Integrating Biological Control into Conservation Practice E-Book

• Herausgeber: John Wiley & Sons
• Kategorie: Wissenschaft und neue Technologien
• Sprache: Englisch

Seitenzahl: 1178

Veröffentlichungsjahr: 2016

Beschreibung

Invasive species have a critical and growing effect upon natural areas. They can modify, degrade, or destroy wildland ecosystem structure and function, and reduce native biodiversity. Landscape-level solutions are needed to address these problems. Conservation biologists seek to limit such damage and restore ecosystems using a variety of approaches. One such approach is biological control: the deliberate importation and establishment of specialized natural enemies, which can address invasive species problems and which should be considered as a possible component of restoration. Biological control can be an effective tool against many invasive insects and plants but it has rarely been successfully employed against other groups. Safety is of paramount concern and requires that the natural enemies used be specialized and that targeted pests be drivers of ecological degradation. While modern approaches allow species to be selected with a high level of security, some risks do remain. However, as in all species introductions, these should be viewed in the context of the risk of failing to reduce the impact of the invasive species.

This unique book identifies the balance among these factors to show how biological control can be integrated into ecosystem restoration as practiced by conservation biologists. Jointly developed by conservation biologists and biological control scientists, it contains chapters on matching tools to management goals; tools in action; measuring and evaluating ecological outcomes of biological control introductions; managing conflict over biological control; and includes case studies as well as an ethical framework for integrating biological control and conservation practice.

Integrating Biological Control into Conservation Practice is suitable for graduate courses in invasive species management and biological control, as well as for research scientists in government and non-profit conservation organizations.

Leseprobe

Table of Contents

Cover

Title Page

List of contributors

Preface

References

CHAPTER 1: Integrating biological control into a conservation context: why is it necessary?

Potential problems if integration is lacking

Book organization

Acknowledgments

References

CHAPTER 2: Designing restoration programs based on understanding the drivers of ecological change

Overview of concepts

Designing a restoration plan using Connecticut River floodplain forests as a model

Acknowledgments

References

CHAPTER 3: Matching tools to management goals

Introduction

Eradication

Limiting spread

Local, or area-wide, temporary suppression of invaders

Area-wide, permanent suppression through modification of ecosystem processes

Area-wide, permanent control through natural enemy introductions

Factors affecting control efficacy

Effects of treatment scale on control operations

When is biological control the right choice?

Key messages

Acknowledgments

References

CHAPTER 4: Tools in action: understanding tradeoffs through case histories

Risk tradeoffs in invasive species management

When is no action the right choice?

Is herbicide use a good option for invasive plant control?

Are biological control projects high or low risk?

Acknowledgments

References

CHAPTER 5: Benefit–risk assessment of biological control in wildlands

Who gets to decide which species are targeted

Evidence needed to justify use of biological control

What principles and processes should guide decision-making?

Risk assessment of biological control agents

Post-release risk evaluation

Conclusions

Acknowledgments

References

CHAPTER 6: Systematics and biological control

Introduction

Identification’s effect on biological control service

Using classifications to make biological predictions

The importance of voucher specimens

Molecular methods

Conclusions

Acknowledgments

References

CHAPTER 7: Forecasting unintended effects of natural enemies used for classical biological control of invasive species

Introduction

Forecasting non-trophic effects

Forecasting unwanted trophic effects

Conclusions

Acknowledgments

References

CHAPTER 8: Measuring and evaluating ecological outcomes of biological control introductions

Introduction

Setting the stage

Defining success in biological control programs

Evidence for success or failure

Population growth rates and demographic modeling

Principles of stewardship

Conclusions

Acknowledgments

References

CHAPTER 9: Methods for evaluation of natural enemy impacts on invasive pests of wildlands

Types of outcomes used as measures of impact

Evaluation methods for invasive insects

Evaluation methods for invasive plants

Conclusions

Acknowledgments

References

CHAPTER 10: Cases of biological control restoring natural systems

The value of case histories

Successes

Failures

Projects still unfolding

Acknowledgments

References

CHAPTER 11: Societal values expressed through policy and regulations concerning biological control releases

Introduction

National regulatory processes and differences

Scientists and regulators: different perspectives

Differences in national approaches and associated cultural values

Where can process improvements be made?

Conclusions

Acknowledgments

References

CHAPTER 12: Managing conflict over biological control: the case of strawberry guava in Hawaii

Introduction

Strawberry guava in Hawaii

Regulatory review of strawberry guava biological control

Lessons learned about our public

Lessons learned about the process

Conclusions

Acknowledgments

References

CHAPTER 13: An ethical framework for integrating biological control into conservation practice

Why integration of biological control into conservation practice is an ethical issue

How environmental thought evolved to encompass classical biological control as an ethical issue

Risk, precaution, prudence, and policy

A framework to improve ethics of biological control decisions

Conclusion

Acknowledgments

References

CHAPTER 14: Economics of biological control for species invading wildlands

Introduction

Losses from wildland invaders

Benefit: cost ratios for biological control

Additional costs of restoration to desired conditions

Effects of economic interests on target selection for biological control

How should biological control in wildlands be financed?

Conclusions

Acknowledgments

References

CHAPTER 15: The future of biological control: a proposal for fundamental reform

Introduction

Ethical and normative foundations of stewardship

Why is biological control a better choice for pest management?

What is the status quo?

Assignment of responsibilities

Biological control program flow structure

Conclusions

Acknowledgments

References

Concluding thoughts on future actions

Index

End User License Agreement

List of Tables

Chapter 03

Table 3.1 Criteria for determining safety and feasibility of invasive species management.

Table 3.2 Summary of management options for invasive species.

Table 3.3 Steps in classical biological control.

Chapter 04

Table 4.1 Biological control projects selected to illustrate a spectrum of risk due to characteristics, not of potential biological control agents, but of the project itself.

Chapter 05

Table 5.1 Factors to be considered in choosing a management method for an invasive population.

Chapter 09

Table 9.1 Projection matrices, dominant eigenvalues, and elasticity matricies for

Carduus nutans

at two sites in New Zealand. Reproduced with permission from Shea and Kelly (1998).

Chapter 12

Table 12.1 Impacts expected from biological control of strawberry guava.

Table 12.2 Significance criteria specified for evaluation in a State of Hawaii Environmental Assessment: an Environmental Impact Statement is required if, to a significant level of expected impact, a project …

List of Illustrations

Chapter 02

Figure 2.1 A chart to classify the ecological role of an invasive species on the spectrum from invasion being a consequence of ecological change to invasion being the cause of ecological change.

Figure 2.2 Decision tree to assist in determining the strategy that is likely to be most effective at restoring an ecosystem.

Figure 2.3 Forlorn TNC intern standing surrounded by Japanese knotweed in a high-gradient river floodplain forest on the Green River in Massachusetts, June 23, 2009.

Figure 2.4 Connecticut River floodplain forest breaking down under a heavy load of invasive oriental bittersweet and turning into a weedy vine thicket, West Springfield, Massachusetts, March 15, 2013.

Figure 2.5 Tree size distribution for American elm (

U. americana

) and silver maple (

A. saccharinum

), the two most common tree species in Connecticut River floodplain forests when surveyed in 2008 to 2011. Note the rapid reduction of elms beyond 20 cm and absence over 60 cm dbh, in contrast to silver maple, likely due to Dutch elm disease.

Chapter 03

Figure 3.1 Mechanical removal of non-native gorse (

Ulex europaeus

) for wildlife habitat improvement at the Coquille Unit of Oregon Islands National Wildlife Refuge, Coos County, Oregon, USA.

Figure 3.2 (a) Flightless female gypsy moth,

Lymantria dispar

; (b) lure traps using female pheromones effectively capture male adult gypsy moths.

Figure 3.3 To target individual trees, cuts can be made into the trunk of a tree and a herbicide applied.

Figure 3.4 Island Conservation, Parks Canada, and Haida Nation worked together in 2013 to protect Ancient Murrelets by removing invasive rats. This image shows an aerial broadcast of cereal pellets on Murchison and Faraday Islands, Haida Gwaii, British Columbia.

Figure 3.5 (a) Treatment area for eradicating

Wasmannia auropunctata

from a 21 ha area on Marchena Island in the Galápagos Islands archipelago. The inset is a closer view of the 3–4 m grid of bait monitoring stations (Causton et al., 2005); (b) preparation of peanut butter bait sticks for monitoring (up to 44,000 baits were put out in grids).

Chapter 04

Figure 4.1 Prescribed fire to control Scotch broom on the Elizabeth Islands, Massachusetts. Scotch broom has invaded maritime grasslands.

Figure 4.2 Herbicide application is a common, usually non-selective form of control used against invasive plants. Here, an aerial herbicide application to kill

Phragmites australis

.

Chapter 05

Figure 5.1 Differences by taxon in rates at which intentionally introduced and accidentally introduced species become pests suggest that, for terrestrial vertebrates, fish, and mollusks, deliberate introduction does not lower risk. In contrast, for insects and plant pathogens, categories for which biological control introductions would dominate the intentionally released groups, risk is highly reduced over random introduction. This reduction suggests that biological agent selection seeks safe agents and avoids release of risky species, which is not the case on average for intentionally released species in other taxa (Office of Technology Assessment, 1993).

Figure 5.2 Biological control is a method of invasive species management that generally shows high levels of target specificity. Shown here are two brown clumps of purple loosestrife (

Lythrum salicaria

) selectively defoliated by

Galerucella

leaf beetles introduced to North America, while all remaining wetland vegetation remains untouched despite food limitation for developing larvae.

Figure 5.3 Larvae of

Cactoblastis cactorum

, an herbivore that attacks various native and pest

Opuntia

cacti.

Figure 5.4

Rhinocyllus conicus

, a polyphagous thistle-head feeding insect that is well known to attack a range of native thistles.

Figure 5.5 Adult polyphemus moth,

Antheraea polyphemus

(Cramer), one of the giant silkmoths strongly attacked by the tachinid

Compsilura concinnata

.

Chapter 06

Figures for Box 6.1 Taxonomic errors can contaminate efforts to evaluate the risks associated with a biological control action (or a decision not to take action).

Manduca brontes

(left) (photo courtesy of Andy Warren) has been treated as a species at risk from the loss of ash, but the species is largely an extralimital, temporary breeder in the continental United States and published reports of the caterpillar feeding on ash were based on no-choice laboratory-reared caterpillars; temporary populations in Florida are believed to use yellow trumpetbush,

Tacoma stans

(L.). Grant’s Hercules beetle (

Dynastes granti

) is the largest beetle in the North America with some males approaching 80 mm in length (right) (photo courtesy of Margarethe Brummermann). It and three other rhinoceros beetles appear to be dependent on ash either as larvae (

Xyloryctes

) or as adults (

Dynastes

) (Wagner and Todd, 2015, 2016). Previous risk assessments for EAB in North America overlooked these charismatic beetles. See text.

Figure 6.1 Haplotype data can be used to identify geographic distributions of invasive organisms. Williams et al. (2007) mapped the probability of having the eastern (left) and western (right) haplotypes of

Schinus terebinthifolius

in Florida. The identification of these genetic and geographic differences also led to the determination that different haplotypes vary in their susceptibility to a candidate biological control agent,

Pseudophilothrips ichnini

, which itself appears to represent a cryptic-species complex (Manrique et al., 2008). Williams et al. 2007.

Figure 6.2 Cladogram for wax scales. The Chinese wax scale

Ceroplastes sinensis

was a cosmopolitan pest of citrus at the time it was described by Del Guercio in 1900 (his holotype was collected in Italy). As indicated by its name, Del Guercio assumed the scale to be of Asian origin, but a cladistic analysis by Qin et al. (1994) based on 57 morphological characters, suggested otherwise. In Qin et al.’s (1998) tree,

C. sinensis

grouped in a clade that contained nine Neotropical wax scales. Subsequent studies suggested the species was native to southern South America, and an expedition to Argentina resulted in the species’ rediscovery as well as parasitoids that were successfully introduced elsewhere. Qin and Gullan 1995.

Figure 6.3 Susceptibility of tropical angiosperms to pathogenic fungi correlates with phylogenetic distance. Such relationships across trophic levels demonstrate the value of using phylogenies and classifications to design non-target/specificity studies that can be used to gauge potential non-target impacts that could accrue from the release of a biological control agent.

Figure 6.4 Ecological niche modeling (ENM) can be integrated with the results from molecular analyses to identify variation in environmental factors influencing the distributions of candidate biological control agents. Using ENM software, Lozier and Mills (2009) identified different distributions for evolutionary significant units (ESUs) of the parasitoid

Aphidius transcaspicus

. They then used these results to identify climatic variables that might restrict each of these ESUs and may influence their establishment as biological control agents. Lozier and Mills 2009. Used under CC-BY-4.0, http://creativecommons.org/licenses/by/4.0/.

Figures for Box 6.2 Improvements in non-destructive DNA extraction methods are allowing museum specimens to be integrated into molecular analyses with little to no morphological damage to the specimens. Andersen and Mills (2012) extracted DNA from individual museum specimens of

Meteorus

sp. On the left is a specimen prior to DNA extraction, in the center the specimen placed in a 1.5 ml microcentrifuge tube, and on the right after articulation following DNA extraction. While not visible in the photo, post-extraction specimens appeared slightly translucent compared to pre-extraction specimens.

Chapter 08

Figure 8.1 Forests in eastern North America have been transformed by excessive browse of white-tailed deer (

Odocoileus virginianus

), eliminating forest understories and tree regeneration (Binghamton University forest preserve in New York State, USA).

Figure 8.2 Leaf beetles (

Galerucella

spp.), biological control agents released against purple loosestrife (

Lythrum salicaria

L.) often eliminate all annual growth, a biological success (here Montezuma wetlands complex in upstate New York). Only if a diverse native flora and fauna replaces purple loosestrife is this biological success also an ecological success.

Figure 8.3 Showy displays of the native spring ephemeral

Trillium grandiflorum

(Michx.) Salisb. were once common in northeastern US forests. The decline in the species has often been associated with non-native plant invasion, but overabundant native deer are a larger demographic threat.

Figure 8.4 Permanently marked (note numbered metal tag)

Trillium grandiflorum

used to assess effects of deer browse on plant demography.

Chapter 09

Figure 9.1 Evaluation methods “before/after” combined with “release/control” to measure impacts of the egg parasitoid

Gonatatocerus ashmeadi

on glassy-wing sharpshooter,

Homalodisca vitripennis

.

Figure 9.2 Changes in density of larch casebearer,

Coleophora laricella

, in Oregon (USA) over a 25-year period, in relation to the introduction of the parasitoid

Agathis pumila

, showing a drop in density of approximately 97% in the “after” period as compared to the “before” interval, demonstrating effective and sustained biological control of this pest.

Figure 9.3 Use of exclusion cages in the native range (Australia) of cottony cushion scale (

Icerya purchasi

) to measure impact of the ladybird beetle

Rodolia cardinalis

.

Figure 9.4 Recovery of

Pisonia grandis

on Cousine Island in the Seychelles after use of ant-baiting to break up mutualism supporting high density of invasive scale: left before baiting; right after baiting.

Figure 9.5 Net population growth rates (

R

0

) of emerald ash borer populations (

Agrilus planipennis

) estimated from life tables of each observed generation across different study sites in Michigan during the seven-year study (2008–14): solid black line represents

R

0

values when all the observed parasitism (by both introduced parasitoid

Tetrastichus planipennisi

and all North American parasitoids); long-dashed line represents

R

0

values when parasitism by

T. planipennisi

is absent, and short-dashed line represents

R

0

values when all the parasitoid species were absent from the life tables.

Figure 9.6 Relationship between seed bank densities and spotted knapweed (

Centaurea stoebe

L.) density, demonstrating the concept of processes acting with thresholds.

Figure 9.7 Relationship over time between density of

Sesbania punicea

at plots with one (A–C), two (D–F or G–L, each species), or three (M–P) biological control agents, in South Africa.

Figure 9.8 Before (left) (1997) and after (right) (1999) control of waterhyacinth (Eichhornia crassipes) at Port Bell on Lake Victoria, Uganda.

Figure 9.9 Sample life cycle diagram of

Carduus nutans

in New Zealand.

Figure 9.10 Increase in native plant richness in plots following biological control of mist flower (

Ageratina riparia

) in New Zealand, where the solid dots are for an invader plot and the hollow dots are for an invaded plot where the weed came under biological control.

Chapter 10

Figure 10.1 Melaleuca island in an invaded wetland.

Figure 10.2 Few to no plants can survive in the sterile habitat under melaleuca stands.

Figure 10.3 The melaleuca weevil,

Oxyops vitiosa

, destroys seed production by more than 95%.

Figure 10.4 (a) The melaleuca psyllid,

Boreioglycaspis melaleucae

Moore (adult) was the second agent released against melaleuca. (b) Psyllid flocculence, showing a colony of the insect.

Figure 10.5 Recovery of native vegetation in a melaleuca-dominated plot, following successful biological control, of melaleuca; note the dead trees and open canopy.

Figure 10.6 Decline in melaleuca density and increase in species diversity as biological control agents exerted their effect at a Florida research site.

Figure 10.7 Control of mist flower (

Ageratina riparia

) by the introduced white smut fungus,

Entyloma ageratinae

, showing the same location (a) about a year after fungus release when plants were infected but not yet killed (January 27, 2000–summer) and (b) two years later when mist flower shrubs had died and been replaced by mostly native vegetation, particularly ferns and grasses (November 8, 2001–spring); site is Brookby on the edge of the Hunua Ranges (an important water catchment and biodiversity area, just southeast of Auckland, New Zealand), and (c) endangered New Zealand endemic plant,

Hebe acutiflora

Cockayne, no longer threated by mist flower.

Figure 10.8 White mangrove infested with cottony cushion scale.

Figure 10.9 Endemic Galápagan plants whose populations were threatened by cottony cushion scale: left

Darwiniothamnus tenuifolius

, and right,

Scalesia

sp.

Figure 10.10 Adult

Rodolia cardinalis

preying on cottony cushion scale.

Figure 10.11 Reduction of cottony cushion scale on white mangrove, on Santa Cruz island, Galápagos, before (2002) and after (2010 and 2011) release of

Rodolia cardinalis.

Hoddle et al. 2013.

Figure 10.12 Walk-in field cage (right) stocked with native plants (left) infested with non-target insects to assess feeding preferences of adult

Rodolia cardinalis

.

Figure 10.13

Rodolia cardinalis

larva faced with a choice: cottony cushion scale or

Ceroplastes

sp. scale (top), attacks the cottony cushion scale (bottom).

Figure 10.14 Remnant patch of lowland dry forest, with wiliwili tree,

Erythrina sandwicensis

, in foreground at Puuwaawaa, Big Island, Hawaii.

Figure 10.15 Adult of erythrina gall wasp,

Quadrastichus erythrinae

(left, photo courtesy of M. Tremblay), and galled leaves from Koko Crater (right, photo credit, Leyla Kaufman).

Figure 10.16 Exploration in Madagascar (left, photo credit, Russell Messing) did not yield parasitoids; but additional collections in South Africa, Kenya, and Tanzania were successful, yielding several parasitoid species – including

Eurytoma erythinae

(right, photo courtesy of Walter Nagamine), which is helping to preserve rare endemic trees in Hawaii.

Figure 10.17 Balsam woolly adelgid on trunk (left) and gouty twigs resulting from twig infestation (right).

Figure 10.18 Fraser fir killed by balsam woolly adelgid, together with regenerating young trees not yet old enough to be susceptible.

Figure 10.19 Lantana invading pastures and native forest understory in southeast Queensland, Australia.

Figure 10.20 Lantana defoliated by the tree hopper

Aconophora compressa

in southeast Queensland, Australia and close up of insect.

Figure 10.21 Adults of emerald ash borer,

Agrilus planipennis

.

Figure 10.22 Dead ash, killed by emerald ash borer, lining Michigan roadside.

Figure 10.23 Emerald ash borer larvae feeding on ash cambium.

Figure 10.24 (a, b)

Tetrastichus planipennisi

adult and larvae inside host larva, an important parasitoid of emerald ash borer, introduced from China.

Figure 10.25

Oobius agrili

ovipositing in an emerald ash borer egg; this is another important parasitoid introduced from China.

Figure 10.26 Increase in rates of parasitism of emerald ash borer larvae by the introduced parasitoid

Tetrastichus planipennisi

in Michigan, following its introduction in 2008.

Figure 10.27 Parasitism by

Tetrastichus planipennisi

is limited by bark thickness relative to ovipositor length, such that only smaller trees are protected.

Figure 10.28 Miconia leaf, showing injury owing to infestation by the fungal pathogen

Colletotrichum gloeosporioides

[Penz] Sacc. f. sp.

miconiae

at Nuku Hiva, Tahiti, 2007.

Figure 10.29 Regrowth of native vegetation, including the rare endemic shrub

Psychotria

(Rubiaceae), in the understory of a miconia invaded forest in Tahiti, with 20% leaf damage on canopy miconia leaves caused by the introduced fungal pathogen

Colletotrichum gloeosporioides

[Penz] Sacc. f. sp.

miconiae

.

Chapter 12

Figure 12.1 Strawberry guava,

Psidium cattleianum

, thickets crowd out native forest species in Hawaii.

Figure 12.2

Tectococcus ovatus

scales make galls on young leaves of strawberry guava.

Figure 12.3 Relative frequency of public comments on 2010 Draft Environmental Assessment of proposed biological control of strawberry guava (State of Hawaii, 2011b).

Chapter 13

Figure 13.1 Distribution by year of citations of five major articles in the biological control controversy literature.

Guide

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Integrating Biological Control into Conservation Practice

EDITED BY

Roy G. Van Driesche

Department of Environmental Conservation, University of Massachusetts, USA

Daniel Simberloff

Department of Ecology & Evolutionary Biology, University of Tennessee, USA

Bernd Blossey

Department of Natural Resources, Cornell University, USA

Charlotte Causton

Charles Darwin Foundation, Galápagos, Ecuador

Mark S. Hoddle

Department of Entomology, University of California, USA

David L. Wagner

Department of Ecology & Evolutionary Biology, University of Connecticut, USA

Christian O. Marks

The Nature Conservancy, Connecticut River Program, USA

Kevin M. Heinz

Department of Entomology, Texas A & M University, USA

Keith D. Warner

Center for Science, Technology, and Society, Santa Clara University, USA

This edition first published 2016 © 2016 by John Wiley & Sons, Ltd.

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List of contributors

Andersen, Jeremy CDepartment of Environmental Science, Policy, & Management,University of California, Berkeley, USA, [email protected]

Blossey, BerndDepartment of Natural Resources, Cornell University, Ithaca,New York, USA, [email protected]

Causton, CharlotteCharles Darwin Foundation, Puerta Ayora, Santa Cruz,Galápagos Islands, Ecuador, [email protected]

Center, Ted DUSDA ARS Invasive Species Laboratory (retired),Ft. Lauderdale, Florida, USA, [email protected]

Duan, Jian JUSDA ARS Beneficial Insects Introduction Research Unit,Newark, Delaware, USA, [email protected]

Fowler, SimonLandcare Research, Manaaki Whenua, New Zealand,[email protected]

Heinz, Kevin MDepartment of Entomology, Texas A & M University,College Station, TX, USA, [email protected]

Hoddle, Mark SDepartment of Entomology, University of California,Riverside, California, USA, [email protected]

Johnson, M. TracyUSDA Forest Service, Pacific Southwest Research Station,Institute of Pacific Islands Forestry, Volcano, Hawaii, USA,[email protected]

Kaufman, LeylaDepartment of Entomology, University of Hawaii, Manoa,Hawaii, USA, [email protected]

Marks, Christian OConnecticut River Program, The Nature Conservancy,Northampton, MA, USA, [email protected]

Messing, Russell HDepartment of Entomology, University of Hawaii, Manoa,Hawaii, USA, [email protected]

Meyer, Jean-YvesDélégation à la Recherche, Tahiti, French Polynesia,[email protected]

Montgomery, Michael ENorthern Research Station, USDA Forest Service (retired),Hamden, Connecticut, USA, [email protected]

Pratt, Paul DUSDA ARS Invasive Species Laboratory, Ft. Lauderdale,Florida, USA, [email protected]

Purcell, MaryUSDA ARS Invasive Species Laboratory, Ft. Lauderdale,Florida, USA, [email protected]

Rayamajhi, Min BUSDA ARS Invasive Species Laboratory, Ft. Lauderdale,Florida, USA, [email protected]

Sheppard, Andy WCommonwealth Scientific and Industrial ResearchOrganisation (CSIRO), ACT, Australia, [email protected]

Simberloff, DanielDepartment of Ecology & Evolutionary Biology,University of Tennessee, Knoxville, TN, USA, [email protected]

Tipping, Phil WUSDA ARS Invasive Species Laboratory, Ft. Lauderdale,Florida, USA, [email protected]

Van Driesche, Roy GDepartment of Environmental Conservation,University of Massachusetts,Amherst, MA, USA, [email protected]

van Klinken, RieksCommonwealth Scientific and Industrial ResearchOrganisation (CSIRO), Brisbane, Queensland, Australia,[email protected]

Wagner, David LDepartment of Ecology & Evolutionary Biology,University of Connecticut, Storrs, Connecticut, USA,[email protected]

Warner, Keith DCenter for Science, Technology, and Society,Santa Clara University, California, USA,[email protected]

Preface

The magnitude of threat posed to native ecosystem function and biodiversity by some invasive vertebrates, insects, pathogens, and plants is enormous and growing. At the landscape level, after damaging invaders are beyond eradication, a variety of habitats and ecosystems, on islands and continents, in all parts of the world may be affected and require some form of restoration. Biological control offers substantial opportunity to reduce the damage from invasive insects and plants, two of the most frequent and damaging groups of invasive species.

The purpose of this book is to address a nearly 25-year-old rift (from the seminal article by Howarth [1991]) that opened between conservation/restoration biologists and biological control scientists, particularly in the United States, so that in the future conservation biologists and biological control scientists might work together better to restore native ecosystems damaged by invasive species. The planning for this book originated in an informal meeting of conservation biologists, invasion biologists, and biological control scientists in October 2009, in Sunapee, New Hampshire, following a meeting that year on biological control for the protection of natural areas, held in Northampton, Massachusetts.

The tension between biological control and conservation biology had two causes. The first was that by the 1960s biological control agents introduced earlier to protect grazing or agricultural interests were found attacking native plants and insects in natural areas. More extensive search found other cases of such non-target impacts (Johnson and Stiling, 1996; Louda et al., 1997; Strong, 1997; Boettner et al., 2000; Kuris, 2003), tarnishing the use of biological control for a generation of conservation biologists and restoration ecologists. Any discussion of potential use of biological control agent to mitigate pest problems prompted the question: “What will it eat next if it controls the target?” This question is today routinely asked by undergraduates, graduate students, and the general public, but fails to recognize the dietary restrictions of many biological control agents. Mechanisms of population dynamics exist that cause insects with specialized diets, unlike vertebrates, to lose host-finding efficiency when the density of their prey or host plant declines, resulting in lower realized fecundity and a decrease in population size. Therefore, for specialized biological control agents, the answer to “what will they eat next” is “the same, just less of it as it becomes harder to find.” Others were concerned that agents would attack non-target species due to evolutionary expansion of their host ranges. However, while host shifts do frequently occur over evolutionary time (Stireman, 2005; Barrett and Heil, 2012), such changes have rarely been documented among insects introduced for biological control.

The second reason for the lack of understanding that developed between biological control and conservation/restoration scientists was research compartmentalization, with each group defining itself into its own sub-disciplines, attending different meetings and publishing in different journals. This is true both for conservation/restoration biologists (who publish in Conservation Biology, Restoration Ecology, Biological Invasions, etc.) and biological control scientists (BioControl, Biological Control, Biological Control Science and Technology, etc.). Opportunities to talk at length between these groups were, therefore, rare.

If invasive species were not one of the most important drivers of ecological degradation across natural ecosystems, the status quo could continue indefinitely. But they are and we must confront them as efficiently as possible. Conservation biologists should no longer leave a good tool unused and biological control scientists should no longer work in isolation from conservation biologists with special knowledge of the invaded ecosystems. The goal of this book is to discuss these issues in ways that make sense to both groups and find ways to work together better.

References

Barrett, L. G. and M. Heil. 2012. Unifying concepts and mechanisms in the specificity of plant-enemy interactions.

Trends in Plant Science

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Boettner, G. H., J. S. Elkinton, and C. J. Boettner. 2000. Effects of a biological control introduction on three nontarget native species of saturniid moths.

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Howarth, F. G. 1991. Environmental impacts of classical biological control.

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Johnson, D. M. and P. D. Stiling. 1996. Host specificity of

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Kuris, A. M. 2003. Did biological control cause extinction of the coconut moth,

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Louda, S. M., D. Kendall, J. Connor, and D. Simberloff. 1997. Ecological effects of an insect introduced for biological control of weeds.

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CHAPTER 1Integrating biological control into a conservation context: why is it necessary?

Kevin M. Heinz1, Roy G. Van Driesche2, and Daniel Simberloff3

1Department of Entomology, Texas A & M University, USA

2Department of Environmental Conservation, University of Massachusetts, USA

3Department of Ecology & Evolutionary Biology, University of Tennessee, USA

Potential problems if integration is lacking

The basic argument of this book is that, for pests of wildlands1, biological control should be one of the tools considered for use. Not to do so would lead to inadequate restoration for many pests because, while they might be controlled in small areas, they would remain uncontrolled over much of the landscape. We further argue that biological control will be done better if integrated into conservation biology because that will force greater consideration of the role of the invader as the true source, or not, of ecosystem degradation (see Chapter 2) and would incorporate into the control program more detailed knowledge of the invaded community’s ecology, which may exist best within the conservation biology community. Finally, we argue that biological control in areas of conservation importance can be done safely with modern methods of evaluation for assessing pest impact and natural enemy host range.

When conservation biologists seek to restore natural communities damaged by invasive species, if they give no thought to biological control, their efforts may be far less successful. Without biological control in the mix of potential tools, restoration efforts move toward eradication if possible, suppression over large areas by changing processes (e.g., fire, flood, or grazing regimes) at the landscape level if relevant, or suppressing the invader on small patches with chemical or mechanical tools if these methods work and money can be found for long-term management. Many invaders, however, cannot be eradicated if they are widespread, or their biology may not be appropriate to control over the long term with pesticides or mechanical tools. Similarly, while some plants or insects may have become highly invasive because people have altered historical landscape processes (MacDougall and Turkington, 2005), this factor surely does not account for the damage caused by some invaders. Certainly, it applies to few if any invasive insects: virtually none of the invasive insects that have so damaged North American forests (Campbell and Schlarbaum, 1994; Van Driesche and Reardon, 2014) could be said to have such factors driving their destructive effects. In contrast, some invasive plants quite likely are augmented in their densities by such forces, but clearly not all are. This leaves many highly damaging insects and plants for which restoration of ecological processes toward historical norms will not lead to restoration of the ecosystem. In such cases, then, restoration efforts are limited to saving fragments through intensive efforts at the preserve rather than the landscape level. While these efforts may protect rare species with small, threatened ranges, they do nothing to preserve average habitat conditions for the bulk of species across the broader landscape. Working with biological control scientists can sometimes provide a solution that can safely (if well conceived and executed) protect the landscape rather than just a few isolated preserves.

To succeed at biological control is not easy and requires cross-disciplinary collaborations to understand fully the implications of releasing natural enemies of the invader. If such collaborations with conservation biologists are lacking, decisions may be taken that undervalue certain native species, miss important ways in which these species are interacting, or fail to consider fully the potential impacts of the introduced biological control agents on the native ecosystem or what other forces may be at work driving ecosystem change. If biological control scientists work within a broader restoration team that includes conservation biologists, these potential pitfalls are more likely to be recognized and avoided.

Carrying out a biological control program typically requires a commitment to travel to the invader’s native range and determine what natural enemies affect the invader’s population dynamics there and which of these are plausibly sufficiently specialized that they might be safe for release in the invaded region. These demands require training in natural enemy biology and population dynamics, as well as knowledge of foreign cultures and geography. If the targeted invader is a plant, the biological control scientist must also have extensive understanding of plant taxonomy, physiology, and how both biotic and abiotic factors affect plant demography. If the invader is an insect, the practitioner must also be familiar with the taxonomy and biology of parasitoids or predators, how to rear them, and how they overcome host defenses. Training in these diverse subjects may leave little time to develop a deep appreciation for the community ecology and details of the particular ecosystems invaded by the pest. This leaves the biological control scientist vulnerable to making decisions that fail to take such information fully into account, and hence underscores the value of collaborative projects within a conservation biology framework, working with specialists on the ecology of the invaded communities.

Book organization

The practices of biological control and ecological restoration can be viewed as large-scale field experiments that unintentionally test many fundamental principles in ecology, as noted previously for both biological control (e.g., Hawkins and Cornell, 1999; Wajnberg et al., 2001; Roderick et al., 2012) and species conservation and habitat restoration (e.g., Young, 2000; Groom et al., 2005). Several issues need addressing when one attempts to integrate biological control of pests of wildlands into the larger framework of conservation biology. In the chapters that follow, experts illustrate some of the problems that can arise when such integration is lacking and provide insights for avoiding problems that may affect the management program or conservation interests.

In Chapter 2, readers are presented with a conceptual framework for confirming whether an invasive species is the primary cause of environmental change and for deciding how to minimize its impacts, potentially as part of a larger package of restoration activities. Approaches potentially able to generate the desired outcomes are discussed and illustrated with the example of conservation threats to floodplain forests in New England. Chapter 3 subsequently addresses the means (tools) available to control invasive species. Depending on circumstances, control goals may be eradication, human-sustained invader suppression with periodic mechanical or chemical control plus monitoring, or permanent area-wide invader suppression through alteration of ecosystem processes or programs of biological control. Once goals are set, a variety of tools may be relevant and are discussed (mechanical, chemical, biological, combinations) in terms of the system or pest attributes that affect efficacy, control cost, and effects on the environment. Chapter 4 examines tradeoffs among risks posed by major control methods using case histories of particular projects. Chapter 5 continues this discussion through an examination of how the risks and benefits of biological control projects against wildland pests can best be recognized and compared, through the planned interaction of biological control scientists and conservation biologists. At the end of these chapters, readers should have a better understanding of when biological control may be the right or wrong option.

The next block of chapters shifts to the practice of biological control within the context of environmental restoration projects. Chapter 6 discusses the importance of systematics and accurate taxonomic identification, both of pests and natural enemies, for biological control programs. The discussion includes recent developments in molecular techniques applicable to modern biological control programs. Chapter 7 addresses our ability to forecast unwanted impacts of biological control, describing the nature of the concern, reviewing the historical record, and ending with a discussion of unresolved issues. Chapters 8 and 9 discuss how to measure and evaluate outcomes of biological control projects. Because biological control is costly in terms of financial and human resources, there is an increasing demand for accountability as to efficacy when biological control is used to restore or protect native ecosystems or species. Addressed directly in these chapters are the difficult tasks associated with delineating the damaged system’s starting conditions and measuring the progress toward achieving restoration goals. Chapter 8 takes a broad conceptual view of the task, while Chapter 9 reviews techniques used for such assessments and their limits and requirements for application. Chapter 10 discusses a series of biological control projects conducted in wildland ecosystems. These cases provide concrete examples of the kinds of damage that can be corrected with biological control, and the discussions of project details highlight the variety of issues that can affect such work.

Concluding chapters address societal and economic matters. Chapter 11 discusses laws and regulations that affect biological control. The evolution of regulations and regulatory agencies from several parts of the world are reviewed, which provides the context for recommendations for improvements in biological control regulations. Chapter 12 describes how conflicts among groups may arise during a biological control project. The focus of the chapter is on methods for setting goals and resolving disagreements that are either initially present or arise during the conduct of the project. Chapter 13 discusses ethical principles related to the introduction of non-native species, focusing on processes and goals that can help resolve disagreements among parties in conflict. In Chapter 14, we discuss economic issues associated with species invasions and their biological control in wildlands. Chapter 15 describes steps to reform the practice of biological control and integrate its use against pests of wildlands into a conservation framework. It also makes recommendations for changes needed to make biological control of agricultural and ornamental pests at least environmentally neutral.

We end by returning to the central message of the book, looking to the future and describing activities likely to further the integration between biological control activities and those of conservation biologists and restoration ecologists.

Acknowledgments

We thank Bernd Blossey, Charlotte Causton, and David Wagner for reviewing Chapter 1.

References

Campbell, F. and S. E. Schlarbaum. 1994.

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Groom, M. J., G. K. Meffe, and C. R. Carroll. 2005.

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, 3rd edn. Sinauer Associates. Amherst, Massachusetts, USA.

Hawkins, B. A. and H. V. Cornell. 1999.

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MacDougall, A. S. and R. Turkington. 2005. Are invasive species the drivers or passengers of change in degraded ecosystems?

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Note

1

For purposes of this book, the term “wildlands” does not equal wilderness nor does the term “natural” mean “pristine.” Rather the term wildlands is taken to mean places, both land and water, that are not intensively managed.

CHAPTER 2Designing restoration programs based on understanding the drivers of ecological change

Christian O. Marks1 and Roy G. Van Driesche2

1Connecticut River Program, The Nature Conservancy, USA

2Department of Environmental Conservation, University of Massachusetts, USA

Overview of concepts

Introduction

The activities of conservation planning and biological control of invasive species are both continuing to evolve, requiring greater collaboration between these disciplines to achieve mutual goals pertaining to invasive species management (Chapter 1). Invasive species can be a factor contributing to ecological degradation (Simberloff, 2011; Kumschick et al., 2015). Even reserves in relatively intact ecosystems in remote regions can be threatened by exotic species invasions. Often this impact is not recognized until after the invasive species has become too abundant and widespread for eradication or even containment (e.g., Herms and McCullough, 2014). Long term, such pervasive invader populations are usually prohibitively expensive to suppress using conventional chemical and mechanical methods, especially as the infested area increases to tens or hundreds of thousands of hectares. Development of an effective biological control program is a potential alternative for managing an invasive pest, but biological control frequently must be integrated into the broader conservation plans of the local ecosystem because invasive species, particularly invasive plants, are rarely the only factor contributing to ecological degradation, as we will illustrate. Even where an invasive species is the leading cause of ecological degradation, its control alone may not accomplish restoration goals, and additional measures may be necessary (Chapter 3). Moreover, funding for conservation is limited, necessitating a strategic approach and a clear vision of what the intended end goal will be for the restoration.

In this chapter, we briefly review the conservation planning process, focusing on the roles invasive species play in ecological change. We pay particular attention to how to determine if an invasive species rises to the level of threat that warrants development of a biological control program, which we illustrate with a representative case study – the restoration of Connecticut River floodplain forests in the northeastern United States. A lack of integration into a wider restoration planning process has sometimes resulted in criticism of past biological control programs. For example, biological control of purple loosestrife (Lythrum salicaria L.) is one of the most widespread biological control programs for weeds in North America (Wilson et al., 2009), yet the necessity of controlling this invader has been questioned by some ecologists (Anderson, 1995) – although some of these concerns have since been rebutted (Blossey et al., 2001). More notably, in another case, a lack of integration of the biological control of saltcedar (Tamarix species) into a wider plan for the ecological restoration of riparian communities in the southwestern United States has resulted in controversy among various interest groups (see Chapter 4; or Dudley and Bean, 2012). Saltcedar is a widespread invader of riparian areas along southwestern rivers with well-known, large negative ecological impacts, but on some rivers it has also become one of the few remaining riparian tree species (Tracy and DeLoach, 1999; Sher and Quigley, 2013). The release of a highly effective biological control agent for saltcedar, without also taking action to increase recruitment of native floodplain tree species like willows (Salix) and cottonwoods (Populus), may have resulted in a loss of some marginal nesting habitat for the federally listed endangered southwestern willow flycatcher (Empidonax traillii extimus Phillips) (Finch et al., 2002; Smith and Finch, 2014). On some southwestern rivers, modifying operations at dams to restore a more natural flood regime downstream, alone or in combination with saltcedar biological control, may be more effective at restoring floodplain function, including natural recruitment of the native riparian trees that the flycatcher prefers for nesting (Cooper et al., 2003; Richard and Julien, 2003; Shafroth et al., 2005; Ahlers and Moore, 2009; Hultine et al., 2009; Merritt and Poff, 2010; Dudley and Bean, 2012). These examples show how important it is to evaluate the factors that are influencing ecosystem function and degradation before irreversible actions are taken. The mere high dominance by an invasive species is not necessarily equivalent to degradation of ecological function. Therefore, it is necessary to rank invasive species not just against each other for control priority, but also to rank their control against other conservation actions that may have a greater positive impact. It is critical to think holistically about how the system functions before designing a plan of action.

Ecological restoration planning process

The motivations for carrying out ecological restoration are diverse and depend on the stakeholders’ values. These motivations can include anything from landscape aesthetics and protection of endangered species to conservation of ecosystem services. The first step in the planning process is to achieve a consensus among stakeholders on what aspects of the ecosystem are valued, as well as what outcomes are desired for the restoration activity. This goal-setting process is subjective, and it is important to achieve a consensus among stakeholders early to avoid conflicts later, when program momentum may be significant, making change difficult or costly (Chapter 12). Next, one needs to understand the threats that have led to past declines in the aspects of the ecosystem where restoration is desired. Specifically, one needs to develop an understanding of system change with the best science available at the time, being aware that our knowledge of the system is usually incomplete. Consequently, it is important to be explicit about one’s assumptions of what is driving change in the system because they could be incorrect (Wilkinson et al., 2005), and scientists should seek to test such assumptions to guide restoration in an adaptive management framework (Westgate et al., 2013).

Invasive species and system change

High abundance of invasive species in wildlands is often associated with dramatic ecosystem alterations, such as eutrophication of soil or water bodies (Green and Galatowitsch, 2002; Perry et al., 2004; Silliman and Bertness, 2004; Kercher et al., 2007), overgrazing (Knight et al., 2009; HilleRisLambers et al., 2010; Dornbush and Hahn, 2013), and altered disturbance regimes such as fire and flooding (Cooper et al., 2003; Katz and Shafroth, 2003; Keeley, 2006; MacDougall and Turkington, 2007; Stromberg et al., 2007; Merritt and Poff, 2010; Metz et al., 2013; Greet et al., 2013; Schmiedel and Tackenberg, 2013; Terwei et al., 2013; Reynolds et al., 2014). However, it is not always immediately obvious to what degree non-native species invasions are the cause or the consequence of the ecological change, or both. Determining the answer to this question is crucial to deciding if the most effective strategy is more likely to be restoring the physical environment and key ecological processes or starting a biological control program, or if both may be necessary.

MacDougal and Turkington (2005) defined invasive species that thrive on ecological change, such as altered ecosystem properties or a shift in disturbance regimes, as passengers (see Figure 2.1). Owing to their high density in degraded ecosystems, passengers appear more damaging than they actually are. If the ecosystem stressor that has allowed the passenger to proliferate is removed, one would expect passenger populations to decline. MacDougal and Turkington (2007) argued, for example, that the Poa pratensis L. invasion of Garry oak (Quercus garryana Douglas ex Hook.) savannas in British Columbia was a consequence of fire suppression. The failure of native vegetation to respond to Poa removal indicated that Poa was not the cause of change, only associated with it. Follow-up experiments found that restoration of fire to these ecosystems reduced invader abundance and promoted native species’ recovery (MacDougall and Turkington, 2007).

Figure 2.1 A chart to classify the ecological role of an invasive species on the spectrum from invasion being a consequence of ecological change to invasion being the cause of ecological change.

Exceptions to the autogenous recovery of native populations following removal of the ecosystem stressor include situations where there are strong feedbacks between biotic factors and the physical environment (Suding et al., 2004). Specifically, once an invasive species is dominant, it might change the environment in ways that would favor its continued dominance even after the factor promoting its initial establishment was removed. For example, marsh disturbances such as ditching create microsites with better soil aeration where invasive common reed (Phragmites australis [Cav.] Trin. ex Steud.) can establish (Bart and Hartman, 2003; Chambers et al., 2003; Lathrop et al., 2003; Silliman and Bertness, 2004). Once established, Phragmites can transfer air within a clone via its hollow stalks, enabling it to spread to the rest of the marsh, forming large monospecific patches (Bart and Hartman, 2000; Lathrop et al., 2003). In another example, native deer herbivory was shown to accelerate forest invasion of garlic mustard (Alliaria petiolata [M. Bieb.] Cavara & Grande), Japanese barberry (Berberis thunbergii DC), and Japanese stiltgrass (Microstegium vimineum [Trin.] A. Camus), but was not as important as canopy disturbance or propagule pressure in explaining different levels of invasive weed abundance (Eschtruth and Battles, 2009). Once these invasive, non-native forest understory plants became abundant, propagule pressure would remain high even if canopy disturbance and deer herbivory were reduced. In such cases, restoration success would require both reducing the ecosystem stressor that had led to ecological degradation and suppressing the invasive species to reduce propagule pressure. Similarly, native plant propagules may be too scarce for native plants to recolonize on their own even after deer and invasive plant populations have been reduced, thus necessitating native plant seed addition or planting (Tanentzap et al., 2009, 2011, 2013; Collard et al., 2010; Royo et al., 2010; Dornbush and Hahn, 2013). Holistic restoration approaches are especially important in urban and suburban areas, where there are usually multiple interacting stressors including invasive plants (Sauer, 1998).

In contrast to ecological passengers, MacDougal and Turkington (2005) defined drivers as invasive species that are both able to proliferate unaided by external ecological change and cause considerable damage. An example of an invasive driver is the fungal pathogen Cryphonectria parasitica (Murrill) Barr, the causal agent of chestnut blight. This fungus was accidentally introduced from Asia into North America, where it killed virtually all mature American chestnut (Castanea dentata [Marshall] Borkh.), the tree that once dominated many eastern North American forests (Braun, 1950). Attempts at biological control of the chestnut blight fungal pathogen with viruses were successful in Europe but not in eastern North America (Anagnostakis, 2001; Milgroom and Cortesi, 2004). Current efforts at restoring American chestnut are instead focused on breeding blight-resistant hybrids (Jacobs, 2007; Anagnostakis, 2012). Other examples of pure drivers of ecological change are the cottony cushion scale (Icerya purchasi Maskell), a phloem-sucking insect that caused many native plant populations in the Galápagos Islands to decline (Chapter 10), and laurel wilt, a disease caused by an invasive fungus vectored by the non-native redbay ambrosia beetle (Xyleborus glabratus Eichhoff), which is causing extensive mortality of redbay (Persea borbonia [L.] Spreng.) in the southeastern United States (Spiegel and Leege, 2013). Clearly, drivers are the most threatening invasive species and thus should receive a high priority on lists of candidate invaders for developing control programs.

Although originally set up as a dichotomy, the distinction between drivers and passengers is more accurately thought of as a spectrum, with many invasive species being intermediate cases where their proliferation has benefited from wider ecosystem change, but their high abundance also affects the ecosystem. Bauer (2012) has called these intermediate cases back-seat drivers, and his review suggests that most invasive plant species are back-seat drivers. Berman et al. (2013) proposed that invasive non-native ants in New Caledonia are back-seat drivers whose initial invasion is associated with disturbance, such as forest clearing, but which subsequently also harms native ant communities. Similarly, experimental manipulations have shown that invasion by the red imported fire ant (Solenopsis invicta Buren.) in the southeastern United States is driven by disturbance (King and Tschinkel, 2008). Many studies have documented large impacts by non-native fire ants on native ants and other native arthropods through competition and predation (Porter and Savignano, 1990; Gotelli and Arnett, 2000; Wojcik et al., 2001; Sanders et al., 2003). Decapitating flies in the genus Pseudacteon (e.g., P. tricuspis Borgmeier) were imported from Argentina and released as biological control agents of the red imported fire ant because the type of disturbance that promotes fire ant invasion has become unavoidable in much of the landscape, resulting in substantial damage to crops, livestock, human health, electrical equipment, and wildlife (Porter et al., 2004). Thus, where system changes that have enabled invasion by a back-seat driver are irreversible, there may be a sufficiently compelling argument for developing a biological control program.

Finally, there are non-native species whose establishment is not associated with significant ecological change either as a cause or consequence. We have labeled these species as pedestrians in Figure 2.1 to highlight the difference in pace of change. It is important to remember that the categories in Figure 2.1 are not immutable; many of today’s invasive driver species were pedestrians receiving little notice during the first century of colonization in their new range (Kowarik et al., 1995; Crooks, 2005). With the right ecological or evolutionary changes, species can quickly switch between these categories. Moreover, local context matters; an invasive species that acts like a back-seat driver or passenger in one area may act like a driver in another part of its invaded range or in a different habitat (Wilson and Pinno, 2013). Therefore, in cases where there are no obvious large impacts by an invader in a particular ecosystem, further study elsewhere may be necessary to make a well-informed assessment of their overall impact in the invaded range.

Ranking invasive species for classical biological control

Central to ranking ecological threats for remediation is a consensus on what level of impact is sufficient to require conservation action. For example, The Nature Conservancy’s conservation planning process ranks threats (both biotic and abiotic) according to scope, severity, and irreversibility (also referred to as permanence). With respect to an invasive species, scope could be the area or percentage of a habitat likely to become threatened by the invader over the coming decade. Severity could be thought of as the level of damage to native biota in the invaded area that can reasonably be expected from the threat given the continuation of current circumstances and trends. Severity is the seriousness of the impact. For example, an insect pest invasion that causes high mortality of its tree host would be considered a more severe threat than one that only reduced the tree’s growth rate. Irreversibility (or permanence) is the degree to which the effects of a threat cannot be reversed by restoration. For instance, the effects of the most damaging non-native species, once they become widespread, are difficult to reverse. Therefore preventing invaders from establishing, through early detection and elimination of incipient populations, generally receives high priority in conservation planning.