Natural Enemies of Wood-Boring Beetles in Northeastern Temperate

Natural Enemies of Wood-boring Beetles in Northeastern Temperate Forests and Implications for Biological Control of the Emerald Ash Borer (Coleoptera: Buprestidae) in North America

Justin Michael Gaudon

A thesis submitted in conformity with the requirements for the degree of Doctor of Philosophy in Forestry

Faculty of Forestry University of Toronto

Natural Enemies of Wood-boring Beetles in Northeastern Temperate Forests and Implications for Biological Control of the Emerald Ash Borer (Coleoptera: Buprestidae) in North America

Justin Michael Gaudon

Doctor of Philosophy in Forestry

Faculty of Forestry University of Toronto

2019

Abstract

The emerald ash borer (hereafter EAB), Agrilus planipennis Fairmaire (Coleoptera: Buprestidae), is a wood-boring beetle accidentally introduced into North America during the 1990s, and has since been killing millions of ash trees, Fraxinus spp. L. (Lamiales: Oleaceae), as it spreads across

Canada and the USA. Native North American natural enemies, especially parasitoid wasps, are important mortality factors of EAB, but little information is available on their arrival and detection in EAB-infested regions, and their feasibility for augmentative biological control against EAB is uncertain. Two important native parasitoid groups, Phasgonophora sulcata Westwood

(Hymenoptera: Chalcididae) and Atanycolus spp. Foerster (Hymenoptera: Braconidae), were investigated to determine (1) factors influencing their capacity to disperse, (2) vegetation and habitat characteristics influencing their local abundance and role in EAB mortality, (3) whether their populations can be augmented to increase parasitism of EAB, and (4) how best to detect and monitor their populations as EAB continues to spread. The weak dispersal capacity of P. sulcata suggests it should be released as pupae close to EAB if used in an augmentative biological control program. Forest vegetation and habitat structure determine local abundance of P. sulcata and

Atanycolus spp., and tree biomass, tree condition, and floral resource availability are important predictors of high parasitism on EAB. Relocating parasitoid-infested ash logs to EAB-infested

sites can significantly augment populations of native parasitoids to increase EAB parasitism by

64.8 ± 18.1 % three years after introduction. Purple prism traps can be used to detect and monitor changes in populations of relocated P. sulcata and Atanycolus spp. These findings improve our understanding of the role of native natural enemies in suppressing EAB population growth and slowing ash tree mortality in North America.

iii

Acknowledgements

First, I thank my academic supervisor, Prof. Sandy Smith, for accepting my application to do graduate work in her laboratory at the Faculty of Forestry, University of Toronto. Her strong mentorship throughout my PhD program has helped me grow as an independent researcher and a team leader in both a field and laboratory setting. She has shown me what it takes to produce rigorous, high-quality scientific research; how to effectively communicate my work; and has even calmed my anxiety during some of those trying times that many of us encounter at some point while completing a degree.

I also thank my supervisory committee members, Profs. Jeremy Allison, Chris Darling, and Sean Thomas for providing valuable advice and feedback throughout all stages of this work, including experimental planning and design, data collection, data analysis, interpretation of my results, and writing. Thank you to Dr. Taylor Scarr for also sitting on my supervisory committee as a “non-voting” member and providing the same support. Special thanks go to Dr. Danijela

Puric-Mladenovic for her immense support throughout my studies, her involvement with my research, and for participating as an external examiner on my qualifying exam committee and both departmental and final oral exam committees. Thank you to Dr. Richard Westwood for participating as the external appraiser on my final oral exam committee.

Early in my PhD program, I received mentorship and help starting one of my research projects from Dr. Lucas Roscoe. Later, I had help growing my knowledge of statistics, especially using R, and collecting parasitoids used for some of this work from Dr. Chris MacQuarrie. Thank you both for your direction and assistance during such critical times in my career.

Much of this research would not have been possible without the various in-kind support received from federal and provincial governments, municipalities, conservation authorities, and

other land owners. Thank you to Kristjan Vitols (City of Toronto), Stacey Bowman (York

Region), Phillip Davies (Lake Simcoe Region Conservation Authority), Tys Theysmeyer and Dr.

David Galbraith (Royal Botanical Gardens), Curtis Marcoux and Candace Karandiuk (Town of

Oakville) for permission to collect data; and Kristjan Vitols (City of Toronto), Mary Orr

(Canadian Food Inspection Agency), and Al Foley and Sarah Drabble (Ontario Tree Seed Plant) for storage space. Thank you to Gene Jones and Drs. Krista Ryall and Barry Lyons at the Canadian

Forest Service for sharing their insight, especially with respect to collecting parasitoids.

It would have been difficult to complete such a large body of work without help from research assistants. Thank you to Jing (Iris) Hu and Charlotte De Keyzer for their help in the laboratory and especially in the field when we worked through hot or rainy weather or began working before sunrise.

Special thanks to my colleague and good friend, Dr. M. Lukas Seehausen. This research would certainly not be what it is without our fruitful conversations about invasive forest insects, parasitoids, and statistics. Many thanks to Richard Dickinson for both resources and assistance with plant identification throughout my thesis work, especially Chapter 3. Thanks to Rhoda

DeJonge, Susan Frye, Janani Sivarajah, Nigel Gale, Dr. Nurul Islam, Eric Davies, Nicolas

Tanguay, Kenneth Dearborn, Ian Jones, and others who have been excellent colleagues that challenged me to do the best that I can. I appreciate that you were there to celebrate the highs and help me through the lows of completing a PhD.

I express my gratitude to my mother-in-law, Susan Ingram; Aunt Geni; parents, Raymond and Judy Gaudon; and grandparents, Donald and Inez Gannett, for their support throughout my academic career; and my dogs, Marley, Macey, and the late Abby, for always giving me the time to take a walk.

Last, my biggest thank you goes to my wife and best friend, Michelle Gaudon, for her continued love and endless support throughout my PhD program and as I proceed throughout my career. She is and will always be my rock. I dedicate this to you.

Table of Contents

Acknowledgements...... iv Table of Contents ...... vii List of Tables ...... ix List of Figures ...... x Appendices ...... xii

Chapter 1 General Introduction...... 1 1.1 Invasive Species and their Success in Novel Environments ...... 1 1.2 Emerald Ash Borer in North America ...... 2 1.3 Parasitoid Definitions and Natural History ...... 5 1.4 North American Natural Enemies of Agrilus (Coleoptera: Buprestidae) ...... 6 1.5 Biological Control Using Parasitoid Wasps ...... 10 1.6 Adult Parasitoid Dispersal ...... 13 1.7 Spatial Distribution of Ash Trees, Emerald Ash Borer, and its North American Parasitoids ...... 15 1.8 Sampling Parasitoid Populations ...... 17 1.9 Thesis Objectives and Outline ...... 18

Chapter 2 Factors Influencing the Dispersal of a Native Parasitoid, Phasgonophora sulcata, Attacking the Emerald Ash Borer: Implications for Biological Control...... 24 2.1 Abstract ...... 24 2.2 Introduction ...... 25 2.3 Materials and Methods ...... 29 2.4 Results ...... 32 2.5 Discussion ...... 36

Chapter 3 Vegetation and Habitat Characteristics Affect the Abundance and Impact of North American Parasitoids Attacking the Emerald Ash Borer ...... 46 3.1 Abstract ...... 46 3.2 Introduction ...... 47 3.3 Materials and Methods ...... 50 3.4 Results ...... 54 3.5 Discussion ...... 56

vii

Chapter 4 Augmenting Populations of North American Parasitoids for Biological Control of the Emerald Ash Borer ...... 67 4.1 Abstract ...... 67 4.2 Introduction ...... 68 4.3 Materials and Methods ...... 72 4.4 Results ...... 75 4.5 Discussion ...... 76

Chapter 5 Evaluating Methods to Detect and Monitor North American Parasitoids of the Emerald Ash Borer ...... 83 5.1 Abstract ...... 83 5.2 Introduction ...... 84 5.3 Materials and Methods ...... 86 5.4 Results ...... 88 5.5 Discussion ...... 90

Chapter 6 General Conclusions and Future Research ...... 96 6.1 Ecological Significance ...... 98 6.2 Limitations and Future Research ...... 103

Literature Cited ...... 107

Appendices ...... 121 Appendix A ...... 121 Appendix B ...... 128

viii

List of Tables

Table 4.1 Mean percentage parasitism (± 1 standard error) by hymenopteran parasitoid species and mean woodpecker predation rates (± 1 standard error) on sampled emerald ash borer in plots where parasitoid releases were made annually (2013: n = 65, 2014: n = 45, 2015: 37, 2016: n = 33) compared to non-release control plots (2013: n = 65, 2014: n = 6, 2016: n = 34) in Toronto, Ontario, Canada.

List of Figures

Figure 1.1 Dolichomitus sp. CNC454923 (top) and Dolichomitus sp. CNC454924 (bottom) collected by J.M. Gaudon in Moreau Trail Park, Toronto, Ontario in 2014. Images were taken by the Canadian National Collection of Insects, Arachnids, and Nematodes.

Figure 1.2 Barcode of Life Data (BOLD) System TaxonID Tree for 35 sequence counts of 11 Dolichomitus spp. Five species were unidentified. Two species of Dolichomitus collected by J.M. Gaudon in Moreau Trail Park, Toronto, Ontario and previously undescribed in BOLD are highlighted in yellow. Sequences were run by the Canadian Centre for DNA barcoding.

Figure 1.3 Recorded distribution of Phasgonophora in North America based on a review of museum collections, scientific literature, and other reputable online databases.

Figure 2.1 Relationship between flight speed and female Phasgonophora sulcata body size tested on a flight mill for 24 h at 21.0, 24.0, 24.5, or 25.0 °C under controlled conditions (i.e. L:D = 16:8 diel period and 50-70 % RH) (n = 30). The solid line shows the fit and the dotted lines show the 95 % confidence intervals.

Figure 2.2 Relationship between the number of flight bouts taken by Phasgonophora sulcata and wasp age tested on a flight mill for 24 h at 21.0, 24.0, 24.5, or 25.0 °C under controlled conditions (i.e. L:D = 16:8 diel period and 50-70 % RH) (n = 30). Log-transformed data were back- transformed for presentation. The solid line shows the fit and the dotted lines show the 95 % confidence intervals.

Figure 2.3 Relationship between Phasgonophora sulcata body mass lost and flight speed tested on a flight mills for 24 h at 21.0, 24.0, 24.5, or 25.0 °C under controlled conditions (i.e. L:D = 16:8 diel period and 50-70 % RH) (n = 30). Log-transformed data were back-transformed for presentation. The solid line shows the fit and the dotted lines show the 95 % confidence intervals.

Figure 2.4 Relationship between potential fecundity in Phasgonophora sulcata after each flight period and wasp age (n = 25) tested on flight mills in the lab. The solid line shows the fit and the dotted lines show the 95 % confidence intervals.

Figure 3.1 The sampling design used in the Vegetation Sampling Protocol (VSP) includes one main plot and five sub-plots. Adapted from Puric-Mladenovic and Kenney (2016).

Figure 3.2 Parasitism of emerald ash borer (EAB) by native parasitoids, Phasgonophora sulcata and Atanycolus spp., within all plots (n = 30) in southern Ontario, Canada. EAB parasitism was inferred from sticky band trap catches of P. sulcata, Atanycolus spp., and EAB between May and September 2017.

Figure 3.3 Relationship between the number of Phasgonophora sulcata captured and (a) mean ash condition, (b) % ash biomass, (c) % alternative host tree (i.e. birch, poplar, and oak) biomass, (d) mean number of stems of flowering herbaceous ground vegetation per subplot, and (e) total number of emerald ash borer (EAB) captured per plot (n = 30) across southern Ontario between May and September 2017.

Figure 3.4 Relationship between the probability of capturing Atanycolus spp. and (a) % alternative host tree (i.e. birch, poplar, and oak) biomass, (b) mean number of stems of flowering herbaceous ground vegetation per subplot, and (c) canopy closure within 30 plots across southern Ontario between May and September 2017.

Figure 3.5 Relationship between overall parasitism of emerald ash borer (EAB) by Phasgonophora sulcata and Atanycolus spp. and (a) mean ash condition, (b) % alternative host tree (i.e. birch, poplar, and oak) biomass, and (c) mean number of stems of flowering herbaceous ground vegetation per subplot within 30 plots across southern Ontario between May and September 2017.

Figure 4.1 Mean parasitism of emerald ash borer (EAB) larvae by native North American parasitoids in parasitoid-release plots (2013: n = 65, 2014: n = 45, 2015: 37, 2016: n = 33) and non-release control plots (2013: n = 65, 2014: n = 6, 2016: n = 34) in Toronto, Ontario between 2013 and 2016. Significant differences between treatments and years at P < 0.05 according to Tukey’s range test are shown by different lowercase letters. Error bars are ± 1 standard error from the mean.

Figure 5.1 Number of Phasgonophora sulcata captured on green prism traps baited with a green leaf volatile (GLV) (n = 3), unbaited green prism traps (n = 3), purple prism traps baited with a GLV (n = 3), unbaited purple prism traps (n = 3), and sticky band traps (n = 30) in the McKeough Conservation Area site in Ontario, Canada during 2010. Significant differences at P < 0.05 between trap type are indicated by different lowercase letters. Error bars are ± 1 standard error from the mean.

Figure 5.2 Number of Phasgonophora sulcata captured on purple prism traps (PPT) (n = 15), sticky band traps (SBT) (n = 15), yellow pan traps (YPT) (n = 15), and emerging from log samples (n = 36) in Oakville, Ontario during 2016 and 2017. Significant differences at P < 0.05 between trap type and year are indicated by different lowercase letters. Error bars are ± 1 standard error from the mean.

List of Appendices

Appendix A (to Chapter 3) Total number of trees (N), relative tree species abundance, and corresponding mean diameter at breast height (DBH) ± 1 standard deviation measurements within all 30 plots across the study area in southern Ontario. Standing dead trees and upright coarse woody debris are classified as snags.

Appendix B (to Chapter 3) Ground vegetation within all 30 plots and 150 subplots across the study area in southern Ontario.

xii

Chapter 1 General Introduction

1.1 Invasive Species and their Success in Novel Environments

Invasive species threaten biodiversity (Pimentel et al. 2004), introduce diseases (Kettunen et al.

2009), and result in large-scale ecological disturbance and economic damage (Olson 2006). While some disagreement surrounds the term ‘invasive’ (Colautti and MacIsaac 2004), here it is defined as a population of a species that is able to spread and become ecologically, culturally, and/or economically damaging to a geographic region, irrespective whether the species is native or introduced to that area. In forest ecosystems, invasive insects can damage or even destroy extensive areas of forest that have important ecological function and may also be commercially valuable and/or culturally significant [e.g. mountain pine beetle, Dendroctonus ponderosae

Hopkins (Coleoptera: Curculionidae) (Kurz et al. 2008; Corbett et al. 2016)].

Several ecological theories that attempt to explain the success of invasive species, especially those non-native to a region, have been proposed, including: (1) biotic resistance, (2) island susceptibility, (3) invasion meltdown, (4) novel weapons, (5) host defence release, and (6) enemy release. Biotic resistance hypothesizes that biodiverse ecosystems are better at resisting invasion by species than ecosystems with low biodiversity (Elton 1958), while the island susceptibility hypothesis suggests that islands are more susceptible to invasive species than continents (Elton 1958). The invasion meltdown hypothesis advances the idea that invasive species are linked and the introduction of one into an ecosystem will promote the establishment of others in the same ecosystem, ultimately leading to ecological collapse of the native ecosystem

(Simberloff and Von Holle 1999). The novel weapons hypothesis suggests that invasive species are successful because they have a novel trait that allows them to outcompete other species in the

2 community (Callaway and Ridenour 2004), while both the defence and enemy release hypotheses maintain that it is the absence of host defences or natural enemies, respectively, that allows for the relatively easy establishment of invasive species (Liebhold et al. 1995; Gandhi and Herms

2010b). Although there are a variety of competing mechanisms, several of these theories may act together to facilitate the success of invasive species. If an invasive species does establish and its establishment results in the damage or destruction of native populations, communities, or ecosystems, human intervention is usually necessary in order to mitigate these impacts and control the invasive population.

1.2 Emerald Ash Borer in North America

The emerald ash borer (EAB), Agrilus planipennis Fairmaire (Coleoptera: Buprestidae), is a metallic wood-boring beetle accidentally introduced from Asia into North America and first discovered in the Detroit-Windsor area during 2002 (Haack et al. 2002). Although discovered in

2002, it was determined to have arrived in North America during the 1990s (Siegert et al. 2014) likely within infested pallets, crating, or dunnage. During its larval stage, EAB can kill all five eastern North American species of ash trees, Fraxinus L. (Lamiales: Oleaceae), including F. pennsylvanica Marshall (green or red ash), F. americana L. (white ash), F. nigra Marshall (black ash), F. quadrangulata Michx. (blue ash) (Anulewicz et al. 2008), and F. profunda (Bush) Bush

(pumpkin ash) (Czerwinski et al. 2007) as it feeds on the inner phloem, cambium, and outer xylem of the tree (Cappaert et al. 2005). Concerns have also been expressed about the possibility of EAB attacking non-ash species, especially species from the family Oleaceae. Agrilus marcopoli

Obenberger and Agrilus marcopoli ulmi Kurosawa (= EAB) (Jendek and Grebenikov 2011) have been observed developing on species of Juglans Linnaeus (Juglandaceae), Pterocarya Kunth

(Juglandaceae), and Ulmus Linnaeus (Ulmaceae) (Ko 1969; Akiyama and Ohmomo 1997).

3 Further, EAB has been recently observed completing development on white fringetree,

Chionanthus virginicus L. (Lamiales: Oleaceae), a novel host in Ohio, USA (Cipollini 2015;

Peterson and Cipollini 2016) and on cultivated olive, Olea europaea L. (Lamiales: Oleaceae)

(Cipollini et al. 2017).

EAB is considered one of the most damaging and costly invasive forest insects to invade

North America to date (Herms and McCullough 2014). There are 282 arthropod species that feed on ash, 43 North American species of which are monophagous and at risk of extinction as a result of the extensive ash mortality caused by EAB (Gandhi and Herms 2010a). Estimated economic impacts are substantial [e.g. up to $2 billion to remove, replace, and treat street and backyard ash trees in their native range in Canadian municipalties (McKenney et al. 2012)]. Further, ash wood is important for making furniture, baseball bats, and guitars (Poland and McCullough 2006), and black ash specifically is used for basketry by indigenous communities in northeastern North

America (Willow 2011).

The abundance of ash trees in the North American landscape, coupled with relaxed host defences and pressure from natural enemies, are important to the successful invasion of EAB, and can be used to predict the widespread ash mortality by EAB observed across five Canadian provinces and 35 US states. Some natural host tree resistance (Anulewicz et al. 2007; Tannis and

McCullough 2012), extreme cold (Crosthwaite et al. 2011) and cold-warm-cold fluctuations in temperature (Sobek-Swant et al. 2012), predators (Rutledge et al. 2013; Jennings et al. 2015), parasitoids (Duan et al. 2009; Duan et al. 2012), and pathogens (Bauer et al. 2004; Kyei-Poku and

Johny 2013) have been reported as mortality factors of EAB in North America, however, to date, none have been able to naturally regulate outbreaking populations of EAB.

Early attempts to control EAB involved quarantines to limit its spread and eradication of smaller, satellite populations (Cappaert et al. 2005). The rapid spread of EAB and its cryptic

4 nature as a wood-boring beetle made eradication efforts difficult and funding eventually ceased leading to the adoption of a ‘slow-the-spread strategy’ to reduce the impact of EAB in North

America (Liu et al. 2003; Herms and McCullough 2014). This strategy involves the early detection of the pest (or weed), the suppression of outbreaking populations, followed by efforts to slow its spread and population growth, as was seen with gypsy moth, Lymantria dispar L.

(Lepidoptera: Erebidae), management in North America during the 20th century (Sharov et al.

2002). Slow the spread strategies include methods to regulate or modify pest (or weed) populations as part of a suppression tactic as well as detect and eradicate the pest on the the leading edge of invasion. To date, management tools that have been used in a slow-the-spread strategy to combat EAB include quarantines of EAB-infested material [i.e. restrictions on the movement of ash trees and ash firewood (Poland and McCullough 2006)], chemical treatment [i.e. with pesticides, such as azadirachtin (TreeAzin) (McKenzie et al. 2010), emamectin benzoate, and imidacloprid (Smitley et al. 2010)], tree removal, and biological control, especially using exotic insect parasitoids (Herms and McCullough 2014). Success in the slow-the-spread gypsy moth program has been measured when low gypsy moth populations are reached (e.g. ≥ 67 % reduction in gypsy moth populations in treated compared to untreated areas) (Tobin and Blackburn 2007).

Similarly, the goal in the EAB system is to slow EAB population growth in order to reduce the rate of overall ash tree mortality and allow ash regeneration (e.g. the SLAM or Slow Ash Mortality project) (Herms and McCullough 2014). In this approach, the amount of ash available for attack by EAB is reduced and beetles are killed prior to adult dispersal and further reproduction (Herms and McCullough). A slow-the-spread approach to managing EAB gives municipalities, conservation authorities, other land owners, and researchers more time to prepare, budget, and respond to EAB infestations compared to reactively managing large numbers of dying and dead trees. Extending the timeline of peak ash mortality allows tree inventories to be conducted or updated so that (1) large/mature street and boulevard ash trees can be identified for protection

5 with systemic insectide, (2) trees can be underplanted in areas where ash will be cut, (3) biological control can be planned in woodlots and naturalized areas, and (4) EAB-resistant ash cultivars can be developed and propagated through hybridization with North American and Asian ash species.

1.3 Parasitoid Definitions and Natural History

In contrast to predators, which usually kill more than one prey, and parasites that feed on one or more hosts without killing their host, parasitoids are organisms whose larvae feed on or within another organism, typically consuming only one host, and ultimately killing them (Godfray 1994).

Parasitoids and their hosts are usually arthropods, and insects (Insecta) are commonly both the parasitoid and host. Principle insect orders wherein the parasitoid habit has evolved include

Hymenoptera and Diptera, but this habit has also evolved in other orders including Coleoptera,

Lepidoptera, and Neuroptera (Godfray 1994).

There are many ways to classify parasitoids including (1) where the host is located (i.e. exposed or concealed parasitoids), (2) where the parasitoid feeds (i.e. ectoparasitoids or endoparasitoids), (3) the number of parasitoids that attack a single host (i.e. solitary or gregarious parasitoids), (4) the life history stage of the host attacked (i.e. egg, egg-larval, larval, pupal, or adult), (5) whether the parasitoid paralyzes the host or allows it to continue developing (i.e. idiobiont or koinobiont parasitoids), (6) the parasitoid’s reproductive strategy (i.e. pro-ovigenic or synovigenic parasitoids), or (7) whether the parasitoid operates as a primary or hyperparasitoid

(Godfray 1994; Quicke 2015). Although these classifications separate parasitoid groups, they are tightly linked. Parasitoids that allow their host to continue to develop after parasitism occurs are usually concealed and develop within the host. Thus, koinobiont parasitoids are usually specialists

[e.g. as found in the New World Braconidae, Althoff (2003)] as they must overcome a specific

6 host immune system as endoparasitoids. Development of idiobiont parasitoids that paralyze their host during parasitism such that the host can no longer develop typically occurs on, rather than within, the host and these parasitoids are exposed to some degree. Thus, idiobiont parasitoids that operate as ectoparasitoids are usually generalists [e.g. Althoff (2003)] since they do not need to overcome the host immune system. Primary parasitoids are those that attack other insects that have not evolved with the parasitoid habit, whereas hyperparasitoids are parasitoids of other parasitoids (Godfray 1994; Quicke 2015).

1.4 North American Natural Enemies of Agrilus beetles

There are more than 180 indigenous Agrilus Curtis (Coleoptera: Buprestidae) beetles in North

America, including common species such as bronze birch borer, A. anxius Gory, and two-lined chestnut borer, A. billineatus (Weber) (Paiero et al. 2012). Agrilus wood-boring beetles have a complex of natural enemies, especially parasitoids, that could possibly switch from native hosts when high populations of a closely-related species, such as EAB, co-occur. Given the lack of co- evolutionary history between these parasitoids and EAB, such a 'jump' in host breadth by native natural enemies to an introduced, non-native host such as EAB would be novel and worth exploring.

Woodpeckers (Aves: Picidae) are often observed attacking EAB throughout its introduced range. Of these, hairy [Picoides villosus (Linnaeus)], downy [Picoides pubescens (Linnaeus)], and red-bellied [Melanerpes carolinus (Linnaeus)] woodpeckers are the most common birds to prey on EAB (Lindell et al. 2008), but all are important mortality factors of EAB populations with as high as 95 % mortality observed on EAB populations from woodpecker predation (e.g.

Cappaert et al. 2005). Although woodpeckers show potential for slowing the spread of EAB

7 population growth and ash mortality, birds are difficult to manipulate for biological control purposes. Thus, it is not surprising that there is little work on augmenting woodpecker populations. However, providing suet in EAB-infested sites may attract and retain woodpeckers and significantly impact EAB populations (Poland and McCullough 2010).

Cerceris fumipennis Say (Hymenoptera: Crabronidae) is a solitary, predatory wasp native to northeastern North America that has been reported to provision its nest preferentially with buprestid beetles and has been observed specifically collecting Agrilus beetles, including EAB

(Hook and Evans 1991; Rutledge et al. 2013). It has recently been proposed as a tool to detect and monitor populations of buprestid beetles (Careless et al. 2014), but it is unlikely that this ground- nesting wasp can regulate EAB populations with such a broad range of prey.

Observations of other predators attacking EAB include the predatory beetles Enoclerus

Gahan (Coleoptera: Cleridae) spp., Catogenus rufus (Fabricius) (Coleoptera: Passandridae), and

Tenebroides Piller and Miterpacher (Coleoptera: Trogossitae) spp., all which feed on EAB larvae

(Liu et al. 2003; Bauer et al. 2004). Little work has explored their use as biological control agents, likely because predators, like C. fumipennis discussed above, are typically generalists and, as such, not often desirable to be manipulated for augmentative biological control.

In North America, EAB have also been observed to be infected by several species of native entomopathogenic fungi, including Beauveria bassiana (Balsamo) Vuillemin (Hypocreales:

Clavicipitaceae), Isaria farinosa (Holmsk.) Fr. (Hypocreales: Cordycipitaceae) (= Paecilomyces farinosus), Isaria fumosorosea Wize (Hypocreales: Cordycipitaceae) (= Paecilomyces fumosoroseus), Lecanicillium lecanii Zare and Gams (Hypocreales: Cordycipitaceae) (=

Verticillium lecanii), and Metarhizium anisopliae (Metchnikoff) Sorokin (Hypocreales:

Clavicipitaceae) (Bauer et al. 2004). Unfortunately, the actual number of individual EAB infected by fungal pathogens is often low [e.g. < 2 % of EAB larvae as observed by Bauer et al. (2004)].

8 Fungal pathogens with a broad host range are also not appealing for biological control because they can have significant non-target effects, although a few may have potential for use against

EAB (e.g. Dean et al. 2012).

Three exotic parasitoids, Tetrastichus planipennisi Yang (Hymenoptera: Eulophidae),

Oobius agrili Zhang and Huang (Hymenoptera: Encyrtidae), and Spathius agrili Yang

(Hymenoptera: Braconidae), were identified in EAB’s native range and observed parasitizing ≥

50 % of EAB in some sites (e.g. Liu et al. 2003; Liu et al. 2007). Since 2007, these parasitoids have been released in various locations across Canada and the USA for biological control against

EAB (Duan et al. 2017). Recent studies show that T. planipennisi and O. agrili have established in some release locations and have promise for regulating EAB populations (e.g. Abell et al. 2014;

Duan et al. 2017). Although S. agrili has been occasionally recovered after releases, less certain is its establishment (Bauer et al. 2015) perhaps due climatic factors (e.g. cold intolerance in northerly latitudes, Hanson et al. (2013)]. Consequently, approval for the import and release of another exotic parasitoid, Spathius galinae Belokobylskij and Strazanac (Hymenoptera:

Braconidae), occurred in 2015 (Federal Register 2015). Although parasitism rates are quite high in both their native and introduced ranges, I do not pursue this aspect of classical biological control in this thesis as I am particularly interested in the novel associations between EAB and its new natural enemies.

My work herein explores North American parasitoids attacking EAB because they are relatively host-specific compared to the predators and pathogens discussed above, and thus desirable as biological control agents (Godfray 1994; Hare and Weseloh 2009). Native natural enemies, such as Phasgonophora sulcata Westwood (Hymenoptera: Chalcididae) in Ontario,

Canada and Atanycolus cappaerti Marsh and Strazanac (Hymenoptera: Braconidae) in Michigan,

USA, are considered important due to high rates of observed parasitism on EAB (e.g. Cappaert

9 and McCullough 2009; Lyons 2010). In particular, P. sulcata has recently been considered for use in augmentative releases against EAB (Roscoe 2014). Other native species of Atanycolus

Foerster (Hymenoptera: Braconidae) that have been reared from EAB in Ontario include A. hicoriae Shenefelt, A. tranquebaricae Shenefelt, A. disputabilis Cresson, and A. longicauda

Shenefelt (Roscoe 2014). Although not yet reported attacking EAB in Ontario, Canada, Spathius floridanus Ashmead (Hymenoptera: Braconidae) has also been shown to have potential for biological control of EAB in North America as it can be reared from EAB larval galleries in ash trees in Michigan, USA (Marsh and Strazanac 2009). Unfortunately, other wasp species, such as

Metapelma spectabile Westwood (Hymenoptera: Eupelmidae), have a relatively rare association with EAB (Lyons 2010) and thus have not been considered for biological control. Interestingly,

Balcha indica (Mani and Kaul) (Hymenoptera: Eupelmidae) has been observed attacking EAB in

North America, although it is a non-native parasitoid established in North America prior to the arrival of EAB (Gibson 2005; Duan et al. 2012). Other native parasitoids attacking EAB include

Leluthia astigma (Ashmead) (Hymenoptera: Braconidae), Eupelmus pini Taylor (Hymenoptera:

Eupelmidae), as well as Cubocephalus Ratzeburg (Hymenoptera: Ichneumonidae), Dolichomitus

(Hymenoptera: Ichneumonidae), Eurytoma Illiger (Hymenoptera: Eurytomidae), and Orthizema

Foerster (Hymenoptera: Ichneumonidae) (Duan et al. 2009, 2010, 2012; Kula et al. 2010). Rates of parasitism on EAB by many of these species are often low, so less emphasis is placed on them in this thesis.

Although the list of parasitoids attacking EAB above may seem comprehensive, novel associations continue to be observed in North America, including those that are likely of undescribed species (e.g. Dolichomitus sp. nov., Figs. 1.1 and 1.2). While surveys of newly- invaded areas show low parasitism of EAB by native parasitoids (e.g. Liu et al. 2003; Duan et al.

2010, 2012), other work suggests some species can parasitize > 40% of EAB larvae at some sites

10 [e.g. A. cappaerti, Cappaert and McCullough (2009); P. sulcata, Lyons (2010)]. These high parasitism rates and the fact that populations appear to span across North America (e.g. P. sulcata,

Fig. 1.3) suggest that there is an opportunity to use native parasitoids in augmentative biological control programs against EAB in North America. Although natural enemies native to North

America have this potential, they have received limited attention to date.

1.5 Biological Control Using Parasitoid Wasps

Biological control involves the use of one or more populations of parasitoids, predators, pathogens, antagonists, or competitors to reduce a pest's population to a new equilibrium position that is less economically damaging than if no action was taken (Van Driesche and Bellows 1996).

Three approaches to biological control are recognized: classical, augmentative, and conservation.

Classical biological control involves the introduction of a natural enemy from the pest's original geographic range into its introduced range, while the other two strategies operate independent of the target pest or natural enemy origin. Augmentative biological control is designed to supplement natural enemies by releasing them into the pest's introduced range while conservation biological control entails the modification or manipulation of the habitat to promote the presence of the pest's natural enemies.

Augmenting native natural enemy populations may be important in areas where they are not already sufficiently effective in controlling pest populations. Depending on the system, this approach may be cost-effective and environmentally sound in combating invasive species compared to traditional pesticide treatments or classical biological control (Van Driesche and

Bellows 1996). In Canada, 32 native (i.e. North American) natural enemy agents have been used in augmentative biological control. Of these, 12 have been released in areas where their

11 populations were already present and 19 were relocated to where agent populations were absent

(MacQuarrie et al. 2016). It is difficult to predict whether biological control of EAB using native natural enemies will be successful from these augmentative biological control programs because the definition of “success” is unique to each program and closely tied to its goal(s) and strategy.

However, augmenting native natural enemy populations to supress a pest population has been considered successful in other systems. Notable examples include the use of Trichogramma minutum Riley (Hymenoptera: Trichogrammatidae) against native spruce budworm,

Choristoneura fumiferana (Clemens) (Lepidoptera: Tortricidae) and native spruce bud moth,

Zeiraphera canadensis Mutt. and Free. (Lepidoptera: Tortricidae) (Smith et al. 1990; West et al.

2002); the use of a native parasitoid, Lathrolestes thomsoni Reshchikov (Hymenoptera:

Ichneumonidae), to control non-native ambermarked birch leafminer populations (MacQuarrie et al. 2013); and the use of introduced biological control agents coupled with native generalist predators to control winter moth, Operophtera brumata (L.) (Lepidoptera: Geometridae) (Roland and Embree 1995).

Two important native parasitoids in terms of their abundance and high parasitism rates of

EAB have been identified, namely Phasgonophora sulcata Westwood (Hymenoptera:

Chalcididae) and Atanycolus cappaerti Marsh and Strazanac (Hymenoptera: Braconidae). Both

P. sulcata and Atanycolus spp. are known to attack other North American Agrilus spp. (e.g.

Britton 1923; Barter 1957, 1965; Nash et al. 1951; Haack et al. 1981) and have recently been recorded on EAB in both Canada and the USA (e.g. Duan et al. 2012, 2015; Roscoe 2014; Roscoe et al. 2016). Phasgonophora sulcata is a solitary larval endoparasitoid known to attack two-lined chestnut borer (Haack et al. 1981), bronze birch borer (Barter 1957), and bronze poplar borer,

Agrilus liragus Barter and Brown (bronze poplar borer) (Barter 1965). Atanycolus cappaerti is a solitary larval ectoparasitoid also known to attack two-lined chestnut borer and bronze poplar

12 borer (Cappaert and McCullough 2009). Some studies report that P. sulcata and A. cappaerti can parasitize up to 40% and 71%, respectively, of EAB larvae at some sites (Cappaert and

McCullough 2009; Lyons 2010). These high parasitism rates, coupled with synchrony between the parasitoids and their preferred host stages (e.g. Roscoe et al. 2016), suggest that there is an opportunity to use native parasitoids in an augmentative biological control program against EAB.

Appropriate natural enemy selection, release, and recovery guidelines are critical for implementing a successful biological control program. For example, when choosing parasitoids to use in a biological control program, it is necessary to determine the effects of a parasitoid’s dispersal capacity on natural enemy efficacy. Parasitoids with intermediate dispersal rates are more likely to establish as biological control agents than parasitoids with low or high dispersal rates (see Heimpel and Asplen 2011). Understanding an agent's dispersal rate is economically relevant because it helps to determine where to focus biological control efforts.

Surrounding environmental conditions may also influence parasitoid populations and parasitism. Thus, ecosystem factors, such as vegetation and habitat characteristics, are important for choosing locations to release biological control agents. For example, parasitism rates of spruce budworm by a few pupal tachinid parasitoids decreased as stand density increased whereas there was no significant association between parasitism by larval hymenopteran parasitoids and stand density (Simmons et al. 1975) suggesting that stand characteristics influence the abundance and perhaps diversity of some natural enemy populations. Similar trends may be observed for native parasitoids of EAB, and this would have implications for EAB parasitism and utilizing parasitoids in a biological control program for EAB.

Developing reliable and cost-effective sampling methods to monitor natural enemies before and after their release is key to assessing their success in establishing and controlling the pest population (Van Driesche and Bellows 1996). Some work has explored effective traps for

13 sampling native EAB parasitoids (e.g. Roscoe 2014), but proof of concept remains to be tested.

To date, no studies on dispersal or ecosystem characteristics have been conducted on native parasitoids of EAB, and more research on trap efficacy is needed in order to accurately detect and monitor their populations.

1.6 Adult Parasitoid Dispersal

Dispersal is a normal feature of many insect life cycles used to escape predation and parasitism, colonize ephemeral habitats and novel geographic regions, avoid crowding, and increase genetic variability to avoid inbreeding (Wellington 1980; Schowalter 1985, 2000). Here, an insect's dispersal capacity is defined as the maximum movement of individuals by migratory, trivial, and passive locomotion. Dispersal through immigration and emigration is in part a function of population growth. Its economic relevance for pest insects is important for forecasting pest pressure, deciding where to focus pest management efforts, and, ultimately, creating an integrated pest management (IPM) program. Dispersal is influenced by both abiotic and biotic factors and responses to these factors greatly vary across different insect groups. This creates a necessity for having a detailed account of the flight capabilities of pest insects for a successful IPM program.

It is also critical to evaluate the dispersal capacity of an insect pest’s natural enemies to forecast their spread relative to the pest and predict negative impacts from Allee effects on a natural enemy’s population if they are to be utilized in a biological control program. Further, given that the dispersal of individuals is the link between fragmented habitat patches and a function of patch quality, knowledge of a population’s dispersal behaviour is key to developing cost-effective strategies for successful biological control, restoration, and conservation programs on a landscape level (see Huxel and Hastings 1999; Heimpel and Asplen 2011; With 2002).

14 Insect dispersal can be measured using various methods, including release-recapture techniques, such as marking a species or releasing a species into an environment where it does not occur; wind tunnels; flight mills; and using a priori knowledge to model dispersal (see

Osborne et al. 2002). The use of flight mills to quantitatively compare the dispersal capacity of insects has been demonstrated in previous work (e.g. Taylor et al. 2010; Fahrner et al. 2014;

Gaudon et al. 2016), and they are particularly valuable tools when trying to control for confounding variables in an experiment.

Relative to other insects, few data are available on parasitoid wasp dispersal. This is likely due to their small size. Of what we know, dispersal capacities vary across parasitoid taxa. For example, the eulophid wasp Tetrastichus planipennisi Yang (Hymenoptera: Eulophidae), an important parasitoid of EAB, flew an average of 1.26 km in 24 h on flight mills (Fahrner et al.

2014) whereas the braconid wasp Microplitis mediator (Haliday) (Hymenoptera: Braconidae) flew > 4 times as far on average [i.e. 5.27 (± 0.51) km] in 24 h on flight mills (Yu et al. 2009).

Data on the dispersal capacity of native North American wood-boring parasitoids are lacking, and, more specifically, little is known about how biotic and abiotic factors affect their dispersal capacity. Further, little information is available on how different modes of dispersal, such as flight and walking, in parasitoids interact, however a few studies have described polymorphic parasitoids with different dispersal behaviours that may dictate the capacity a species has to disperse [e.g. Melittobia spp. (Hymenoptera: Eulophidae), (Matthews et al. 2009)]. Such polymorphism is different from any innate variation in dispersal among individuals of the same population and is brought about by environmental cues as discussed by Clark (1976). If native parasitoids are to be used in a biological control program against EAB, we must study their dispersal capacities to guide biological control efforts.

15 1.7 Spatial Distribution of Ash Trees, Emerald Ash Borer, and its North

American Parasitoids

The spatial distribution of individuals within a population can be characterized as random, uniform, or clumped [as discussed by Schowalter (2000)]. If populations are distributed evenly throughout an area (i.e. uniform), it is likely that they will remain at an equilibrium or grow at a steady rate. In situations where populations are unevenly distributed (i.e. random or clumped), populations will likely grow faster in more dense areas and slower in sparsely distributed areas.

The spatial distribution of a population is not solely influenced by environmental conditions or its habitat but by species interactions as well.

Many field studies have investigated the spatial distribution of parasitism to determine how spatial heterogeneity influences host-parasitoid population dynamics (see Hassell and Pacala

1990). For example, a habitat patch with high host availability might be characterized as a high- quality patch that supports a stable or growing population whereas the opposite may be true for a patch with low host availability. Habitat patch quality is often indirectly measured. For example, the number of reproducing females or the total number of individuals within the habitat patch may be used as a proxy for patch quality, where many reproducing females or individuals reflect a high-quality patch and few reproducing females or individuals reflect a low-quality patch. Some animals, including parasitoid wasps, can use information that they obtain from their environment to make decisions about habitat patch quality and optimize their foraging for food and/or hosts such that they maximize their mean rate of gain of fitness (Charnov 1976). Information that the animal collects to determine whether it will remain within or disperse from a habitat patch includes the time it takes to travel between patches as well as the amount of resource(s) (e.g. the number of hosts, the number of host plants, and/or floral resource availability) within the patches

[e.g. Lysiphlebus testaceipes Cresson (Hymenoptera: Braconidae), Tentelier et al. 2006]. Thus,

16 these environmental conditions may dictate where native EAB parasitoids are present and absent across the landscape.

Prior to European settlement, the landscape in southern Ontario was largely comprised of forest. Pre-settlement forests were cleared and converted to mostly agriculatural and urban land.

Now, second and third-growth forest fragments, both structurally and compositionally different from pre-settlement forests, are scattered throughout the landscape (Riley 1999) and dominated by early-successional species, such as Betula spp. L. (Fagales: Betulaceae) (birch) and Populus spp. L. (Malpighiales: Salicaceae) (poplar), and common associated species, such as ash (Ontario

Ministry of Natural Resources 2000). Observations of ash can be dated to the early- to mid-1800s

(Lundy 1832), and it has become a prevalent tree genus and occurs in many stands across southern

Ontario as species within the genus can tolerate a range of soil moisture/drainage conditions

(Maycock 1963; Puric-Mladenovic 2003). Moreover, after the arrival of Dutch elm disease in

Canada in 1945 and subsequent death of numerous elm trees, Ulmus spp. L. (Rosales: Ulmaceae)

(Hubbes 1999), ash (typically green ash) was planted extensively due to its rapid growth and ability to survive as a street tree and fill canopy gaps across many urban forests. In extreme cases, the urban forests of many municipalities can have ≥ 25 % ash (e.g. cities of Montreal, Thunder

Bay, and Winnipeg). Such dominance by ash has likely increased the connectivity between natural forests, woodlots, plantations, and urban forests, and possibly facilitated the rapid spread of EAB throughout the landscape. Thus, connectivity between these habitat patches is valuable to EAB as a habitat corridor and enables it to disperse along with its native parasitoids.

Within ash trees, Timms et al. (2006) observed that EAB larval galleries were not distributed randomly or uniformly, but generally clumped on the southwest side of the trees. Thus, it could be hypothesized that EAB parasitoids would then have a clumped within-tree distribution as well. Interestingly, pro-ovigeny in parasitoid wasps is a life history strategy that is largely an

17 adaptation to a uniform distribution of hosts whereas synovigeny in parasitoid wasps can be attributed to a clumped distribution of hosts (Ellers and Jervis 2004), such as EAB. Although P. sulcata is a pro-ovigenic species, eggs in adult female wasps have been observed undergoing oosorption (Roscoe 2014), which may be a mechanism to gain energy for dispersal needed to overcome the patchy within-tree distribution of EAB or dispersal between patches. Further,

Roscoe et al. (2016) observed no difference in proportion of EAB parasitized by P. sulcata among different tree heights in young ash trees. However, the bark thickness of ash trees has a strong correlation with their size (Timms et al. 2006), so EAB parasitism by P. sulcata in the lower boles of large ash trees will likely be significantly less than in young ash trees due to its short ovipositor.

Similarly, the Asian EAB parasitoid, Tetrastichus planipennisi Yang (Hymenoptera: Eulophidae) cannot parasitize EAB in ash trees with a bark thickness > 3.2 mm (Abell et al. 2012). In these instances, Atanycolus spp. can fill this niche and parasitize EAB larvae in much larger ash trees with bark thicknesses up to 8.8 mm thick (Abell et al. 2012). Thus, differences in tree bark thickness among varying sizes of ash trees or between tree species of similar size [e.g. ash vs. birch, J.M. Gaudon (personal observation)] may lead to differences in the native EAB parasitoid community and ultimately parasitism of EAB.

1.8 Sampling Parasitoid Populations

Malaise traps, yellow pan traps, or a combination of both are often used to sample the parasitoid community within a region (e.g. Darling and Packer 1988; Noyes 1989). The detection and continuous monitoring of parasitoid populations is a critical component for determining the success of biological control efforts, especially the establishment of released enemy agents (Van

Driesche and Bellow 1996). Thus, these sampling tools and techniques are needed to monitor an

18 augmentative biological control program using native EAB parasitoids to combat invasive EAB populations.

Many sampling methods used for detecting and quantifying EAB populations provide no information on the presence of the parasitoid community. For example, signs and symptoms exhibited by EAB-infested ash trees have yet to be associated with native EAB parasitoids and girdled ash or birch trap trees do not appear to attract these parasitoids in the absence of EAB or without high numbers of a bronze birch borer host (J.M. Gaudon, personal observation). However,

Roscoe (2014) found that traps used to sample EAB show promise for also sampling native

Agrilus parasitoids (e.g. purple prism traps). Although this is a useful contribution, there are still more trapping methods to explore.

1.9 Thesis Objectives and Outline

Extensive work has been carried out on EAB over the past ~15 years since it was first introduced, but only a few studies have focused on native natural enemies. The long-term role of natural enemies in controlling EAB and similar wood-boring beetle populations in North America is not known, yet this is key to its long-term management. An important step in understanding the potential for these natural enemies to slow the spread of EAB across North America is collecting information about the behavioural and dispersal capacities of these native parasitoid species as well as the natural variation in the density and diversity of native parasitoids between forested sites. Thus, my research examines these aspects to understand the ecological drivers associated with the arrival and detection of native natural enemies attacking EAB in Canada. Based on this information, I attempt to augment these native parasitoid populations to combat EAB through transport of parasitoid-infested ash material. The research conducted in the course of this thesis

19 work adds to existing information on how to address the challenge of invasion by EAB into North

America’s northeastern temperate forests by using native parasitoids as a tool that can help in the long-term suppression and management of EAB as part of a larger slow-the-spread strategy.

The overall goal of my dissertation is to study the biology and ecology of native natural enemies of wood-boring beetles attacking trees in North America's northeastern temperate forests in order to determine their potential for regulating introduced Agrilus spp., especially EAB. Top- down regulation of EAB populations by a responding community of parasitoids may constrain

EAB populations from continuing to grow exponentially, although not eradicating EAB from its introduced range. It is largely accepted that EAB will continue to spread throughout the contiguous range of ash in North America, and this method of slowing its spread may bring EAB populations below the level of ecological and economic damage needed to allow ash populations to naturally regenerate.

Specifically, I study the flight capacity, walking activity, and fecundity of the native larval parasitoid, P. sulcata, and implications for biological control of EAB to identify key parameters for a release protocol if this wasp is to be used for augmentive biological control (Chapter 2). This is followed by Chapter 3 wherein the question of what forest vegetation and habitat characteristics best predict the abundance and impact of native EAB parasitoids so that necessary parasitoid- release locations can be determined. In this chapter, I tested the hypotheses that tree biomass, tree condition, floral resource availability, and canopy closure are mechanisms driving variability in native EAB parasitoid abundance and EAB parasitism across the landscape. Chapter 4 examines the efficacy of augmenting populations of native North American parasitoids in order to slow the spread of introduced EAB where I tested the hypothesis that native EAB parasitoids could be successfully relocated to increase parasitism of EAB. Chapter 5 explores differences between sampling methods to monitor the establishment and impact of relocated native North American

20 EAB parasitoids that can facilitate future sampling of native parasitoid populations. These studies will contribute to our overall understanding of the biology and ecology of important native parasitoids attacking EAB to improve their use in biological control programs. Chapter 6 provides an overall summary of the main findings and outlines the significance of these results in the context of existing literature in order to make recommendations for future research.

Chapters 2 through 5 are arranged as publishable manuscripts, with Chapter 2 already published in BioControl; as a result, repetition in the text may occur as each chapter is made to be succinct, especially within the Introduction sections. Multiple authors may be associated with each data chapter; however, I am the first author on all papers having conceived and designed the experiments, conducted the experiments, analyzed the data, and written each manuscript. All references throughout this thesis have been compiled and are listed in the Literatured Cited section at the end of the thesis. Appendices are also provided at the end for Chapter 3 so that raw data describing the study plots in greater detail are available.

21 Figures

2 % Dolichomitus dolichosoma|CNCHYM 013883|632[0n]bp|Canada.Ontario

Dolichomitus populneus|CNCHYM 014255|282[0n]bp|Canada.Manitoba

Dolichomitus messor|CNCHYM 09783|633[0n]bp|Canada.Manitoba

Dolichomitus populneus|CNCHYM 014256|249[0n]bp

Dolichomitus messor|CNCHYM 014247|623[2n]bp|Canada.Ontario

Dolichomitus messor|CNCHYM 014248|633[3n]bp|Canada.Ontario

Dolichomitus messor|CNCHYM 014251|244[0n]bp|Canada.Manitoba

Dolichomitus messor|CNCHYM 014250|632[0n]bp|Canada.Manitoba

Dolichomitus messor|CNCHYM 014246|164[0n]bp|Canada.British Columbia

Dolichomitus messor|CNCHYM 014249|262[0n]bp|Canada.British Columbia

Dolichomitus|CNCH0134|657[0n]bp|Canada.Manitoba

Dolichomitus californicus|CNCHYM 013894|613[1n]bp|Canada.British Columbia

Dolichomitus|07PROBE-22449|657[0n]bp|Canada.Manitoba

Dolichomitus foxleei|CNCHYM 013898|164[0n]bp|Canada.British Columbia

Dolichomitus foxleei|CNCHYM 013901|627[0n]bp|Canada.Ontario

Dolichomitus foxleei|CNCHYM 013902|634[0n]bp|Canada.Ontario

Dolichomitus irritator|CNCHYM 013909|634[0n]bp|Canada.Ontario

Dolichomitus irritator|CNCHYM 013910|634[0n]bp|Canada.Ontario

Dolichomitus irritator|CNCHYM 013911|279[0n]bp|United States.Iowa

Dolichomitus irritator|CNCHYM 013912|625[0n]bp|Canada.Quebec

Dolichomitus imperator|CNCHYM 013903|257[0n]bp|Canada.British Columbia

Dolichomitus terebrans|CNCHYM 015619|164[0n]bp|Canada.New Brunswick

Dolichomitus terebrans|CNCHYM 015620|164[0n]bp|Canada.New Brunswick

Dolichomitus terebrans|CNCHYM 015621|164[0n]bp|Canada.New Brunswick

Dolichomitus terebrans|CNCHYM 015622|164[0n]bp|Canada.New Brunswick

Dolichomitus terebrans|CNCHYM 015623|634[0n]bp|Canada.Ontario

Dolichomitus terebrans|CNCHYM 015625|633[0n]bp|Canada.Northwest Territories

Dolichomitus pygmaeus|CNCHYM 014257|238[0n]bp|Canada.Ontario

Dolichomitus|CNCHYM 013884|639[0n]bp|Canada.Ontario

Dolichomitus sericeus|CNCHYM 014258|164[1n]bp|Canada.British Columbia

Dolichomitus tuberculatus|CNCHYM 015631|633[0n]bp|Canada.Nova Scotia

Dolichomitus sericeus|CNCHYM 014260|273[0n]bp|Canada.Ontario

Dolichomitus tuberculatus|CNCHYM 015632|632[0n]bp|Canada.Nova Scotia

Dolichomitus|CNC454923|658[0n]bp|Canada.Ontario Moreau Trail Park, Toronto Dolichomitus|CNC454924|596[0n]bp|Canada.Ontario

Figure 1.3 Recorded distribution of Phasgonophora in North America based on a review of museum collections, scientific literature, and other reputable online databases.

24 Chapter 2 Factors Influencing the Dispersal of a Native Parasitoid, Phasgonophora sulcata, Attacking the Emerald Ash Borer: Implications for Biological Control1

2.1 Abstract

High parasitism by a native parasitoid, Phasgonophora sulcata Westwood (Hymenoptera:

Chalcididae), has been reported on emerald ash borer (hereafter EAB), Agrilus planipennis

Fairmaire (Coleoptera: Buprestidae), in North America. Use of this parasitoid in an augmentative biological control program has been proposed to slow the spread of EAB, yet information is lacking on key aspects of this parasitoid’s dispersal. We document the flight capacity and walking activity of P. sulcata, its potential fecundity, and describe how age, body size, temperature, and time of day affect these parameters. Wasp flight capacity, measured using flight mills, increased with temperature and decreased with age. Unexpectedly, age and body size did not affect wasp walking activity, and we saw no relationship between walking activity and flight capacity. Older wasps had lower potential fecundity than younger wasps. These results suggest that P. sulcata should be released as pupae near EAB-infested ash trees to improve efficacy and potential biological control success.

1 Reprinted by permission from Springer Nature: Gaudon, J.M., Allison, J.D., and Smith, S.M. (2018) Factors affecting the dispersal of a native parasitoid, Phasgonophora sulcata, attacking the emerald ash borer: implications for biological control. BioControl 63: 751-761. 25 2.2 Introduction

Understanding the dispersal of an insect parasitoid is an important consideration for successful biological control (Hopper and Roush 1993; Heimpel and Asplen 2011; Mills and Heimpel 2018).

Knowing the dispersal of a biological control agent can help optimize distances needed between release sites and the appropriate number of agents released in order to avoid negative impacts from Allee effects (Hopper and Roush 1993; Shea and Possingham 2000). Generally, parasitoids with an intermediate dispersal rate are more likely to establish than species with low or high dispersal rates because, for example, those with low dispersal rates may have low success locating hosts while those with high dispersal rates have increased risks of Allee effects (i.e. mate limitation at low population densities) (Heimpel and Asplen 2011). Active dispersal capacities, or a species’ propensity and potential ability to disperse (i.e. speed, distance, and activity) without assistance, vary among parasitoid taxa. For example, on a flight mill, Cotesia glomerata (L.)

(Hymenoptera: Braconidae) was observed to have a mean flight distance between 0.05 (± 0.01) km and 0.75 (± 0.20) km depending on the food source provided prior to flight (Wanner et al.

2006) whereas Microplitis mediator (Haliday) (Hymenoptera: Braconidae) had a mean flight distance of 5.27 (± 0.51) km after 24 h (Yu et al. 2009). The level of parasitism exerted on the target pest population, which ultimately will determine its ability to establish and control that population, is in part driven by its dispersal capacity. Therefore, characterizing the dispersal capacity of a biological control agent is an important consideration when developing any biological control program.

Morphological and life history traits, such as age and body size, as well as abiotic factors, such as temperature, can affect insect dispersal capacity. For example, in Sirex noctilio F.

(Hymenoptera: Siricidae), body size, temperature, and sex all influence flight speed, distance flown, and frequency of flight (Gaudon et al. 2016). Flight capacity has been reported to decrease

26 with increasing age in many insect orders, including Coleoptera [e.g. Tribolium castaneum

(Herbst) (Tenebrionidae), Perez-Mendoza et al. (2011)], Diptera [e.g. Aedes aegypti L.

(Culicidae), Rowley and Graham (1968)], Hemiptera [e.g. female Oncopeltus fasciatus (Dallas)

(Lygaeidae), Dingle (1965)], and Hymenoptera [e.g. Tetrastichus planipennisi Yang

(Eulophidae), Fahrner et al. (2014)]. In general, larger wasps have a greater flight capacity than smaller ones (e.g. Bruzzone et al. 2009; Fahrner et al. 2014), perhaps related to their greater energy stores and muscle mass. Other studies have shown that older parasitoids, especially those without access to carbohydrates and water, have reduced flight capacity and are less likely to survive than younger ones after being tethered to a flight mill for 24 h (e.g. Fahrner et al. 2014). Thus, body size, access to food, and other traits may all impact parasitoid performance, and having more information as to how they affect an individual species’ ability to disperse will help to better select the most appropriate biological control agent.

Similarly, insect fecundity is well documented to be affected by body size, age, and temperature. Within many parasitoid species, fecundity is known to increase with body size (e.g.

Waage and Ng 1984; Rosenheim and Rosen 1991), although this is often much more complex because other factors, such as longevity, can affect it and lead to an unclear relationship between these parameters (Leather 1988). Further, maximum fecundity occurs at an optimal temperature and decreases toward the upper and lower limits around that optimum (Ratte 1985). A parasitoid with high maximum lifetime (i.e. potential) fecundity should be able to parasitize more hosts than a parasitoid with low potential fecundity, and thus could be predicted to have higher realized fecundity under field conditions.

The emerald ash borer (EAB), Agrilus planipennis Fairmaire (Coleoptera: Buprestidae), was accidentally introduced into North America from Asia, and was first discovered in 2002 around the Detroit-Windsor area (Haack et al. 2002). Since then, it has become one of the most

27 damaging and costly insects invading North American forests (Herms and McCullough 2014), with an estimated impact of up to $2 billion for the removal, replacement, and treatment of street and backyard ash trees [Fraxinus L. (Lamiales: Oleaceae)] in Canadian urban areas throughout the natural distribution of native ash (McKenney et al. 2012). During its larval stage, EAB can kill all five eastern North American ash species it attacks, including F. pennsylvanica Marshall

(green or red ash), F. americana L. (white ash), F. nigra Marshall (black ash), F. quadrangulata

Michx. (blue ash) (Anulewicz et al. 2008), and F. profunda (Bush) Bush (pumpkin ash)

(Czerwinski et al. 2007). Concerns have also been expressed about the possibility of it attacking non-ash tree species in North America, given that it has been shown to complete development on white fringetree, Chionanthus virginicus L. (Lamiales: Oleaceae), a novel host in Ohio, USA

(Cipollini 2015), and in the laboratory on cultivated olive, Olea europaea L. (Lamiales: Oleaceae)

(Cipollini et al. 2017).

Eradication is no longer considered a viable option for managing EAB, so ongoing efforts aim to slow its spread across North America. Biological control is one of the few long-term tools available to incorporate into such a management strategy (Herms and McCullough 2014).

Phasgonophora sulcata Westwood (Hymenoptera: Chalcididae), a solitary larval parasitoid native to North America, is known to attack native Agrilus spp. (Coleoptera: Buprestidae) (e.g.

Barter 1957, 1965; Haack et al. 1981), and has been proposed as a candidate for augmentative biological control (e.g. Roscoe 2014), using it in supplemental releases across the introduced range of EAB. It is thought to have significant potential because it has been observed parasitizing up to 40.7 % of EAB larvae at some sites in southern Ontario, Canada (Lyons 2010).

The biology of many native North American parasitoids is poorly documented, although biotic and abiotic factors clearly affect a parasitoid’s ability to disperse in its habitat, and ultimately to locate and parasitize hosts. Thus, it is extremely important to understand the

28 dispersal capacity and fecundity of any parasitoid being assessed for potential use in a biological control program. Critical predictors of a parasitoid’s activity, and ultimately successful parasitism in the field, are its natural modes of dispersal (e.g. walking and flying). The basic biology of P. sulcata has been recently described by Roscoe (2014), but its dispersal capacity has not been studied nor has its fecundity been verified even though these parameters are important to evaluate its potential as a suitable candidate for augmentative biological control.

Here, we explore the flight capacity and walking activity of P. sulcata in relation to biological and environmental conditions in order to predict its potential for dispersal when used augmentatively against EAB. Specifically, we: (1) examine the relationship between parasitoid age and body size, temperature, and time of day on parasitoid flight capacity and walking activity;

(2) determine how flight, parasitoid age, and body size affect potential fecundity; and (3) investigate whether there is a trade-off between flight capacity and walking activity in this parasitoid species. We predict that at warm temperatures larger and younger wasps will have a greater dispersal capacity (i.e. fly and walk faster, farther, and more frequently) than smaller and older ones, and that potential fecundity will decrease with increasing wasp age and decreasing body size. Less clear is the relationship between parasitoid flight and walking capacities, which could be either positive, in that wasps walking more also fly more (i.e. active vs inactive parasitoids), or negative, in that wasps walking more fly less (i.e. a trade-off between the energy expended on flying and walking).

29 2.3 Materials and Methods

Wasp collection and rearing

EAB-infested ash trees were identified in three woodlots in southern Ontario: two from properties managed by the Ausable Bayfield Conservation Authority (43.34573, -81.5572 and 43.38295, -

81.54295) and another on a private ash plantation near Brooke Line, Alvinston, ON (42.84389, -

81.85438). These trees were felled, cut into ~50-cm lengths, brought to the Great Lakes Forestry

Centre (Sault Ste. Marie, Ontario, Canada), and put into rearing cabinets at temperatures ranging from 23 to 28 °C depending on cabinet height. Relative humidity (RH) was maintained at 45 % and the diel period at L:D = 16:8. Adult wasps were collected daily from the rearing cages and housed together, separated by sex, in ventilated, clear plastic cups (375 ml) at 24 °C,

60-70 % RH, and a diel period of L:D = 16:8 (the latter starting at 0700 and ending at 2300 h) until they were used in the experiments. Wasps were fed using a streak of pure honey on duct tape attached to the inside of each cup. Water was provided in each cup by saturated cotton inside a

12-ml vial.

Experimental procedures

Wasps were walked and flown 1 to 26 days following adult emergence. The body mass of each wasp was recorded as a proxy for body size taking pre- and post-walking and pre- and post-flight measurements to the nearest 0.1 mg using a digital analytical balance (Mettler Toledo AG285).

After weighing, wasps were gently moved into KIMAX (USA) test tubes (150 mm in length, 18- mm opening diameter) with 1 drop of a honey-water solution (i.e. 50 % honey, 50 % water), and the tubes laid out horizontally in the laboratory at ~24.5 °C. Walking activity was observed for

70 wasps, with an observation event occurring every 5 minutes over 2.5 h, from 0900 to 1130 h, recording whether the wasp was ‘walking’ or ‘resting’. After the walking period, wasps were chilled to slow their movement and the head of an insect pin (# 1) was glued (Quick Grip

30 Permanent Adhesive, Beacon Adhesives, Mt. Veron, NY) to the prothorax of each to allow it to be tethered to a flight mill. The flight mills were similar to those used by Jones et al. (2010) and

Wiman et al. (2014) [see Haavik et al. (2016) for more detail]. All wasps were then flown for 24 h at 21.0, 24.0, 24.5, or 25.0 °C under controlled conditions (i.e. L:D = 16:8 diel period and 50-

70 % RH). Female wasps were dissected after each flight period to count their total number of eggs.

Data processing and analyses

LabVIEW Full Development System software was used to record in-flight data and Scout 1.6.0.0

(Signal.X Technologies LLC, Commerce Township, MI) to export these data to Microsoft Excel.

The R software package, ‘flightmillR’ (developed by CJK MacQuarrie), was used to calculate summary statistics, including mean flight bout speed, total distance flown, and number of flight bouts taken. A successful flight bout was defined as ≥ 30 s of continuous flight; bouts of flight <

30 s were excluded from the analysis. Time of day was specified as either ‘light’ if wasps were flying during the photoperiod or ‘dark’ if wasps were flying during the scotoperiod. The proportion of time spent walking (i.e. the number of walking events over the total count of walking and resting events) was a binomial outcome (i.e. walking or resting), so we specified the number of ‘walking’ and ‘resting’ counts observed in a two-vector response variable. In all cases, female and male wasps were analyzed separately. All possible interactions were considered. In cases where interaction terms were not significant, they were removed to use the simplest model that, at minimum, tested main effects. All data were analyzed using the R statistical environment (R

Development Core Team 2018).

Flight capacity

The effects of wasp age, wasp body size, and temperature on mean flight bout speed, total distance flown, and the number of flight bouts taken were fit to linear models for multi-factor analysis.

31 Each model was assessed using graphical methods for homogeneity of variance and normality of the residuals. Models for total distance flown and number of flight bouts taken violated our assumptions, so both dependent variables were log-transformed to improve the model fit.

The effects of time of day and temperature on distance flown were also tested using a linear mixed-effects model (LMM), with a random effect on each wasp accounting for repeated measures on individuals in the photoperiod and scotoperiod. Because there was an unequal L:D ratio across the 24-h flight period, we analyzed the mean distance flown per hour in the photoperiod and scotoperiod instead of total distance flown. This LMM violated our assumptions as before, so mean distance flown per hour was log-transformed, which improved the model fit.

Pearson's χ2 test with Yates' continuity correction was used to determine whether an abrupt change from photoperiod to scotoperiod or scotoperiod to photoperiod stimulated flight in the wasps.

A paired t-test examined whether there was a difference in wasp body mass before and after the 24-h flight period. The effects of mean flight bout speed, total distance flown, and the number of flight bouts taken on body mass lost were tested using multiple regression. We observed that residuals had a non-normal distribution and heterogeneity in the variance of mass lost among individual wasps, so body mass lost was log-transformed to improve the model fit.

The same effects on post-flight survival were tested using logistic regression with a logit link function, which we assessed for overdispersion or underdispersion by examining the residual deviance, which was considered similar to the residual degrees of freedom.

Walking activity

A generalized linear model (GLM) with binomial errors and logit link function was fit to test the effects of wasp age and body size on the proportion of time spent walking. We diagnosed overdispersion using the same method as above and re-fit the model using quasibinomal errors.

This did not improve the model, so a generalized linear mixed model (GLMM) with binomial

32 errors and logit link function and observation-level random effect was fit to account for the overdispersion by modelling the excess variance. Because this is a relatively novel approach (see

Harrison 2015), we also tested the same by adjusting the covariance matrix and fit statistics using the R package ‘dispmod’ and compared the two models using a likelihood ratio test.

Relationship between flight capacity and walking activity

Simple linear models were used to test the effect of walking activity on both total distance flown and the number of flight bouts taken. Both models violated our assumptions, so total distance flown and the number of flight bouts taken were log-tranformed, which improved the models.

Fecundity

A linear model was used to test the effects of wasp age, body size, and whether a wasp flew on egg counts. This model met our assumptions of constancy of variance and normally-distributed residuals.

2.4 Results

Body size

Female P. sulcata body mass ranged from 6.2 to 22.2 mg, with a mean ± SE of 11.3 ± 0.3 mg (n

= 83), whereas male P. sulcata body mass ranged from 2.1 to 10.2 mg, with a mean ± SE of 6.8

± 0.4 mg (n = 23). Female wasps were significantly larger than males in terms of body mass

(Welch Two-Sample t-test: t = 10.13, df = 48.49, P < 0.001), with the largest female P. sulcata being 2.18 times heavier than the largest male.

33 Flight capacity

For female P. sulcata, successful flights were recorded for 30 of 83 individuals, with the total distance flown by females averaging 0.32 ± 0.15 km. The farthest flight by a female wasp was

4.05 km, with each taking on average 6 ± 1 bouts of flight. The maximum number of flight bouts taken by a female wasp was 24. Female wasps had a maximum flight speed of 1.18 ± 0.08 km/h.

Mean flight bout speed was not affected by the age of female wasps (F = 0.06; df = 1, 26;

P = 0.798) or temperature (F = 1.30; df = 1, 26; P = 0.265), however body size had a significant effect on mean speed for female wasps, with larger females observed flying significantly faster than smaller ones (F = 5.71; df = 1, 26; P = 0.024) (Fig. 2.1). Total distance flown was not affected by the age of female P. sulcata (F = 0.85; df = 1, 26; P = 0.365), or body size (F = 1.17; df = 1,

26; P = 0.290), however increasing temperature did significantly increase total distance flown (F

= 4.59; df = 1, 26; P = 0.042). As age of female P. sulcata increased, their number of bouts of flight taken significantly decreased (F = 5.75; df = 1, 26; P = 0.024) (Fig. 2.2), but the number of bouts of flight was not affected by female body size and temperature (F = 0.26; df = 1, 26; P =

0.613 and F = 3.04; df = 1, 26; P = 0.093, respectively).

There was no effect of time of day on the distance flown for female P. sulcata (F = 3.04; df = 1, 29; P = 0.092). We continued to observe a significant effect of temperature on distance flown by wasps, where increasing temperature increased the total distance flown (F = 5.99; df =

1, 28; P = 0.021), although this distance was not affected for females flying immediately after an abrupt shift in light intensity (between photoperiod and scotoperiod and vice versa) (χ2 = 1.70, df

= 1, P = 0.193).

Significant body mass was lost by female wasps after being tethered to the flight mill for

24 h (t = 6.42; df = 28; P < 0.001). Wasps lost significantly more mass as their mean flight speed increased (F = 11.34; df = 1, 26; P = 0.002) (Fig. 2.3), but no relationship was observed between

34 mass lost and total distance flown (F = 0.19; df = 1, 26; P = 0.665) or the number of flight bouts taken (F = 0.07; df = 1, 26; P = 0.796). Mean flight bout speed, total distance flown, and number of flight bouts taken were not significant predictors of post-flight survival (χ2 = 0.76, df = 1, P =

0.38 and χ2 = 0.09, df = 1, P = 0.763 and χ2 = 1.29, df = 1, P = 0.256, respectively).

Successful flights were only recorded for 5 of 23 males, with a total distance flown averaging 0.13 ± 0.07 km. The farthest distance flown by a male wasp was 0.39 km, with an average of 20 ± 11 flight bouts. The maximum number of flight bouts taken by a male was 63.

Male wasps had a maximum flight speed of 1.14 ± 0.25 km/h. No analysis was made of age, body mass, temperature, and time of day on males as successful flights were recorded from too few wasps (5).

Walking activity

Individual wasps displayed considerable range in walking activity. Female wasps were observed walking between 3 and 100 % of the observation events, and none were observed inactive for the entire assay. One 16-day old female weighing 8.0 mg walked 3 % of the observation events whereas a 15-day old female weighing 10.0 mg walked 100 % of the observation events. On average, female wasps walked 62 ± 3 % of the observation events.

Male wasp walking activity was similarly variable as female wasp walking activity. Male wasps spent between 3 and 80 % of the observation events walking, and none were observed inactive for the entire assay. One 16-day old male weighing 8.9 mg walked 3 % of the observation events while a 13-day old male weighing 8.0 mg walked 80 % of the observation events. On average, male wasps spent 29 ± 7 % of the time they were observed walking.

The time spent walking by female and male wasps was not affected by age (χ2 = 1.73, df

= 1, P = 0.188 and χ2 = 0.10, df = 1, P = 0.750, respectively) or body size (χ2 = 0.27, df = 1, P =

0.602 and χ2 = 0.21, df = 1, P = 0.643, respectively). Both female and male wasps lost significant

35 body mass after the 2.5-h walking period (Paired t-test: t = 2.64, df = 57, P = 0.011 and t = 5.41, df = 11, P < 0.001, respectively). On average, female body mass was reduced from 11.3 ± 0.3 mg to 10.7 ± 0.3 mg and male body mass was reduced from 7.0 ± 0.4 mg to 6.4 ± 0.4 mg.

Results of the model with the adjusted covariance matrix and fit statistics were similar to that of the GLMM (age: χ2 = 1.46, df = 1, P = 0.226 and body size: χ2 = 0.24, df = 1, P = 0.625).

Thus, it was expected that these two models would be similar (χ2 = 0.00, df = 1, P = 1.000).

Relationship between flight capacity and walking activity

There was no evidence of a significant trade-off between the walking activity and total distance flown (F = 0.41; df = 1, 22; P = 0.527) or number of flight bouts taken (F = 0.34; df = 1, 22; P =

0.567) by female wasps. This was not surprising considering the large variation in wasp walking activity coupled with the low flight capacity of these same wasps. For example, of the wasps that both walked and flew during the experiments, two walking 93 % of the observation events also flew 0.02 km over 6 flight bouts and 0.38 km over 3 flight bouts, while one wasp walking 3 % of the observation events then flew < 0.01 km over 1 flight bout.

Potential fecundity

The maximum number of eggs observed in a single wasp was 85 eggs at 0.5 days old while the minimum number was 29 eggs at 24 days old. On average, females had 60 ± 3 eggs upon dissection after flying. Whether a wasp flew did not affect its egg count (F = 2.62; df = 1, 19; P =

0.122). The relationship between P. sulcata body size and egg count was not significant (F = 2.29; df = 1, 19; P = 0.147), although egg count decreased significantly with wasp age (F = 6.22; df =

1, 19; P = 0.022) (Fig. 2.4).

36 2.5 Discussion

Dispersal capacity in Phasgonophora sulcata

When developing any biological control program, flight capacity and the factors that influence it are important to understand as these will help identify the most appropriate parasitoid(s) and protocol for their release. Our work is the first to explore the flight capacity of P. sulcata, a parasitoid native to North America being considered for augmentative release against introduced

EAB. Flight capacity has also been examined for the Asian parasitoid species, T. planipennisi, used in classical biological control of EAB. Fahrner et al. (2014) found that female T. planipennisi on flight mills in the lab flew 1.26 ± 0.17 km on average, ~3.9 times farther than we observed here for native female P. sulcata. This suggests that P. sulcata has limited dispersal capacity relative to other EAB biological control agents and will spread less rapidly than T. planipennisi when released against EAB. It also implies that P. sulcata will have a relatively localized impact on EAB since it cannot spread as far as other EAB parasitoids, such as T. planipennisi.

The dispersal capacity of a biological control agent is also important relative to its target host. If the target host is able to disperse farther than the biological control agent, it may be that the target host will “escape” parasitism (or predation) and its population will grow beyond a level of ecological or economic damage [e.g. Sericothrips staphylinus Haliday (Thysanoptera:

Thripidae) has a limited dispersal capacity and thus is not likely to spread and control large areas of gorse, Ulex europaeus L. (Fabales: Fabaceae), in Australia (Ireson et al. 2008)]. Taylor et al.

(2010) found that EAB flew on average 1.3 km per day (over four days) on a flight mill, with 10

% of beetles flying over 7 km per day. This is farther relative to P. sulcata, however experimental conditions do differ between these two studies. For example, room temperature was as high as 29

°C during beetle flight assays (Taylor et al. 2010), which is 4 °C higher than the warmest temperature during P. sulcata flight assays. The difference in temperature may account for

37 differences in the distances flown between P. sulcata and EAB. However, our results suggest that

P. sulcata will impact local EAB populations (i.e. those nearby release locations) following its release because of its limited dispersal capacity.

Parasitoid age and body size appear to have a variable impact on parasitoid activity, and this will also have implications for the release of P. sulcata. We found that younger female P. sulcata were more active in terms of flight than older females and that larger female wasps had a greater flight capacity compared to smaller ones as measured by flight bout speed. Fahrner et al.

(2014) examined the effect of wasp age and body size on flight capacity in T. planipennisi, observing no effect of age, however they did not measure the number of flight bouts taken.

Further, we found no effect of body size on the distance flown or the number of flight bouts taken by female wasps, which suggests that body size alone would not reduce the impact of P. sulcata in a biological control program against EAB as these parameters likely influence host location more than flight speed.

It is well established that ambient temperature affects the flight behaviour of many poikilothermic organisms, including insects (Taylor 1963), and this effect, along with parasitoid age and body size, should be considered when releasing P. sulcata in the field to select the optimal location for dispersal and parasitism of EAB. For example, for Trichogramma minutum Riley

(Hymenoptera: Trichogrammatidae), maximum flight propensity occurred at 25 and 30 °C, whereas wasps experienced a reduced flight capacity at lower temperatures (Forsse et al. 1992).

Similarly, we found that P. sulcata wasps flew significantly farther as the ambient temperature increased suggesting that releases will support parasitoid movement and be more effective if made in the summer, especially during periods of warm temperature. In contrast, time of day (as measured by ‘light’ or ‘dark’ period) did not influence wasp flight in our system, and this would mean that P. sulcata populations could be augmented and equally effective when released at any

38 point during daylight hours. The implications of parasitoid flight activity as measured here is difficult to interpret in all cases since tethered wasps have no landing cues on a flight mill, and thus, our results may vary somewhat from realized parasitoid activity in the field.

The fact that wasps lost significant body mass after being tethered to flight mills for 24 h and also when their mean flight bout speed increased implies that access to carbohydrates and water is important for parasitoid maintenance, especially for wasps with increased flight capacity.

This is true for T. planipennisi (Fahrner et al. 2014), C. glomerata (Wanner et al. 2006), and many other insects in the order Hymenoptera (Beenakkers et al. 1984). Carbohydrate sources, such as floral resources, can provide important nourishment for parasitoid maintenance and dispersal in the field (Wäckers 2005), and may be necessary for P. sulcata to achieve its maximum flight capacity.

After a parasitoid has located the habitat of its host, walking becomes the next component of its searching behaviour. We observed high variation in walking activity among P. sulcata as did Suverkrop et al. (2001) for the egg parasitoid Trichogramma brassicae Bezdenko

(Hymenoptera: Trichogrammatidae) at low temperature. The impact of this variation on the ability of parasitoids to establish and control an invasive insect pest population, such as EAB, is not clear.

If the time spent walking is positively correlated with the ability to locate hosts, such variation in dispersal could result in low rates of establishment and host location. Although we observed no differences in the proportion of time spent walking between wasps of varying age and body size, it is possible that younger and larger wasps can walk farther and/or faster than older and smaller ones. Here, the time spent walking was not affected by female wasp age or body size, suggesting that parasitism rates for younger wasps of all sizes would remain relatively consistent across the reproductive season; we did, however, observe reduced potential fecundity in older wasps, which might result in a lower realized fecundity in the field, especially later in the season.

39 Relationship between flight capacity and walking activity

The relationship between flight capacity and walking activity in the parasitic Hymenoptera is not well understood, possibly due to their small size and the difficulty in measuring these behaviours, especially under field conditions. Distinct polymorphic differences may explain varying dispersal capacities for some insects as in Melittobia spp. (Hymenoptera: Eulophidae), parasitoids attacking solitary bees and wasps. Brachypterous females within this genus develop quickly and remain within their natal patch to mate, lay eggs, and have offspring, while their macropterous counterparts are those that develop slowly, mate, and disperse from the natal patch to look for other hosts (Matthews et al. 2009). Further, host size can also impact the dispersal capacity of parasitoids, where individuals developing in particularly small hosts may emerge as smaller, and even wingless, adults than those developing in larger hosts (Salt 1941). We show that the frequency of walking over a 2.5-h walking period did not affect the total distance flown or flight activity (i.e. number of flight bouts taken) during a 24-h flight period. Although the walking assay was limited in time compared to the flight assay, we did observe a reduction in body mass after the 2.5-h walking period, which may indicate that this time interval is sufficient to show depletion of energetic resources in P. sulcata. For parasitoids such as P. sulcata, where no such morphological differences are apparent, it is less obvious there is a trade-off between flight and walking and more likely that host size and quality impacts their dispersal capacity.

Potential fecundity

Our work shows that P. sulcata is pro-ovigenic and emerges with a large complement of eggs as suggested by Roscoe et al. (2016). Similar to other pro-ovigenic parasitoids, eggs are developed from nutritional resources gained during larval feeding, but these resources must be split, in part, between the parasitoid’s fitness parameters (i.e. egg development, dispersal capacity, and longevity) (e.g. Innocent et al. 2010; Venkateswaran et al. 2017). Thus, it is expected that pro-

40 ovigenic parasitoids with increased egg loads at emergence will have decreased dispersal capacities and reduced longevity (Jervis et al. 2001). Consequently, pro-ovigenic parasitoids are recommended for environments with high host densities, where they can quickly locate and oviposit on or in hosts following emergence. In contrast to Roscoe (2014), we observed a trade- off between wasp age and potential fecundity in P. sulcata during the first 26 days of its life, and a decrease in flight capacity as wasps aged. Thus, we suspect that maximum EAB parasitism by

P. sulcata would occur soon after adult parasitoid emergence and that wasps consequently should be released before their emergence (i.e. possibly as pupae in EAB hosts) in order to maximize their egg-load complement and potential for parasitism.

Several metrics can be used to estimate parasitoid fecundity, with female body size often being a good predictor, however we observed no correlation between potential fecundity and body size in P. sulcata. Few studies have found no relationship between female body size and fecundity

(e.g. Boggs 1986; Johnson 1990). The fact that we did not see such a relationship may be partly explained by differences in parasitoid age at the time of dissection or possibly by some environmental constraint, such as differences in host quality, where less fecund parasitoids developed from poor-quality hosts and more fecund parasitoids developed from high-quality hosts irrespective of size. As such, it might be important to provision wasps with a food source at release locations in order to increase their potential fecundity and longevity.

Application to biological control

Flight capacity, walking activity, and fecundity are all important determinants of the ability of a parasitoid to locate and parasitize its host. Our results with P. sulcata may be useful to optimize a release protocol against EAB. Poor capacity to fly or walk limits the effectiveness of a parasitoid as a biological control agent. Thus, our findings that flight capacity (i.e. number of flight bouts taken) and fecundity decrease with wasp age suggest that this parasitoid should be released in the

41 field as soon as possible after emergence where mass-rearing in the laboratory is possible or as pupae if it will be released by transporting parasitoid-infested ash material. Further, the weak dispersal capacity of P. sulcata observed on flight mills, combined with the reduction in flight capacity and fecundity with increased age, suggests that augmentative releases should occur near

EAB-infested ash trees for optimal host location and parasitism. Given personal observations with this system in the laboratory and the fact that wasps walked but not all flew, it appears that P. sulcata uses walking and hopping more than flight for dispersal and host-finding, and this, combined with the fact that we saw no effect of wasp age or body size on P. sulcata walking activity, suggests that the most successful approach for implementation in an augmentative biological control program would be to release it in close proximity to the target host, EAB.

Figures

46 Chapter 3 Vegetation and Habitat Characteristics Affect the Abundance and Impact of North American Parasitoids Attacking the Emerald Ash Borer

3.1 Abstract

The structural and compositional variability of forest vegetation influences the probability of establishment of invasive forest insects as well as the abundance and diversity of parasitoids and ultimately parasitism of those invasive insects. Using the Vegetation Sampling Protocol (VSP), I identified various forest stand characteristics and explored their relationship with the abundance and impact of two important native North American parasitoid groups, Phasgonophora sulcata

Westwood (Hymenoptera: Chalcididae) and Atanycolus spp. Foerster (Hymenoptera:

Braconidae), attacking the emerald ash borer (EAB), Agrilus planipennis Fairmaire (Coleoptera:

Buprestidae). I found that tree biomass, tree condition, canopy closure, floral resource availability, and the abundance of a novel host, EAB, all were important predictors of parasitoid abundance in southern Ontario. Overall parasitism of EAB by P. sulcata and Atanycolus spp. exhibited similar relationships with tree biomass, tree condition, and floral resource availability. These results can be used to identify forest stands that are more likely to have large native parasitoid populations, be more efficient in supressing EAB populations, and provide insight into where future biological control efforts against EAB should be focused.

47 3.2 Introduction

The emerald ash borer (EAB), Agrilus planipennis Fairmaire (Coleoptera: Buprestidae), is an invasive forest insect accidentally introduced into North America in the 1990s (Siegert et al.

2014). Since its discovery in 2002 in the Detroit-Windsor area (Haack et al. 2002), this saproxylic beetle has been rapidly spreading and attacking all ash tree species, Fraxinus L. (Lamiales:

Oleaceae), it has encountered to date (Czerwinski et al. 2007; Anulewicz et al. 2008). More recently, it has been observed completing development on white fringetree, Chionanthus virginicus L. (Lamiales: Oleaceae) (Cipollini 2015), and on cultivated olive, Olea europaea L.

(Lamiales: Oleaceae) (Cipollini 2017) in North America. This puts many forests, especially municipal urban forests with large numbers of ash trees, across North America at high risk of invasion and damage by EAB.

The variability among these forest communities, such as changes in species composition, can influence the probability of establishment of invasive forest insects (e.g. Mangels et al. 2015).

For example, diverse forests are thought to provide more host plants for an invasive insect to attack and promote establishment than simple monoculture plantations (as hypothesized by

Niemelä and Mattson 1996). However, this does not appear true for EAB as studies to date suggest that species composition and other stand-level forest characteristics have no effect on EAB populations (e.g. Smith et al. 2015). As such, it is possible that EAB can invade a wide range of forest types leaving them and all North American ash trees susceptible to infestation.

In contrast with Niemelä and Mattson (1996), forest stand diversity along with other structural characteristics are thought to support greater natural enemy diversity that can attack an invasive insect and thus prevent its establishment [as proposed by Jactel et al. (2002) for European stem borer, Dioryctria sylvestrella Ratz. (Lepidoptera: Pyralidae)]. Parasitoid populations have been extensively studied at the ecosystem level, and ecosystem characteristics that can influence

48 parasitoid abundance and parasitism include plant species composition, tree health, canopy closure, and floral subsidies. Harrington and Barbosa (1978) found that oviposition by

Parasetigena silvestris (R-D) (Diptera: Tachinidae) was influenced by the tree on which gypsy moth, Lymantria dispar (L.) (Lepidoptera: Lymantriidae), was located while Weseloh (1972) observed that oviposition by Brachymeria intermedia (Nees) (Hymenoptera: Chalcididae) occurred mainly in open and sunny locations. These canopy gaps may alter forest microclimate by increasing temperature and lead to increases in parasitoid abundance and activity in open areas

(Hébert et al. 1990). Similarly, the flight activity of native EAB parasitoids increased with temperature (e.g. P. sulcata, Gaudon et al. 2018), suggesting that increased host searching will occur in areas with low canopy closure, although other parasitoids have been found to preferentially attack their hosts in shady areas (e.g. Apantales spp. attacking gypsy moth, Weseloh

1979). Flaherty et al. (2011) observed that parasitism of Tetropium fuscum (F.) (Coleoptera:

Cerambycidae) was lower on healthy trees than on cut and girdled (i.e. stressed) trees suggesting that parasitoids of this wood-boring beetle will exploit volatiles emitted by the stressed host tree to use to locate its host. Further complicating this relationship, Wäckers (2004) found that flowers varied in their olfactory attractiveness and nectar accessibility to parasitoids, with some flowers attracting, not attracting, and others even repelling parasitoids with or without providing an accessible food source. Thus, tree composition, tree condition, canopy gaps/closure, and floral resources may affect parasitoid populations and could partially explain why EAB parasitism by native parasitoids has been observed to vary greatly (i.e. < 1 % to 71 %) between different forest sites (Liu et al. 2003; Cappaert and McCullough 2009; Duan et al. 2009, 2010; Lyons 2010;

Roscoe et al. 2016). To date, many studies explore the effects of vegetation and habitat characteristics on parasitoid diversity and parasitism, but fewer studies have investigated these relationships between parasitoid abundance and, more importantly, no such information is available in forest ecosystems for native North American parasitoids attacking EAB.

49 Two important groups of native parasitoids in terms of their abundance and high parasitism rates of EAB have been identified, namely Phasgonophora sulcata Westwood

(Hymenoptera: Chalcididae) and Atanycolus spp. Foerster (Hymenoptera: Braconidae). These parasitoids may switch from native Agrilus hosts, specifically bronze birch borer (Agrilus anxius

Gory) on birch trees [Betula spp. L. (Fagales: Betulaceae)] (Barter 1957), bronze poplar borer

(Agrilus liragus Barter and Brown) on poplar trees [Populus spp. L. (Malpighiales: Salicaceae)]

(Barter 1965), and two-lined chestnut borer (Agrilus bilineatus Weber) on oak trees [Quercus spp.

L. (Fagales: Fagaceae)] (Haack et al. 1981), when high populations of non-native EAB co-occur as seen in other systems. For example, the native parasitoid Lathrolestes thomsoni (= luteolator)

Reshchikov (Hymenoptera: Ichneumonidae) was highly abundant on the non-native birch leafmining sawfly Profenusa thomsoni (Konow) (Hymenoptera: Tenthredinidae) (Digweed

1998). Such a host switch by native parasitoids to an introduced, novel host such as EAB would be worth exploring for potential application in conservation biological control. Regions recently invaded by EAB typically show very low parasitism of EAB by native North American parasitoids

(e.g. Liu et al. 2003; Duan et al. 2010, 2012), however, other work suggests populations of P. sulcata and A. cappaerti will increase and these species can parasitize up to 40% and 71% of EAB larvae, respectively, at some EAB-infested sites (Cappaert and McCullough 2009; Lyons 2010).

Parasitoids, especially hymenopteran parasitoids, are rarely considered in conservation planning likely because they are a diverse taxanomic group that is time-consuming to survey and difficult to identify (Shaw and Hochberg 2001). However, parasitoids are important in terms of their potential to provide resistance to insect disturbances, including the introduction of non-native and invasive insects, in forest communities. Thus, linking hymenopteran parasitoids with vegetation and habitat characteristics would enable predicting their abundance and impact in a given area. Measuring and understanding this relationship would also serve as a tool that

50 conservation authorities, municipalities, and other land planners can use to maintain or enhance parasitoid populations to establish such resilient forests.

The objective of this chapter is to investigate the influence of stand-level forest characteristics on the abundance of native Agrilus parasitoids and EAB parasitism in forest stands, specifically: (1) the amount of ash and alternative host tree biomass, (2) tree health condition, (3) canopy closure, (4) floral resource availability, and (5) the availability of a novel host, EAB. I tested the hypotheses that: (1) stands dominated by host trees for native Agrilus spp. (i.e. birch, poplar, and oak trees) will have the largest populations of native Agrilus parasitoids and higher rates of EAB parasitism, (2) native parasitoids will be more abundant on ash trees in poor health and parasitism in these stands will be higher than stands with healthy ash, (3) parasitoids will be more abundant in stands with some gaps in the canopy leading to an increase in EAB parasitism rates, (4) the availability of floral resources within stands will promote parasitoid populations and

EAB parasitism, and (5) an increase in available hosts, such as EAB, will increase parasitoid abundance and EAB parasitism.

3.3 Materials and Methods

Experimental design

The current study was conducted in southern Ontario between 24 April and 1 September 2017. I identified 30 forest stands containing varying amounts (i.e. % biomass) of ash, birch, poplar, and oak trees across southern Ontario in the City of Kitchener, Town of Oakville, Royal Botanical

Gardens (City of Hamilton), Scanlon Creek Conservation Area (Town of Bradford West

Gwillimbury), and Koffler Scientific Reserve at Jokers Hill (Township of King). Stands were kept within a 150-km distance to reduce variation in site factors such as climate. In each stand, I set up

51 a 400-m2 circular plot using the Vegetation Sampling Protocol (VSP), a rigorous yet efficient approach to characterize forest plots (Puric-Mladenovic and Kenney 2016), to measure specific site conditions that may affect the abundance and impact of P. sulcata and Atanycolus spp. attacking EAB.

First, the center of each plot was identified, and then I measured 11.28 m in the four main cardinal directions (i.e. north, east, south, and west) to identify the plot boundaries (Fig. 3.1).

Within each plot, I identified and measured the diameter at breast height (DBH) for all trees ≥ 5 cm DBH at 1.3 m up from the base of the tree, whether alive or dead (Appendix A). For trees where a fork occurred below 1.3 m DBH, each stem was measured separately. The DBH of each ash, birch, poplar, and oak tree was used to calculate percentage aboveground biomass using the formula by Lambert et al. (2005). The total biomass of birch, poplar, and oak trees were added together to investigate the overall effect of percentage alternative host tree biomass on the abundance and impact of P. sulcata and Atanycolus spp. populations.

To classify ash tree condition, I rated ash canopy decline for ash trees within the plots from 1 (healthy) to 5 (dead), with ratings 2, 3, and 4 being successive degrees of decline (Smith

2006). Canopy decline 1 represented an apparently healthy ash tree with no detectable canopy decline; 2 showed slight canopy decline although all branches exposed to sunlight had leaves; 3 displayed < 50 % canopy decline; 4 exhibited > 50 % canopy decline; and 5 had no visible leaves in its canopy but may or may not have had epicormic shoots on the lower bole of the tree. Canopy decline measurements were taken between 18 and 20 July 2017. I averaged the rating for canopy decline of ash trees in each plot to determine the overall condition by plot.

Using VSP within the plots, I also set up five 1-m2 subplots starting with one in the plot center and then the other four subplots 5.64 m in all cardinal directions to measure canopy closure and floral resource availability (Fig. 3.1). I used visual assessments to estimate the amount (i.e. 0

52 to 100 %) of canopy closure for all vegetation above two meters. I recorded stem counts of flowering herbaceous ground vegetation independent of whether they were flowering at the time of the survey (Appendix B). Trees and saplings (as part of the ground vegetation layer) and non- angiosperms were not counted as they generally do not flower and thus do not provide floral resources as part of the ground layer. Graminoids were also excluded. Ground vegetation cover inventories were conducted during a short sampling window between 9 and 17 August 2017 to reduce variation (i.e. as ground cover will generally increase through the season). The number of stems counted became the proxy for floral resource availability.

Sticky band traps, which have been shown to effectively capture both EAB and its parasitoids (e.g. Lyons 2010; Roscoe et al. 2016), were wrapped around the trunks of 5 ash trees in each of the plots. If a plot contained less than 5 ash trees, sticky band traps were wrapped around the most dominant broadleaf tree. Sticky band traps were made from ~51-cm wide plastic wrap (Item # 498385, Staples Canada Inc.) coated with Pestick™ insect glue (Catalogue # 01-

3522-2, Hummert International). Traps (n = 150) were sampled every two weeks from May to

September 2017. All P. sulcata and Atanycolus spp. as well as EAB were removed from the traps using forceps and put into vials containing Histo-Clear II, a nontoxic and biodegradable solvent to dissolve any glue on the insects collected. Vials were taken back to the laboratory so that the identification of all insects could be confirmed, and insects pinned or stored in vials containing

70 % ethanol to create a reference collection. EAB and its parasitoids were identified using guides

(Paiero et al. 2012; Roscoe 2014). EAB parasitism was calculated by dividing the total number of wasps by the sum of the total number of EAB and wasps captured on sticky band traps (Lyons

2010; Roscoe et al. 2016).

53 Data analyses

All data were analyzed using the R software (R Development Core Team 2018). Interaction terms were excluded from all models because they led to problems with model convergence. In all models, I tested for non-linearity by comparing linear models with quadratic models using the anova function. Specifically, ash tree condition was better explained as a quadratic term than as a linear term. The effects of percentage ash biomass, percentage other alternative host tree biomass, ash condition, canopy closure, amount of floral resources, and number of EAB captured within plots on the number of P. sulcata captured were fit to a generalized linear model (GLM) with Poisson errors and log link function.

The same effects on the number of Atanycolus spp. were also fit to a Poisson regression with log link function, but I identified overdispersion. All efforts to account for the overdispersion still led to poorly-fitted models, so I opted to analyze the presence/absence of Atanycolus spp. using a logistic regression with quasibinomial errors and logit link function.

The effects of percentage ash biomass, percentage alternative host tree biomass, ash condition, canopy closure, floral resource availability, and number of EAB captured within plots on EAB parasitism were also analyzed. Because analysis of the parasitism rate by each hymenopteran group led to problems with model convergence, I tested the above effects on overall

EAB parasitism (i.e. EAB parasitism by all P. sulcata and Atanycolus spp.) within the plots first by using a logistic regression with binomial errors and logit link function. I again identified overdispersion that could not be corrected using quasibinomial errors (and logit link function), so

I refit the model to account for the excess variation using the R package, ‘dispmod’, to adjust the covariance matrix and overall fit statistics. Statistical significance was set at P < 0.05 and all errors reported are ± 1 SE from the mean.

54 3.4 Results

Across the 30 plots, percentage ash biomass ranged from 1.9 to 100.0 % with a mean of 31.2 ±

4.5 % ash biomass, while percentage alternative host tree biomass ranged from 0 to 87.2 % with a mean of 23.7 ± 5.1 % biomass, mostly comprised of paper birch, Betula papyrifera Marshall, and trembling aspen, Populus tremuloides Michx. Mean ash condition ranged from 1 (healthy) to

5 (dead), with the average tree exhibiting slightly > 50 % canopy decline and a mean canopy decline rating of 3.3 ± 0.2. Canopy closure ranged from 16.3 to 84.0 % closure with a mean of

61.0 ± 3.4 % closure. The average number of stems of flowering herbaceous ground vegetation in each subplot ranged from 0 to 89 with a mean of 21 ± 4 stems in each plot. In some plots, non- native species greatly contributed to floral resource availability [e.g. garlic mustard, Alliaria petiolata (M. Bieb) Cavara and Grande (Brassicales: Brassicaceae) in VSPPLOTid 6 (see

Appendix B)], however native species, such as white trillium [Trillium grandiflorum (Michx.)

Salisb. (Liliales: Melanthiaceae], Canada mayflower [Maianthemum canadense Desf.

(Asparagales: Asparagaceae], and mayapple [Podophyllum peltatum L. (Ranunculales:

Berberidaceae], were exclusively observed in other plots (e.g. VSPPLOTid 1). The total number of EAB captured ranged from 0 to 82 beetles with a mean of 8 ± 4 EAB captured across all plots.

Fewer native EAB parasitoids were captured than EAB: the number of P. sulcata captured ranged from 0 to 23 wasps per plot with a mean of 2 ± 1 wasps across the plots, and even fewer Atanycolus spp. were captured with a minimum of 0 individuals, a maximum of 13 individuals, and a mean of 1 ± 0 wasps across the plots. I observed large variation in EAB parasitism by P. sulcata, ranging from 0.0 to 53.6 % with a mean of 9.2 ± 4.0 % parasitism and for Atanycolus spp. from 0.0 to

100.0 % with a mean of 11.6 ± 5.3 % parasitism. Overall EAB parasitism (i.e. by both parasitoid groups) across the plots ranged from 0.0 to 100.0 % with a mean of 16.6 ± 5.7 % parasitism per

55 plot (Fig. 3.2). Despite some high parasitism rates observed during our study, it should be noted that some stands had very low capture rates of P. sulcata, Atanycolus spp., and EAB.

The abundance of P. sulcata within plots increased significantly with the percentage of

2 2 ash biomass (χ = 22.09; df = 1; P < 0.001), mean number of flowering plants (χ = 74.57; df = 1;

2 P < 0.001), and the total number of EAB captured within plots (χ = 16.74; df = 1; P < 0.001). As the percentage of alternative host tree biomass increased, the abundance of P. sulcata surprisingly

2 decreased (χ = 17.30; df = 1; P < 0.001). There was a curvilinear relationship between the number of P. sulcata captured and mean ash condition, where an increase of P. sulcata occurred initially as ash condition started to decrease up until ash condition reached ~3, where P. sulcata abundance

2 then decreased (χ = 6.29; df = 1; P = 0.043) (Fig. 3.3). There was no relationship between P.

2 sulcata abundance and mean canopy closure within the plots (χ = 0.02; df = 1; P = 0.901).

The probability of Atanycolus spp. occurring within a plot increased significantly with the

2 mean abundance of flowering ground vegetation (χ = 5.69; df = 1; P = 0.017). Different from that for P. sulcata, the probability of Atanycolus spp. occurring with in a plot increased with

2 percentage alternative host tree biomass (χ = 15.52; df = 1; P < 0.001). The probability of

2 Atanycolus spp. occurring within a plot increased significantly as canopy closure decreased (χ =

6.19; df = 1; P = 0.013) (Fig. 3.4), unlike with P. sulcata. The probability of Atanycolus spp.

2 occurring within a plot was not affected by mean ash condition (χ = 5.78; df = 1; P = 0.056),

2 percentage ash biomass (χ = 0.18; df = 1; P = 0.673), nor the total number of EAB captured

2 within the plots (χ = 0.17; df = 1; P = 0.676).

EAB parasitism across the plots also exhibited a curvilinear relationship with mean ash condition, where an increase in parasitism occurred initially as ash condition started to decrease up until ash condition ~4 where the proportion of parasitized EAB began to slightly decrease as

2 trees exhibited between 50 % and complete canopy decline (χ = 7.01; df = 1; P = 0.030). EAB

2 parasitism was also affected by the mean abundance of flowering ground vegetation (χ = 4.02; df

2 = 1; P = 0.045) and percentage alternative host tree biomass (χ = 13.92; df = 1; P < 0.001) (Fig.

2 3.5). EAB parasitism was not affected by percentage ash biomass (χ = 2.20; df = 1; P = 0.138),

2 2 mean canopy closure (χ = 3.68; df = 1; P = 0.055), or total EAB captured within a plot (χ = 0.21; df = 1; P =0.650).

3.5 Discussion

I observed that stand vegetation and habitat characteristics clearly had an effect on the abundance of native Agrilus parasitoid populations, and this varied between the parasitoid groups, namely P. sulcata and Atanycolus spp. The stand characteristics also impacted rates of EAB parasitism by these parasitoids. These results help explain the variability in distribution of native Agrilus parasitoids and their parasitism of EAB by way of differences in tree biomass, tree condition, canopy closure, floral resource availability, and host availability across forested sites. This new knowledge of where large populations of these native parasitoids exist help strategize where to focus future biological control efforts against EAB or other, possibly newly-introduced, Agrilus pests, and provides insight on how one might begin to design or restore forests to foster natural enemies to slow the spread of EAB through conservation biological control.

Parasitoid populations may vary in size on different host plant species, and, as such, insect herbivores may experience differential parasitism. For example, oviposition by P. silvestris was more prevalent on red oak, Quercus rubra L., than white pine, Pinus strobus L. (Pinales:

Pinaceae), perhaps because tree species composition within a region influences parasitism of gypsy moth (Harrington and Barbosa 1978). On a larger scale, Sperber et al. (2004) showed an increase in parasitoid diversity with tree species richness during the warm seasons in Brazilian cacao agroforestry systems, which may result in more natural enemies and increased parasitism

57 of cacao pests than in areas with fewer tree species. Thus, generalizing patterns in distribution between parasitoid species may be difficult as also seen here. In my study, P. sulcata abundance increased as the percentage of ash biomass increased, yet no effect was seen for Atanycolus spp.

In contrast, as the percentage biomass of alternative host trees increased, P. sulcata abundance decreased while that of Atanycolus spp. increased. This shows that parasitoids, such as P. sulcata, have plasticity to parasitize a novel host and suggests that there may have been rapid and intense selection on their population(s) to switch host populations and shows plasticity in this species.

Both P. sulcata and Atanycolus spp. have been commonly found parasitizing EAB on ash trees, but Atanycolus spp. have also been commonly found parasitizing Agrilus beetles on alternative host tree species, especially birch; it has been less common to observe P. sulcata attacking native

Agrilus beetles (Lyons 2010). The highest EAB parasitism rates occurred in stands with the highest percentage alternative host tree biomass indicating the importance of tree diversity in these forest ecosystems. These opposing patterns may mean parasitism of EAB will naturally occur across a range of vegetation and habitat characteristics and perhaps be strongly influenced by one

(or few) parasitoid group(s) in each forest stand type.

In response to insect herbivory, stressed trees have been shown to release larger quantities of volatile organic compounds (VOCs) than under normal conditions (Paré and Tumlinson 1999).

Parasitoids use these chemical cues (i.e. allelochemicals such as synomones and kairomones) to locate their host’s habitat and host insects (Godfray 1994). A laboratory study by Pettersson

(2001) found that the bark beetle (subfamily Scolytinae) parasitoid, Rhopalicus tutela (Walker)

(Hymenoptera: Pteromalidae), was attracted to chemicals associated with damaged host trees.

Rojas et al. (2006) observed a parasitoid of the coffee berry borer, Phymastichus coffea LaSalle

(Hymenoptera: Eulophidae), to be attracted to both mechanically-damaged and infested coffee beans also in the laboratory. Further, a field study by Flaherty et al. (2011) showed that parasitism

58 of T. fuscum was high on cut and girdled (i.e. stressed) trees. Not surprisingly, I observed a similar relationship between ash condition and native Agrilus parasitoids, where both P. sulcata abundance and overall EAB parasitism by both P. sulcata and Atanycolus spp. increased as the overall condition of ash in the stands declined up until a certain threshold; as complete tree mortality neared, their abundance and rates of parasitism decreased. Although I observed no relationship between ash condition and Atanycolus spp. abundance, these parasitoids may be following similar patterns with the health/condition of alternative host trees, especially on birch where these parasitoids are also commonly found (Lyons 2010).

Large gaps in the forest canopy can influence parasitoid abundance and ultimately their parasitism. As canopy gaps occur, so do changes in abiotic conditions, such as temperature, and this altered microclimate may increase parasitoid populations and help explain variation in parasitism, such as with Winthemia fumiferanae Tothill (Diptera: Tachinidae), a parasitoid of spruce budworm [Choristoneura fumiferana (Clemens) (Lepidoptera: Tortricidae)] (Hébert et al.

1990). Similarly, Weseloh (1972) found that oviposition by B. intermedia was typical in open, sunny areas rather than shady locations. I observed that the probability of capturing Atanycolus spp. in stands with less canopy closure increased. This may be due to microclimate preferences, where these parasitoids prefer open and sunny areas that are warmer than shady locations.

Although there was no effect of canopy closure on P. sulcata populations, it should be noted that canopy gaps will occur as the number of dead and dying ash in a stand increase, and this was found to be associated with large populations of P. sulcata. Many canopy gaps may increase the distance a parasitoid must travel to reach the nearest host patch and hinder parasitism.

Interestingly, I observed similar population sizes of P. sulcata and consistent parasitism across a range of canopy closures. However, since P. sulcata is a weak disperser (Gaudon et al. 2018), these wasps may not be able to disperse to locate new host habitat or host larvae in the presence

59 of very large canopy gaps. Thus, sites with intermittent gaps in the canopy might yield the greatest abundance of both P. sulcata and Atanycolus spp. and highest rates of EAB parasitism.

Floral subsidies can provide important nourishment by way of sugars for parasitoid maintenance and dispersal making it difficult to assess their true impact on parasitoid populations in the field (Wäckers 2005). I found that the abundance of P. sulcata and Atanycolus spp. along with EAB parasitism also increased with increasing numbers of flowering plants within the plots and this suggests that adult wasps could potentially use flowers as a food source. In addition to increases in parasitoid abundance, parasitism has been shown to increase with an increase in flowers. Parasitism of Lygus lineolaris by Leiophron pallipes Curtis [= Peristenus pseudopallipes

(Loan)] (Hymenoptera: Braconidae) was found to be higher (i.e. 30-40 %) on Erigeron spp. than other weedy species because L. pallipes was attracted to flowers on these plants (Shahjahan 1974).

Similarly, high parasitism rates have been observed on EAB by P. sulcata and Atanycolus spp. in my work possibly due to the availablability of floral resources, and it may be that flowers are attractive to these native Agrilus parasitoids and provide important nourishment for adult wasps.

However, flowering herbaceous ground vegetation cover may increase as canopy gaps occur as a result of dead and dying ash, both which I found to be mechanisms driving the high abundance of native EAB parasitoids and parasitism of EAB. Including such an interaction term in the models used in my analysis led to problems with model convergence, but such an interaction should be considered as similar understory responses immediately following disturbance are known to occur

(e.g. Kovacic et al. 1985; Pec et al. 2015).

The attack rate or rate of parasitism of any given parasitoid species depends on how its population is distributed relative to its host(s). I observed that the abundance of P. sulcata, but not

Atanycolus spp., increased with the abundance of its EAB host in forest stands, while overall EAB parasitism rates remained relatively constant. Similarly, Waage (1983) observed that abundance

60 of Diadegma eucerophaga Horstmann (Hymenoptera: Ichneumonidae) increased with host density although parasitism of hosts remained similar at different host densities. Further, the time spent in a patch by Venturia (= Nemeritis) canescens (Grav.) (Hymenoptera: Ichneumonidae) and its host encounter rate both increased with the density of its host (Waage 1979). Given that Roscoe

(2014) found positive responses by P. sulcata to EAB volatiles [e.g. 3-(Z)-lactone], it is not surprising that an increase in EAB abundance led to an increase P. sulcata abundance in forest stands. Again, this shows plasticity in P. sulcata or suggests that there may have been rapid and intense selection on this species to switch hosts, and perhaps indicates the importance of at least one alternative native host for Atanycolus spp.

Implications for parasitoid conservation and EAB management

Parasitoid wasps are of economic importance because of their ability to regulate insect pests. Here,

I show that diverse forest types are important to support native Agrilus parasitoid populations in terms of their high abundance and parasitism rates on their EAB host. Differences in these stands appear to provide parasitoids with diversified host habitats and host resources, carbohydrates for adults, and suitable microclimates for development and parasitism to occur. It is difficult to make silvicultural recommendations to foster native EAB parasitoid populations. For example, I found no consistent positive pattern with respect to increased biomass of host trees (i.e. ash, birch, poplar, and oak) and large populations of P. sulcata or Atanycolus spp. However, I did find high levels of EAB parasitism in areas with high birch, poplar, and oak biomass and flowering ground vegetation suggesting that companion planting with alternative host tree species and flowering ground vegetation may create forests in southern Ontario that are better able to resist EAB invasion. This approach to combating EAB through conservation biological control operates under the assumption that low levels of native Agrilus populations will support native parasitoid

61 populations on trees that can “spill over” onto EAB populations, although it remains uncertain as to whether this will ultimate reduce EAB populations overall.

Substantial ash regeneration has been observed in EAB-infested sites, but the probability of these ash trees surviving to a seed-bearing age seems low (Aubin et al. 2015). However, it may be possible to delay mortality of ash saplings through site manipulation (i.e. conservation biological control) that fosters native EAB parasitoids as discussed above. Native EAB parasitoids have been found to significantly decrease EAB population growth rates (e.g. Duan et al. 2015).

Thus, such conservation efforts should in turn aim to increase parasitism of EAB. Although many studies test the effects of habitat characteristics on parasitoid communities, few have replicated the results by designing habitats that foster parasitoids. This would be a unique opportunity to test the accuracy of the habitat predictors of EAB parasitism observed in this study, and such projects are currently underway.

62 Figures

Figure 3.1 The sampling design used in the Vegetation Sampling Protocol (VSP) includes one main plot and five sub-plots. Adapted from Puric-Mladenovic and Kenney (2016).

0 % parasitism 100 % parasitism

67 Chapter 4 Augmenting Populations of North American Parasitoids for Biological Control of the Emerald Ash Borer

4.1 Abstract

Since its introduction into North America, Agrilus planipennis Fairmaire (Coleoptera:

Buprestidae) (emerald ash borer or EAB) has been spreading rapidly, killing millions of ash trees,

Fraxinus spp. L. (Lamiales: Oleaceae). Eradication is not viable, so biological control is now a leading management strategy. I moved parasitoid-infested ash material to EAB-infested sites to augment EAB parasitism by native North American parasitoids. Changes in EAB parasitism, EAB density, and woodpecker predation were monitored over three years following transport of logs and the release of parasitoids. Significantly higher EAB parasitism was observed after three years in plots where native parasitoids were released, however no changes were seen in EAB density between the treatment and control plots or over time. Predation by woodpeckers was also examined in the plots and showed no relationship with EAB density, year, and, more importantly, woodpecker predation did not differ significantly between parasitoid-release and control plots.

The release of native parasitoids through transport of infested ash logs may be used as one component to increase natural mortality of EAB and in its long-term management.

4.2 Introduction

The emerald ash borer (EAB), Agrilus planipennis Fairmaire (Coleoptera: Buprestidae), is an invasive jewel beetle introduced into North America from Asia in the 1990s (Siegert et al. 2014) and discovered in North America in 2002 (Haack et al. 2002). Since its discovery, it has shown a continue pattern of spread in all directions from the epicentre of the intial invasion in the Detroit-

Windsor region and is now present in five Canadian provinces and 35 US states. EAB is considered one of the most damaging and costly invasive forest insects to invade North America to date because of its ability to infest and kill healthy North American ash trees, Fraxinus spp. L.

(Lamiales: Oleaceae), with up to 99 % tree mortality observed shortly after a region is infested by this saproxylic beetle (Knight et al. 2013).

Early management efforts against EAB involved strategies to isolate known populations and eradicate satellite populations (Cappaert et al. 2005). In this way, areas were essentially put under quarantine to limit the spread of EAB-infested ash material to new, uninfested areas

(Cappaert et al. 2005). Eradication efforts eventually ceased in both Canada and the USA because of the difficulties associated with its rapid spread and cryptic nature as a wood-boring beetle as well as an overall lack of funding for this approach to EAB management (Liu et al. 2003; Herms and McCullough 2014). Thus, slowing the spread of EAB across North America, rather than eradication, has become the most viable management option for reducing its impact. Slow-the- spread strategies involve early detection followed by efforts to slow the pest’s spread and population growth [e.g. gypsy moth, Lymantria dispar L. (Lepidoptera: Erebidae), Sharov et al.

(2002)]. To combat EAB, both short- and long-term management tools are being used in a slow- the-spread strategy, including quarantines of EAB-infested material, chemical treatment, tree removal, and girdled “trap trees” as short-term and biological control as long-term management approaches (Herms and McCullough 2014).

69 Although there are many abiotic and biotic mortality factors affecting EAB, including natural host tree resistance (Anulewicz et al. 2007; Tannis and McCullough 2012), extreme cold

(Crosthwaite et al. 2011), cold-warm-cold fluctuations in temperature (Sobek-Swant et al. 2012), predators (Rutledge et al. 2013; Jennings et al. 2015), parasitoids (Duan et al. 2009, 2012), and pathogens (Bauer et al. 2004; Kyei-Poku and Johny 2013), none of these factors have been able to provide sufficient mortality to naturally suppress EAB populations below an ecological- or economically-damaging threshold in North America. Thus, human intervention through classical biological control has been proposed and is currently underway.

Classical biological control of EAB has involved the introduction of a number of important non-native natural enemies, including Tetrastichus planipennisi Yang (Hymenoptera:

Eulophidae), Oobius agrili Zhang and Huang (Hymenoptera: Encyrtidae), Spathius agrili Yang

(Hymenoptera: Braconidae), and Spathius galinae Belokobylskij and Strazanac (Hymenoptera:

Braconidae), from Asia, where EAB is native, into its introduced range. Recent studies show that introductions of some of these parasitoids, especially T. planipennisi on ash saplings (e.g. Duan et al. 2017), can cause significant EAB mortality. However, these co-evolved parasitoids alone have not yet been shown to suppress EAB populations below an ecologically- and economically- damaging threshold long term in North America. Although Asian parasitoids hold promise to do so, it would be worthwhile to explore other options for managing the spread of EAB throughout its introduced range.

A cost-effective and environmentally-sound approach to combat invasive species such as

EAB may be through augmenting native natural enemy populations where they are absent or present only in low numbers. In Canada, 32 native natural enemy agents have been used for augmentative biological control (MacQuarrie et al. 2016). Of these, studies that relocate native

70 natural enemies to control a non-native pest show promise for this EAB system (e.g. MacQuarrie et al. 2013).

Records show that there are several North American Agrilus Curtis (Coleoptera:

Buprestidae) species, all which have a complex of native natural enemies that could possibly switch from native hosts when high populations of a closely-related species, such as EAB, co- occur. Woodpeckers (Aves: Picidae) attacking EAB commonly include hairy [Picoides villosus

(Linnaeus)], downy [Picoides pubescens (Linnaeus)], and red-bellied [Melanerpes carolinus

(Linnaeus)] woodpeckers (Lindell et al. 2008) and are important mortality factors affecting EAB populations in North America (e.g. Cappaert et al. 2005). Woodpeckers have been found to increase in abundance as a response to outbreaking saproxylic beetles, including mountain pine beetle, Dendroctonus ponderosae Hopkins (Coleoptera: Curculionidae), and Asian longhorned beetle, Anoplophora glabripennis Motschulsky (Coleoptera: Cerambycidae) (Jiao et al. 2008;

Edworthy et al. 2011), and similar increases in woodpecker populations have been seen in this

EAB system (e.g. Lindell et al. 2008; Jennings et al. 2013). For example, up to 95 % EAB mortality has been observed as a result of woodpecker predation (Cappaert et al. 2005).

Additionally, high EAB parasitism rates of > 40% by Atanycolus cappaerti Marsh and

Strazanac (Hymenoptera: Braconidae) (Cappaert and McCullough 2009) and Phasgonophora sulcata Westwood (Hymenoptera: Chalcididae) (Lyons 2010) have also been observed and suggest that there is an opportunity to use native parasitoids in an augmentative biological control program against EAB. Although natural enemies native to North America have the potential to be used in an augmentative biological control program against EAB, they have received limited attention to date. Successful augmentative biological control using these native parasitoids will involve the appropriate timing and location of release and knowledge of factors that affect their dispersal (Chapters 2 and 3), determining how to appropriately monitor populations after releases

71 are made (Chapter 5), and an effective method to distribute large quantities of parasitoids (Chapter

4).

A system for the rearing and release of a biological control agent is important, especially for commercial application. Successful and well-studied examples of mass rearing for use in both agricultural and forest systems include those that involve Trichogramma spp. (Hymenoptera:

Trichogrammatidae) egg parasitoids (see Smith 1996). Although several native parasitoids show promise as effective biological control agents against EAB, some of their life-history traits are unfavourable for laboratory rearing. For example, P. sulcata is both solitary and univoltine wasp

(Roscoe 2014); its 1:1 parasitoid:host ratio and long generation time makes insect culturing impractical. Thus, it is necessary to identify better methods for augmenting and disseminating their populations in the field if they are to be used to combat EAB populations. Field collection and transport of parasitoid-infested plant material may be one approach. Although this has proven to be a cost-effective approach for disseminating natural enemies to combat invasive forest insects

(see Haugen and Underdown 1990), this method of augmentative biological control has been attempted in very few forest insect systems to date, so the proof of concept and criteria for long- term establishment and spread of parasitoids remains to be shown.

The objective of this study is to determine whether the collection and transport of parasitoid-infested ash material to sites where their native populations are absent can be used as a means of developing an augmentative biological control program against EAB to help slow its spread. The relationship between EAB parasitism by native North American parasitoids and EAB density is examined until 2016 to assess the relative rate of increase in parasitism by native North

American parasitoids. Specifically, I (1) investigate the influence of augmenting parasitoid populations via ash log transport on EAB parasitism, EAB density, and woodpecker predation over multiple years, and (2) determine how woodpecker predation affects EAB density over time

72 and native parasitoid spread. I predict that (1) transported parasitoids will establish on EAB in the new sites and EAB parasitism will increase over time, and (2) woodpecker predation will increase with EAB density irrespective of time.

4.3 Materials and Methods

Experimental design

In 2013, woodlots near the epicentre of the initial EAB invasion in southwestern Ontario, Canada were surveyed for a source of native parasitoids attacking EAB. Additionally, data obtained from ash logs collected by the Canadian Forest Service in 2012 were used to locate suitable collection sites. Three collection sites were identified, two in Middlesex County, Ontario and another in

Elgin County, Ontario, both which had ≥ 50 % EAB parasitism by P. sulcata and Atanycolus spp. combined. At the same time, parasitoid-release and non-release control plots were established in

3 mixed hardwood forests within the Toronto, Ontario, Canada that had at least ≥ 25 % ash trees with a diameter at breast height (DBH) of 9.5 ± 1.0 cm and where no current EAB parasitism was recorded. Over three years, parasitoid-infested ash trees within each collection site were cut into

60-cm lengths and the logs transported to the parasitoid-release plots each year in early spring.

Parasitoid-infested ash logs were divided such that there were a similar number of logs from each collection site in terms of both surface area and quantity placed adjacent to living ash trees in each parasitoid-release plot. Non-release control plots without parasitoid-infested ash logs were located at least 500 m away from the release plots based on our observations [later confirmed by Gaudon et al. (2018)] that P. sulcata has a relatively poor dispersal capacity. These plots were of similar stand composition, age, and structure and were used to monitor comparative changes in EAB parasitism, density, and woodpecker predation.

73 Ash trees were sampled in 2013 before parasitoid-infested logs were transported to the sites, as well as one and three years after the initial transport. In 2013, I removed one to three logs per sample tree from two to three trees per plot to measure the number of native EAB parasitoids, number of EAB, and woodpecker predation rates. In the following year (2014), I sampled fewer trees (i.e. one to two trees) per plot, removing more logs (i.e. two to four logs per tree) to conserve the number of living trees remaining in each plot. Due to the destructive nature of sampling and rapid decline of ash trees in these EAB-infested plots, a complete sample of trees in the parasitoid- release and non-release control plots was not done in 2015, but rather all plots were examined to ensure that sufficient living ash and viable phloem was available for EAB and its parasitoids to reinfest. In the final sampling year (2016), a much larger sample was taken, using two to four logs per tree from two to seven trees per plot. Once removed from the sites, logs were fully enclosed with emergence netting and hung on hooks in a temperature- and humidity-controlled rearing chamber with a mean temperature and humidity of 22.8 ± 0.5 °C and 51.4 ± 2.9 %, respectively.

The length and diameter of both ends of all logs were measured, and a mean surface area for each log was calculated.

Counts were made of the number of adult EAB and parasitoids emerging from the ash logs collected in order to measure the change in EAB parasitism and EAB density between the parasitoid-release and non-release control plots. Logs were dissected after 4 months to ensure that all EAB and parasitoids had been accounted for, including those that had failed to emerge. EAB parasitism in the release plots were calculated for solitary parasitoids by dividing the total number of wasps by the sum of the total number of EAB and wasps collected from the logs (Lyons 2010;

Roscoe et al. 2016) and then compared with control plots using the same criteria.

Woodpecker predation rate was also determined by visually inspecting each log for damage to the bark and/or sapwood leading directly to an EAB larval gallery or pupal chamber

74 inside each ash log. EAB density was measured as the sum of the number of D-shaped exit holes and woodpecker predation holes on each ash log.

Data analyses

I tested the effects of treatment and year on EAB parasitism using logistic regression with a binomial distribution. I assessed this model’s fit by looking for overdispersion and underdispersion by comparing the residual deviance and residual degrees of freedom, which were considered similar. Tukey’s range test was used to explore significant differences between means of significant effects.

Multiple linear regression was used to analyze EAB density and its relationship with treatment and year. I graphically assessed my assumptions of this model and observed both heteroscedasticity of residuals and normality of errors, which were assumed.

Finally, I used logistic regression to test the effects treatment, year, and EAB density on woodpecker predation. First, a binomial distribution was assumed; however, this model did not meet the assumptions using the same approach to diagnose overdispersion and underdispersion above, so I refit a model using a quasibinomial distribution, resulting in a better model fit.

All data were analyzed in the R statistical environment (R Development Core Team 2018).

Interaction terms were considered for all models; if there were no significant interactions between model terms, they were removed to use the simplest model that considered all main effects.

Statistical significance was set at P < 0.05 and all errors reported are ± 1 standard error from the mean.

75 4.4 Results

Data from the Canadian Forest Service showed 49.4 % EAB parasitism by P. sulcata and only

0.6 % EAB parasitism by Atanycolus spp. (n = 2063 EAB) in parasitoid collection sites. In 2013,

I observed 46.15 % EAB parasitism by P. sulcata and 30.77 % EAB parasitism by Atanycolus spp., as well as 7.69 % EAB parasitism by Balcha indica (Mani and Kaul) (Hymenoptera:

Eupelmidae) from a small sample of the ash logs being transported to the parasitoid-release plots in 2013 (n =

13 EAB). Of 13 EAB collected, 11 were parasitized by native parasitoids suggesting that only a small fraction of EAB will be transported into parasitoid-release sites.

Phasgonophora sulcata, Atanycolus spp., and Metapelma spectabile Westwood

(Hymenoptera: Eupelmidae) were all found emerging from ash trees in sites that had no previous record of parasitism after only one year of transporting parasitoid-rich ash logs to these areas.

Spathius flordianus was also recorded emerging from ash logs collected from parasitoid-release plots, however percentage EAB parasitism by this species could not be determined as it is a gregarious wasp where more than one wasp emerges from a single EAB host. As well, B. indica was found emerging from the sample logs being transported to the release plots, however it was not recovered in either of the release or non-release control plots over any of the sampling years

(Table 4.1).

Overall EAB parasitism by native parasitoids was influenced significantly by treatment

(χ2 = 6.24; df = 1; P = 0.012) and year (χ2 = 54.68; df = 1; P < 0.001). Overall parasitism was significantly higher in parasitoid-release plots than non-release control plots. Further, parasitism levels were significantly higher in 2016 than in years 2013 or 2014, with 64.8 ± 18.1 % of EAB parasitized in the three release plots (Fig. 4.1).

76 EAB density ranged from 0 to 7 EAB per m2 of ash surface area across the parasitoid- release and non-release control plots. EAB density was highest (i.e. 7 EAB per m2 of ash surface area) at one of the release plots in 2014. There were no differences in EAB density between treatment plots (F = 0.43; df = 1, 8; P = 0.529) or treatment year (F = 0.80; df = 1, 8; P = 0.398).

I also found a range of woodpecker predation across parasitoid-release and non-release control plots, with a low of 0 % to a high of 70.3 % EAB predation by woodpeckers, and woodpecker predation was highest in one of the control plots in 2016. Despite large variation in woodpecker predation rates, woodpecker predation did not differ significantly between the release and control plots (χ2 = 1.29; df = 1; P = 0.256), between years (χ2 = 0.24; df = 1; P = 0.623), or across the range of EAB density (χ2 = 1.97; df = 1; P = 0.161).

4.5 Discussion

Studies on native parasitoids attacking EAB usually report low parasitism (i.e. < 1 %) of EAB in newly-invaded regions (e.g. Liu et al. 2003; Duan et al. 2010, 2012), however, high parasitism by

P. sulcata and Atanycolus spp., notably A. cappaerti, have been observed in some Canadian and

American sites (e.g. Cappaert and McCullough 2009; Lyons 2010). Chapter 3 linked abundance of these two genera and their impact on EAB (i.e. EAB parasitism) to select vegetation and habitat characteristics found in sites from southern Ontario, notably tree biomass, tree condition, canopy closure, and floral resource availability. Additional evidence suggests that these parasitoids may not have strong dispersal capacities [e.g. P. sulcata, Gaudon et al. (2018)], and thus their populations would not be expected to move quickly with the spread of EAB. Although vegetation, habitat characteristics, and dispersal capacities in part drive the natural variation of native parasitoids and their impact on EAB, I show that EAB parasitism by North American parasitoids

77 can be significantly increased by augmenting their populations through transport of parasitoid- infested ash logs irrespective of site characteristics.

The EAB parasitism rates achieved after augmentative releases here were similar to those parasitism rates seen in existing classical biological control programs with non-native T. planipennisi and O. agrili. A study by Duan et al. (2018) found that EAB parasitism by T. planipennisi was ~1 to 6 % from 2008 to 2011 and increased to ~30 % parasitism by 2014 in both parasitoid-release and non-release control plots, while O. agrili averaged 1 to 4 % EAB parasitism from 2008 to 2011 and then increased to ~28 % by 2014 in release plots. I found a comparable steady increase in overall EAB parasitism by Atanycolus spp. in parasitoid-release plots, however,

I observed decreases in EAB parasitism by both M. spectabile and P. sulcata in the release plots by the end of the three-year study. It is possible that too few M. spectabile individuals were transported and mate limitation negatively affected their population growth and establishment, and that the decrease in EAB parasitism by P. sulcata that was observed here was due to an increase in host refuges for EAB, in particular as the infestation progressed from the canopy and branches downward toward the lower trunks in trees where bark thickness would limit parasitism.

The short ovipositor length of P. sulcata [i.e. approximately 6.4 mm (Roscoe 2014)] suggests that these wasps would be unable to parasitize all EAB larvae within the plots, especially those at the lower tree levels because ash bark thicknesses reached approximately 12.0 mm in some of the trees sampled (J.M. Gaudon, unpublished data). Abell et al. (2012) observed similar refuges for

EAB from T. planipennisi, where this introduced parasitoid was unable to successfully parasitize

EAB in ash trees with a bark thickness > 3.2 mm. In contrast with T. planipennisi and P. sulcata,

Atanycolus spp. have much longer ovipositors and can successfully parasitize EAB in ash trees with a bark thickness up to 8.8 mm (Abell et al. 2012). Such refuges for EAB would limit the effectiveness of P. sulcata but not Atanycolus spp. as biological control agents against EAB.

78 Here, I report the first observation of S. floridanus attacking EAB in Canada. Until now,

S. floridanus has been on its native hosts in eastern North America (Lyons 2010) and only observed attacking EAB in the USA (e.g. Duan et al. 2012). I suspect this is an artefact of my biological control efforts as it was undetected during my baseline sampling for native EAB parasitoids in the parasitoid-release and non-release control plots, and remained absent in the control plots over the entire study period.

Although EAB parasitism was significantly higher in the parasitoid-release plots than the non-release control plots, there was no difference in EAB density between the treatment plots or between treatment years. Duan et al. (2015) also found no difference in EAB densities (including all larvae and adults) between parasitoid-release and non-release control plots over a 7-year study with T. planipennisi. Unlike here, Duan et al. (2015) found that EAB density differed significantly by year, but their study was > 2 times longer than the current one allowing more time to see an effect of biological control or natural year-to-year variation. Duan et al. (2015) showed that EAB density was first decreased by native parasitism then by introduced T. planipennisi suggesting that both are important mortality factors of EAB. Thus, it may be beneficial to attempt to establish populations of both native and introduced parasitoids in biological control programs against EAB.

My work shows no difference in woodpecker predation between parasitoid-release and non-release plots, treatment years, and most interestingly, no relationship between woodpecker predation and EAB density. While some studies have found predation by woodpeckers has a positive linear relationship with EAB larval density (e.g. Lindell et al. 2008; Jennings et al. 2013), others have not (e.g. Duan et al. 2010). Woodpecker foraging behaviour differs in the presence of parasitized EAB larvae such that fewer EAB larvae are preyed upon than if they were not parasitized, although woodpecker predation does not seem to affect rates of EAB parasitism

(Murphy et al. 2018). Murphy et al. (2018) postulate that parasitism reduces the quality of the

79 food source (i.e. EAB larvae) such that it is not rewarding relative to foraging effort. Thus, woodpeckers leave the low-quality patch to locate a high-quality patch (i.e. with less parasitized

EAB). This may be the case, however it is also possible that woodpeckers are unable to locate

EAB larvae parasitized by idiobiont parasitoids, such as Atanycolus spp., because their host is paralyzed prior to oviposition and therefore not active and difficult to detect. This would also mean that woodpecker predation would be unlikely to affect the rates of EAB parasitism seen here because increasing rates of EAB parasitism was driven by Atanycolus spp. Interestingly, I found that woodpecker predation rates decreased, although not significantly, as EAB parasitism by

Atanycolus spp. increased in the parasitoid-release plots. Overall, it is expected that EAB populations would decrease (i.e. the population growth rate of EAB would be r < 1) if ~95 %

EAB mortality occurred, 60 % of which is caused by woodpecker predation and 35 % of which is a result of parasitism (Duan et al. 2018). My research suggests that this may be possible to achieve by native natural enemies alone.

The transport of logs appears to be an effective method to augment parasitoid populations, and this may be an easier and cost-effective approach compared to mass rearing parasitoids in the laboratory or a production facility although there may be some challenge of locating parasitoid- infested logs. Other studies have shown that redistributing infested plant material is both effective in disseminating natural enemies and cost-effective compared to rearing cultures (e.g. Haugen and Underdown 1990). In their study, these authors recommended transporting parasitoid-infested logs collected in the field as part of a strategy to combat another invasive forest insect, Sirex noctilio F. (Hymenoptera: Siricidae), in South Australia. They found this to be a successful option for increasing populations of Megarhyssa nortoni (Cresson) (Hymenoptera: Ichneumonidae), an important parasitoid of S. noctilio (Haugen and Underdown 1990). Using this approach, they released greater numbers of M. nortoni through log transport than in previous efforts to rear

80 parasitoids in the laboratory (Haugen and Underdown 1990). Although I observed no difference in EAB densities between treatment plots, transport of parasitoid-infested ash logs may be incorporated as part of a strategy to slow the spread of EAB as field-collected logs can supply large numbers of parasitoids to significantly increase parasitism of this invasive forest insect under field conditions, and natural enemy populations can be extremely abundant at some sites.

81 Tables

% parasitism / predation % parasitism / predation in control plots (n = 3) in release plots (n =3) Mortality factor 2013 2014 2015* 2016 2013 2014 2015* 2016 Chalcididae Phasgonophora 0.00 0.00 - 0.00 0.00 10.32 ± 5.20 0.00 16.67 ± 16.67 sulcata Braconidae Atanycolus spp. 0.00 16.67 ± 16.67 - 5.88 ± 5.88 0.00 0.00 8.11 48.18 ± 4.29 Spathius ------floridanus Eupelmidae Balcha indica 0.00 0.00 - 0.00 0.00 0.00 0.00 0.00 Metapelma 0.00 0.00 - 0.00 0.00 4.76 ± 4.76 0.00 0.00 spectabile Picidae - 27.78 ± 14.70 - 49.32 ± 20.98 - 30.99 ± 16.67 - 25.39 ± 4.57 (woodpeckers) *Sampled from Centennial Park only

82 Figures

83 Chapter 5 Evaluating Methods to Detect and Monitor North American Parasitoids of the Emerald Ash Borer

5.1 Abstract

Populations of native North American parasitoids attacking Agrilus spp. Curtis (Coleoptera:

Buprestidae) have recently been used in an augmentative biological control program in an attempt to slow the spread of the emerald ash borer (EAB), Agrilus planipennis Fairmaire, a destructive wood-boring beetle that was discovered in North America in 2002. I evaluate trapping methods to detect and monitor populations of two important parasitoid groups, Phasgonophora sulcata

Westwood (Hymenoptera: Chalcididae) and Atanycolus spp. Foerster (Hymenoptera:

Braconidae), attacking EAB in its introduced range. I found that purple prism traps captured more

P. sulcata than green prism traps, yellow pan traps, and log samples, and thus are better for detecting and monitoring P. sulcata populations. Although there was no difference between captures of P. sulcata on purple prism traps or sticky band traps at one study site, purple prism traps performed significantly better at another. Trap type did not affect the number of captures of

Atanycolus spp. Surprisingly, baiting prism traps with a green leaf volatile did not significantly increase captures of P. sulcata or Atanycolus spp. Based on these results, purple prism traps would be optimal for sampling these important native parasitoids attacking EAB.

5.2 Introduction

There are a number of parasitoids that attack native Agrilus spp. Curtis (Coleoptera:

Buprestidae) in North America that have been observed attacking the introduced emerald ash borer (EAB), Agrilus plannipennis Fairmaire (Coleoptera: Buprestidae). Specifically, native larval parasitoids Phasgonophora sulcata Westwood (Hymenoptera: Chalcididae) and Atanycolus spp. Foerster (Hymenoptera: Braconidae) are particularly important mortality factors of EAB.

Parasitism by these two parasitoids has been observed as high as 40 % and 71 %, respectively, in some sites (Cappaert and McCullough 2009; Lyons 2010). These parasitoids are weak dispersers

(Gaudon et al. 2018), and both their population sizes and EAB parasitism are correlated with ecosystem characteristics (e.g. tree condition, floral resource availability) and at least partly explain the strong localized impact on EAB populations observed in some sites (Chapter 3). The variation in native EAB parasitoid abundance across the North American landscape necessitates the discovery of an insect trap that can detect and monitor these parasitoids if they are used in a biological control program against EAB.

Several EAB parasitoids, including species from Asia and North America, have been identified and are currently being used in biological control programs to slow the spread of EAB

(see Bauer et al. 2015; Chapter 4 herein). In general, sampling for parasitoids in this EAB system is labour-intensive, involving the cutting, sampling, and movement of many large-sized logs from the field to the laboratory where elaborate cardboard rearing and emergence tubes or wooden rearing and emergence cabinets are needed to collect the beetle and its natural enemies. This approach is destructive, time consuming, and requires considerable space and resources. Thus, there is a need to develop a more non-destructive and cost-effective sampling technique to survey and monitor the establishment of native EAB parasitoids either before augmentative releases or to follow their establishment and movement after release (Chapter 4).

85 Parasitoids can be sampled using a wide variety of techniques (e.g. Malaise traps, yellow pan traps, sweep netting), and different sampling methods will collect different groups of parasitoids (e.g. parasitic Hymenoptera, Darling and Packer 1988; Aguiar and Santos 2010).

These recovery techniques can be implemented to monitor the establishment and impact of parasitoid populations in biological control projects. To date, methods for surveying and monitoring EAB parasitoid populations have focused on the introduced Asian parasitoids (e.g.

Parisio et al. 2017) and not on native species. Thus, it is crucial to determine which traps are most effective for sampling native EAB parasitoids.

Traps designed to capture host insects have been able to simultaneously detect their parasitoids (e.g. Derocles et al. 2014). Such an approach, where native EAB parasitoids are also collected on traps designed to detect and monitor their wood-boring beetle host, would be extremely useful and appears promising (e.g. Roscoe et al. 2014). A recent review and meta- analysis by Allison and Redak (2017) found general patterns in trap design features for surveying wood-boring beetles. For example, panel traps generally captured more bark and wood-boring beetles than multiple-funnel traps, while traps treated with a slippery coating increased captures of wood-boring beetles compared to those left untreated (Allison and Redak 2017). Work on trap colours has shown that black or purple traps are superior in capturing insects to white and green traps, respectively, with purple being an important visual stimulus for trapping Agrilus spp.

(Allison and Redak 2017). Currently, attractant-based traps are considered the most effective for detecting (low-density) populations of non-native forest beetles (Coleoptera) (Allison and Redak

2017), however it is unknown whether traps now being used to sample EAB (e.g. purple or green prism traps) should also be implemented for native parasitoids of EAB and similar beetles.

The objective of this chapter is to identify the optimal insect trap that will sample native parasitoids along with EAB. Specifically, I evaluate the efficacy of trap type, visual stimulus, and

86 bait to determine the optimal sampling strategy for monitoring native North American EAB parasitoids following augmentative release. I predict that dark-coloured traps (i.e. purple) treated with a bait (i.e. an attractant from the host plant and/or host beetle) will capture the greatest numbers of native EAB parasitoids, as these traps best resemble the silhouette of a tree that a native Agrilus parasitoid may use as a visual stimulus to orient themselves toward their host insects (e.g. Lindgren 1983).

5.3 Materials and Methods

Experimental field trials

Two experimental field trials were used to test different trap designs with varying stimuli over three years. In the McKeough Conservation Area (Wilkesport, Ontario), comparisons were made between trap types, visual stimuli, and baits using 6 purple prism traps, 6 green prism traps, and

30 sticky band traps; half of the purple (n = 3) and half of the green (n = 3) prism traps were baited with a green leaf volatile (GLV), (Z)-3-hexenol. The traps were deployed by the Canadian Forest

Service (CFS) between 2-4 June 2010, and P. sulcata and Atanycolus spp. were collected from them approximately every two weeks until 9-12 September 2010. Individuals were collected from the traps, brought to the Great Lakes Forestry Centre (Sault Ste. Marie, Ontario, Canada), and identified by the CFS.

The second site in the study was a naturalized area (Merchants Trail) in Oakville, Ontario, where 15 yellow pan traps, 15 sticky band traps, and 15 purple prism traps were installed on ash trees on 25 May 2016. Phasgonophora sulcata and Atanycolus spp. were collected from the yellow pan traps every week and from the sticky band traps and purple prism traps every two weeks until 18 August 2016. The following year (2017), the same number of yellow pan traps,

87 sticky band traps, and purple prism traps were installed on ash trees and collected as in 2016 to examine population variability between years. These parasitoids were also sampled in this site by cutting three ~60-cm logs from 12 ash trees (n = 36) during 2017 in order to rear P. sulcata and

Atanycolus spp. to adulthood and compare to the number of parasitoids caught in the pan, sticky, and prism traps that same year with the expectation that log samples will provide an accurate representation of parasitoid activity in the site.

Data processing and analyses

Counts over the sampling period were summed and the total catch per trap analyzed. Data collected in the McKeough Conservation Area in 2010 and in Oakville during 2016 and 2017 were analyzed separately.

Counts of the number of P. sulcata captured in the McKeough Conservation Area were analyzed using a linear model (lm) because total numbers of wasps captured per trap were quite large. I assessed this model for homogeneity of variance and normality of errors, and this model did not meet these assumptions, so I log-transformed the number of P. sulcata captured. The lm with the log-transformed response met my assumptions of homogeneity of variance and normality of errors.

Fewer Atanycolus spp. were captured in the McKeough Conservation Area. As such, a generalized linear model (GLM) with Poisson errors and a log link function was fit to the effects of trap type on the counts of Atanycolus spp. captured. I diagnosed over/underdispersion by examining the residual deviance and degrees of freedom. Efforts to account for this over/underdispersion by refiting a GLM with quasipoisson errors and a log link function were not successful, so I used a negative binomial model, which was a better fit.

A GLM with Poisson errors and log link function was first fit to the effects of trap type and year on the number of P. sulcata captured in the Oakville site. I examined the GLM fit by

88 looking for overdispersion and underdispersion by comparing the residual deviance and degrees of freedom, which were considered dissimilar. I attempted to improve the fit by using quasipoisson errors but did not succeed. As such, I used a negative binomial model to test the effects of trap type and year on the number of P. sulcata captured. I explored the fit of this model using the methods above and was satisfied with the fit.

A GLM with Poisson errors and log link function was also fit to the effects of trap type and year on the number of Atanycolus spp. captured in the Oakville site. I again examined the

GLM fit by using methods described above, and the residual deviance and degrees of freedom were dissimilar. The fit could not be improved by using quasipoisson errors. As such, I again used a negative binomial model to test the effects of trap type and year on the number of Atanycolus spp. captured. I explored the fit of this model by looking for over/underdispersion as described above and was satisfied with the fit.

Data were analyzed in the R statistical environment (R Development Core Team 2018).

All interactions were initially considered, but in cases where models would not converge if interactions were included, I was only able to explore main effects. In all other cases, non- significant interaction terms were sequentially removed from models to use the simplest model that at minimum explored main effects (i.e. trap type and year). Tukey’s range test was used to explore significant difference between main effects at alpha 0.05. Significance was set at P < 0.05 and all errors reports are ± 1 standard error from the mean unless otherwise specified.

5.4 Results

The maximum number of P. sulcata captured on a trap in the McKeough Conservation Area was

191 wasps on one baited purple prism trap. Baited green prism traps captured 0 P. sulcata during

89 the 2010 study. Fewer Atanycolus spp. were captured than P. sulcata, with a maximum number of 22 wasps captured on one sticky band trap at this site.

The number of P. sulcata captured was significantly influenced by trap type (F = 25.61; df = 4, 37; P < 0.001) in the McKeough Conservation Area (Fig 5.1). Baited and unbaited purple prism traps and sticky band traps caught significantly more P. sulcata than unbaited green prism traps, and unbaited green prism traps caught significantly more P. sulcata than baited green prism traps (see Fig. 5.1). The number of Atanycolus spp. captured was not influenced by trap type (χ2

= 7.50, df = 4, P = 0.112).

Fewer parasitoids were captured at the Oakville site. The maximum number of P. sulcata captured was 21 wasps on two purple prism traps in 2016, while the maximum number of

Atanycolus spp. captured was 10 wasps on one sticky band trap in 2016.

The number of P. sulcata captured was significantly influenced by trap type (χ2 = 90.02, df = 3, P < 0.001) and year (χ2 = 4.18, df = 1, P = 0.041) in Oakville (Fig. 5.2). Significantly greater numbers of P. sulcata were captured using purple prism traps than all other traps types

(i.e. sticky band traps, yellow pan traps) or by taking log samples (see Fig. 5.2). Captures of P. sulcata were significantly higher in 2016 than 2017 (see Fig. 5.2).

The number of Atanycolus spp. captured was not influenced by trap type (χ2 = 6.23, df =

3, P = 0.101), however year significantly influenced the number of Atanycolus spp. captured at the Oakville site (χ2 = 21.55, df = 1, P < 0.001). Similar to P. sulcata, more Atanycolus spp. were captured in 2016 than 2017.

5.5 Discussion

I show that purple prism traps perform significantly better for capturing P. sulcata than green prism traps, sticky band traps, or yellow pan traps, however the total catch of Atanycolus spp. was not affected by any of the trap types examined and these wasps were captured on all trap types.

Purple prism traps are particularly useful as monitoring devices as these are also used to sample

EAB populations, and thus one trap can sample both EAB and these two groups of native parasitoids. Sampling infested ash logs did not appear to provide a ‘real’ measure EAB parasitism.

Annual variation in P. sulcata and Atanycolus spp. population sizes also affects total trap catch irrespective of trap type and it appears important to consider the stage of EAB infestation while sampling within sites. For example, more native parasitoids will likely be caught on traps installed on ash trees that display ~50 % canopy decline, while traps installed on healthy or dead ash trees will likely capture fewer parasitoids if any at all (Chapter 3).

Parisio et al. (2017) assessed different methods to recover and monitor populations of

Tetrastichus planipennisi Yang (Hymenoptera: Eulophidae), Oobius agrili Zhang and Huang

(Hymenoptera: Encyrtidae), and Spathius agrili Yang (Hymenoptera: Braconidae), all which are

Asian EAB parasitoids first released in North America in 2007 to combat EAB, which is especially important work needed in order to accurately confirm their establishment and impact on invasive EAB populations. Yellow pan traps were found to be effective for sampling populations of all introduced EAB parasitoids (Parisio et al. 2017). Here, yellow pan traps took more time install and service and did not appear to reflect the number of native EAB parasitoids within the site, especially when compared to purple prism traps. Thus, it may be that multiple trap types are needed to sample for both introduced and native EAB parasitoids in North America.

Branch sampling has been shown to be an important technique to sample EAB in open grown ash trees in newly-infested urban regions (Ryall et al. 2011) and the use of larval sentinel

91 logs and destructively sampling ash trees are not necessarily significantly different compared to yellow pan traps in terms of their efficacy for sampling introduced EAB parasitoids (see Parisio et al. 2017). Surprisingly, logs were unable to detect P. sulcata and Atanycolus spp. in the Oakville site. This may be an artefact of the stage of the EAB infestation at the Oakville site, where ash trees in new or old infestations contain few EAB and native parasitoids, and the mature ash trees in this Oakville site were dead while the young ash were apparently healthy indicating an old EAB infestation.

In the laboratory, P. sulcata elicited an antennal response to a GLV, (Z)-3-hexenol

(Roscoe 2014). The same GLV was used here to bait the traps. Surprisingly, there was no difference in captures of P. sulcata or Atanycolus spp. between traps with or without the GLV bait. In the field, these parasitoids may be using this GLV in combination with other semiochemicals, especially those from the host [e.g. 3-(Z)-lactone, a pheromone released by EAB when it feeds on ash foliage, Bartelt et al. (2007)] to locate EAB in the field.

Trap placement is an important consideration when attempting to detect or monitor insect populations. For example, Ulyshen and Sheehan (2017) found that optimal trap height for beetles is dependent on their feeding guild, with greater numbers of phloem and wood-boring beetle captures on traps installed high above the ground while greater number of ambrosia beetle captures on traps set up near the ground. Further, Trichogramma minutum Riley (Hymenoptera:

Trichogrammatidae) have been observed to disperse only 18.5 m over 5 days in forest stands

(Smith 1988) whereas Trichogramma ostriniae Pang and Chen (Hymenoptera:

Trichogrammatidae) was found to disperse ~175 m after 4 to 7 days in corn fields (Gardner et al.

2012), demonstrating that trap distance from where a biological control agent is released is important. It also demonstrates that structure complexity of the ecosystem is important and that there are differences between forests and fields. Here, traps close to the emergence site of the

92 parasitoid (i.e. the purple prism traps deployed in tree canopies) were more likely to be effective, especially for P. sulcata likely in part due to its relatively poor dispersal capacity (Gaudon et al.

2018). Future implementation might mean that traps should be hung no farther than 300 m away from the nearest infested site in order to detect native parasitoids.

Of interest to land managers, purple prism traps that were initially designed to capture

EAB can also be used to detect if native EAB parasitoids are also present in the region. This might be a cost-effective approach that can be used to determine how best to combat the EAB invasion in an area, especially if funds are limited. For example, if there are high numbers of native EAB parasitoids and low numbers of EAB on a purple prism traps, one may focus biological control efforts elsewhere. Studies that investigate traps that can be used to sample both a parasitoid and its host are rare. However, a study by Derocles et al. (2014) showed that aphidiine parasitoids

(Hymenoptera: Aphiidinae) were indirectly captured on the same traps used to capture their aphid hosts. Interestingly, it appears that adult P. sulcata will preferentially orient themselves toward and land on purple prism traps. Traps that can be used to sample both a host and its parasitoid

(e.g. purple prism traps discussed here) would be particularly cost-effective and time-efficient if used in a biological control program.

Although the surface area of the purple prism traps used in this study was much greater than those of the yellow pan traps, they were similar to the surface area of the green prism traps and sticky band traps, which still captured significantly fewer P. sulcata than purple prism traps.

Further, captures of Atanycolus spp. were similar on all traps regardless of trap surface area. It is likely that trap colour and placement (i.e. traps in the canopy vs. traps close to the ground) are more important factors than trap surface area for some parasitoid species. Both colour and placement may work synergistically such that more native EAB parasitoids, such as P. sulcata, may be captured on purple prism traps in the canopy.

93 Moving forward, studies that investigate trap placement (i.e. height) are necessary to determine the most useful trapping protocol for sampling EAB parasitoids. Additionally, the

GLV-baited prism traps did not perform significantly better than non-baited traps; investigating traps that include an alternative bait, such as 3-(Z)-lactone, or a combination of baits, such as 3-

(Z)-lactone plus a GLV, would be particularly useful to determine whether purple prism traps can be used and improved for greater trap catches of P. sulcata, Atanycolus spp., and even other native

EAB parasitoids. Such research will build a better trapping protocol to detect and monitor populations of EAB parasitoids and to integrate with augmentative releases against EAB.

94 Figures

96 Chapter 6 General Conclusions and Future Research

Natural enemies are an important tool for the long-term management of emerald ash borer (EAB),

Agrilus planipennis Fairmaire (Coleoptera: Buprestidae), in its introduced range. Although various areas of North America have been surveyed for natural enemies attacking EAB (e.g. Liu et al. 2003; Bauer et al. 2004; Cappaert and McCullough 2009; Duan et al. 2009; Lyons 2010), relatively few studies have investigated the biology and ecology of North American parasitoids attacking EAB, especially in the context of their potential for biological control (but see Lyons

2010; Roscoe 2014; Duan and Schmude 2016). The purpose of this study was to investigate the role native natural enemies may have in supressing EAB populations in North America. My research explored the biological and ecological drivers associated with the arrival and detection of two important groups of native parasitoids, Phasgonophora sulcata Westwood (Hymenoptera:

Chalcididae) and Atanycolus spp. Foerster (Hymenoptera: Braconidae), attacking EAB in Ontario

(Chapters 2 and 3) to use this knowledge to improve augmentative releases of native parasitoid populations through transporting parasitoid-infested ash [Fraxinus L. (Lamiales: Oleaceae)] material (Chapter 4) and develop a trapping protocol for important native parasitoids attacking

EAB (Chapter 5).

In this thesis, I evaluated the dispersal capacity of one important native parasitoid, P. sulcata, to determine its efficacy if it were to be used in an augmentative biological control program against EAB. Overall, P. sulcata was a weak disperser, and its flight capacity (i.e. number of flight bouts) decreased with age suggesting the capacity to search and attack EAB will similarly decrease with age. Consistent with existing literature, the flight capacity of female wasps increased with temperature. I also explored the potential fecundity of this parasitoid wasp to verify

97 that it is pro-ovigenic and found that oosorption of eggs occured before the females reach 30 days of age, much sooner than reported in the existing literature (see Roscoe 2014), yet flight distance or activity did not significantly affect the number of eggs resorbed. Based on this work, I recommend releasing P. sulcata close to EAB if this species is used for augmentative biological control (Chapter 4).

In Chapter 3, I explored the vegetation and habitat characteristics that best predict where native parasitoids of EAB can be found in high abundance. Several factors influence the abundance of P. sulcata and Atanycolus spp. attacking EAB, including tree biomass, tree condition, canopy closure, floral resource availability, and the abundance of EAB. Perhaps more importantly, overall parasitism of EAB by P. sulcata and Atanycolus spp. was influenced by tree biomass, tree condition, and floral resource availability. Using this information, augmentative releases of Atanycolus spp. may be important in forests where few native EAB parasitoids are present, such as those with (1) ash trees on the leading edge of an EAB invasion; (2) a low biomass of birch, poplar, and oak trees; or (3) have limited floral resource availability, since this genus is an important source of EAB mortality (Duan et al. 2015; Chapter 4).

In Chapter 4, I tested whether native EAB parasitoid populations can be augmented by transporting parasitoid-infested ash material over multiple years. This approach of transporting parasitoid-infested logs collected in the field is a relatively novel method of parasitoid dissemination compared to other biological control strategies and has only used in only one other study system (see Haugen and Underdown 1990). This approach was used because of the difficulties associated with rearing native EAB parasitoids in a laboratory. After three years of transporting parasitoid-infested ash material into parasitoid-release plots, EAB parasitism by native parasitoids was significantly higher in release compared to non-release plots. It remains to be determined whether higher parasitism actually leads to reduced EAB population growth or

98 whether it will allow ash to persist in the landscape over the long term, however one might assume that this additional mortality would have some beneficial effect.

Finally, I evaluated several trap types to determine which are most effective for detecting and monitoring native parasitoid populations attacking EAB. I found that purple prism traps can be used to detect and monitor the establishment of both EAB and native EAB parasitoids and are particularly effective for capturing P. sulcata compared to other traps used in this study. Purple prism traps can be used to determine the establishment and monitor population dynamics of augmentated native EAB parasitoid populations (Chapter 4). However, given that Asian EAB parasitoids are best sampled using yellow pan traps or sentinel logs (see Parisio et al. 2017), it is likely that multiple trap types (e.g. both purple prism traps and yellow pan traps) will need to be deployed to determine the establishment and monitor population dynamics of Asian and North

American EAB parasitoids if both groups are to be used in a biological control program against

EAB.

6.1 Ecological Significance

EAB has killed and damaged ecologically-important and economically-valuable ash trees across eastern North America since its accidental introduction, and the ongoing devastation continues to cost municipalities, conservation authorities, and other land owners millions of dollars. Ash trees are an important riparian species (see Nisbet 2014) and as street and boulevard trees in the urban environment (e.g. City of Montreal, City of Thunder Bay, and City of Winnipeg). After trees are infested with EAB, they quickly die (see Knight et al. 2013). Dead trees need to be removed from public spaces, along trail edges, and in backyards because they become very brittle and, as such, a liability for property damage, human injury, or even death. The mortality of ash trees leaves at least 43 arthropod species that are exclusively associated with ash trees in North America at risk

99 of extinction (Gandhi and Herms 2010a). Further, the loss of ash in the landscape will likely reduce other biodiversity and lead to “invasional meltdown” where only a few, likely invasive, species replace ash. Such species observed replacing ash in southern Ontario include European buckthorn, Rhamnus cathartica L. (Rosales: Rhamnaceae) and garlic mustard, Alliaria petiolata

(M. Bieb) Cavara and Grande (Brassicales: Brassicaceae) (J.M. Gaudon, personal observations).

The research herein can be used to help create strategies to retain part of the function ash trees have in forest ecosystems across eastern North America.

My work significantly expands upon the little-known biology and ecology of North

American parasitoids of Agrilus beetles, especially to improve releases of native parasitoids against EAB. Very few studies have examined the dispersal of insect parasitoids relative to the number of species belonging to this group. My work is the first to look at the dispersal capacity of Phasgonophora. These results have important implications for using P. sulcata in a biological control program against EAB as they suggest that these wasps should be released as pupae close to EAB to improve efficacy and potential biological control success. Further, my work is the first to explore vegetation and habitat characteristics influencing the abundance of native EAB parasitoid populations in southern Ontario and their impact against an invasive insect species. My results show that few native parasitoids are attacking EAB in forests with low flowering herbaceous ground vegetation and completely healthy or dead ash trees, and that native parasitoid populations are small in areas with high canopy closure. This understanding can assist in the selection of future release locations of both native or Asian EAB parasitoids.

Mass production of a biological control agent is important for the success of any biological control program. The inability to release large numbers of biological control agents may result in negative impacts on their population(s) from Allee effects (e.g. mate limitation or inbreeding).

Further, large-scale production of agents allows for their use across many localities. This is true

100 for Trichogramma spp. (Hymenoptera: Trichogrammatidae) that are used to combat both agricultural and forest pests worldwide (Smith 1996). Over 1,000,000 Asian parasitoids [i.e.

Tetristichus planipennsi Yang (Hymenoptera: Eulophidae), Oobius agrili Zhang and Huang

(Hymenoptera: Encyrtidae), and Spathius agrili Yang (Hymenoptera: Braconidae)] have been reared at the United States Department of Agriculture Emerald Ash Borer Biocontrol Facility in

Brighton, Michigan and released in sites in 19 U.S. states and 2 Canadian provinces (Ontario and

Quebec) (Bauer et al. 2015), and T. planipennsi is now being reared by the Canadian Forest

Service (CFS) at the Great Lakes Forestry Centre in Sault Ste. Marie, Ontario. A laboratory rearing protocol for native parasitoids has not been established. In light of the difficulties associated with rearing these wasps, I transported native parasitoid populations via a relatively novel approach moving parasitoid-infested logs and have shown this is a successful approach to disseminate natural enemy populations to increase their abundance and parasitism of a target host

(Chapter 4). Transport of parasitoid-infested ash material should occur prior to P. sulcata emergence and ash material should be positioned adjacent to EAB-infested ash trees.

This dissertation builds a strong biological and ecological basis on which to create augmentative biological control programs using native insect parasitoids more generally for managing future invasive species. Results from my work provide an example of native natural enemies operating as an important source of EAB mortality after augmentative biological control efforts. Few attempts have been made to augment natural enemy populations for biological control in Canada (MacQuarrie et al. 2016), but my work suggests that this can be effective to increase parasitism over only a few years. The addition of four groups of native natural enemy agents [P. sulcata, Atanycolus spp., Spathius floridanus Ashmead (Hymenoptera: Braconidae), and

Metapelma spectabile Westwood (Hymenoptera: Eupelmidae)] and one naturalized but non- native natural enemy agent [Balcha indica (Mani and Kaul) (Hymenoptera: Eupelmidae)] can

101 now be included in the list of those relocated for augmentative biological control in Canada, all five of which have been relocated to where their populations were absent. Although five agents were released, I consider at least two to have established (P. sulcata and Atanycolus spp.).

My research will assist biological control researchers and practitioners in developing research and biological control programs against EAB and other invasive species. For example, I found that traps used for sampling EAB can also be used to detect and monitor populations of its parasitoids, which is rarely reported in the literature but evidently possible. Thus, purple prism traps may be a cost-effective tool that can be used to survey for both EAB and its parasitoids to provide information on where to focus biological control efforts.

Temperature and photoperiod are important factors that can induce or terminate diapause in insects (Schowalter 2000). Temporal asynchrony between adult parasitoids and their preferred host stage may occur if the parasitoid and host respond and develop differently at changing climatic conditions. As EAB spreads to northerly latitudes or as the effects of climate change intensify, EAB may experience a two-year life cycle (Tluczek et al. 2011) or modify its oviposition behaviour (Wetherington et al. 2017). Consequently, native EAB parasitoids may have limited effectiveness in an augmentative biological control program. Wetherington et al.

(2017) observed a similar scenario with O. agrili and its EAB host under only slight temperature variation in the laboratory. Although the O. agrili life history strategy differs from those of the native parasitoids P. sulcata and Atanycolus spp. studied here, native larval parasitoids may experience desynchronization with their preferred EAB host stage under comparable climatic pressure as did this introduced egg parasitoid. This could lead to reduced EAB parasitism and survival of native EAB parasitoids in the future.

Enemy-free space may exist for EAB in the relatively short period of time it has been in southern Ontario, especially in newly-infested areas. According to the Enemy Release Hypothesis

102 (see Keane and Crawley 2002), such an enemy-free period would have allowed EAB to easily establish within the invaded area and its population to rapidly increase (although this mechanism may also operate in combination with a lack of host tree defenses in North American ash species).

Gratton and Welter (1999) found that a leafmining fly, Liriomyza helianthi Spencer (Diptera:

Agromyzidae), experienced enemy-free space following a host plant shift (i.e. when flies were introduced to novel host plants), however, such enemy-free space was better determined by the population variability of the fly parasitoids both within and between years over the course of their study. Native EAB parasitoids appear to show the same population variability, and thus annual releases that augment their populations may yield the greatest impact possible by these natural enemies on EAB populations.

Indeed, there is not a complete absence of natural enemies attacking EAB in its introduced range, however these natural enemies appear to slowly transition from native hosts to attack EAB and this may explain why EAB mortality by natural enemies, especially parasitoids, is low in newly-infested regions. Such relaxed natural enemy pressure coupled with low ash tree resistance to EAB attack drives rapid EAB population growth. Based on my findings, natural enemy pressure appears to be relaxed in part due to their weak dispersal capabilities (Chapter 2) and unfavourable environmental factors (Chapter 3), and this necessitates augmentation of their populations

(Chapter 4). However, my research also suggests that that these natural enemies should play a key role in the long-term suppression of EAB populations, especially at a local/ecosystem scale as I had observed in some sites (Chapter 2) and after augmenting their populations (Chapter 4). I hope that this will slow ash tree mortality such that “pockets” of ash trees will persist in the North

American landscape although at a lesser extent than previously seen. We now have an opportunity to rethink how we plan urban forests, especially to design forests that are more resilient to invasive

103 forest insects than before. These results remain consistent with existing literature that there is no

“silver bullet” to eradicate invasive EAB populations in North America.

6.2 Limitations and Future Research

This thesis adds to the collection of work on EAB and significantly expands our knowledge of the biology and ecology of North American parasitoids of Agrilus spp., however questions remain surrounding the long-term role of parasitoids in slowing the spread of EAB and ash tree mortality, including in aftermath forests, and if it is possible to design urban forests that foster large populations of natural enemies.

Indeed, it appears possible to augment populations of native parasitoids attacking EAB and consequently increase EAB parasitism (Chapter 4), however, it is still undetermined whether increased numbers of natural enemies and higher levels of EAB parasitism will slow ash tree mortality. The role indigenous parasitoids have in regulating EAB populations over the long term is also not yet clear. It would be useful to monitor changes in ash tree decline between parasitoid- release and non-release control plots and test for significant differences among the two treatments.

Proxies for ash tree decline may include canopy dieback (using a rating system), number of epicormic shoots, and number of bark splits, and would be particularly useful in evaluation the success of biological control efforts.

In the short term (i.e. ≤ 7 years), releases of both native and introduced parasitoids appear important for augmenting EAB mortality (Duan et al. 2015, 2017; Chapter 5). The long-term role of native and introduced EAB parasitoids in controlling EAB populations is less clear. It may be that T. planipennisi becomes an important mortality factor of EAB attacking ash saplings in aftermath forests across North America (e.g. Duan et al. 2017). Similarly, P. sulcata populations

104 could be an important source of EAB mortality on ash saplings and on small branches in the canopy of large ash trees, whereas Atanycolus spp. may be important for keeping EAB populations low in forests with large/mature ash (Duan et al. 2015; Chapter 5). Overall, biological control using native and/or introduced EAB parasitoids may be most successful in areas with less preferred or resistant ash species (Herms and McCullough 2014).

There is also a need to explore the interspecific interactions within this novel parasitoid community of both introduced Asian and native North American EAB parasitoids. Functional redundancy is an important attribute in resilient systems and, as such, is an important aspect of a biological control program in case one enemy agent fails to establish in a release site (e.g. as seen with S. agrili in this EAB system). In this EAB system, niche partitioning exists between parasitoids attacking EAB. However, interspecific competition may lead to multiparasitism and result in only one or few species establishing over others. For example, idiobiont parasitoids that paralyze their host typically outcompete koinobiont parasitoids because host development is crucial for the success of koinobiont parasitoids that require their host to continue development.

Thus, studying the community ecology of these parasitoids may be necessary to create a successful biological control program across the different forest types throughout North America.

Testing the efficacy of different traps, especially compared to purple prism traps, may continue to improve our current understanding of how to best detect and monitor populations of

P. sulcata, Atanycolus spp., and other native EAB parasitoids. Specifically, testing purple prism traps with a different lure [e.g. 3-(Z)-lactone, a pheromone emitted by EAB when it feeds on ash foliage (Bartelt et al. 2007), or a combination of 3-(Z)-lactone and (Z)-3-hexenol, both of which elicit antennal responses in P. suclata (Roscoe 2014) and perhaps other EAB/Agrilus parasitoids] at different heights and in sites at different stages of EAB infestation (i.e. new vs. old infestations) will provide key information on how to best deploy traps in sites. This research is particularly

105 important to assess the success of a biological control program against EAB using these parasitoids.

Conservation authorities, municipalities, and other land owners face the challenge of determining their planting needs after the initial invasion of EAB and in aftermath forests.

Specifically, the widespread mortality of ash trees due to their abundance in the landscape and extensive distribution throughout North America are key considerations for why we must explore tree planting needs for after EAB invasion and impact. Future research should investigate whether planting high diversity of native trees, including trees that support alternative hosts for EAB natural enemies, in areas in southern Ontario that have been previously infested by EAB will create a forest more resilient to this invasive insect. Landscape context (e.g. surrounding vegetation types) should also be considered and all trees planted should be over 4 cm diameter at breast height so that they can be readily infested by Agrilus beetles with the idea that natural host tree resistance will in part regulate these native populations, avoid a beetle outbreak, and prevent widespread tree mortatlity. For example, bronze birch borer, A. anxius Gory, and two-lined chestnut borer, A. billineatus (Weber), attack stressed and dying birch [Betula spp. L. (Fagales:

Betulaceae)] and oak [Quercus spp. L. (Fagales: Fagaceae)] trees, respectively, over healthy trees

(Anderson 1944; Dunn et al. 1986). Future research exploring habitat manipulation to attract native EAB parasitoids is particularly important to allow for ash tree reproduction and the persistence of ash trees in the landscape.

Finally, to conserve other parasitoid populations, it would be useful if vegetation and habitat characteristics could be used as proxies as discussed in Chapter 3 for P. sulcata and

Atanycolus spp. Parasitoids are often overlooked in terms of conservation efforts although they provide important resilience to disturbances by insect pests (but see Lassau and Hochuli 2005,

2007; Fraser et al. 2007, 2009; Vance et al. 2007). However, being such a speciose group of

106 insects, it is difficult to generalize such proxies for all parasitoids, and different studies exploring the link between parasitoid species over many different habitat types are required. Such studies would be particularly useful to create a diverse ‘natural’ reserve of parasitoids that could provide resistance to future invasive species, especially those that are insects, and I am now investigating this for forest types in southern Ontario, Canada.

107 Literature Cited

Abell, K.J., Bauer, L.S., Duan, J.J., and Van Driesche, R. (2014) Long-term monitoring of the introduced emerald ash borer (Coleoptera: Buprestidae) egg parasitoid, Oobius agrili (Hymenoptera: Encyrtidae), in Michigan, USA and evaluation of a newly developed monitoring technique. Biological Control 79: 36-42. Abell, K.J., Duan, J.J., Bauer, L., Lelito, J.P., and Van Driesche, R.G. (2012) The effect of bark thickness on host partitioning between Tetrastichus planipennisi (Hymen: Eulophidae) and Atanycolus spp. (Hymen: Braconidae), two parasitoids of emerald ash borer (Coleop: Buprestidae). Biological Control 63: 320-325. Aguiar, A.P., and Santos, B.F. (2010) Discovery of potent, unsuspected sampling disparities for Malaise and Möricke traps, as shown for neotropical Cryptini (Hymenoptera, Ichneumonidae). Journal of Insect Conservation 14: 199-206. Akiyama, K., and Ohmomo, S. (1997) A checklist of the Japanese Buprestidae. Gekkan-Mushi Supplement 1: 1-67. Allison, J.D., and Redak, R.A. (2017) The impact of trap type and design features on survey and detection of bark and woodboring beetles and their associates: a review and meta-analysis. Annual Review of Entomology 62: 127-146. Althoff, D.M. (2003) Does parasitoid attack strategy influence host specificity? A test with New World braconids. Ecological Entomology 28: 500-502. Anderson, R.F. (1944) The relation between host condition and attacks by bronze birch borer. Journal of Economic Entomology 37: 588-596. Anulewicz, A.C., McCullough, D.G., and Cappaert, D.L. (2007) Emerald ash borer (Agrilus planipennis) density and canopy dieback in three North American ash species. Arboriculture and Urban Forestry 33: 338-349. Anulewicz, A.C., McCullough, D.G., Cappaert, D.L., and Poland, T.M. (2008) Host range of the emerald ash borer (Agrilus planipennis Fairmaire) (Coleoptera: Buprestidae) in North America: results of multiple-choice field experiments. Environmental Entomology 37: 230-241. Aubin, I., Cardou, F., Ryall, K., Kreutzweiser, D., and Scarr, T. (2015) Ash regeneration capacity after emerald ash borer (EAB) outbreaks: some early results. The Forestry Chronicle 91: 291-298. Bartelt, R.J., Cossé, A.A., Zilkowski, B.W., Fraser, I. (2007) Antennally active macrolide from the emerald ash borer Agrilus planipennis emitted predominately by females. Journal of Chemical Ecology 33: 1299-1302. Barter, G.W. (1957) Studies of the bronze birch borer, Agrilus anxius Gory, in New Brunswick. The Canadian Entomologist 89: 12-36. Barter, G.W. (1965) Survival and development of the bronze poplar borer Agrilus liragus Barter and Brown (Coleoptera: Buprestidae). The Canadian Entomologist 97: 1063-1068.

108 Bauer, L.S., Duan, J.J., Gould, J.R., Van Driesche, R. (2015) Progress in the classical biological control of Agrilus planipennis Fairmaire (Coleoptera: Buprestidae) in North America. The Canadian Entomologist 147: 300-317. Bauer, L.S., Liu, H., Haack, R.A., Petrice, T.R., and Miller, D.L. (2004) Natural enemies of emerald ash borer in southeastern Michigan. In Emerald ash borer research and technology development meeting, 30 September-1 October 2004, Port Huron, Michigan, United States of America. Compiled by V. Mastro, R. Reardon. United States Department of Agriculture, Morgantown, West Virginia, United States of America. Pp 33-34. Beenakkers, A.M.T., van der Horst, D.J., and van Marrewijk, W.J.A. (1984) Insect flight muscle metabolism. Insect Biochemistry 14: 243-260. Boggs, C.L. (1986) Reproductive strategies of female butter flies: variation in and constraints on fecundity. Environmental Entomolology 11: 7-15. Britton, W.E. (1923) 22nd report of the state entomologist for 1922. Conn. Agr. Sta. Bull. 247: 269-381. Bruzzone, O.A., Villacide, J.M., Bernstein, C., and Corley, J.C. (2009) Flight variability in the woodwasp Sirex noctilio (Hymenoptera: Siricidae): an analysis of flight data using wavelets. Journal of Experimental Biology 212: 731–737. Callaway, R.M., and Ridenour, W.M. (2004) Novel weapons: invasive success and the evolution of increased competitive ability. Frontiers in Ecology and the Environment 2: 436–443. Cappaert, D., and McCullough, D.G. (2009) Occurrence and seasonal abundance of Atanycolus cappaerti (Hymenoptera: Braconidae) a native parastioid of emerald ash borer, Agrilus planipennisi (Coleoptera: Buprestidae). Great Lakes Entomologist 51: 152-165. Cappaert, D., McCullough, D.G., Poland, T.M., and Siegert, N.W. (2005) Emerald ash borer in North America: a research and regulatory challenge. American Entomologist 51: 152-165. Careless, P.D., Marshall, S.A., Gill, B.D. (2014) The use of Cerceris fumipennis (Hymenoptera: Crabronidae) for surveying and monitoring emerald ash borer (Coleoptera: Buprestidae) infestations in eastern North America. The Canadian Entomologist 146: 90-105. Clark, W.C. (1976) The environment and the genotype in polymorphism. Zoological Journal of the Linnaen Society 58: 255-262. Charnov, E.L. (1976) Optimal foraging: the marginal value theorem. Theoretical Population Biology 9: 129-136. Chen, Y., Ulyshen, M.D., and Poland, T.M. (2016) Abundance of volatile organic compounds in white ash phloem and emerald ash borer larval frass does not attract Tetrastichus planipennisi in a Y-tube olfactometer. Insect Science 23: 712-719. Cipollini, D. (2015) White fringetree as a novel larval host for emerald ash borer. Journal of Economic Entomology 1-6, doi: 10.1093/jee/tou026. Cipollini, D., Rigsby, C.M., and Peterson, D.L. (2017) Feeding and development of emerald ash borer (Coleoptera: Buprestidae) on cultivated olive, Olea europaea. Journal of Economic Entomology 110: 1935-1937.

109 Colautti, R.I., and MacIsaac, H.J. (2004) A neutral terminology to define 'invasive' species. Diversity and Distributions 10: 135-141. Corbett, L.J., Withey, P., Lantz, V.A., and Ochuodho, T.O. (2016) The economic impact of the mountain pine beetle infestation in British Columbia: provincial estimates from a CGE analysis. Forestry 89: 100-105. Crosthwaite, J.C., Sobek, S., Lyons, D.B., Bernards, M.A., and Sinclair, B.J. (2011) The overwintering physiology of the emerald ash borer, Agrilus planipennis Fairmaire (Coleoptera: Buprestidae). Journal of Insect Physiology 57: 166-173. Czerwinski, E.J., Tucker, L., Evans, H.J., Biggs, W.D., and Rowlinson, D.T. (2007) Invasive species in Ontario’s forest. In Forest health conditions in Ontario, 2006. Edited by T. Scarr, A. Hopkin, J. Pollard. Ontario Ministry of Natural Resources and Natural Resources Canada, Sault Ste. Marie, Ontario. Darling, D.C., and Packer, L. (1988) Effectiveness of Malaise traps in collecting Hymenoptera: the influence of trap design, mesh size, and location. The Canadian Entomologist 120: 787-796. Dean, K.M., Vandenberg, J.D., Griggs, M.H., Bauer, L.S., and Fierke, M.K. (2012) Susceptibility of two hymenopteran parasitoids of Agrilus planipennis (Coleoptera: Buprestidae) to the entomopathogenic fungus Beauvaria bassiana (Ascomycota: Hypocreales). Journal of Inverebrate Pathology 109: 303-306. Derocles, S.A., Plantegenest, M., Ait-Ighil, E.E.T., Chaubet, B., Dedryver, C.A., and Le Ralec, A. (2014) Larval hitch-hiking and adult flight are two ways of Aphidiinae parasitoids long- range dispersal. Environmental Entomology 43: 1327-1332. Digweed, S.C. (1998) Mortality of birch leafmining sawflies (Hymenoptera: Tenthredinidae): impacts of natural enemies on introduced pests. Environmental Entomology 27: 1357- 1367. Dingle, H. (1965) The relation between age and flight activity in the milkweed bug, Oncopeltus. Journal of Experimental Biology 42: 269-283. Duan, J.J., Bauer, L.S., Abell, K.J., Ulyshen, M.D., and Van Driesche, R.G. (2015) Population dynamics of an invasive forest insect and associated natural enemies in the aftermath of invasion: implication for biological control. Journal of Applied Ecology 52: 1246-1254. Duan, J.J., Bauer, L.S., Abell, K.J., and Van Driesche, R. (2012) Population responses of hymenopteran parasitoids to the emerald ash borer (Coleoptera: Buprestidae) in recently invaded areas in north central United States. BioControl 57: 199-209. Duan, J.J., Bauer, L.S., and Van Driesche, R.G. (2017) Emerald ash borer biocontrol in ash saplings: the potential for early stage recovery of North American ash trees. Forest Ecology and Management 394: 64-72. Duan, J.J., Bauer, L.S., Van Driesche, R.G., and Gould, J.R. (2018) Progress and challenges of protecting North American ash trees from the emerald ash borer using biological control. Forests 1-17, doi: 10.3390/f9030142.

110 Duan, J.J., Fuester, R.W., Wildonger, J., Taylor, P.B., Barth, S., and Spichiger, S.E. (2009) Parasitoids attacking the emerald ash borer (Coleoptera: Buprestidae) in western Pennsylvania. Florida Entomologist 92: 588-592. Duan, J.J., and Schmude, J. (2016) Biology and life history of Atanycolus cappaerti (Hymenoptera: Braconidae), a North American larval parasitoid attacking the invasive emerald ash borer (Coleoptera: Buprestidae). Florida Entomologist 99: 722-728. Duan, J.J., Ulyshen, M.D., Bauer, L.S., Gould, J., and Van Driesche, R. (2010) Measuring the impact of biotic factors on populations of immature emerald ash borers (Coleoptera: Buprestidae). Environmental Entomology 39: 1513-1522. Dunn, J.P., Kimmerer T.W., and Nordin, G.L. (1986) The role of host tree condition in attack of white oaks by the two lined chestnut borer, Agrilus bilineatus (Weber) (Coleoptera: Buprestidae). Oecologia 70: 596-600. Edworthy, A.B., Drever, M.C., and Martin, K. (2011) Woodpeckers increase in abundance but maintain fecundity in response to an outbreak of mountain pine bark beetles. Forest Ecology and Management 261: 203-210. Ellers, J., and Jervis, M.A. (2004) Why are so few parasitoid wasp species pro-ovigenic? Evolutionary Ecology Research 6: 993-1002. Elton, C.S. (1958) The ecology of invasions by animals and plants. Methuen, London, 181 pp. Fahrner, S.J., Lelito, J.P., Blaedow, K., Heimpel, G.E., and Aukema, B.H. (2014) Factors affecting the flight capacity of Tetrastichus planipennisi (Hymenoptera: Eulophidae), a classical biological control agent of Agrilus planipennis (Coleoptera: Buprestidae). Environmental Entomology 43: 1603-1612. Federal Register (2015) Availability of an environmental assessment for field release of the parasitoid Spathius galinae for the biological control of the emerald ash borer (Agrilus planipennis) in the contiguous United States. Fed. Reg. 80 (19), 7827 [Docket No. APHIS- 2014-0094]. https://www.regulations.gov/docket?D=APHIS-2014-0094 (accessed 15 August 2018). Flaherty, L., Sweeney, J.D., Pureswaran, D., and Quiring, D.T. (2011) Influence of host tree condition on the performance of Tetropium fuscum (Coleoptera: Cerambycidae). Environmental Entomology 40: 1200-1209. Forsse, E., Smith, S.M., and Bourchier, R.S. (1992) Flight initiation in the egg parasitoid Trichogramma minutum: Effects of ambient temperature, mates, food, and host eggs. Entomologia Experimentalis et Applicata 62: 147-154. Fraser, S.E.M., Beresford, A.E., Peters, J., Redhead, J.W., Welch, A.J., Mayhew, P.J., and Dytham, C. (2009) Effectiveness of vegetation surrogates for parasitoid wasps in reserve selection. Conservation Biology 23: 142-150. Fraser, S.E.M., Dytham, C., and Mayhew, P.J. (2007) Determinants of parasitoid abundance and diversity in woodland habitats. Journal of Applied Ecology 44: 352-361. Gandhi, J.K.J., and Herms, D.A. (2010a) North American arthropods at risk due to widespread Fraxinus mortality caused by the alien emerald ash borer. Biological Invasions 12: 1839– 1846.

111 Gandhi, K.J.K., and Herms, D.A. (2010b) Direct and indirect effects of alien insect herbivores on ecological processes and interactions in forests of eastern North America. Biological Invasions 12: 389-405. Gardner, J., Wright, M.G., Kuhar, T.P., Pitcher, S.A., and Hoffmann, M.P. (2012) Dispersal of Trichogramma ostriniae in field corn. Biocontrol Science and Technology 22: 1221-1223. Gaudon, J.M., Allison, J.D., and Smith, S.M. 2018. Factors affecting the dispersal of a native parasitoid, Phasgonophora sulcata, attacking the emerald ash borer: implications for biological control. BioControl 63: 751-761. Gaudon, J.M., Haavik, L.J., MacQuarrie, C.J.K., Smith, S.M., and Allison JD (2016) Influence of nematode parasitism, body size, temperature, and diel periodicity on the flight capacity of Sirex noctilio F. (Hymenoptera: Siricidae). Journal of Insect Behavior 29: 301-314. Gibson, G.A.P. (2005) The world species of Balcha Walker (Hymenoptera: Chalcidoidea: Eupelmidae), parasitoids of wood-boring beetles. Zootaxa 1033: 1-62. Godfray, H.C.J. (1994) Parasitoids: Behavioural and Evolutionary Ecology. Princeton University Press, Princeton, 473 pp. Gratton, C., and Welter, S.C. (1999) Does “enemy-free space” exist? Experimental host shifts of an herbivorous fly. Ecology 80: 773-785. Haack, R.A., Benjamin, D.M., and Schuh, B.A. (1981) Observations on the biology of Phasgonophora sulcata (Hymenoptera: Chalcididae), a larval parasitoid of the twolined chestnut borer, Agrilus bilineatus (Coleoptera: Buprestidae), in Wisconsin. The Great Lakes Entomologist 14: 42-45. Haack, R.A., Jendek, E., Liu, H., Marchant, K.R., Petrice, T.R., Poland, T.M., and Ye, H. (2002) The emerald ash borer: a new exotic pest in North America. Newsletter of the Michigan Entomological Society 47: 1–5. Haavik, L.J., Allison, J.D., MacQuarrie, C.J.K., Nott, R.W., Ryan, K., de Groot, P., and Turgeon, J.J. (2016) Non-lethal effects of nematode infection on Sirex noctilio and Sirex nigricornis (Hymenoptera: Siricidae). Environmental Entomology 45: 320–327. Hanson, A.A., Venette, R.C., and Lelito, J.P. (2013) Cold tolerance of Chinese emerald ash borer parasitoids: Spathius agrili Yang (Hymenoptera: Braconidae), Tetrastichus planipennisi Yang (Hymenoptera: Eulophidae), and Oobius agrili Zhang and Huang (Hymenoptera: Encyrtidae). Biological Control 67: 516-529. Hare, J.D., and Weseloh, R.M. (2009) Host seeking, by parasitoids. In Encyclopedia of insects (second edition). Edited by V.H. Resh and R.T. Cardé. Pp. 463-466. Harrington, E.A., and Barbosa, P. (1978) Host habitat influences on oviposition by Parasetigena silvestris (R.-D.), a larval parasite of the gypsy moth. Environmental Entomology 7: 466- 468. Harrison, X.A. (2015) A comparison of observation-level random effect and Beta-Binomial models for modelling overdispersion in Binomial data in ecology and evolution. PeerJ 3e1114, doi:10.7717/peerj.1114. Hassell, M.P., and Pacala, S.W. (1990) Heterogeneity and the dynamics of host-parasitoid interactions. Philosophical Transactions of the Royal Society of London B 330: 203-220.

112 Haugen, D.A., and Underdown, M.G. (1990) Release of parasitoids for Sirex noctilio by transporting infested logs. Australian Forestry 53: 266-270. Hébert, C., Cloutier, C., and Régnière, J. (1990) Factors affecting the flight activity of Winthemia fumiferanae (Diptera: Tachinidae). Environmental Entomology 19: 293–302. Heimpel, G.E., and Asplen, M.K. (2011) A 'Goldilocks' hypothesis for dispersal of biological control agents. BioControl 56: 441-450. Herms, D.A., and McCullough, D.G. (2014) Emerald ash borer invasion of North America: history, biology, ecology, impacts, and management. Annual Review of Entomology 59: 13-30. Hilszczajski, J., Gibb, H., Hjältén, J., Atlegrim, O., Johansson, T., Pettersson, R.B., Ball, J.P., and Danell, K. (2005) Parasitoids (Hymenoptera: Ichneumonoidea) of saproxylic beetles are affected by forest successional stage and dead wood characteristics in a boreal spruce forest. Biological Conservation 126: 456-464. Hook, A.W., and Evans, H.E. (1991) Prey and parasites of Cerceris fumipennis (Hymenoptera: Sphecidae) from Central Texas, with description of the larva of Dasymutilla scaevola (Hymenoptera: Mutillidae). Journal of the Kansas Entomological Society 64: 257-264. Hopper, K.R., and Roush, R.T. (1993) Mate finding, dispersal, number released, and the success of biological control introductions. Ecological Entomology 18: 321-331. Hubbes, M. (1999) The American elm and Dutch elm disease. The Forestry Chronicle 75: 265- 273. Huxel, G.R., and Hastings, A. (1999) Habitat loss, fragmentation, and restoration. Restoration Ecology 7: 309-315. Innocent, T.M., Abe, J., West, S.A., and Reece, S.E. (2010) Competition between relatives and the evolution of dispersal in a parasitoid wasp. Journal of Evolutionary Biology 23: 1374- 1385. Jactel, H., Goulard, M., Menassieu, P., and Goujon, G. (2002) Habitat diversity in forest plantations reduces infestations of the pine stem borer Dioryctria sylvestrella. Journal of Applied Ecology 39: 618-628. Jendek, E., and Grebennikov, V.V. (2011) Agrilus (Coleoptera: Buprestidae) of East Asia. Jan Farkac, Prague, Czech Republic. Jennings, D.E., Duan, J.J., Bauer, L.S., Schmude, J.M., Wetherington, M.T., and Shrewsbury, P.M. (2015) Temporal dynamics of woodpecker predation on emerald ash borer (Agrilus planipennis) in northeastern U.S.A. Agricultural and Forest Entomology 1-8, doi: 10.1111/afe.12142. Jennings, D.E., Gould, J.R., Vandenberg, J.D., Duan, J.J., and Shrewsbury, P.M. (2013) Quantifying the impact of woodpecker predation on population dyanmics of the emerald ash borer (Agrilus planipennis). PLoS ONE 8: e83491. Jervis, M.A., Heimpel, G.E., Ferns, P.N., Harvey, J.A., and Kidd, N.A.C. (2001) Life-history strategies in parasitoid wasps: a comparative analysis of ‘ovigeny’. Journal of Animal Ecology 70: 442-458.

113 Jiao, Z.-B., Wan, T., Wen, J.-B., Hu, J.-F., Luo, Y.-Q., Zhang, L.-S., and Fu, L.-J. (2008) Functional response and numerical response of great spotted woodpecker Picoides major on Asian longhorned beetle Anoplophora glabripennis larvae. Acta Zoologica Sinica 54: 1106-1111. Johnson, M.D. (1990) Female size and fecundity in the small carpenter bee, Ceratina calcarata (Robertson) (Hymenoptera: Anthophoridae). Journal of the Kansas Entomological Society 63: 414-419. Johnson, T.D., Lelito, J.P., and Raffa, K.F. (2014) Responses of two parasitoids, the exotic Spathius agrili Yang and the native Spathius floridanus Ashmead, to volatile cues associated with the emerald ash borer, Agrilus planipennis Fairmaire. Biological Control 79: 110-117. Jones, V.P., Smith, T.J., and Naranjo, S. (2010) Insect ecology and behavior: laboratory flight mill studies. http://entomology.tfrec.wsu.edu/VPJ_Lab/Flight-Mill.html. Accessed 11 July 2015. Keane, R.M., and Crawley, M.J. (2002) Exotic plant invasions and the enemy release hypothesis. Trends in Ecology & Evolution 17: 164-170. Kettunen, M., Genovesi, P., Gollasch, S., Pagad, S., Starfinger, U., ten Brink, P., and Shine, C. (2009) Technical support to EU strategy on invasive alien species (IAS) – Assessment of the impacts of IAS in Europe and the EU. Institute for European Environmental Policy, Brussels, 44 pp. Knight, K.S., Brown, J.P., Long, R.P. (2013) Factors affecting the survival of ash (Fraxinus spp.) trees infested by emerald ash borer (Agrilus planipennis). Biological Invasions 15: 371- 383. Ko, J.H. (1969) A list of forest insect pests in Korea. Forest Research Institute, Seoul, South Korea. Kovacic, D.A., Dyer, M.I., and Cringan, A.T. (1985) Understory biomass in ponderosa pine following mountain pine beetle infestation. Forest Ecology and Management 13: 53-67. Kula, R.R., Knight, K.S., Rebbek, J., Bauer, L.S., Cappaert, D.L., and Gandhi, K.J.K. (2010) Leluthia astigma (Ashmead) (Hymenoptera: Braconidae: Doryctinae) as a parasitoid of Agrilus planipennis Fairmaire (Coleoptera: Buprestidae: Agrilinae), with an assessment of host associations for Nearctic species of Leluthia Cameron. Proceedings of the Entomological Society of Washington 112: 246-257. Kurz, W.A., Dymond, C.C., Stinson, G., Rampley, G.J., Neilson, E.T., Carroll, A.L., Ebata, T., Safranyik, L. (2008) Mountain pine beetle and forest carbon feedback to climate change. Nature 452: 987-990. Kyei-Poku, G., and Johny, S. (2013) Evaluation of indigenous Beauveria isolates as potential agents for emerald ash borer management and the development of a diagnostic marker to monitor a post-release isolate. International Organisation for Biological and Integrated Control-West Palaearctic Regional Section Bulletin 90: 199-124. Lambert, M.C., Ung, C.H., and Raulier, F. (2005) Canadian national tree aboveground biomass equations. Canadian Journal of Forest Research 35: 1996-2018.

114 Lassau, S.A., and Hochuli, D.F. (2005) Wasp community responses to habitat complexity in Sydney sandstone forests. Austral Ecology 30: 179-187. Lassau, S.A., and Hochuli, D.F. (2007) Associations between wasp communities and forest structure: do strong local patterns hold across landscapes? Austral Ecology 32: 656-662. Leather, S.R. (1988) Reproductive potential and fecundity in insects: things aren’t as simple as they seem. Oikos 51: 386-389. Levins, R. (1969) Some demographic and genetic consequences of environmental heterogeneity for biological control. Bulletin of the Entomological Society of America 15: 237-240. Liebhold, A.M., Macdonald, W.L., Bergdahl, D., and Mastro, V.C. (1995) Invasion by exotic forest pests: a threat to forest ecosystems. Forest Science, Monograph 30: 1-58. Lindell, C.A., McCullough, D.G., Cappaert, D., Apostolou, N.M., and Roth, M.B. (2008) Factors influencing woodpecker predation on emerald ash borer. The American Midland Naturalist 159: 434-444. Lindgren, B.S. (1983) A multiple funnel trap for scolytid beetles (Coleoptera). The Canadian Entomologist 115: 299-302. Liu, H., Bauer, L.S., Gao, R., Zhao, T., Petrice, T.R., and Haack, R.E. (2003) Exploratory survey for the emerald ash borer, Agrilus planipennis (Coleoptera: Buprestidae), and its natural enemies in China. The Great Lakes Entomologist 36: 191-204. Liu, H., Bauer, L.S., Miller, D.L., Zhao, T., Gao, R., Song, L., Luan, Q., Jin, R., and Gao, C. (2007) Seasonal abundance of Agrilus planipennis (Coleoptera: Buprestidae) and its natural enemies Oobius agrili (Hymenoptera: Encyrtidae) and Tetrastichus planipennisi (Hymenoptera: Eulophidae) in China. Biological Control 42: 61-71. Lundy, B. (1832) The diary of Benhamin Lundy, written during his journey through Upper Canada, January, 1832. (Edited by Fred Landon. Ontario History Society Papers and Records 19: 110-133. Toronto. 1922.) Lyons, D.B. (2010) Biological control of emerald ash borer. In Guiding Principles for Managing the Emerald Ash Borer in Urban Environments. Edited by D.B. Lyons, T.A. Scarr. Natural Resources Canada and Ontario Ministry of Natural Resources, Burlington, Ontario. Pp. 29-34. MacQuarrie, C.J.K., Lyons, D.B., Seehausen, M.L., and Smith, S.M. (2016) A history of biological control in Canadian forests, 1882-2014. The Canadian Entomologist 148(S1): S239-S269. MacQuarrie, C.J.K., Williams, D.J., and Langor, D.W. (2013) Update on the establishment of birch leafminer parasitoids in western Canada. Journal of the Entomological Society of British Columbia 110: 35-37. Marsh, P.M., and Strazanac, J.S. (2009) A taxonomic review of the genus Spathius Nees (Hymenoptera: Braconidae) in North America and comments on the biological control of the emerald ash borer (Coleoptera: Buprestidae). Journal of Hymenoptera Research 18: 80-112.

115 Matthews, R.W., González, J.M., Matthews, J.R., and Deyrup, L.D. (2009) Biology of the parasitoid Melittobia (Hymenoptera: Eulophidae). Annual Review of Entomology 54: 251-266. Maycock, P.F. (1963) The phytosociology of the deciduous forests of extreme southern Ontario. Canadian Journal of Botany 41: 379-438. McKenney, D.W., Pedlar, J.H., Yemshanov, D., Lyons, D.B., Campbell, K.L., and Lawrence, K. (2012) Estimates of the potential cost of emerald ash borer (Agrilus planipennis Fairmaire) in Canadian municipalities. Arboriculture and Urban Forestry 38: 81-91. Mills, N.J., and Heimpel, G.E. (2018) Could increased understanding of foraging behavior help to predict the success of biological control? Current Opinion of Insect Science 27: 26-31. Murphy, T.C., Gould, J.R., Van Driesche, R.G., and Elkinton, J.S. (2018) Interactions between woodpecker attack and parasitism by introduced parasitoids of the emerald ash borer. Biological Control 122: 109-117. Nash, R.W., Duda, E.J., and Gray, N.H. (1951) Studies on extensive dying, regeneration, and management of birch. Maine Forest Service Bulletin 15. Niemelä, P., and Mattson, W.J. (1996) Invasion of North American forests by phytophagous insects. BioScience 46: 741-753. Nisbet, D.J. (2014) Indirect ecological impacts of emerald ash borer and associated ash decline in southern Ontario riparian areas: risks to aquatic communities. MSc thesis, University of Guelph, Canada. Noyes, J.S. (1989) A study of five methods for sampling Hymenoptera (Insecta) in a tropical rainforest, with special reference to the Parasitica. Journal of Natural History 23: 285-298. Ontario Ministry of Natural Resources. (2000) A silvicultural guide to managing southern Ontario Forests. Ontario Ministry of Natural Resources. Queen’s Printer for Ontario, 654 pp. Osborne, J.L., Loxdale, H.D., and Woiwod, I.P. (2002) Monitoring insect dispersal: methods and approachs. In Dispersal ecology. Edited by J.M. Bullock, R.E. Kenward, R.S. Hails. Blackwell Publishing, Oxford. Pp. 24-49. Paiero, S.M., Jackson, M.D., Jewiss-Gaines, A., Gill, B.D., Marshall, S.A. (2012) Field guide to the jewel beetles (Coleoptera: Buprestidae) of northeastern North America. Canadian Food Inspection Agency, 411 pp. Paré, P.W., and Tumlinson, J.H. (1999) Plant volatiles as defense against insect herbivores. Plant Physiology 121: 325-331. Parisio, M.S., Gould, J.R., Vandenberg, J.D., Bauer, L.S., and Fierke, M.K. (2017) Evaluation of recovery and monitoring methods for parasitoids released against emerald ash borer. Biological Control 106: 45-53. Pec, G.J., Karst, J., Sywenky, A.N., Cigan, P.W., Erbilgin, N., Simard, S.W., and Cahill Jr., J.F. (2015) Rapid increases in forest understory diversity and productivity following a mountain pine beetle (Dendroctonus ponderosae) outbreak in pine forests. PLoS ONE 10(4): e0124691.

116 Perez-Mendoza, J., Campbell, J.F., and Throne, J.E. (2011) Influence of age, mating status, sex, quantity of food, and long-term food deprivation on red flour beetle (Coleoptera: Tenebrionidae) flight initiation. Journal of Economic Entomology 104: 2078-2086. Peterson, D.L., and Cipollini, D. (2016) Distribution, predictors, and impacts of emerald ash borer (Agrilus planipennis) (Coleoptera: Buprestidae) infestation of white fringetree (Chionanthus virginicus). Environmental Entomology 46: 50-57. Pettersson, E.M. (2001) Volatiles from potential hosts of Rhopalicus tutela a bark beetle parasitoid. Journal of Chemical Ecology 27: 2219-2231. Pimentel, D., Zuniga, R., and Morrison, D. (2004) Update on the environmental and economic costs associated with alien-invasive species in the United States. Ecological Economics 52: 273-288. Poland, T.M., and McCullough, D.G. (2006) Emerald ash borer: Invasion of the urban forest and the threat to North America’s ash resource. Journal of Forestry 104: 118-124. Poland, T.M., and McCullough, D.G. (2010) SLAM: a multi-agency pilot project to SL.ow A.sh M.ortalitycaused by emerald ash borer in outlier sights. Newsletter of the Michigan Entomological Society 55: 4-8. Puric-Mladenovic, D. (2003) Predictive vegetation modeling for forest conservation and management in settled landscapes. PhD thesis, University of Toronto, Canada. Puric-Mladenovic, D., and Kenney, A. (2016) The VSP field inventory and monitoring pocket guide. University of Toronto, Toronto, Canada. Quick, D.L.J. (2015) The braconid and ichneumonid parasitoid wasps: biology, systematics, evolution and ecology. John Wiley & Sons, Ltd., Chichester, 688 pp. R Development Core Team (2018) R: a language and environment for statistical computing. R Foundation for Statistical Computing, Vienna, Austria. URL http://www.r-project.org Ratte, H.T. (1985) Temperature and insect development. In Environmental physiology and biochemistry of insects. Edited by K.H. Hoffman. Springer, Berlin. Pp. 33-66. Riley, J.L. (1999) Southern Ontario woodlands: the conservation challenge. In Southern Ontario woodlands: the conservation challenge. Federation of Ontario Naturalists, Don Mills. Pp. 9-22. Rojas, J.C., Castillo, A., and Virgen, A. (2006) Chemical cues used in host location by Phymastichus coffea, a parasitoid of coffee berry borer adults, Hypothenemus hampei. Biological Control 37: 141-147. Roland, J., and Embree, D.G. (1995) Biological control of the winter moth. Annual Review of Entomology 40: 475-492. Roscoe, L.E. (2014) Phasgonophora sulcata Westwood (Hymenoptera: Chalcididae): a potential augmentative biological control agent for the invasive Agrilus planipennis (Fairmaire) (Coleoptera: Buprestidae) in Canada. PhD thesis, University of Toronto, Canada.

117 Roscoe, L.E., Lyons, D.B., and Smith, S.M. (2016) Observations on the life-history traits of the North American parasitoid Phasgonophora sulcata Westwood (Hymenoptera: Chalcididae) attacking Agrilus planipennis (Coleoptera: Buprestidae) in Ontario, Canada. The Canadian Entomologist 148: 294-306. Rosenheim, J.A., and Rosen, D. (1991) Foraging and oviposition decisions in the parasitoid Aphytis lingnanensis: distinguishing the influences of egg load and experience. Journal of Animal Ecology 60: 873-894. Rowley, W.A., and Graham, C.L. (1968) The effect of age on the flight performance of female Aedes aegypti mosquitoes. Journal of Insect Physiology 14: 719-728. Rutledge, C.E., Fierke, M.K., Careless, P.D., and Worthley, T. (2013) First detection of Agrilus planipennis in Connecticut made by monitoring Cerceris fumipennis (Crabonidae) colonies. Journal of Hymenoptera Research 32: 75-81. Ryall, K.L., Fidgen, J.G., and Turgeon, J.J. 2011. Detectability of the emerald ash borer (Coleoptera: Buprestidae) in asymptomatic urban trees by using branch samples. Environmental Entomology 40: 679-688. Salt, G. (1941) The effects of hosts upon their insect parasites. Biological Reviews 16: 239-264. Shahjahan, M. (1974) Erigeron flowers as a food and attractive odor source for Peristenus pseudopallipes, a braconid parasitoid of the tarnished plant bug. Environmental Entomology 3: 69-72. Sharov, A.A., Leonard, D., Liebhold, A.M., Roberts, E.A., and Dickerson, W. (2002) “Slow the spread”: a national program to contain the gypsy moth. Journal of Forestry 100: 30-35. Shaw, M.R., and Hochberg, M.E. (2001) The neglect of parasitic Hymenoptera in insect conservation strategies: the British fauna as a prime example. Journal of Insect Conservation 5: 253-263. Shea, K., and Possingham, H.P. (2000) Optimal release strategies for biological control agents: an application of stochastic dynamic programming to population management. Journal of Applied Ecology 37: 77-86. Schowalter, T.D. (1985) Adaptations of insects to disturbance. In The ecology of natural disturbance and patch dynamics. Edited by S.T.A. Pickett, P.S. White. Academic Press, Orlando, Florida. Pp. 235-252. Schowalter, T.D. (2000) Insect ecology: an ecosystem approach. Academic Press, San Diego, 483 pp. Siegert, N.W., McCullough, D.G., Liebhold, A.M., and Telewski, F.W. (2007) Resurrected from the ashes: a historical reconstruction of emerald ash borer dynamics through dendrochronological analyses. Proc. 2006 Emerald Ash Borer Res. Technol. Dev. Meet., ed. V. Mastro, D. Lance, R. Reardon, G. Parra, pp. 18–19. FHTET-2007-04. Morgantown, WV: USDA For. Serv. Siegert, N.W., McCullough, D.G., Liebhold, A.M. and Telewski, F.W. (2014). Dendrochronological reconstructions of the epicentre and early spread of emerald ash borer in North America. Diversity and Distributions 20: 847-858.

118 Simberloff, D., and Von Holle, B. (1999) Positive interactions of nonindigenous species: invasional meltdown? Biological Invasions 1: 21–32. Simmons, G.A., Leonard, D.E., and Chen, C.W. (1975) Influence of tree species density and composition on parasitism of the spruce budworm, Choristoneura fumiferana (Clem.). Environmental Entomology 4: 832-836. Smith, A. (2006) Effects of community structure on forest susceptibility and response to the emerald ash borer invasion of the Huron River watershed in southeast Michigan. M.Sc. thesis. The Ohio State University, Columbus, USA. Smith, A., Herms, D.A., Long, R.P., and Gandhi, K.J.K. (2015) Community composition and structure had no effect on forest susceptibility to invasion by the emerald ash borer (Coleoptera: Buprestidae). The Canadian Entomologist 147: 318-328. Smith, S.M. (1988) Pattern of attack on spruce budworm egg masses by Trichogramma minutum (Hymenoptera: Trichogrammatidae) released in forest stands. Environmental Entomology 17: 1009-1015. Smith, S.M. (1996) Biological control with Trichogramma: advances, successes, and potential of their use. Annual Review of Entomology 41: 375-406. Smith, S.M., Wallace, D.R., Howse, G., and Meating, J. (1990) 3.5 Suppresion of spruce budworm populations by Trichogramma minutum Riley, 1982-1986. The Memoirs of the Entomological Society of Canada 122: 56-81. Smitley, D.R., Doccola, J.J., Cox, D.L. (2010) Multiple-year protection of ash trees from emerald ash borer with a single trunk injection of emamectin benzoate, and single-year protection with an imidacloprid basal drench. Arboriculture and Urban Forestry 36: 206-211. Sobek-Swant, S., Crosthwaite, J.C., Lyons, D.B., and Sinclair, B.J. (2012) Could phenotypic plasticity limit an invasive species? Incomplete reversibility of mid-winter deacclimation in emerald ash borer. Biological Invasions 14: 115-125. Sperber, C.F., Nakayama, K., Valverde, M.J., and Neves, F.S. (2004) Tree species richness and diversity affect parasitoid diversity in cacao agroforestry. Basic and Applied Ecology 5: 241-251. Tannis, S.R., and McCullough, D.G. (2012) Differential persistence of blue ash and white ash following emerald ash borer invasion. Canadian Journal of Forest Research 42: 1541- 1550. Taylor, L.R. (1963) Analysis of the effect of temperature on insects in flight. Journal of Animal Ecology 32: 99-117. Taylor, R.A.J., Bauer, L.S., Poland, T.M., and Windell, K.N. (2010) Flight performance of Agrilus planipennis (Coleoptera: Buprestidae) on a flight mill and in free flight. Journal of Insect of Behavior 23: 128-148. Tentelier, C., Desouhant, E., and Fauvergue, X. (2006) Habitat assessment by parasitoids: mechanisms for patch use behavior. Behavioral Ecology 17: 515-521. Timms, L.L., Smith, S.M., and de Groot, P. (2006) Patterns in the within-tree distribution of the emerald ash borer Agrilus planipennis (Fairmaire) in young, green-ash plantations of south-western Ontario, Canada. Agricultural and Forest Entomology 8: 313-321.

119 Tluczek, A.R., McCullough, D.G., and Poland, T.M. (2011) Influence of host stress on emerald ash borer (Coleoptera: Buprestidae) adult density, development, and distribution in Fraxinus pennsylvanica trees. Environmental Entomology 40: 357-366. Tobin, P.C., and Blackburn L.M., eds. (2007) Slow the Spread: a national program to manage the gypsy moth. Gen. Tech. Rep. NRS-6. Newtown Square, PA: U.S. Department of Agriculture, Forest Service, Northern Research Station. 109 pp. Ulyshen, M.D., and Sheehan, T.N. (2017) Trap height considerations for detecting two economically important forest beetle guilds in southeastern US forests. Journal of Pest Science 1-13, doi:10.1007/s10340-017-0883-7. Vance, C.C., Smith, S.M., Malcolm, J.R., Huber, J., and Bellocq, M.I. (2007) Differences between forest type and vertical strata in the diversity and composition of hymenopteran families and mymarid genera in northeastern temperate forests. Environmental Entomology 36: 1073-1083. Van Driesche, R.G., and Bellows, T.S. (1996) Biological control. Chapman and Hall, New York, 539 pp. Venkateswaran, V., Shrivastava, A., Kumble, A.L.K., and Borges, R.M. (2017) Life-history strategy, resource dispersion and phylogenetic associations shape dispersal of a fig wasp community. Movement Ecology 1-11, doi:10.1186/s40462-017-0117-x. Waage, J.K. (1979) Foraging for patchily-distributed hosts by the parasitoid, Nemeritis canescens. Journal of Animal Ecology 48: 353-371. Waage, J.K. (1983) Aggregation in field parasitoid populations: foraging time allocation by a population of Diadegma (Hymenoptera: Ichneumonidae). Ecological Entomology 8: 447- 453. Waage, J.K., and Ng, S.M. (1984) The reproductive strategy of a parasitic wasp. I. Optimal progeny allocation in Trichogramma evanescens. Journal of Animal Ecology 53: 401-415. Wäckers, F.L. (2004) Assessing the suitability of flowering herbs as parasitoid food sources: flower attractiveness and nectar accessibility. Biological Control 29: 307-314. Wäckers, F.L. (2005) Suitability of (extra-) floral nectar, pollen and honeydew as insect food sources. Plant-Provided Food for Carnivorous Insects: A Protective Mutualism and its Applications (ed. by Wäckers, F.L., van Rijn, P.C.J. and Bruin, J.), pp. 17-74. Cambridge University Press, Cambridge, UK. Wanner, H., Gu, H., and Dorn, S. (2006) Nutritional value of floral nectar sources for flight in the parasitoid wasp, Cotesia glomerata. Physiological Entomology 31: 127-133. Wellington, W.G. (1980) Dispersal and population change. In Dispersal of forest insects: evaluation, theory, and management. Edited by A.A. Berryman, L. Safranyik. Proceedings of the International Union of Forest Research Organizations Conference, Washington State University Cooperative Extension Service, Pullman, Washington. Pp. 11-24. Weseloh, R.M. (1972) Spatial distribution of gypsy moth (Lepidoptera: Lymantriidae) and some of its parasitoids within a forest environment. Entomophaga 17: 339-351.

120 Weseloh, R.M. (1979) Comparative behavioral responses to three Brachymeria species and other gypsy moth parasitoids to humidty and temperature. Environmental Entomology 8: 670- 675. West, R.J., Kenis, M., Bouchier, R.S., Smith, S.M., and Butt, G.W. (2002) Zeiraphera canadensis Mutuura and Freeman, spruce bud moth (Lepidoptera: Tortricidae). In Biological control programmes in Canada, 1981-2000. Edited by P.G. Mason and J.T. Huber. CABI Publishing, Wallingford, United Kingdom. Pp. 279-283. Wetherington, M.T., Jennings, D.E., Shrewsbury, P.M., and Duan, J.J. (2017) Climate variation alters the synchrony of host-parasitoid interactions. Ecology and Evolution 7: 8578-8587. Willow, A.J. (2011) Indigenizing invasive species management: native North Americans and the emerald ash borer (EAB) beetle. Culture, Agriculture, Food and Environment 33: 70-82. Wiman, N.G., Walton, V.M., Shearer, P.W., Rondon, S.I., and Lee, J.C. (2014) Factors affecting flight capacity of brown marmorated stink bug, Halyomorpha halys (Hemiptera: Pentatomidae). Journal of Pest Science 88: 37–47. With, K.A. (2002) The landscape ecology of invasive spread. Conservation Biology 16: 1192- 1203. Yu, H., Zhang, Y., Wu, K., Wyckhuys, K.A.G., and Guo, Y. (2009) Flight potential of Microplitis mediator, a parasitoid of various lepidopteran pests. BioControl 54: 183-193.

121 Appendix A (to Chapter 3) Total number of trees (N), relative tree species abundance, and corresponding mean diameter at breast height (DBH) ± 1 standard deviation measurements within all 30 plots across the study area in southern Ontario. Standing dead trees and upright coarse woody debris are classified as snags.

± 1 Species Relative Mean DBH VSPPLOTid N standard name/description abundance (cm) deviation 1 78 Acer saccharum 0.58 9.42 4.62 Ostrya virginiana 0.10 11.51 5.76 Populus sp. 0.09 31.40 4.43 Snag 0.05 17.18 13.85 Thuja occidentalis 0.05 14.23 2.50 Betula papyrifera 0.04 20.37 3.80 Fraxinus sp. 0.04 19.00 1.31 Prunus serotina 0.04 10.83 1.01 Pinus strobus 0.01 30.80 - 2 30 Thuja occidentalis 0.40 26.75 13.78 Cornus alternifolia 0.27 9.78 5.86 Snag 0.17 7.00 0.95 Fraxinus sp. 0.13 11.10 6.38 Prunus serotina 0.03 48.30 - 3 60 Populus sp. 0.32 25.18 5.63 Cornus alternifolia 0.28 7.59 3.10 Thuja occidentalis 0.13 10.18 4.96 Snag 0.10 6.82 1.30 Crataegus punctata 0.07 11.23 3.08 Fraxinus sp. 0.07 9.23 3.43 Betula papyrifera 0.02 9.90 - Pinus strobus 0.02 7.60 - 4 42 Acer saccharum 0.52 11.11 12.63 Fraxinus sp. 0.24 27.83 12.59 Snag 0.12 9.18 3.25 Betula papyrifera 0.07 15.47 3.15 Tilia americana 0.02 28.80 - Ostrya virginiana 0.02 5.20 - 5 39 Pinus strobus 0.26 14.38 4.33 Picea abies 0.26 13.09 6.12 Populus tremuloides 0.21 39.30 13.70 Fraxinus 0.15 35.98 20.48 pennsylvanica Betula papyrifera 0.10 26.13 3.98

122

Carya cordiformis 0.03 23.50 - 6 34 Juglans nigra 0.24 34.75 17.59 Crataegus punctata 0.21 15.39 8.86 Fraxinus americana 0.21 12.07 11.21 Acer negundo 0.15 10.30 3.92 Rhamnus sp. 0.15 6.82 2.34 Snag 0.06 15.65 5.59 7 71 Rhamnus sp. 0.56 6.56 1.43 Fraxinus americana 0.21 7.61 2.17 Carya ovata 0.18 19.16 15.84 Snag 0.04 7.87 1.97 8 51 Acer saccharum 0.37 8.26 3.08 Snag 0.27 18.45 15.86 Fraxinus americana 0.24 19.68 15.38 Crataegus punctata 0.04 11.55 2.90 Acer rubrum 0.04 8.00 3.82 Rhamnus sp. 0.02 8.60 - Quercus rubra 0.02 5.20 - 9 38 Fraxinus americana 0.29 12.05 3.70 Acer saccharum 0.24 10.43 4.64 Amelanchier laevis 0.13 5.42 0.48 Snag 0.08 7.60 3.44 Rhus typhina 0.05 9.10 3.11 Acer saccharinum 0.05 6.50 0.28 Acer rubrum 0.05 5.50 0.28 Populus deltoides 0.03 55.20 - Picea pungens 0.03 12.70 - Salix sp. 0.03 11.20 - Rhamnus sp. 0.03 5.30 - 10 55 Tilia americana 0.25 11.36 4.51 Fagus grandifolia 0.20 8.82 2.76 Ostrya virginiana 0.20 7.80 1.85 Fraxinus americana 0.18 13.93 4.71 Ulmus americana 0.09 7.44 1.66 Rhamnus cathartica 0.04 5.05 0.07 Prunus serotina 0.02 14.50 - Juglans cinerea 0.02 8.50 - 11 42 Rhamnus cathartica 0.40 7.78 1.78 Fraxinus 0.17 29.31 14.05 pennsylvanica Prunus serotina 0.12 10.44 4.72 Acer rubrum 0.10 15.85 5.83

123

Ulmus americana 0.10 10.18 4.22 Populus tremuloides 0.05 35.80 3.11 Amelanchier sp. 0.02 27.00 - Quercus rubra 0.02 12.00 - Prunus virginiana 0.02 6.00 - 12 34 Acer saccharum 0.62 12.71 12.89 Fraxinus americana 0.26 14.56 17.05 Juglans cinerea 0.03 10.00 - Ostrya virginiana 0.03 9.50 - Fagus grandifolia 0.03 8.00 - Rhamnus cathartica 0.03 6.00 - 13 52 Fraxinus americana 0.29 15.44 10.30 Populus tremuloides 0.17 25.08 4.62 Carpinus caroliniana 0.15 6.85 1.65 Rhamnus cathartica 0.08 6.85 3.11 Fraxinus nigra 0.06 8.47 1.65 Quercus rubra 0.04 44.50 11.46 Acer rubrum 0.04 16.55 0.92 Thuja occidentalis 0.04 6.65 1.91 Ulmus americana 0.04 6.50 1.84 Picea glauca 0.02 19.50 - Betula papyrifera 0.02 8.60 - Prunus serotina 0.02 7.80 - Acer saccharum 0.02 7.60 - Rhamnus frangula 0.02 5.10 - 14 70 Acer saccharum 0.23 7.42 2.08 Populus sp. 0.19 36.89 7.94 Prunus serotina 0.13 8.82 3.23 Betula papyrifera 0.10 20.16 4.64 Snag 0.10 12.50 9.55 Fraxinus sp. 0.07 15.28 4.37 Fraxinus americana 0.06 6.28 0.74 Fagus grandifolia 0.04 7.67 1.50 Pinus strobus 0.03 11.55 7.99 Ostrya virginiana 0.03 8.40 0.14 Quercus rubra 0.03 6.35 0.78 15 45 Acer saccharum 0.27 8.53 6.92 Fraxinus sp. 0.18 14.19 4.57 Populus sp. 0.16 20.70 9.18 Ostrya virginiana 0.13 10.42 3.89 Snag 0.07 23.63 17.63 Prunus serotina 0.07 16.00 16.04

124

Fraxinus americana 0.07 9.40 6.85 Betula papyrifera 0.04 33.45 9.55 Quercus rubra 0.02 10.80 - 16 64 Acer saccharum 0.56 9.77 4.29 Fraxinus sp. 0.17 21.97 9.01 Snag 0.14 6.06 1.50 Carya cordiformis 0.05 26.77 14.94 Betula papyrifera 0.05 16.93 0.51 Prunus serotina 0.02 25.90 - Ostrya virginiana 0.02 9.00 - 17 53 Fraxinus sp. 0.28 15.22 7.36 Acer saccharum 0.21 9.54 6.83 Snag 0.15 18.29 18.24 Ostrya virginiana 0.15 11.01 4.24 Pinus strobus 0.09 30.42 44.09 Carya cordiformis 0.08 22.35 5.27 Populus sp. 0.02 42.50 - Quercus rubra 0.02 32.80 - 18 68 Fraxinus sp. 0.29 24.04 9.82 Acer saccharum 0.18 7.02 1.93 Ostrya virginiana 0.15 7.98 1.00 Pinus strobus 0.10 12.57 4.21 Snag 0.10 12.20 4.65 Carya cordiformis 0.09 16.77 7.36 Tilia americana 0.03 18.70 17.11 Populus sp. 0.01 27.40 - Quercus rubra 0.01 22.20 - Betula papyrifera 0.01 18.30 - Fagus grandifolia 0.01 5.00 - 19 68 Fraxinus sp. 0.26 17.77 9.74 Snag 0.21 13.43 7.12 Ostrya virginiana 0.15 11.16 3.47 Betula papyrifera 0.12 20.19 6.87 Populus sp. 0.10 32.13 2.84 Prunus serotina 0.04 15.80 9.16 Fraxinus americana 0.03 12.35 5.87 Quercus rubra 0.03 10.60 3.25 Acer saccharum 0.03 8.05 0.92 Pinus strobus 0.01 25.10 - Cornus alternifolia 0.01 6.80 - 20 35 Acer saccharum 0.34 8.03 2.24 Snag 0.20 24.27 8.64

125

Fraxinus sp. 0.14 21.18 8.05 Acer rubrum 0.14 17.40 8.16 Crataegus punctata 0.06 14.00 2.40 Tilia americana 0.06 10.05 3.75 Pinus strobus 0.03 20.80 - Ulmus americana 0.03 13.20 - 21 27 Quercus rubra 0.37 11.93 3.35 Acer rubrum 0.22 6.07 0.66 Fraxinus sp. 0.15 15.40 4.25 Acer platanoides 0.07 13.15 1.20 Rhamnus sp. 0.07 6.40 0.28 Populus sp. 0.04 12.90 - Snag 0.04 7.80 - Fraxinus 0.04 5.80 - pennsylvanica 22 28 Acer saccharum 0.32 9.64 1.82 Fraxinus sp. 0.21 10.90 3.07 Snag 0.14 7.55 2.24 Rhamnus sp. 0.11 5.53 0.61 Cornus sp. 0.11 5.50 0.44 Populus sp. 0.04 19.80 - Acer saccharinum 0.04 12.90 - Rhus typhina 0.04 9.70 - 23 29 Acer saccharum 0.41 8.03 2.56 Carya ovata 0.24 20.90 7.59 Quercus rubra 0.17 31.36 9.99 Quercus macrocarpa 0.03 45.80 - Snag 0.03 39.70 - Fraxinus sp. 0.03 24.40 - Ostrya virginiana 0.03 7.60 - Rhamnus sp. 0.03 5.00 - 24 64 Snag 0.31 10.85 3.65 Carya ovata 0.14 14.91 7.44 Ostrya virginiana 0.13 7.61 1.81 Fraxinus sp. 0.11 12.61 5.31 Rhamnus sp. 0.11 6.57 1.48 Carya cordiformis 0.05 13.37 6.55 Quercus macrocarpa 0.05 9.80 3.34 Juglans nigra 0.03 17.50 13.72 Ulmus americana 0.03 13.15 9.83 Tilia americana 0.02 13.00 - Acer saccharum 0.02 12.30 -

126

Quercus rubra 0.02 5.80 - 25 67 Fraxinus americana 0.54 8.95 3.95 Juglans nigra 0.15 14.12 6.07 Snag 0.13 6.99 1.93 Populus sp. 0.09 20.77 9.70 Fraxinus sp. 0.06 10.35 1.94 Rhamnus sp. 0.03 5.35 0.07 26 53 Fraxinus americana 0.43 10.34 4.40 Rhamnus sp. 0.25 6.36 0.95 Fraxinus sp. 0.09 14.48 6.45 Snag 0.08 6.20 1.24 Cornus sp. 0.08 5.43 0.42 Thuja occidentalis 0.04 14.50 9.90 Ulmus americana 0.02 26.00 - Tilia americana 0.02 10.20 - 27 29 Pinus strobus 0.21 36.53 17.44 Fraxinus sp. 0.17 14.20 1.97 Betula papyrifera 0.14 22.30 2.94 Fraxinus americana 0.14 12.18 11.29 Prunus serotina 0.14 7.58 2.17 Acer platanoides 0.10 10.83 2.71 Acer negundo 0.03 17.70 - Rhamnus sp. 0.03 7.00 - Snag 0.03 5.60 - 28 50 Fraxinus sp. 0.24 18.98 7.36 Fagus grandifolia 0.22 11.83 2.18 Snag 0.18 13.19 3.98 Prunus serotina 0.14 11.94 6.05 Ostrya virginiana 0.10 24.48 8.26 Acer saccharum 0.06 11.27 0.93 Fraxinus americana 0.04 10.85 0.49 Quercus rubra 0.02 5.00 - 29 42 Fraxinus americana 0.19 13.75 14.39 Populus tremuloides 0.17 27.56 3.48 Snag 0.17 10.40 4.39 Acer rubrum 0.14 25.17 15.14 Fraxinus sp. 0.10 19.23 6.75 Quercus rubra 0.05 49.30 7.35 Tsuga canadensis 0.05 21.95 15.91 Thuja occidentalis 0.05 10.50 2.26 Rhamnus cathartica 0.05 8.00 3.82 Populus sp. 0.02 22.20 -

127

Ulmus americana 0.02 8.10 - 30 33 Fraxinus 1.00 22.64 12.44 pennsylvanica

128 Appendix B (to Chapter 3) Ground vegetation within all 30 plots and 150 subplots across the study area in southern Ontario.

VSPPLOTid VSPSUBPLOTid Species name/description # of stems 1 Center Maianthemum canadense 60 Trillium grandiflorum 31 North Maianthemum canadense 3 Podophyllum peltatum 2 East Maianthemum canadense 38 Trillium grandiflorum 7 South Trillium grandiflorum 2 Podophyllum peltatum 1 West Maianthemum canadense 13 Trillium grandiflorum 13 Podophyllum peltatum 2 2 Center Geranium robertianum 11 Circaea lutetiana 9 Alliaria petiolata 2 Podophyllum peltatum 1 North - 0 East Parthenocissus quinquefolia 12 Circaea lutetiana 9 Geranium robertianum 3 Ribes cynosbati 3 Solidago altissima 1 South Maianthemum canadense 47 Podophyllum peltatum 7 Alliaria petiolata 5 Geranium robertianum 3 Parthenocissus quinquefolia 3 Circaea lutetiana 1 West Circaea lutetiana 15 Alliaria petiolata 7 Fragaria sp. 3 Podophyllum peltatum 2 Symphytotricum lateriflorum 2 Geranium robertianum 1 Ranunculus acris 1 3 Center Parthenocissus quinquefolia 10 Podophyllum peltatum 1 North Maianthemum canadense 66 Parthenocissus quinquefolia 2 East Parthenocissus quinquefolia 14

129

Circaea lutetiana 8 Podophyllum peltatum 5 South Trail 0 West Circaea lutetiana 7 Geum urbanum 4 4 Center Trillium grandiflorum 12 Solidago flexicaulis 6 Hydrophyllum virginianum 5 Asarum canadense 2 North Ribes cynosbati 5 Asarum canadense 4 Circaea lutetiana 3 Trillium grandiflorum 2 Hydrophyllum virginianum 1 East Caulophyllum thalictroides 12 Asarum canadense 11 Trillium grandiflorum 11 South Trillium grandiflorum 23 Hydrophyllum virginianum 13 West Trillium grandiflorum 36 Actaea pachypoda 1 5 Center Alliaria petiolata 8 Maianthemum canadense 4 Circaea lutetiana 2 Galium triflorum 2 Geranium robertianum 1 North Trillium grandiflorum 2 Circaea lutetiana 1 East - 0 South Trillium grandiflorum 9 Ranunculus acris 3 Actaea pachypoda 1 Hieracium caespitosum 1 Maianthemum canadense 1 West Trillium grandiflorum 9 Alliaria petiolata 4 Circaea lutetiana 3 Actaea pachypoda 1 Geranium robertianum 1 6 Center Alliaria petiolata 198 Galium aparine 8 Unidentified sp. 1 North Alliaria petiolata 53 Impatiens capensis 9

130

Geum urbanum 2 East Solidago altissima 20 Alliaria petiolata 19 Impatiens capensis 6 Hesperis matronalis 5 South Galium aparine 9 Alliaria petiolata 5 Geum urbanum 3 Hesperis matronalis 2 Impatiens capensis 2 West Alliaria petiolata 91 Geum urbanum 13 7 Center - 0 North Impatiens capensis 6 Geum urbanum 1 East - 0 South - 0 West Impatiens capensis 3 8 Center Impatiens capensis 37 Veronica officinalis 19 Alliaria petiolata 7 Geum urbanum 7 Circaea lutetiana 1 North Impatiens capensis 81 Alliaria petiolata 30 Podophyllum peltatum 5 Maianthemum racemosum 2 Trillium grandiflorum 2 Geranium robertianum 1 Parthenocissus quinquefolia 1 East Impatiens capensis 73 Trillium grandiflorum 4 Parthenocissus quinquefolia 3 South Impatiens capensis 18 Podophyllum peltatum 4 Geum urbanum 1 West Impatiens capensis 48 Trillium grandiflorum 9 Alliaria petiolata 5 Geum urbanum 4 9 Center Geranium robertianum 1 North Solidago altissima 14 East Galium aparine 1 Solidago altissima 1

131

South Water 0 West Solidago altissima 6 Geum urbanum 5 Alliaria petiolata 1 10 Center Alliaria petiolata 12 Geum urbanum 5 Circaea lutetiana 4 Caulophyllum thalictroides 1 Rubus idaeaus 1 Trillium grandiflorum 1 North Circaea lutetiana 9 Asarum canadense 6 Geum urbanum 4 Maianthemum racemosum 2 East Circaea lutetiana 12 Geum urbanum 9 Latuca sp. 4 Chelidonium majus 1 Maianthemum racemosum 1 Trillium grandiflorum 1 South Maianthemum canadense 109 Geum urbanum 12 Circaea lutetiana 4 West Geum urbanum 24 Veronica officinalis 14 Circaea lutetiana 12 Asarum canadense 5 11 Center Fragaria sp. 17 Viola sororia 3 Trillium grandiflorum 2 North Geranium robertianum 3 Trillium grandiflorum 1 Veronica officinalis 1 East Viola sororia 5 Fragaria sp. 3 South Viola sororia 5 Geum canadense 2 Fragaria sp. 1 West Trillium grandiflorum 4 Veronica officinalis 2 Viola sororia 1 12 Center Maianthemum racemosum 1 Trillium grandiflorum 1 North Trillium grandiflorum 14

132

Circaea lutetiana 2 Alliaria petiolata 1 Asarum canadense 1 East Trillium grandiflorum 6 Maianthemum racemosum 5 South Trillium grandiflorum 2 West Maianthemum racemosum 13 Maianthemum canadense 6 Trillium grandiflorum 2 Actaea pachypoda 1 Caulophyllum thalictroides 1 13 Center - 0 North - 0 East Epipactus helleborine 1 South Trillium grandiflorum 1 West - 0 14 Center Maianthemum canadense 22 North - 0 East Maianthemum canadense 32 South - 0 West Maianthemum canadense 5 Actaea pachypoda 1 15 Center - 0 North - 0 East Streptopus lanceolatus 3 Trillium 1 South Maianthemum canadense 26 Podophyllum peltatum 3 West Desmodium glutinosum 3 Trillium 1 16 Center Trillium 9 Anemone acutiloba 1 North Unidentifiable species 1 East Caulophyllum thalictroides 2 Trillium 2 South Trillium 6 Polygonatum biflorum 1 West Circaea lutetiana 2 17 Center Maianthemum canadense 2 North Anemone acutiloba 3 Trillium 1 East Caulophyllum thalictroides 1 South Anemone acutiloba 4 Maianthemum canadense 3

133

Actaea pachypoda 1 Trillium 1 West Caulophyllum thalictroides 8 Actaea pachypoda 2 Anemone acutiloba 1 18 Center Arisaema triphyllum 3 North Maianthemum canadense 3 Actaea pachypoda 2 Viola sororia 1 East Maianthemum canadense 27 Circaea lutetiana 5 South Maianthemum canadense 35 Impatiens capensis 14 Viola sororia 3 Geranium robertianum 1 West - 0 19 Center Actaea pachypoda 4 North Maianthemum canadense 6 Actaea pachypoda 1 East Caulophyllum thalictroides 3 Trillium 1 South Circaea lutetiana 5 Actaea pachypoda 2 Maianthemum racemosum 1 West Polygonatum biflorum 2 Trillium 2 20 Center Impatiens capensis 32 Alliaria petiolata 29 Geum urbanum 2 North Impatiens capensis 4 Alliaria petiolata 3 Trillium 2 East Circaea lutetiana 5 Impatiens capensis 5 Trillium 3 Alliaria petiolata 2 Geranium maculatum 1 South Impatiens capensis 132 Trillium 3 West Impatiens capensis 74 Alliaria petiolata 6 Geum urbanum 2 21 Center - 0 North Dipsacus sp. 6

134

East Trail 0 South Water 0 West Lythrum salicaria 12 Solidago altissima 5 22 Center - 0 North Solidago altissima 14 East Trail 0 South Geum urbanum 5 Lamiastrum galeobdolon 1 West Lythrum salicaria 6 Solidago altissima 5 Geum urbanum 2 Erigeron annuus 1 23 Centre Impatiens capensis 1 East Impatiens capensis 6 Alliaria petiolata 3 Geranium maculatum 2 Circaea lutetiana 1 South Alliaria petiolata 3 Circaea lutetiana 2 Geranium robertianum 2 West Circaea lutetiana 10 Impatiens capensis 10 Alliaria petiolata 2 Arisaema triphyllum 1 24 Center Impatiens capensis 11 North Impatiens capensis 13 Geranium robertianum 7 Alliaria petiolata 2 East Circaea lutetiana 14 Oxalis sp. 4 Alliaria petiolata 3 Impatiens capensis 3 Solidago altissima 1 South Geranium robertianum 5 Alliaria petiolata 3 West Geranium robertianum 9 Alliaria petiolata 1 Impatiens capensis 1 25 Center Alliaria petiolata 79 Circaea lutetiana 6 Geum canadense 5 North Alliaria petiolata 53 Hieracium caespitosum 9

135

Geum canadense 5 Circaea lutetiana 1 East Alliaria petiolata 62 Geum canadense 2 South Alliaria petiolata 11 West Geranium maculatum 4 Geum canadense 1 Taraxacum officinale 1 26 Center Solanum dulcamara 10 Impatiens capensis 9 North Circaea lutetiana 5 Impatiens capensis 4 Solidago altissima 4 Geranium maculatum 3 East Galium sp. Uncertain Impatiens capensis 10 Lythrum salicaria 2 Myosotis scorpioides 1 South Myosotis scorpioides 11 Impatiens capensis 8 Campanula aparinoides 3 West Impatiens capensis 2 Campanula aparinoides 1 27 Center Arisaema triphyllum 1 North Circaea lutetiana 11 Arisaema triphyllum 1 East Alliaria petiolata 8 Rubus idaeus 1 South Arisaema triphyllum 1 Trillium 1 West Alliaria petiolata 5 28 Center Fragaria sp. 15 Circaea lutetiana 5 Viola sororia 1 North Circaea lutetiana 21 Arisaema triphyllum 3 Fragaria sp. 1 East Fragaria sp. 7 Arisaema triphyllum 2 Geranium robertianum 1 South Fragaria sp. 19 Circaea lutetiana 8 West Fragaria sp. 10 Arisaema triphyllum 3

136

29 Center Arisaema triphyllum 2 North - 0 East - 0 South - 0 West Arisaema triphyllum 2 30 Center Flooded 0 North Flooded 0 East Flooded 0 South Flooded 0 West Flooded 0