SupPreSSing Over-abundant invaSive plantS and

inSectS in natural areaS

by uSe Of their Specialized natural enemieS

Edited by Roy G. Van Driesche and Richard C. Reardon

United States Department of Agriculture

Forest Health Technology

Enterprise Team

FHTET-2017-02

April 2017
a

b

c d

e

f g

h i j

Cover Photos: (a) Rainbow falls, Hawaii (Credit: Kyle Hawton); (b) healthy brood of 

vegetarian ground finches, Platyspiza crassirostris (Credit: George Heimpel); (c) Philornis 

downsi, an invasive parasitic fly (Credit: David L. Hansen, University of Minnesota); (d) 

dead vegetarian ground finch chick killed by P. downsi (Credit: Jody O’Connor and Sonia 

Kleindorfer); (e) Conura annulifera, a parasitoid of P. downsi (Credit: David L. Hansen, 

University of Minnesota); (f) the melaleuca weevil, Oxyops vitiosa (Credit: Ellen C. Lake, 

USDA ARS); (g) O. vitiosa adult and feeding scars on melaleuca foliage (Credit: Willey 

Durden, USDA ARS, bugwood.org); (h) the melaleuca psyllid, Boreioglycaspis melaleucae

(Credit: Paul D. Pratt, USDA ARS); (i) waxy flocculence produced by B. melaleucae 

nymphs on melaleuca foliage (Credit: Susan A. Wright, USDA ARS); (j) melaleuca 

infestation (Credit: Randy Westbrooks, bugwood.org)

The Forest Health Technology Enterprise Team (FHTET) was created in 1995 by the Deputy Chief for State 

and Private Forestry, USDA, Forest Service, to develop and deliver technologies to protect and improve the 

health of American forests. This book was published by FHTET as part of the technology transfer series.

http://www.fs.fed.us/foresthealth/technology/

For additional copies of this publication, contact: 

Richard Reardon FHTET, USDA Forest Service, 180 Canfield Street, Morgantown, WV 26505

304.285.1566; rreardon@fs.fed.us

This publication is available online at: 

https://www.fs.fed.us/foresthealth/technology/pdfs/FHTET-2017-02_Biocontrol_Natural_Areas.pdf

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of any product or service to the exclusion of others that may be suitable.

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Federal Recycling Program

Printed on Recycled Paper
Suppressing Over-abundant

Invasive Plants and Insects in

Natural Areas by Use of Their

Specialized Natural Enemies

Edited by 

Roy G. Van Driesche 

Department of Environmental Conservation, University of Massachusetts, Amherst

and 

Richard C. Reardon

USDA Forest Service, Morgantown, West Virginia

Acknowledgments

We thank the authors of the individual chapters for their expertise in reviewing and summarizing the 

literature and providing current information on biological control in natural areas. The layout of this guide 

was designed by Rachel Winston (MIA Consulting). We extend our gratitude to all of the photographers 

who granted permission for the use of photos. We would also like to thank the U. S. Department of 

Agriculture Forest Service for technical and financial support and Forest Health Technology Enterprise 

Team for providing funding for the preparation and printing of this publication. 

FHTET-2017-02

April 2017
Biological Control in Natural Areas

TABLE OF CONTENTS

Introduction........................................................................................................................ 1

Chapter 1. History of Biocontrol in Natural Areas and Issues Affecting Its Use

R. G. Van Driesche .........................................................................................................................................2

Chapter 2. Improving Health of Native Galápagos Plants by Introducing a Specialized Predator

of the Invasive Cottony Cushion Scale

C. E. Causton1, C. Calderón Alvaréz1, C. D. Hoddle2, M. S. Hoddle2, M. P. Lincango1, T. G. A. Poulsom1, and 

R. G. Van Driesche3 ......................................................................................................................................16

Chapter 3. Potential for Biological Control of a Parasite of Darwin’s Finches

R. A. Boulton and G. E. Heimpel ................................................................................................................23

Chapter 4. Restoring Gumwood Forests on St Helena: The Contribution of Biocontrol

of Orthezia Scale

S. V. Fowler ..................................................................................................................................................29

Chapter 5. Protection of Wiliwili Trees, the Foundation of Hawaiian Dry Forests, through

Suppression of an Invasive Gall Wasp

L. V. Kaufman1 and J. Yalemar2 .....................................................................................................................36

Chapter 6. Can Ash Communities and Their Dependent Species be Partially Protected through

Biological Control of Emerald Ash Borer?

J. J. Duan......................................................................................................................................................41

Chapter 7. Biological Control of Waterhyacinth Restores Ecosystem Functioning to an

Impoundment in South Africa

M. P. Hill1 and J. A. Coetzee2 ........................................................................................................................48

Chapter 8. The Role of Melaleuca Control in Everglades Restoration: Accomplishments

and Future Plans

E. C. Lake1, P. W. Tipping1, M. B. Rayamajhi1, P. D. Pratt2, F. A. Dray Jr.1, G. S. Wheeler1, M. F. Purcell3, and 

T. D. Center1 ...............................................................................................................................................53

Chapter 9. Partial Restoration of Native Rainforests in the Island ofTahiti (French Polynesia) 

after Introduction of a Fungal Pathogen to Control the Invasive Tree Miconia calvescens

J.-Y. Meyer....................................................................................................................................................59

Chapter 10. Strategic Restoration of Saltcedar-affected Riparian Ecosystems of the U.S. 

Southwest: Integration of Biocontrol and Ecohydrological Conditions in Restoration

Planning

T. L. Dudley1, D. W. Bean2 and C. J. DeLoach3 ............................................................................................64

Chapter 11. Future Trends in Use of Natural-Area Biocontrol

R. G. Van Driesche .......................................................................................................................................74

References .......................................................................................................................... 77
Biological Control in Natural Areas

INTRODUCTION

Summary. Biological control has long been used against invasive plants and, less often, invasive insects, 

in forests, aquatic systems, grasslands, coastal areas, riparian areas, islands, and other biomes or habitats of 

conservation concern. Historically, motivation for such efforts has shifted gradually away from economic 

interests toward ecological protection for its own sake. Future projects of this type will require better 

integration of biological control into conservation practice, a better understanding by societies of the 

reasons for such work and its possible risks and benefits, as well as continued, consistent public funding. 

Examples of successful projects in high-value conservations areas (Everglades, Galápagos, oceanic islands, 

wetlands, forests) are discussed, and future challenges and caveats mentioned.

Keywords biological control, natural areas, conservation, parasitoids, historical trends

Introduction 1
Biological Control in Natural Areas

CHAPTER 1. HISTORY OF BIOCONTROL IN NATURAL 

AREAS AND ISSUES AFFECTING ITS USE

R. G. Van Driesche

Department of Environmental Conservation, University of Massachusetts, Amherst

Like crops, plant communities in natural areas are increasingly subject to pressures from invasive insects 

and plants that directly attack native species or displace them through competition for resources (Foxcroft 

et al., 2013). While some of these problems can be relieved at the local level through traditional methods 

for control of invasive species – cutting, pulling, burning, flooding, use of herbicides, or use of poison 

baits – the limitations of such approaches are becoming increasingly obvious: (1) often they do not work, 

(2) many species cannot be eradicated and so treatments, even if effective in the year of application, must 

be repeated, causing an unsustainable burden on management budgets, and (3) treatment is economically 

feasible only for small areas (hundreds of hectares), while many damaging invasive species require control 

at much larger scales over whole landscapes, which also increase risks to nontarget species.

Consequently, there is a need for more frequent consideration of biocontrol as an option for managing 

invasive species in natural areas (see Van Driesche et al., 2016a). To determine if biological control is an 

appropriate response to any given invasive species problem, the following points need to be considered: 

(1) Natural areas are often strongly affected at the landscape level by invasive plants and insects, and many 

species in these groups are suitable for suppression with biological control. (2) However, before considering 

biological control, it should be determined whether high densities of the invasive species are promoted by 

anthropogenic disruptions of habitat-shaping processes such as changes in fire or hydrological regimes, 

changes in grazing or nutrient levels in soils or water bodies (the “drivers vs. passengers” argument). If 

so, ecologists should determine if those influences can be reversed. Anthropogenic eutrophication of water 

bodies, for example, is commonly linked to proliferation of fast-growing invasive plants and reduction of 

nutrient input is important to improve the balance among native and invasive plants in such water bodies. 

Similarly, lack of large grazing animals may affect competitive balances between native and introduced 

plants in grasslands. (3) The physical scale over which control of the invasive species is desired should 

be determined. If control is only desired at the preserve level, then active controls (cutting, burning, use 

of herbicides, traps, poison baits, etc.) may suffice and biological control may not be needed. (4) For the 

residual group of invasive species that are not driven by disturbances to habitat-forming processes, that 

cannot be eradicated, and whose landscape-level control is desired, biocontrol should be considered.

2 Chapter 1: History of Biocontrol in Natural Areas and Issues Affecting Its Use
Biological Control in Natural Areas

A few biological control agents have strongly damaged populations of conservation concern (Stiling 

and Moon, 2001; Russell and Louda, 2004; Elkinton and Boettner, 2012), and frequent attention to these 

cases (Strong, 1997; Stiling, 2000; Louda et al., 2003) has conveyed the impression that risk is common 

and significant. This perception of high, general risk has lingered and has damaged relations between 

conservation biologists and biological control scientists in some parts of the world. However, reviews of 

the level of non-target impacts have shown that weed biocontrol is quite safe (Pemberton, 2000; Suckling 

and Sforza, 2014) and has become increasingly so over the last 30 years. The risk from insect biological 

control, while lower than generally assumed, is somewhat higher than for weed biocontrol (Van Driesche 

and Hoddle, 2017); recent efforts have been made to improve this situation (Van Driesche et al., 2016a). To 

promote such improvement, the following actions have been advocated:

• Step 1. Agree on good targets. The goal of the use of biological control in a conservation context is 

ecological restoration, not invasive species suppression per se. Suppression of the invasive species 

may be required, or not, depending on circumstances, and if suppression is required, biocontrol 

may or may not be the best approach. Such decisions need to be based on discussion among all 

concerned parties before initiation of projects against particular invaders.

• Step 2. Agree on which tools to use. If invasive species need to be controlled, all possible options, 

including biological control (or not) and its integration with any other activities as appropriate (a 

strategy sometimes called Integrated Ecosystem Pest Management) should be considered.

• Step 3. Build a common set of information by sharing experiences. To promote more frequent 

consideration of the use of biological control in natural areas, conservation biologists need 

opportunities to observe restoration of systems that incorporate biological control, through the 

sharing of information and visits to field sites. 

• Step 4. Plan projects holistically with active conservation partners. Once a decision has been taken 

to use biological control, a system is needed to holistically plan the project. Several questions will 

need to be answered. For example, is reduction of the pest alone enough to restore the ecosystem, 

or will other actions also be needed (before, during or after) to mitigate aspects of the biocontrol 

project or to enhance ecosystem recovery as the invasive species’ density decreases? 

• Step 5. Anticipate possible conflicts. Conflicts should be anticipated and consensus developed in 

support of desired actions. This may involve enlisting partners to undertake information-sharing or 

consensus-building activities. Additional time and effort will likely be needed to convince external 

audiences rather than just knowledgeable internal ones of the merits of the proposed projects or 

introductions. The active support of partners often will play a critical role in developing consensus.

HISTORICAL PERSPECTIVE

With the above as a rationale for the use of biological control in natural areas, here we summarize past 

experiences.

Chapter 1: History of Biocontrol in Natural Areas and Issues Affecting Its Use 3
Biological Control in Natural Areas

Previous Summaries 

While benefits from biological control projects have accrued to natural areas since the earliest days of 

biological control, particularly of weeds invasive in forests and grasslands (e.g., Dodd, 1940), projects were 

justified to societies in economic rather than ecological terms until about the 1990s. Chronologically, some 

key publications highlighting use of biological control in natural areas include the following:

In 2010, Van Driesche et al. reviewed previous benefits of biological control in natural areas and 

documented benefits to many biomes, including aquatic and wetland habitats (Buckingham, 2002; Landis 

et al., 2003; Ajuonu and Neuenschwander, 2003; Pratt et al., 2005; Denoth and Myers, 2005; Grevstad, 

2006; Rayamajhi et al., 2008; Coetzee et al., 2009; Heard and Paynter, 2009; Hill and McConnachie, 2009; 

Julien et al., 2009; Neuenschwander et al., 2009; Tipping et al., 2009; Boughton and Pemberton, 2009, 

2011 [see Smith et al., 2014]) for update of previous reference], boreal and temperate forests (Embree, 

1971; Embree and Otvos, 1984; Ryan, 1990; Fielding and Evans, 1997; Hough-Goldstein et al., 2009; 

see also Van Driesche and Reardon [2014] for an update on North American forest projects), subtropical 

and tropical forests (Barton et al., 2007), grasslands (Huffaker and Kennett, 1959; McEvoy et al., 1991; 

McCaffrey et al., 1995; Vogler and Lindsay, 2002; Tomley and Evans, 2004; Samuel et al., 2008), deserts 

and arid shrublands (Dodd, 1940; Zimmermann et al., 2009), coastal areas and dunes (Holtkamp, 2002; 

Edwards et al., 2009), and oceanic islands (Fowler, 2004; Grandgirard et al., 2008; Meyer et al., 2008a; 

Calderón Alvarez et al., 2012; Hoddle et al., 2013), with benefits occurring to biodiversity or ecosystem 

services. 

In 2012, a special issue of BioControl “Conserving Nature with Biological Control” published 

selected papers from a meeting on this topic held in Northampton, Massachusetts, 3-7 October, 2010. This 

issue highlighted the benefits of biological control to such world premier conservation sites as the South 

African fynbos (Moran and Hoffman, 2012), the Florida Everglades (Center et al., 2012), the Galápagos 

National Park (Calderón Alvarez et al., 2012), and the native forests of Tahiti (Meyer et al., 2012). Other 

papers discussed risks to natural areas or native species from introduced natural enemies (Simberloff, 2012, 

Elkinton and Boettner, 2012) and management of conflicts that arise in some cases (Dudley and Bean, 2012). 

In 2013 in the forward to Plant Invasions in Protected Areas (Foxcroft et al., 2013), Ian Macdonald 

stated “Who could have foreseen that the greatest long-term threat to the integrity of the ecosystems 

“protected” in these national parks, game reserves and state forests would turn out not to be the axeman, 

the hunter or the land developer, but rather the inexorable spread of alien species. But this is the situation 

that protected area managers throughout the world find themselves in today.” A chapter in this book (Van 

Driesche and Center, 2013) laid out the role for plant biological control in conserving such protected areas, 

including a table documenting the contributions to natural area protection of biocontrol projects against 71 

invasive plant species, worldwide. 

Following the 2010 meeting in Northampton, Massachusetts, on protecting natural areas through 

the use of biological control, a group of the meeting’s participants began collaboration on a book on this 

topic, eventually entitled Integrating Biological Control into Conservation Practice (Van Driesche et al., 

2016a), produced by an editorial group consisting of three biological control scientists (Van Driesche, 

Hoddle, and Heinz), five conservationists (Simberloff, Blossey, Causton, Marks, and Wagner), and one 

ethicist (Warner). This collaboration forced each discipline to consider the point of view of the other and 

engage in conversation until commonly agreed positions emerged. This book discusses issues relevant to 

biological control’s use in natural areas, including (1) design of restoration programs based on understanding 

of drivers of ecological degradation, (2) review of the tools available for invasive species suppression 

(mechanical, chemical, biological, etc) and how to manage tradeoffs among them, (4) risks and benefits of 

4 Chapter 1: History of Biocontrol in Natural Areas and Issues Affecting Its Use
Biological Control in Natural Areas

biological control and forecasting of risks, (5) use of modern systematics for agent and pest identification, 

(6) evaluation of natural enemy impacts on target pests, (7) case histories of biological control in natural 

areas written by original researchers carrying out the work, (8) regulation of biological control in various 

parts of the world, (9) the ethics and economics of use of biocontrol, and (10) proposals for future reforms 

of biological control practice. 

Critical to growth in the use of biological control in natural areas is better information on nontarget 

impacts of agents used in past biological control projects, so that the real risks and their likelihood can 

be more accurately judged by conservationists considering the issue before beginning biological control 

projects. Recently, this topic has been addressed for both weed (Suckling and Sforza, 2014) and arthropod 

targets (Van Driesche and Hoddle, 2017). Such summaries replace speculation or generalizations based 

on atypical cases. Suckling and Sforza, (2014) found the risk of use of non-target plants to be very low 

currently and to have occurred historically only for two groups of projects, one directed at thistles and the 

other at cacti. Van Driesche and Hoddle (2017) found that the parasitoids and predators used against invasive 

insects and mites historically have been less specific than weed biocontrol agents, and these agents can be 

expected to attack some non-target hosts. Population-level impacts on non-target species are relatively 

uncommon but do occur, and the average level of host specificity of parasitoids and predators released has 

narrowed since 1985.

Historical Development of Biocontrol for Protection of Nature

1880-1975: Spill-over benefits from economic projects. Classical biological control of pest 

insects is generally traced to California in the 1880s when it was used successfully to protect citrus crops 

(Caltagirone and Doutt, 1989). Weed biocontrol was effectively launched (through its first large scale 

success) in Australia in the 1920s when vast areas of grazing and forest lands were cleared of dense stands 

of invasive Opuntia cacti through introductions of natural enemies (Dodd, 1940). The goal of pest control 

through release of natural enemies in most of the method’s first century (up to ca 1975) was to improve the 

economic circumstances of groups or regions. Economics was the measure of value of most activities of 

societies in this period of history. Protection of intact natural areas as national parks or wildlife refuges had 

begun in this period and the ideas that their protection embodied slowly broadened societies’ views on the 

conservation and protection of natural areas, especially in the 1960s when many key environmental laws 

were first enacted. Nevertheless, for biological control projects done before 1975, economic gain was their 

purpose in nearly all cases. That said, even in this early period, some biological control efforts had spill 

over benefits to natural areas. This was particularly true for weed biological control projects. Removal of 

dense thickets of exotic cacti in New South Wales in the 1920s was done to promote ranching and forestry, 

but perforce permitted native plants to reclaim much lost ground, as not all the 20 million infested hectares 

were converted to farms or tree plantations as cactus stands declined. On land that remained undeveloped, 

reduction in cactus density must have resulted in better growth of native plants, although no one thought it 

worthwhile to document it at the time. Similarly, pre-1975 projects of biocontrol of European invasive forbs 

in North America, Australia, and other parts of the world also benefitted native grassland plants, although 

these project’s goals were to improve grazing. In the United States, these rangeland projects started with 

the control of St. Johnswort (Hypericum perforatum L.) in California in the 1940s, followed by control 

of tansy ragwort (Senecio jacobaea L., currently given as Jacobaea vulgaris Gaertn.) in California and 

Oregon (US), and puncture vine (Tribulus terrestris L.) in California and the southwestern US (Andres 

et al., 1976). Further projects against other grassland species with similar dual benefits (economic and 

ecological) were initiated in subsequent decades in North America (Nechols et al., 1995; see also Winston 

Chapter 1: History of Biocontrol in Natural Areas and Issues Affecting Its Use 5
Biological Control in Natural Areas

et al., 2014). During this same general period, programs were also started in the eastern United States 

against two floating aquatic invasive plants: alligator weed (Alternanthera philoxeroides [Mart.] Griseb.) 

and waterhyacinth (Eichhornia crassipes [Mart.] Solms.-Laubach) (Andres et al., 1976), and later expanded 

to other pests of water bodies, wetlands, or surrounding habitats in the southern United States (Van Driesche 

et al., 2002). These projects, which were justified in terms of economics (reduction in herbicide use, 

preservation of navigation, protection of electrical generating facilities on rivers), produced benefits that 

were overwhelmingly ecological. 

Historically, insect biological control was most frequently done in support of agriculture and as such 

did not often provide benefits to natural areas, except for projects directed against invasive pests of native 

forest trees, especially in North America (e.g., Embree, 1971; Embree and Otvos, 1984; Ryan, 1990; see 

also Van Driesche and Reardon, 2014 for a review of the use of classical biological control in North America 

to protect native forests). Also, because some insect pests that feed on crops are polyphagous, they may also 

attack native plants in natural areas. In this case, release of natural enemies against them, for their control 

in crops, may lead to the natural spread of the controlling natural enemies into areas of native vegetation, 

providing benefits there, as occurred in Australia when parasitoids of Ceroplastes wax scales moved into 

forests and controlled the scales there, as discussed by Sands in Van Driesche et al. (2010). 

1975-present: Infrastructure expansion and benefits of new projects. By 1971, in the United 

States, three agencies of the Department of Agriculture were involved in carrying out biological control 

projects: ARS (Agricultural Research Service), APHIS (Animal and Plant Health Inspection Service) and 

FS (Forest Service) (Vail et al., 2001). Overseas laboratories to support collection of natural enemies were 

established in various locations, including most enduringly in Europe (France, Greece, and Italy), South 

America, Australia, and China (Vail et al., 2001). Establishment of USDA weed biocontrol laboratories in 

Albany, California, and Sidney (initially in Bozeman), Montana, supported expansion of work on invasive 

plants in grasslands, and one in Ft. Lauderdale (first in Gainesville), Florida, supported work on aquatic 

weeds. Insect biological control occurred at a USDA laboratory in Newark, Delaware (with work on both 

crop and forest pests), and in various U.S. land-grant universities (especially those in California, Hawaii, 

Florida, and Texas). Other facilities supplemented these, including weed biocontrol work at the Beltsville, 

Maryland laboratory, and various others constructed to support specific projects. 

Facility development also occurred in the other major countries involved in classical biological 

control, including various laboratories in the UK’s CABI system and a multi-country facility developed 

in Montpelier, France. These new or refurbished locations increased the capacity of countries to address 

invasive insect and plant problems through natural enemy introductions. 

In Australia, biological control research is undertaken by CSIRO and the Queensland, New South 

Wales, and Victorian Departments of Agriculture, with quarantine facilities in each state. From the 1960s 

onward, temporary overseas laboratories were established in various countries to support collection of natural 

enemies for projects. The first opened in 1960 in the UK and focused on Sirex wood wasp biological control. 

Others followed: Brazil, 1965-1982 (weeds of South American origin); Mexico, 1984-2010; Montpelier, 

France, 1966 (skeleton weed, Chondrilla juncea L.); Iran, 1978-79; Portugal (millipedes) and Spain (dung 

beetles) in the 1980s; South Africa, 1970-1994 (dung beetles and weeds); and Argentina, 1996-1999 (South 

American weeds). Of these, all have closed except the CSIRO European Laboratory at Montpellier, France, 

which is based on a joint biocontrol research campus with the USDA at that location (Andrew Sheppard, 

CSIRO, pers. comm.).

In South Africa, before 1970, several successful but isolated weed biological control projects were 

carried out against cacti, with little institutional development or broad organizational support. In the early 

1970s, the Weed Research Division of the Plant Protection Research Institute (PPRI) was established under 

the Department of Agriculture and Technical Services (later the Agricultural Research Council), and, in 

6 Chapter 1: History of Biocontrol in Natural Areas and Issues Affecting Its Use
Biological Control in Natural Areas

collaboration with Rhodes University, this Division began to more systematically target invasive cacti, 

waterweeds, and woody invasive plants (especially species of Acacia and Hakea). In the 1980s, the University 

of Cape Town established a biological control research group targeting invasive woody plants. In 1995, with 

the formation of a governmental program called Working for Water (initially in the Department of Water 

Affairs but later the Department of Environmental Affairs), funding available for weed biocontrol in South 

Africa increased greatly. PPRI currently has three research laboratories (Pretoria, Cedara – KwaZulu-Natal 

Province, and Stellenbosch). Separately, the University of Cape Town, Rhodes University, the University 

of the Witwatersrand, and the University of KwaZulu-Natal all engage in biological control research. The 

Working for Water Program employs biological control implementation officers in all nine of the provinces 

of the country (Martin Hill, pers. comm.). All weed biocontrol projects in South Africa are directed against 

environmental weeds.

In New Zealand, weed biocontrol was carried out from 1975 by the Department of Scientific and 

Industrial Research until its disestablishment in 1992. Since then Landcare Research (a new crown research 

institute) has been largely responsible for weed biocontrol (Hayes et al., 2013). New Zealand does not 

maintain any overseas biocontrol laboratories, but rather partners with other organizations such as CABI to 

do overseas work. Funding for weed biocontrol comes from both national and local governments and has 

been relatively stable. New quarantine containment facilities have been built. Regulation of biocontrol is 

through Environmental Protection Authority (EPA) under the Hazardous Substances and New Organisms 

Act of 1996, which provides a clear framework for decision-making on applications to release new weed 

biocontrol agents, uses formal risk assessment and cost-benefit analyses, has statutory time frames for 

decision-making, ensures public participation, and maintains the independence of the EPA (Hill et al., 

2013). Biocontrol targeting arthropod pests affecting the natural environment has been limited to one 

project against yellow jackets (Vespula wasps) (Ward and Schnitzler, 2013). However, further interest 

has emerged recently as part of a new multi-agency Bioheritage National Science Challenge (http://www.

biologicalheritage.nz/home).

In Canada, no system of overseas laboratories was established. Rather, many projects were 

conducted using the resources of the Commonwealth Agricultural Bureaux (CAB), now CABI, in the UK, 

an organization that had laboratories of its own in Europe, India, the Caribbean, and other parts of the 

world. In addition, some Canadian projects were continuations of projects begun in the United States. 

Collectively, these new facilities supported many new projects after 1975, many of which benefited 

natural areas. For details on projects on western US weeds and insects see Nechols et al. (1995), for projects 

against invasive plants in the eastern United States see Van Driesche et al. (2002), for invasive forest insect 

pests see Van Driesche and Reardon (2014), and for weeds in South Africa see Impson et al. (2011). While 

no clear line exists, gradually projects’ goals for weed biocontrol, and to a lesser degree for insect biocontrol, 

became increasingly focused on ecological benefits, either alone or in conjunction with economic goals. A 

partial sampling of these projects is briefly noted here, arranged by the biome gaining the project’s benefits 

(Table 1-1). Nine examples are presented at greater length in the next section as case histories and are not 

included in Table 1-1.

Chapter 1: History of Biocontrol in Natural Areas and Issues Affecting Its Use 7
Biological Control in Natural Areas Table 1-1. Selected examples of additional biological control projects producing benefits for natural areas, not including those presented as case studies.

Biome/ locationbenefitted

Target pest(P = plant; I = Insect)

Principal biocontrolagents

Examples of species or processes benefitted

References

Grasslands (G)

Western US

Tansy ragwort, 

P

Senecio jacobaea L. (Asteraceae)

Two Chrysolina spp. (Chrysomelidae)

In Oregon coastal prairies, lower tansy ragwort numbers allowed a rare plant (Sidalcea hirtipes C.L. Hitchc.) to increase by 40% (Gruber and Whytemare 

1997). In California, native plant populations increased in coastal prairies following biocontrol, benefitting California poppy (Eschscholzia californica

Cham.), beach strawberry (Fragaria chiloensis [L.] Mill.), Alameda County 

thistle (Cirsium quercetorum [A. Gray] Jeps.), coyote mint (Monardella 

villosa Benth.), and coyotebrush (Baccharis pilularis DC.) (Pemberton and Turner 1990).

Coombs et al., 1996; Gruber and Whytemare 1997; McEvoy et al., 1991; Pemberton and Turner 1990; Turner and McEvoy, 1995 

(G)Western US

Leafy spurge, 

P

Euphorbia esula L. (Euphorbiaceae)

Various species of Aphthona (Chrysomelidae)

Butler and Wacker (2010) found that in an experimental context in Montana, biocontrol agents reduced leafy spurge density by 80-90%. However, the vegetation that took its place was dominated by nonnative grasses rather than the mix of native forbs and grasses representative of non-infested areas. Setter and Lym (2013) found similar or higher reductions in leafy spurge stems and seedbank numbers at two sites in North Dakota over a ten-year interval, which led to partial recovery of native species. Recovery was slowed, however, by increases in non-native grasses, and seeding of native species may be needed to enhance restoration. Anderson et al. (2015) found that over an 18-yr period in the mountain foothills of Utah, leafy spurge declined 96% due to biocontrol agents, but it was replaced largely by European grasses. This change, while not benefitting native plants, did increase forage quality for native elk (Cervus canadensis) in the region. Lesica et al. (2016) identified suitable native forbs for reseeding into leafy spurge-infested areas during or following its biocontrol. Other noted benefits include the recovery of the threatened western prairie fringed orchid (Plantathera praeclara Cheviak and Bowles) (Lym, 2005) 

and of a native Euphorbia (E. robusta Engelm.), predicted to be within the host range of one of the leafy spurge control agents (Aphthona nigriscutisFoudras). In the latter case, increases in population of the native species occurred because lowered competition pressure from leafy spurge was a stronger effect on the native species than feeding by the biocontrol agent (Baker and Webber, 2008).

Anderson et al. 2015; Baker and Webber, 2008; Butler and Wacker 2010; Lesica et al. 2016; Lym, 2005; Setter and Lym 2013

(G)Queensland, Australia

Rubber vine, 

P

Cryptostegia grandiflora (Roxb.) R. Br. (Apocynaceae)

Maravalia cryptostegiae

(Cummins) Ono (Pucciniales)

Increased growth of local grasses. Evans and Tomley, 

1994; Vogler and Lindsay, 2002

8 Chapter 1: History of Biocontrol in Natural Areas and Issues Affecting Its Use
Biological Control in Natural Areas

 Table 1-1 (continued). Selected examples of additional biological control projects producing benefits for natural areas, not including those presented as case studies.

Biome/ locationbenefitted

Target pest(P = plant; I = Insect)

Principal biocontrolagents

Examples of species or processes benefitted

References

Aquatic or wetland habitats (AW)

Purple loosestrife, 

P

Lythrum salicaria L. (Lythraceae)

Two Galerucella spp. (Chrysomelidae); Hylobius transversovittatusGoeze (Curculionidae); Nanophyes marmoratus(Goeze) (Brentidae)

In many areas where loosestrife has been suppressed, native species have returned (Landis et al., 2003). Suppression of loosestrife and increase in native plants occurred at one of three sites in Virginia subjected to long term monitoring (McAvoy et al., 2016).

Boag and Eckert, 2013; Denoth and Myers, 2005; Grevstad, 2006; Landis et al., 2003; McAvoy et al., 2016; St. Louis, 2013

(AW) Various, East Africa Lake Victoria

Waterhyacinth, 

P

Eichhornia crassipes(Mart.) Solms(Pontederiaceae)

Two Neochetina spp.(Curculionidae)

In Benin, populations of fish in the genus Citarhinus returned to normal 

levels after biological control of waterhyacinth (Ajuonu et al., 2003; P. 

Neuenschwander, pers. comm.).

Anon. 2000; Beshir and Bennett 1985; Center et al., 2002; Coetzee et al., 2009; Wilson et al., 2007 

(AW)Various countries, includingPapua New Guinea and Australia

Giant salvinia, 

P

Salvinia molesta D. Mitchell 

(Salviniaceae)

Cyrtobagous salviniaeCalder and Sands (Curculionidae)

Better light penetration into aquatic systems;Reduced use of herbicides for aquatic weed management in conservation areas (Julien and Storrs, 1996)

Diop and Hill, 2009; Mbati and Neuenschwander, 2005; Room et al., 1981; Thomas and Room, 1986 

(AW)South Africa and Zimbabwe

Red fern, 

P

Azolla filiculoides Lam. 

(Azollaceae)

Stenopelmus rufinasusGyllenhal (Curculionidae)

Red fern invaded much of South Africa by the 1990s, matting over still water bodies (Hill, 1999), which affected native biodiversity in various parts of southern Africa (Gratwicke and Marshall, 2001). The rare fish Sandelia bainsii Castelnau was nearly driven to extinction due to alteration of it last habitat (J. Cambray, Albany Museum, pers. comm.). Stenopelmus rufinasuswas released in 1997, established and dispersed widely, and rapidly suppressed the weed (McConnachie et al., 2004), allowing recovery (Hill and McConnachie, 2009).

Gratwicke and Marshall, 2001; Hill, 1999; Hill and McConnachie, 2009; McConnachie et al., 2004

(AW)Northern Territory, Australia

Mimosa, 

P

Mimosa pigra L.(Fabaceae)

Various, but especially Carmenta mimosa Eichlin & Passoa (Sesiidae) and Neurostrota gunniella

(Busck) (Gracillariidae)

In some areas, agents have reduced seed rain and seedling regeneration, lowered seed banks, and caused the retreat of mimosa at stand edges in wetlands (Heard and Paynter, 2009), opening space for native plant regeneration.

Heard and Paynter, 2009

Temperate forests (TempF)

US mid-Atlantic 

Mile-a-minute weed, 

P

Persicaria perfoliata(L.) H. Gross (Polygonaceae)

Rhinoncomimus latipes

Korotyaev (Curculionidae)

With reseeding, areas freed of mile-a-minute weed by biological control reverted to native vegetation, both seeded and spontaneous species (Cutting and Hough-Goldstein, 2013; Lake et al., 2014)

Cutting and HoughGoldstein, 2013; Hough-Goldstein et al., 2009, 2012; Lake et al., 2011, 2014a

Chapter 1: History of Biocontrol in Natural Areas and Issues Affecting Its Use 9
Biological Control in Natural Areas Table 1-1 (continued). Selected examples of additional biological control projects producing benefits for natural areas, not including those presented as case studies.

Biome/ locationbenefitted

Target pest(P = plant; I = Insect)

Principal biocontrolagents

Examples of species or processes benefitted

References

(TempF) New 

Zealand 

Mist flower, 

P

Ageratina riparia(Regal) R. King and H. Robinson (Asteraceae)

Entyloma ageratinae

Barreto and Evans (Entylomatales) and Procecidochares alaniSteyskal(Tephritidae)

Native plant species richness and cover increased after decline of mist flower due to biological control, with no increase in invasive plants, except a weak response by African club moss (Selaginella kraussiana [Kunze] A. Braun). Many plants colonizing plots were native mid- or late-successional shrubs or trees (Barton et al., 2007)

Barton et al., 2007

(TempF) Eastern Canada, Western USA

Larch casebearer, IColeophora laricella(Hübner)(Coleophoridae)

Agathis pumila (Ratz.) (Braconidae) and Chrysocharis laricinellae(Ratz.) (Eulophidae)

Repeated defoliation of western larch (Larix occidentalis Nutt.) in northern Idaho reduced tree growth by 80%, but growth returned to normal after successful biological control (Long, 1988). Other biodiversity impacts were not investigated

Long, 1988; Ryan, 1990; Webb and Quednau 1971 

Tropical or subtropical forest (TropF)

Florida, USA

Old World climbing fern, P Lygodium microphyllum(Cav.) R.Br. (Lygodiaceae)

Neomusotima conspurcatalis (Warren) (Pyralidae) and Floracarus perrepaeKnihinicki & Boczek (Eriophyidae)

Some areas defoliated by N. conspurcatalis have been re-occupied by native plants (Boughton and Pemberton, 2009)

Boughton and Pemberton 2009, 2011; Smith et al. 2014; Lake et al. 2014b

Dunes or coastal habitats

(DCH) New South Wales and Western Australia

Bridal creeper, P Asparagus asparagoides (L.) Druce (Asparagaceae)

Undescribed leafhopper (Tribe Erythroneurini, formerly referred to as Zygina sp.) (Cicadellidae) and Puccinia myrsiphylli(Thuem.) Wint.(Pucciniales)

Bridal creeper is a threat to four endangered ecological communities in New South Wales – littoral rainforest, river-flat eucalypt forest on coastal floodplains, swamp-oak floodplain forest, and sub-tropical coastal floodplain forest (Downey 2006) and many native plants (Sorensen and Jusaitis 1995, Willis et al., 2003). The rust fungus is currently effectively controlling bridal creeper in many areas, but dead tuber mats persist and can prevent native plant root growth, unless additional measures are taken. Therefore, current benefits are largely in reducing rates of spread.

Downey 2006; Morin and Edwards 2006; Morin et al. 2006; Sorensen and Jusaitis 1995; Willis et al., 2003

Riparian Areas

(RE) Rivers and adjacent area in Cape and Fynbos regions, South Africa

Rattlebox, P Sesbania punicea (Cav.) Benth.(Fabaceae)

Trichapion lativentre(Bèguin-Billecocq) (Brentidae); Rhyssomatus marginatus Fåhraeus (Curculionidae);Neodiplogrammus quadrivittatus (Olivier) (Curculionidae)

Following more than >95% decline in sesbania density due to the biological control agents (Hoffmann and Moran, 1998), many stretches of rivers once clogged with the plant returned to a free-flowing state (Hoffmann, personal observations).

Hoffmann and Moran 1991, 1998 

(RE)South Africa

Long-leaved wattle, P Acacia longifolia(Andr.) Willd. (Fabaceae)

Trichilogaster acaciaelongifoliae(Frogatt) (Pteromalidae)and Melanterius ventralisLea (Curculionidae)

Agents reduced seed production of A. longifolia >95%, allowing increased in stream water flow (Dennill et al 1999).

Dennill et al. 1999

10 Chapter 1: History of Biocontrol in Natural Areas and Issues Affecting Its Use
Biological Control in Natural Areas

 Table 1-1 (continued). Selected examples of additional biological control projects producing benefits for natural areas, not including those presented as case studies.

Biome/ locationbenefitted

Target pest(P = plant; I = Insect)

Principal biocontrolagents

Examples of species or processes benefitted

References

(RE) South Africa

Port Jackson willow, P Acacia saligna (Labill.) H. Wendl. (Fabaceae)

Uromycladium tepperianum (Sacc.) McAlpine (Pucciniales)

From 1991 to 2005, A. saligna declined 87-98% at five monitored sites (Wood and Morris 2007). In many areas, dead A. saligna trees are being replaced by native fynbos plants, or other weeds and grasses (Lennox et al. 2004)

Wood 2012; Wood and Morris 2007

(RE) South Africa

Golden wattle, P Acacia pycnantha Benth.(Fabaceae)

Trichilogaster signiventris(Girault) (Pteromalidae)

This plant was brought under biological control before it had caused great damage to biodiversity, based on previous experience with the similar but more extensive infestation of A. longifolia, as a preventative measure. As such, its benefits are not in the way of correction of actual existing damage, but in its prevention.

Hoffmann et al. 2002; Impson et al. 2011

Oceanic Islands (OI)

Seychelles

Icerya seychellarum

(Westwood), I(Monophlebidae)

Rodolia chermesinaMulsant(Coccinellidae)

Following this scale’s invasion of Aldabra around 1968, native plants, including the endemic Sideroxylon inerme ssp. cryptophlebium (Baker) J. H. Hemsl., were heavily damaged (Newbery and Hill 1985; Newbery 1988; Gery 1991) and some rare species were on the point of extinction (Friedmann 1994). Following the 1989 release of R. chermesina (Gery 1991), scale density was shown to decline and native plants recover over a 5 yr monitoring effort (Johnson and Threadgold 1999.).

Friedmann 1994; Gery 1991; Johnson and Threadgold 1999; Newbery 1988; Newbery and Hill 1985

(OI) Ascension Island, South Atlantic

Icerya purchasi Maskell, I (Monophlebidae)

Rodolia cardinalis(Mulsant); released in 1976(Coccinellidae)

The ladybird beetle protected a rare endemic spurge, Euphorbia origanoidesL., whose populations were observed by Cronk (1980) to be declining due to cottony cushion scale (Cronk, 1980). By 1995 (19-yr after release of predator), the scale was not present on sampled spurge populations, including one site documented as infested by Cronk (1980) (Fowler, unpub.). A second survey in 1997 confirmed that E. origanoides was no longer attacked (Ashmole and Ashmole, 1997).

Ashmole and Ashmole 1997; Cronk 1980

(OI) Western Pacific atolls of Kiribati, Federated States of Micronesia, and Marshall Islands

Icerya aegyptiaca(Douglas), I(Monophlebidae)

Rodolia limbata(Blackburn)(Coccinellidae)

Icerya aegyptiaca invaded various western Pacific atolls (Beardsley 1955, 

Waterhouse 1993) and attacked various native plants, including Casuarina equisetifolia L., Calophyllum inophyllum L., and the sand dune plants 

Scaevola koenigii Vahl. and Scaevola frutescens K. Kraus (Manser 1974, 

Maddison 1976). Following its 1994-99, introduction, R. limbata controlled the scale (Brancatini and Sands 1997, Sands et al., 1999) allowing recovery of affected plants.

Beardsley 1955; Brancatini and Sands 1997; Maddison 1976; Manser 1974; Sands et al. 1999; Waterhouse 1993

Chapter 1: History of Biocontrol in Natural Areas and Issues Affecting Its Use 11
Biological Control in Natural Areas Table 1-1 (continued). Selected examples of additional biological control projects producing benefits for natural areas, not including those presented as case studies.

Biome/ locationbenefitted

Target pest(P = plant; I = Insect)

Principal biocontrolagents

Examples of species or processes benefitted

References

(OI) Guam and other Pacfic Islands

Cycad scale, IAulacaspis yasumatsuiTakagi (Diaspididae)

Rhyzobius lophanthae Blaisdell (Coccinellidae)

The scale Aulacaspis yasumatsui Takagi, along with several other nonnative insects, has killed 94% of the cycads in Guam’s limestone forests. Consequently, Cycas micronesica K.D. Hill has been listed as Endangered by IUCN since 2006, and Threatened by U.S. Fish & Wildlife Service since 2015. The coccinellid Rhyzobius lophanthae Blaisdell was introduced 

from Hawaii, established widely in 2005, and suppressed the scale on mature cycads. This predator was also established on Rota, Koror, and Babeldaup as the scale spread in Micronesia. However, this predator did not suppress the scale adequately on seedlings. The parasitoids Aphytis lingnanensis Compere and Coccobius fulvus (Compere and Annecke) have 

been introduced to Guam on numerous occasions. Establishment of these parasitoids has not been confirmed, and cycad plants continue to die every year on Guam and Rota. (Marler, pers. comm.). It is predicted that the population of cycads in western Guam will soon be extinct unless this trend is reversed (Marler and Lawrence 2012). More effective agents are needed, requiring work in the scale’s native range in southeast Asia.

Marler and Lawrence 2012

12 Chapter 1: History of Biocontrol in Natural Areas and Issues Affecting Its Use
Biological Control in Natural Areas

In addition to projects listed in Table 1-1 or discussed below, other projects initiated more recently have 

ecological goals but have not yet suppressed the pest or achieved demonstrable ecological benefits. Some of 

these projects likely will provide important benefits in the future and are mentioned here for completeness. 

While we do not list all such unfinished or in-progress projects, we call attention to the following: (1) hydrilla, 

Hydrilla verticillata (L.f.) Royle (Hydrocharitaceae) (Grodowitz et al., 2003; Bownes, 2015); (2) Japanese 

knotweed, Fallopia japonica (Houtt.) Ronse Decraene (Polygonaceae) and related species and hybrids (Shaw 

et al., 2009; Djeddour and Shaw, 2010; Grevstad et al., 2013) (3) cat’s claw creeper, Macfadyena unguis-cati 

A.H. Gentry (currently given as Dolichandra unguis-cati [L.] L. G. Lohmann) (Bignoniaceae), in Australia 

and South Africa (L.) (Dhileepan et al., 2007ab, 2013; King et al., 2011); (4) Brazilian peppertree, Schinus 

terebinthifolius Raddi (Anacardiaceae) in Florida (Cuda et al., 2012; Wheeler et al., 2016); (5) air potato, 

Dioscorea bulbifera L. (Dioscoreaceae), in the southeastern US (Overholt et al., 2016); (6) Madeira vine, 

Anredera cordifolia (Ten.) Steenis (Basellaceae), in Australia (Westhuizen, 2011); (7) diffuse (Centaurea 

diffusa Larmarck) and spotted (Centaurea stoebe L.) knapweeds (Asteraceae) in western North America 

(Smith, 2004; Story et al., 2000, 2006, 2008; Seastedt et al., 2007; Myers et al., 2009; Harris, 2011; Newman 

et al., 2011; Gayton and Miller, 2012); (8) Parthenium hysterophorus L. in Queensland, Australia (Dhileepan, 

2003); (9) saltcedars, Tamarix ramosissima Ledeb., other Tamarix species, and hybrids (Tamaricaceae), in 

the southwestern USA (Hudgeons et al., 2007; Carruthers et al., 2008; DeLoach et al., 2008; Tracy and 

Robbins, 2009; Pattison et al., 2011; Dudley and Bean, 2012); (10) hemlock woolly adelgid, Adelges tsugae 

Annand (Adelgidae), in the eastern United States (Havill et al., 2014; Mausel and Salom, 2014); (11) winter 

moth, Operophtera brumata (L ) (Geometridae), in New England (Elkinton and Boettner, 2014; Elkinton et 

al., 2015); and (12) emerald ash borer, Agrilus planipennis Fairmaire (Buprestidae) (Bauer et al., 2014; Van 

Driesche and Reardon, 2015) in eastern North America.

ISSUES AFFECTING DEVELOPMENT OF BIOCONTROL PROGRAMS 

FOR NATURAL AREAS

The further development of biological control projects against pests of natural areas will be affected by 

several social and financial influences that vary strongly among the various countries developing such 

projects, including the ability of biological control scientists and conservationists to agree on goals, the 

willingness of governments to pay for such projects as part of their support for the common good, social 

beliefs about the merits or offenses of biological control, and the goals and efficiency of whatever regulatory 

agency must approve the introduction of biological control agents. 

Common Cause with Conservation Biologists 

To succeed, natural-area biocontrol projects must in the future have support from local conservation 

biologists. Currently, the degree to which this occurs varies widely and can increase or decrease depending 

on local events. In Galápagos, for example, an area with no previous history of use of biological control, 

the invasion of a scale (Icerya purchasi Maskell) that proved highly destructive to local plants, stimulated 

the park service to consider the use of biological control. Successful completion of the project, with large 

benefits and high safety to local species (see details below in case history) cemented the view that use of 

biocontrol in the park was potentially suitable, stimulating consideration of further consideration of its use 

against invasive blackberry (Rubus niveus Thunb.) and a parasitic fly (Philornis downsi Dodge & Aitken) 

destroying populations of native land birds. In Hawaii, conservationists recognized the destructive influence 

Chapter 1: History of Biocontrol in Natural Areas and Issues Affecting Its Use 13
Biological Control in Natural Areas

of strawberry guava (Psidium cattleyanum Sabine) on native forests and strongly supported a biological 

control project against the invasive tree despite social controversy against it arising from an alliance of 

pig hunters and anti-government, anti-biological control individuals (Johnson, 2016). In contrast, in the 

southwestern United States, efforts to suppress vast stands of invasive saltcedars (Tamarix spp.) dominating 

native riparian areas were severely impeded by controversy despite wide agreement of the high level of 

damage to native riparian plant communities due to Tamarix (Van Driesche et al., 2016b). On the surface, 

this controversy was based on the use of Tamarix as nesting habitat by an endangered subspecies of bird 

(Empidonax traillii extimus Phillips); however, this alone does not account for the degree of rancor that 

developed since (1) the needs of the bird were, in the long term, consonant with saltcedar reduction, that 

plant being poor habitat compared to willow/popular stands found in uninvaded riparian stretches in the 

region and (2) the short term effects of biological control could likely be managed by replanting native 

trees, fencing, and better management of river dam releases and grazing intensity. Rather, it seems that the 

rancor reached the level of intensity it did due to dislike and distrust between conservation groups and those 

carrying out the biocontrol project. With that as motive, the bird’s status under the Endangered Species Act 

was used to stop the project. 

The degree of trust, rancor, or mutual ignorance of each other between conservation biologists and 

biological control scientists varies by location. In South Africa, a relatively small group of conservation 

biologists and biological control scientists have shown a high level of cooperation resulting in many 

effective projects against environmentally damaging plant invaders, in part because of a clear national 

understanding of the importance of weed biocontrol to protect water supplies. 

In New Zealand, a rigorous new-organisms regulatory act (HSNO) required extensive assessments 

for biological control introductions. While one might suppose that this would impede the introduction 

of new biological control agents, it has not. Rather, it has allowed introductions of higher quality agents 

(safer, more likely to be effective) to proceed. This favorable outcome likely stems from several factors, 

including the relatively smaller bureaucracy of a small country, social and cultural conditions that allow 

rapid development of consensus on values, and the small insect and plant biotas requiring assessment. 

The success of HSNO has stimulated the belief that a similar law might be effective in other countries. 

This extrapolation, however, may fail in practice in countries like the United States, due to more complex 

bureaucracy, greater social and political diversity, and larger continental insect and plant biotas, which 

likely would make it more difficult to reach consensus during safety assessment of biocontrol agents. 

In Canada, historically there has been a similar divide between conservation biologists and biocontrol 

researchers, although the increasing focus on urban areas and issues of invasive species in recent years has 

raised awareness for the need of ecological restoration with minimal disruption. Consequently, conservation 

authorities and municipalities have started to see biological control as a viable management option; a case in 

point is the recent introduction of Hypena opulenta Christoph (Lep.: Noctuidae) for biocontrol of invasive 

swallow-worts (Vincetoxicum spp.) (MacQuarrie et al., 2016) (Sandy Smith, pers. comm.). 

Governmental Policies on Funding Projects 

The study and introduction of biological control agents is of necessity a government-funded activity done 

in support of the common good. As such, its support requires both a social and governmental belief that 

the money would be well spent, producing tangible benefits. The use of biological control in natural areas 

competes poorly for funds against projects aimed to reduce losses from invasive agricultural or forestry 

pests because there is no obvious interest group to promote the project and plead for funds. For projects in 

natural areas, this role can potentially be played by NGOs or “friends” groups, but often such influence is 

not brought adequately to bear. 

14 Chapter 1: History of Biocontrol in Natural Areas and Issues Affecting Its Use
Biological Control in Natural Areas

Trends in funding and support for scientists and laboratories dedicated for biological control work 

in various countries are not rigorously compiled. However, in several key countries, disinvestment from 

public sector activity, particularly research, has been a dominant theme in politics over the last several 

decades. In contrast, in South Africa, the Working for Water Program has provided considerable, stable 

funding for weed biological control (Martin Hill, pers. comm.).

Future funding for biological control is not knowable, but if a broad social consensus on the utility of 

biological control for management of certain invasive pests at the landscape level can be developed through 

a better understanding between conservation biologists and biological control scientists, and then conveyed 

successfully to the public, prospects for funding would be improved. 

Social Beliefs about Merits and Offenses of Biological Control Agents

For biological control to be widely applied to conservation biology problems, the social understanding of the 

process, its risks, and the potential to alleviate risks from invasive species, must be improved, as controversy 

among groups is itself, in many cases, enough to prevent its use. A particular problem affecting this concern 

is that many people reason by reference to categories (e.g.: “Pesticides are bad. This material is a pesticide. 

Therefore, it is bad.”). Given the tendency of many people to think in this style, being able to impose a label 

on a thing defines its perception. Category labels are often based on familiar examples. Therefore, whether 

groups perceive biological control as good or bad will depend on the examples they know. Knowledge of 

biocontrol successes fade over time, while its failures or mistakes remain visible and influential. 

Given these ways in which groups perceive and judge biological control, the future of biological control’s 

use in natural areas will be affected by the degree to which agencies, NGOs, conservation groups, or others 

disseminate messages and stories about biological control – either as a force for ecological restoration or as 

an ecologically risky practice to be shunned. In Hawaii, efforts to communicate reasons for biological control 

of strawberry guava were initially not well and widely conveyed to the broader public, and, consequently, 

difficulties arose getting permits for release of the agent (Tectococcus ovatus Hempel) (Hemiptera: 

Eriococcidae) on state land (Johnson, 2016). When local conservation groups in favor of suppression of this 

invasive tree became engaged, communication about the project – its need, its methods, its benefits and risks – 

better reached more people, and views of the project became more widely favorable and permits were issued. 

Countries and agencies that partner with such conservation groups to undertake such public outreach on their 

behalf will benefit from improved social understanding and acceptance of their projects. 

Goals and Efficiency of Regulatory Agencies

Regulatory mandates for the agencies that pass judgment on biological control agent introductions matter 

to outcomes. All such agencies seek to ensure safety but may not be charged to also actively seek the 

potential benefits from such introductions (performing a risk/benefit analysis of introductions). Mandates 

for natural enemy introduction reviews vary considerably among countries (Sheppard and Warner, 2016). 

Features of good systems include (1) clear goals, (2) clear and not overly complex procedures for reviews 

and consultations, (3) mandated time periods for completion of reviews, and (4) ability to obtain scientific 

input from all relevant sources. Rigorous systems with prompt time frames (e.g., HSNO, New Zealand) 

appear to function smoothly, while vague, complex systems with no time frames (e.g., United States) may 

find themselves seriously confounded when required to change procedures or adopt new goals. Indeed, 

in the United States, recent efforts to include food web effects in weed biocontrol agent reviews caused 

authorities charged with protection of wildlife to virtually cease reviews for several or more years. Ideas for 

improvement have been articulated by Blossey (2016).

Chapter 1: History of Biocontrol in Natural Areas and Issues Affecting Its Use 15
Biological Control in Natural Areas

16 Chapter 2: Improving Health of Native Galapagos Plants with Biological Control

CHAPTER 2. IMPROVING HEALTH OF NATIVE 

GALÁPAGOS PLANTS BY INTRODUCING A SPECIALIZED 

PREDATOR OF THE INVASIVE COTTONY CUSHION SCALE

C. E. Causton1, C. Calderón Alvaréz1, C. D. Hoddle2, M. S. Hoddle2, M. P. Lincango1, 

T. G. A. Poulsom1, and R. G. Van Driesche3

1

Charles Darwin Foundation, Santa Cruz, Galápagos, Ecuador

2

Department of Entomology, University of California, Riverside

3

Department of Environmental Conservation, University of Massachusetts, Amherst

WAS THE INVADER A KEY SOURCE OF ECOSYSTEM DEGRADATION 

OR A THREAT TO NATIVE SPECIES?

The Galápagos archipelago, a UNESCO World Heritage Site and Biosphere Reserve, is renowned for its 

high levels of endemic flora and fauna, a result of its isolation from other land masses and late colonization 

by humans. These ecosystems are so unique that they have recently been recognized as one of the 137 

most irreplaceable protected areas in the world (Le Saout et al., 2013). A fast-growing tourist destination, 

this archipelago has seen a rapid a rise in visitors and residents since the 1990s, and the Galápagos are 

increasingly becoming more connected to the rest of the world. With this, pathways for alien species have 

multiplied, and the threats associated with these bio-invasions are increasing. 

The accidental introduction on plant material 

of the cottony cushion scale, Icerya purchasi Maskell 

(Hemiptera: Monophlebidae) (Figure 2-1), a cosmopolitan 

plant pest native to Australia (and possibly New Zealand), 

brought new threats to endemic Galápagos plants and 

other indigenous plants of conservation value. This insect 

is known to feed on over 200 different plant species, 

although it is best known as a pest of citrus (Caltagirone 

and Doutt, 1989; Causton, 2001). Icerya purchasi feeds 

on plant phloem, removing nutrients required for plant 

growth, and its honey dew promotes sooty mold that 

reduces photosynthesis. 

Figure 2-1. Cottony cushion scale, Icerya purchasi

(Credit: Mark Hoddle)
Biological Control in Natural Areas

Chapter 2: Improving Health of Native Galapagos Plants with Biological Control 17

Icerya purchasi was first recorded in the Galápagos Islands in 1982 and by 1996 had spread to 15 

of the 18 larger islands either by wind currents or movement of plants between islands (Causton, 2001; 

Causton et al., 2004). By 1996, the Charles Darwin Foundation (CDF) and the Galápagos National Park 

Service (GNPS) had received reports that I. purchasi appeared to be affecting the growth and survival 

of plant species of importance, including endemic plants with restricted ranges. In response, a Technical 

Advisory Committee was formed to oversee research to confirm observations that I. purchasi was causing 

substantial harm to Galápagos flora, and if so, to develop a control plan.

a b

Figure 2-2. (a) Cottony cushion scale, Icerya purchasi, on seagrape, Scaevola plumieri, a dunes dominant much affected by 

the scale; (b) assessing scale abundance (Credits: a,b Mark Hoddle)

Surveys found that at least 62 native and endemic plant 

species were fed on by the scale (Figures 2-2 and 2-3), of which 

16 were on the IUCN Red List of Threatened Species, including 

six in the two highest threat categories (Endangered and 

Critically Endangered) (Causton, 2001; Causton et al., 2004). 

Many introduced and cultivated species were also identified 

as hosts, some of which increased scale density in natural 

areas. Scale density and damage varied among host plants, and 

high infestations caused plant die back or death. One heavily 

infested native plant was white mangrove (Laguncularia 

racemosa [L.] Gaertn. F.), a dominant species along the islands’ 

coastline that creates important habitat for native invertebrates, 

fish, and birds, including the critically endangered mangrove 

finch, Camarhynchus heliobates (Snodgrass & Heller), which 

is restricted to a small stand of mangroves on the western coast 

of Isabela Island (Fessl et al., 2010). Other plants that were 

heavily infested included species of the endemic daisy genera 

Darwiniothamnus and Scalesia (Family Asteraceae), many of 

which are red listed because of their vulnerable status (Causton, 

2001; Calderón Alvarez et al., 2012). The rapid decline of 

Darwiniothamnus tenuifolius (Hook. f.) Harling (Darwin’s 

Figure 2-3. Assessing density of cottony cushion 

scale on Acacia macracantha in the arid zone 

(Credit: Mark Hoddle)
Biological Control in Natural Areas

18 Chapter 2: Improving Health of Native Galapagos Plants with Biological Control

Aster) on Alcedo Volcano on Isabela Island was associated with heavy I. purchasi infestations, and this 

plant’s decline may also have harmed its specialist feeders; three endemic Lepidoptera (Platyptilia vilema

B. Landry, Semiothisa cerussata Herbulot, Tebenna galapagoensis Heppner & B. Landry), previously 

found on the flanks of the volcano, were not recorded in the years following the invasion of I. purchasi 

and may have become locally extinct (Roque-Albelo, 2003). Also, honey dew produced by the scale may 

have facilitated population expansion of invasive ants (including Wasmannia auropunctata [Roger] and 

Solenopsis geminata [Fabricius]). These ants defend scale colonies from predators and transport scales 

between plants (Causton 2001; Hoddle et al., 2013).

Despite many reports of plants being affected by I. purchasi, proving that scale infestations harmed 

or killed field-grown plants was difficult because of the co-occurrence of other stress factors such as drought 

or poor nutrition. To resolve this uncertainty, in 1999 experimental trials were set up using potted plants 

to compare growth rates of scale-free vs scale-infested plants. Four native plants known to be hosts of I. 

purchasi in the wild were tested. Growth of leaves, shoots, and roots of three species (white mangrove, Acacia 

macracantha Humb. & Bonpl. ex Willd., and Phaseolus mollis Hook f.) was reduced by medium or high 

scale infestations. In contrast, growth of the endemic species Scalesia helleri B. L. Rob was not affected 

significantly by I. purchasi, and this plant consistently had only low densities of scale in the wild, suggesting it 

has some level of resistance to attack compared to other Scalesia (Causton, 2001; ML Johnson et al., unpub.).

WERE OTHER CONTROLS FEASIBLE?

With the data from these trials and plant surveys, the Technical Advisory Committee concluded that there 

was enough evidence to show that I. purchasi was having a significant impact on Galápagos ecosystems 

and measures should be undertaken archipelago-wide to reduce I. purchasi populations. The extensive 

distribution of the scale insect and the inaccessibility of some invaded sites meant that mechanical or 

chemical control methods (Causton and Van Driesche, 2016) were not feasible. Consequently, the committee 

decided that classical biological control was the best option as it was the only management tool that was 

likely to cause permanent and widespread suppression of this pest on both inhabited and uninhabited 

islands. Fortuitously, biological control had been used against I. purchasi previously to protect citrus plants, 

eliminating the need to conduct exploratory surveys and providing relevant information for evaluating the 

safety of importing natural enemies of the pest to the Galápagos. 

WHAT AGENTS WERE RELEASED AND WERE THEY SAFE?

The coccinellid beetle Rodolia cardinalis Mulsant (Figure 2-4a,b), a natural enemy of Icerya in its native 

range and the most commonly used species in earlier biological control programs against the pest, appeared 

best matched to survive the conditions of Galápagos. Furthermore, studies in agricultural systems where R. 

cardinalis had been deliberately introduced (e.g., California) (Quezada and DeBach, 1973) and in its native 

range in Australia (Prasad, 1989) suggested that the beetle would pose little risk to non-target species and 

effectively suppress I. purchasi. Nevertheless, most control programs with R. cardinalis were conducted 

before strict screening protocols were put in place to evaluate the risks of introducing an alien species to 

control an alien pest, and no or few post-release evaluations had been conducted to evaluate non-target 

impacts. Additional studies were therefore considered necessary to determine whether the importation of this 

beetle would pose risks to native and endemic insects or jeopardize Galápagos ecosystems in other ways. 
Biological Control in Natural Areas

Chapter 2: Improving Health of Native Galapagos Plants with Biological Control 19

a b

Figure 2-4. An adult (a) and larva (b) of Rodolia cardinalis feeding on cottony cushion scale, Icerya purchasi (Credits: a,b 

Mark Hoddle)

In 1999, with the help of CSIRO Entomology, R. cardinalis beetles were collected from Brisbane, 

Australia and imported into a quarantine facility at the Charles Darwin Research Station (CDRS) on Santa 

Cruz Island, Galápagos. This facility was specially built to conduct tests after a cost analysis demonstrated 

that it would be prohibitively expensive and logistically difficult to do testing outside Galápagos. From 

1999 to 2000, 16 insect species from three orders (Hemiptera, Coleoptera, and Neuroptera) and nine 

families (Aphididae, Coccidae, Coccinellidae, Chrysopidae, Diaspididae, Eriococcidae, Monophlebidae, 

Ortheziidae, Pseudococcidae) were tested as prey for R. cardinalis larvae or adults. Species tested were 

those presumed to be at high risk of being preyed upon by R. cardinalis in the event of its release (Causton 

et al., 2004). For all trials, no-choice tests were used in which adults or larvae were exposed only to nontarget prey (but not I. purchasi).

Non-target impact studies were later extended to evaluate the potential risk of R. cardinalis, if eaten, 

to endemic, insectivorous vertebrates, especially birds, because when attacked coccinellids produce a 

repellent fluid containing an alkaloid that might be toxic to some species. Although, no evidence exists 

to suggest that fluids secreted by R. cardinalis are toxic, the precautionary principle was applied and 

experimental studies were carried out to test the consequences of feeding immature stages or adults of R. 

cardinalis to two species of Darwin’s finches (Lincango et al., 2011). 

The results of these studies suggested that R. cardinalis would not present any significant threat to 

non-target insects or insectivorous birds (Causton et al., 2004; Lincango et al., 2011). In the host-specificity 

trials, immature and adult stages of R. cardinalis did not feed on any of the insect species presented to 

them except for an endemic ground pearl, Margarodes similis Morrison, a scale in the same family as I. 

purchasi. This species, however, is typically encased in a hard shell and found 40 cm below ground and 

was unlikely to be at risk from predation. Furthermore, R. cardinalis was not able to complete its life cycle 

on this species. When evaluating the effects of feeding R. cardinalis adults to species higher up the food 

chain, we found that beetles were distasteful to the two species of finch tested and were actively avoided or 

rejected. Symptoms of toxicity (mortality, deterioration, critical weight loss) of birds that had been handfed R. cardinalis were not observed in either species while in captivity; long-term monitoring could not be 

conducted because of ethics issues regarding time to keep the birds in captivity. 

In 2001, a risk assessment was submitted to the Technical Advisory Committee, and subsequently 

GNPS approved release of R. cardinalis from quarantine. From 2002 to 2005, over 2200, R. cardinalis
Biological Control in Natural Areas

20 Chapter 2: Improving Health of Native Galapagos Plants with Biological Control

were released on 10 islands. A strong emphasis was placed on community involvement in the release and 

post-introduction program to raise awareness about biological control and how it works and to emphasize 

the years of studies that are needed before importation to ensure the safety of the biological control agent.

WERE THERE UNANTICIPATED SYSTEM RESPONSES?

A monitoring program was initiated on Santa Cruz Island soon after the release of R. cardinalis, and we 

found that R. cardinalis quickly established in urban, rural, and natural areas. Within 10 weeks, it was 

found on at least 10 plant species in Puerto Ayora, and within three months, R. cardinalis had dispersed 

without human assistance some 45 km across Santa Cruz Island and reached the neighboring island of 

Baltra (separated from Santa Cruz by a 200 m channel) (Calderón Alvarez et al., 2012), demonstrating 

a strong dispersal capacity. Direct observations and predator-exclusion cages were used to evaluate the 

efficacy of the beetle in reducing scale populations on the badly damaged white mangroves on Santa Cruz 

Island, and these studies showed a rapid decline of cottony cushion scale density, together with an increase 

in R. cardinalis (Calderón Alvarez et al., 2012). Within three months, scale on mangrove in the experiment 

declined >99% (Calderón Alvarez et al., 2012). These results suggested that R. cardinalis by itself was 

sufficient to reduce scale insect numbers, at least on some host plants, and that other restoration actions (see 

Causton and Van Driesche, 2016) were not needed in the short-term. A decision was made to carry out a 

more extensive evaluation after enough time had passed to allow outcomes to become apparent. 

WAS INVADER SUPPRESSION SUFFICIENT FOR RESTORATION?

Seven years after the first release of R. cardinalis, a second evaluation was initiated through a collaborative 

effort of scientists from CDF, UMASS Amherst, University of California Riverside, technical staff from 

Galápagos Biosecurity Agency and the Galápagos National Park Directorate, and naturalist guides. The 

objectives of the evaluation were (1) to determine how widely R. cardinalis had established and whether 

additional releases were needed, (2) to evaluate whether R. cardinalis was successfully limiting I. purchasi 

densities, (3) to determine if damaged plant species had recovered, and (4) to see if R. cardinalis was 

feeding on any non-target species. In 2009, nine islands were surveyed for the presence of I. purchasi 

and R. cardinalis, in urban, agricultural, or natural areas. In addition, from late 2009 through late 2011 

selected species of plants were monitored monthly at six study sites on Santa Cruz or San Cristobal islands 

to measure the degree of suppression of I. purchasi by R. cardinalis. Results indicated that R. cardinalis 

had survived and spread after its introduction in 2002 and that it was found in a wide range of habitats. 

Establishment of R. cardinalis was confirmed for six of the ten islands (Fernandina, Floreana, Isabela, 

Marchena, San Cristobal, and Santa Cruz) where it had been released and on two additional islands (Baltra 

and Champion). In 2016, R. cardinalis was also found on Genovesa Island, where only 22 beetles had been 

released (in 2003). Genovesa, is a small (14 km2), isolated, low elevation, arid island, and this recovery 

suggests that R. cardinalis can survive for long periods of time under harsh conditions. 

In 2009, R. cardinalis was found on 48 of 112 plant species reported as hosts of I. purchasi (Calderón 

Alvarez et al., 2012). Monthly counts on plants at four sites on Santa Cruz and two sites on San Cristobal 

islands during 2009-2011 found that I. purchasi populations were at relatively low densities on most plant 

species, and overall, these were lower than levels observed before R. cardinalis was released (Hoddle et 

al., 2013). On Santa Cruz Island, densities of I. purchasi were extremely low on white mangrove over the 

whole 26-month survey, and scale insects were rarely detected during one-minute timed searches (Hoddle 
Biological Control in Natural Areas

Chapter 2: Improving Health of Native Galapagos Plants with Biological Control 21

et al., 2013) compared to more than 100 adult females per branch in 2002 before the predator’s release 

(Calderón Alvarez et al., 2012). Results were similar though less dramatic on three plant species monitored 

pre- and post-liberation in two arid-zone sites on Santa Cruz Island, and varied with host plant and habitat 

(Hoddle et al., 2013). Consistently low I. purchasi densities were found on palo verde (Parkinsonia 

aculeata L.) and Acacia insulae-iacobi L. Riley; however, infestation rates on A. macracantha varied 

between sites, suggesting that differences in vegetation diversity or density, and the level of moisture in the 

habitat influenced the efficacy of R. cardinalis. The highest I. purchasi densities were found in dunes on 

Scaevola plumieri (L.) Vahl, at Tortuga Beach, Santa Cruz Island, where over the 2-yr survey, infestation 

rates ranged from 20 to 90%. However, plant condition remained consistently good (Hoddle et al., 2013), 

despite strong predator-prey cycling. Scale densities also remained relatively high in urban areas on San 

Cristobal, on pigeon pea (Cajanus cajan [L.] Millsp.) and gooseberry tree (Phyllanthus acidus [L.] Skeels), 

where scales were heavily tended by invasive ants that interfered with R. cardinalis foraging (Hoddle et al., 

2013) (Figure 2-5).

Figure 2-5. Tending of cottony cushion scale by invasive ants promotes high scale density by interfering with predation by 

the biocontrol agent, Rodolia cardinalis (Credit: Mark Hoddle)

Lack of resources prevented us from monitoring the response of the threatened D. tenuifolius

populations on the flanks of Alcedo Volcano to the 

introduction of R. cardinalis; however, in 2011 park 

rangers confirmed the presence of R. cardinalis in the 

area and reported healthy populations of D. tenuifolius. 

To determine if R. cardinalis fed on nontarget 

species, we conducted some field observations, but 

most observations were made by exposing R. cardinalis 

adults to potential prey in large, outdoor cages (ca 

1.5 x 2 x 2 m) at CDRS in 2009. Potted endemic or 

native plants (e.g., A. macracantha, Gossypium sp., 

P. aculeata, and Waltheria ovata Cav.) were placed in 

the cages and infested with I. purchasi (the target) and 

nontarget prey species (e.g., Coccus viridis [Green], 

Ceroplastes spp., local mealybugs, aphids, and spider 

mites, as available) (Figure 2-6). Thirty-one fieldcollected or laboratory-reared Rodolia that had been 

starved for 24 hrs were released individually into 

Figure 2-6. Field cage used to measure reaction of 

Rodolia cardinalis to nontarget insects offered as prey, 

when presented to starved adult beetles in field cages 

stocked with native plants infested with potential 

prey species (Credit: Mark Hoddle)
Biological Control in Natural Areas

22 Chapter 2: Improving Health of Native Galapagos Plants with Biological Control

these cages and each beetle was observed for ~1 hr. Behaviors that were noted included encounters with 

nontargets and foraging activity. During more than 22 h of observations, free-ranging R. cardinalis adult 

beetles only fed on the target, I. purchasi, and were not observed to feed on non-target insects, despite 

frequent encounters. Furthermore, we found no evidence of R. cardinalis attacking non-Icerya prey species 

during field surveys, even when non-target insect species were close to I. purchasi (Hoddle et al., 2013). 

These results support pre-introduction work conducted under quarantine showing that R. cardinalis has 

high host fidelity and is unlikely to feed on nontarget species in the Galápagos. Finally, we evaluated the 

search efficiency of R. cardinalis by placing 15 sticky trap cards in Icerya-infested plants and 15 sticky 

traps in non-infested plants. Very few Rodolia beetles were trapped on non-infested plants, suggesting that 

R. cardinalis readily locates I. purchasi-infested plants from a distance and that encounters are not random. 

This suggests that R. cardinalis can detect isolated populations of Icerya in the wild.

FUTURE OF THE SYSTEM

Additional surveys are needed to confirm establishment of R. cardinalis on all islands and to determine 

whether threatened plant species on some of the outer islands have recovered. In areas that have been 

surveyed, results indicate (1) that the importation of a single natural enemy of the invasive scale I. purchasi 

has been sufficient to restore plant communities and (2) that additional interventions, such as efforts to 

replant native species, are not needed. The level of control, however, varies with the plant species, habitat, 

and season, and may also be influenced by the presence of invasive ants. 

ACKNOWLEDGMENTS

The success of this program was due to a large collaborative effort (see participants listed in Calderón 

Alvaréz et al., 2012 and Hoddle et al., 2013), and we thank everybody who contributed to this project. We 

would like to express our special appreciation and thanks to the Galapágos National Park Directorate for 

their support and help with this project. This article is contribution number 2153 of the Charles Darwin 

Foundation for the Galapágos Islands.
Biological Control in Natural Areas

CHAPTER 3. POTENTIAL FOR BIOLOGICAL CONTROL 

OF A PARASITE OF DARWIN’S FINCHES 

R. A. Boulton and G. E. Heimpel 

Department of Entomology, University of Minnesota

WAS THE INVADER A KEY SOURCE OF ECOSYSTEM DEGRADATION 

OR A THREAT TO NATIVE SPECIES?

Philornis downsi (Figure 3-1) is a parasitic fly (Diptera: Muscidae) that was first recorded in the Galápagos 

archipelago in 1964 where it likely invaded from mainland Ecuador (Bulgarella et al. 2015). Philornis 

downsi poses a major threat to the passerine fauna of the Galápagos Islands due to the way in which its 

larvae feed. Adult females lay their eggs in bird nests, and first-instar larvae feed in the nares of nestlings 

while the second and third instars feed on nestlings externally by rasping at the skin and ingesting blood and 

fluid from the wounds, causing death or deformity of nestlings (Figure 3-2) (Fessl et al., 2006a). 

Figure 3-1. Philornis downsi is an invasive parasite of 

nestling birds threatening to cause extinction of many 

of Darwin’s finches in the Galápagos Islands (Credit: 

David L. Hansen, University of Minnesota)

Chapter 3: Potential for Biological Control of a Parasite of Darwin’s Finches 23

Parasitism by P. downsi causes delayed growth 

and development, anemia, and death of nestlings 

(Fessl et al., 2006a). Nestling mortality of up to 

100% has been reported for some of the ten species 

of Darwin’s finches affected by P. downsi in the 

Galápagos (O’Connor et al., 2010; Kleindorfer et 

al., 2014; Heimpel et al., 2016; Koop et al., 2016). 

Depending on host species and year, 64-98% of nests 

are infested by P. downsi, and on average there are 

over 40 pupae per nest (Fessl and Tebbich, 2002). 

This is a stark contrast to the same estimates for P. 

downsi recorded on mainland Ecuador, where, in 

any given year, 34.6% of nests were infested and on 

average there were 12.4 P. downsi pupae per infested 

nest (Bulgarella et al., 2015). 
Biological Control in Natural Areas

a

b

c

Figure 3-2. When the nestlings of ground finches are attacked by Philornis downsi, healthy brood, as in this vegetarian 

ground finch, Platyspiza crassirostris (a) are replaced by dead chicks (b) or ones with gross deformities (c) (Credits: 

a George Heimpel; b,c Jody O’Connor and Sonia Kleindorfer)

24 Chapter 3: Potential for Biological Control of a Parasite of Darwin’s Finches

Philornis downsi is now considered the greatest threat to the persistence of the adaptive radiation of 

passerines that was instrumental to the development of evolutionary theory by Charles Darwin (Causton 

et al., 2013). Until now, none of the 13 species of Darwin’s finches have gone extinct in Galápagos since 

human colonization (Grant and Grant, 2008). However, extinction is a very real threat for several of 

these species as a result of the invasion of P. downsi, in particular the critically endangered Mangrove 

Finch (Camarhynchus heliobates [Snodgrass & Heller]) and Medium Tree Finch (Camarhynchus pauper

Ridgway) (Fessl et al., 2010; O’Connor et al., 2010). Currently, Mangrove Finches are being protected from 

exposure to P. downsi by a practice known as ‘head-starting’, in which eggs are removed from nests in the 

field and reared in a biosecure facility. The reared fledglings are released into their native habitat. This is a 

very time- and labor-intensive procedure and is being used as a stop-gap measure while long-term solutions 

are developed (Cunninghame et al., 2013; Fessl et al., 2017). 

WERE OTHER CONTROLS FEASIBLE?

Several strategies for the suppression of P. downsi populations are currently being considered. These 

include short-term measures, such as the treatment of nests with insecticides and mass trapping using lures, 

and long-term measures, such as importation (classical) biological control, release of sterile males, and 

pheromone disruption. As with head-starting, the short-term control tactics are considered primarily stopgap measures while long-term measures are developed. Before discussing biological control, we briefly 

outline developments in strategies employing insecticides and the release of sterile males.

Pyrethroid insecticides such as permethrin have been used to treat the nests of several species of 

finches. Such treatments can greatly reduce the intensity of P. downsi infestation, resulting in increased 

nestling growth and fledging success (e.g., Fessl et al., 2006b; Koop et al., 2013). This work has been 

extended to include a novel study by Knutie et al. (2014) that demonstrated that birds can self-fumigate 

nests when permethrin-treated cotton is made available to birds. Permethrin is considered safe to birds as 
Biological Control in Natural Areas

LD50 doses are exceptionally high for various species including chickens, ducks, pheasants, and starlings 

(Causton and Lincango, 2014). Furthermore, Causton and Lincango (2014) carried out a case-specific 

risk assessment using data from Galápagos and concluded that there is no evidence that permethrin has 

detrimental effects on the mortality of adult and juvenile birds, but the authors point out that the potential 

for longer-term effects on reproductive function or fitness needs to be evaluated. This work is currently 

being extended using the zebra finch (Taeniopygia guttata Reichenbach) as a model species with the hope 

that the long-term safety of permethrin for birds as small as Darwin’s finches will be corroborated (M. 

Bulgarella, unpublished).

The sterile insect-release technique (SIT) is based on the release of large numbers of sterile male 

insects, and it has been successfully implemented for the control of several fly species including the new 

world screw worm (Cochliomyia hominivorax [Coquerel]) (Diptera: Calliphoridae) and several species of 

tropical fruit flies (Klassen and Curtis, 2005). The basic premise of SIT is that introducing many sterile males 

into a population prevents the production of offspring by any female that mates with such a male. Although 

the use of SIT has been suggested for P. downsi, the feasibility and likely efficacy of this technique for this 

species are questionable. The main impediment is that P. downsi flies do not mate readily under laboratory 

conditions. Fertile eggs can be obtained by capturing mated female flies in the field, however, and methods 

have been developed to raise the resulting larvae on a chicken blood-based diet (Lahuatte et al., 2016). 

Unfortunately, these procedures together are far too labor-intensive to offer a realistic basis for SIT at this 

time. Due to the difficulty of rearing P. downsi in the laboratory, it has not been determined how P. downsi

males respond to the ionizing radiation typically used to obtain male sterility in SIT programs. If ionizing 

radiation has a detrimental impact on male mating success, then success of SIT releases would be reduced 

(see Calkins and Parker, 2005). Furthermore, the mating behavior of P. downsi females is likely to strongly 

influence the success of the sterile insect technique. Indeed, polyandry (mating with multiple males) would 

not only dilute the effectiveness of releases but would place sperm of irradiated males in competition with 

that of normal males. Although there are few data on the mating system of P. downsi, Dudaniec et al. (2010) 

demonstrated a moderate degree of polyandry in P. downsi (females produced progeny from 1.9 different 

sires on average). 

Among the methods of intervention currently being used or contemplated, only head-starting and 

insecticide use have demonstrated protection of Darwin’s finches from P. downsi parasitism (Fessl et 

al., 2017). However, neither of these measures has any potential for long-term suppression of P. downsi

populations. Currently, the most promising long-term management strategy is importation of specialized 

parasitoids of the fly. 

WHAT AGENTS WERE RELEASED AND WERE THEY SAFE?

No biological control agents have yet been released into the Galápagos for the control of P. downsi. However, 

recent studies in the native range of P. downsi (mainland Ecuador and Trinidad & Tobago) have revealed 

parasitism of Philornis puparia (of several species) by five parasitoid species. Ecuadorian parasitoids include 

Conura annulifera (Walker) (Hymenoptera: Chalcididae) (Figure 3-3), a Brachymeria species (Hymenoptera: 

Chalcididae), a Trichopria species (Hymenoptera: Diapriidae), a Spalangia species (Hymenoptera: 

Pteromalidae), and an Exoristobia species (Hymenoptera: Encyrtidae) (Bulgarella et al., 2015, 2017). In 

Trinidad & Tobago, the parasitoid Brachymeria philornisae n. sp. Delvare was recently recovered from 

Philornis trinitensis Dodge & Aitken in Tobago and the species described (Delvare et al., 2017). 

Of these species, the most information is available for C. annulifera, which has been recovered 

exclusively from Philornis species (including P. downsi) in Trinidad, Brazil (Burks, 1960; Delvare, 1992; 

Chapter 3: Potential for Biological Control of a Parasite of Darwin’s Finches 25
Biological Control in Natural Areas

Figure 3-3. Parasitoids of Philornis downsi, 

such as the chalcidid Conura annulifera, 

attack larvae or pupae of the fly and occur 

in South America, where the fly is native, 

but are absent from the invaded range of the 

fly (Galápagos) (Credit: David L. Hansen, 

University of Minnesota)

26 Chapter 3: Potential for Biological Control of a Parasite of Darwin’s Finches

Couri et al., 2006), and Ecuador (Bulgarella et al., 2017). This suggests that it may be a specialist on 

Philornis species. Indeed, laboratory host-specificity studies done in quarantine in the U.S. support the 

hypothesis that C. annulifera is a Philornis specialist. While puparia of P. downsi were readily attacked, 

the puparia of five non-target fly species including other muscids, sarcophagids, and calliphorids were not 

(Bulgarella et al., 2017). 

Before the release of C. annulifera can be considered, host-use tests must be conducted on the nontarget species that would be most at risk in the Galápagos. The greatest priority is to test endemic species 

belonging to the cyclorrhapan Diptera – a taxonomic level that contains various superfamilies, including the 

Muscoidea, Oestroidea, Hippoboscoidea, Syrphoidea, and Platypezoidea (Griffiths, 1972). The life history 

of C. annulifera suggests that it is these species that represent the most likely non-target hosts. Conura 

annulifera is a solitary pupal ectoparasitoid that lays a single egg in the gap between the hard external 

puparium and the soft body of the developing Philornis pupa. The Cyclorrhapha are the only group of insects 

that exhibit this gap (Whitten, 1957), and comparative analyses suggest that parasitoids that exploit this 

niche are restricted to ovipositing there (Bulgarella et al., 2017). In summary, our empirical investigations 

and the historical record suggest that C. annulifera is a specialist on the genus Philornis. Furthermore, the 

life history traits of this species suggest that it is physiologically restricted to the cyclorrhaphan Diptera. 

However, to move the project forward, it is critical that the host specificity of C. annulifera be assessed 

more extensively on endemic species in the Galápagos. 

WERE THERE UNANTICIPATED SYSTEM RESPONSES?

Whether unanticipated problems might occur from releases of Philornis parasitoids cannot be foretold now, 

since none have yet occurred. 

WOULD INVADER SUPPRESSION BE SUFFICIENT FOR RESTORATION?

Since releases have not been made, we can only speculate about the potential efficacy of C. annulifera

or other species as biological control agents of P. downsi. Currently, it is unknown how much they might 

depress fly populations or if any measures would need to be integrated with a biological control program. 

With respect to the potential effects of release of C. annulifera, we can only say that limited field sampling 
Biological Control in Natural Areas

done to date has revealed a maximum of 11% parasitism of Philornis at one field site in mainland Ecuador 

(Bulgarella et al., 2017). While successful biological control agents used in other projects have often caused 

higher parasitism than this in their native ranges (Hawkins and Cornell, 1994), it would be premature to 

conclude that C. annulifera would be ineffective.

Mortality of the exact host stage attacked by C. annulifera (the pupa) is likely to have the strongest 

negative effect on P. downsi populations. Philornis downsi larvae presumably experience strong densitydependent, intraspecific competition (Kleindorfer et al., 2014), and so adding mortality to the post-larval 

stages (i.e., of pupae) avoids problems of potential redundant mortality or alleviation of host densitydependence. Population models have illustrated this principle (May et al., 1981; Abrams, 2009), and a 

population model developed for P. downsi confirmed that raising mortality of pupae should indeed be more 

effective in suppressing populations than raising mortality of eggs or larvae (G.E. Heimpel unpublished). 

Additional parasitoids beyond C. annulifera are currently being considered for biological control 

against P. downsi, should they prove necessary. An unidentified species of Trichopria has been found to 

parasitize Philornis in mainland Ecuador (Bulgarella et al., 2017). This species attacks and emerges from 

the pupal stage, and preliminary studies suggest specificity to Philornis (M. Bulgarella, I. Ramirez and 

G.E. Heimpel, unpub. data). Some characteristics of the life history of this species suggest that it may 

be an efficient biological control agent if used alongside C. annulifera. Unlike C. annulifera, Trichopria

sp. is a gregarious species, laying multiple eggs per host. Theoretical work by Heimpel (2000) highlights 

how gregarious vs. solitary development can influence host population suppression by parasitoids. These 

models suggest that egg limitation is more likely to reduce the efficacy of control by gregarious species 

such as Trichopria sp., whereas host resources are more likely to become limiting for solitary species like C. 

annulifera. Although it has been suggested that gregarious species are likely to be more effective biological 

control agents due to their higher fecundity and rapid recruitment (Heimpel, 2000), it may be that egg 

limitation results in lower population suppression when hosts are abundant. Host limitation may, on the other 

hand, become more of a problem for a solitary biological control agent if the host species is successfully 

suppressed. As such, it may be that the complementary introduction of a solitary and a gregarious parasitoid 

would buffer against failure of biological control should either hosts or eggs be limiting at any given time. 

Evolution should also be at least considered as a theoretical solution to this problem. Several studies 

suggest that the potential exists for at least some bird species to become resistant or tolerant to P. downsi

through natural selection (see Huber et al., 2010; Koop et al., 2013; Knutie et al., 2016). Coevolution of 

hosts and their parasites is, however, a slow process, and one that usually favors the faster-reproducing 

parasites (Damore and Gore, 2012). But birds can also adapt behaviorally, and this may allow resistance to 

develop more rapidly if these behaviors are socially transmitted. For example, adult and nestling Darwin’s 

finches engage in preening behaviors that may result in removal of P. downsi larvae from nestlings (Koop 

et al., 2013). In addition, Cimadom et al. (2016) have shown that three finch species preen using the leaves 

of an endemic Galápagos plant, Psidium galapageium Hook. f., and experimental work has demonstrated 

that extracts of this plant repel adult P. downsi. The possibility of natural resistance, whether it is conferred 

genetically or socially is appealing, but relying on this possibility alone to protect vulnerable populations 

seems unwarranted, particularly given the rapidity with which P. downsi is causing population declines in 

finch species (Koop et al., 2016). 

FUTURE OF THE SYSTEM

The reasons behind the greater abundance of P. downsi on birds in the Galápagos Islands compared to 

the mainland are likely complex and multifaceted, but a paucity of natural enemies in the introduced 

Chapter 3: Potential for Biological Control of a Parasite of Darwin’s Finches 27
Biological Control in Natural Areas

range is almost certainly a contributing factor. In mainland Ecuador, overall parasitism of P. downsi by 

parasitoids typically ranges from 10-20% (Bulgarella et al., 2015, 2017), but parasitism in the Galápagos 

has never exceeded 5%, and then only by introduced generalist species (Fessl et al., 2017), suggesting that 

eventually, parasitoid introductions, if safe, may be beneficial. Classical biological control of P. downsi

in the Galápagos Islands is a promising means of suppressing the population of this invasive species. 

Although the immediate aim would be to suppress the current population of P. downsi, introducing a highly 

specific biological control agent could also serve to protect vulnerable bird populations from invasions 

of other similar brood parasites, which may occur in the future. In addition to P. downsi there are at least 

three other Philornis species (P. niger Dodge & Aitken, P. falsificus Dodge & Aitken, and an additional 

unidentified species) in mainland Ecuador (Bulgarella et al., 2017). Ensuring that appropriate specialized 

natural enemies are present on the archipelago may minimize the impacts of future invasions of more 

Philornis species. Preventative measures such as this may be particularly prudent in this case, given the 

increasing popularity of the Galápagos as a tourist destination and the demand for transport links between 

the islands and the mainland (Causton et al., 2006).

ACKNOWLEDGMENTS

We thank Mariana Bulgarella and Jody O’Connor for their help. We would also like to thank our collaborators, 

the Galápagos National Parks Directorate (Project: PC-10-15 and 18-16) and the Charles Darwin Foundation 

for their continued help and support. We are grateful to the Galápagos Conservancy, International Community 

Foundation (with a grant awarded by The Leona M. and Harry B. Helmsley Charitable Trust), the National 

Geographic Society (award no. C257-13), the University of Minnesota Institute on the Environment, and the 

McKnight Foundation for funding.

28 Chapter 3: Potential for Biological Control of a Parasite of Darwin’s Finches
Biological Control in Natural Areas

CHAPTER 4. RESTORING GUMWOOD FORESTS ON 

ST HELENA: THE CONTRIBUTION OF BIOCONTROL 

OF ORTHEZIA SCALE

S. V. Fowler 

Landcare Research, Lincoln, New Zealand

WAS THE INVADER A KEY SOURCE OF ECOSYSTEM DEGRADATION 

OR A THREAT TO NATIVE SPECIES?

The small South Atlantic island of St Helena has 

a highly degraded but internationally significant 

terrestrial biota. For example, St Helena’s 

flowering plants include 30 endemic species 

in 23 genera, of which 10 genera are endemic 

(Pearce-Kelly and Cronk, 1990). However, this 

indigenous flora now covers only 1% of the 

island’s land area of 122 km2 (Cronk, 1989). 

In the past there would have been extensive 

forests of arborescent daisies such as the 8-10 m 

high gumwood tree, Commidendrum robustum

(Roxb.) DC. (Asteraceae). By the 1980s, this 

important part of the relic flora was represented 

by just 2500 trees, mostly in just two natural 

stands (Cronk, 2000) (Figure 4-1).

Figure 4-1. Last gumwood, Commidendrum robustum, relic 

arborescent daisy woodland at Peak Dale – about 2500 trees 

(Credit: Simon Fowler)

Chapter 4: Restoring Gumwood Forests on St Helena: Biocontrol of Orthezia Scale 29
Biological Control in Natural Areas

A scale insect was noticed 

on these gumwoods in 1991 and 

identified as Orthezia insignis 

Browne (Hemiptera: Ortheziidae) 

(Figure 4-2). This was the first 

record on St Helena of this South 

American scale insect, which is 

a pest throughout the continental 

tropics. Tree deaths started 

occurring in 1992. By 1993, the 

scale’s infestation of the gumwoods 

was a source of extreme concern to 

island conservationists, with over 

100 gumwood trees having died 

(ca 5% of all existing individuals 

in two years) (Fowler, 2004). If 

the observed exponential increase 

in the rate of tree deaths had been 

unchecked, it was estimated that 

all 2500 trees would have died by 1995 (Figure 4-3) (Fowler, 2004). This outcome was likely given the lack 

of natural enemies and the abundance of alternative hosts for the scale, such as the abundant weed Lantana 

camara L. (Verbenaceae) (Fowler, 2004).

This potential extinction of the gumwood trees of St Helena followed a long history of environmental 

degradation that began with the island’s discovery in the 16th century by Portuguese mariners and their 

release of goats, which by the early 18th century was thought responsible for the near complete destruction 

of the low altitude forests (as recorded by Darwin [1845]). Human settlement, from 1659 on, led to overexploitation of all but the very high altitude vegetation zones for timber and firewood, and unrestrained 

grazing by domestic and feral animals prevented regeneration (Cronk, 2000). Finally, from the late 1800s large 

areas of the highest altitude vegetation 

were cleared for plantations of New 

Zealand flax, Phormium tenax Forst. 

& Forst. (Asphodelaceae), a fiber crop 

(Cronk, 2000). The scale O. insignis

was therefore acting on an already 

highly degraded landscape, threatening 

the last wild stands of the endemic C. 

robustum and other Commidendrum 

species with extinction (Fowler, 2004).

Because St Helena was 

geopolitically important to colonial 

powers in the 17 through 19th centuries, 

detailed studies on its invertebrates 

were made incidentally by Europeans 

based there (Ashmole and Ashmole, 

2000). The island’s invertebrates 

showed remarkable evolutionary 

a b

Figure 4-2. Orthezia insignis (a) and its damage on gumwoods (b) (Credits: 

a United States National Collection of Scale Insects Photographs, USDA 

ARS, bugwood.org; b Simon Fowler)

30 Chapter 4: Restoring Gumwood Forests on St Helena: Biocontrol of Orthezia Scale

Figure 4-3. Rapid loss of gumwood trees following invasion of Orthezia 

scale (1991-1993) occurred but losses stabilized following biological 

control of scale (1993 and following) (Simon Fowler, unpublished data)
Biological Control in Natural Areas

radiation, furthered detailed in the 20th and 21st centuries (e.g., Basilewsky, 1972, 1976). As a result, the 

invertebrates that are dependent on Commidendrum species on St Helena are well known, and their 

loss would be predicted if Commidendrum species were to go extinct in the wild. Three insect orders 

(Coleoptera, Lepidoptera, and Hemiptera) on St Helena include families that have particularly impressive 

evolutionary radiations. These groups are well studied and include most of the Commidendrum-dependent 

species (Figure 4-4). For analysis of the impacts of Orthezia scale, only fully identified, well described 

species were included here, and any species likely to have become extinct before the invasion of O. insignis 

were excluded. Detailed collections made from 1800 to present give reasonable confidence in assessing 

the island’s insect fauna. Species were considered extinct pre-Orthezia if they had not been collected 

since 1960 and were in a group that was sufficiently well known, with available experts, for species 

to be identified/determined. The host specificity of the insects was based on observations, deductions, 

and host-records in published sources (Melliss, 1875; Wollaston, 1877; Wollaston TV (Mrs.), 1879; 

Basilewsky, 1972, 1976; Ashmole and Ashmole, 2000; Mendel et al., 2008). Some modifications were 

based on modern findings where this was deemed authoritative and thorough, e.g., Longitarsus helenae 

Wollaston (Coleoptera; Chrysomelidae), originally thought to be from “cabbage trees” (Wollaston 1877) 

but determined by the Belgian expeditions (Basilewsky, 1972) to be host specific on Trimeris scaevolifolia 

(Roxb.) Mabb. (Campanulaceae). Species only recorded rarely on single host plants were assumed to be 

host specific: in this case, it was assumed that the recent, very thorough surveys would have recorded 

these insect species from other host plants if the insects were not host specific. Commidendrum-dependent 

species that were oligophagous, i.e., ones recorded from more than one Commidendrum species, were also 

included in Table 4-1 as these would also have become extinct if all Commidendrum species on the island 

were driven to extinction in the wild by O. insignis.

Failure to suppress O. insignis through biological control would not only have driven a major 

component of the endemic flora extinct in the wild, but would have caused the co-extinction of a further 

14% of the endemic insect fauna of St Helena in the orders Hemiptera, Coleoptera, and Lepidoptera (Table 

4-1). Given the likely loss of further endemic species, such as associated parasitoids or other specialized 

natural enemies of the plant-associated insects in Table 4-1, this is likely to be a low estimate of the total 

predicted co-extinctions if Commidendrum species had become extinct in the wild on St Helena.

a b

Figure 4-4. The endemic insects of St Helena that are dependent on Commidendrum species that would have been destroyed 

by Orthezia scale include 12 beetles, 15 moths, and 7 hemipterans; among them, the two weevils illustrated here: 

Lamprochrus cossonoides commidendri Decelle (a) and Acanthinomerus sp. (b). All of this biodiversity was threatened by the 

pending destruction of their host plants by the invasive scale. (Credits: a,b Roger Key)

Chapter 4: Restoring Gumwood Forests on St Helena: Biocontrol of Orthezia Scale 31
Biological Control in Natural Areas

Table 4-1. The predicted loss of dependent species in the main insect orders associated with Commidendrum spp. on St 

Helena should these plant species become extinct in the wild as a result of the invasion of the exotic scale insect, O. 

insignis. Percentage losses used total numbers of endemic species from Ashmole and Ashmole (2000).

Taxonomic group Predicted number of species lost Percentage loss of endemic species

Hemiptera 7 18%

Coleoptera 12 8%

Lepidoptera 15 28%

Total 34 14%

32 Chapter 4: Restoring Gumwood Forests on St Helena: Biocontrol of Orthezia Scale

WERE OTHER CONTROLS FEASIBLE?

Foliage-applied insecticides can control O. insignis on individual gumwood trees, but effective use over 

whole stands in steep terrain subject to persistent southeast trade winds would have been extremely difficult. 

Perhaps more probably, soil-applied insecticides, which can be used to control O. insignis (Crop and Plant 

Protection Unit, 2009), might be applied, but the abundance of alternative host plants for the pest would 

have promoted frequent re-invasion that would have required an ongoing program of pesticide application. 

Furthermore, use of chemical control, while conceivably protective to the plants, would have caused severe 

non-target effects on the many endemic arthropods feeding on treated gumwoods (Table 4-1). Finally, the 

option of natural control by local native natural enemies was discarded as there appeared to be no existing 

natural enemies of O. insignis on St Helena, as shown by the fact that none were seen in a two-year study 

of the scale in gumwood stands on the island (Fowler, 2004).

WHAT AGENTS WERE RELEASED AND WERE THEY SAFE?

The biological control agent released on St Helena from 1993-95 was Hyperaspis pantherina Fürsch 

(Coleoptera: Coccinellidae) (Figure 4-5), which had previously been used successfully to control O. 

insignis in Hawaii and in several countries in Africa (Booth et al., 1995). The only report of any prey other 

than O. insignis being attacked by H. pantherina is from the host range testing carried out before its release 

on St Helena: when deprived of O. insignis, 

adult H. pantherina did attack mealybugs 

(Pseudococcus and Planococcus spp.). 

However, oviposition and successful 

development only occurred in the presence 

of O. insignis prey (Booth et al., 1995). At 

the time of the release of H. pantherina on 

St Helena, there were no known indigenous 

mealybugs or scale insects on St Helena 

(Fowler, 2004). The mealybug Ripersiella 

mediatlantica Matile-Ferrero (Hemiptera: 

Pseudococcidae) is now regarded as an 

endemic species on St Helena (Ashmole 

Figure 4-5. Hyperaspis pantherina, the predator that controls 

Orthezia scale on St Helena (Credit: www.discoverlife.org)
Biological Control in Natural Areas

and Ashmole, 2000; Key, 2014), but the Belgian entomologists who collected and described the only four 

specimens ever collected (found together with an exotic mealybug and an exotic scale on dead cabbage 

trees on the High Central Ridge in the 1960s) state it was not possible to say whether it was native or not 

(Matile-Ferrero, 1976). Whether H. pantherina has any impact on endemic mealybugs on St Helena is 

unknown, but it is highly unlikely to establish a breeding population on such prey. In contrast, the generalist 

mealybug predator Cryptolaemus montrouzieri (Mulsant) (Coleoptera; Coccinellidae), which was released 

in 1971 (St Helena Government, 2013) would be a greater threat to any indigenous mealybugs on St Helena, 

as it has a wide prey range (Kairo et al., 2012). However, any harm C. montrouzieri might do to putative 

indigenous mealybugs would have to be balanced against the beneficial effects it has by attacking exotic 

pest species of mealybugs on St Helena: these pest mealybugs severely damage crop and ornamental plants 

(St Helena Government, 2016) and critically endangered endemic plants such as Mellissia begoniifolia 

(Roxb.) Hook.f. (Solanaceae), the St Helena boxwood (Lambdon and Ellick, 2016).

WERE THERE UNANTICIPATED SYSTEM RESPONSES?

There have not been any specific surveys to assess whether the introduction of H. pantherina and the 

successful suppression of the exotic scale O. insignis have had any unexpected ecosystem effects. However, 

there have been substantial surveys of invertebrates as part of recent conservation initiatives, and none of 

these reports show any unexpected presence of the predator in native habitats in the absence of O. insignis 

prey (Mendel et al., 2008).

There were concerns that the weed L. camara might become more abundant once O. insignis was 

suppressed as this was one of the scale insect’s preferred host plants. Suppression of L. camara on St Helena 

has been suggested as an essential prerequisite for the natural regeneration of scrubwoods, Commidendrum 

rugosum (Dryand.) DC. (Asteraceae), following the near-eradication of goats (Ashmole and Ashmole, 

2000). Fortunately, other introduced biocontrol agents (e.g., lantana lacebug, Teleonemia scrupulosa 

Stål [Hemiptera: Tingidae]) appear to be maintaining L. camara under sufficient biological suppression 

(Ashmole and Ashmole, 2000; Jill Key, pers. comm.).

WAS INVADER SUPPRESSION SUFFICIENT FOR RESTORATION?

Reduction of O. insignis by H. pantherina was >95% (Figure 4-6), but this is probably an underestimate 

because it is hard in the field to distinguish live scales from fragments left after attack by the beetle (Fowler, 

2004). Indeed, rearing of H. pantherina ended on St Helena in 1995 because sufficient O. insignis could 

not be collected on the island to support the colony, presumably because biocontrol had been so successful 

(Fowler, 2004). More recently, occasional outbreaks of the scale on the island have been reported, but 

these outbreak patches are rapidly colonized by H. pantherina and the scale suppressed (Mendel et al., 

2008). The only exceptions to this successful suppression have been noted at some low-altitude restoration 

sites where tending ants appear to be protecting O. insignis from predation by H. pantherina, resulting in 

the scale persisting as a nuisance pest (Roger Key, pers. comm. 2013). After a visit to St Helena in 2015, 

Quentin Cronk, an authority of the endemic flora of St Helena (Cronk, 2000) commented “Orthezia is still 

about, as is Hyperaspis – the two seem to have settled down into a stable co-existence with Orthezia no 

longer a serious problem, just a minor nuisance. This is an extraordinary success story” (Q. Cronk pers. 

comm. 2015).

Chapter 4: Restoring Gumwood Forests on St Helena: Biocontrol of Orthezia Scale 33
Biological Control in Natural Areas

WERE OTHER MEASURES NEEDED?

Controls over domestic and feral stock, rabbits, and weeds have all been essential for protection of both 

naturally regenerating stands of Commidendrum and for planted stands established under the Millennium 

Forest Project, which aims to expand the area of natural forest on the island (Thompson, 2008; IUCN, 2011). 

Some invertebrates still cause minor problems on new plantings, such as mollusks, lepidopteran larvae, and 

the root mealybug Pseudococcus viburni Signoret (Hemiptera: Pseudococcidae), but O. insignis has never 

caused any damage to the Millennium Forest (Figure 4-7) (Rebecca Cairns-Wick, pers. comm., 2013). 

FUTURE OF SYSTEM

St Helena has a dismal history of failed conservation efforts stretching back to at least 1709 (Ashmole 

and Ashmole, 2000; Cronk, 2000). Recent conservation initiatives include the Millennium Forest (IUCN, 

2011), which is a substantial re-planting project aiming to re-create part of the Great Wood that once 

existed in mid-altitude parts of St Helena. The Millennium Forest, which predominately is comprised of 

restoration plantings of gumwood trees, currently covers 38 ha, with a total planned area of 250 ha. The 

main difference between the modern-era programs and past failures seems to be the current high level of 

community engagement, which was often markedly absent in the past (Ashmole and Ashmole, 2000; Cronk, 

2000; IUCN, 2011). For example, virtually every islander has purchased at least one tree to plant in the 

Millennium Forest (Thompson, 2008), and there are extensive volunteer and education programs focused 

on this restoration project (IUCN, 2011). Substantial financial support for this work has been contributed by 

the UK government (e.g., Darwin Initiative). After centuries of biodiversity loss on St Helena, interspersed 

by many failed conservation attempts, there finally seems some hope. Although continued pest and weed 

control efforts, as well as restoration planting, are still needed, many restorations are succeeding that would 

Figure 4-6. Control of Orthezia scale by the 

ladybird beetle Hyperaspis pantherina on 

the island of St Helena

34 Chapter 4: Restoring Gumwood Forests on St Helena: Biocontrol of Orthezia Scale
Biological Control in Natural Areas

have failed had biological control of O. insignis not been achieved (Fowler, 2004). To quote a past president 

of the St Helena National Trust “The introduction of Hyperaspis to St Helena and its biocontrol of Orthezia

is a conservation success story; we would be in a very different situation had Orthezia not been brought 

under control” (Rebecca Cairns-Wick, pers. comm. 2013).

ACKNOWLEDGMENTS

Rebecca Cairns-Wick, Quentin Cronk, Jill Key, and Roger Key provided up-to-date information from St 

Helena. The Agriculture and Forestry Dept., St Helena Government, funded and operated the project on St 

Helena. The project was started by CAB Bioscience using quarantine facilities licensed by the Ministry of 

Agriculture, Fisheries and Food, UK (Licence PHF 873A/667/73) and funding from the British Government 

(ODA Project OMC 92/94 349/053/001A). Landcare Research, New Zealand, provided funding for analysis 

and writing up.

a b

Figure 4-7. New restoration initiatives to expand gumwood stands are now possible due to suppression of Orthezia scale 

by the introduced biocontrol agent, allowing for recovery of surviving gumwoods (a) or planting of gumwoods in 

additional areas (b) (Credits: a,b Simon Fowler)

Chapter 4: Restoring Gumwood Forests on St Helena: Biocontrol of Orthezia Scale 35
Biological Control in Natural Areas

CHAPTER 5. PROTECTION OF WILIWILI TREES, 

THE FOUNDATION OF HAWAIIAN DRY FORESTS, 

THROUGH SUPPRESSION OF AN INVASIVE GALL WASP

L. V. Kaufman1 and J. Yalemar2

1

Department of Entomology, University of Hawaii 2

Hawaii Department of Agriculture

WAS THE INVADER A KEY SOURCE OF ECOSYSTEM DEGRADATION 

OR A THREAT TO NATIVE SPECIES?

The endemic tree Erythrina sandwicensis O. Deg. (Fabaceae), locally known as wiliwili, is one of the few 

native deciduous species in Hawaii (Rock, 1913). It is considered a keystone species of lowland dry forests, 

one of the most endangered habitats in the Hawaiian archipelago (Cabin et al., 2000). It grows on leeward 

slopes on all of the main islands, from sea level to 600 m in elevation (Wagner et al. 1990). 

This tree has faced many threats over the last several decades, which have reduced its abundance 

(Bruegmann, 1996). The most recent of these has been the invasion of the erythrina gall wasp (EGW), 

Quadrastichus erythrinae Kim (Hymenoptera: Eulophidae) (Figure 5-1a), which put E. sandwicensis at 

risk of extinction. The gall wasp was first reported from Reunion Island, Mauritius, and Singapore in 2003, 

and was formally described by Kim et al. (2004). The gall wasp was first reported on Oahu, Hawaii in April 

2005, and by August it was found on all the main Hawaiian Islands (Heu et al., 2006). The gall wasp was 

first observed infesting the ornamental and windbreak forms of the exotic species Erythrina variegata L. 

Both forms of this tree species were extremely susceptible to the gall wasp, and they suffered nearly 100% 

mortality statewide. The Department of Parks and Recreation of the City and County of Honolulu removed 

over 1,000 dead E. variegata trees (Vorsino, 2006). Soon thereafter, the gall wasp was reported infesting 

several other exotic species of Erythrina growing in botanical gardens around the islands (Messing et al., 

2009), as well as the endemic E. sandwicensis in both urban and natural areas. Nearly 100% of endemic 

wiliwili trees had some level of galling in urban and natural areas. Over 65% of shoots inspected during 

pre-release months were rated as heavily galled (infestation ratings of 2 or 3 out of a 4-point scale from 0-3). 

Female EGW insert their eggs into young plant tissues such as leaves, petioles, stems, flowers, and 

seedpods. Larval feeding causes galling of infested tissues (Figure 5-1b). Severe galling induces defoliation, 

36 Chapter 5: Protection of Hawaiian Wiliwili Trees through Suppression of an Invasive Gall Wasp
Biological Control in Natural Areas

a b

Figure 5-1. Adults of the erythrina gall wasp, Quadrastichus erythrinae (a); galled foliage of the wiliwili tree (b) (Credits: 

a,b Hawaii Department of Agriculture)

Chapter 5: Protection of Hawaiian Wiliwili Trees through Suppression of an Invasive Gall Wasp 37

which in turn reduces growth and plant vigor, and potentially causes tree mortality (Kim et al., 2004; Heu 

et al., 2006). Within months of the invasion of EGW, heavy infestations and tree mortality were observed 

in natural habitats, which caused great concern among conservationists and local communities. A statewide 

seed bank program was established in 2007 to preserve the species in the event of extinction in the field 

(Hollier, 2007). Concurrently, feasible short and long-term control methods were sought.

WERE OTHER CONTROLS FEASIBLE?

Initial efforts to slow the spread of the gall wasp included removal and disposal of infested material, but 

this was quickly recognized as impractical due to the small size of the wasp, its rapid dispersal, and the 

occurrence of overlapping generations. 

Applications of systemic insecticides, as trunk injections or soil drenches, were assessed for 

preservation of ornamental trees at irrigated urban sites (Xu et al., 2008). It was concluded that trunk 

injection with imidacloprid was the most effective treatment under those conditions. Daccola et al. (2009) 

tested the effectiveness of two stem-injected insecticides (imidacloprid and emamectin benzoate) in a 

natural, dryland site on Maui; they concluded that imidacloprid was the better treatment and that control 

lasted about one year. 

Insecticide treatments were used mainly in urban settings, golf courses, and botanical gardens. 

Occasionally this approach was used in natural areas. Even though pesticides were found to be effective 

in some cases, they were not considered useful as a cost-effective, long-term, self-sustaining option to 

suppress EGW in natural dryland areas in Hawaii. Consequently, a classical biological control program was 

initiated in 2005.

WHAT AGENTS WERE RELEASED AND WERE THEY SAFE?

The Hawaii Department of Agriculture (HDOA) and the University of Hawaii undertook a search for 

possible biological control agents in several parts of Africa, the most plausible area of origin of the gall 

wasp. The parasitoid Eurytoma erythrinae Gates and Delaware (Hymenoptera: Eurytomidae) (Figure 5-2a) 
Biological Control in Natural Areas

was reared from galled leaves of Erythrina abyssinica Lam. ex DC., which had been collected in Tanzania 

in 2006. A colony of this parasitoid was successfully established in quarantine in Hawaii.

Eurytoma erythrinae was also unknown to science at the time of collection and was formally described 

by Gates and Delvare (2008). Studies on the basic biology and host specificity of this species were done 

at the HDOA Insect Quarantine Facility. Females of E. erythrinae lay their eggs inside EGW galls, and 

parasitoid larvae feed externally on up to five EGW larvae or pupae to complete their development. 

Host specific studies examined potential risks to seven gall-forming non-target species: one Hawaiian 

endemic psyllid, four beneficial species (three tephritids and one eriococcid used for weed biological 

control), and two immigrant wasps (one agaonid and one eulophid) (HDOA, unpublished data). Studies 

concluded that E. erythrinae did not pose a risk to any of the non-target species tested, and the parasitoid 

was approved for release in late 2008, which occurred at several sites on all main Hawaiian Islands. 

a b

Figure 5-2. Two parasitoids have been studied for biocontrol of the erythrina gall wasp in Hawaii: (a) the eurytomid 

Eurytoma erythrinae and (b) the eulophid Aprostocitus nites (Credits: a,b Hawaii Department of Agriculture)

38 Chapter 5: Protection of Hawaiian Wiliwili Trees through Suppression of an Invasive Gall Wasp

WERE THERE UNANTICIPATED SYSTEM RESPONSES?

Eurytoma erythrinae quickly established in natural habitats and now occurs on all the main Hawaiian 

Islands. Pre- and post-release monitoring data show that this biocontrol agent has reduced infestation levels 

in foliage and improved the health of wiliwili trees in both urban settings and natural stands. Only a year 

after the release of the biocontrol agent, trees showed a significant reduction of infestation level in leaves, 

stems, and petioles. An increase in flowering was also observed after release. Nevertheless, galling of 

inflorescences is still high and reduces seed production (Figure 5-3). Galling is particularly high on young 

inflorescences and seedpods at drier sites on Maui and Hawaii. Monitoring showed that heavily infested 

young inflorescences die before blossoming. Germination studies show that seeds from heavily infested 

pods have lower germination rates compared to seeds from lightly infested or uninfested pods. 

WAS INVADER SUPPRESSION SUFFICIENT FOR RESTORATION?

A multi-site survey throughout the state done in 2012 found that approximately 30-35% of the wild wiliwili 

trees had been killed by the gall wasp, and little-to-no recruitment of seedlings was observed. 
Biological Control in Natural Areas

a b

Figure 5-3. Uninfested flowers of the wiliwili tree (a) versus a galled inflorescence (b) (Credits: a,b Hawaii Department of 

Agriculture)

Chapter 5: Protection of Hawaiian Wiliwili Trees through Suppression of an Invasive Gall Wasp 39

Release in 2008 of the parasitoid E. erythrinae reduced the risk of mortality for wiliwili trees and 

supported the continued use of the wiliwili tree in dryland restoration efforts in Hawaii. Nevertheless, 

these efforts are not yet self-sustaining because of the continued reduction in tree seed production due to 

the gall wasp, coupled with pre-existing threats to the tree such as habitat destruction, changes in the fire 

regime, seed predation by the bruchid Specularis impressithorax (Pic), competition with invasive weeds 

such as fountain grass, Pennisetum setaceum (Forssk.) Chiov., as well as severe browsing pressure by feral 

ungulates. The combined effect of these threats, together with poor control of the EGW in inflorescences 

may be preventing plant recruitment in natural areas in Hawaii. Monitoring showed that over 50% of 

seedpods collected only three weeks after maturity have seeds already infested with S. impressithorax eggs. 

Germination tests indicated that uninfested seeds have an average germination rate of about 45-50%. Seeds 

with at least one bruchid exit hole have less than 10% chance of germination, whereas seeds with two or 

more exit holes do not germinate. More research is needed to elucidate how these threats affect wiliwili 

recruitment in dryland forests in Hawaii. 

WERE OTHER MEASURES NEEDED?

Currently, E. eryhrinae is the only control agent that has been released in Hawaii against EGW. Nevertheless, 

it is evident that additional control measures are needed to prevent heavy galling of inflorescences 

and seedpods. Another biological control agent, the parasitoid Aprostocitus nites Prisloo and Kelly 

(Hymenoptera: Eulophidae) (Figure 5-2b), which has potential to provide additional control, is currently 

being considered for release. The HDOA has finished studies on the biology and host specificity on this 

prospective agent. Additional studies were also conducted to address any potential negative interactions of 

this second parasitoid with the first parasitoid, E. erythrinae. Results from these studies suggest that there is 

no risk to potential nontarget species and that there is no evidence for negative interactions between the two 

biocontrol agents. Results also suggested that the two agents together could provide additional protection 

to inflorescences and seedpods (Yalemar, pers. comm.). 

The HDOA has completed an Environmental Assessment for A. nites to propose its release into the 

environment, and concluded that this species poses no significant impacts to the environment. Pending state 

and federal approvals, this biocontrol agent could be released soon.
Biological Control in Natural Areas

FUTURE OF SYSTEM

The first biological control agent released against the gall wasp has provided enough control to prevent 

mortality of existing wiliwili trees in Hawaii. This measure alone, however, is not enough to restore selfsustaining tree populations. The second biological control agent is expected to provide better suppression 

of galling on inflorescences, which should increase production of viable seeds. This and the reduction of 

other threats to the wiliwili tree are needed to ensure the long-term survival of this endemic Hawaiian tree. 

ACKNOWLEDGMENTS

This study was funded by USDA T-STAR, USDA-APHIS (CAPS) and USDA Forest Service. We thank 

DLNR DOFAW staff, Rob Hauff and Cynthia King, for providing travel funds. We also thank the many 

field assistants that helped with data collection.

40 Chapter 5: Protection of Hawaiian Wiliwili Trees through Suppression of an Invasive Gall Wasp
Biological Control in Natural Areas

CHAPTER 6. CAN ASH COMMUNITIES AND THEIR 

DEPENDENT SPECIES BE PARTIALLY PROTECTED 

THROUGH BIOLOGICAL CONTROL OF 

EMERALD ASH BORER?

J. J. Duan

USDA ARS, BIIRU, Newark, Delaware, USA

WAS THE INVADER A KEY SOURCE OF ECOSYSTEM DEGRADATION 

OR A THREAT TO NATIVE SPECIES?

Ash trees (Fraxinus spp.) are important components in many hardwood forests in the United States and 

Canada (Federal Register, 2003; Nowak et al., 2003). There are 16 species of Fraxinus native to North 

America (Harlow et al., 1996; USGS, 2014), each adapted to different ecological habitats across a range 

of climates zones, soil types, and moisture gradients (Eyre, 1980). These diverse ash communities provide 

not only economic benefits, with US ash timber alone valued at $282 billion (Nowak et al., 2003), but 

also serve as food, cover, nesting sites, and habitat for mammals, birds, arthropods, and other organisms 

(Poland and McCullough, 2006; Gandhi and Herms, 2010; Koenig et al., 2013, Wagner and Todd, 2016). 

Any significant reduction in the abundance and/or diversity of the ash community would inevitably result in 

both economic losses (e.g., reduction of lumber production) and cascading ecological degradation in North 

American natural forests.

a b

Figure 6-1. Emerald ash borer, perhaps the most destructive invasive pest affecting North American forests today (a) adult; 

(b) typical gallery showing a mature larva digging into wood for pupation (Credits: a Debbie Miller, USDA Forest 

Service, bugwood.org; b Jian Duan)

Chapter 6: Protecting Ash Communities through Biological Control of Emerald Ash Borer 41
Biological Control in Natural Areas

Emerald ash borer (EAB), Agrilus planipennis

Fairmaire (Coleoptera: Buprestidae) (Figure 6-1a), 

is an invasive forest insect native to northeastern 

Asia (China, the Korean Peninsula, and the Russian 

Far East), whose larvae kill trees by tunneling in the 

phloem (Figure 6-1b). Since emerald ash borer was 

first detected in southern Michigan and Ontario in 

2002 (Haack et al., 2002), it has killed hundreds of 

millions of North American ash trees in 32 US states 

(Herms and McCullough, 2014; Emerald Ash Borer 

Information, 2016) and two Canadian provinces. 

While the economic cost of the EAB invasion is 

estimated to be $1 billion per year for the next 

decade (Kovacs et al., 2010), the ecological impacts 

of the EAB invasion on North American forests are 

already severe and widespread (Figure 6-2). EAB 

has killed 99% of healthy ash trees in some earlyinfested forests and has the potential to functionally 

extirpate ash from the continent (Klooster et al., 

2014). The loss of ash diversity and abundance in 

natural forests in the earliest-invaded regions (e.g., 

Midwestern and Mid-Atlantic States) has already 

harmed native plants and has the potential to 

affect populations of native, ash-dependent insects 

(Figure 6-3), as well as changing nutrient cycling and other ecological proecesses (Ulyshen et al., 2011, 

2012; Flower et al., 2013; Stephens et al., 2013; Gandhi et al., 2014; Jennings et al., 2016a). 

WERE OTHER CONTROLS FEASIBLE?

As discussed above, EAB is driving the ecological degradation in ash-dominated forests in North America, 

and taking no action against it is not a sensible option. Initial efforts to contain EAB damage focused on 

eradication of incipient populations from newly infested areas by destroying all infested trees within these 

areas, while at the same time restricting the movements of EAB-infested live trees or firewood in or out 

of the eradication area. However, the eradication effort was abandoned by 2009 because EAB populations 

in many infested areas (such as Michigan, Ohio, and Maryland) were already too high and too widespread 

(Herms and McCullough, 2014). Subsequently, efforts were shifted towards slowing the spread of EAB 

into new areas by regulatory restriction of movement of EAB-infested wood or plant materials, insecticide 

treatment of susceptible trees (either artificially girdled or naturally stressed as trap trees), and biological 

control via introduction and releases of natural enemies collected from EAB’s native range. Because it is 

impractical to treat all susceptible trees with insecticides in natural forests, classical biological control via 

establishment of introduced (non-native) natural enemies is needed to protect ash resources against EAB in 

North American (Bauer et al., 2015).

Figure 6-2. Dead and dying ash trees in heavily damaged 

forest in south central Michigan, showing the ecological 

impact of emerald ash borer (Credit: Jian Duan)

42 Chapter 6: Protecting Ash Communities through Biological Control of Emerald Ash Borer
Biological Control in Natural Areas

a b

Chapter 6: Protecting Ash Communities through Biological Control of Emerald Ash Borer 43

WHAT AGENTS WERE RELEASED AND WERE THEY SAFE?

Foreign exploration for natural enemies in northeastern China (part of the native range of the pest) began 

in 2003 (Liu et al., 2003, 2007) and was later extended into the Russian Far East (Duan et al., 2012a; Yao 

et al., 2016). Between 2003 and 2007, four species of hymenopteran parasitoids were found in northeastern 

China: (1) Sclerodermus pupariae Yang et Yao (Bethylidae), an ectoparasitoid of larvae, prepupae, and 

pupae (Wu et al., 2008, Yang et al., 2012); (2) Spathius agrili Yang (Braconidae), an ectoparasitoid of 

late instar larvae (Liu et al., 2003; Yang et al., 2005); (3) Oobius agrili Zhang and Huang (Encyrtidae), an 

egg parasitoid (Figure 6-4a) (Zhang et al., 2005; Liu et al., 2007), and (4) Tetrastichus planipennisi Yang 

(Eulophidae) (Figure 6-4b,c), an endoparasitoid of late instar larvae (Liu et al., 2003; Yang et al., 2006). 

Field surveys conducted in the Russian Far East between 2008 and 2012 found three additional species 

of hymenopteran parasitoids: (5) Spathius galinae Belokobylskij and Strazanac (Braconidae) (Figure 

6-4d), an ectoparasitoid of late instar larvae (Belokobylskij et al., 2012; Duan et al., 2012a); (6) Atanycolus 

nigriventris Vojnovskaja-Krieger (Braconidae), an ectoparasitoid of late instar larvae (Belokobylskij et al., 

2012; Duan et al., 2012a); and (7) Oobius primorskyensis Yao & Duan, a recently described new species of 

egg parasitoid (Yao et al., 2016). 

a b c d

Figure 6-3. Two of nearly 100 species of ash-dependent insects threatened by loss of ash: (a) the scarabaeid beetle Dynastes 

granti and (b) the sphingid moth Sphinx franckii (Credits: a Margarethe Brummerman; b Colin Gillette)

Figure 6-4. Parasitoids: (a) Oobius agrili, a parasitoid of emerald ash borer eggs imported from China; (b) Tetrastichus 

plannipenisi adult, an internal larval parasitoid of emerald ash borer imported from China; (c) brood of T. plannipenisi 

parasitoid larvae in gallery of host; (d) Spathius galinae, an external parasitoid of emerald ash borer larvae imported from 

the Russian Far East (Credits: a,c,d Jian Duan; b David Cappaert, bugwood.org) 
Biological Control in Natural Areas

Of these, the host range of S. pupariae was deemed too broad based on studies in China (Yang et 

al., 2012), and it was not considered further as a potential biological control agent. The other three Chinese 

parasitoids (T. planipennisi, S. agrili, and O. agrili) were proposed for introduction, and after extensive 

host range testing and safety evaluation, a petition for their release was approved in 2007 (USDA APHIS, 

2007). Separately, similar work was done with the Russian parasitoids S. galinae and O. primorskyensis. 

The petition for release of S. galinae was approved in 2015 and one for release of O. primorskyensis is 

under review (as of November, 2016). Although importation of A. nigriventris from the Russian Far East 

was considered, the difficulty in maintaining a colony of this species under quarantine laboratory conditions 

prevented conducting host range studies for further evaluation (JJD, unpublished data). 

The safety of Asiatic parasitoids petitioned for environmental releases in North America was 

assessed with data collected from both field surveys of other wood-boring insects in the parasitoids’ native 

range (China and Russian Far East) and quarantine testing with North American wood-boring insects. Data 

from these studies shows that host specificity of these Asian parasitoids in general is highly constrained by 

the phylogenetic relationship of nontarget hosts to their co-evolved host, A. planipennis (USDA APHIS, 

2007, 2015; Yang et al., 2008; Duan et al., 2015a). Field data from China and the Russian Far East show 

that these parasitoids do not attack other wood-boring insects in ash, such as bark beetles (Scolytinae) 

or longhorned beetles (Cerambycidae) (Yang et al., 2008; Duan et al., 2015a). However, host specificity 

studies in quarantine laboratories further showed that three of these introduced parasitoids – O. agrili, S. 

agrili, and S. galinae – do attack some North American Agrilus species (Table 6-1). In contrast to the attack 

on EAB, however, their attack rates in these nontarget Agrilus species is much lower, even under laboratory 

conditions that promote maximum parasitism (USDA APHIS, 2007, 2015; Yang et al., 2008; Duan et al., 

2012a, 2015a). Based on both laboratory and field host range studies, the potential non-target impact from 

introduction of these Asian parasitoids for EAB biocontrol, if any, would be only a low level of parasitism 

of some non-target Agrilus species in North America. 

WERE THERE UNANTICIPATED SYSTEM RESPONSES?

There are approximately 800 Agrilus species in North America, with 175 species in the United States 

(Bellamy, 2008; Nelson et al., 2008). It would be extremely difficult, if not impossible, to test every single 

North American Agrilus species with the introduced EAB parasitoids for safety assessment. It may be 

possible that these introduced EAB parasitoids (e.g., O. agrili and S. agrili) may occasionally attack some 

Table 6-1. Numbers of non-target insect taxa tested with Asian parasitoids petitioned in North America for environmental 

release against emerald ash borer (Agrilus planipennis). 

Parasitoids Insect

orders

Families Non-target

species

Non-target

Agrilus

Non-target species

attacked

Oobius agrili 1 2 6 18 6 3/18

Tetrastichus planipennisi 1 3 6 13 5 0/13

Spathius agrili 1 2 6 19 9 4/19

Spathius galinae 2 3 6 15 6 1/15

1

 Data compiled from USDA APHIS (2007)

2

 Data compiled from USDA APHIS (2015) and Duan et al. (2015a)

44 Chapter 6: Protecting Ash Communities through Biological Control of Emerald Ash Borer
Biological Control in Natural Areas

of the non-target Agrilus species in North America. However, recent host-finding studies also show that 

these parasitoids are strongly attracted to volatiles from ash trees (Yang et al., 2008; Johnson et al., 2014), 

indicating their strong affinities to the host’s food plants. Thus, it can be reasonably predicted that the level 

of these attacks of non-ash feeding Agrilus species, if any, would be limited. Field surveys of non-target 

insects associated with ash trees following field releases of introduced EAB parasitoids in Maryland have 

not found any evidence of non-target attack from these introduced parasitoids (Jennings et al., 2016a,b; JJD 

unpub. data). In contrast, arthropod diversity associated with ash trees is significantly reduced because of 

EAB invasion in Maryland (Jennings et al., 2016b). Therefore, introduction of EAB parasitoids from Asia 

is not likely to cause any unexpected, negative system responses, but may be expected to induce positive 

system responses from the protection of ash trees against EAB infestation in North America. 

WAS INVADER SUPPRESSION SUFFICIENT FOR RESTORATION?

Following regulatory approval, environmental releases of the three Chinese parasitoids (S. agrili, T. 

planipennisi, and O. agrili) began with very small numbers (a few hundred per species) in 2007 in Michigan, 

the epicenter of EAB invasion. However, large releases (tens of thousands per species) did not start until 

2010 when the USDA APHIS mass-rearing facility was established in Brighton, Michigan (Gould et al., 

2015). By the summer of 2016, 24 EAB-infested states had received releases of one or more species of these 

agents (MapBiocontrol, 2016). All three biocontrol agents have been recovered following their releases in 

most release areas in the U.S.; however, only T. planipennisi and O. agrili have been consistently recovered 

more than one year following their release. 

At six study sites in Michigan, EAB egg parasitism by O. agrili increased from 1-4% in the first few 

years (2007–2011) during or after their release (which occurred thru 2010) to 20-28% in 2014 (Abell et al., 

2014). However, the role of the introduced egg parasitoid O. agrili in suppressing EAB population growth 

remains to be determined because of the difficulty of measuring EAB egg densities under the loose bark 

or bark crevices of infested ash trees in the field. In contrast, T. planipennisi appears to have spread much 

faster and became a dominant larval parasitoid of EAB in the Michigan study sites shortly after its release 

(Duan et al., 2012b, 2013, 2015b). The parasitism rate of EAB larvae by T. planipennisi increased from 

1.0-5.6% in the first three years (2008-2011) after the start of field releases to 26-30% in both release and 

control sites four or five years later (by 2014). Recent life table analyses of seven years of data collected 

from the six Michigan study sites following the releases of these biocontrol agents show that parasitism by 

T. planipennisi has contributed significantly to the reduction of net EAB population growth rates in these 

sites since 2012, approximately four years after the initial release (Duan et al., 2015b). With additional 

mortality from woodpeckers and native parasitoids (primarily braconids in the genus Atanycolus), EAB 

larval densities declined nearly 90% in infested ash trees at both parasitoid-release and control plots 

between 2009 and 2014 (Duan et al., 2015b). Findings of the life table study in Michigan further indicate 

that successful biocontrol of EAB may involve suppression of EAB abundance both by local, oligophagous 

natural enemies (such as Atanycolus spp.) and by introduced specialist parasitoids (such as T. planipennisi) 

(Figure 6-5). 

However, the effectiveness of T. planipennisi in protecting ash trees against EAB is affected by bark 

thickness, which increases with tree size. Data from field surveys in T. planipennisi’s native range (China 

and Russian Far East) and in the U.S. show that parasitism by T. planipennisi is concentrated in smaller ash 

trees (Liu et al., 2007; Duan et al., 2012a; Jennings et al., 2016b). A recent survey of randomly selected ash 

saplings (2-5 cm DBH [Diameter at Breast Height]) at the Michigan study sites found that T. planipennisi

was the dominant mortality factor, causing 36-85% parasitism of older (3rd-JL) EAB larvae and that this 

Chapter 6: Protecting Ash Communities through Biological Control of Emerald Ash Borer 45
Biological Control in Natural Areas

parasitoid protects ash saplings effectively from attack by EAB (Duan et al., 2017). The reason for this 

higher parasitism rate in smaller trees is the better match between the parasitoid’s ovipositor length (average 

2-2.5 mm) and bark thickness, which prevents wasps from attacking EAB larvae in the lower boles of trees 

that exceed 12 cm DBH and a bark thickness of 3.2 mm (Abell et al., 2012). 

To control EAB successfully in larger ash trees, efforts are currently focused on establishing the 

Russian parasitoid S. galinae in North America. This braconid has a longer ovipositor (4-6 mm) than T. 

planipennisi, and it can attack EAB larvae feeding on trees with DBHs up to 25-30 cm (Duan et al., 

2012a). Release of S. galiane in the U.S. began in the summer of 2016, and more time is needed to assess 

its effectiveness in protecting large ash trees. In theory, highly effective egg parasitoids from EAB’s native 

range might also be able to protect both small and large ash trees against EAB, as they can kill the pest 

before larvae bore into the ash cambium. However, the current level of egg parasitism by O. agrili (<29%) 

by itself is not sufficient to protect ash trees. Introduction of other egg parasitoids, such as O. primorskyensis 

may enhance egg parasitism (Larson and Duan, 2016).

WERE OTHER MEASURES NEEDED?

Host tree resistance can have significant effects on EAB survival and development, and hence population 

growth rates. Thus, levels of ash tree resistance can also affect the degree of success of EAB biocontrol. 

Field studies in the pest’s native range show that EAB is rarely a serious pest on many Asian ash species 

such as Fraxinus mandschurica Rupr. and Fraxinus rhynchophylla Hance, largely because of putative tree 

resistance as well as the prevalence of EAB egg and larval parasitoids (Liu et al., 2007; Duan, 2012a; Wang 

et al., 2016). In North America, the common species of Fraxinus, such as F. pennsylvanica Marshall, F. 

americana L., and F. nigra Marshall, have little resistance to EAB and are readily killed by even moderate 

EAB infestations (Rebek et al., 2008; Tanis and McCullough, 2015). However, some North American ash, 

such as F. quadrangulata Michx., appear to have largely escaped EAB infestation in Michigan (Tanis and 

McCullough, 2012). In addition, some “lingering” healthy (often younger) trees of the susceptible ash 

Figure 6-5. Reduction in emerald 

ash borer intensity (rate of 

attack on surviving trees) over 

a seven year period at 12 plots 

in Michigan, reflecting lower 

attack rates by the borer. Arrows 

represent the time and number 

of parasitoid releases, with big 

arrows being major (>11,00 

wasps) and small arrows being 

minor (<700 wasps) releases 

(from Duan et al., 2015)

46 Chapter 6: Protecting Ash Communities through Biological Control of Emerald Ash Borer
Biological Control in Natural Areas

species (F. pennsylvanica and F. americana) have been frequently observed to survive EAB infestation in 

the epicenter of EAB invasion such as Michigan and Ohio (Tanis and McCullough, 2012). A population 

dynamics model parameterized with combined parasitism rates of 60% parasitism of EAB eggs and 30% 

larval parasitism shows that EAB population growth rate in North America can be reduced to the belowreplacement level (<1) when accompanied with moderate to high levels of host plant resistance (Duan 

et al., 2015c). Together, these observations provide hope that some level of host tree resistance in North 

American ash may facilitate biocontrol in protecting ash and their dependent communities in North America. 

Resistance in North American ash currently is low but may be higher in the future ash forest due to natural 

selection (selective survival of partly resistance ash) or outplanting of lines of ash bred for higher levels of 

resistance (see review in Villari et al., 2015).

FUTURE OF SYSTEM

Following its accidental introduction into the U.S, probably in the 1990s (Siegert et al., 2014), EAB 

continues to spread and degrade ash communities in North America. The EAB biocontrol program, started 

nearly a decade ago via introduction and establishment of co-evolved natural enemies from the pest’s native 

range, appears to hold promise to protect North American ash communities. To date, this program has 

successfully established the egg parasitoid O. agrili and the larval parasitoid T. planipennisi introduced from 

China, in most of the areas in the U.S. where these species have been released. While the role of O. agrili

in reducing EAB population growth requires continued evaluation, the larval parasitoid T. planipennisi

currently plays a significantly suppressive role in saplings and smaller trees (DBH <12 cm) in aftermath 

forests in Michigan (Duan et al., 2015b, Duan et al., 2017). This suppressive effect is likely to spread 

geographically as populations of O. agrili and T. planipennis increase in density in more EAB-infested 

areas where they have been released, protecting saplings and small trees. To protect larger trees, however, 

establishment of more effective egg parasitoids and larval parasitoids with long ovipositors is needed. The 

most recent introduction of S. galinae from the Russian Far East, which has a much longer ovipositor (4-6 

mm) than T. planipennisi, is one such species. Initial recoveries in New York and Connecticut have been 

made at promising levels (2016, RGVD and JJD, unpub. data). As EAB densities decline, future studies will 

be required on (1) growth and survival rates of ash in aftermath forests with self-sustaining and increasing 

biocontrol agent populations, (2) persistence of key ash-dependent invertebrates (Wagner and Todd, 2016), 

and (3) repetition of basic population work in new regions as EAB invades southern and western areas in 

the United States, which will be ecologically distinct from the north central and northeastern US areas on 

which this discussion is based. 

ACKNOWLEDGMENTS

Roy Van Driesche (University of Massachusetts) provided constructive comments and edits on the 

manuscript. Collaborative work with Leah S. Bauer (U.S. Forest Service), Juli Gould (USDA APHIS), 

David Jennings and Paula Shrewsburry (both at University of Maryland), Scott Salom (Virginia Tech), 

Clifford Sidoff (Purdue University), Deborah McCullough (Michigan State University), and Roy Van 

Driesche and Joe Elkinton (both at University of Massachusetts) motivated the write-up of this manuscript. 

Many technicians and student helpers from the University of Massachusetts, University of Maryland, 

Michigan State University, and USDA Agricultural Research Service, Forest Services, and APHIS have 

assisted in the emerald ash borer control work reviewed in this paper.

Chapter 6: Protecting Ash Communities through Biological Control of Emerald Ash Borer 47
Biological Control in Natural Areas

Figure 7-1. Waterhyacinth, Eichhornia crassipes, is a South American plant that has 

become invasive worldwide in tropical and subtropical standing or slow moving 

water, shown here at New Year’s Dam in South Africa (Credit: Julie Coetzee)

48 Chapter 7: Biological Control of Waterhyacinth Restores Ecosystem Function in South Africa

CHAPTER 7. BIOLOGICAL CONTROL OF WATERHYACINTH 

RESTORES ECOSYSTEM FUNCTIONING TO AN 

IMPOUNDMENT IN SOUTH AFRICA

M. P. Hill1 and J. A. Coetzee2

1

Department of Zoology, 2Department of Botany; Rhodes University, Grahamstown South Africa 

Rhodes University, Grahamstown South Africa

WAS THE INVADER A KEY SOURCE OF ECOSYSTEM DEGRADATION 

OR THREAT TO NATIVE SPECIES?

Eichhornia crassipes (Mart.) Solms-Laub. (Pontederiaceae) (waterhyacinth) (Figure 7-1), from South 

America, has been considered the most damaging aquatic macrophyte in many of the rivers, manmade 

impoundments, lakes, and 

wetlands of South Africa 

(Hill, 2003). The lack of 

natural enemies, coupled with 

eutrophication of many aquatic 

habitats in South Africa, has 

facilitated the invasion by 

this weed. These nutrientenrichment disturbances are 

mainly anthropogenic, including 

agricultural activities and an 

increase in urbanization due to 

accelerated human population 

growth. The management of E. 

crassipes requires site-specific 

applications where priorities 

depend on the prevailing 

climate, the nature and use of 
Biological Control in Natural Areas

the water body in question, extent of the infestation, and the resources available for control activities. Here 

we present a case study in which we discuss the biodiversity impacts of E. crassipes on the environment, 

including direct effects on aquatic biodiversity and indirect effects of increased evapotranspiration leading 

to greater water loss, and how these impacts were lessened by biological control of the plant.

Our case system is that of the plant’s impact on a small water body created by the New Year’s Dam, 

behind which there was an 80.8 ha impoundment. This site is near the town of Alicedale in the Eastern 

Cape Province of South Africa. This impoundment predominantly supplies water to the local population for 

domestic use and irrigation (Doudenski, 2004). Waterhyacinth was first recorded on the reservoir in 1988, 

and by 1990 the reservoir was 80% covered (Figure 7-2a), affecting water removal and recreational fishing 

(Hill, 2003), but the infestation’s impact on biodiversity and functioning of the system was largely unknown.

The impacts of E. crassipes are considerable and have been well reviewed (Villamagna and Murphy, 

2010). While the direct and indirect (cost of control) economic impacts of E. crassipes are readily measured 

(Mailu, 2001), its biodiversity impacts are more difficult to define, and very few studies have attempted to 

do so (Villamagna and Murphy, 2010). At New Year’s Dam, Midgley et al. (2006) showed that infestations 

of E. crassipes significantly reduced aquatic benthic biodiversity in comparison to waterhyacinth-free 

areas. This study, carried out over some 18 months, showed that the number of families and the number 

of individuals of benthic invertebrates were significantly lower under the mats of E. crassipes than areas 

where the weed had been controlled. Further, standard measures of biodiversity, including the ShannonWeiner Diversity Index, Margalef’s Richness Index, Pielou’s Evenness Index, and chlorophyll a were all 

significantly lower under waterhyacinth mats than in waterhyacinth-free zones. Eichhornia crassipesinfested areas were characterized by poor water quality specialists such as leeches and midge larvae, while 

E. crassipes-free zones had a significantly higher abundance and diversity of species that are less tolerant 

of poor water quality, such as mayflies, damselflies, and dragonflies. 

WERE OTHER CONTROLS FEASIBLE?

Mechanical control or the manual removal of E. crassipes is invariably the initial control option tried. 

While manual removal using rakes and pitchforks can be successful for really small infestations, it is labora b

Figure 7-2. A small reservoir in South Africa (New Year’s Dam) completely covered by waterhyacinth before biological control 

in 1990 (a) and the same site in 1999 after introduction of a biological control agent suppressed weed density (b) (Credits: 

a,b Martin Hill)

Chapter 7: Biological Control of Waterhyacinth Restores Ecosystem Function in South Africa 49
Biological Control in Natural Areas

intensive and ineffective in larger infestations. Mechanical harvesters have been used on some of the larger 

rivers, dams, and wetlands in South Africa, and, while harvesters can be successful, they are expensive 

(Hill and Olckers, 2001). These control options are, however, impractical for infestations larger than 

about one hectare because of the weed’s rapid rate of increase and the fact that it is more than 90% water 

(Hill and Cilliers, 1999). Herbicidal control is widely used in South Africa, commonly with application 

of glyphosate, glyphosphate trimesium, diquat, or terbutryn. Herbicidal control of waterhyacinth depends 

on skilled operators who maintain a long-term follow-up program to control re-infestation by scattered 

plants and those germinating from seed. Therefore, any herbicide program against the weed requires a 

commitment to an ongoing operation of unlimited duration. It is the lack of follow-up that often causes 

herbicidal programs to fail (Hill et al., 2012). 

WHAT AGENTS WERE RELEASED AND WERE THEY SAFE?

The biological control program against waterhyacinth in South Africa began in 1973. Nine biological control 

agents have been introduced in the country: (1) the weevil Neochetina eichhorniae Warner (Figure 7-3) in 

1974 (Cilliers, 1991); (2) the fungus Cercospora piaropi Tharp. (Mycosphaerellales: Mycosphaerellaceae) 

in 1987 (Coetzee et al., 2011); (3) the galumnid mite Orthogalumna terebrantis Wallwork in 1989; (4) the 

weevil Neochetina bruchi Hustache in 1990; (5) the crambid moth Niphograpta albiguttalis Warren (= 

Sameodes albiguttalis [Warren]) in 1990; (6) the mirid bug Eccritotarsus catarinensis (Carvalho) in 1996; 

(7) the mirid bug Eccritotarsus sp. in 2007 (Paterson et al., 2016), (8) the acridid grasshopper Cornops 

aquaticum (Brüner) in 2011 (Winston et al., 2014); and (9) the delphacid Megamelus scutellaris Berg in 

2013. All these agents have been subjected to extensive host-specificity testing in several countries around 

the world (Center et al., 2002; Coetzee et al., 2009), and to date no non-target effects have been recorded 

(Louda et al. 2003) despite the occurrence of closely related species in all the countries where these agents 

have been released.

Even though there were five biological control agents available for release on E. crassipes in South 

Africa at the time of the infestation on New Year’s Dam, only the weevil N. eichhorniae was introduced. 

In January 1990, 200 adults were released, with 1000 more in November of the same year. In contrast, 

Conlong et al. (2009) suggested that 100,000 adult Neochetina weevils were needed per hectare to suppress 

E. crassipes in 2-4 years. Annual field trips were made to New Year’s Dam from 1990 to 2016 to evaluate 

the impact of the biological control. By 1994, weevils had spread throughout the infestation on the reservoir, 

a b

Figure 7-3. The weevil Neochetina eichhorniae, the agent responsible for biological control of waterhyacinth at the New 

Years Dam site in South Africa: (a) adult; (b) adult feeding damage (Credits: a Anthony King; b Julie Coetzee)

50 Chapter 7: Biological Control of Waterhyacinth Restores Ecosystem Function in South Africa
Biological Control in Natural Areas

and they had reduced E. crassipes coverage to <20%, and average plant height had been reduced from 60 

to 20 cm (Hill and Cilliers, 1999). The percentage cover of the weed on the impoundment was less than 5% 

by 2000 and has remained at this level (Figure 7-2b) (Coetzee et al., 2011). 

WERE THERE UNANTICIPATED SYSTEM RESPONSES?

No unexpected, harmful ecosystem responses to biological control of the invasive plant were observed.

Results were beneficial to biodiversity and the general habitat. Indeed, the only unexpected ecosystem 

response was the rapidity of control. Successful biological control of E. crassipes usually takes 3-5 years in 

tropical areas (Julien et al., 1999), but considerably longer periods are usually required (8-10 years) under 

more temperate climates such as that at New Year’s Dam (Hill and Olckers, 2001). Thus, the complete 

control of the weed on New Year’s Dam in four years was unexpected, especially given the low inoculation 

density of the agent used. This rapid control has been ascribed to the fact that the system is oligotrophic 

(see below).

WAS INVADER SUPPRESSION SUFFICIENT FOR RESTORATION?

The significant reduction of E. crassipes coverage on New Year’s Dam has resulted in the recovery of 

the benthic invertebrate community (Midgley et al., 2006), and it has also resulted in water savings for 

the area. Invasive alien plants have significant environmental and ecological impacts in South Africa, 

especially in habitats such as riparian zones and freshwater systems (Chamier et al., 2012). The increased 

evapotranspiration rates associated with invasive alien plants are estimated to result in a loss of about 

3,300 million m3 of water (about 7% of the national total) (Le Maitre et al. 2000), and this impact was the 

main reason for the establishment of the Working for Water Program (WfW) in 1995, a national program 

created to reduce the impacts of alien invasive plants on South Africa’s critical water resources, while 

creating jobs at the same time. Eichhornia crassipes increases water loss from aquatic ecosystems up 

to 40% (Penfound and Earl, 1948). More recently, Fraser et al. (2016) conducted analysis of the water 

loss saving due to the biological control program on New Year’s Dam. Published estimates of the water 

loss due to evapotranspiration from E. crassipes vary significantly, and therefore the study used three 

different rates from the literature, a high, medium and low rate (Allen et al., 1997). It is estimated that 

without biological control, and thus 80% plant cover of the impoundment’s surface, the total annual 

evaporation rate of the impoundment would vary between 1,477,946 m3 and 6,545,188 m3, depending on 

the evapotranspiration rate estimate used. However, this water loss significantly decreased with biological 

control once plant cover was reduced to 5% of the impoundment’s surface, with the total evaporation 

rate estimated to be between 1,082,067 m3 and 1,398,770 m3 for the low and high evapotranspiration rate 

estimates, respectively, and thus a water loss saving of nearly 400,000 m3 at the low evapotranspiration 

rate and 5 million m3 at the high rate. 

FUTURE OF THE SYSTEM

Many studies have shown that the successful biological control of E. crassipes is often limited if water is 

eutrophic, and the first step towards control should be to reduce nutrient inputs into the system (Coetzee 

et al., 2007; Coetzee and Hill, 2012). A long-term monitoring program assessing the factors affecting the 

Chapter 7: Biological Control of Waterhyacinth Restores Ecosystem Function in South Africa 51
Biological Control in Natural Areas

52 Chapter 7: Biological Control of Waterhyacinth Restores Ecosystem Function in South Africa

success of biological control of E. crassipes in South Africa suggested that additional control measures are 

needed in systems where water phosphorus levels are above 0.1 mg l-1 (Byrne et al., 2010). Unfortunately, 

biological control efforts of E. crassipes in South Africa are hampered by the adoption of an excessively 

high water P regulatory standard of 1 mg l-1 for all water treatment works (DWAF, 1996). Every invaded 

system in the country is characterized by P levels greater than this level, all except New Year’s Dam (Byrne 

et al. 2010). Thus, the success of biological control at New Year’s Dam using a single agent can be attributed 

to the fact that the level of P in the water has remained below 0.1 mg l-1 (Byrne et al. 2010).

ACKNOWLEDGMENTS

This research was funded through the Department of Environmental Affairs, Natural Resource Management 

Programme’s “Working for Water Programme.” Further funding for this work was provided by the South 

African Research Chairs Initiative of the Department of Science and Technology and the National Research 

Foundation of South Africa. Any opinion, finding, conclusion, or recommendation expressed in this material 

is that of the authors, and the National Research Foundation does not accept any liability in this regard.
Biological Control in Natural Areas

CHAPTER 8. THE ROLE OF MELALEUCA CONTROL 

IN EVERGLADES RESTORATION: ACCOMPLISHMENTS 

AND FUTURE PLANS

E. C. Lake1, P. W. Tipping1, M. B. Rayamajhi1, P. D. Pratt2, F. A. Dray Jr.1, 

G. S. Wheeler1, M. F. Purcell3, and T. D. Center1

1

USDA ARS Invasive Plant Research Laboratory in Fort Lauderdale, Florida,

2

USDA ARS Exotics and Invasive Weeds Research Unit in Albany, California,

3

USDA ARS Australian Biological Control Laboratory in Brisbane, Australia

WAS THE INVADER A KEY SOURCE OF ECOSYSTEM DEGRADATION 

OR A THREAT TO NATIVE SPECIES?

The Australian tree Melaleuca quinquenervia (Cav.) S.T. Blake (Myrtales: Myrtaceae) was first introduced 

to south Florida in 1886 (Dray et al., 2006). The tree was sold by nurseries, planted as agricultural 

windbreaks and forestry crops, and used to stabilize levees (Bodle et al., 1994; Dray et al., 2006). Melaleuca 

quinquenervia, hereafter melaleuca, flowers primarily in the fall and winter months, although it can flower 

at other times. Flowering spikes are produced on branches that vary between growth and reproductive 

phases, resulting in an alternating pattern of leaves and flowers or persistent woody seed capsules on a 

given branch. These seed capsules contain a great many small seeds that are released when the capsule dries 

(Tomlinson, 1980). Rayachhetry et al. (1998) estimated that a 21-m tall tree could produce 51 million seeds, 

of which 15%, or 5.6 million, would likely be viable. 

Melaleuca was first reported as naturalized in Florida in the mid-1920s (Dray et al., 2006). It invades 

a variety of areas from disturbed sites such as the sides of roads and ditches to natural habitats, including 

cypress stands, sawgrass marshes, pine flatwoods, and mesic prairies (Bodle et al., 1994). Melaleuca can 

outcompete native plants and form dense monocultures (Figure 8-1), changing the vertical structure of 

communities and converting wetland prairies and marshes into melaleuca forests (Gordon, 1998; Turner et 

al., 1998). Melaleuca can also alter fire frequency and intensity, affect water table depth and surface flow, 

and modify above and belowground ecosystem components (Gordon, 1998; Serbesoff-King, 2003; Martin 

et al., 2009; Center et al., 2012). Laroche and Ferriter (1992) estimated that a small melaleuca population 

could grow exponentially to infest 95% of a 250 ha patch in just 25 years. By 1998, melaleuca had invaded 

nearly 400,000 hectares in Florida (Laroche 1998).

Chapter 8: Melaleuca Control in Everglades Restoration: Accomplishments and Future Plans 53
Biological Control in Natural Areas

Figure 8-1. Melaleuca quinquenervia can displace native vegetation in invaded wetlands and 

forests in Florida (Credit: Paul D. Pratt, USDA ARS)

54 Chapter 8: Melaleuca Control in Everglades Restoration: Accomplishments and Future Plans

The hydrology of the Everglades historically consisted of a dry season from November to April 

and a wet season from May through October (Rader and Richardson, 1992). However, this pattern has 

been largely reversed by drainage and water conservation programs, which discharge water after high 

rainfall events and hold it during dry periods (Kushlan, 1987; Rader and Richardson, 1992). Changes in 

hydroperiod can reduce native flora and fauna and increase susceptibility to invasive species, including 

melaleuca, which can then cause additional changes in ecosystem structure and function (Myers, 1975; 

Kushlan, 1987; Gordon, 1998; Turner et al., 1998 and references therein). 

The spread of melaleuca and its ecological damage prompted the formation of a Melaleuca Task 

Force in 1990. This group included local, state, and federal scientists and land managers and was charged 

with developing a comprehensive plan to manage melaleuca. The first plan was issued in 1990 and it has 

been periodically updated (Laroche, 1999). 

WERE OTHER CONTROLS FEASIBLE?

Removing melaleuca with heavy equipment can be effective in rights-of-way for canals and utilities, 

but stumps must be treated with herbicide to prevent re-growth. Mechanical removal causes too much 

disturbance and non-target damage to be useful in natural areas, where hand removal of seedlings is the only 

appropriate mechanical tool (Laroche, 1998). Herbicides can provide effective control of melaleuca. Hack 

and squirt and cut-stump treatments are the most effective way to treat individual trees, but these techniques 

are labor intensive. Aerial applications are more appropriate for large-scale infestations (Laroche, 1998). 

Melaleuca is adapted to fire, which will kill small saplings but not large trees (Turner et al., 1998). 

Melaleuca responds to stress from herbicides, physical damage, fire, frost, or desiccation by 

synchronously releasing seed held in the capsules (Meskimen, 1962). This can lead to establishment of 

dense stands of melaleuca seedlings. These seedlings outcompete native vegetation, and after a fire, they 

can promote the conversion of sawgrass prairies (Cladium jamaicense Crantz) (Cyperales: Cyperaceae) into 

melaleuca forests (Turner et al., 1998). Thus, follow-up treatments are needed to eliminate the seedlings 

before they reproduce, which can occur in less than three years (Laroche, 1998; Meskimen, 1962). 
Biological Control in Natural Areas

WHAT AGENTS WERE RELEASED AND WERE THEY SAFE?

Development of a biological control program to complement the removal of standing biomass with 

herbicidal, mechanical, and physical controls was a high priority for the Melaleuca Task Force (Laroche, 

1998; Center et al., 2012). The goals of the biological control program were to find agents that were hostspecific to the genus Melaleuca, or below, that would limit stand regeneration by reducing flowering or 

seed production (targeting inflorescences or seeds) or increasing mortality of seedlings, saplings, or stump 

sprouts (Center et al., 2012).

Three agents have established on melaleuca in Florida. Releases of the first agent, Oxyops vitiosa

Pascoe (Coleoptera: Curculionidae), began in 1997 (Figure 8-2a). Adults and larvae feed on melaleuca 

leaves; larvae cause extensive damage, but prefer young foliage and are most common in the field from 

October through May (Purcell and Balciunas, 1994; Center et al., 2000). The weevil pupates in the soil and 

does not establish at permanently flooded sites (Center et al., 2000). The melaleuca psyllid, Boreioglycaspis 

melaleucae Moore (Hemiptera: Psyllidae) (Figure 8-2b), feeds on melaleuca leaves but will also attack 

stem tissue if population densities are high (Wineriter et al., 2003). This phloem-feeder completes its entire 

life cycle on the plant and is not affected by site hydrology (Wineriter et al., 2003; Center et al., 2006). 

The psyllid was first released in 2002 and it established at all release sites (Center et al., 2006). The third 

agent, Lophodiplosis trifida Gagné (Diptera: Cecidomyiidae), is a stem-galling midge that disrupts apical 

and axillary growth of melaleuca (Figure 8-2c). It completes its entire life cycle within galls on melaleuca 

and is most active during the wet-warm season and in wet habitats in the native range (Pratt et al., 2013). 

Releases of L. trifida began in 2008, and it established readily and dispersed rapidly (Pratt et al. 2013). The 

ecological host ranges of all three agents were validated in field experiments after their release, confirming 

that non-target species in Florida were at no risk of attack (Center et al., 2007; Pratt et al., 2009, 2013). 

WERE THERE UNANTICIPATED SYSTEM RESPONSES?

Melaleuca compensates for damage by O. vitiosa by producing replacement foliage and new stems, but this 

comes at the cost of reproduction. One year following O. vitiosa herbivory, damaged trees had only 20% as 

a b c

Figure 8-2. (a) The melaleuca weevil, Oxyops vitiosa, was the first biocontrol agent released against melaleuca. Feeding by weevil 

adults and larvae on melaleuca leaves decreases melaleuca’s reproductive capacity; (b) the melaleuca psyllid, Boreioglycaspis 

melaleucae, is a phloem feeder that attacks melaleuca leaves and stems; (c) the stem-galling midge Lophodiplosis trifida

disrupts apical and axillary growth of melaleuca (Credits: a Ellen C. Lake, USDA ARS; b,c Paul D. Pratt, USDA ARS)

Chapter 8: Melaleuca Control in Everglades Restoration: Accomplishments and Future Plans 55
Biological Control in Natural Areas

many reproductive structures as undamaged trees (Pratt et al., 2005). However, in permanently wet habitats, 

melaleuca escaped persistent attack by the weevil due to insufficient dry substrate for pupation (Center et 

al., 2000). Boreioglycaspis melaleucae attacks both young and old melaleuca leaves, and leaves damaged 

by the psyllid are 4.7 times more likely to fall off than are undamaged leaves (Morath et al., 2006). Trees 

attacked by O. vitiosa and B. melaleucae produced only 1% of the seed formed by control trees protected by 

insecticide treatments (Tipping et al., 2008). High densities of psyllids also decreased growth and increased 

mortality of melaleuca seedlings (Franks et al., 2006). Tipping et al. (2012) monitored recruitment and 

survival of melaleuca seedlings in a seasonally inundated wetland and documented negative population 

growth of melaleuca at the site. The decline in this population was more likely due to reduction in seed 

inputs due to agent herbivory than to seedling mortality (Tipping et al., 2012). An insecticide check study 

that limited herbivory on melaleuca saplings demonstrated that galling by the newest agent, L. trifida, 

significantly reduced above- and below-ground biomass (Tipping et al., 2016). 

WAS INVADER SUPPRESSION SUFFICIENT FOR RESTORATION?

Melaleuca invasion alters above- and below-ground ecosystem components, including decreasing nutrient 

storage and the concentration of nutrients available in the litter layer. These changes may impede succession 

of the native plant community after melaleuca is removed (Martin et al., 2009). Rayamajhi et al. (2009) 

monitored mature melaleuca stands in seasonally flooded and non-flooded sites from 1997, the year the 

first biological control agent was released, through 2005. A two to four-fold increase in plant diversity was 

correlated with declines in melaleuca within these sites (Figure 8-3), with greater decreases in melaleuca 

density and increases in plant diversity at the non-flooded sites (Rayamajhi et al. 2009). Foliar damage by 

the biological control agents may have accelerated thinning of the melaleuca canopy and increased light 

penetration to the forest floor, thus facilitating native plant recovery (Rayamajhi et al., 2009).

In an insecticide-check field study, population-level suppression of melaleuca by O. vitiosa and 

B. melaleucae was documented 

(Tipping et al., 2009); however, 

there was no significant recovery of 

the native plant community. Several 

factors, including reduced nutrient 

levels, reduced soil moisture due to 

a reduction in leaf litter, or distance 

from propagules of native plants 

could account for the lack of recovery 

at this site (Tipping et al. 2009). The 

restoration of some invaded sites 

may require replanting of native 

vegetation rather than solely relying 

on passive recolonization by native 

plants (Rayamajhi et al., 2009; 

Martin et al., 2011). 

Removal of melaleuca can 

benefit native wildlife. Radiotelemetry data of endangered Florida 

Figure 8-3. As melaleuca stands decline under pressure from the introduced 

biological control agents, native plants recover (Credit: Min B. Rayamajhi, 

USDA ARS)

56 Chapter 8: Melaleuca Control in Everglades Restoration: Accomplishments and Future Plans
Biological Control in Natural Areas

panthers, Puma concolor coryi Bangs, within Big Cypress National Preserve indicated a shift in habitat use 

and a 16% reduction in panther home range size correlated with landscape-level removal of melaleuca and 

presumably an increase in habitat quality (Julian et al., 2012).

WERE OTHER MEASURES NEEDED?

Rodgers et al. (2014) conducted aerial surveys in the Everglades between 2010 and 2012 and determined 

that melaleuca occupied 17,802 out of 728,000 ha mapped. This represented 15% of the 1 km grid cells; 

however, the infestation level within these cells was less than 10% in 92% of the cells and less than 1% in 

79% of the cells (Rodgers et al., 2014). Reductions in the density of melaleuca achieved through integrated 

weed management now enable targeted treatment of high density patches as well as large areas with low 

density infestations under a maintenance control program (LeRoy Rodgers, pers. comm.). Herbivory 

from the suite of biological control agents increases the susceptibility of remaining melaleuca trees to 

herbicides, fire, and frost, and decreases the ability of melaleuca to recover from these events (Center et al., 

2012). Although biological control helps to reduce the impact of melaleuca infestations located on private 

land adjacent to conservation areas, these infestations remain of concern (LeRoy Rodgers, pers. comm.). 

Host-range studies are underway on another potential agent, Lophodiplosis indentata Gagné (Diptera: 

Cecidomyiidae), a leaf-galling midge (Center et al., 2012).

FUTURE OF SYSTEM

Interagency collaboration, including The Areawide Management and Evaluation of Melaleuca program 

(TAME Melaleuca), contributed to the progress in controlling melaleuca (Silvers et al., 2007). TAME 

Melaleuca promoted regional implementation of biological control as part of the integrated weed 

management effort to control melaleuca. It included demonstrations of different removal techniques singly 

and within an integrated strategy, assessing the efficacy of these techniques, and transferring management 

technologies to land managers and property owners (Silvers et al., 2007).

A new collaborative project is underway between the USDA ARS Invasive Plant Research 

Laboratory (IPRL), the South Florida Water Management District, and the US Army Corps of Engineers. 

The “Melaleuca Eradication and Other Exotic Plants (2016) – Implement Biological Controls Project” is 

a component of the Comprehensive Everglades Restoration Plan (CERP 2016), which was signed into law 

in the Water Resources Development Act (WRDA) in 2000. This multibillion dollar multi-decade effort 

outlines a plan to simultaneously restore the ecosystem of south Florida while ensuring a sufficient water 

supply for agricultural and urban use and providing flood protection. CERP has been called the world’s 

largest ecosystem restoration effort, encompassing 18,000 square miles over 16 counties in south and central 

Florida. Construction of a mass-rearing annex at the USDA ARS Invasive Plant Research Laboratory was 

the first component of CERP to be completed. The role of IPRL within CERP is to mass rear, release, 

establish, and evaluate the efficacy of biological control agents throughout the area in southern Florida 

being restored. This biological control effort currently targets additional species, including air potato – 

Dioscorea bulbifera L. (Liliales: Dioscoreaceae), waterhyacinth – Eichhornia crassipes (Mart.) Solms 

(Commelinales: Pontederiaceae), and Old World climbing fern – Lygodium microphyllum (Cav.) R. Br. 

(Schizaeales: Lygodiaceae). The agents and activities will change as additional biological control agents 

become available for release and new weed problems emerge. 

Chapter 8: Melaleuca Control in Everglades Restoration: Accomplishments and Future Plans 57
Biological Control in Natural Areas

ACKNOWLEDGMENTS

We thank LeRoy Rodgers, Jon Lane, Melissa C. Smith, and Carey R. Minteer who contributed to this work. 

Funding was provided by the USDA Agricultural Research Service, the US Army Corps of Engineers, the 

South Florida Water Management District, and the Florida Fish and Wildlife Conservation Commission. 

Mention of trade names or commercial products in this publication is solely for the purpose of providing 

specific information and does not imply recommendation or endorsement by the U.S. Department of 

Agriculture. USDA is an equal opportunity employer and provider.

58 Chapter 8: Melaleuca Control in Everglades Restoration: Accomplishments and Future Plans
Biological Control in Natural Areas

CHAPTER 9. PARTIAL RESTORATION OF NATIVE 

RAINFORESTS IN THE ISLAND OF TAHITI 

(FRENCH POLYNESIA) AFTER INTRODUCTION 

OF A FUNGAL PATHOGEN TO CONTROL 

THE INVASIVE TREE MICONIA CALVESCENS

J.-Y. Meyer

Délégation à la Recherche de la Polynésie française

WAS THE INVADER A KEY SOURCE OF ECOSYSTEM DEGRADATION 

OR A THREAT TO NATIVE SPECIES?

The tropical forest ecosystems of the high 

volcanic islands of Hawaii (USA), the Society, 

and the Marquesas Islands (French Polynesia) in 

the Pacific Ocean are recognized as biodiversity 

hotspots (Myers et al., 2012), especially for their 

plant richness and uniqueness (Olson et al., 2001). 

Their vascular flora include 1030, 553, and 333 

native species, respectively, with an endemism 

rate of 48 to 88% (Meyer et al., 2014; Lorence et 

al., 2016). Between 60 and 70% of the endemic 

taxa found in the Society and the Marquesas 

Islands are restricted to tropical montane cloud 

forests (Meyer, 2010a). An alien plant species 

introduced in these three archipelagoes, Miconia 

calvescens DC. (Melastomataceae) (Figure 9-1), 

is a direct and important threat to this unique 

island biodiversity, and its control is viewed as 

an important priority for conservation.

Figure 9-1. Miconia calvescens is a New World tree with large 

leaves that forms an overtopping, dense canopy on Tahiti, 

shading out many native plant species (Credit: Eloise Killgore)

Chapter 9: Partial Restoration of Tahitian Rainforests after Biocontrol of Miconia calvescens 59
Biological Control in Natural Areas

Miconia calvescens (hereafter Miconia) is a small tree (typically 6-12 m in height, but up to 16 m) 

native to rainforests of Central and South America that became very popular for horticulture because of 

its large, brightly colored leaves. It was first introduced to Tahiti in the Society Islands as an ornamental 

in a private garden in 1937. Within 50 years, Miconia had invaded all the native rainforest habitats on the 

island, from nearly sea-level to montane cloudforests, up to 1,400 m elevation, habitats that cover about 

70% of the island (about 80,000 hectares). This fast growing (up to 1.5 m/year) and early reproducing 

(fruiting after 4-5 years of vegetative growth) species produces thousands of small fleshy berries, each of 

them containing hundreds of tiny seeds which are dispersed over long distances by frugivorous birds. A 

single large reproductive tree can produce over 50 million seeds per year, which can germinate in low light 

conditions such as in undisturbed native rainforest understory. Within a few generations, Miconia can form 

dense almost monotypic stands (up to 3 to 6 individuals per m2), causing a dramatic reduction in light. 

By the 1990s, 40 to 50 of the 107 endemic plants of Tahiti were directly threatened by Miconia

(Meyer and Florence, 1996). Most of these endemic plants (classified as Critically Endangered, Endangered, 

or Vulnerable in the IUCN Red List) are understory herbs, subshrubs, shrubs, or small trees that cannot 

survive or reproduce under the shade of a closed canopy of Miconia. Studies conducted on Tahiti have 

demonstrated that reproduction of several endangered endemic shrubs and subshrubs (Psychotria spp. and 

Ophiorrhiza spp., Rubiaceae), and the seedling recruitment of the small endangered endemic tree Rapanea 

(Myrsine) longifolia (Nadeaud) Mez (Myrsinaceae) both decrease with increase in Miconia cover (Meyer 

et al., 2003, 2008b). The replacement of native forests by dense monotypic stands of Miconia, coupled with 

potentially severe impacts on watersheds (e.g., increase of runoff and reduction in groundwater recharge) 

and increased soil erosion (Kaiser, 2006), makes the species an “ecosystem transformer” (Richardson 

et al., 2000). Miconia also became invasive in tropical rainforests following later introductions to other 

islands of the Society archipelago (Moorea, Raiatea, and Tahaa), the Marquesas (Nuku Hiva and Fatu Iva), 

the Hawaiian Islands (Hawai’i, Kauai, Maui, Oahu), New Caledonia (Province Sud), and the Queensland 

region in Australia, and it remains a threat to many other tropical islands in the Pacific and Indian Oceans, 

as well as in the Caribbean (Meyer, 2010b).

WERE OTHER CONTROLS FEASIBLE?

In the Society and the Marquesas Islands, different control strategies were used, depending on the intensity 

of the Miconia invasion. Manual and chemical controls (by hand removal of seedlings and saplings, and 

cut-stump treatment of large reproductive trees) were carried out in small localized areas of high ecological 

value (e.g., areas with remnant populations of endangered endemic plants) in the heavily invaded islands 

of Tahiti and Moorea (>3,500 infested ha). Containment was done on the island of Raiatea where Miconia 

populations were less important (ca 500 infested ha, i.e., less than 3% of the island surface). Eradication 

was attempted on the islands of Tahaa, Nuku Hiva, and Fatu Iva, where the number of individuals was 

low and confined to a small area (<10 ha), and with few reproductive trees. On Raiatea where “Miconia 

removal campaigns” were organized from 1992 to 2008, with total of ca. 3,500 people involved (including 

the French Army, local nature protection groups and schoolchildren), more than 2.2 million Miconia plants 

were destroyed, including more than 4,500 reproductive trees. There was, however, no significant decrease 

in the number of plants being removed annually over this period, likely due to the short duration of each 

control campaign (one or two weeks per year), the relatively low number of participants and volunteers, a 

long lasting soil seed bank (more than 15 years), and the discovery of new populations in remote areas that 

were difficult for control teams to reach (Meyer, 2010b).

60 Chapter 9: Partial Restoration of Tahitian Rainforests after Biocontrol of Miconia calvescens
Biological Control in Natural Areas

WHAT AGENTS WERE RELEASED AND WERE THEY SAFE?

Because manual and chemical control methods had limited effect in heavily invaded islands, especially 

where M. calvescens had already formed dense monospecific stands on steep mountain slopes, investigations 

on biological control of the plant were begun. In addition, the low genetic diversity of the species suggested 

it would be unlikely to rapidly evolve resistance to biological control agents (Le Roux et al., 2008).

Exploratory field trips were conducted by the Hawaii Department of Agriculture (HDOA) between 

1993-1995 in the native range of Miconia (mainly in Costa Rica, Brazil, and Trinidad) in search of its natural 

enemies. Among the biological control candidates, a fungal pathogen – Colletotrichum gloeosporioides 

forma specialis miconiae (Order: Melanconiales, Class: Coelomycetes, Subdivision: Deuteromycetinae, 

hereafter Cgm) – was found in Brazil in 1997, causing Miconia leaf anthracnose and plant necrosis (Figure 

9-2a,b). Rigorous host-specificity testing conducted at the HDOA quarantine facilities to assess its safety 

for introduction involved testing species in the same order (Myrtales) and included native and endemic 

melastomes (Melastoma denticulatum D. Don and Astronidium spp.) found in the Society Islands. Results 

indicated that this strain of C. gloeosporioides was highly specific to M. calvescens (Killgore et al., 1999). 

Cgm causes leaf spots, defoliation, and eventually death of young seedlings under laboratory conditions 

(up to 74% mortality one month after inoculation) (Meyer et al., 2008a). This fungus was successfully 

introduced from HDOA quarantine facilities to Tahiti in 2000, with approval of the French Polynesian 

government, and it was released in two permanent plots located at about 600-620 m elevation with a mean 

annual rainfall ranging from 3,300 mm to 7,000 mm. About one hundred Miconia seedlings or saplings 

(between 10 cm and 2.8 m in height) were monitored for a period of six years after the inoculation to assess 

the pathogen’s dispersal and its impacts on Miconia in the wild. Leaf spots were observed about 30 days 

after inoculation, and the percentage of plants infected reached 100% after 3 months, with lesions on 90-

99% of all leaves. The mortality rate of plants in the monitored group was 15% as a whole, but it reached 

30% for seedlings <50 cm tall (Meyer et al., 2008a). Within three years, the fungus had disseminated 

throughout the island of Tahiti and had infected nearly all the Miconia plants up to 1,400 m in montane 

rainforests. It was also found on the neighbouring island of Moorea located nearly 20 km away, without 

any intentional release there. Cgm has succeeded in infecting nearly all the Miconia plants on both islands, 

particularly at high elevation montane rainforests or cloudforests, by natural means without causing any 

apparent harm to non-target plant species (Meyer et al., 2008a).

a b

Figure 9-2. (a) Inoculating Miconia with the introduced pathogen Colletotrichum gloeosporioides forma specialis miconiae; 

(b) the leaf spot fungus causing deterioration of leaves (Credits: a,b Jean-Yves Meyer)

Chapter 9: Partial Restoration of Tahitian Rainforests after Biocontrol of Miconia calvescens 61
Biological Control in Natural Areas

WERE THERE UNANTICIPATED SYSTEM RESPONSES?

An unanticipated response of the forest ecosystem after 

the release and dispersal of the Cgm in Tahiti was the 

partial defoliation (Figure 9-3) of large reproductive 

Miconia trees. Field studies showed that leaf damage 

on Miconia canopy trees increased from 5% to 45% 

with elevation in study plots set up between 600 and 

1,020 m, suggesting that temperature and moisture 

(as humidity or free water) are important factors for 

disease development, reproduction and dissemination. 

This partial defoliation of Miconia forests favoured 

the recruitment of native plants (Figures 9-4, 9-5), 

including rare threatened endemic plants such as the 

tree R. longifolia (Myrsinaceae) and the subshrub 

Ophiorrhiza subumbellata G. Forst. (Rubiaceae), 

by enhancing the light availability in the understory 

(Meyer et al., 2008a; Meyer and Fourdrigniez, 2011).

Figure 9-3. Partial defoliation of Miconia calvescens

canopy trees caused by the fungal pathogen (Cgm) 

in montane cloud forest in Tahiti (Credit: Jean-Yves 

Meyer)

62 Chapter 9: Partial Restoration of Tahitian Rainforests after Biocontrol of Miconia calvescens

WAS INVADER SUPPRESSION SUFFICIENT FOR RESTORATION?

Monitoring plots (10 x 10 m) were set up along elevation gradients located between 600 and 1200 m to 

study forest dynamics and plant succession over time, noting the taxonomical types (flowering plants, ferns), 

biogeographical status (native, endemic or alien) and ecological groups (shade-tolerant or light-demanding) 

of each species. Results collected over 5 years (2005-2010) showed that the number and cover of understory 

plants increased with time, an effect which was more pronounced for light-demanding pioneer species. Most 

colonizing understory species (80%) were native trees, shrubs, or ferns, with little reinvasion by other alien 

plants, which was mainly restricted to the lowest altitude plots (Meyer et al., 2012). Seedlings of rare endemic 

plants – such as the tree Pittosporum taitense Putterlick (Pittosporaceae) and the small terrestrial orchid 

a b c

Figure 9-4. Threatened endemic plant species such as the shrub Psychotria speciosa (a), the small tree Myrsine longifolia (b), 

and the small orchid Liparis clypeolum (c) have benefited from the partial defoliation of Miconia calvescens caused by the 

introduced fungal pathogen (Credits: a-c Jean-Yves Meyer)
Biological Control in Natural Areas

Figure 9-5. Recruitment of native ferns 

and shrubs in the understory of a 

Miconia calvescens-invaded rainforest 

where Miconia has been partially 

defoliated by the introduced fungal 

pathogen (Credit: Jean-Yves Meyer)

Chapter 9: Partial Restoration of Tahitian Rainforests after Biocontrol of Miconia calvescens 63

Liparis clypeolum (G. Forst.) Lindl. (Figure 9-4c) – were found in the permanent plot at 600 m elevation in 

2014 although they had not been seen in the past 20 years at that site (J-Y Meyer, unpub. data). However, 

these plant assemblages are certainly different from the pre-invasion stage (which remains unknown), and 

long-term monitoring of forest dynamics will be needed to study the trajectory of these “novel habitats”. 

FUTURE OF SYSTEM

Sixteen years after its successful release in Tahiti, the fungal pathogen Cgm did not provide complete 

control of Miconia, especially in low elevation rainforests (<600 m) where warmer temperatures seem to 

decrease the fungus effectiveness. However, long-term monitoring demonstrates that this biological control 

agent can be considered a tool for partial restoration of heavily invaded rainforests at higher elevation 

(including montane cloudforests) and contribute to the recovery of native and endemic plant species (Van 

Driesche et al., 2016c). Other highly specific biocontrol agents are being tested in Brazil, such as the fungus 

Coccodiella miconiae (Duby) I. Hino & Katum. (Ascomycota: Phyllachoraceae), which attacks younger 

leaves than Cgm, and so could be a useful complementary agent (Alves et al., 2014). 

In summary, although these fungal pathogens may slow the growth of established Miconia plants and 

cause the dieback of young seedlings, it is unlikely that they will reverse the massive invasion of Miconia.

On the islands of Raiatea and Nuku Hiva where Cgm was released in 2004 and 2007 respectively, as well as 

in the Hawaiian Islands, manual and chemical control efforts will be still necessary to eradicate or contain 

this extremely invasive alien plant in tropical island forest ecosystems.

ACKNOWLEDGMENTS

I thank the graduate students and field assistants for technical assistance and data collection in the field and 

the laboratory for the past 15 years of monitoring, including Anne Duplouy, Marie Fourdrigniez, Meryl 

Jordan, and Ravahere Taputuarai. I thank the plant pathologist Eloise M. Killgore (formerly at the Hawaii 

Department of Agriculture, Honolulu, now retired) for her invaluable contribution to Miconia biological 

control both in Hawaii and French Polynesia.
Biological Control in Natural Areas

CHAPTER 10. STRATEGIC RESTORATION OF SALTCEDARAFFECTED RIPARIAN ECOSYSTEMS OF THE U.S. 

SOUTHWEST: INTEGRATION OF BIOCONTROL AND 

ECOHYDROLOGICAL CONDITIONS 

IN RESTORATION PLANNING

T. L. Dudley1, D. W. Bean2 and C. J. DeLoach3

1

Marine Science Institute, University of California, Santa Barbara 2

Colorado Department of Agriculture, Palisade Insectary

3

USDA ARS, Temple, Texas

WAS THE INVADER A KEY SOURCE OF ECOSYSTEM DEGRADATION 

OR A THREAT TO NATIVE SPECIES?

Tamarix spp. (saltcedar, or Tamarix hereafter) is a complex of semi-deciduous shrub species originating in 

Eurasia and currently established in wetlands and riparian ecosystems of the arid and semi-arid parts of western 

North America (Figure 10-1) and other regions with similar climates (Argentina, South Africa, Australia). 

Although introduced for horticultural purposes (erosion control and shade in desert regions), these shrubs escaped 

cultivation early in the 20th century in association with the widespread regulation of western US rivers, which 

inhibited the natural flooding regimes that 

promote regeneration of native cottonwoodwillow riparian woodlands (Horton, 1977). By 

the mid-1900s, several Tamarix species and 

their hybrids (Gaskin and Schaal, 2002) had 

invaded roughly 1 million ha of the western 

United States (Robinson, 1965; Everitt, 1998) 

and replaced these riparian woodlands, as well 

as other vegetation types such as mesquite 

bosque, riparian scrub, and saltgrass/halophytic 

scrublands, in part as a consequence of reduced 

water availability, soil salinization, and other 

degradation that reduced these ecosystems’ 

capacity to sustain native vegetation (Busch 

and Smith, 1995). At this time, water managers 

also removed large areas of riparian vegetation 

Figure 10-1. Several species of invasive saltcedars (Tamarix spp.) 

dominate riparian areas over much of the southwestern United 

States (Credit: Steve Dewey, Utah State University, bugwood.org)

64 Chapter 10: Restoration of Saltcedar-affected Ecosystems of the U.S. Southwest
Biological Control in Natural Areas

for presumed water ‘salvage’ in places like the Gila River (Orr et al., 2014), which further facilitated Tamarix’s 

expansion in human-altered habitats due to its tolerance of degraded conditions and its capacity to outcompete 

native plants under such conditions (Sher and Marshall, 2003; Shafroth et al., 2005; Hultine and Dudley, 2013). 

Tamarix invasion, however, is also observed in many largely non-degraded systems where hydrology or fire 

regimes allow its expansion (Dudley et al., 2000; Whiteman, 2006; Mortenson and Weisberg, 2010). 

Once Tamarix becomes a dominant element in a system, it causes several environmental and economic 

problems, including high evapotranspiration rates that deplete groundwater (Smith et al., 1998), inhibition of 

native plant recruitment as a consequence of depleted soil moisture and soil salinization by throughfall and 

deposition of salt-laden litter (Shafroth et al., 1995), and increased sedimentation and erosion where dense 

Tamarix stands slow water flow and re-direct water into narrower open areas that are scoured (Graf, 1978; Birken 

and Cooper, 2006). Tamarix offers poorer quality habitat to wildlife than native vegetation, in part by presenting 

a simplified vegetative structure and altering understory conditions for birds, reptiles, and other taxa (Ellis, 

1995; Shafroth et al., 2005). Arthropods are also affected, which may reduce habitat suitability for insectivorous 

vertebrates (Durst et al., 2008; Longland and Dudley, 2008). 

Of special concern is a federally listed sub-species, the 

southwestern willow flycatcher (SWFL, Empidonax traillii 

extimus Phillips) (Figure 10-2), which has declined globally, 

in part due to the Tamarix invasion. However, the bird’s use 

of Tamarix as a nesting substrate in a limited portion of its 

range (Sogge et al., 2003) led to legal actions to modify or halt 

measures, including biocontrol, to reduce Tamarix abundance 

(Dudley and DeLoach, 2004). 

Aquatic systems benefit from riparian litter, which 

forms the trophic base for many aquatic invertebrates, as well as 

the fish that then feed on these invertebrates, but Tamarix litter 

is relatively unsuitable for detritus-feeding organisms (Going 

and Dudley, 2008; Moline and Poff, 2008), with subsequent 

effects on fish (Kennedy et al., 2005). Salt marsh habitats are 

also affected by Tamarix invasion, with trophic assemblages 

being degraded by low nutritional quality of organic inputs 

(Whitcraft et al., 2008).

An increasingly serious impact of Tamarix invasion is 

the promotion of wildfire (Busch, 1995; Drus, 2013) (Figure 10-

3). Fire is uncommon in native-dominated riparian vegetation 

in the invaded region, but Tamarix ignites readily and burns 

with high intensity whether foliage is senescent and dry, or 

green and healthy (Drus et al., 2013), thus turning riparian 

areas into promoters of wildfire spread, rather than acting as 

barriers (Lambert et al., 2010). Burn severity increases with 

Tamarix density, as does mortality of associated native woody 

plants such that fire drives a positive feedback that can lead 

to Tamarix monocultures (Drus, 2013). Tamarix is also poor 

quality forage for wildlife and livestock, albeit the nitrogenrich flowering parts can be useful for livestock when present 

(Cliven Bundy, pers. comm.). 

Figure 10-2. The endangered subspecies of willow 

flycatcher, Empidonax traillii extimus, in tamarisk 

in Tonto National Forest, Arizona (Credit: 

Andre Silva, USDA Forest Service)

Chapter 10: Restoration of Saltcedar-affected Ecosystems of the U.S. Southwest 65

Figure 10-3. In absence of biocontrol, native 

cottonwood trees, Populus fremontii, were 

killed by Tamarix-fueled wildfire at the San 

Pedro River, Arizona (Credit: G. Drus)
Biological Control in Natural Areas

WERE OTHER CONTROLS FEASIBLE?

Western U.S. land managers have long attempted to reduce Tamarix densities by one means or another, 

primarily to save water that otherwise would be lost to the atmosphere through Tamarix transpiration 

(Cleverly et al., 2002), as well as for fuel reduction and habitat or range improvement. Control measures 

have included herbicide treatments to wood or foliage using hand, ground rig, or aerial application 

methods, particularly with triclopyr formulations. Other methods have been use of prescribed fire and direct 

mechanical removal with handtools, bulldozers, or other tools (Taylor and McDaniel, 1998; Kaufmann, 

2005; Shafroth et al., 2005). These methods can be effective under many circumstances, but control is hard 

to sustain because chemical applications generally require repeated follow-up treatments, and re-sprouting 

following fire or mechanical clearing is typical. 

Substrate disturbance associated with mechanical treatments also tends to promote secondary 

invasion of disturbance-associated weeds such as Russian thistles (Salsola spp.), knapweeds (Centaurea

spp., Rhaponticum [Acroptilon] repens [L.] Hidalgo), and pepperweeds (Lepidium spp.), which are also 

management concerns (Bay and Sher, 2008; Ostoja et al., 2014). Also, Tamarix-infested sites tend to be of 

marginal economic value and benefits rarely justify the exorbitant costs of conventional treatments, while 

collateral damage to the landscape and remnant native vegetation are not readily mitigated (Dudley and 

Brooks, 2011). Furthermore, the anticipated water conservation believed to follow Tamarix removal has 

been largely unattained. This difficulty is in part due to the ambiguities inherent in measuring water yields 

under arid conditions and in part because groundwater retained in one river reach by Tamarix clearing can 

be negated by increased uptake and transpiration further downstream (Shafroth et al., 2005).

WHAT AGENTS WERE RELEASED AND WERE THEY SAFE?

Since there are no native species of Tamaricaceae (to which Tamarix belongs) in the New World, and Tamarix

species in North America are of little economic or ecological value (other than the evergreen athel, Tamarix 

aphylla [L.] Karst., which is used as a shade tree in warm, arid regions), this group of invasive shrubs 

seemed an appropriate target for biological control. A biocontrol program against Tamarix was started by 

the USDA-ARS in the 1960s by Lloyd Andres and subsequently led by Jack DeLoach, who with many 

overseas cooperators identified over 300 specialist herbivores with potential for suppressing this invasive 

plant in North America (Pemberton and Hoover, 1980; DeLoach et al., 1996). At this phase, the U.S. Fish & 

Wildlife Service was an active participant in the biocontrol program (Stenquist, 2000). Of candidates tested 

for specificity and efficacy, by 1996 three were approved by the USDA-APHIS TAG (Technical Advisory 

Group on the Biological Control Agents of Weeds) review process as safe for open release (DeLoach et al., 

2004). These three were a mealybug (Trabutina mannipara [Hemprich & Ehrenberg]), a foliage-feeding 

weevil (Coniatus tamarisci F.), and a leaf beetle (originally referred to as Diorhabda elongata deserticola

Chen, and subsequently reconsidered as D. carinulata [Desbrochers] within a complex of Tamarix-feeding 

Diorhabda species) (Tracy and Robbins, 2009). 

Four of the Diorhabda species (D. carinulata, D. elongata Chen, D. carinata [Faldermann], and 

D. sublineata [Lucas]) and various geographic ecotypes of these species eventually underwent quarantine 

testing in USDA quarantine facilities in Albany, California or Temple, Texas (Lewis et al., 2003; Milbrath and 

DeLoach, 2006). Additional taxa have received attention in overseas or U.S. quarantine studies (DeLoach 

et al., 1996), but these additional species have not yet been submitted to APHIS for TAG consideration. 

On the other hand, several unintentionally introduced Tamarix specialists, including a leafhopper (Opsius 

66 Chapter 10: Restoration of Saltcedar-affected Ecosystems of the U.S. Southwest
Biological Control in Natural Areas

stactogalus Fieber) and two diaspidid 

scales in Chionaspis are present in North 

America (Wiesenborn, 2005) and can 

interact with agents intended for release 

(Louden, 2010). Finally, the weevil 

Coniatus splendidulus F. (Figure 10-

4a,b), of unknown origin, was recently 

detected in Arizona and quickly spread to 

surrounding states (Eckberg and Foster, 

2011; Dudley and Bean, 2012), and this 

species is similar ecologically to one 

approved previously by TAG (Sohbian et 

al. 1998) but not released. 

In the 1990s, the Tamarix biocontrol 

program came under increased scrutiny 

when it was found that the SWFL was 

nesting in Tamarix in some regions, 

particularly in Arizona and New Mexico 

(Sogge et al., 2003). Informal consultation 

with the US Fish & Wildlife Service 

delayed releases until 1998 when a compromise was reached between USFWS and USDA under which one 

agent, Diorhabda sp., could proceed to open release, but with the limitation that releases would be made at 

least 200 miles from known locations where SWFL was known to use Tamarix for nesting (Kaufman, 2005). 

Thus, in 1999, D. carinulata (formerly known as D. elongata deserticola) from northwestern China and 

Kazakhstan was released into secure cages in seven western states to evaluate its growth and survival, while 

avoiding the southwestern region where SWFL was present (Dudley et al., 2001). Where overwintering of 

the beetle was successful, beetles were released into the open in 2001, and evidence of substantial defoliation 

of ca 2 ha was first observed in the Humboldt River drainage of northern Nevada in late 2002 associated 

with high densities of larvae (Figure 10-5a-c) (DeLoach et al., 2004). In subsequent years, the area of 

defoliation expanded by several orders of magnitude, and establishment was also documented in other 

locations (e.g., the lower stretches of the Walker River, Nevada, the Sevier River, Utah, and the Dolores and 

a b

Figure 10-4. The tamarisk weevil, Coniatus splendidulus (a), is a 

self-introduced herbivore affecting Tamarix in the United States, 

which may damage Tamarix ramosissima, as seen here on the Bill 

Williams River in Arizona (b) (Credits: a Z. Ozsoy; b Tom Dudley)

Chapter 10: Restoration of Saltcedar-affected Ecosystems of the U.S. Southwest 67

a b c

Figure 10-5. Diorhabda carinulata: (a) adults scraping Tamarix photosynthetic stem tissue; (b) larvae on Tamarix; (c) larvae 

reaching high densities that defoliate saltcedar (Credits: a A. Abela; b Dan Bean, Colorado Department of Agriculture; 

c Tom Dudley)
Biological Control in Natural Areas

Colorado Rivers in the Colorado/Utah border region, while another form, D. sublineata, established along 

the Rio Grande in Texas and New Mexico) (Figures 10-6a,b). Here, the focus is on the Colorado River 

Basin and Great Basin where most of Dudley’s work has taken place.

The released beetle caused no significant impacts to non-target plants. The potential for feeding by 

the biocontrol agent on native species of Frankenia (Tamaricales: Frankeniaceae) was evaluated in the field 

because Frankenia species were fed on to a limited extent in quarantine studies (Lewis et al., 2003; Herr 

et al., 2009; Moran et al., 2009). At two field sites where Diorhabda beetle densities were high and host 

plants were defoliated, larvae left plants in search of food but still caused insignificant damage to F. salina

experimentally planted adjacent to the infested Tamarix stands (Dudley and Kazmer, 2005). 

Furthermore, this central Asian beetle selectively oviposited on forms of Tamarix from central Asia 

but did not recognize the Mediterranean Tamarix species (Tamarix parviflora DC., T. aphylla) (Figure 10-

7) as suitable oviposition hosts, while middle eastern or Mediterranean Diorhabda species, presumed to be 

a b

Figure 10-6. (a) Tamarix defoliation and native plant recovery in 2008, Dolores River, Colorado; (b) C. J. DeLoach, 

USDA-ARS, lead scientist for the Tamarix biocontrol program, in front of defoliated Tamarix ramosissima on the Rio 

Grande in Texas (Credits: a,b Dan Bean, Colorado Department of Agriculture)

68 Chapter 10: Restoration of Saltcedar-affected Ecosystems of the U.S. Southwest

Figure 10-7. Defoliation of Tamarix ramosissima next to Tamarix 

parviflora, which is largely unaffected by Diorhabda feeding, in 

2011 on the Virgin River in Arizona (Credit: Tom Dudley)

Figure 10-8. Traps baited with aggregation 

pheromone are used for detection of 

Diorhabda colonization, showing beetles 

caught on trap (Credit: Tom Dudley)
Biological Control in Natural Areas

sympatric in origin with these plants, did oviposit on them (Dudley et al., 2012; Moran et al., 2009). When 

herbivore densities were at epidemic levels causing complete defoliation of T. ramosissima/T. chinensis

Lour. forms, some spillover onto T. aphylla occurred and was a concern to landowners, especially in 

northern Mexico. But affected trees readily recovered and little impact was observed in subsequent years 

when insect densities were more moderate (DeLoach et al., 2012). This specificity within the genus Tamarix

was a strong indication that non-target effects outside this taxon are unlikely. 

WERE THERE UNANTICIPATED SYSTEM RESPONSES?

In the Herbivore-Plant System

Initial expectations were that Tamarix biocontrol would be gradual, and thus the rapid insect population 

growth and expansion of the area of defoliation at key locations (Bean et al., 2013; Pattison et al., 2011) was 

unexpected (Hultine et al. 2010), although this should reasonably be considered a positive sign of pending 

biocontrol success. At the same time, however, releases of Diorhabda beetles did not lead to establishment 

in most locations, for reasons that were unclear at the time (Dudley et al., 2012). It was later found out 

that these failures were due to factors such as release on the wrong species of Tamarix, consumption by 

generalist predators (Knutson et al., 2012, Strudley and Dalin, 2013), and, most importantly, a mismatch 

between the beetle’s diapause induction requirements and the local day length patterns (Bean et al., 2007). 

Thus, at latitudes south of ca 38oN (the area used by SWFL), adult insects entered diapause too early (in 

mid-summer), resulting in mortality because metabolic reserves were exhausted before the following spring 

(Dalin et al., 2010). Other genotypes of Diorhabda were introduced that were more effective at lower 

latitudes, particularly in Texas (Knutson et al., 2012; Michels et al., 2013) or against T. parviflora (Tracy 

and Robbins, 2009). Sentinel traps baited with an aggregation pheromone were used to detect populations 

before defoliation occurs (Figure 10-8), followed by long-term monitoring to determine impacts on Tamarix 

plants (Figure 10-9a,b).

a b

Figure 10-9. Monitoring the progress of the Tamarix biocontrol program at Bedrock Creek, in Colorado, showing 

the impact of Diorhabda carinulata on Tamarix ramosissima between 2010 (a) and 2014 (b) (Credits: a,b Colorado 

Department of Agriculture)

Chapter 10: Restoration of Saltcedar-affected Ecosystems of the U.S. Southwest 69
Biological Control in Natural Areas

70 Chapter 10: Restoration of Saltcedar-affected Ecosystems of the U.S. Southwest

Subsequently, D. carinulata was observed to have become established south of the former latitudinal 

limit along the Virgin River in southern Utah downstream to the lower Colorado River (Hultine et al., 

2015a). Growth chamber studies with beetles collected from these field sites showed that in only 4-5 years 

natural selection for delayed diapause induction had enabled this southern expansion (Bean et al., 2012). By 

late 2016, beetles had become established through the lower Colorado River Basin, south to almost 34oN 

(unpub. data). It is likely these beetles will continue to expand southward to the Mexican border and into 

the Colorado Delta where Tamarix is also of conservation concern (Nagler et al., 2009). 

On the other hand, we have typically observed that the initial epidemic levels of Diorhabda

population increase and defoliation were followed a few years later by unanticipated population declines, 

the mechanism for which is not clear. This decline probably involves a combination of reductions in the host 

plant biomass (necessarily supporting fewer herbivores) and the impacts of generalist predators (arthropods 

as well as vertebrates), which respond positively to this new prey resource (Longland and Dudley, 2008; 

Bateman et al., 2010; Bean et al., 2013; Strudley and Dalin, 2013). 

Ecosystem Responses

Major alteration of any dominant vegetation type must be assumed to have secondary or interactive effects, 

many of which can only be anticipated in general terms (Denslow and D’Antonio, 2005). Indeed, such 

indirect secondary effects as competitive release of associated flora, improvements in litter and soil traits or 

water relations, and increased use of native vegetation by wildlife are primary objectives of ‘conservation 

biocontrol,’ although some short-term negative effects may occur. Some positive responses occurred 

immediately; for instance, over 2,500 acre-feet of water were estimated to have been conserved annually 

in groundwater due to reduced evapotranspiration after defoliation of Tamarix (Pattison et al., 2011). But 

given the diverse objectives of Tamarix biocontrol, none of the responses in affected ecosystems were 

unanticipated, even if their magnitudes were unknown.

The consequences of Tamarix biocontrol for vertebrate wildlife are unclear (Bateman et al., 2013), but 

early projections of major mortality of nesting birds from defoliation (Sogge et al., 2008) have not been borne 

out. Inferential data (Paxton et al., 2011) suggest that biocontrol was implicated in some nesting failures. 

Certainly, Tamarix dieback reduced habitat quality for some taxa in the short term (Bateman et al., 2013). 

While long-term outcomes are still not known, emerging data in defoliated areas indicate that as riparian 

systems trend back toward native plant composition (Figure 10-10) due to the biocontrol program, reptiles 

and amphibians increase in diversity and abundance, but in heavily altered systems some measure of active 

restoration may be needed to accelerate this response (Mosher and Bateman, 2016). Wildlife benefits from 

substantial native vegetation cover (> ca 30%) within mixed Tamarix stands (van Riper et al., 2008), so the 

key question is whether a given system has capacity for recovery in a timeframe that sustains wildlife.

WAS INVADER SUPPRESSION SUFFICIENT FOR RESTORATION?

The initial few years of Tamarix suppression by Diorhabda was impressive, and reduction in aerial cover 

of this competitive invader was associated with measureable increases in native or desired vegetation in 

several locations (N. Louden, unpub. data). Mortality of 60% or more was noted in some areas (Bean et 

al., 2012; Hultine et al., 2015a) due to gradual depletion of stored metabolites by repeated defoliation that 

prevented the plants from replenishing their carbohydrates (Hudgeons et al., 2007). 
Biological Control in Natural Areas

Figure 10-10. Tamarix defoliation and mortality, with recovery of native willow, Salix gooddingii, and cottonwood, Populus 

fremontii, in 2010 at the Virgin River, Arizona (Credit: Tom Dudley)

Chapter 10: Restoration of Saltcedar-affected Ecosystems of the U.S. Southwest 71

Whether this suppression was sufficient to promote ecosystem recovery is a more complex question. In 

systems where native woody plant abundance is moderately high and environmental conditions (moisture, soil 

salinity, etc.) are not excessively degraded, such as the Dolores River near Moab, Utah or the Walker River in 

central Nevada, we find that recovery may be slow but is taking place without major intervention. Recovery of 

important natural resource features is likely to be the case in many other systems where ecosystem processes 

still favor growth of native woody plants (Shafroth et al., 2007), but recovery will be slower where degradation 

levels are high or multiple stress factors inhibit regeneration. It is important to note that SWFL is absent and 

other sensitive riparian species rare under such degraded conditions (Dudley and DeLoach, 2004).

The Virgin River is an interesting case, in that it largely retains the hydrologic integrity (flood/drought 

sequences) that normally favors native species over Tamarix, and indeed where active restoration measures 

were implemented in areas along this river that had been affected by Tamarix biocontrol, recovery of native 

vegetation appears to be good (Leverich et al., 2014; Mosher and Bateman, 2016). But in this system, we 

have observed that periodic wildfires, promoted by Tamarix, have nearly eliminated native trees (Fremont 

cottonwood, Populus fremontii S. Watson and Goodding’s willow, Salix gooddingii C.R. Ball) from much 

of the system such that propagules are lacking that could regenerate native riparian woodlands (Dudley and 

Brooks, 2011). That is, recruitment limitation circumvents the process by which natural flooding creates 

scoured substrates that facilitate germination and establishment of these native taxa (Stella et al., 2006; 

Merritt and Poff, 2010). In such cases, a restoration strategy is to re-introduce native trees (manual seeding, 

container-stock outplanting, pole cuttings etc.) in patches distributed at intervals across the floodplain. 

These ‘propagule islands’ serve as seed sources so that in the future when hydrological conditions favor 

recruitment, there will be sufficient reproductive plants to broadly disperse seed across the floodplain. This 

approach can be done with relatively low investment, while enhancing riparian recovery because natural 

recruitment tends to yield better plant growth and survival than in large-scale active re-vegetation (Dudley 

and Bean, 2012, Leverich et al., 2014). In addition, because biocontrol reduces Tamarix fuel loads, it also 

can improve native plant survival by reducing the frequency, extent, and severity of wildfires (Drus, 2013). 

In other cases, the same environmentally damaging processes that facilitated conversion from native 

woodlands to Tamarix stands (water diversion, salinization, livestock grazing, etc.) must be mitigated along 

with Tamarix suppression if recovery is to take place. At the Humboldt River, Nevada, diversion and salinization 

are intense and grazing ubiquitous, so that cottonwood-willow recovery has low probability of success, 

although some native taxa that tolerate such conditions (e.g., Atriplex, Sarcobatus, Distichlis) may still be able 
Biological Control in Natural Areas

to increase with Tamarix suppression, but so do various other secondary weeds. Likewise, the lower Colorado 

River and many of its tributaries such as the Gila River have been so altered, particularly in curtailing natural 

flooding, that Tamarix suppression alone will not lead to passive recovery of riparian vegetation (Shafroth et 

al., 2007). Except for narrow bands immediately adjacent to the water’s edge, restoration in such areas will 

likely require active re-vegetation to compensate for the loss of factors that promote native tree recruitment 

under more natural hydrologic regimes. A useful approach to improve recovery potential is manipulation 

of flow regimes to mimic natural processes (Konrad et al., 2011), which has been implemented at another 

Colorado River tributary, the Bill Williams River, where simulated flow fluctuations from the Alamo Reservoir 

were successful in sustaining diverse native woodlands despite presence of Tamarix (Shafroth et al., 2010). 

Diorhabda beetles established along the Bill Williams River in 2016; however, it is expected that there will be 

little or no need for active measures during the incremental decline of Tamarix because flow conditions and 

native propagules at the sites appear suitable for natural vegetation recovery. 

For systems in which sufficient ecosystem function exists to support native vegetation, the probability 

for recovery can be improved by a strategic approach to restoration which integrates hydrology and fluvial 

geomorphology, soils, and current vegetation condition and wildlife status, a process termed “Ecohydrological 

Assessment.” These factors are used along with historical and current aerial imagery and remote sensing to 

direct effort where it will be most productive, at relatively low cost, for improving the chances of riparian 

recovery (Leverich et al., 2014). This is being applied in several western streams, and in the context of 

Tamarix suppression by biocontrol, we are implementing restoration in the Virgin River (Nevada) and upper 

Gila River (eastern Arizona) based on these principles (Orr et al., 2014; Stillwater Sciences, 2014). 

While it is too early to claim success, native vegetation is establishing well at Virgin River sites 

where SWFL populations had previously declined. Once Diorhabda beetles are established in a location, 

the resulting Tamarix defoliation occurs early in the season, before SWFL migrants return from their overwintering grounds in Central America. Returning birds have been observed to pass over defoliated Tamarix 

and occupy the remaining patches of native willows (BOR, unpub data). The time frame for vegetation 

recovery at a site necessarily requires several years, but cottonwood-willow replacement of Tamarix in 

other systems, e.g., at Elephant Butte Reservoir on the Rio Grande (NM), was associated with dramatic 

increases of this endangered bird over a period of ca 10 years (Ahlers and Moore, 2009). Riparian plants are 

effectively native weedy species, dependent on natural disturbance to promote regrowth, and the SWFL and 

other sensitive riparian-associated species such as Bell’s vireo (Vireo bellii Audobon) and yellow warbler 

(Setophaga petechia [L.]) are associated with early phases of riparian succession (Ellis, 1995; Shafroth et 

al., 2005), so it is unsurprising that they would respond positively to enhancement of native vegetation, 

essentially a “build it and they will come” relationship (Dudley and Bean, 2012). 

FUTURE OF SYSTEM

There are many Tamarix-affected ecosystems where biocontrol would provide a useful ‘tool’ for resource 

management, and in California an Alliance supported by the state EPA was recently formed to work with 

regional stakeholders to consider if and how biocontrol implementation would help meet their resource 

management goals. Implementation should consider how experiences to-date allow us to objectively 

evaluate the costs, benefits, and risks of Tamarix biocontrol, with the realization that to-date, neither the dire 

predictions of endangered species managers nor the early hopes of biocontrol practitioners have come to pass. 

With respect to wildlife, riparian areas are dynamic ecosystems and the biota, even including federally listed 

taxa, are adapted to periodic disturbances, so it is important to put short-term alteration of Tamarix vegetation 

into the context of shifting faunal populations as suitable habitat is transitive and routinely lost, whether 

72 Chapter 10: Restoration of Saltcedar-affected Ecosystems of the U.S. Southwest
Biological Control in Natural Areas

from biocontrol, wildfire, or community 

maturation into a less suitable state. At 

the same time, we would be naïve to 

ignore the landscape-level implications 

of both large-scale Tamarix decline 

(Figure 10-11) or of its continued 

domination (Mortenson and Weisberg, 

2010), particularly in the context of the 

anticipated effects of climate change in 

western North America (Hultine et al., 

2015b). That debate will continue, but 

recent discussion among federal agencies 

suggests that at least the legal situation that 

has constrained the Tamarix biocontrol 

program and even the monitoring of its 

consequences may soon be resolved. 

Such resolution would allow the 

program to be re-initiated, and the mixed results of the current situation can be addressed in a more constructive 

manner. That would include evaluation of new agents given the moderate level of suppression and 

plateauing of host mortality from the Diorhabda beetle introduction alone. A potential agent is the already 

approved mealybug T. mannipara, considered appropriate for high temperature regions, while evaluation of 

the naturally established weevil C. splendidulus may ultimately provide the same benefit as that anticipated 

earlier from the related, TAG-approved related weevil C. tamarisci, which has not been introduced. Overseas 

cooperators have identified and tested other organisms that could be used to affect other plant parts, such as 

stem-feeding Psectrosema spp. (Cecidomyiidae), the gall-forming weevil C. tamarisci, a defoliating moth 

Agdistis tamaricis (Zeller) (Pterophoridae) and several other Tamarix specialists (Sohbian et al., 1998) if 

pre-release evaluation concludes that multiple agents will be synergistic rather than agonistic.

ACKNOWLEDGMENTS

The Tamarix Biological Control Program has involved many researchers and specialists over its 25+ year 

lifespan, too many to name but particular appreciation is given for the commitment to this program by 

long term colleagues, including Li Baoping (Nanjing Agricultural University), Roman Jashenko and ID 

Mityaev (Al-Farabi Kazakh National University), Rouhollah Sobhian (USDA-EBCL), from USDA-ARS 

& APHIS - James Tracy, John Gaskin, Bob Bartelt, Allard Cossé, Earl Andress, Phil Lewis, R. Carruthers 

and Julie Gould, Debra Eberts and Ken Lair (Bureau of Reclamation), Allen Knutson and Jerry Michels 

(Texas A&M University), David Thompson (New Mexico State University), Zeynep Ozsoy (Colorado 

Mesa University), Nina Louden (Colorado Department of Agriculture), Richard Reardon and Elizabeth 

Hebertson (USFS Forest Health Protection), Bruce Orr and Glen Leverich (Stillwater Sciences), Jeff 

Knight (Nevada Department of Agriculture), Gail Drus (St. Francis University), Tim Higgs (Grand County, 

Utah) and Mike Pitcairn (California Department of Food & Agriculture). Sorely missed but not forgotten 

are David Kazmer, Brian Cardall, and from USFWS, Scott Stenquist and John Taylor. Support has been 

from USDA-NIFA, Clark County, Nevada Desert Conservation Program, USFS Forest Health Protection, 

USDA-APHIS, and other sundry sources.

Figure 10-11. Defoliated Tamarix on Colorado River near Moab, 

Utah in 2008, following release of Diorhabda carinulata beetles, 

with patches of remnant native willow stands in the far background 

(Credit: Tom Dudley)

Chapter 10: Restoration of Saltcedar-affected Ecosystems of the U.S. Southwest 73
Biological Control in Natural Areas

74 Chapter 11: Future Trends in Use of Natural-area Biocontrol

CHAPTER 11. FUTURE TRENDS IN USE OF 

NATURAL-AREA BIOCONTROL

R. G. Van Driesche

Department of Environmental Conservation, 

University of Massachusetts, Amherst

CONCLUSION FROM THE CASE STUDIES

Finally, for the Tamarix case and the other examples discussed above, we need to upgrade our collective 

perception of biological control and differentiate the conceptual framework of conservation biocontrol 

(against pests of natural areas) from agricultural biocontrol, with its simpler objective to suppress or 

eliminate pest organisms (Van Driesche et al., 2016a). Conservation biocontrol is a holistic approach to 

re-assembling a weed’s or insect’s co-evolved mosaic of species, based on what existed in the pest’s area of 

origin so that this group of species functions in a more complex and less damaging manner than when the 

pest occurs in invaded areas without such associates. Post-biocontrol, the former pest exists in the invaded 

ecosystem, restrained by many such trophic links and becomes just another point in the local foodweb, with 

various feedbacks within the biotic assemblage, in place of acting as a source of biotic stress for the system. 

CHALLENGES AHEAD

Future use of biological control as an emerging tool for conservation biology (Van Driesche et al., 2016a) 

will depend on how well several currently visible problems are dealt with, including the following:

Funding and Scientist-power Going Down, While Need Goes Up

The costs of natural enemy discovery, assessment of safety, petitioning for release, mass rearing and release, 

and post-release efficacy and safety assessments all must be paid for by government, or these activities do 

not happen. Since the 1970s, there have been strong swings in the willingness of governments in the major 

biocontrol-conducting countries to pay such bills. Because many insect and even some weed biocontrol 

projects are done to provide economic benefits to grazers, farmers, or other private groups, the view has 

arisen, particularly in Australia and New Zealand, that those groups should pay these bills. That has worked 
Biological Control in Natural Areas

Chapter 11: Future Trends in Use of Natural-area Biocontrol 75

relatively badly because the finer points of good biological control practice – taxonomic revisions, nontarget 

impact studies, post-release evaluations, development of new techniques, etc – may either be unknown to 

these groups or be little valued. Also, the projects of great need for protection of natural systems may be 

of only moderate economic importance and vice versa, mis-aligning selection of projects with ecological 

need. The need for strong governmental support, with private supplemental funds, remains – and when not 

provided, outputs decline. Similarly, continuity of biocontrol/natural enemy knowledge is critical, with new 

scientists learning much from the previous generation before they retire. Bottlenecking human resources in 

this science, by periodic steep reductions in hiring in the face of continued retirements of older scientists, 

degrades institutional capabilities, with loss of scientific competency. Furthermore, while the trend in 

funding and human resources for biocontrol in most of the major practicing countries is negative or neutral 

at best, the extent of invasive pest problems in natural areas continues to increase. Only in South Africa does 

an absolute increase in funding – via the Working for Water Program – seem to have occurred over the last 

several decades (see Moran et al. [2005] for comments on Working for Water Program). 

Lingering Distrust of Biocontrol by Conservation Biologists

While some degree of improved understanding between conservation biologists and biological control 

scientists may be underway in the United States (Van Driesche et al., 2016a), much disagreement remains. 

This will likely take a full generation to alter. As such, this factor certainly will affect the choice of projects, 

speed of their adoption, and wide understanding of their consequences in the United States for the foreseeable 

future. This problem seems smaller or non-existence in other countries (e.g., South Africa). 

Failure to Distinguish Use from Population-level Impacts

Conservation biologists focus on the welfare of populations of the species of concern. Animal-welfare 

groups (such as PETA), in contrast, focus on suffering or health of individual animals. This pits the two 

groups against each other in ways that may be unresolvable. For example, groups concerned with vegetation 

destruction due to overgrazing by introduced ungulates or equines, may be prevented from reducing such 

populations by animal right groups that demand no harm to any individual animals, despite them being the 

source of important ecological damage and suffering by native animals. For biological control, a similar 

difference exists between those focused on risks to non-target species populations versus others for whom 

the death of some individuals of non-target insects or plants is unacceptable. 

Demands by Regulators to Identify Food Web-based Risks

Previously, the standard that new biological control agents had to meet was to demonstrate low risk of 

directly attacking native species at population-damaging levels. However, recently the US Department 

of Interior (responsible for protection of wildlife) has expanded this standard to require consideration of 

possible indirect, food web-mediated effects. While doing so has some potential benefits, two points need 

to be considered. First, introductions for no other purpose other than biocontrol are required to meet such 

a standard – not introductions made of horticultural plants, pets, species used in aquaculture, or other 

groups. Second, while host ranges of biocontrol agents are susceptible to being estimated by following an 

understandable and predictable set of tests, the same is not true of indirect effects. Indirect effects, rather, 

have no common thread and can only be sought and tested for one by one and only if the type of interaction 

can be imagined. That said, there are some categories of indirect effects that might be reasonable to test 
Biological Control in Natural Areas

76 Chapter 11: Future Trends in Use of Natural-area Biocontrol

for, including potential toxicity of the agent to native species that might eat it. This risk, for example was 

considered in the case of the introduction of Rodolia cardinalis (Mulsant) to the Galápagos to ensure safety 

to native land birds, should they prey on the agent (Lincango et al., 2011). More comprehensive evaluation 

of indirect effects may be so unending as to be not a standard but a total barrier to agent introduction, 

depending on the inclination of the reviewing agency.

New Players with Old Values

A further complication likely to affect whether and how well biological control will be used for the protection 

of nature is that the group of countries practicing biocontrol introductions may increase in the future, 

bringing new regulatory systems (or countries with no system at all) and values (either highly conservative 

or rather careless) into play. Countries facing threats to food supplies that are only marginally secure or 

threats of important economic losses can be expected to weigh those things heavily, perhaps not being as 

concerned about potential for non-target effects. 

REQUIRED ACTIONS

To enhance the frequency and quality of biocontrol projects in natural areas, several actions would be 

useful, including the following:

• Use high quality taxonomy to characterize the agent correctly BEFORE its release. 

• When several agents are known at the same time, release them, if possible, in order of their likely 

impact.

• Expect some non-target host use, but work to avoid population-level impacts.

• Develop on-line summaries of relevant information on natural enemy releases. 

• Plan biological control projects jointly with conservation partners.

• Conduct post-release studies of agents on the target and on non-target species believed to potentially 

be attacked or plausible negative indirect effect.
Biological Control in Natural Areas

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