and along the Gulf Coast in the continental U.S. Fungal

Fig. 2 Soybean rust (Phakopsora pachyrhizi) infecting living leaves. Photo courtesy of Dr. Jose Hernandez

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After three years the virulent soybean rust fungi has spread north as far as Illinois although damage to soybean has not been as devastating as anticipated. Grower readiness may be a factor as well as weather conditions not conducive to the disease.

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nursery stock. Because a fungal pathogen may not be visible on the plant surface, nursery stock must be inspected prior to shipment or held on quarantine for extended periods of time after arrival in the U.S. Only then will fungal pathogens such as P. tracheiphila, the cause of mal secco disease, be detected.

Citrus diseases Citrus black spot, caused by the ascomycetous fungus Guignardia citricarpa Kiely, is a disease that has been considered of plant quarantine importance for Europe and the U.S. for several decades. Detection of this fungus is difficult because it often remains latent until late in the growing season. Identification of this fungus is difficult because of confusion with a similar species that occurs on the same host. Recently this systematic puzzle was unraveled after the look-alike fungus was mistaken for real black spot on citrus fruits exported to Europe from the U.S., where Citrus black spot does not occur. This potential trade issue was averted when the fungus was determined to be the non-pathogenic species. Additional systematic research showed that, while similar morphologically to the citrus pathogen G. citricarpa, an endophytic species, G. mangiferae A.J. Roy, can be distinguished upon close scrutiny and using molecular sequence data (Baayen et al. 2002). Once the taxonomic and biological distinction was made between these two species, a molecular test was developed to distinguish them (Meyer et al. 2006). In order to protect U.S. agriculture from invasive fungi, knowledge of the fungi normally associated with crop plants as well as the biological characteristics of the plant pathogen is essential. Another potentially harmful pathogen on Citrus causes a tracheomycotic disease called mal secco. It is caused by Phoma tracheiphila (Petri) Kantachveli & Gikachvili and is known from the Mediterranean countries from which the U.S. imports significant amounts of Citrus products (Smith et al. 1992). The systematics of this group is relatively unknown as they generally are non-descript and lack distinctive morphological characteristics. In fact, P. tracheiphila may be more closely related to genera other than Phoma. A molecular diagnostic test has been developed for the cause of mal secco disease (Licciardello et al. 2006) but this test was not based on accurate systematic knowledge of species closely related to the causal organism. The most likely pathway by which pathogens of Citrus may enter the U.S. is via infected

Diseases of specialty crops—daylily rust, chrysanthemum white rust, and gladiolus rust Specialty crops that include many beloved horticultural plants are extremely valuable to commercial growers and home gardeners. New varieties and cultivars are continuously developed based on novel, often exotic germplasm. With importation of germplasm from outside the U.S. comes the risk of new invasive fungi. One recently introduced unwanted fungus is the cause of daylily rust, Puccinia hemerocallidis Thu¨m., on daylily (Hemerocallis spp.) (Wise et al. 2004). Nursery stock of daylily is grown in some Central American countries during the winter. In early spring daylilies are imported for distribution to nurseries throughout the U.S. The plant material is usually trimmed to include only the basal living rootstock and may appear to be free of disease. In 2001 P. hemerocallidis was found on daylily in the U.S. The rust fungus was previously known mainly in eastern Asia where most daylilies are resistant to that fungus. It is difficult to know how a fungus known only from Asia came to infect rootstocks of daylilies in Central America and why it was not detected in situ prior to export. Unfortunately, infected nursery stock was distributed throughout the U.S. before the disease was detected and the fungus was spread quickly. Initially there was a question of whether the rust on daylily in the U.S. was the same as that in Asia. Systematic research demonstrated that the fungus in the U.S., Costa Rica, and Japan is the same (Hernandez et al. 2002). At that time, so little was known about the life history of daylily rust that it was unclear if an alternate host existed. Recent research demonstrates that P. hemerocallidis and a closely related species develop additional spore states on the horticultural plant Patrinia (Chatasiri et al. 2006). Knowing the complete life history, specifically that daylily rust has an alternate host on which this fungus could have been transported from Asia to the western hemisphere, provides a clue to the mysterious appearance of this invasive species. Although daylily growers enamored of these lovely

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

flowers are worried about the future of their favorite plant, resistance exists in cultivars of daylily. Other rust fungi pose a significant threat to specialty crops especially because their pathways of entry continue to elude detection. Outbreaks of Chrysanthemum white rust caused by Puccinia horiana P. Henn. have occurred repeatedly in greenhouses throughout the U.S. apparently brought in with infected germplasm. This fungus was transported from its native range in Asia to Europe, especially England and The Netherlands, in the 1960’s and more recently to South America. Despite the recurring outbreaks around the U.S., Chrysanthemum white rust has been contained and eradicated every time it has been encountered. Given its widespread occurrence in other parts of the world and the vast quantities of imported germplasm, continuous vigilance and rigorous inspection of plants and cut flowers prior to shipping is required to keep this invasive fungus from becoming established in the U.S. (Wise et al. 2004). Rust fungi also threaten gladiolus and other Iridaceae. An outbreak of the rust Uromyces transversalis (Thu¨men) Winter in April–May 2006, then again in 2007, occurred in Florida and later in California (Blomquist et al 2007; Schubert et al. 2007) as well as in Mexico (Rodrı´guez-Alvarado et al. 2006). Prior to discovery in the U.S. this rust fungus was widely distributed throughout the world but not known from the northern half of the western hemisphere. In this case both federal and state plant quarantine officials were aware of the potential problem and responded quickly when the disease was detected. This awareness resulted in containment of this invasive fungus. A second rust fungus on gladiolus, U. gladioli P. Henn., also poses a threat (Wise et al. 2004) but is not as widespread nor as close to U.S. borders as U. transversalis. Despite knowledge of the existence and threat posed by the rust fungi on horticultural commodities, these invasive fungi continue to be introduced.

Fungal diseases influence exchange of agricultural commodities Karnal bunt Karnal bunt is a disease primarily of wheat (Triticum aestivum L.) caused by the obligately parasitic fungus Tilletia indica Mitra (Ustilaginales-smut and bunt

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fungi). This fungus was described in Karnal, India, in the 1930’s and spread to Mexico. For decades the U.S. was concerned that the disease might be detected in this country. Just one diseased kernel can generate hundred of spores that can be carried by wind or on equipment and humans. Although the disease does little to damage crop yield, the quality of the grain is degraded due primarily to a fishy odor. Because of the decrease in seed quality and proximity of this disease to the U.S., the federal government in the 1980’s placed a zero tolerance for Karnal bunt on imported wheat. Other countries throughout the world followed the U.S.’s lead and also placed a zero tolerance for Karnal bunt on imported wheat. Knowing that this invasive fungus was close to our borders, the USDA developed a response plan in case Karnal bunt was discovered in the U.S. Identification of most bunt fungi is difficult based on the patterns of ornamentation as seen on the outside of the microscopic teliospores (Fig. 3). In addition to detecting the teliospores by visual inspection of seed washes, a molecular test was developed for rapid identification of the germinated teliospores (Smith et al. 1996). The basic systematics of bunt fungi had not been well studied partly because most agricultural crops have been bred for resistance to bunt and smut diseases. At the time that Karnal bunt was initially recognized as a potentially invasive fungus, there was only one scientist working on the systematics of bunt and smut fungi in the U.S. In 1996, the Arizona Department of Agriculture detected Karnal bunt on durum wheat planted in Arizona, California, New Mexico and Texas (Ykema et al. 1996). Because the fungus can spread rapidly, a survey plan was put into effect and government agencies including APHIS (Animal and Plant Health Inspection Service) issued a quarantine for these states. A survey of wheat for Karnal bunt throughout the U.S. was initiated by visual inspection of teliospores in seed washes and using the molecular test to verify the identity of purported Karnal bunt teliospores. To the surprise of everyone, both the inspection and molecular test indicated that Karnal bunt was present throughout much of the U.S. including the Pacific Northwest, the southwestern, and the southeastern states. What a shock! Countries that imported wheat from the U.S. immediately closed their ports to U.S. wheat and the annual $5.5 billion wheat export market was threatened.

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D.W. Langor & J. Sweeney (eds.)

Fig. 3 Karnal bunt (Tilletia indica) (top) and a similar looking species (Tilletia walkeri) (bottom) that were confused causing a trade crisis in the export of U.S. wheat. Photo courtesy of Dr. Lisa Castlebury

Dr. Mary Palm, the APHIS National Mycologist who examined and verified the identification of the teliospores of Karnal bunt, noticed that some of the teliospores were very slightly different and could not be identified with certainty as T. indica. With Drs. Lori Carris and Lisa Castlebury and others it was discovered that these odd teliospores were not Karnal bunt but were those of a previously unknown bunt species that occurred on annual and perennial rye grass (Lolium multiflora L.and L. perenne L.), often inadvertently mixed with wheat upon harvest. After detailed examination with light and scanning electron microscopes, it was determined that these similar-appearing teliospores had a slightly different ornamentation that could be distinguished by an experienced, practiced eye (Fig. 3). The slightly different teliospores that gave the positive test to the molecular test for Karnal bunt represented a new species described as Tilletia walkeri and now known on rye grass from Australia as well as the U.S. (Castlebury and Carris 1999). Throughout the survey the molecular test was giving a false positive response to the rye grass bunt fungus because the test had been developed based on a wrong assumption about the relationship of the Karnal bunt fungus to its next closest relative. The test was developed assuming that rice bunt (Tilletia horrida Takah.) was the next closest relative. Later molecular studies demonstrated that the newly discovered

T. walkeri on perennial ryegrass was more closely related to Karnal bunt than rice bunt, thus resulting in the false positive response (Levy et al. 2001). This situation illustrates the dire need for accurate phylogenetic information upon which to base molecular diagnostic tests. Such tests are not accurate without the essential underpinning of systematic knowledge. Once the ryegrass bunt fungus was discovered, new criteria were established for positive identification of Karnal bunt. In reality the Karnal bunt disease had spread to only a few limited locations. The U.S. could prove that most wheat for export was grown in areas without Karnal bunt and importing countries throughout the world were again willing to accept wheat from the U.S. This crucial agricultural export market was saved. The U.S. has contained the spread of Karnal bunt and most U.S. wheat is grown in areas free of this disease.

Given the tremendous challenge posed by invasive fungi, how can we combat them? Legal recognition of the seriousness of invasive organisms was demonstrated in the Plant Quarantine Act of 1912 (Campbell et al. 1999; Palm 2001). Since then, scientists and policy makers have worked to prevent entry and damage caused by invasive species.

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

Some progress has been made in the last decade given modern tools for communication and molecular tools for identification of cryptic organisms such as fungi. One of the most basic requirements is to know which organisms threaten U.S. agriculture. For fungal pathogens this is not an easy task because, as mentioned above, most fungi have not yet been discovered and are difficult to detect. APHIS has regulated a number of plant pests through inclusion in federal regulations. The fungi on this Regulated Plant Pest List (RPPL) were recently reviewed to determine the accuracy of the systematic information about them including their host range and geographic distribution (Cline and Farr 2006). Once current knowledge was synthesized, each fungal species on the RPPL was evaluated for the threat it posed to U.S. agriculture (Rossman et al. 2006). About half of the fungi on the RPPL were determined to already exist in the U.S., occur on crops not grown in the U.S., or be too unknown to evaluate their threat e.g., known only from their type specimen. This exercise demonstrates the need to periodically evaluate which fungi pose a threat to U.S. agriculture. As new fungal pathogens outside the U.S. are described on agricultural commodities, their threat to U.S. agriculture should be evaluated immediately. Knowledge of the plant pathogenic organisms that already occur in the U.S. is also essential. A comprehensive account of the fungi reported in the U.S. was published (Farr et al. 1989) including about 12,000 fungal species on 6,000 plant hosts. Since then, the number has almost doubled to 23,162 species of fungi now reported in the U.S. (D.F. Farr, pers. comm. 2007). In addition, progress has been made in providing access to reports of fungi on plant hosts from throughout the entire world. Farr et al. (2007) currently maintains a database of literature reports of fungi on plant hosts throughout the world that includes over 94,000 fungal species. Most plant quarantine officials agree there is a need for increased knowledge of potentially invasive organisms that exist throughout the world including their movement in real-time. Such knowledge was essential in preparing for the invasion by Asian soybean rust and Karnal bunt. In the last decade the emphasis has begun to shift from quarantine policies centered on individual invasive species to determining the most dangerous pathways through which invasive organisms enter the country. For fungi these pathways appear to be

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primarily nursery stock and raw logs (Ivors et al. 2006; Palm and Rossman 2003). The heightened awareness of the need to prevent the introduction of invasive species through greater scrutiny of most likely pathways of entry has resulted in new regulations governing the import of wood packing material. Nursery stock as a pathway has been implicated in many serious diseases including daylily rust, gladiolus rust, and dogwood anthracnose. Plans are in progress to increase scrutiny of nursery stock as a pathway for invasive species. For agricultural, horticultural and forest pathogens, this pathway should include all germplasm for planting including seeds and tubers. Given the vast quantities of agricultural goods that are shipped around the world, the most efficacious approach to detect invasive species is to inspect commodities prior to shipment. This would be especially useful for agricultural goods growing outside of their natural environment. Ideally, such inspections would be done by local authorities trained in plant pathology and able to detect potential invasive fungi. This requires that the scientific infrastructure in developing countries be created or enhanced. Developing these capabilities in the countries that ship agricultural commodities to the U.S. will be expensive but will pay off in preventing the entry of invasive species. Continuous scrutiny by a local plant pathologist who would notice diseased plants in the field prior to harvesting will be more effective than sending a U.S. inspector to look superficially at the commodity prior to shipment. All of the above approaches require a strong systematics safety net. By safety net I mean a body of knowledge about invasive species useful for their identification, developing accurate methods of detection, and determining pathways of entry for pathogenic fungi outside and inside the U.S. Such knowledge will serve as the basis for quarantine policy decisions to prevent the entry of invasive fungi and respond when they are found in the U.S. Knowledge needed about invasive fungi includes accurate scientific names, with descriptions and illustrations; biology such as host range and pathogenicity; geographic distribution and the ability to track their movement; and phylogeny to predict the disease potential. This information must be readily available on the Internet. In addition, rapid molecular diagnostic tools for invasive fungi should be developed. As mentioned in the Karnal bunt example, these tools must be based on sound systematic knowledge of the phylogenetic relationships among

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species. While molecular diagnostics may seem to be the solution for rapid identification of invasive fungi, the Karnal bunt situation dramatically exemplifies the crucial need for accurate systematic understanding of potentially invasive fungi. The organisms most destructive to agricultural systems are fungi, insects and nematodes. Ironically, these are the groups of organisms that are still relatively unknown. The systematics safety net must include a body of experts who know these organisms well. For example, scientists who are experts in the microscopic fungi are crucial for identification of invasive species once they are detected in the U.S. These experts also are essential for determining relationships among species essential for designing accurate molecular diagnostic tools.

Conclusion Historically, plant pathogenic fungi have caused considerable damage to agricultural crops. Many of these fungi still pose serious problems. Because of the enormous damage caused by invasive fungi, great effort is made to identify and understand these organisms once they have entered the country. However, knowledge of invasive fungi that threaten the U.S. is still relatively meager. Often even the most rudimentary systematic understanding required for their identification is lacking. More emphasis should be placed on (1) knowing specifically what species of invasive fungi exist on agricultural plants outside the U.S., (2) preventing their entry through inspection of agricultural commodities before export, and (3) being able to detect and eradicate invasive fungi once they are found within our borders. In addition, there is a need for the scientific infrastructure able to respond quickly to invasive fungi especially those not previously described. This requires a cadre of systematic experts who can characterize newly discovered invasive fungi, whether previously known or new to science, not just in the U.S. but in countries around the world.

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D.W. Langor & J. Sweeney (eds.) Baayen RP, Bonants PJM, Verkley G, Carroll GC, Van Der Aa HA, De Weerdt M, van Brouwershaven IR, Schutte GC, Maccheroni W Jr, de Blanco CG, Azevedo JL (2002) Nonpathogenic isolates of the citrus black spot fungus, Guignardia citricarpa, identified as a cosmopolitan endophyte of woody plants, G. mangiferae (Phyllosticta capitalensis). Phytopathology 92:464–477 Barnes CW, Szabo LJ, Johnson JL, Bowersox VC, Harlin KS (2006) Detection of Phakopsora pachyrhizi spores in rain using a real-time PCR assay. Phytopathology 96 (6 suppl):S9 Black BD, Padgett GB, Russin JS, Griffin JL, Snow JP, Gerggren GT Jr (1996) Potential weed hosts for Diaporthe phaseolorum var. caulivora, causal agent for soybean stem canker. Plant Dis 80:763–765 Blomquist CL, Thomas SL, McKemy JM, Nolan PA, LuqueWilliams M (2007) First report of Uromyces transversalis, causal agent of Gladiolus rust, in San Diego County, California. Plant Dis 91:1202 Bromfield KR (1984) Soybean Rust Monograph No. 11. American Phytopathological Society, St. Paul, Minnesota, p 63 Campbell CL, Peterson PD, Griffith CS (1999) The formative years of plant pathology in the United States. APS Press, St Paul, Minnesota Castlebury LA, Carris LM (1999) Tilletia walkeri, a new species on Lolium multiflorum and L perenne. Mycologia 91:121–131 Castlebury LA, Farr DF, Rossman AY (2001) Phylogenetic distinction of Phomopsis isolates from cucurbits. Inoculum 52:25 Chatasiri S, Kitade O, Ono Y (2006) Phylogenetic relationships among Puccinia hemerocallidis, P. funkiae, and P. patriniae (Uredinales) inferred from ITS sequence data. Mycoscience 47:123–129 Cline ET, Farr DF (2006) Synopsis of fungi listed as regulated plant pests by the USDA Animal and Plant Health Inspection Service: Notes on nomenclature, disease, plant hosts, and geographic distribution. Online Plant Health Progr 05-05-06. Farr DF, Bills GF, Chamuris GP, Rossman AY (1989) Fungi on plants and plant products in the United States. American Phytopathological Society, St Paul, Minnesota Farr DF, Rossman AY, Palm ME, McCray EB (2007) Fungal Databases, Systematic Botany & Mycology Laboratory, ARS, USDA. Retrieved from http://nt.ars-grin.gov/ fungaldatabases/ Farr DF, Castlebury LA, Rossman AY (2002) Morphological and molecular characterization of Phomopsis vaccinii and additional isolates of Phomopsis from blueberry and cranberry in the eastern United States. Mycologia 94:494–504 Fry WE, Goodwin SB (1997) Resurgence of the Irish potato famine fungus. Bioscience 47:363–371 Fry WE, Goodwin SB, Dyer AT, Matuszak JM, Drenth A, Tooley PW, Sujkowski LS, Koh YJ, Cohen BA, Spielman LJ, Deahl KL, Inglis DA, Sandlan KP (1993) Historical and recent migrations of Phytophthora infestans: chronology, pathways, and implications. Plant Dis 77:653–661 Go´mez-Alpizar L, Carbone I, Ristaino JB (2007) An Andean origin of Phytophthora infestans inferred from mitochondrial and nuclear gene genealogies. Proc Natl Acad Sci 104:3306–3311

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Responses of Chalcidoidea (Hymenoptera) parasitoids to invasion of the cabbage seedpod weevil (Coleoptera: Curculionidae) in western Canada Lloyd M. Dosdall Æ Gary A. P. Gibson Æ Owen O. Olfert Æ Peter G. Mason

Originally published in the Journal Biological Invasions, Volume 11, No 1, 109–125. DOI: 10.1007/s10530-008-9323-1 Ó Springer Science+Business Media B.V. 2008

G. A. P. Gibson
P. G. Mason Agriculture and Agri-Food Canada, Research Centre, 960 Carling Avenue, Ottawa, ON, Canada K1A 0C6

Chalcidoidea fauna of C. obstrictus include species with both Nearctic and Holarctic distributions, with some species having restricted host ranges and others that are more niche than taxon-specific. These Chalcidoidea species appear to have expanded their host ranges to parasitize C. obstrictus in the region. Most parasitism is attributable to Trichomalus lucidus (Walker), Chlorocytus sp., and Pteromalus sp. (Pteromalidae), and Necremnus tidius (Walker) (Eulophidae). Parasitism levels varied considerably over the three years of this study. From 2003 to 2005 increases in parasitism occurred among all four of the species dominating the parasitoid fauna of C. obstrictus, but greater increases were observed for Chlorocytus sp. and Pteromalus sp. than for T. lucidus. Parasitoid species have sometimes caused substantial levels of host mortality, although current levels are usually less than 15% for all species combined and so are not sufficient to control weevil populations. Implementing a classical biological control program for C. obstrictus by reconstructing its European natural enemy complex is being considered, but it is still uncertain whether parasitism levels by native Chalcidoidea will increase over time since considerable yearto-year variation has been found. Parasitism levels of C. obstrictus should therefore continue to be monitored to assess whether a classical biological control program should be implemented.

O. O. Olfert Agriculture and Agri-Food Canada, Saskatoon Research Centre, 107 Science Place, Saskatoon, SK, Canada S7N 0X2

Keywords Ceutorhynchus obstrictus
Invasive species
Canola
Chalcidoidea
Geographical distribution patterns

Abstract Invasion of the European weevil, Ceutorhynchus obstrictus (Marsham), was investigated through surveys of its range and population densities in Alberta and Saskatchewan from 2001 to 2005. After it was first reported in southern Alberta, C. obstrictus rapidly expanded its range and abundance. Our more recent surveys indicate that its northward expansion has slowed, but that it has continued to extend its range eastward to southcentral Saskatchewan. The distribution and abundance of parasitoids of C. obstrictus in Alberta and Saskatchewan were investigated from 2003 to 2005 by mass rearing canola pods infested with C. obstrictus larvae. Although weevil populations were not parasitized for several years immediately following its introduction to southern Alberta, a surprisingly diverse assemblage of Chalcidoidea parasitoids, comprising 12 species from four families, were recently reared from weevil-infested canola siliques in Alberta and Saskatchewan. The

L. M. Dosdall (&) Department of Agricultural, Food and Nutritional Science, 4-16B Agriculture-Forestry Centre, University of Alberta, Edmonton, AB, Canada T6G 2P5 e-mail: [email protected]

D.W. Langor & J. Sweeney (eds.), Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems. DOI: 10.1007/978-1-4020-9680-8_8

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Introduction The cabbage seedpod weevil, Ceutorhynchus obstrictus (Marsham) (Coleoptera: Curculionidae), is native to Europe and is a serious pest of brassicaceous seed crops, especially canola (Brassica napus L. and Brassica rapa L.) and brown mustard (Brassica juncea (L.) Czern). It was first discovered in North America in the southwestern coastal region of British Columbia in 1931 (McLeod 1962). During the next 25 years it was recorded from several additional localities in the Pacific Northwest including Washington (Baker 1936), Idaho (Walz 1957), Oregon (Crowell 1952), and California (Hagen 1946), but it was not discovered in Alberta until 1995 (Dosdall et al. 2001). It has also been found in eastern North America, in Maryland (Anonymous 1977), Tennessee (Boyd and Lentz 1994), and Georgia (Buntin et al. 1995), and most recently in Que´bec (Brodeur et al. 2001) and Ontario (Mason et al. 2004). Based on analysis of variations in the mitochondrial DNA COI gene among geographically distinct populations, Laffin et al. (2005) concluded that North American populations of C. obstrictus likely resulted from at least two separate introductions. An introduction to western North America appears to have occurred from source populations in western or northern Europe, and a second introduction was made to northeastern North America in Que´bec from source populations in northern Europe. The cabbage seedpod weevil is univoltine. Sexually immature adults diapause during winter in protected areas such as beneath leaf litter in shelterbelts, and emerge in spring when soil temperatures increase to approximately 15°C (Ulmer and Dosdall 2006). After feeding on early-flowering brassicaceous plants, adults invade canola crops in the bud to early flowering stages where mating and additional feeding occur (Fox and Dosdall 2003; Dosdall and Moisey 2004). After flowering and as soon as pistil enlargement occurs on canola racemes, a mated female can excavate a cavity in an immature pod with her mouthparts, and may deposit a white, cylindrical egg into the puncture. Eggs hatch in approximately one week, and each larva consumes about 5–6 seeds during three instars (Dmoch 1965; Dosdall and McFarlane 2004). Mature larvae chew exit holes in the walls of the siliques, drop to the soil surface, burrow in, and pupate. New generation adults emerge

D.W. Langor & J. Sweeney (eds.)

approximately 10 days later, and feed on canola or other brassicaceous plants until late in the season when temperatures decline and they migrate to overwintering sites (Bonnemaison 1957; Dmoch 1965; Dosdall and Moisey 2004). The invasion of western Canada by the cabbage seedpod weevil has garnered considerable attention within the agricultural community because of its economic impact on canola crops. Ceutorhynchus obstrictus can cause severe economic losses in canola at several stages of crop development. When overwintered adults invade crops primarily in the bud to earlyflowering stages, they feed on buds and often destroy their vascular tissue causing ‘‘bud-blasting’’ (Dosdall et al. 2001). Plants with severe bud-blasting produce racemes with few pods, and may even fail to flower when weevil densities are high. Feeding by larvae within siliques causes seed losses of approximately 18–20% (Dmoch 1965). If environmental conditions are humid after larvae bore exit holes, the pods can be invaded by fungal pathogens that damage additional seeds within the pods. When new generation adults emerge late in the season, they feed upon seeds within green pods to build up fat stores for overwintering, causing further reductions in crop yield and quality (Buntin et al. 1995; Dosdall et al. 2001). The rapid range expansions that have occurred historically in North America indicate that C. obstrictus has a remarkable ability to disperse and establish new populations. Surveys conducted in western Canada from 1997 to 2000 inclusive to monitor geographical changes in its distribution and abundance indicated rapid dispersal of C. obstrictus from southern Alberta to the north and east at a rate of approximately 55 km per year (Dosdall et al. 2002). When meteorological data from 193 sites across Alberta, Saskatchewan, and Manitoba were analyzed with data on the biological requirements of C. obstrictus using the CLIMEXÒ model, it was predicted that the distribution of the species would eventually encompass the entire region of canola production in western Canada (Dosdall et al. 2002). However, several species of Chalcidoidea parasitoids were recently found to attack larval populations of the weevil (Gibson et al. 2005; Dosdall et al. 2006). The current study was undertaken to monitor the distribution and abundance of the cabbage seedpod weevil in western Canada since earlier surveys established its initial dispersal patterns, and to

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

determine its ecological impact on populations of the Chalcidoidea parasitoids.

Materials and methods Weevil geographical distribution patterns, 2001–2005 In 2001–2005, respectively, 258, 167, 419, 420, and 396 canola fields were sampled from 25 June to 10 July, when most canola crops in southern and central Alberta and Saskatchewan were in full flower (Growth Stages 4.2–4.3 of Harper and Berkenkamp 1975). A single sample, consisting of 25, 180° sweeps with a standard insect sweep net, was collected from each field. Sampling commenced at the field edge and progressed inward over a distance of about 30 m. Fields were sampled in an area extending from Edmonton, AB (53°340 N; 113°310 W) and North Battleford, SK (52°460 N; 108°150 W) south to the United States border, east to Swift Current (50°170 N; 107°410 W) in southcentral Saskatchewan, and west to the limits of canola production in the foothills of the Rocky Mountains in Alberta. The geographical location of each field was determined using high resolution maps or hand-held global positioning system units (Magellan Trailblazer XLÒ Personal Satellite Navigation System). Entire sweep net contents were either stored in plastic bags and frozen, or were preserved in jars containing 70% ethanol until they were sorted, and C. obstrictus specimens counted and recorded. The numbers of C. obstrictus adults collected at each site were used to generate surface maps of cabbage seedpod weevil abundance with the Potential Mapping routine in SPANS GISTM v. 5.31 for OS/2 (PCI Geomatics). This procedure calculates a value for square cells which form the map as a weighted average of the values of the point data from the survey. A classification scheme was used to group cells according to the calculated weighted averages. Geographical distribution and abundance of Chalcidoidea parasitoids The distribution of Chalcidoidea parasitoids and their levels of parasitism of C. obstrictus larvae were assessed within infested areas of southern Alberta and Saskatchewan from 2003–2005. Commercial canola

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crops were sampled from late July to mid August each year when the majority of plants in the fields were in the pod ripening stage of development and seeds in the lower pods were translucent to green (Growth Stages 5.1–5.2 of Harper and Berkenkamp [1975]). A total of 147, 202, and 165 fields were sampled in 2003–2005, respectively. A single sample of approximately 450–850 canola siliques was collected from each field. Pods were collected randomly from canola racemes by sampling plants first at the field edges, and progressing inward over a distance of about 35–50 m. The geographical location of each field was determined using high resolution maps or hand-held global positioning system units as described previously. Pod samples were taken to the laboratory and placed into cardboard rearing containers, as described by Dosdall et al. (2006). Boxes were sealed with tape to prevent escape of C. obstrictus larvae or parasitoids, and were maintained under continuous light at room temperature for six weeks until all parasitoids had emerged. The boxes were then opened, and pod samples were examined under a stereoscopic microscope for the presence of any additional parasitoid adults that had not moved to the plastic collecting vials. Parasitoids were identified using published taxonomic keys and by comparison with type material or other authoritatively identified specimens. Voucher specimens of parasitoids have been deposited in the Canadian National Collection of Insects and Arachnids (CNC), Ottawa, ON, Canada. Pods from each box were examined and counts made of larval and parasitoid exit holes, which could be distinguished because exit holes of C. obstrictus larvae are circular but irregular along their margins, whereas those of weevil parasitoids are smaller and uniformly rounded along their margins (Dosdall et al. 2006). Although it is possible that some weevil and parasitoid larvae died before producing exit holes, our determinations of parasitism levels assumed no larval mortality. For each site, the level of parasitism was determined by dividing the number of weevil larvae parasitized by a particular species of Chalcidoidea by the total number of weevil hosts in the sample. The total number of specimens of a particular parasitoid species was determined by subtracting the number of individuals of that species in the sample from the total number of parasitoids that emerged, including both the species in question and other species. Samples were discarded when total numbers of emerged adult

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parasitoids did not equal total parasitoid exit holes from the pods. Parasitism levels of Chalcidoidea at each site were used to generate surface maps of abundance using ArcMapÒ Geographical Information Systems software with Spatial Analyst Extension (ESRI Services, Redlands, CA). The process examines point (site) data in relation to values from its nearest neighbor to estimate how values decline (decay) with distance from the point in question. The ‘‘interpolate to raster function’’ was used with the ‘‘inverse distance weighted option’’. A fixed search radius of 500 km and an output cell size of 1 km were set. For each year, the spatial distributions of Chalcidoidea adults were presented as contour maps of parasitism levels of C. obstrictus larvae.

Results Weevil geographical distribution patterns In 2001, C. obstrictus was widespread throughout fields of flowering canola in southern Alberta but uncommon in southern Saskatchewan (Fig. 1). The northern limit of its range was near Red Deer, AB (52°110 N; 113°540 W) where densities averaged only about 0.1 adult per sweep; however, further south near Calgary densities increased to nearly 0.4 per sweep. Greatest population densities were recorded in fields between Lethbridge (49°380 N; 112°480 W) and Medicine Hat, AB (50°10 N; 110°430 W), where mean densities exceeded 3.6 per sweep and in some fields exceeded 11.0 weevils per sweep. In 2002, adult weevil densities in southern Alberta declined somewhat to the east of Calgary; however, mean densities throughout large areas of canola cropland from Calgary to Lethbridge and east to the Saskatchewan border remained at approximately 1.0–4.0 adults per sweep (Fig. 1). Greatest densities of 8.4 per sweep were observed near Lethbridge. In Saskatchewan, weevil densities increased relative to those observed in 2001, with as many as 1.0 per sweep found in some fields west of Swift Current. Infestations of C. obstrictus declined substantially in Alberta from 2002 to 2003, with few adults collected in fields from Red Deer south to Lethbridge (Fig. 1). The most densely populated fields spanned an area extending from Lethbridge to Medicine Hat where populations were occasionally as high as 1.0 per sweep, but often

D.W. Langor & J. Sweeney (eds.)

only 0.2 per sweep. In Saskatchewan, the most densely populated fields were near its western border, with densities as high as 3.0 per sweep recorded. In 2004, infestations of the cabbage seedpod weevil in canola fields from Lethbridge north to Calgary increased relative to those found in 2003, with most fields having densities of approximately 0.3 per sweep (Fig. 1). As in previous years, the highest densities occurred from Lethbridge to Medicine Hat where mean densities ranged from 0.4 to 1.0 per sweep. In Saskatchewan, weevil abundance in 2004 was similar to that observed in 2003. Infestations of C. obstrictus increased substantially from 2004 to 2005, especially throughout southern Saskatchewan (Fig. 1). From Calgary south to Lethbridge and Medicine Hat, infestations ranged from approximately 0.4 to 3.6 per sweep. In southwestern Saskatchewan, densities of 3.6 per sweep were recorded near Maple Creek (49°550 N; 109°290 W), and vast areas of canola cropland from Swift Current to the U.S.A. border had populations of approximately 0.4 per sweep. Distribution and abundance of Chalcidoidea parasitoids In total, 12 species of Chalcidoidea representing four families were reared from canola pods infested with C. obstrictus during the three years of this study. The species reared were Necremnus tidius (Walker) and Euderus albitarsus (Zetterstedt) (Eulophidae), Trichomalus lucidus (Walker), Chlorocytus sp., Pteromalus sp., Mesopolobus moryoides Gibson, Lyrcus maculatus (Gahan), Lyrcus incertus (Ashmead), and Catolaccus aeneoviridis (Girault) (Pteromalidae), Conura torvina (Cresson) and Conura albifrons (Walsh) (Chalcididae), and Eurytoma tylodermatis Ashmead (Eurytomidae). Of these, N. tidius, T. lucidus, Chlorocytus sp., and Pteromalus sp. were the species most frequently reared, collectively comprising 53.8, 95.5, and 80.6% of all parasitoids reared in 2003–2005, respectively (Table 1). Of the remaining species, most were collected each year in low numbers, with the exception of L. incertus which was collected only in 2005 and E. tylodermatis which was collected only in 2004. In 2003, T. lucidus was reared from five canola fields (n = 147) infested with C. obstrictus larvae north of Lethbridge and west of Medicine Hat in

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Fig. 1 Distribution and abundance of C. obstrictus in canola in Alberta and Saskatchewan, Canada, from 2001 to 2005. Maps were developed from field survey data (25 sweep net samples per field) using SPANSÒ

Alberta, but it was not found in Saskatchewan (Fig. 2). The maximum levels of parasitism of C. obstrictus larvae by T. lucidus were 2–3%. In 2004, T. lucidus was found in seven fields (n = 202) in southern Alberta spanning a range extending from approximately 75 km north of the U.S.A. border to near Calgary. Parasitism levels ranged from 0–7%. In 2005, T. lucidus was reared from 14 sites (n = 165) throughout southern Alberta and from six fields in

southwestern Saskatchewan. Its range in Alberta extended from the foothills of the Rocky Mountains, east to Medicine Hat and south to the U.S.A. border. In Saskatchewan, the species was collected most commonly in fields north of Swift Current. Its maximum parasitism level reached 2–3% (Fig. 2). Chlorocytus sp. emerged from pod collections made in three canola fields infested with C. obstrictus larvae in Alberta in 2003, and from collections made

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Table 1 Species composition of Chalcidoidea parasitoids of Ceutorhynchus obstrictus reared from collections of siliques of Brassica napus and B. rapa from southern Alberta and Saskatchewan in 2003 to 2005 Parasitoid species

Species composition (%) 2003

2004

2005

8.1 6.7

4.8 7.5

12.0 26.9

13.8

8.2

13.8

Mesopolobus moryoides

2.4

0.2

1.1

Lyrcus maculatus

0.0

0.9

3.2

Pteromalidae Trichomalus lucidus Chlorocytus sp. Pteromalus sp.

Lyrcus incertus

0.0

0.0

0.4

13.8

0.2

1.1

25.2

75.0

27.9

0.8

1.1

1.1

Conura torvina

16.2

0.5

10.4

Conura albifrons

13.0

0.9

2.1

0.0

0.7

0.0

Catolaccus aenoviridis Eulophidae Necremnus tidius Euderus albitarsus Chalcididae

Eurytomidae Eurytoma tylodermatis

in four fields in Saskatchewan (Fig. 3). The maximum parasitism level was 1.5%. In 2004, Chlorocytus sp. emerged from 12 fields throughout southern Alberta and from six fields in Saskatchewan, spanning a range extending from Lethbridge north nearly to Calgary and east to Medicine Hat and Swift Current. Parasitism levels ranged from 0–6%. In 2005, Chlorocytus sp. was reared from 16 sites throughout southern Alberta and from 18 fields in southwestern Saskatchewan. Its range extended from the foothills of the Rocky Mountains, east to Swift Current, and south to the U.S.A. border. Its maximum parasitism level reached 10% (Fig. 3). Specimens of Pteromalus sp. were reared from eight canola fields infested with C. obstrictus larvae in Alberta and from two fields in Saskatchewan in 2003 (Fig. 4). The range of Pteromalus sp. extended from the U.S.A. border north to near Medicine Hat and Swift Current, but its densities were low in all fields with maximum parasitism levels of 1.5%. In 2004, Pteromalus sp. was reared from canola pods collected in eight fields in Alberta and six in Saskatchewan. In Alberta it occurred in fields near Lethbridge and Medicine Hat, but it was more abundant throughout southern Saskatchewan where

its maximum parasitism level was 7%. In 2005 the abundance and range of Pteromalus sp. increased in Saskatchewan, but not in Alberta. Specimens were reared from eight fields in Alberta in 2005 and from 17 fields in southern Saskatchewan. Maximum parasitism levels were 9% in Saskatchewan and 2% in Alberta (Fig. 4).

Discussion Our surveys to monitor geographical changes in the distribution and abundance of the cabbage seedpod weevil indicate that dispersal rate of the species has slowed considerably since the years immediately following its discovery in southern Alberta. Dosdall et al. (2002) reported that from 1997–2000 C. obstrictus extended its range northward and eastward at a rate of approximately 55 km per year. In the outbreak of 2000, hundreds of thousands of hectares throughout southern Alberta harbored weevil densities that equaled or exceeded the economic threshold for insecticidal application (Dosdall et al. 2001). Since then, populations have never met or exceeded 2000 densities, and our surveys have indicated considerable elasticity in how the species range has varied over time. The most dramatic change in weevil population densities occurred from 2002–2003 when the geographical area with populations that equaled or exceeded the economic threshold density in canola (3.0–4.0 per sweep) declined to zero throughout all of southern Alberta. One likely explanation for this weevil population decline between years is related to lack of moisture. During 2001 and 2002, precipitation levels throughout much of western Canada and all of southern Alberta and Saskatchewan were substantially lower than long-term normal values; in fact, precipitation levels at several sites within this region were the lowest ever recorded (Wheaton et al. 2005). Drought could have severely affected C. obstrictus populations in several ways. Larval development and survival within siliques could have been reduced because canola host plants were unable to absorb and translocate sufficient quantities of water from the soil. In addition, drought can leave the soil surface dry, crusted, and hard, making it difficult for final instars that have exited from siliques to burrow below the soil surface for pupation. Lengthening the time required by larvae to burrow beneath the soil surface

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems Fig. 2 Changes from 2003 to 2005 in the distribution and level of parasitism of T. lucidus on C. obstrictus. The legend indicates the percentages of weevilinfested canola pods parasitized by T. lucidus in southern Alberta and Saskatchewan

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Trichomalus lucidus Alberta

Saskatchewan

Legend 0.0 - 0.5

Calgary

0.5 - 1.0 1.0 - 1.5 Swift Current Medicine Hat

1.5 - 2.0 2.0 - 3.0

Lethbridge

3.0 - 4.0 4.0 - 5.0 5.0 - 6.0

2003

6.0 - 7.0 Alberta

Saskatchewan

Calgary

Swift Current Medicine Hat Lethbridge

2004

Alberta

Saskatchewan

Calgary

Swift Current Medicine Hat Lethbridge

2005

would expose them to greater potential mortality from opportunistic carnivores like species of Carabidae (Coleoptera). Finally, drought conditions would have reduced the availability of host plants for feeding by new generation adults in late summer prior to their migration to overwintering sites. With fewer late-season hosts available, it would be more difficult for weevils to build up the fat stores needed for survival during winter, resulting in lower populations in the following year. Although the northward expansion of the geographical range of C. obstrictus has been minimal

from 2001 to 2005, the weevil has extended its range eastward to currently include much of southern Saskatchewan. Since 2005, C. obstrictus continued to spread east in Saskatchewan and was collected in a sample near Moose Jaw (50°200 N; 105°300 W) in 2006 (Olfert et al. 2007). The species range is expected to eventually encompass southern Manitoba because there are no geographical or biotic barriers to hamper its eastward dispersal. Once in Manitoba, however, it is probable that crop losses from weevil infestation will exceed those experienced in southern Alberta and Saskatchewan because ecoclimatic index

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Fig. 3 Changes from 2003 to 2005 in the distribution and level of parasitism of Chlorocytus sp. on C. obstrictus. The legend indicates the percentages of weevil-infested canola pods parasitized by Chlorocytus sp. in southern Alberta and Saskatchewan

Chlorocytus sp. Alberta

Saskatchewan

Legend 0.0 - 0.5

Calgary

0.5 - 1.0 1.0 - 1.5 Swift Current Medicine Hat

1.5 - 2.0 2.0 - 3.0

Lethbridge

3.0 - 4.0 4.0 - 6.0 6.0 - 8.0

2003

8.0 - 10.0 Alberta

Saskatchewan

Calgary

Swift Current Medicine Hat Lethbridge

2004

Alberta

Saskatchewan

Calgary

Swift Current Medicine Hat Lethbridge

2005

values developed with the CLIMEXÒ model were considerably higher for southern Manitoba than for other regions of western Canada (Dosdall et al. 2002). In addition, changes in environmental conditions (particularly temperature and precipitation) throughout western Canada associated with climate change scenarios are predicted to result in a northerly expansion of the geographical range of C. obstrictus by 400–600 km (Olfert and Weiss 2006). The economic costs of the invasion of western Canada by C. obstrictus have been substantial. Although a central problem in deriving accurate and comprehensive assessments of damage caused by

invasive species is the paucity of available data to systematically track direct and indirect, market and non-market costs for species (Colautti et al. 2006), data for the impact of the cabbage seedpod weevil on canola producers can at least be estimated broadly based on costs incurred to apply insecticide on cropland. On this basis alone, it is estimated that a total of 180,000 ha of canola and brown mustard crops were sprayed with insecticide to reduce C. obstrictus infestations from 1999 to 2005 inclusive, at an estimated cost of $4.5 million (CAD) (Western Committee on Crop Pests 2001, 2003, 2004; Dosdall unpublished data). This cost is an underestimate of total economic loss

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems Fig. 4 Changes from 2003 to 2005 in the distribution and level of parasitism of Pteromalus sp. on C. obstrictus. The legend indicates the percentages of weevil-infested canola pods parasitized by Pteromalus sp. in southern Alberta and Saskatchewan

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Pteromalus sp. Alberta

Saskatchewan

Legend 0.0 - 0.5

Calgary

0.5 - 1.0 1.0 - 1.5 Swift Current Medicine Hat

1.5 - 2.0 2.0 - 3.0

Lethbridge

3.0 - 4.0 4.0 - 5.0 5.0 - 7.0

2003

7.0 - 9.0

Alberta

Saskatchewan

Calgary

Swift Current Medicine Hat Lethbridge

2004

Alberta

Saskatchewan

Calgary

Swift Current Medicine Hat Lethbridge

2005

because not all farmers sprayed insecticide when their crops had weevil populations at or near the economic threshold, and crops with densities below the threshold still incurred some damage. Dosdall and Dolinski (2001) reported that many canola pod samples were dissected in 1999 and 2000 and examined for the presence of larval parasitoids, but none were discovered. Fox et al. (2004) found that the braconid endoparasitoid, Microctonus melanopus (Ruthe), occasionally infested adult weevils in southern Alberta, but its occurrence was rare. Parasitoids

were first discovered attacking C. obstrictus larvae in southern Alberta in 2001 (Dosdall unpublished data). A low incidence of larval parasitism of C. obstrictus was reported in 2002 in a field near Lethbridge, and both the numbers of parasitoid species recorded and the incidence of parasitism increased during 2003 and 2004 (Dosdall et al. 2006). Parasitism rates increased from 0.1 to 5.0% during the three-year period from 2002 to 2004 (Dosdall et al. 2006). Our current study involved sampling many additional sites across a broad geographic scale and confirmed that parasitism of

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Table 2 Species of Chalcidoidea reared from canola siliques in western Canada, their known distributions and host records Chalcidoidea species

Distribution

Host(s)

Holarctic

Coleoptera: Curculionidae

Citations

Pteromalidae Trichomalus lucidus (Walker)

Ceutorhynchus obstrictus (Marsham)

Gibson et al. (2005), Dosdall et al. (2006), Present study

Ceutorhynchus neglectus Blatchley

Dosdall et al. (2007)

Ceutorhynchus alliariae Brisout

Kuhlmann et al. (2000)

Ceutorhynchus pallidactylus (Marsham)

Kuhlmann et al. (2000)

Ceutorhynchus roberti Gyllenhal

Klander (2001)

Ceutorhynchus scrobicollis Nerensheimer & Wagner

Muller et al. (2003)

Microplontus rugulosus (Herbst)

Muller et al. (2004)

Coleoptera: Chrysomelidae—1 species Chlorocytus sp.

Unknown

Coleoptera: Curculionidae C. obstrictus

Pteromalus sp.

Unknown

Coleoptera: Curculionidae C. obstrictus

Lyrcus maculatus (Gahan)

Nearctic

Dosdall et al. (2006), Present study Dosdall et al. (2006), Present study

Coleoptera: Curculionidae C. obstrictus

Gibson et al. (2006b), Dosdall et al. (2006), Present study

Desmoris fulvus LeConte Rhinocyllus conicus Froehlich Smicronyx fulvus LeConte Miccotrogus picirostris (Fabricius) Lyrcus perdubius (Girault)

Nearctic

Coleoptera: Curculionidae C. obstrictus

Dosdall et al. (2006)

Anthonomus grandis Boheman Anthonomus musculus Say Anthonomus nubiloides Fall Anthonomus signatus Say Smicronyx tychoides LeConte Lyrcus incertus (Ashmead)

Nearctic

Coleoptera: Curculionidae C. obstrictus Anthonomus aeneolus Dietz Anthonomus albopilosus Dietz Anthonomus aphanostephi Pierce Anthonomus disjunctus LeConte Anthonomus eugenii Cano Anthonomus fulvus LeConte A. grandis Anthonomus nigrinus Boheman A. signatus

Gibson et al. (2006b), Present study

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119

Table 2 continued Chalcidoidea species

Distribution

Host(s)

Citations

Auleutes tenuipes (LeConte) Baris sp. Baris cuneipennis Casey Chalcodermus aenus Boheman Chalcodermus inaequicollis Horn D. fulvus Eutinobothrus brasiliensis (Hambleton) Rhynchaenus pallicornis (Say) Tachypterellus quadrigibbus Say Coleoptera: Dryophthoridae—2 species Lepidoptera: Gelechiidae—2 species Coleoptera: Apionidae—5 species Coleoptera: Bruchidae—4 species Diptera: 1 species in 1 family Hymenoptera: 4 species in 1 family Mesopolobus moryoides Gibson

Nearctic

Mesopolobus bruchophagi Gahan

Nearctic

Coleoptera: Curculionidae C. obstrictus

Gibson et al. (2005, 2006a, b), Dosdall et al. (2006), Present study

Coleoptera: Curculionidae C. obstrictus

Dosdall et al. (2006)

Coleoptera: Bruchidae—1 species Diptera: 2 species in 2 families

Catolaccus aeneoviridis (Girault)

Nearctic

Hymenoptera: 9 species in 3 families Coleoptera: Curculionidae C. obstrictus

Present study

Hypera postica (Gyllenhal) Microlarinus lypriformis (Wollaston) Lepidoptera: 36 species in 13 families Hymenoptera: 26 species in 3 families

Stenomalina gracilis (Walker)

Palearctic

Coleoptera: Curculionidae C. obstrictus

Gibson et al. (2006a), Gillespie et al. (2006), Muller et al. (2004), Haye et al. (2005)

Ceutorhynchus napi (Gyllenhal)

Klukowski and Kelm (2000)

C. pallidactyllus

Kuhlmann and Mason (2002)

C. roberti

Kuhlmann and Mason (2002)

Hadroplontus litura (Fabricius)

Kuhlmann and Mason (2002)

Microplontus edentulus (Schultze)

Kuhlmann and Mason (2002)

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Table 2 continued Chalcidoidea species

Distribution

Host(s) M. rugulosus

Citations Muller et al. (2004)

Lepidoptera: 1 species in 1 family Hymenoptera: 1 species in 1 family Diptera: 15 species in 5 families Eulophidae Necremnus tidius (Walker)

Holarctic

Coleoptera: Curculionidae C. obstrictus

Gibson et al. (2005), Gibson et al. (2006b), Gillespie et al. (2006), Dosdall et al. (2006), Dosdall et al. (2007), Present study

Ceutorhynchus floralis (Paykull)

Kuhlmann and Mason (2002)

Ceutorhynchus sp.

Gomez and Zamora (1994)

Coleoptera: Apionidae—1 species Coleoptera: Chrysomelidae—2 species Diptera: 2 species in 2 families Lepidoptera: 2 species in 1 family Euderus Holarctic albitarsis (Zetterstedt)

Coleoptera: Curculionidae C. obstrictus

Dosdall et al. (2006), Present study

Coleoptera: Cerambycidae—1 species Coleoptera: Scolytidae—1 species Diptera: 3 species in 2 families Lepidoptera: 16 species in 7 families Hemiptera: 1 species in 1 family Hymenoptera: 2 species in 2 families Chalcididae Conura albifrons (Walsh)

Nearctic

Coleoptera: Curculionidae C. obstrictus

Dosdall et al. (2006), Present study

Hypera eximia (LeConte) Hympera paludicola Warner H. postica Coleoptera: Buprestidae—1 species Coleoptera: Chrysomelidae—7 species Diptera: 1 species in 1 family Lepidoptera: 36 species in 11 families Hymenoptera: 12 species in 2 families Conura torvina (Cresson)a

Nearctic

Coleoptera: Curculionidae C. obstrictus H. postica Coleoptera: Chrysomelidae—1 species Lepidoptera: 2 species in 2 families Hymenoptera: 4 species in 2 families

Gibson et al. (2006b), Dosdall et al. (2006), Present study

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121

Table 2 continued Chalcidoidea species

Distribution

Host(s)

Citations

Nearctic

Coleoptera: Curculionidae

Eurytomidae Eurytoma tylodermatis Ashmead

C. obstrictus

Gibson et al. (2006b), Present study

Anthonomus consors (Dietz) Anthonomus disjunctus LeConte Anthonomus grandis (Boheman) Anthonomus heterothecae Pierce Anthonomus quadrigibbus Say Anthonomus squamosus LeConte Anthonomus vestitus (Boheman) Baris sp. Chalcodermus aeneus (Boheman) Conotrachelus elegans (Say) Cylindrocopturus adspersus (LeConte) Cylindrocopturus longulus (LeConte) Lixus musculus Say Lixus scrobicollis Boheman Macrorhoptus sphaeralciae (Pierce) Orthoris crotchii (LeConte) Smicraulax tuberculatus (Pierce) Trichobaris texana LeConte Trichobaris trinotata (Say) Tyloderma foveolatum (Say) Coleoptera: Anthribidae—1 species Coleoptera: Apionidae—1 species Coleoptera: Bruchidae—5 species Diptera: 1 species in 1 family Lepidoptera: 12 species in 5 families Hymenoptera: 7 species in 4 families Host species and citations documenting records are provided only for Ceutorhynchinae species; other host taxa listed are summarized from Noyes (2003) a The host list given for C. torvina by Noyes (2003) is inaccurate because this species was consistently misidentified as Conura side (Walker) prior to Delvare (1992) and most of the 86 species (in 24 families of 3 orders) listed as hosts of C. side by Noyes (2003) likely refer to C. torvina (Gibson et al. 2006a)

C. obstrictus populations was dominated by larval ectoparasitoid species, most of which appear to be native to North America. The larval parasitoid species, comprising T. lucidus, Chlorocytus sp., Pteromalus sp., and N. tidius, were identified as principal components of the parasitoid fauna of C. obstrictus as early as 2002 (Dosdall et al. 2006), and they remain as the key parasitoids attacking this species.

Of the 12 Chalcidoidea species reared from siliques infested with C. obstrictus, only C. aeneoviridis has not been reported previously as reared from canola. This species is most commonly reared as a hyperparasitoid of Lepidoptera cocoons through species of Braconidae and Ichneumonidae (Hymenoptera), but it has also been reared as a hyperparasitoid of the alfalfa weevil, Hypera postica

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(Gyllenhal) (Coleoptera: Curculionidae), as has C. albifrons (Putler 1966; Rethwisch and Manglitz 1986) and C. torvina (Rethwisch and Manglitz 1986). Rearings of parasitoids from mass collections of canola siliques can be associated with introduction of contaminant insect hosts, and in this study larvae of the diamondback moth, Plutella xylostella (L.) (Lepidoptera: Plutellidae), occurred occasionally in some pod collections. Specimens of C. aeneoviridis, C. albifrons, and C. torvina reared in this study are likely hyperparasitoids of C. obstrictus or perhaps hyperparasitoids of P. xylostella. The Chalcidoidea parasitoid fauna of C. obstrictus in western Canada consists of some species with restricted host ranges, like T. lucidus which but for a single exception is only known from Ceutorhynchinae weevil hosts (Table 2). The only reported nonceutorhynchine host of T. lucidus is the cabbage stem flea beetle, Psylliodes chrysocephala (L.) (Chrysomelidae), another pest of canola in Europe (Vidal 1997). Other species, like the presumed hyperparasitoids and E. albitarsis, E. tylodermatis, and L. maculatus are known from a wider range of hosts from several insect orders and families (Table 2). These species are probably more niche- than taxonspecific. The invasion of western Canadian canola agroecosystems by C. obstrictus has evidently provided abundant populations of weevil larvae in siliques that enabled parasitoid species to switch from their native host insects to exploit this new resource. The Chalcidoidea fauna of C. obstrictus includes species with both Nearctic and Holarctic distributions, but in general, the hyperparasitoids are Nearctic and the primary parasitoids are Holarctic (Table 2). It is uncertain whether the Holarctic distributions of T. lucidus and N. tidius are natural or whether they were inadvertently introduced from Europe with unknown host insects. Neither T. lucidus nor N. tidius is a principal parasitoid of C. obstrictus in Europe (Williams 2003). Ceutorhynchus obstrictus is the only known host of M. moryoides, a species currently known only from the Nearctic region. Mesopolobus moryoides could be a native species with an unknown native host, but it also remains possible that it is an unrecognized adventive European species. The origins of the Chlorocytus and Pteromalus species will remain unknown until species determinations can be made. Europe and North

D.W. Langor & J. Sweeney (eds.)

America both have a rich fauna of Pteromalus, whereas Chlorocytus is much more speciose in Europe than in North America (Graham 1969). The taxonomic status of Chlorocytus sp. and Pteromalus sp. should be clarified to determine whether these species are native, naturally Holarctic, or one or both are introduced European species. The levels of parasitism of T. lucidus, Chlorocytus sp., and Pteromalus sp. on C. obstrictus varied considerably over the three years of this study. In 2003, all three species were rarely collected, probably because the weevil population was reduced from drought during the preceding years. In 2004, however, populations of both Pteromalus sp. and Chlorocytus sp. increased, especially in western Saskatchewan, reaching parasitism levels of approximately 7% in some sites. Further population increases were observed in 2005. Increased levels of parasitism by Pteromalus sp. occurred in eastern Alberta and western Saskatchewan, and an even greater increase in the level of parasitism of Chlorocytus sp. occurred throughout both provinces. Increases in parasitism levels of T. lucidus also occurred, but not to the same magnitude as for Chlorocytus sp. and Pteromalus sp. These observations concur with those reported by Dosdall et al. (2007) for N. tidius, who found that low parasitism levels in 2003 increased in 2004 and 2005 for this species. The large fluctuations both in the levels of parasitism of C. obstrictus and the dominance of the species comprising the parasitoid fauna could represent effects of competition among Chalcidoidea for C. obstrictus larvae. Invasion of western Canada by C. obstrictus has followed the generalized steps for invasive species outlined by Sakai et al. (2001) where initial transport and establishment in a new area (in this case southwestern British Columbia) was followed by a lag period of population build-up before further dispersal to new habitats (first throughout the Pacific Northwest, then to eastern United States, and later northward to Canada). As with other successful invaders, the cabbage seedpod weevil possesses several traits like small body size, rapid time to maturity, and a high intrinsic growth rate that facilitated establishment of viable, self-sustaining populations (Crawley 1986; Lawton and Brown 1986; Dosdall et al. 2001; Dosdall and Moisey 2004). But perhaps most importantly, its invasion was successful because it could occupy an

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

‘‘empty niche’’ (Mack 1996; Levine and D’Antonio 1999). When C. obstrictus dispersed to southern Alberta and Saskatchewan it encountered a vast resource of many billions of siliques of B. napus and B. rapa that was largely unexploited by other herbivores and that provided a nutritious, high energy food resource for its larvae where they were protected from surface-dwelling predators. This study and those of Gibson et al. (2005, 2006a) and Dosdall et al. (2006) documented a surprisingly diverse fauna of Chalcidoidea species that have responded to the invasion of C. obstrictus, so that these parasitoids have now become, in turn, invasive of canola agroecosystems infested with C. obstrictus larvae. Along with their hosts, a lag period of low parasitoid population densities (Dosdall and Dolinski 2001) was followed by years of increased numbers (Dosdall et al. 2006), presumably as their populations expanded both numerically and geographically. It is doubtful, however, that the community structure in these systems has yet stabilized. For instance, the fauna of hyperparasitoid species, like C. aenoviridis and Conura spp., may now perhaps increase to exploit the parasitoids. Implementing a classical biological control program for the cabbage seedpod weevil could enable partial reconstruction of the natural enemy complex of this insect, and so prevent or reduce its continued dispersal across western Canada. The ability of introduced natural enemies to persist in the environment, to reproduce, and to spread are important advantages of the biological control approach (Greathead 1986). A first attempt at classical biological control of C. obstrictus in North America involved releases of three species of pteromalid larval ectoparasitoids in southwesten British Columbia in 1949 (Gibson et al. 2006a; Gillespie et al. 2006). Examination of voucher specimens from the releases determined that the species introduced included Mesopolobus morys (Walker), Trichomalus perfectus (Walker), and Stenomalina gracilis (Walker) (Pteromalidae) (Gibson et al. 2006a). An extensive series of surveys conducted near the original release sites found that of these, only S. gracilis appears to have established (Gibson et al. 2006a; Gillespie et al. 2006). However, M. morys and T. perfectus are considered to be the most effective of the released parasitoids for controlling C. obstrictus in Europe (Williams 2003), and therefore hold the greatest potential for facilitating sustainable biological control

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of the weevil in North America. Gillespie et al. (2006) proposed that re-releases of one or both of T. perfectus and M. morys should be considered after taking into account the host ranges of these species. But the question arises of whether additional releases of European parasitoids are warranted when several species of native natural enemies already attack the weevil. Results reported here and those of Dosdall et al. (2006, 2007) indicate that the present parasitoid fauna has sometimes caused substantial levels of parasitism, but current parasitism levels (usually less than 15% for all species combined) are not sufficiently high to adequately control the weevil population. It is not clear whether parasitism levels are still likely to increase over time, since the current situation has shown considerable year-to-year variation (Figs. 2, 3, 4). Parasitism levels should therefore continue to be monitored, along with the dispersal biology of the C. obstrictus host, to assess whether a classical biological control program should be implemented. Acknowledgments We gratefully acknowledge funding awarded for this project from the Alberta Agricultural Research Institute, the canola producer commissions of Alberta, Saskatchewan, and Manitoba, the University of Alberta, Alberta Agriculture and Food, Agriculture and AgriFood Canada, and the Agriculture and Agri-Food Canada Pest Management Centre Improving Farming Systems Program Project PRR03370. We gratefully acknowledge technical assistance from R. Adams, N. Cowle, B. Peake, and K. Peake. Special thanks are extended to M. Braun for conducting the Saskatchewan surveys, and to M. Herbut and D. Giffen for assistance with preparing the distribution maps.

References Anonymous (1977) Cabbage seedpod weevil (Ceutorhynchus assimilis)—Maryland—new state record. United States Department of Agriculture, Cooperative Plant Pest Report 2, 812 Baker WW (1936) Notes on the European weevil, Ceutorhynchus assimilis Payk., recently found in the state of Washington. Can Entomol 8:191–193 Bonnemaison L (1957) Le charanc¸on des siliques (Ceutorhynchus assimilis Payk.) biologie et me´thode de lutte. Annales des E´piphytes 4:387–543 Boyd ML, Lentz GL (1994) Seasonal incidence of the cabbage seedpod weevil (Coleoptera: Curculionidae) on rapeseed in west Tennessee. Environ Entomol 23:900–905 Brodeur J, Leclerc LA, Fournier M, Roy M (2001) Cabbage seedpod weevil (Coleoptera: Curculionidae): a new pest of canola in northeastern North America. Can Entomol 133:709–711

124 Buntin GD, McCaffrey JP, Raymer PL, Romero J (1995) Quality and germination of rapeseed and canola seed damaged by adult cabbage seedpod weevil, Ceutorhynchus assimilis Paykull (Coleoptera: Curculionidae). Can J Plant Sci 75:539–541 Colautti RI, Bailey SA, van Overdijk CDA, Amundsen K, MacIsaac HJ (2006) Characterised and projected costs of nonindigenous species in Canada. Biol Invasions 8:45–59 Crawley MJ (1986) The population biology of invaders. Philos Trans R Soc Lond Ser B 314:711–29 Crowell HH (1952) Cabbage seedpod weevil control with parathion. J Econ Entomol 45:545–546 Delvare G (1992) A reclassification of the Chalcicini with a checklist of the New World species. Mem Am Entomol Inst 53:119–457 Dmoch J (1965) The dynamics of a population of the cabbage seedpod weevil (Ceutorhynchus assimilis Payk.) and the development of winter rape. Part I. Ekologia Polska Seria A 13:249–287 Dosdall LM, Dolinski MG (2001) Biology and control of the cabbage seedpod weevil, a new pest of canola in Alberta. Alberta Agriculture, Food and Rural Development Technical Report, Edmonton, AB, 65 pp Dosdall LM, Moisey D, Ca´rcamo H, Dunn R (2001) Cabbage seedpod weevil factsheet. Alberta Agriculture, Food and Rural Development Agdex 622-21, 4 pp Dosdall LM, Weiss RM, Olfert O, Ca´rcamo HA (2002) Temporal and geographical distribution patterns of cabbage seedpod weevil (Coleoptera: Curculionidae) in canola. Can Entomol 134:403–418 Dosdall LM, McFarlane MA (2004) Morphology of the preimaginal life stages of the cabbage seedpod weevil, Ceutorhynchus obstrictus (Marsham) (Coleoptera: Curculionidae). Coleopt Bull 58:45–52 Dosdall LM, Moisey DWA (2004) Developmental biology of the cabbage seedpod weevil, Ceutorhynchus obstrictus (Coleoptera: Curculionidae), in spring canola, Brassica napus, in western Canada. Ann Entomol Soc Am 97:458–465 Dosdall LM, Ulmer BJ, Gibson GAP, Ca´rcamo HA (2006) The spatio-temporal distribution dynamics of the cabbage seedpod weevil, Ceutorhynchus obstrictus (Coleoptera: Curculionidae), and its larval parasitoids in canola in western Canada. Biocont Sci Technol 16:987–1006 Dosdall LM, Gibson GAP, Olfert O, Keddie BA, Ulmer BJ (2007) Contributions to the life history, host range, and distribution of Necremnus tidius (Walker) (Hymenoptera: Eulophidae). Ann Entomol Soc Am 100:861–868 Fox AS, Dosdall LM (2003) Reproductive biology of Ceutorhynchus obstrictus (Coleoptera: Curculionidae) on wild and cultivated Brassicaceae in southern Alberta. J Entomol Sci 38:365–376 Fox AS, Shaw SR, Dosdall LM, Lee B (2004) Microctonus melanopus (Ruthe) (Hymenoptera: Braconidae), a parasitoid of adult cabbage seedpod weevil (Coleoptera: Curculionidae): Distribution in southern Alberta and female diagnosis. J Entomol Sci 39:350–361 Gibson GAP, Baur H, Ulmer B, Dosdall L, Muller F (2005) On the misidentification of chalcid (Hymenoptera: Chalcidoidea) parasitoids of the cabbage seedpod weevil (Coleoptera: Curculionidae) in North America. Can Entomol 137:381–403

D.W. Langor & J. Sweeney (eds.) Gibson GAP, Gillespie DR, Dosdall L (2006a) The species of Chalcidoidea (Hymenoptera) introduced to North America for biological control of the cabbage seedpod weevil, and the first recovery of Stenomalina gracilis (Chalcidoidea: Pteromalidae). Can Entomol 138:285–291 Gibson GAP, Gates MW, Buntin GD (2006b) Parasitoids (Hymenoptera: Chalcidoidea) of the cabbage seedpod weevil (Coleoptera: Curculionidae) in Georgia, USA. J Hymenopt Res 15:187–207 Gillespie DR, Mason PG, Dosdall LM, Bouchard P, Gibson GAP (2006) Importance of long-term research in classical biological control: An analytical review of a release against the cabbage seedpod weevil in North America. J Appl Entomol 130:401–409 Gomez JM, Zamora R (1994) Top-down effects in a tritrophic system: parasitoids enhance plant fitness. Ecology 75: 1023–1030 Graham MWRdeV (1969) The Pteromalidae of North-western Europe (Hymenoptera: Chalcidoidea). Bull Br Mus (Nat His) Entomol, 908 pp Greathead DJ (1986) Parasitoids in classical biological control. In: Waage J, Greathead D (eds) Insect Parasitoids. Academic Press, London, pp 289–318 Hagen KS (1946) The occurrence of Ceutorhynchus assimilis (Paykull) in California. Pan-Pac Entomol 22:73 Harper FR, Berkenkamp B (1975) Revised growth-stage key for Brassica campestris and B napus. Can J Plant Sci 55:657–658 Haye T, Muller F, Kuhlmann U (2005) Biological control of cabbage seedpod weevil, Ceutorhynchus obstrictus. Annual Report 2005–2006. Unpublished report, CABI Bioscience Switzerland Centre, Dele´mont Klander B (2001) Die Ru¨sselka¨fer der Unterfamillie Ceutorhynchinae und deren Parasitoide auf Winterraps und begleitenden Unkra¨utern in Schleswig-Holstein. MSc thesis. Christian Albrechts-University, Kiel Klukowski Z, Kelm M (2000) Stenomalina gracilis (Walker), a new parasitoid reared from Ceutorhynchus napi Gyll. in Poland. Bulletin de la Section Re´gionale de l’Ouest-Palaearctique, Organisation Internationale de Lutte Biologique 23:135–138 Kuhlmann U, Mason PG (2002) Use of field host range surveys for selecting candidate non-target species for physiological host specificity testing of entomophagous biological control agents. In: van Driesche RG (ed) Proceedings of the first international symposium on biological control of arthropods, Honolulu, Hawaii. United States Department of Agriculture, Forest Service, Washington, USA, pp 370–377 Kuhlmann U, Toepfer S, Grossrieder M, White H, Klander B, Reimer L, Hunt E, Higginson D, Jenner W (2000) Summary Report—Progress in 2000. Agricultural Pest Research, Unpublished Report, CABI Bioscience Switzerland Centre, Dele´mont Laffin RD, Dosdall LM, Sperling FAH (2005) Population structure of the cabbage seedpod weevil, Ceutorhynchus obstrictus (Marsham) (Coleoptera: Curculionidae): origins of North American introductions. Environ Entomol 34:504–510 Lawton JH, Brown KC (1986) The population and community ecology of invading insects. Philos Trans R Soc Lon Ser B 314:607–616

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Putler B (1966) Biological notes on some hyperparasites of Bathyplectes curculionis (Thomson). J Econ Entomol 59:483–484 Rethwisch MD, Manglitz R (1986) Parasitoids of Bathyplectes curculionis (Hymenoptera: Ichneumonidae) in southeastern Nebraska. J Kans Entomol Soc 59:648–652 Sakai AK, Allendorf FW, Holt JS, Lodge DM, Molofsky J, With KA, Baughman S, Cabin RJ, Cohen JE, Ellstrand NC, McCauley DE, O’Neil P, Parker IM, Thompson JN, Weller SG (2001) The population biology of invasive species. Annu Rev Ecol Syst 32:305–332 Ulmer BJ, Dosdall LM (2006) Spring emergence biology of the cabbage seedpod weevil (Coleoptera: Curculionidae). Ann Entomol Soc Am 99:64–69 Vidal S (Ed.) (1997) Determination list of entomophagous insects. Nr. 13. Bulletin. Section Re´gionale Ouest Palaearctique, Organisation Internationale de Lutte Biologique 20: 26 Walz AJ (1957) Observations on the biologies of some hymenopterous parasites of the cabbage seedpod weevil in northern Idaho. Ann Entomol Soc Am 50:219–220 Western Committee on Crop Pests (2001) Minutes of the 41st annual meeting, October 15. Banff, Alberta, 71 pp Western Committee on Crop Pests (2003) Minutes of the 43rd annual meeting, October 31. Kelowna, British Columbia, 88 pp Western Committee on Crop Pests (2004) Minutes of the 44th annual meeting, October 13. Saskatoon, Saskatchewan, 76 pp Wheaton E, Wittrock V, Kulshreshtha S, Koshida G, Grant C, Chipanski A, Bonsal B (2005) Lessons learned from the Canadian drought years of 2001 and 2002: synthesis report. Saskatchewan Research Council Publication 11602-46E03, 30 pp Williams IH (2003) Parasitoids of cabbage seed weevil. In: Alford DV (ed) Biocontrol of oilseed rape pests. Blackwell, Oxford, pp 97–112

Non-native insects in agriculture: strategies to manage the economic and environmental impact of wheat midge, Sitodiplosis mosellana, in Saskatchewan O. Olfert Æ R. H. Elliott Æ S. Hartley

Originally published in the Journal Biological Invasions, Volume 11, No 1, 127–133. DOI: 10.1007/s10530-008-9324-0 Ó Her Majesty the Queen in Right of Canada 2008

Abstract Wheat midge, Sitodiplosis mosellana (Ge´hin) (Diptera: Cecidomyiidae), was first detected as early as 1901 in western Canada. The first major outbreak in Saskatchewan was recorded in 1983. Today wheat midge infests much of the wheat-growing area of Manitoba, Saskatchewan and North Dakota (USA), and is beginning to invade Alberta and Montana (USA). In 1984, Saskatchewan wheat midge populations were found to be parasitized by the egg-larval parasitoid, Macroglenes penetrans (Kirby) (Hymenoptera). Through the successful implementation of conservation techniques, this parasitoid now controls an average of 31.5% of the wheat midge across Saskatchewan. Estimated value of the parasitoid, due to reduction in insecticide costs in Saskatchewan alone, was estimated to be in excess of $248.3 million in the 1990s. The environmental benefits of not having to apply this amount of chemical insecticide are a bonus. To minimize the economic and ecological impact of S. mosellana today, wheat producers in western Canada have access to one of the most comprehensive management programs of any insect pest of field crops.

O. Olfert (&)
R. H. Elliott Saskatoon Research Centre Agriculture & Agri-Food Canada, 107 Science Place, Saskatoon, SK, Canada S7N 0X2 e-mail: [email protected] S. Hartley Saskatchewan Agriculture and Food, 133-3085 Albert Street, Regina, SK, Canada S4S 0B1

Forecasts and risk warnings, monitoring tools, cultural control, agronomic practices, chemical control, biological control and plant resistance are all available for producers to manage wheat midge. Keywords Biological control
Conservation
Spatial analysis
Wheat midge

Introduction Wheat midge, Sitodiplosis mosellana (Ge´hin) (Diptera: Cecidomyiidae), is Palearctic in origin and was accidentally introduced into North America in the 1800s (Felt 1912). It is now a chronic pest of wheat (Triticum spp.) in the northern Great Plains, including the Canadian prairies and is widely distributed in many parts of the world where wheat production occurs, especially between the 42nd and 62nd parallels (Affolter 1990). In western Canada, S. mosellana was first reported in Manitoba (Fletcher 1902) but was not considered to be a pest until the 1950s (Allen 1955). The first major outbreak of wheat midge in Saskatchewan was recorded in 1983 causing an estimated loss in yield of $30 million that year (Olfert et al. 1985). The life cycle of S. mosellana has been reviewed by Barnes (1956) for the British Isles and Europe, by Reeher (1945) for British Columbia and Washington, and by Doane et al. (1987) and Mukerji et al. (1988)

D.W. Langor & J. Sweeney (eds.), Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems. DOI: 10.1007/978-1-4020-9680-8_9

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for Saskatchewan. On the northern Great Plains, adults emerge over a six-week period beginning in late June or early July. The highest populations usually occur during the second or third week of July. Adults are relatively poor fliers and may be distributed over long distances by thermal updrafts and wind. They are difficult to detect during the day because they remain within the crop canopy close to ground level where it is more humid. Females become more active in the evening. Most egg-laying occurs at dusk when conditions are calm and temperatures are above 10–11°C (Pivnick and Labbe´ 1993). Females live 3–7 days and lay an average of 80 eggs underneath the glumes or on grooves on the floret surface. Eggs are laid singly or in clusters of up to four eggs on the florets of emerging wheat heads. Each larva crawls into a floret and feeds on the kernel surface for 2–3 weeks. Mature larvae remain within their cast skin in the wheat head when conditions are dry. Once moist conditions are detected, larvae drop to the ground, burrow into the soil, spin a larval cocoon and overwinter. The following spring, further larval development depends on temperature and soil moisture. If conditions are dry during May and June, larvae remain dormant until the following year; if moist, larvae leave their cocoons and move to the soil surface to pupate (Basedow and Schu¨tte 1973; Doane et al. 1987; Elliott and Mann 1996). Canadian varieties of hard red spring wheat, durum wheat and soft spring wheat differ in their susceptibility to damage (Elliott et al. 2000a, b). The extent of crop damage due to wheat midge depends on its population density, spatial distribution, and timing of oviposition relative to heading and anthesis (Wright and Doane 1987; Elliott and Mann 1996). In addition to causing yield loss, lesser damage to kernels also results in a lower grade for grain (Dexter et al. 1987; Elliott 1988; Lamb et al. 2000). Currently, S. mosellana infests much of the wheat-growing area of Manitoba, Saskatchewan, North Dakota and Idaho, and is beginning to invade Alberta and Montana (Lamb et al. 1999; Bechinski and Loftus 2000; Olfert et al. 2004; Shanower 2005). Wheat midge also occurs in wheatgrowing areas of Nova Scotia, Ontario, Quebec, and British Columbia (Doane et al. 2002). Starting in 1984, a research team began comprehensive biological, ecological and agronomic studies on this new invasive pest species. That year, S. mosellana populations in Saskatchewan were

D.W. Langor & J. Sweeney (eds.)

found to be parasitized by an egg-larval parasitoid, Macroglenes penetrans (Kirby) (Hymenoptera) (Doane et al. 1985, 1989). The female wasp is 1–2 mm long and lays an egg inside the egg of its host. Despite the presence of the parasitoid, the wheat midge larva completes its development and overwinters in the soil (Doane et al. 2002). The next spring, the parasitoid larva consumes its host, and emerges as an adult in July. As with many non-native agricultural insect pests, the ecological impact of S. mosellana was more associated with efforts to control infestations using chemical insecticides than with impacts that this nonnative insect pest had on native plant species (Wright and Doane 1987) and habitat (Floate et al. 1990). In the case of S. mosellana, insecticidal sprays were applied to 300,000–500,000 ha of wheat annually to minimize crop damage during a major outbreak in the 1990s (WCCP 1997). This paper describes the increasing pest status of a non-native, agricultural insect pest (S. mosellana), and the positive influence that its natural enemy (M. penetrans) and management technologies have had on mitigating the potential economic and environmental impact of wheat midge.

Materials and methods Surveys of S. mosellana and M. penetrans abundance and distribution were conducted annually in Saskatchewan in August 1991–2000. The overwintering larval cocoon is the best stage for comparing populations from one year to the next because it is sessile and is present throughout the autumn and early spring when sampling can be done to identify potentially damaging populations for the coming crop year. Four wheat fields per Rural Municipality were selected at random and sampled by taking 10 soil cores (15 cm deep, 5.06 cm2 in area) in each field. The 10 cores per field were bulked to represent that field. Soil cores were processed by wet sieving as described by Doane et al. (1987). Wheat midge larvae were counted and then dissected to determine whether they were parasitized. Data from approximately 200–300 fields were compiled each year, and summarized to produce a forecast of the damage potential from S. mosellana for the subsequent growing season. The field locations (latitude and longitude) were entered into a database

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

compatible for spatial analyses. Sitodiplosis mosellana and M. penetrans counts (larvae/m2) were used to generate surface maps of abundance with the ‘Potential Mapping’ procedure in SPANS GISÒ for OS/2—PCI Geomatics. This procedure calculates a weighted average of the point data based on the corresponding values of its ‘nearest neighbors’, selects the specified risk category and produces a contoured map with different colours representing five population level categories. The area (km2) of each population level was calculated annually (Table 1). To estimate wheat production within the areas infested with S. mosellana, as calculated by the SPANS GISÒ software, we obtained estimates of: (i) the proportion of total land area that was in crop production in any given year; and (ii) the proportion of total crop production area that was in wheat production (Saskatchewan Agriculture and Food 2007a). Estimates of wheat production (% area) were then applied to calculate the impact of the parasitoid populations in reducing S. mosellana numbers below action threshold levels (Saskatchewan Agriculture and Food 2007b). Areas of reduced S. mosellana populations were then regarded as not requiring an application of insecticide for control of this invasive species. The invasive potential of S. mosellana was estimated by calculating the dispersal of populations from

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northeast Saskatchewan (Melfort) from 1991 to 2000 in a southwest direction to the edge of new infestations, as depicted by the SPANSÒ GIS-produced distribution maps.

Results & discussion Between 1991 and 2000, Saskatchewan experienced one major outbreak of S. mosellana, that lasted from 1994 to 1998 (Fig. 1). The outbreak resulted due to several consecutive years where weather conditions were highly suitable for oviposition during peak adult emergence. Wheat midge populations advanced approximately 360 km spreading from Melfort (52.50° N; 104.37° W) to Swift Current (50.28° N; 107.8° W). Populations advanced in south and westerly directions an average of 36 km/year; the greatest distance occurring in 1993 (90 km) and the least in 1997 (10 km) (Fig. 1). Populations of M. penetrans successfully followed the advance of S. mosellana, never lagging far behind (Fig. 2). For the agricultural areas of the province that were surveyed annually for wheat midge, the estimated area under crop production was 51% of the total land base (Saskatchewan Agriculture and Food 2007a). Non-crop production areas included summer fallow, improved pasture, unimproved pasture and wooded

Table 1 Area (km2) of wheat midge, Sitodiplosis mosellana (Ge´hin), infestation in Saskatchewan, 1991–2000 1991

1992

1993

1994

1995

1996

1997

1998

1999

2000

Total midge larvae (m 2) None \600 600 B 1200 1200 B 1800 [1800

287

2376

22495

22439

25613

20981

17739

5058

11423

7496

67667 18699

107544 12703

89419 29352

81585 46438

138821 44615

143843 38222

86289 54993

103223 52868

143324 70891

202616 48073

1226

5300

5865

18024

10218

10663

26621

37529

21272

7467

540

3131

9271

19135

7135

8456

45011

37828

12494

6966

Viable midge larvae (m2) None

287

4776

22682

22687

28305

21569

18681

5086

11576

11855

\600

81168

115631

113934

125836

162047

164892

117778

133841

198494

243398

600 B 1200

6771

7086

14508

23891

26647

25145

49787

58735

39925

12764

1200 B 1800

125

1429

1977

8558

4395

5944

19623

19528

6408

3866

68

2132

3302

6648

5008

4616

24785

19316

3001

740

20465

21134

44488

83597

61968

57341

126625

128225

104657

62506

6964

10647

19787

39097

36050

35705

94195

97579

49334

17370

13501

10487

24701

44500

25918

21636

32430

30646

55323

45136

[1800 C600 midge larvae (m2) Total midge Viable midge Difference (area)

130

D.W. Langor & J. Sweeney (eds.)

Fig. 1 Larval populations of wheat midge, Sitodiplosis mosellana (Ge´hin), estimated from soil core samples in Saskatchewan from 1991 to 2001

Fig. 2 Area infested with wheat midge, Sitodiplosis mosellana (Ge´hin) (total shaded regions) and area where wheat midge larvae were parasitized with Macroglenes penetrans Kirby (light shaded region) in Saskatchewan from 1992 to 1997

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

areas. The total area that was seeded to wheat in any given year ranged from a high of 8.5 million ha (68.6% of total production) in 1991 to a low of 5.9 million ha (43.9% of total) in 1999 (Saskatchewan Agriculture and Food 2007a). The extent of wheat midge infestations and parasitoid populations in wheat in Saskatchewan is detailed in Table 1. In any given year,[95% of the area infested with wheat midge also had some level of parasitism. Of this total wheat production during the 1990s, there were approximately 15.5 million hectares of wheat in Saskatchewan that did not require a pesticide application because parasitism reduced the density of viable wheat midge larvae to less than 600/m2 (Saskatchewan Agriculture and Food 2007b). At a cost of $16.00/ha (Saskatchewan Agriculture and Food 2007c), the total saving in pesticide costs alone were $248.3 million. The environmental benefits of not having to apply this amount of chemical insecticide are additional. Annually, M. penetrans controlled an average of 20–45% of the wheat midge population in Saskatchewan (Fig. 3). In some individual fields, the rate of parasitism was as high as 80%. In 2000, the average rate of parasitism across Saskatchewan was 45%, which is still below parasitism levels (64%) commonly recorded in Europe for S. mosellana (Carl and Raps 1992). Researchers at Saskatoon Research Centre, collaborating with scientists in Europe, successfully introduced a second parasitoid, Platygaster tuberosula Kieffer (Hymenoptera), to further improve biological control of wheat midge (Olfert et al. 2003). To protect this valuable resource, new pest management tools must continue to preserve parasitoid populations into the future. For example, the recent

% Parasitism

50 40 30 20 10 0 1991 1992 1993 1994 1995 1996 1997 1998 1999 2000

Years

Fig. 3 Annual provincial average rates of parasitism of wheat midge, Sitodiplosis mosellana (Ge´hin), in Saskatchewan in 1991–2000. Solid line indicates overall mean rate of parasitism

131

development of wheat cultivars that are resistant to wheat midge feeding (Lamb et al. 2002a) will negatively impact localized populations of the parasitoid in situations where the host is decimated. Smith et al. (2004) designed an agronomic practice of providing a refuge for virulent S. mosellana which allows producers to grow resistant wheats while at the same time protecting the resistance gene from being overcome by S. mosellana, and providing a host population for preserving parasitoid populations. In relation to chemical control options, Elliott et al. (1991) and Elliott and Mann (1997) developed reduced-risk strategies for insecticides to control S. mosellana while minimizing the impact of insecticides on parasitoid populations. In conclusion, to minimize the economic and ecological impact of S. mosellana, wheat producers in western Canada have access to a comprehensive management program (Elliott et al. 2007). Forecasts and risk warnings, monitoring tools, cultural control, agronomic practices, chemical control, biological control and plant resistance are all available for the industry to manage S. mosellana. Prior to the growing season, forecast maps predict high risk areas (Olfert et al. 2004). If the rotation allows, the producer may choose not to grow wheat, grow a resistant variety of wheat (Lamb et al. 2002a), or grow an alternate resistant crop instead. If a lower degree of infestation is predicted, producers may stick to their plans to grow wheat, but may choose a less susceptible wheat cultivar and plant early to avoid high midge populations during heading (Elliott et al. 2000a). Producers are urged to monitor crops closely in all areas where S. mosellana is present during the susceptible period (emergence of the wheat head from the boot until anthesis begins). Field scouting tools, including visual counts (Elliott and Mann 1996), sticky cards (Lamb et al. 2002b), and pheromone traps (Gries et al. 2000) are readily available for producers to utilize. An insecticide application is recommended when the crop is heading but not yet flowering and wheat midge density is one adult per 4–5 wheat heads. To maintain optimum grade, insecticide should be used when the pest population reaches one adult per 8–10 heads (Elliott et al. 2000b). Late insecticide applications should be avoided as it is not cost effective and may adversely affect biological control agents (Elliott and Mann 1996).

132 Acknowledgements Special thanks are extended to M. Braun and L. Mann for their entomological assistance and to D. Giffen for preparation of the maps. Funding for the surveys was provided by Saskatchewan Crop Insurance.

References Affolter F (1990) Structure and dynamics of the parasitoid complex of the wheat midges Sitodiplosis mosellana (Ge´hin) and Contarinia tritici (Kirby). International Institute for Biological Control, Dele´mont, Switzerland, June 1990 Report, 108 pp Allen WR (1955) The wheat midge, Sitodiplosis mosellana (Ge´hin). In: Annual conference of Manitoba agronomists 1955, pp 28–29 Barnes HF (1956) Gall midges of economic importance. VII. Gall midges of cereal crops. London, 261 pp Basedow T, Schu¨tte F (1973) Neue Untersuchungen u¨ber Eiablage, wirtschaftliche Schadensschwelle und Beka¨mpfung der Weizengallmu¨cken (Dipt.: Cecidomyidae). Zeits fu¨r Angew Entomol 73:238–251 Bechinski E, Loftus B (2000) UI experts monitor orange midge, a potential threat to Idaho wheat. Resources for Idaho— News. University of Idaho, Moscow, ID, May 26, 2000 Carl K, Raps A (1992) Wheat midge (Sitodiplosis mosellana). Annual report to agriculture Canada, International Institute of Biological Control, De´lemont, Switzerland, pp 15–19 Dexter JE, Preston KR, Cooke LA, Morgan BC, Kruger JE, Kilborn RH, Elliott RH (1987) The influence of orange wheat blossom midge, Sitodiplosis mosellana (Ge´hin), damage on hard red spring wheat quality and the effectiveness of insecticide treatments. Can J Plant Sci 67:697–712 Doane JF, Mukerji MK, Olfert O (1985) Biological research on the wheat midge. Research Highlights—1984, Saskatoon Research Station, pp 51–53 Doane JF, Olfert O, Mukerji MK (1987) Extraction precision of sieving and brine flotation for removal of wheat midge, Sitodiplosis mosellana (Diptera: Cecidomyiidae), from soil. J Econ Entomol 80:268–271 Doane JF, DeClerck-Floate R, Arthur AP, Affolter F (1989) Description of the life stages of Macroglenes penetrans (Kirby) (Hymenoptera: Chalcidoidea, Pteroma´lidae), a parasitoid of the wheat midge, Sitodiplosis mosellana (Ge´hin) (Diptera: Cecidomyiidae). Can Entomol 121: 1041–1048 Doane JF, Braun MP, Olfert O, Affolter K, Carl K (2002) Sitodiplosis mosellana (Ge´hin), orange wheat blossom midge (Diptera: Cecidomyiidae). In: Mason PG, Huber JH (eds) Biological control programmes in Canada 1981– 2000. Oxford University Press, pp 246–249 Elliott RH (1988) Factors influencing the efficacy and economic returns of aerial sprays against the wheat midge, Sitodiplosis mosellana (Ge´hin) (Diptera: Cecidomyiidae). Can Entomol 120:941–954 Elliott RH, Mann LM (1996) Susceptibility of red spring wheat, Triticum aestivum L. cv. Katepwa, during heading and anthesis to damage by wheat midge, Sitidiplosis mosellana (Ge´hin) (Diptera: Cecidomyiidae). Can Entomol 128:367–375

D.W. Langor & J. Sweeney (eds.) Elliott RH, Mann LM (1997) Control of wheat midge, Sitidiplosis mosellana (Ge´hin) at lower chemical rates with small-capacity sprayer nozzles. Crop Prot 16:235–242 Elliott RH, Mann LM, Spurr DT, Sacher ER (1991) Distribution of dye deposits from a small-plot track sprayer on biological and artificial targets. Crop Prot 10:129–135 Elliott RH, Mann L, Olfert O (2000a) Susceptibility of hard red spring wheats to damage by high populations of wheat midge. SRC-Saskatoon Research Letter #2000–09, 3 pp Elliott RH, Mann L, Olfert O (2000b) Susceptibility of durum and CPS wheats to damage by high populations of wheat midge. SRC-Saskatoon Research Letter #2000–06, 3 pp Elliott RH, Olfert O, Lamb RJ (2007) Wheat midge. In: Buntin GD, Pike KS, Weiss MJ, Webster JA (eds) Handbook of small grain insects. Entomological Society of America, pp 73–75 Felt EP (1912) Observations on the identity of the wheat midge. J Econ Entomol 5:286–289 Fletcher J (1902) Experimental farms reports for 1901, No. 16. Government of Canada, Ottawa, ON, Canada, p 212 Floate KD, Doane JF, Gillott C (1990) Carabid predators of the wheat midge, Sitodiplosis mosellana (Diptera: Cecidomyiidae), in Saskatchewan. Environ Entomol 19: 1503–1511 Gries R, Gries G, Khaskin G, King S, Olfert O, Kaminski L, Lamb R, Bennett R (2000) Sex pheromone of orange wheat blossom midge, Sitodiplosis mosellana (Ge´hin). Naturwiss 87:450–454 Lamb RJ, Wise IL, Olfert O, Gavloski J, Barker PS (1999) Distribution and seasonal abundance of Sitodiplosis mosellana (Diptera: Cecidomyiidae) in spring wheat. Can Entomol 131:387–397 Lamb RJ, Tucker JR, Wise IL, Smith MAH (2000) Trophic interaction between Sitodiplosis mosellana (Diptera: Cecidomyiidae) and spring wheat: Implication for yield and seed quality. Can Entomol 132:607–625 Lamb RJ, Wise IL, Smith MA, Mackenzie RIH, Thomas J, Olfert O (2002a) Oviposition deterrence against Sitodiplosis mosellana (Diptera: Cecidomyiidae) in spring wheat (Gramineae). Can Entomol 134:85–96 Lamb RJ, Wise IL, Gavloski J, Kaminski L, Olfert O (2002b) Making control decisions for Sitodiplosis mosellana in wheat using sticky traps. Can Entomol 134:851–854 Mukerji MK, Olfert O, Doane JF (1988) Development of sampling designs for egg and larval populations of the wheat midge, Sitodiplosis mosellana (Ge´hin) (Diptera: Cecidomyiidae), in wheat. Can Entomol 120:497–505 Olfert O, Mukerji MK, Doane JF (1985) Relationship between infestation levels and yield loss caused by wheat midge, Sitodiplosis mosellana (Ge´hin) (Diptera: Cecidomyiidae), in spring wheat in Saskatchewan. Can Entomol 117: 593–598 Olfert O, Doane JF, Braun MP (2003) Establishment of Platygaster tuberosula, an introduced parasitoid of the wheat midge Sitodiplosis mosellana, in Saskatchewan. Can Entomol 135:303–308 Olfert O, Elliott RH, Hartley S, Zeleny K (2004) Forecast of wheat midge in Alberta and Saskatchewan for 2004. 2003 Crop variety highlights and insect pest forecasts. Saskatoon Research Centre Technical Bulletin No. 2004–01, pp 14–15

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems Pivnick KA, Labbe´ E (1993) Daily patterns of activity of females of the orange wheat blossom midge, Sitodiplosis mosellana (Ge´hin) (Diptera: Cecidomyiidae). Can Entomol 125:725–736 Reeher MM (1945) The wheat midge in the Pacific Northwest. U.S. Department of Agriculture Circular # 732, 8 pp Saskatchewan Agriculture and Food (2007a) Agriculture statistics: grains and oilseeds. Website: http://www.agr.gov. sk.ca/apps/agriculture_statistics/HBV5_Crop1.asp. Cited 15 March 2007 Saskatchewan Agriculture and Food (2007b) Production— insects—wheat midge. Website: http://www.agr.gov.sk. ca/docs/production/midge0106.asp. Cited 15 March, 2007 Saskatchewan Agriculture and Food (2007c) 2007 Guide to crop protection. Saskatchewan Agriculture and Food, Regina, SK, pp 340

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Shanower TG (2005) Occurrence of Sitodiplosis mosellana (Diptera: Cecidomyiidae) and its parasitoid, Macroglenes penetrans (Hymenoptera: Platygasteridae), in northeastern Montana. Can Entomol 137:753–755 Smith MAH, Lamb RJ, Wise IL, Olfert O (2004) An interspersed refuge for Sitodiplosis mosellana (Diptera: Cecidomyiidae) to protect crop resistance and biocontrol in wheat. Bull Entomol Res 94:179–188 WCCP (1997) Western committee on crop pests—minutes of 36th annual meeting, Edmonton, AB. October 9–10, 1997, 56 pp Wright AT, Doane JF (1987) Wheat midge infestation of spring cereals in northeastern Saskatchewan. Can J Plant Sci 67:117–120

Climate suitability and management of the gypsy moth invasion into Canada Jacques Re´gnie`re Æ Vince Nealis Æ Kevin Porter

Originally published in the Journal Biological Invasions, Volume 11, No 1, 135–148. DOI: 10.1007/s10530-008-9325-z Ó Her Majesty the Queen in Right of Canada 2008

Abstract The gypsy moth has become established throughout southern Canada east of Lake Superior where the climate is suitable for the completion of its univoltine life cycle. The spread of the gypsy moth to the north and west in Canada has so far been prevented by climatic barriers and host plant availability as well as by aggressive eradication of incipient populations. Climate change is expected to increase the area of climatic suitability and result in greater overlap with susceptible forest types throughout Canada, especially in the west. At the same time, the gypsy moth is spreading west in the USA into states bordering western Canadian provinces. These circumstances all lead to a greatly increased risk of further invasion into Canadian forests by the gypsy moth. Management actions need to be intensified in different ways in different parts of the country to reduce the impacts of spread in eastern Canada and to prevent the gypsy moth from invading western regions. J. Re´gnie`re (&) Natural Resources Canada, Canadian Forest Service, PO Box 10380, Stn. Ste. Foy, Quebec, QC, Canada G1V 4C7 e-mail: [email protected] V. Nealis Natural Resources Canada, Canadian Forest Service, 506 West Burnside Rd, Victoria, BC, Canada V8Z 1M5 K. Porter Natural Resources Canada, Canadian Forest Service, PO Box 4000, Fredericton, NB, Canada E3B 5P7

Keywords Invasive alien species
Lymantria dispar
Risk assessment

Introduction The gypsy moth, Lymantria dispar L. (Lepidoptera: Lymantriidae), is a notorious insect defoliator of trees. It is native to the temperate forests of Europe and Asia where occasional outbreaks occur. A European strain of the gypsy moth was introduced accidentally to North America near Boston, USA, in 1869 and has subsequently invaded much of the susceptible forest of northeastern North America from the Atlantic Coast westward to the Great Lakes Basin, northward into the Canadian Maritime Provinces, Quebec and Ontario (Liebhold et al. 1992) and as far south in the USA as North Carolina (Hastings et al. 2002, Fig. 1). In the USA, expansion of the insect’s range continues southward along the Appalachian mountains, threatening to reach the hardwood forests of the southeast down to northern Florida (Allen et al. 1993) and westward through the transition ecozones between the eastern forests and the Great Plains. This invasion has been countered by over a century of aggressive and often controversial pest control (Nealis and Erb 1993) which has, in some cases, slowed but not halted the rate of spread (Sharov et al. 2002). There have been repeated introductions of gypsy moth to western North

D.W. Langor & J. Sweeney (eds.), Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems. DOI: 10.1007/978-1-4020-9680-8_10

135

136

D.W. Langor & J. Sweeney (eds.)

Fig. 1 Regulated area for the gypsy moth in the USA (2006, pale grey area), and Canada in 1974 (dark hatched area), 1989 (medium hatched area) and 2006 (pale hatched area)

America since the 1970s but to date (2007) these have been eradicated successfully or have otherwise failed to result in the establishment of viable populations (Nealis 2002; Weseloh 2003; Logan et al. 2007). Adult females of the gypsy moth in North America are flightless so natural dispersal is limited to relatively short-distance ballooning of newly-hatched larvae on silk threads (Fosberg and Peterson 1986; Weseloh 1997). The rapid invasion into North America, and in particular the periodic infestations in western regions, has been aided by the insect’s surreptitious association with humans. Many of the insect’s numerous preferred host plants flourish in disturbed habitats and female moths tend to oviposit in protected shelters at ground level near the host trees (Lyons and Liebhold 1992). This often results in egg masses on outdoor articles such as lawn furniture, firewood and the underside of vehicles. When these articles are moved, the eggs are moved as well. The gypsy moth was intercepted several times in Canada near the USA border during the first half of the 20th century as populations in the USA spread north and west. Eradication programs were carried out in southern Quebec in the 1920s and in southwestern New Brunswick in the 1930s. Significant increases in gypsy moth infestations in the USA in the mid-1950s

(Doane and McManus 1981) coincided with repeated infestations in southern Quebec and Ontario leading to spray programs that were conducted every year beginning in 1960 (Brown 1975). By the late 1970s, defoliation was observed over a large area of southwestern Quebec and a few years later in southeastern Ontario (Jobin 1995), again following an increase in the intensity and area of a gypsy moth outbreak in the northeastern United States. By this time, authorities in Quebec and Ontario conceded that eradication was no longer feasible and pest management objectives shifted from eradication to suppression (Nealis and Erb 1993). In spite of these control efforts, the gypsy moth has continued to expand its range in eastern North America. Thus, the source area for new introductions of the gypsy moth has increased steadily for nearly 135 years. The increasing frequency of new introductions to western Canada since 1990 reflects both this increase in the source area and the high risk of successful transport via the pathway of westward movement of people and goods in North America (Nealis 2002). The current policy of the Canadian federal government for management of the gypsy moth in Canada distinguishes two different situations: areas where the insect is considered established and therefore regulated by the federal Canadian Food Inspection Agency

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

(CFIA), and areas located outside this regulated zone. Within the regulated zone, the CFIA focuses on the management of pathways through inspection of specific commodities as defined by multilateral phytosanitary agreements. Once an area is regulated, response to outbreaks within the zone becomes the purview of provincial and municipal governments or private landowners and is determined essentially by their sensitivity to damage caused by outbreaks. The goal here is suppression and mitigation of damage. Outside of the regulated zone, the CFIA maintains a network of pheromone traps to update the boundaries of the regulated zone and to detect new isolated introductions far removed from the regulated area. The CFIA does not necessarily attempt to eradicate these isolated populations. The intent of the monitoring is to reveal where the gypsy moth has become established and, if necessary, regulate these areas. Jurisdictions wanting to avoid the potential economic penalty of regulation may carry out their own eradication programs to comply with the North American Plant Protection Organization’s (NAPPO) definition of a pest-free state as ‘‘an officially identified area in which a target pest does not occur and is maintained as such’’. This has been, for example, the situation in British Columbia on Canada’s West Coast since 1999 when the provincial government assumed responsibility for the eradication of the gypsy moth. The maintenance of an annual network of traps to detect gypsy moth throughout the vast, unregulated area of Canada and to update continuously the boundaries along the several thousand km of the regulated zone is a formidable undertaking. Furthermore, regulation may be an effective instrument to facilitate the movement of goods and reduce pathways between countries but it seems to have had little practical influence on the invasion process within Canada. A re-evaluation of the strategies and policies for managing the invasion of Canada by the gypsy moth is needed. In this paper, we apply our knowledge of Canada’s climatic suitability for the gypsy moth to the design of a strategy for management of invasion. Our argument is guided by the following premises: (1)

The gypsy moth requires a suitable place to live, including an adequate supply of host plants and a climate suitable for the completion of its life cycle over several generations.

(2)

(3)

(4)

(5)

137

Process modeling of temperature-dependent life history events can be used to map climatically suitable areas for gypsy moth (Re´gnie`re and Nealis 2002; Gray 2004) and to examine the influence of climate on historical patterns of invasion. Climate is changing and the distribution of areas climatically favourable to the gypsy moth can be expected to change accordingly (Logan et al. 2003; Pitt et al. 2007). Process-modeling is well suited to examining the consequences of this change on the likelihood of gypsy moth populations persisting in a particular area and the associated change in risk (Logan et al. 2007). Natural spread of gypsy moth along the broad contiguous frontier between regulated and unregulated areas cannot be stopped but can it be slowed, albeit at considerable effort and cost (Mayo et al. 2003). Benefits can be weighed against costs for the particular areas of concern (Sharov and Liebhold 1998a, b); Sharov et al. 1998). Eradication of isolated populations resulting from long-distance transport by humans is demonstrably feasible and cost-effective (Nealis 2002; Mayo et al. 2003).

We first evaluate the historical invasion and current range of the gypsy moth in Canada in the context of climate suitability. Then, we use a plausible, conservative 1% per year CO2 increase climate-change scenario (Price et al. 2001) to drive a model of gypsy moth seasonality and to predict the potential future distribution of the gypsy moth. We discuss the relationship between these future potential distributions and the known ecology of the gypsy moth and propose management options that could be implemented in Canada.

Methods Historical gypsy moth data Regulated area The Canadian Food Inspection Agency (CFIA) monitors exotic pests such as the gypsy moth in Canada, defines the zone where these organisms are

138

established, and regulates the movement of potentially infested materials to un-infested regions both within and outside the country. To identify and demarcate areas requiring regulation, the CFIA deploys an extensive network of pheromone traps in unregulated areas. Thus the CFIA regulatory maps, updated annually, reflect accurately the insect’s current range. We obtained maps of the CFIAregulated area for the gypsy moth in Canada as of 1974, 1989, and the most recent update, 2006. Until 1995, the Canadian Forest Service’s Forest Insect and Disease Survey (FIDS) also deployed traps and reported annually on the status of populations. Data included evidence of tree damage, population density estimates and pheromone trap catch records and are included in a geo-referenced historical database called the FIDS INFOBASE (Power 1986). The database includes both positive and negative pheromone trap capture records and therefore provides a very good record of the actual distribution of finds. Verification of doubtful records and supplementary information was obtained by searching a digitized library of annual reports produced by the FIDS for the same period and assigning a geo-reference on the basis of maps and descriptive narrative. Monitoring inside the regulated zone is less systematic and carried out by provincial and municipal governments. Quebec began gypsy moth surveys in 1986 and the Maritime Provinces in 1993. Monitoring of the gypsy moth populations in Ontario remained a joint venture of the Canadian Forest Service and the Ontario Ministry of Natural Resources until 2003. We obtained geo-referenced records of gypsy moth occurrence from New Brunswick, Nova Scotia and Quebec up to 2006 and from Ontario up to 2003 to complete the history started in the FIDS INFOBASE database described earlier. Information on the area of moderate to severe defoliation by the gypsy moth from 1975–1980 in Ontario and 1975–1990 in Quebec was obtained from Jobin (1995). For comparison, the total area under moderate to severe defoliation in the American states bordering eastern Canada was obtained for the period 1960–2006 from the Gypsy Moth Digest Web site (http://www.na.fs.fed.us/fhp/gm). The gypsy moth is regularly detected in pheromone traps in western Canada (from Manitoba to British Columbia). Incipient populations have been

D.W. Langor & J. Sweeney (eds.)

successfully eradicated on several occasions over the past 20 years in British Columbia (Nealis 2002). Data from the CFIA, the CFS and the various provincial reports have been used to compile the history of the gypsy moth in western Canada. Climate suitability maps We used a detailed model of gypsy moth phenology (Gray et al. 1991, 1995, 2001; Logan et al. 1991; Sheehan 1992; Re´gnie`re and Sharov 1998; Re´gnie`re and Nealis 2002), hereafter the Gypsy Moth Life Stage (GMLS) model, to estimate the probability distribution of climatic suitability to support persistent gypsy moth populations throughout Canada under past, current and future climate conditions. Our method is similar to Logan et al. (2007) and Pitt et al. (2007). The approach consisted of determining whether the insect can successfully complete its life cycle under a specific climate. This was done in two steps. First, the GMLS model was run for 25,000 locations over the area of interest (Canada south of 608N), using daily temperature time series generated from monthly normals as input (Re´gnie`re and St-Amant 2007). The process was replicated 50 times. Each run returned an establishment likelihood value F = 0 (non-viable seasonality) or F = 1 (viable). Because each time the model was run the weather data provided are slightly different (daily stochastic variation being restored to the monthly normals), the probability (p) of establishment in a given location was calculated as the proportion of replicates where the model returned F = 1 (successful completion of the life cycle). Second, an establishment probability map was generated by universal kriging with elevation as an external drift variable (Isaaks and Srivastava 1989), using a Digital Elevation Model of Canada at 30 arc-second resolution assembled from http://edc.usgs.gov/. This entire process was automated by the BioSIM system (Re´gnie`re 1996). The climate change scenario was generated by the Canadian Global Circulation Model version 1 (CGCM-1), driven by a 1% CO2 increase per year, over the time frame 2000–2070. The output of the CGCM-1 is a series of daily values from 1931 to 2070 over a low-resolution grid covering the planet. From these, deviations of monthly average minimum and maximum temperature from the standard

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

1961–1990 reference period were calculated for each year and grid point in Canada. The spatial resolution of these monthly deviations from the reference period, also called monthly anomalies, was increased through a spatial interpolation technique taking topography into account (Price et al. 2001). Daily observations of minimum and maximum temperature from all Environment Canada weather stations for years 1961 to 1990 were then modified by adding, by year and month, the monthly anomalies from the nearest high-resolution grid cell. For example, monthly anomalies for year 2041 and daily temperature observations from 1961 were used to produce climate-changed daily observations for year 2041. The same was done to 1962 daily observations using 2042 monthly anomalies, and so on. From these climate-changed daily values, normals were calculated in 10-year increments from 1981 (1981–2010 normals) to 2041 (2041–2070 normals). Actual normals were calculated from Environment Canada daily records for each decade from 1931 (1931–1960 normals) to 1971 (1971–2000 normals). Annual climate fluctuation and establishment probability in Ontario An analysis of the annual fluctuations in the geographical distribution of viable seasonality in Ontario, Canada was carried out using 250 simulation points located randomly across the province. The GMLS model was run for each point for each year between 1981 and 2006, using as input observed daily minimum and maximum air temperature records from Environment Canada weather stations. Temperature data were interpolated from the four recording stations nearest to each simulation point using the GIDS method (Nalder and Wein 1998). The relationship between the establishment flag F output by the GMLS model (described above) and latitude (L) was determined by binary logistic regression using the model: gðpÞ ¼ a þ bL þ e

ð1Þ

where g(p) is the logit link function, p is the probability of establishment, e is a binomiallydistributed error term, and a, b are regression parameters. By rearranging Eq. 1, the latitude L0.5 at which the probability of establishment p = 0.5, or g(p) = 0, can be calculated:

L0:5 ¼ 

a b

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ð2Þ

The probability of establishment p (proportion of simulation outputs with an establishment flag value of F = 1) was calculated by year. The annual fluctuations and time trend of L0.5 from 1981–2006 as determined by Eq. 2 were also examined. Forest resources and area at risk Consideration of area at risk in Canada depends on the location and abundance of susceptible host species. The gypsy moth is a polyphagous species with a large number of deciduous and coniferous host tree species (Liebhold et al. 1995). Spatial forest inventories from provincial governments and forest companies contain much of the required data but often do not adequately account for the many noncommercial host tree species of gypsy moth and would require a monumental effort to assemble due to differing protocols, varying formats and restricted access. Maps of the distribution of gypsy mothsusceptible forests in Canada, grouped by image analysis into three broad forest types were obtained from Beaubien et al. (2002): deciduous, closed (dense) and open (thin) mixed stands. The area of overlap between these forest types and the zone of high establishment probability (p [ 0.5) of the gypsy moth, by decade under past and future climate conditions were calculated by GIS analysis.

Results The historical record of recoveries of gypsy moth adults in pheromone traps and of other life stages in various samples from provincial surveys in Canada is illustrated in Fig. 2, and the area defoliated in Canada and the bordering USA is shown in Fig. 3. Until the late 1960s, the insect was confined to southwestern Quebec and the area around Kingston, Ontario (Fig. 2a), having crossed into Canada from adjacent New York State. In the early 1970s, the insect was also recovered from several locations in New Brunswick (Fig. 2b). Between 1975 and 1980, pheromone trap records indicated a significant increase in the gypsy moth’s distribution in Quebec and New Brunswick, as well as new finds in Nova Scotia and Prince Edward Island (Fig. 2c). In the late 1970s,

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Fig. 2 Historical distribution of gypsy moth recoveries from pheromone traps (blue d) and from sampling of other life stages (yellow j) in Canada between 1964 and 2006. Background: probability of establishment based on climate

suitability for an adaptive life cycle using climate normals for 1971–2000. (a) 1964–1970; (b) 1971–1975; (c) 1976–1980; (d) 1981–1985; (e) 1986–1990; (f) 1991–1995; (g) 1996–2000; (h) 2001–2006

defoliation by the gypsy moth in Quebec exceeded 5,000 km2 (Fig. 3a). By the early 1980s, the insect had become widely distributed in southeastern Ontario (Fig. 2d) with nearly 3,000 km2 of forest defoliated in 1985 (Fig. 3a). It seems likely that the rapid invasion of Ontario by the gypsy moth was driven by spread west from outbreak populations in Quebec into eastern Ontario as well as by northward expansion of the very large outbreak of 1980–1982 in the northeastern USA (Fig. 3b). In 1989, the CFIA determined that the area generally infested with the gypsy moth in eastern Canada covered 151,000 km2 between southeastern Ontario and Nova Scotia (Fig. 1). By 1990, with a second defoliation episode covering 4,000 km2 in Ontario (Fig. 3a), the insect was already occupying most of the area of medium to high climatic suitability in eastern Canada (Fig. 2e). There has

been little change in its geographical distribution since then (Fig. 2f, g, h). The CFIA-regulated area in 2006 covered 326,000 km2 from the eastern tip of Lake Superior to Nova Scotia (Fig. 1). Adult gypsy moths have been recovered regularly in pheromone traps from northwestern Ontario to British Columbia since the late 1980s, but so far there is no evidence of established populations. This situation is partly the result of numerous successful eradication campaigns as well as the fact that areas of climate suitability in the west are fragmented (Fig. 2), as is the distribution of host plants (see later). The change over time in the range of coordinates at which moths and other life stages of the gypsy moth have been recovered in pheromone traps or other samples suggests a rapid invasion of the climatically suitable parts of Canada followed by a

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

Fig. 3 Area (km2) with moderate to severe defoliation by the gypsy moth in (a) Ontario (s) and Quebec (d) and (b) the bordering areas infested in the United States (eastern: from Maine to New York, d; western: from Ohio to Wisconsin, s) between 1960 and 2006

period of relative stasis (Fig. 4). In Quebec, the spread northward and from east to west occurred very rapidly prior to 1980 at the time of the widespread Fig. 4 North-south (left column) and east-west (right column) range (km) of gypsy moth recovered in pheromone traps by FIDS (s) and various other life stages (d) found over time. (a, b) Quebec. (c, d) Maritime Provinces. (e, f) Ontario

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defoliation episode depicted in Fig. 3a, and has since stopped (Fig. 4a, b). Nor has there been a resurgence of defoliation in the Maritime Provinces of New Brunswick and Nova Scotia. This spatio-temporal pattern of invasion has an interesting additional feature: there is a marked reduction in both the north-south and east-west ranges from a maximum in the early 1980s (Fig. 4c, d) such that the range of the gypsy moth has become ‘‘compressed’’ over time. This compression was most dramatic in Ontario, where the gypsy moth spread rapidly between 1980 and 1990 to reach a maximum north-south and eastwest range between 1000 km and 1600 km, respectively. A significant compression of the gypsy moth’s range then occurred in the early to mid-1990s (Fig. 4e, f). This apparent compression pattern could, for example, be an artifact if sampling effort was reduced in the outlying areas where risk was considered low. However, in both the Maritimes and Ontario, pheromone trapping intensity actually increased during the 1990s as a result of the sharp increase in positive finds in the late 1980s. There is better evidence that the distribution of the gypsy moth was indeed very broad following the initial outbreak phase as potential migrants were abundant (1977–1979 in Quebec and New Brunswick, and 1984–1992 in Ontario), and the

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Fig. 5 Annual fluctuations of P, the probability of gypsy moth establishment (d), and of the latitude L0.5 at which P = 0.5 (s) as determined by Eqs. 1 and 2 (straight dotted line: regression of L0.5 over time)

risk of long-distance transport, even to inhospitable areas, increased. Positive traps in northern Ontario, for example, are all associated with campgrounds. As outbreaks in favorable areas waned, the abundance of migrants decreased, and the apparent range of the gypsy moth as determined from pheromone trap catches compressed as depicted in Fig. 4. There is also evidence of a strong association between these historical patterns of the gypsy moth distribution and historical patterns of climatic suitability in Ontario, as defined by the probability of completing its life cycle. First, there has been a steady increase in the area of climatic suitability for the gypsy moth across Ontario since 1980 (regression of L0.5 against time, F1,24 = 4.82, P = 0.038; Fig. 5). This trend would have been more pronounced had it not been for a sharp decrease in suitability for five of the six years between 1992 and 1997. These years also correspond to a pronounced reduction in the area of defoliation in Ontario (710 km2 compared to the 8,491 km2 defoliated in the previous 10 years) and before the resurgence of defoliation (2,873 km2) since 1998 (Fig. 6). The temporary decline in climatic suitability across Ontario between 1992 and 1997 (Fig. 5), led to a reduction in population density and therefore in potential migrants in the formerly favorable climatic zones, and reduced even further the likelihood of persistence in marginal climatic zones. Recovery of the trend in increasing climatic suitability since 1998 (Fig. 5) has been correlated with resurgence in defoliation (Fig. 6c) and increased frequency of moths in pheromone traps to the north and west in Ontario (Fig. 2). Reductions in population densities within the established range of the gypsy moth may also have involved the belated arrival of natural enemies. The

Fig. 6 Geographical distribution of moderate to severe defoliation by the gypsy moth in Ontario. Background: probability of establishment based on climate suitability for an adaptive life cycle using climate normals for 1971–2000. (a) 1981– 1991; (b) 1992–1997; (c) 1998–2003

fungal pathogen Entomophaga maimaiga Humber, Shimazu and Soper spread across the North American range of the gypsy moth, including Ontario, during

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Fig. 7 Potential distribution of the gypsy moth in Canada based on its probability of establishment as a function of climate. Prepared by applying the output of the Canadian

Global Circulation Model (version I) under a 1%/year increase in atmospheric CO2. Normals updated each decade, from (a) 1971–2000 to (h) 2041–2070

the first half of the 1990s (Nealis et al. 1999; Villedieu and van Frankenhuyzen 2004). Natural enemies may dampen the amplitude of outbreaks in favorable areas and therefore reduce the risk of migrants but will have no effect on the likelihood of establishment in new areas. The northern and western expansion of the gypsy moth in Ontario in the 1990s is reflected in the spatial distribution of defoliation in that province. From 1970 to 1991, most of the defoliation occurred in the southeastern portion of the province near the area of original introduction (Fig. 6a). Since that time, defoliation has been farther north and west (Fig. 6b, c). These sequences of defoliation correspond to the familiar pattern of severe defoliation on the ‘‘leading edge’’ of the invasion process (Elkinton and Liebhold 1990; Liebhold and Tobin 2006). This historical information shows that the gypsy moth has spread rapidly throughout the area in

eastern Canada where the climate is suitable for its establishment and that the most severe defoliation occurred in a wave that followed shortly after the insect first invaded and became established in Quebec and then in Ontario. In Quebec, the gypsy moth has not caused damage of concern since the late 1980s but in Ontario, defoliation has occurred in a series of waves (Fig. 3a). Climate change is expected to increase the future area of suitable climate for gypsy moth establishment throughout Canada as increasingly warm summers allow the insect to complete its univoltine life consistently in areas that are now marginal for an adaptive life cycle (Fig. 7). The change in suitable area for the gypsy moth over the next 50 years will be greatest in western Canada, especially the Prairie Provinces. In British Columbia, the steep topography and the cold climate of the mountains and central plateau restrict the total area of climatic susceptibility

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Fig. 8 Overlap between high climatic suitability for the gypsy moth (pale shading: probability of establishment [0.5 calculated from climatechanged normals 2041– 2070) and susceptible forest types in Canada (red: deciduous; blue: closed mixed; green: open mixed). (a) Western Canada (note the patchy distribution of high climatic suitability in BC). (b) Eastern Canada

to gypsy moth invasion over the period considered here. The overlap between the area of climatic suitability for the gypsy moth (where the probability of establishment[0.5) and the area of distribution of the most susceptible forest types in Canada (deciduous, closed) represents the potential ecological and economic risks posed by invasion of this species. This risk (Fig. 8), depicted here as a surface area of forested landscape, has been relatively low historically, mostly because the portion of eastern Canada that was climatically suitable to the insect was also the least forested as the result of human occupation (e.g. southwestern Ontario in Fig. 8). However, this area at risk is expected to double or triple over the next 50 years as areas with susceptible forest species become more climatically suitable for the gypsy moth as a result of climate change (Fig. 9). Much of this increased risk will occur in eastern Canada, as a result of the vast extent of susceptible forest types in that part of the country that have so far been protected from the gypsy moth by unfavorable climate (Fig. 8). In western Canada, a large proportion of the climatically suitable area will remain dominated either by prairie vegetation or coniferous forests (spruce, pine) that are not expected to support significant gypsy moth populations or suffer much

Fig. 9 Change in area of overlap between high climatic suitability for the gypsy moth (probability of establishment [0.5) and susceptible forest types in Canada, in response to climate change. Dark tone: deciduous forest type (most susceptible). Medium tone: closed mixed forests; Light tone: open mixed forests (least susceptible)

damage. Nonetheless, the trees that will constitute the future hosts for the gypsy moth in an infested prairie region are trees in shelter belts, river valleys and urban settings. These trees are highly valued and they occupy areas associated with humans and are therefore at a greater risk of receiving the gypsy moth. At the most distant reaches of our projections (after 2010 in Fig. 7), the area of climatic suitability for the gypsy moth in the Prairies will overlap with the

Ecological Impacts of Non-Native Invertebrates and Fungi on Terrestrial Ecosystems

southern range of several suitable hosts in the birch (Betula) and poplar (Populus) genera. Although sustained outbreaks of the gypsy moth in forest stands dominated by these species may be, at present, uncommon, the congruence of suitable climate and suitable hosts will surely result in a native forest sustaining an alien insect species with proven potential for damage.

Discussion The gypsy moth repeatedly invaded eastern Canada from adjacent infested areas of the northeastern USA. Once established in the upper St. Lawrence River region in the 1960s, the gypsy moth rapidly spread through suitable habitats in Quebec, Ontario and the Maritime Provinces until its distribution in eastern Canada stabilized in the mid-1990s. Our seasonalitybased modeling indicates that this apparent stability of range of the past 10 years in eastern Canada corresponds to the area where the climate most suitable for completion of the gypsy moth life cycle overlaps with susceptible forest types. To the north and west of the current established range in Canada (i.e. eastern shore of Lake Superior near the city of Sault Ste. Marie, ON), the natural spread of gypsy moth has been limited by a combination of unsuitable climate and lack of susceptible forest types. The relative contribution of these two factors in limiting the insect’s northern spread has been controversial (Sullivan and Wallace 1972; Williams and Liebhold 1995; Sharov et al. 1999). The GMLS model used here predicts closely the current distribution of the insect in eastern Canada on the basis of climate suitability alone indicating climate plays the dominant role. Aggressive management of the gypsy moth associated with the Slow the Spread program in Wisconsin and Minnesota (Reardon et al. 1998; Sharov et al. 2002; Thorpe et al. 2006) contributes to the slow rate of spread in adjacent northwestern Ontario. Eradication of spot infestations in Pacific coastal regions in both the USA and Canada have kept that region free of established gypsy moth populations despite the presence of areas with both suitable climate and host species (Nealis 2002). Where the area of greatest climate suitability (i.e. probability of establishment [0.5) overlaps with areas of high host abundance in Quebec and Ontario,

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extensive and severe defoliation has occurred. A distinct warming trend since the mid-1980s in Canada is expanding this zone of climatic suitability both in eastern and western Canada. There is evidence that gypsy moth populations are invading these new areas further north and causing severe defoliation (Fig. 7). We now consider the consequences of these patterns in terms of future risk and possible management options. Risk to forest resources from the gypsy moth in Canada is a function of two factors: (1) the risk of introduction and (2) the overlap between the area of climate suitability and the geographical distribution of host trees. Areas already infested by the gypsy moth meet both of these conditions and will remain sources of future invasions. As zones immediately adjacent to these infested areas become more suitable because of climate change, invasion will be rapid if suitable hosts are present. The impact will be dependent upon the actual nature of the host plant distribution in these areas, with closed forests dominated by preferred hosts experiencing the greatest impacts. Given the broad host range of the gypsy moth, including abundant northern hardwoods such as birches, aspens and a variety of ornamental plants, defoliation is likely but the ecological and economic impacts are less predictable. On the one hand, the historical pattern in Ontario and elsewhere suggests that leading-edge populations can reach damaging levels quickly. On the other hand, the apparent rapid movement of significant mortality agents, especially the pathogens Nuclear Polyhedrosis Virus (NPV) and Entomophaga maimaiga, that accompany the invasion of the gypsy moth (Nealis et al. 1999; Hastings et al. 2002; Villedieu and van Frankenhuyzen 2004; Dwyer et al. 2004) could dampen those damaging effects. Regardless of the dynamic behaviour in this expanding distribution, the source area for gypsy moth invasions further west will continue to increase in eastern Canada and in the USA (Williams and Liebhold 1995) by a factor of two to three times in the next 50 years as will the area of suitable climate and susceptible host plants in the potential receiving areas further west. With the ever-increasing movement of people and commodities, movement of gypsy moth life stages across otherwise inhospitable habitats into susceptible areas is inevitable. These

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circumstances have several important implications for the management of the forest resources at risk. Our analysis suggests that distinct but complementary management strategies are required in different parts of Canada. First, the potential for damaging outbreaks will persist for most of the deciduous forest east of the Great Lakes. Climate change will result in those areas currently at the margin of the range to become more suitable and the gypsy moth can be expected to invade quickly in spite of regulatory efforts. Accordingly, management priorities in eastern Canada should address resource protection. Areas at greatest risk of damaging defoliation will be those for which models such as the one used here show that climate is most suitable and for which inventories show that host plants are most abundant. Monitoring gypsy moth populations in these areas will permit resource managers to identify regions of potential damage and implement effective suppression options. This monitoring will also identify zones where the risk of transport of life stages outside the regulated area is greatest as there are strong correlations between extensive eastern outbreaks and the frequency of trap captures in the west (Phero Tech Inc. 1994). Identification of increased risk resulting from elevated populations in source areas will trigger increased focus on the well-known pathways for gypsy moth transport; movement of people and their associated goods. Regulation of these pathways then becomes feasible via existing regulations pertaining to transportation of goods within Canada. In western Canada, information from monitoring gypsy moth in source areas will trigger monitoring efforts to detect new invasions. The risk factors with respect to pathways and susceptible host plants are well known. This analysis adds the capability of adding climate suitability to the assessment to provide an optimal monitoring scheme that balances cost and risk (Logan et al. 2007). While the expanded potential distribution of the gypsy moth in the Prairie Provinces appears largely contiguous under climate change, the potential distribution in British Columbia will continue to be discontinuous because of topographical and forest heterogeneity. In fact, the actual susceptible area in both western regions may be mostly fragmented because in the Prairies, most of the area at risk (based on climate) will remain agricultural or natural grassland with trees mostly isolated near urban

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centres, river bottoms and windbreaks. Thus, it should be possible to prevent establishment of the gypsy moth in all of western Canada by maintaining a strategic monitoring network and undertaking occasional eradication programs where the climate profile indicates that immigrant populations threaten to persist in an area. There are several sound reasons to maintain the gypsy moth-free status of western Canada. The first and most obvious is that the gypsy moth is a demonstrably destructive insect in North America. Wherever it has become established, significant damage has occurred to native and to high-value amenity trees. Infested areas then become subject to regulation with the associated negative economic impacts on trade. Further, this damage leads to the expensive and unpopular use of pesticides, often in urban areas or in habitats that are, ironically, set aside for conservation purposes and that are now at risk of a biological invasion. A less obvious reason for maintaining as large a gypsy-moth free area in Canada as possible is the additional threat of invasion by the Asian strain of the gypsy moth. Female moths of this strain are capable of flying and the larvae have a much broader host plant range, including conifer species, than does the European strain. The Asian strain of the gypsy moth is not currently present anywhere in Canada but constitutes a much higher risk to conifer forests, especially on the Pacific coast. At present, it is monitored with the same pheromone lure as the European gypsy moth. Captured male moths are distinguished as European or Asian only by subsequent DNA analysis. As long as western areas remain free of the gypsy moth, the few male moths trapped there each year can all be screened using molecular technology. However, if the European gypsy moth becomes established and common in western regions, the necessity of sub-sampling trap catches could severely compromise the ability of phytosanitary measures to detect the Asian gypsy moth invasion in Canada. The management of risk of invasion by alien species has far-reaching socio-economic and environmental benefits. An effective policy and response can be greatly aided by the application of scientific knowledge to interpret historical patterns and then to model likely future scenarios. Such an approach allows us to identify variability in the risk profile

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associated with different factors in different areas and thereby to determine variable and appropriate policies and responses. In our case, extensive knowledge of gypsy moth biology and ecology enabled us to design an effective management strategy to achieve realistic objectives based on the analysis of the risk factors discussed in this paper. Acknowledgements We thank Vince Waring, Ron Fournier and Pierre Duval, all of CFS, for help in collecting and verifying information and for preparing the maps.

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Hastings FL, Hain FP, Smith HR, Cook SP, Monahan JF (2002) Predation of gypsy moth (Lepidoptera: Lymantriidae) pupae in three ecosystems along the southern edge of infestation. Environ Entomol 31:668–675 Isaaks EH, Srivastava RM (1989) An introduction to applied geostatistics. Oxford University Press, New York Jobin L (1995) Gypsy moth, Lymantria dispar. In: Armstrong JA, Ives WGH (eds) Forest insect pests in Canada. Natural Resources Canada, Canadian Forest Service, Ottawa, pp 133–139, Fo24-235/1995E Liebhold AM, Tobin PC (2006) Growth of newly established alien populations: comparison of North American gypsy moth colonies with invasion theory. Popul Ecol 48: 253–262 Liebhold AM, Halverson JA, Elmes GA (1992) Gypsy moth invasion in North America: a quantitative analysis. J Biogeogr 19:513–520 Liebhold AM, Gottschalk KW, Muzika R-M, Montgomery ME, Young R, O’Day K, Kelley B (1995) Suitability of North American tree species to the gypsy moth: a summary of field and laboratory tests. USDA Forest Service, Northeastern Forest Experiment Station, General Technical Report NE-211 Logan JA, Casagrande RA, Liebhold AM (1991) Modeling environment for simulation of gypsy moth (Lepidoptera: Lymantriidae) larval phenology. Environ Entomol 20:1516–1525 Logan JA, Re´gnie`re J, Powell JA (2003) Assessing the impacts of global warming on forest pest dynamics. Front Ecol Environ 1:130–137 Logan JA, Re´gnie`re J, Gray DR, Munson AS (2007) Risk assessment in face of a changing environment: Gypsy moth and climate change in Utah. Ecol Appl 17:101–117 Lyons DB, Liebhold AM (1992) Spatial distribution and hatch times of egg masses of gypsy moth (Lepidoptera: Lymantriidae). Environ Entomol 21:354–358 Mayo JH, Straka TJ, Leonard DS (2003) The cost of slowing the spread of the gypsy moth (Lepidoptera: Lymantriidae). J Econ Entomol 96:1448–1454 Nalder IA, Wein RW (1998) Spatial interpolation of climatic normals: test of a new method in the Canadian boreal forest. Agric For Meteorol 92:211–225 Nealis VG (2002) Gypsy moth in Canada: case study of an invasive insect. In: Claudi R, Nantel P, Muckle-Jeffs E (eds) Alien invaders in Canada’s waters, wetlands, and forests. Natural Resources Canada, pp 151–159, Fo42329/2002E Nealis VG, Erb S (1993) A sourcebook for management of the gypsy moth. Canadian Forestry Service, Great Lakes Forestry Centre, Fo42-193/1993E Nealis VG, Roden PM, Ortiz DA (1999) Natural mortality of the gypsy moth along a gradient of infestation. Can Entomol 131:507–519 Phero Tech Inc (1994) A risk assessment of European gypsy moth in British Columbia. Prepared by Deloitte and Touche Management Consultants, Guelph, Ontario Pitt JP, Re´gnie`re J, Worner S (2007) Risk assessment of gypsy moth, Lymantria dispar (L), in New Zealand based on phenology modeling. Int J Biometeorol 51:295–305 Power JM (1986) FIDSINFOBASE: the forest insect and disease survey information system. Canadian Forest Service,

148 Petawawa National Forestry Institute, Information Report PI-X-65 Price DT, McKenney DW, Caya D, Coˆte´ H (2001) Transient climate change scenarios for high resolution assessment of impacts on Canada’s forest ecosystems. Report to Climate Change Action Fund (CCAF) and CCIS Reardon RC, Leonard DS, Mastro VC, McLane W, Leonhardt BA, Talley S, Thorpe K, Webb RE (1998) Using mating distruption to manage gypsy moth: a review. USDA Forest Service, Forest Health Technology Enterprise Team, FHTET-98-01 Re´gnie`re J (1996) Generalized approach to landscape-wide seasonal forecasting with temperature-driven simulation models. Environ Entomol 25:869–881 Re´gnie`re J, Nealis VG (2002) Modelling seasonality of gypsy moth, Lymantria dispar (Lepidoptera: Lymantriidae), to evaluate probability of its persistence in novel environments. Can Entomol 134:805–824 Re´gnie`re J, Sharov A (1998) Phenology of Lymantria dispar (Lepidoptera: Lymantriidae) male flight and the effect of moth dispersal in heterogeous landscapes. Int J Biometeorol 41:161–168 Re´gnie`re J, St-Amant R (2007) Stochastic simulation of daily air temperature and precipitation from monthly normals in North America north of Mexico. Int J Biometeorol 51:415–430 Sharov AA, Liebhold AM (1998a) Bioeconomics of managing the spread of exotic pest species with barrier zones. Ecol Appl 8:833–845 Sharov AA, Liebhold AM (1998b) Model of slowing the spread of gypsy moth (Lepidoptera: Lymantriidae) with a barrier zone. Ecol Appl 8:1170–1179 Sharov AA, Liebhold AM, Roberts EA (1998) Optimizing the use of barrier zones to slow the spread of gypsy moth (Lepidoptera: Lymantriidae) in North America. J Econ Entomol 91:165–174

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We can eliminate invasions or live with them. Successful management projects Daniel Simberloff

Originally published in the Journal Biological Invasions, Volume 11, No 1, 149–157. DOI: 10.1007/s10530-008-9317-z Ó Springer Science+Business Media B.V. 2008

Abstract Pessimism about the prospects of eradicating invasive species or managing them at acceptably low densities is fostered by publicity about failures, but it is often unwarranted. Many species, including insects, plants, and aquatic invaders of various taxa, have been eradicated, and a variety of management techniques have maintained others at low densities for long periods. Many of these projects entailed low-tech, scorched-earth approaches, often using mechanical, physical, or chemical means. Others required sophisticated scientific research. There are at best rough guidelines rather than general rules about what approach to undertake. It is therefore important in each case to consider the full range of possible methods for eradication or long-term control, including the possibility of marshalling a massive amount of physical labor. Keywords Biological control
Eradication
Maintenance management

Introduction Ironically, just when the public finally recognizes the scope of biological invasions, there is widespread

D. Simberloff (&) Department of Ecology and Evolutionary Biology, University of Tennessee, Knoxville, TN 37996, USA e-mail: [email protected]

pessimism that much can be done to alleviate the problem, both because of the inevitable increase in trade and travel and the seemingly endless ways in which some invaders circumvent attempts to control them. David Quammen (1998) expresses this view in his article ‘‘Planet of Weeds’’, envisioning a globally dominant biota comprising the animal and plant weeds of the world. In fact, there are many successful examples of eradication and long-term management of invasives at low densities, but these are often not well-known to the public, and sometimes even to scientists. Partly this is a function of the news media. Ex-Secretary of Defense Rumsfeld (U.S.) frequently complained about the invasion of Iraq that newspapers and television quickly publicize the bad news and largely ignore all the good things happening there. One can say the same about invasions generally. To take just one example, the invasion of the northwest Mediterranean by the tropical Pacific ‘‘killer’’ alga Caulerpa taxifolia (M. Vahl) C. Agardh generated headlines throughout Europe and even in the United States, featuring failed attempts to control the invader (Simons 1997; Naik 2001; cf. Meinesz 2001). Its arrival in California waters was publicized with alarm, and early failures to eradicate it highlighted (Perry and Mehta 2000; Scoch 2000). Yet the successful eradication in 2006 of C. taxifolia in California, in spite of its far wider distribution than the original invasion in the Mediterranean, was reported only on an interior page of the local press (Lin 2006) and not at all in Europe. That the media

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tend to report disasters and not triumphs has contributed to the sense of doom about biological invasions. A second reason why successful management of biological invasions is often not well known is that managers are usually rewarded for managing, not for publishing (Simberloff 1999). Many successful eradication and maintenance projects entail quite lowtech science, often just basic natural history (Simberloff 2003), and many resource management agencies and organizations have little tradition of publication in refereed journals, so their efforts— both successes and failures—are in the gray literature at best: agency reports and the like. However, a number of programs have succeeded in eliminating or reducing populations of introduced species, even in circumstances that appeared quite hopeless at the outset. Of course, the most effective way to deal with potential invaders is to exclude them, but I will treat only invasions that are already underway.

Eradication The possibility of eradicating introduced species is seductive, but scientists are often pessimistic (e.g., Dahlsten 1986; Whitten and Mahon 2005), especially for invertebrates, plants, and aquatic species in general. Part of the pessimism is due to a few highly visible, expensive, failed eradication campaigns with notable non-target damage, such as efforts to eradicate the red imported fire ant Solenopsis invicta Buren (Buhs 2004) and the gypsy moth Lymantria dispar L. (Spear 2005) in the United States. Nevertheless, many pests have been eradicated, including many invertebrates and some plants (Simberloff 2002a). Tse-tse flies (Glossina spp.), probably native, were first eradicated on the island of Principe between 1911 and 1914, and again in 1956 after reinvasion (Lapeyssonie 1988). By the use of hybrid sterility, native tse-tses were also eradicated from a discrete 26 km2 area in east Africa (Klassen and Curtis 2005). Subsequently, various species of Glossina have been eradicated from several African areas over 1,000 km2 by variants of the sterile-male technique (SIT) in combination with trapping and insecticides (Klassen and Curtis 2005). Other early insect eradications included the native screw-worm (Cochliomyia hominivorax (Coquerel))

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from Curac¸ao by SIT in 1954–1955 (Baumhover et al. 1955), and more recently from Puerto Rico, the Virgin Islands and the southern U.S. through Mexico and all the way to Panama by 2001 (Klassen and Curtis 2005). Introduced fruit flies have also been eradicated from many islands and some mainland areas by various combinations of SIT, trapping, male annihilation, and insecticides: for instance, the Oriental fruit fly (Dacus dorsalis (Hendel)) was eradicated from Rota and Guam between 1962 and 1965 (Steiner et al. 1965, 1970) and the melon fly (Bactrocera cucurbitae (Coquillett)) by 1993 from the Ryukyu Archipelago, including Okinawa and Amami-oshima (Iwahashi 1996; Kuba et al. 1996). The Queensland fruit fly (B. tryoni (Frogatt)) was eradicated from an incipient infestation of 125 km2 in Western Australia, while the Mediterranean fruit fly (Ceratitis capitata (Weidemann)) has been eradicated from 20 Florida counties (Ayers 1957), as well as small areas in western Australia and South Australia and a large part of the US, all of Mexico, and half of Guatemala (Klassen and Curtis 2005). The giant African snail (Achatina fulica Bowdich) was eradicated by manual and chemical means from small areas in south Florida (Mead 1979) and Queensland, Australia (Colman 1978). A recent dramatic molluscan eradication was that of the Caribbean black-striped mussel, Mytilopsis salei (Recluz), discovered in 1999 in Cullen Bay (12.5 ha), Darwin Harbor, probably within 6 months of arrival (Wittenberg and Cock 2001). The bay was quickly quarantined and treated with bleach and copper sulfate, and the mussel has not been seen since. It is often said that plants cannot be eradicated, particularly if they have a soil seed bank (Rejma´nek and Pitcairn 2002), but there have been a few notable successes, such as elimination of sandbur (Cenchrus echinatus L.) from Laysan (Flint and Rehkemper 2002; E. Flint, pers. comm. 2007) and Bassia scoparia (L.) A. J. Scott (kochia) from a large area in Western Australia (Randall 2001). Mack and Lonsdale (2002) cite several eradications of plants from small areas of Australia, New Zealand, and the United States, but they observe that there are few such examples. Many eradications are from small islands, but there have been notable successes on much larger islands. For instance, nutria (Myocaster coypus Kerr) has been eradicated from Great Britain (Gosling 1989) and the

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yellow fever mosquito (Aedes aegypti (L)) from Cuba (Fenner et al. 1988). Rat eradication has succeeded on islands of increasing size (Towns and Broome 2003), with planning now underway for an attempt to eradicate Rattus rattus (L) and R. exulans Peale from 274 km2 Great Barrier Island, New Zealand (J. Ogden, pers. comm. 2007). Having eradicated various combinations of introduced mammals from many islands in the Gulf of California (Tershy et al. 2002), Island Conservation, a California non-governmental organization, has undertaken, with collaborators, eradication of pigs and goats from much larger islands in the Galapagos. They have already eliminated pigs and goats from Santiago Island (58,000 ha) and are close to success with goats on 400,000 ha Isabella Island (Campbell et al. 2005; J. Donlan, pers. comm. 2006). At a continental scale, the African malaria vector, Anopheles gambiae Giles, was eradicated from 31,000 km2 of northeastern Brazil (Soper and Wilson 1943; Davis and Garcia 1989). What common features can we draw from these and other examples to suggest when to mount an eradication attempt and how to proceed? Five main features characterize most successful eradications (cf. Myers et al. 2000; Mack and Lonsdale 2002; Simberloff 2002b; Mack and Foster 2004): a)

Detecting an invasion early, and acting quickly to eradicate it. b) Sufficient resources allocated at the start to finish the project, including post-eradication surveys and follow-up, if necessary. c) Existence of a person or agency with the authority to enforce cooperation. Eradication cannot succeed even if the great majority of stakeholders cooperate in the campaign so long as a small minority allow the invader to persist on property they control. d) The target species must be studied well enough to suggest vulnerabilities. Often basic natural history suffices. e) Project leaders must be energetic, optimistic, and persistent in the face of occasional setbacks. The importance of acting quickly and decisively and the need for post-eradication surveys and followup are exemplified by the successful eradication of the Pacific alga Caulerpa taxifolia from two substantial areas in southern California (Merkel & Associates, Inc 2006) and the failure to eradicate

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the same species in Europe (Meinesz 2001). A small infestation of the alga first appeared in 1984 just offshore of the Oceanographic Museum of Monaco. The French and Monacan government agencies argued first about who was responsible for the infestation, then about whether it would become invasive, and finally, as it became apparent that an invasion was underway, about how to deal with it. The upshot is that C. taxifolia now infests over 10,000 ha off the coasts of Spain, France, Monaco, Italy, Croatia, and Tunisia. In California, by contrast, within six months after the alga was discovered, the Southern California Caulerpa Action Team (a public–private partnership) was established and proceeded to garner public support for a brute-force eradication campaign. The effort entailed extensive diving to locate all infestations, then covering them with tarpaulins and pumping in chlorine. When monitoring uncovered infestations that had escaped initial detection, these too were covered and poisoned. No further Caulerpa was seen after 2 years of intensive monitoring, and on July, 12, 2006, the State of California celebrated the success of the effort, which cost ca. US$7 million. Finding invasions early and acting promptly is not an absolute sine qua non for eradication, as witness both the global eradication of smallpox (Fenner et al. 1988) and great progress in attempting to eradicate witchweed (Striga asiatica (L.) Kuntze) from the United States after a long, widespread establishment (Eplee 2001). It is therefore important not to have hard-and-fast thresholds of establishment beyond which eradication will never be undertaken. However, the cost of eradication is likely to rise dramatically when invasions are widespread; both the smallpox and the witchweed campaigns cost over US$100 million. One would therefore expect great interest in early-warning systems. However, even New Zealand, the nation with the most comprehensive strategy on introduced species, recognizes monitoring for rapid detection of invaders and an adequate structure for a rapid response as two weaknesses in their approach (Parliamentary Commissioner for the Environment 2000). In the United States, the 2001 draft action plan of the Federal Interagency Committee for the Management of Noxious and Exotic Weeds for a national early warning and rapid response system for invasive plants has yet to be incorporated in a revision of

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the National Invasive Species Management Plan, approved by the National Invasive Species Council in 2001. It is frustrating not to be able to provide very specific guidance about when and how to undertake an eradication campaign. However, it is a fact that ecological knowledge is generally accrued through amassing a catalog of case studies, each of which is, to an extent, sui generis, but which together can give guidance about the ways to approach study and management of a new system (Simberloff 2004). Ecology also lacks general laws, except at levels so high (e.g., the laws of thermodynamics) that they cannot provide management guidance (Simberloff 2004). Invasion biology, as a branch of ecology, is similarly often criticized as just a collection of special cases (see Williamson 1999), but it is simply the nature of the beast that invasions are sufficiently complex and different from one another that there will likely be many exceptions to any general rule that might be formulated by examination of the growing catalog of cases. As an example, many invasion biologists and managers, mostly working with plants, advocate attacking peripheral, nascent foci first, then gradually attacking large, central populations. This notion is based on common sense (e.g., Fuller and Barbe 1985), several empirical experiences (see Mack and Lonsdale 2002), and two theoretical treatments (Moody and Mack 1988; Higgins et al. 2000). However, recent, spatially explicit models for Himalayan balsam (Impatiens glandulifera Royle), giant hogweed (Heracleum mantegazzianum (Sommier & Levier), and Old World climbing fern (Lygodium microphyllum (Cav.) R. Br.) suggest that the specifics of particular invasions and costs of various management options could lead to exactly the opposite recommendation (Wadsworth et al. 2000; Duke-Sylvester 2006). A model similar to that of Moody and Mack (Whittle et al. 2007) also leads to a result different from theirs. As noted above, failed eradications are often widely reported, while successful campaigns may be noted only in the gray literature. This is not a phenomenon restricted to Caulerpa eradications. A successful 9-year campaign to eradicate a major part of the Chicago invasion of the Asian longhorned beetle (Anoplophora glabripennis (Motschulsky)) was barely noted, and then only in the local press. By contrast, the failed campaign to eradicate northern

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pike (Esox lucius L.) from Lake Davis (California) was reported in national newspapers (e.g., Nieves 2002) and is even the subject of a documentary film (Elmendorf et al. 2005). Even more damaging to the development of a technology of eradication, less dramatic failed eradication attempts are often not reported except perhaps in the grayest of literature, hindering the ability of scientists to learn from previous efforts.

Maintenance management Many introduced species have been controlled for long periods at low densities, but, as for eradication, there is often great pessimism that such an approach can succeed, and some efforts, both successes and failures, are not published in the scientific literature. For instance, musk thistle (Carduus nutans L.) is well controlled in Kentucky by supervised crews of drunkdriving convicts, a program that required optimism and persistence in the face of a soil seed bank (J. Bender, pers. comm. 2000). Nature Conservancy policies precluding herbicides led to a successful campaign to control European beachgrass (Ammophila arenaria (L.)) manually at the LanphereChristensen Dunes Preserve, California (now part of the Humboldt Bay National Wildlife Refuge). This program used a state public works program plus a convict alternative work plan and has kept the site clear of beachgrass for a decade save for small, easily removed infestations (Pickart and Sawyer 1998; A. Pickart, pers. comm. 2005). This plant can also be kept at low density with herbicides (Aptekar 2000). A collective in Tennessee, The Farm, with a policy of not using herbicides, managed to control kudzu (Pueraria montana var. lobata (Willd.) Maesen & S. Almeida) using prescribed burns plus manual labor, followed by annual monitoring for survivors and recruits (Rohrbach 2003). Very substantial areas can be kept clear of invaders for the long term, as witness the control of water hyacinth (Eichhornia crassipes (Mart.) Solms) in Florida. In 1960, this plant infested 51,000 ha. After failed attempts at mechanical and biological control, the state of Florida reduced water hyacinth to ca. 2,000 ha (with locations changing annually) primarily by using the herbicide 2,4-D (Schardt 1997). Importantly, the cost and health issues

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associated with 2,4-D waned as the amount required quickly fell to less than 1% of the early usage. Florida has recently engaged in an ambitious campaign to reduce Australian paperbark (Melaleuca quinquenervia (Cav.) S. F. Blake) from its peak of ca. 202,000 ha in the late 1980s (Ferriter 2004). The main tool is aerial spraying of herbicides with manual application to sprouts and survivors, and with some manual and physical methods (Laroche 1994; Anonymous 2005). As paperbark on public lands has dwindled, a weevil and a psyllid from Australia have been released with the aim of preventing re-establishment and controlling this tree on private lands. Though more than a third of infested areas have been cleared, there is also substantial research on more effective, less costly methods (e.g., Myers et al. 2001; Laroche and McKim 2004). An important aspect of the success of the Florida programs for water hyacinth and paperbark was consistent, adequate funding, mostly from the state, to prevent these plants from re-establishing during a lull in management activity. Also noteworthy is the Alberta Rat Patrol, a provincial program that has successfully maintained Norway rats at very low levels through use of anticoagulant baits and aggressive hunting by a team of officers (Bourne 2000). Successful maintenance management for both agriculture and conservation sometimes entails classical biological control. For instance, on the island of St. Helena, the introduced lady beetle Hyperaspis pantherina Fu¨rsch probably saved the endemic gumwood tree, Commidendrum robustum (Roxb.), from extinction by the South American scale insect Orthezia insignis Browne (Wittenberg and Cock 2001). In Florida, rapidly spreading aquatic alligatorweed (Alternanthera phyloxeroides (Mart.) Griseb.) is now well controlled by introduced natural enemies, especially the flea beetle Agasicles hygrophila Selman & Vogt (Center et al. 1997). Such successes have led to a popular view that biological control is a panacea for metastatic invasions, especially because, when it works, biological control may operate in perpetuity without further human intervention. However, biological control is not without limitations (Simberloff and Stiling 1996). It usually fails to control the target species (Williamson 1996), and some introduced natural enemies threaten species of

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conservation or commercial importance. A recent worrisome example is the spread of the South American cactus moth Cactoblastis cactorum (Berg) through the southeastern United States towards Mexico, after it island-hopped from the Lesser Antilles to the Florida Keys long after introduction to Nevis for control of pest cactus. This moth may now affect many species of Opuntia prickly pear, including species of agricultural, horticultural, and conservation concern (Zimmermann et al. 2001). Such mishaps can be rendered less likely by extensive host-testing and avoiding release of any species that is not narrowly host-specific, but biological control will always entail risks, and, unlike chemical and mechanical control, an errant biological control release is generally irreversible. For this reason, the full panoply of possible approaches should be considered for any introduced species requiring management, and biological control should perhaps be seen as a last resort.

Discussion Many successful eradication and maintenance management campaigns are decidedly low-tech and rely on massive use of manual labor (Simberloff 2003). In an age when manual labor is viewed as old-fashioned and governments as well as industry struggle to minimize labor costs, it is important to note two sources of low-cost labor that have already been widely used in invasive species management and could doubtless be employed even more extensively. Many projects use convict labor, including those to control musk thistle, beachgrass, and paperbark (Campbell and Carter 1999) discussed above. Most convicts pull, dig, or cut invasive plants, but other uses are possible. In Oregon, convicts construct pheromone traps used in campaigns to eradicate the Asian race of the gypsy moth, L. dispar (Manzano 1995). Volunteers can be enlisted in wealthy nations to provide a huge manual labor force for both eradication and maintenance management. In the United States, the Nature Conservancy uses volunteers frequently for such purposes (Randall et al. 1997), while in the both Canada and the United States schoolchildren often engage in projects to manage invasive plants, thereby sensitizing the populace as a

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whole to the issue of biological invasions. Ecotourists pay over US$1,000 to rid scenic areas (such as national parks) of introduced plants (Newman 2004). Probably the major limitation on use of volunteers, as for convicts, is the requirement for supervision and training. However, it is evident that approaches that might have been rejected on grounds of sheer personpower requirements are, in fact, feasible under many circumstances. Paid labor is also possible but seems difficult to justify for environmental rather than agricultural uses in the dominant political climate nowadays. However, such climates change, and the example of the South African Working for Water Programme suggests how a public works program can greatly aid the battle against introduced species while fulfilling other socioeconomic goals as well. Initiated in 1995, this project has employed tens of thousands of previously destitute citizens and effectively controlled many invasive plants (McQueen et al. 2000; van Wilgen et al. 2000). Critical to the success of projects using massive manual labor, whether paid, volunteer, or convict, is the fact that often highly detailed knowledge of the target species is unnecessary, and a crude, scorched earth approach suffices (Simberloff 2003; Krajick 2005). However, not every problem will yield to such methods, and some high-tech methods employing sophisticated science have been remarkably effective. For example, efforts to eradicate goats from the Galapagos Islands with the Judas goat technique were stymied on large islands when females typically became pregnant within a month of release and went off to be alone to deliver and nurse their kids. To solve this problem, Karl Campbell developed a method of producing super-Judas goats by capturing females, inducing abortion, sterilizing them by Fallopian tube clips, and giving them a hormone injection that produces estrous for over 100 days (Campbell et al. 2005). This procedure is performed in the field in minutes. A recent exciting project entailing sophisticated science concerns the sea lamprey (Petromyzon marinus L.) in the Great Lakes. Management has focused to date on use of lampricides and dams, and although there has been moderate success in decreasing lamprey numbers, these methods are expensive and have non-target impacts. Peter Sorensen and his colleagues (Sorensen and Vrieze 2003; Sorensen and

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Stacey 2004; Sorensen et al. 2005) have developed an entirely new approach. The sea lamprey is anadromous, and adults returning from the ocean or large lakes to breed avoid streams with no larvae. Sorensen suspected larvae emit an attraction pheromone, and his team has concentrated it from water containing larvae, characterized it chemically, and synthesized it. Larvae can be lured into traps or diverted into blocked or poor-quality streams with this pheromone. Widespread use of the pheromone is currently prevented by a patent fight between two organizations that funded the research. Many effective maintenance management projects and successful eradications employ chemicals, alone or in concert with mechanical or physical methods. Expense of chemicals, especially when used for environmental purposes over large areas, and the evolution of resistance are frequent and valid concerns, but chemicals also generate opposition that seems ideological (Williams 1997). All environmentalists have been heavily influenced by Rachel Carson’s masterpiece, Silent Spring (1962), which aroused the public about non-target impacts of pesticides. However, pesticides and herbicides registered today, if used properly, have far more limited non-target impacts, and in some instances they may be the only means currently available to stop irreversible damage from an invasion, at least until some other method is developed. Such is the case with the explosion in the late 1980s of the introduced yellow crazy ant (Anoplolepis gracilipes F. Smith) on Christmas Island, threatening the ecological keystone species of the island, the native red land crab (Gecarcoidea natalis (Pocock)), as well as many other aspects of the island ecosystem (O’Dowd et al. 2003; Abbott 2004). What appeared to be a hopeless case and a conservation disaster was controlled by the use of the bait Fipronil in a fish protein base (\http://www.deh.gov.au/parks/ christmas/fauna/crazy.html[, accessed 8/29/06). Part of the pessimism about controlling invasions arises from widely publicized management failures, especially failed eradications. Worse, many cases of successful eradication and maintenance management are unpublicized or barely mentioned in the popular press or scientific literature. Nevertheless, the catalog of successful cases, of which the above cases are just a sample, is now sufficiently large and known at least to professional managers and invasion biologists that

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one would expect the public to demand more implementation and policymakers to respond. However, the fraction of invasions for which eradication is seriously attempted is minuscule, and many maintenance management programs are so sporadically funded that damage remains high. Why is the effort not greater? One possibility is that scientists and managers with relevant expertise have simply not been assiduous or effective enough at publicizing the success stories. It is a common lament in conservation biology and environmental science generally that scientists lack the expertise and sometimes the desire to interact with the public and policymakers. Without judging the merits of this complaint, I suggest that the problem for invasion management in general is somewhat different, and the failure in particular to implement eradication and ongoing control problems sufficiently frequently is part of this problem. The effective approaches outlined above all require preventive action of some sort, at some expense. The most effective way to prevent damage from invasions would be better exclusion; although exclusion is not addressed in this paper, it is widely known by scientists and the public that exclusionary measures are quite weak. The single biggest difficulty of implementing the next line of defense—eradication— is absence of an ongoing early warning system; that is, a rigorous, comprehensive, ongoing monitoring program to detect new invasions. The next issue is that, for the great majority of invaders whose presence is already known, there is little evidence of current harm, but the frequent occurrence of a lag time before major invasions, often a substantial lag time, is well known (e.g., Crooks 2005). Yet, because of the great complexity of the forces controlling invaders, and the many idiosyncrasies of each invader, we cannot predict with much assurance which established introductions are going to lead ultimately to damaging invasions. Finally, with many eradication campaigns, the harmful impact disappears in the initial stages of the campaign, even before the target species has been eradicated. Similar, many successful maintenance management campaigns reduce the impact of a pest to a level at which it is not in the public eye. In each of the above matters, what is required is preventive measures against events that will happen only probabilistically: exclusionary rules and machinery to keep out species, only some of which would

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become problematic; ongoing surveys that will certainly detect many introductions that may never become pests; eradication campaigns against species that are not currently problematic and may never be; continued maintenance measures against species that are currently controlled and that may not bounce back if the ongoing funding level and resulting control measures were decreased. Yet it is far more difficult to generate public fervor and support from policymakers for prevention than for cures to existing problems. Insurance companies balk at paying for vaccines, despite studies showing ultimate cost savings. Bridges collapse for lack of infrastructure maintenance, to be rebuilt at great cost on a crash schedule. It appears to be a fundamental aspect of human nature that preventive measures, even those demonstrated to be cost-effective, are not easily implemented if they are expensive, and sometimes even if they are not very expensive. It is unlikely that preventive measures for damage from biological invasions will be implemented nearly as widely as they might be until the public has fully grasped the overwhelming, daily cost imposed by some invasions, plus the notion of probabilistic events. This will be a hard sell, and it is unfair to indict invasion specialists for not having had great success to date. Acknowledgment Richard Mack and an anonymous referee provided insightful comments on an early draft of this manuscript.

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