To Variation in Woody Debris Supply in Boreal Northeastern Ontario

Responses of ground beetles (Coleoptera: Carabidae) to variation in woody debris supply in boreal northeastern Ontario

Paul Wojciech Piascik

A thesis submitted in conformity with the requirements for the degree of Master of Science in Forestry

Faculty of Forestry University of Toronto

Responses of ground beetles (Coleoptera: Carabidae) to variation in woody debris supply in boreal northeastern Ontario

Paul Piascik

Master of Science in Forestry

Faculty of Forestry University of Toronto

2013

Piascik, Paul. 2013. Responses of ground beetles (Coleoptera: Carabidae) to variation in woody debris supply in boreal northeastern Ontario. Master of Science in Forestry, Faculty of Forestry, University of Toronto.

Abstract

The maintenance of downed woody debris supplies is increasingly being recognized as an integral part of forest management. In order to better manage this resource, it is important to assess its role in supporting biodiversity. In this thesis, I investigate the responses of carabid communities to variation in woody debris availability in an experimental manipulation of woody debris volume in closed-canopy forests and following a biomass harvest in a clearcut. Within closed-canopy forests, total carabid abundance and the abundances of eight species increased significantly with increasing volumes of various types of woody debris, particularly large-diameter, late-decay conifer wood. Similarly, a strong affinity with woody debris was observed in the clearcut. These findings suggest that reductions in woody debris will have negative consequences for carabids and indicate the need to ensure a diverse and abundant supply of woody debris during stand development.

Acknowledgements

I would like to thank my advisors, Dr. Jay Malcolm and Dr. Sandy Smith, for the opportunity to take on this study and for their continuous support throughout its duration. I thank my committee member Dr. Chris Darling for his guidance and valuable inputs.

For invaluable help in the initial stages of learning carabid identification, I would like to thank Kathleen Ryan and Nurul Islam. I thank Henri Goulet at the Canadian National Collection for his hospitality and for helping to enhance my identification skills. Thanks to Brad Hubley at the Royal Ontario Museum for providing me access to study the carabid collection.

Thanks to the Kapuskasing field crews for their efforts, enthusiasm, and for making each season very enjoyable and memorable. I am very grateful to all those who spent long hours in the lab tediously processing my samples.

I would like to thank everyone in the Wildlife Ecology and Forest Entomology labs for their support, encouragement, and insightful discussions. I would also like to thank my friends and colleagues in the faculty for a great atmosphere and continued inspiration that was an integral part of my experience.

Funding for this project has been provided by the Sustainable Forest Management Network, Natural Science and Engineering Research Council of Canada, Ivey Foundation, Tembec, Canadian Forest Service, Ontario Ministry of Natural Resources, and the Faculty of Forestry. A special thanks to Ian Thompson and Dave Morris for logistic support.

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Table of Contents Abstract ...... ii Acknowledgements ...... iii List of Tables ...... v List of Figures ...... vi List of Appendices ...... vii General Introduction ...... 1 Chapter 1: The response of ground beetles (Coleoptera: Carabidae) to a large-scale downed woody debris manipulation in boreal northeastern Ontario ...... 7 Introduction ...... 7 Methods ...... 9 Study Sites ...... 9 Experimental Design ...... 10 Carabid Sampling ...... 11 Downed Woody Debris Sampling ...... 13 Statistical Analyses ...... 15 Results ...... 17 Discussion ...... 26 Chapter 2: The importance of slash for ground beetles (Coleoptera: Carabidae) in a biomass clearcut ...... 35 Introduction ...... 35 Methods ...... 38 Study Site ...... 38 Experimental Design ...... 41 Carabid Sampling ...... 41 Statistical Analyses ...... 42 Results ...... 43 Discussion ...... 50 General Conclusions ...... 58 Literature Cited ...... 65 Appendices ...... 77

List of Tables

Table 1.1. P values of downed woody debris (DWD) volume effects from mixed-model ANCOVAs on total carabid abundance, species richness, and the abundances of the 20 most abundant carabid species as a function of variation in nine DWD variables. Carabids were collected between 2010 and 2011 on a DWD manipulation experiment in northeastern Ontario (see text for details) (vtot represents total volume of DWD, in all other cases the first letter of the acronym represents size [s = small-diameter, l = large-diameter], the second letter the taxon [c = conifer, d = deciduous], and the last two numbers the decay class [12=early, 35=late]).

Table 2.1. Mean abundance (standardized to 100 bucket-nights), total abundance, and species richness of carabids collected in 2010 and 2011 in a biomass clearcut in northeastern Ontario near (<5 m away) and far (>83 m away) from slash piles at three distances from forest edge (Near = 34-40 m, Medium = 66-84 m, Far = 181-268 m). Each mean represents abundances from two pitfall arrays; means are over three collection periods (August of 2010 and June and August of 2011).

List of Figures

Figure 1.1. Most significant positive relationships between abundances of carabids and volume of various downed woody debris types in a downed woody debris manipulation experiment in mature mixedwood forests of northeastern Ontario.

Figure 1.2. Most significant negative relationships between abundances of carabids and volume of various downed woody debris types in a downed woody debris manipulation experiment in mature mixedwood forests of northeastern Ontario.

Figure 1.3. Individual-based rarefaction on total carabid captures in 2010 and 2011 from study plots divided into three classes of downed woody debris volumes (low = white, medium = grey, and high = black) for each of nine downed woody debris variables measured in a woody-debris manipulation experiment in northeastern Ontario. Shown are 95% confidence intervals.

Figure 2.1. Map of the biomass clearcut in northeastern Ontario sampled for carabid beetles in 2010 and 2011. Pitfall trap arrays were close to slash piles (<5 m; circles) or far from slash piles (> 83 m; squares). White = forest; grey with black border = clearcut; solid black lines = roads and major skid trails; dashed black lines = transects sampled for downed woody debris.

Figure 2.2. First two axes from a Principle Component Analysis on the covariance matrix of carabid species ln-transformed abundances in a biomass clearcut in boreal northeastern Ontario. Carabid species acronyms consist of the first four letters of the genus and the first four letters of the species. Total carabid abundance and species richness were passive variables (totabun and Richness, respectively). Symbol shapes and colours represent proximity to slash piles and to forest edge (circles represent samples near slash piles [<5 m away]; squares represent samples away from slash piles [>83 m away]; white indicates samples 34-40 m from forest edge; grey indicates samples 66-84 m from forest edge; black indicates samples 181-268 m from forest edge).

Figure 2.3. Relationship between total carabid abundance and distance from forest edge using A. samples located near slash (< 5 m away; circles) and B. samples located away from slash (> 83 m away; squares) in a biomass clearcut in northeastern Ontario. Each regression based on six pitfall samples collected in 2010 and 2011.

Figure 2.4. Individual-based rarefaction of carabid species for A. samples near slash piles (black) and away from slash piles (white) (<5 m or >83 m away, respectively), B. samples at three distances from forest edge (white = 34-40 m, gray = 66-84 m, and black = 181-268 m), and C. samples near (<5 m away; circles) and away (>83 m away; squares) from slash piles at three distances from forest edge (white = 34-40 m, gray = 66-84 m, and black = 181-268 m) in a biomass clearcut in northeastern Ontario. Error bars indicate 95% confidence intervals.

List of Appendices

Appendix 1. Effective length of drift fence arms for carabid captures in a three-arm drift fence pitfall array

Appendix 2. Principal component analysis (PCA) on the correlation matrix of volumes of downed woody debris in the five decay classes [acronym definition: vdc = volume of decay class; 1-5 = decay class 1-5]

Appendix 3. Schematic plot showing the method of partialling out site effects without removing wood volume effects. In this example, the two sites (represented by white and black circles) have both wood volume effects (i.e., each has a positive slope) and extraneous effects (i.e., differences in regression elevations). As shown by the arrows, extraneous effects were removed by partialling out the least square means (i.e., the differences in elevation)

Appendix 4. Non-standardized number of carabid beetle individuals by species collected in June and August sampling sessions in 2010 and 2011 from experimentally-manipulated boreal mixedwood stands in northeastern Ontario. In total, 27 plots were sampled, each with three pitfall arrays (see text for details)

Appendix 5. Rank abundance curve for carabid species collected in closed-canopy boreal mixedwood stands in northeastern Ontario

Appendix 6. Rank abundance curve for carabid species collected in a biomass clearcut that was formerly a boreal mixedwood stand in northeastern Ontario

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General Introduction

Conserving biodiversity, especially through ensuring critical habitat, is of central importance to sustainable forest management (Attiwill 1994, Graham et al. 1994, Freedman et al. 1996, Lindenmayer et al. 2000). In recognition of this goal, a major target of boreal forest management in Ontario is emulating natural disturbance through clearcutting (OMNR 2001).

The underlying idea behind this strategy is to mimic the post-disturbance conditions and landscape mosaics that boreal biota are adapted to. Included in the guidelines set out for achieving this objective is the recognition of the importance of maintaining sufficient woody debris supplies (OMNR 2001). In Ontario, the dominant natural disturbance events that occur in boreal forests are fire and insect outbreaks (Bergeron et al. 2001, Bergeron et al. 2007). These events typically result in large influxes of woody debris to the landscape (Pearce et al. 2005,

Brassard and Chen 2008). A key difference that typically separates managed boreal landscapes from those shaped by natural disturbance is the reduced amount of woody debris (e.g., Siitonen

2001, Pedlar et al. 2002). As boreal forests become increasingly managed, it is important to have a sound understanding of the ecological role of woody debris and how management-related reductions in availability may affect local biodiversity.

Growing concerns over climate change have led to an increased interest in the development of renewable energy sources, including use of forest biomass as a fuel. Sources for this biomass include logging residues and stems of low commercial value. Development of this resource has the potential to result in even more removal of fibre from forests during harvesting, and a further reduction in woody debris supplies in managed forests. In many forest regions, such management-associated reductions in woody debris are considered a main threat to

2 biodiversity (Siitonen 2001, Grove 2002). For example, in Europe, where most of the forested area has been heavily managed for long periods of time, many organisms are threatened by the reductions in woody debris (Berg et al. 1994, Esseen et al. 1997, Martikainen et al. 2000).

Understanding the response of forest biota to variation in woody debris availability provides the opportunity to minimize these negative ecological consequences in Canadian boreal forests where the logging industry is younger and where management-related disturbances have been relatively less-intense.

Woody debris is an important structural feature of forest ecosystems and comprises a relatively large proportion of dead organic matter (Freedman et al. 1996). Decaying wood plays an ecological role in nutrient cycling by providing a slow-release, long-term source of nutrients

(Hagan and Grove 1999) and as an eventual component of forest soils (Siitonen 2001). In addition to its functional role within forest ecosystems, woody debris is recognised as a critical habitat feature for many forest organisms (Harmon et al. 1986, Graham et al. 1994, Freedman et al. 1996, Hagan and Grove 1999, Carey and Harrington 2001, Ehnstrom 2001) and an important characteristic of old-growth forests (Niemelä 1997). A variety of forest biota are associated with woody debris, including mosses, lichens, fungi, herbaceous and woody plants, and many vertebrate and invertebrate animals (reviewed in Harmon et al. 1986). The maintenance of woody debris is considered a principle strategy to sustain wildlife habitat and ecological function in forest management (Graham et al. 1994, Hagan and Grove 1999).

In addition to the amount of woody debris available in an ecosystem, qualitative features of the wood also are an important consideration (Esseen et al. 1997, Sturtevant et al. 1997,

Langor et al. 2008, Nieto and Alexander 2010). Many different microhabitats exist in decaying wood and are highly variable based on wood characteristics such as tree species, size, state of

3 decay, and the types of fungi colonizing the wood (Jonsson 2000, Siitonen 2001). Studies in western Canadian boreal stands reveal a distinct succession of saproxylic species as woody debris changes physically, chemically, and biologically through the decomposition process

(Hammond et al. 2004, Jacobs et al. 2007).

The quantity and diversity of woody debris within managed ecosystems are generally different from those that have not been managed (Siitonen 2001, Pedlar et al. 2002, Brassard and

Chen 2008) and the amount of woody debris present will vary with the intensity of the harvest

(e.g., Green and Peterken 1997). Sources of woody debris in unmanaged stands include tree mortality and self-thinning as well as natural disturbance such as wind-throw, snow breakage, and fire (Hansen et al. 1991, Lee et al. 1997, Siitonen 2001). Volumes of woody debris within an unmanaged forest are in a continual flux (Graham et al. 1994) with the highest inputs attributed to stand senescence and natural disturbance (Lee et al. 1997, Sturtevant et al. 1997). In disturbance-driven ecosystems, woody debris dynamics have been described to follow a

“U-shaped” pattern (Spies et al. 1988, Sturtevant et al. 1997, Clark et al. 1998) consisting of an initial influx of wood after disturbance followed by decreasing levels resulting from the decay of the initial input and, as the stand matures, a subsequent increase in woody debris levels.

Managed boreal stands generally follow a similar pattern; however, they have lower overall volumes of woody debris relative to unmanaged stands, especially of large-diameter woody debris (Sturtevant et al. 1997, Fridman and Walheim 2000, Siitonen 2001, Rouvinen et al. 2002,

Ekbom et al. 2006). This is particularly true after clearcutting as a result of harvesting large diameter trees and the relatively low volume of post-harvest residues as compared with natural disturbance (Rouvinen et al. 2002). The amount of highly decayed wood is particularly limited

4 as it is largely destroyed by machinery or exposed to drying conditions following logging operations (Hautala et al. 2004).

Because it is difficult to determine the volume of woody debris required to support all of the woody debris-dependant organisms in an ecosystem, use of indicator taxa is a valuable management tool. The use of indicator taxa can be an effective method for measuring the impacts of habitat change by providing an efficient and economical approach to assessing the effects of disturbance on an ecosystem (Pearce and Venier 2006). In North America, carabids

(Coleoptera: Carabidae) are a commonly used insect group to study environmental change in terrestrial ecosystems (e.g., Beaudry et al. 1997, Burke and Goulet 1998, Pearce et al. 2003,

Moore et al. 2004, Pearce et al. 2005, Cobb et al. 2007). They exhibit many qualities of an effective indicator group, such as high abundance and species diversity, varying habitat demands among species, wide distributions, well-known taxonomy, and high sensitivity and rapid responses to habitat change (Lindroth 1961, 1963, 1966, 1968, 1969a, 1969b, Thiele 1977,

Niemelä et al. 1988, Rainio and Niemelä 2003, Pearce and Venier 2006). As a largely predatory group, carabids play an important role in ecosystem dynamics and trophic interactions between plants and other ground-dwelling organisms such as spiders and springtails (Snyder and Wise

2001, Larochelle and Lariviere 2003). Changes in microhabitat conditions, such that may be associated with structural features such as woody debris, are considered to be among the most important factors in determining carabid assemblages (Thiele 1977, Lövei and Sunderland 1996).

Carabids are also relatively easy to sample with pitfall traps, a common method for sampling ground-active invertebrates (Krebs 1999).

Woody debris is recognized to be among the most important substrates for sustaining arthropod diversity in boreal forests (Ehnstrom 2001). Numerous studies have found the

5 abundance of many arthropod taxa to increase near woody debris (Evans et al. 2003, Jabin et al.

2004, Ulyshen et al. 2004, Ulyshen and Hanula 2009a). Indeed, this association with woody debris is widely recognized specifically for carabids in both clearcuts and closed-canopy forests

(Carcamo and Parkinson 1999, Pearce et al. 2003, Work et al. 2004, Latty et al. 2006, Cobb et al. 2007, Ulyshen and Hanula 2009b). Carabids have been reported to utilize woody debris as sites for oviposition, larval development, shelter, and overwintering (Goulet 1974, Larochelle and Lariviere 2003, Bousquet 2010).

Unfortunately carabids are relatively understudied from a woody debris perspective in several Canadian forest types, including mixedwood forests, and specific associations with various quantities and types of wood are poorly understood. Current knowledge on associations with woody debris is typically based on correlations and has not been examined through experimental manipulations that could detect causal relationships. Relationships of Canadian carabid assemblages with woody debris in clearcuts are similarly understudied and have not been examined for slash piles, a dominant structural feature of clearcuts. A number of European studies have found carabids to be positively associated with slash piles in clearcuts (Koivula and

Niemelä 2003, Nittérus and Gunnarsson 2006, Nittérus et al. 2007); however, they were based largely on conifer-dominated stands that contain relatively lower volumes of post-harvest woody debris than mixedwood stands (e.g., Pedlar et al. 2002). Although positive correlations with woody debris have been observed for several carabid species in Canadian mixedwood clearcuts

(Pearce et al. 2003), specific associations with slash piles have not been examined.

In this thesis, I explore associations of carabid beetles with woody debris in mixedwood forests of boreal northeastern Ontario, including forests at two ages in the cutting cycle: shortly after harvest and closed-canopy forests some 36-68 years old. In chapter 1, I examine the

6 response of carabid beetles to variation in the availability of downed woody debris in closed-canopy stands through a large-scale manipulation experiment. I specifically look at changes in carabid community composition, richness, and abundance as a function of variation in quantity, size, species, and decay states of downed woody debris. In chapter 2, I examine whether remnant slash piles contribute to maintaining carabid populations in a clearcut where, in addition to traditional harvesting, much of the logging residues and non-commercial trees were removed from the site for biofuel. I specifically look at differences in local abundances and species richness of carabids near and far from slash piles and as a function of distance from forest edge.

Chapter 1: The response of ground beetles (Coleoptera: Carabidae) to a large-scale downed woody debris manipulation in boreal northeastern Ontario

Introduction

Downed woody debris (DWD) is increasingly being recognized as an important structural feature of forest ecosystems that provides critical habitat for numerous species (Harmon et al.

1986, Freedman et al. 1996, Esseen et al. 1997, Hagan and Grove 1999, Ekbom et al. 2006). In addition to the quantity of DWD, its qualitative features such as the size, species, and state of decay have been identified as key functional characteristics providing diverse microhabitats

(Harmon et al. 1986, Økland et al. 1996, Siitonen 2001, Simila et al. 2003, Hammond et al.

2004, Langor and Spence 2006, Jacobs et al. 2007, Lassauce et al. 2012). A higher volume and qualitative variety of DWD will provide a wider variety of niches and therefore is expected to support greater numbers of saproxylic species (Siitonen 2001).

Forest management can have significant impacts on the availability of DWD in an ecosystem (Sturtevant et al. 1997, Siitonen 2001) and in many forest regions, management- associated reductions in woody debris are considered a main threat to biodiversity (Grove 2002).

In Europe, where most of the forested area has been intensively managed for the past 100 – 150 years, many organisms are threatened by the associated reductions in DWD, including saproxylic insects (Berg et al. 1994, Esseen et al. 1997, Martikainen et al. 2000, Siitonen 2001). In contrast, forest harvesting in the Canadian boreal region is a relatively young industry and there is an opportunity to minimize negative ecological impacts associated with forest management. In

8 order to assess the potential impacts associated with reductions in DWD in Canadian boreal forests, it is important to understand the response of forest biota to variation in DWD availability. Of particular value are manipulation studies that extricate causation from correlation

(Thompson 2006).

Carabids (Coleoptera: Carabidae) are commonly used to evaluate changes associated with disturbances in forest ecosystems as they are ubiquitous, diverse, abundant, well-known taxonomically, and highly sensitive to environmental change; they also respond rapidly to habitat change (Lindroth 1969a, Niemelä et al. 1988, Rainio and Niemelä 2003, Pearce and Venier

2006). They also are an integral part of trophic interactions on the forest floor in that they are largely predatory, feeding on a variety of different prey (Lövei and Sunderland 1996) and simultaneously provide an abundant source of food for other predators. Microhabitat conditions are considered to be among the most important factors in determining carabid assemblages

(Thiele 1977) and therefore structural features, such as DWD that provide relatively stable microclimates, are presumably important habitat features. Downed woody debris contributes to habitat heterogeneity and structural complexity within an ecosystem and may also be an important factor in carabid prey distributions. The abundance of numerous arthropod taxa have been found to increase near woody debris (e.g., Evans et al. 2003, Jabin et al. 2004, Ulyshen and

Hanula 2009). A number of studies in a variety of habitats have reported correlations between carabids abundances and variation in woody debris supplies (Pearce et al. 2003, Latty et al.

2006, Cobb et al. 2007). Downed woody debris is thought to serve as sites for oviposition, diurnal shelter, and overwintering (Goulet 1974, Bousquet 2010). Unfortunately, however, carabids are relatively understudied from a DWD perspective in Canada, especially in closed-canopy mixedwood forests. Although recognized as an important habitat component for

9 carabids (eg. Work et al. 2004), habitat associations with DWD are poorly understood and responses to experimental manipulations of DWD supplies have not been examined. Examining the response of carabids to variation in the availability of DWD has the potential to provide valuable insights into the importance of ensuring a sustainable supply of DWD during forest management.

Here, I report on the effects of a large-scale DWD manipulation experiment on carabid communities in closed-canopy, boreal mixedwood stands. My specific objective was to examine changes in carabid communities as a function of variation in quantity, size, species, and decay states of DWD.

Methods

Study Sites

The study was conducted in northeastern Ontario, Canada, in the Gorden Cosens Forest

Management Unit within 80 km of Kapuskasing (49°25’0” N, 82°26’0” W). Nine study sites were established in the oldest, post-clearcut forests available in the study area, including 1) forests clearcut mechanically using skidders between 1967 and 1975 (n = 4 sites) and 2) forests clearcut using horses between 1943 and 1959 (n = 5 sites). A range of stand ages were used in the experiment because it allowed indirect investigation of qualitative variation in DWD supplies. The younger, mechanically-logged sites had relatively large quantities of highly decayed DWD remnant from logging operations, but relatively lower quantities of recent DWD inputs from stand development (Fischer et al. in press). By contrast, in the older, horse-logged stands, DWD resulting from the logging operations had largely disappeared, but recent DWD was relatively common (Fischer et al. in press) due to higher inputs associated with more mature

10 stands such as self-thinning (Sturtevant et al. 1997). All sites were closed-canopy mixedwoods; that is, they had a mixture of both deciduous and coniferous tree species. Typically, such forests are dominated by poplars, balsam fir, and white spruce. Based on 16 prism sweeps per experimental plot (see below) undertaken in 2006 and 2007, the plots on average had 52% deciduous composition by basal area and 48% conifer composition by basal area. All had at least some Populus (mean = 40%, range = 6-78%) and at least some Picea (mean = 17%, range = 1-

48%).

Experimental Design

Within each site, which consisted of either a single stand (n = 6) or two nearby stands of the same age and species composition (n = 3) according to Forest Resource Inventory maps

(Ontario Ministry of Natural Resouces, unplublished), we established three 2.25 ha plots (150 x

150 m) separated by at least 150 m and located at least 100 m from roads. In the central portion of each plot, we established a 90 × 90 m grid with 15-m spacing to serve as a focus of sampling.

Within each site, three treatments were assigned, one at random to each plot. In the "full- removal" treatment, all DWD ≥ 7 cm diameter was removed; in the "half-removal" treatment, one-half of the DWD was removed; and in the control treatment, no DWD was removed. Three sites were treated in 2006; the others were treated in 2007. In each plot, three parallel skid trails separated by approximately 50 m were established to facilitate removal of the DWD and to permit skid distances ≤ 25 m. In order to minimize site damage, skid trails were laid out to minimize tree cutting and to avoid low-lying areas. Wood removal was undertaken using a small

(cable) skidder, with the exception of wood too decayed to skid, which was instead broken apart with chainsaws and axes (it was completely removed from the full-removal plots at 4 years post-treatment). In the full removal treatment plots, all new DWD accumulation was removed at

11 two year intervals. All removed wood was piled at least 75 m away from the plot edge. Skid trails in control grids were lightly bladed to emulate the repeated travel on skid trails in the removal plots. Wood removals occurred in August-October.

Detailed sampling of DWD was undertaken prior to and following the manipulation

(respectively, at 1.0 year pre-manipulation on average [range: 0.2 - 2.2 years] and at 1.1 years post-manipulation on average [range: 0.6 - 2.0 years]). Full details on methods are provided below in the section on additional DWD sampling undertaken in support of the carabid research.

This before-and-after sampling (1680 m of line intersect per plot) revealed the expected reduction in DWD volumes as well as the expected change in DWD correlations before and after the manipulation. DWD volume remained more-or-less constant in the control grids before and after the manipulation (49 and 45 m3/ha on average, respectively) and the correlation between the before and after measurements across these plots was high (R2 = 84%, p = 0.0004). In the half-removals, DWD volume was reduced by 45% (from 61 to 33 m3/ha on average) and in the full-removals, by 81% (67 to 13 m3/ha on average). As expected, due to the manipulation, the correlation between pre- and post-manipulation DWD volume across the 27 sites was not significant (R2 = 6%, p = 0.22).

Carabid Sampling

Carabids were sampled using pitfall buckets in combination with drift-fences formed into

“Y”-shaped arrays. The 90 × 90 m grid was divided into four 45 × 45 m quadrants, and an array was placed in the centre of three of the four quadrants. Drift fences were built of polypropylene geotextile ("green line") with one pitfall bucket placed at the end of each of the three arms and one in the centre of the array (i.e., 4 buckets per array and 12 buckets per grid). Each arm was

0.5 m high and 3.65 m long, with the bottom 10 cm of the geotextile buried in the soil. Buckets

12 were 15 cm in diameter by 15 cm deep (with the exception of one sampling session where a number of smaller buckets were used; see below) and were buried flush with the ground surface and the edge(s) of the pitfall arms. During trapping, the buckets were filled with 5 cm of a 5% saline solution to which a small amount of soap had been added to break surface tension.

Carabids were sampled twice per year in 2010 and 2011; 4 – 12 June and 24 August – 1

September in 2010 and 12 – 19 June and 15 – 22 August in 2011. During each sampling session, traps were active for five consecutive nights; buckets were covered with lids and leaf litter at other times. Sampling among the various plots was conducted as near to simultaneously as possible in each sampling session (i.e., all traps were set within 3 days of one another and the three plots per site were set simultaneously). Pitfall samples were collected at the end of each five-night session and preserved in 70% ethanol. Carabids were identified to species using keys from Lindroth (1961, 1963, 1966, 1968, 1969a, 1969b) and Bousquet (2010). Nomenclature follows that of Bousquet (2010). A voucher collection was authenticated by H. Goulet at the

Canadian National Collection, Ottawa, and is located at the Faculty of Forestry, University of

Toronto, Ontario.

0 % of centre buckets in 2010, and 7.41 % of end buckets and 3.70 % of centre buckets in 2011.

In order to calculate effort, I accounted for the fact that the centre and end buckets did not represent the same sampling effort: centre buckets can be expected to catch relatively more carabids than end buckets because they are fed by three drift fence arms instead of one. To determine the appropriate conversion factor, I counted the number of carabids in centre and end

13 buckets for 219 arrays over both sessions in 2010. On average, centre buckets captured 2.236 times as many individuals as end buckets, which was therefore used in calculating bucket-nights and in standardizing for missing effort (specifically, one array set for 5 nights with no missing effort was assumed to represent [2.236 · 1 centre bucket + 3 end buckets] · 5 nights = 26.18 bucket-nights). Interestingly, this conversion factor can be used to estimate the length of an array arm that was effective in directing carabids into a bucket (Appendix 1).

An additional complication arose during the June 2010 trapping session in that 69 of the end buckets across 37 pitfall arrays had slightly smaller buckets (11 cm diameter × 13.3 cm high instead of 15 × 15 cm). To calculate the appropriate conversion factor, I counted the number of carabids in small versus large buckets for 28 arrays, and found that a small bucket captured 60.1

% of the number of individuals as a large one. This was not significantly different from expectations based on the ratio of perimeters of the small:large buckets (0.733; one-sample t-test p > 0.05) and is consistent with the findings of Lange et al. (2011) who found that the number of carabid individuals captured increased with the diameter of the pitfall trap.

Downed Woody Debris Sampling

Detailed surveys of DWD were carried out in 2010 in each plot using the line-intersect method (Van Wagner 1968; 3 years after the manipulation on average [range: 2.6 - 3.8 years]).

Fourteen transects, each 120 m long (extending 15 m beyond the 90 × 90 m grid) were sampled in each plot for a total of 1680 m per plot. Transects were spaced 15 m apart with seven transects spanning each dimension of the square grid (and thereby intersecting 49 times). For every piece of downed wood ≥ 7 cm diameter intercepted on the transect, the diameter, state of decay

(“early” or “late”), and whether it was coniferous or deciduous was recorded. Early decay wood had firm outer layers, was largely intact, and could not be kicked apart; those in a late stage of

14 decay had soft to substantially decayed or missing outer layers and could be kicked apart with either some or little effort. These groupings approximately correspond to decomposition classes

I-II and III-V respectively as described by Maser et al. (1979). A principle component analysis

(PCA) on the correlation matrix of volumes in decay classes I-V revealed that volume of decomposition class III was approximately equally correlated with both classes I-II and IV-V

(Appendix 2); based on the substantial decay of sapwood in decay class III logs, I grouped it with the “late decay” category. From these data, I calculated wood volumes for nine variables: 1) total downed wood volume, 2) small-diameter, early-decay conifer, 3) small-diameter, late-decay conifer, 4) large-diameter, early-decay conifer, 5) large-diameter, late-decay conifer, 6) small-diameter, early-decay deciduous, 7) small-diameter, late-decay deciduous, 8) large-diameter, early-decay deciduous, and 9) large-diameter, late-decay deciduous.

Classification as either "small-diameter" or "large-diameter was based on the median diameter

(small = ≤ 12 cm diameter, large = > 12 cm diameter). When downed wood could not be identified as either deciduous or coniferous with certainty, I assumed that it was distributed between the two taxonomic groupings in the same ratio that was obtained for identified pieces of the same size and decay class. The mean volumes across the 27 sites for the various DWD variables were: 1) total downed wood volume = 43.5 m3/ha [range:6.2 - 133.4 m3/ha], 2) small-diameter, early-decay conifer = 1.4 m3/ha [range: 0.1 - 4.6 m3/ha], 3) small-diameter, late-decay conifer = 2 m3/ha [range: 0.2 – 9.1 m3/ha], 4) large-diameter, early-decay conifer =

2.5 m3/ha [range: 0.2 – 8.8 m3/ha], 5) large-diameter, late-decay conifer = 5.4 m3/ha [range:

0 - 19 m3/ha], 6) small-diameter, early-decay deciduous = 3.3 m3/ha [range: 0.6 – 9.7 m3/ha], 7) small-diameter, late-decay deciduous = 2.3 m3/ha [range: 0.6 – 8.4 m3/ha], 8) large-diameter,

15 early-decay deciduous = 6.8 m3/ha [range: 0 – 37.5 m3/ha], and 9) large-diameter, late-decay deciduous = 19.7 m3/ha [range: 1.4 – 67.4 m3/ha].

Statistical Analyses

The wood removal treatments in this study were designed to result in a net reduction in wood volume among the sites with two rationales: 1) species DWD relationships were more likely to be more evident when wood volume was limiting and 2) such relationships would reflect causation rather than correlation; that is, effects due to variation in DWD volumes per se rather than other factors that might have originally been correlated with variation in DWD volumes. Because the reductions in wood volumes were relative rather than absolute, I took a regression approach rather than an analysis of variance approach (that is, the volumes of DWD in the plots as measured in 2010 were was used as a continuous independent variable rather than classifying grids as being from one or another of the removal treatments). I did this because half-removal treatments were relative to what was originally there; for example, after treatment, a half removal plot may have a higher volume of DWD than a control plot. In addition, control plots were assigned at random, so it is possible that a control plot might not only have less DWD than a half removal plot, but it might even have less DWD than a full removal plot (given that steady, albeit slight, additions of DWD over time occurred in the full removal plots through natural events such as snag falls and windthrow).

For those species that strongly varied in abundance between the June and August sessions, I reasoned that from a statistical standpoint, the period of peak abundance would best reveal DWD relationships because sampling error at other times would play a larger role in determining whether or not a trap contained individuals. Accordingly, if > 80% of the standardized captures for a given species were in either June or August, I analyzed standardized

16 abundance from only their most abundant month (except for rare species; see below).

Standardized species abundances were ln(x + 1) transformed to better meet the assumptions of normality and homogeneity, which were evaluated from plots of residuals in species-specific tests (see below). The 20 most abundant species (species occurring in ≥ 50% of samples) and rare species were examined separately for species-specific responses to variation in DWD volume.

Prior to species-specific analyses, a permutation test using redundancy analyses (RDA;

Lepš and Šmilauer 2003) on the 20 abundant species was performed to assess whether the nine

DWD variables explained significant variation in the correlation matrix of carabid standardized abundances.

Species-specific analyses for the 20 most abundant species and analyses of total abundance and species richness were performed using analysis of covariance in the SAS Mixed procedure (SAS v. 9.2) with trapping session as a repeated measure, site as a random effect, and

DWD volume as a continuous variable. DWD variables included in the model were checked for possible curvilinear relationships via binomial regression. In order to present significant relationships graphically, I removed date and site effects and plotted mean standardized abundance over all sessions against the relevant DWD variable. A complication that may arise in this situation is that site effects might consist of both DWD volume effects (i.e., the possibility that some sites might have more DWD on average than others) and extraneous effects (i.e., the possibility that some sites might have more individuals of a certain species than others, irrespective of DWD volumes). Accordingly, to partial out the latter, but retain the former, I used Analysis of Covariance to remove elevation (intercept) effects (see Appendix 3). This ensured that DWD volume effects among sites remained intact, but that extraneous site-related

17 variation was removed. For rarer species, I classified abundances as above or below the species-specific median, and undertook contingency table analyses on counts in the two categories. In these analyses, standardized total species abundance across all four sampling sessions was used and each plot was categorized as having either “low” or “high” volumes of each DWD variable.

I also undertook rarefaction analyses to test whether species richness corrected for abundance varied as a function of variation in DWD volume. Rarefaction curves were generated

(before standardization for sampling effort) for each DWD variable by categorizing each plot as having either "low" "medium" or "high" volumes of that variable. In this way, I could examine if species accumulation varied between the three wood categories. Wood categories were determined using tri-tiles; i.e., approximately one-third of plots were in each class. Rarefaction estimates were calculated using the formula for individual-based rarefaction (mean and variance) from Coleman et al. (1982).

Results

A total of 11,604 individuals from 22 genera and 43 species were collected over the four trapping sessions in 2010 and 2011 (Appendix 4; Rank abundance curve: Appendix 5). The 20 most abundant species (each represented in at least 50% of the samples) were Agonum gratiosum

(Mannerheim), Agonum retractum LeConte, Agonum sordens Kirby, Bradycellus lugubris

(LeConte), Calathus ingratus Dejean, Harpalus fulvilabris Mannerheim, Loricera pilicornis

(Fabricius), Patrobus foveocollis (Eschscholtz), Platynus decens (Say), Platynus mannerheimii

(Dejean), Pterostichus adstrictus Eschscholtz, Pterostichus coracinus (Newman), Pterostichus melanarius (Illiger), Pterostichus pensylvanicus LeConte, Pterostichus punctatissimus (Randall),

Scaphinotus bilobus (Say), Sphaeroderus nitidicollis Guerin-Meneville, Sphaeroderus

18 stenostomus Dejean, Synuchus impunctatus (Say), and Trechus apicalis Motschulsky.

Abundances varied strongly between sampling years and months; 2011 represented 72% of the total captures, and June 2011 itself represented 60% of the total captures. Of the 20 abundant species, 13 were much more abundant in June than in August (Agonum gratiosum [95% of captures in June], Agonum retractum [98% of captures in June], Agonum sordens [97% of captures in June], Bradycellus lugubris [99% of captures in June], Calathus ingratus [87% of captures in June], Harpalus fulvilabris [80% of captures in June], Loricera pilicornis [96% of captures in June], Patrobus foveocollis [98% of captures in June], Platynus decens [96% of captures in June], Platynus mannerheimii [96% of captures in June], Pterostichus adstrictus

[94% of captures in June], Pterostichus pensylvanicus [95% of captures in June], and

Pterostichus punctatissimus [97% of captures in June]); two were more abundant in August than

June (Sphaeroderus nitidicollis [87% of captures in August], and Synuchus impunctatus [100% of captures in August]); and five did not show strong seasonal variation (Pterostichus coracinus

[63% of captures in June, 37% in August], Pterostichus melanarius [58% of captures in June,

42% in August], Scaphinotus bilobus [50% of captures in June, 50% in August], Sphaeroderus stenostomus [46% of captures in June, 54% in August], and Trechus apicalis [70% of captures in

June, 30% in August ]).

Redundancy analysis indicated that 40.8 % of the variance in the abundances of the 20 species was explained by the nine DWD variables (p = 0.0127). Species-specific tests revealed that 12 of the 20 abundant species were significantly correlated (p < 0.05) with variation in one or more of the nine DWD variables (Table 1.1). Of these, six showed significant relationships with multiple DWD variables, and all except one showed at least one relationship that was highly significant (p < 0.01). In all cases, the sign of the relationships for a given species were either

19 always positive or always negative. Agonum sordens was positively associated with small and large-diameter early-decay deciduous DWD (p = 0.0102 and 0.0251, respectively) and both coniferous and deciduous small-diameter late-decay DWD (p = 0.0201 and 0.0016, respectively).

Bradycellus lugubris was negatively associated with total DWD volume (p = 0.0038), small- diameter early-decay conifer DWD (p = 0.0052), and both small and large-diameter late-decay deciduous DWD (p = 0.0408 and 0.0021, respectively). Loricera pilicornis was positively associated with both early and late-decay small-diameter coniferous DWD (p = 0.0146 and

0.0260, respectively) and small-diameter, late-decay deciduous DWD (p = 0.0093). Pterostichus adstrictus was negatively associated with total DWD volume (p = 0.0324), both small and large- diameter late-decay deciduous DWD (p = 0.0318 and 0.0351, respectively), and large-diameter, early-decay deciduous DWD (p = 0.0048). Scaphinotus bilobus was positively associated with total DWD volume (p = 0.0028), small and large-diameter early-decay deciduous DWD (p =

0.0181 and <0.0001, respectively), and small-diameter late-decay coniferous and deciduous

DWD (p = 0.0259 and 0.0003, respectively). Synuchus impunctatus was negatively associated with both early and late-decay small-diameter conifer (p = 0.0400 and 0.0267, respectively). Six species showed significant relationships with just one DWD variable; two of these showed a relationship that was highly significant. Calathus ingratus, Platynus decens, Pterostichus coracinus, and Pterostichus melanarius were positively associated with large-diameter, late- decay conifer DWD (p = 0.0344, 0.0002, 0.0465, and < 0.0001 respectively). Sphaeroderus stenostomus was negatively associated with large-diameter, late-decay conifer DWD (p =

0.0275), and Trechus apicalis was positively associated with small-diameter, late-decay DWD (p

= 0.0314).

Table 1.1. P values of downed woody debris (DWD) volume effects from mixed-model ANCOVAs on total carabid abundance, species richness, and the abundances of the 20 most abundant carabid species as a function of variation in nine DWD variables (significant values in boldface type (α = 0.05)). Carabids were collected between 2010 and 2011 on a DWD manipulation experiment in northeastern Ontario (see text for details) (vtot represents total volume of DWD, in all other cases the first letter of the acronym represents size [s = small-diameter, l = large-diameter], the second letter the taxon [c = conifer, d = deciduous], and the last two numbers the decay class [12=early, 35=late]).

Woody debris variableb Response variablea vtot sc_12 sd_12 sc_35 sd_35 lc_12 ld_12 lc_35 ld_35 Total abundance 0.8611 0.7632 0.7345 0.8100 0.5256 0.1794 0.7817 0.0004 (+) 0.2033 Species richness 0.7582 0.6544 0.9003 0.8614 0.1968 0.4268 0.4795 0.1825 0.6733 Agonum gratiosum2 0.7760 0.3396 0.4977 0.8652 0.6613 0.8680 0.5179 0.6333 0.2085 Agonum retractum2 0.3329 0.0804 0.5485 0.6133 0.6873 0.6062 0.4111 0.1567 0.4812 Agonum sordens2 0.3005 0.6465 0.0102 (+) 0.0201 (+) 0.0016 (+) 0.8864 0.0251 (+) 0.6966 0.7155 Bradycellus lugubris2 0.0038 (- ) 0.0052 (- ) 0.2187 0.0644 0.0408 (-) 0.0559 0.0975 0.8165 0.0021 (- ) Calathus ingratus2 0.2016 0.7618 0.6786 0.6899 0.8163 0.1262 0.9588 0.0344 (+) 0.1189 Harpalus fulvilabris2 0.4190 0.4141 0.2316 0.8871 0.2597 0.3832 0.2582 0.5436 0.6041 Loricera pilicornis2 0.5409 0.0146 (+) 0.2571 0.0260 (+) 0.0093 (+) 0.4704 0.2625 0.7092 0.0624 Patrobus foveocollis2 0.4360 0.9023 0.3897 0.9349 0.2818 0.6163 0.3410 0.0941 0.9613 Platynus decens2 0.1924 0.1858 0.3316 0.2523 0.1098 0.8816 0.4923 0.0002 (+) 0.7135 Platynus mannerheimii2 0.3827 0.0716 0.1490 0.3207 0.1943 0.4345 0.5261 0.2538 0.5553 Pterostichus adstrictus2 0.0324 (- ) 0.2972 0.2231 0.1389 0.0318 (- ) 0.8657 0.0048 (- ) 0.8004 0.0351 (- ) Pterostichus coracinus1 0.5956 0.7493 0.3371 0.9303 0.6676 0.0796 0.4240 0.0465 (+) 0.1174 Pterostichus melanarius1 0.8721 0.6491 0.2815 0.5299 0.5443 0.4424 0.2699 <0.0001 (+) 0.2647 Pterostichus pensylvanicus2 0.9076 0.6797 0.2071 0.4943 0.6885 0.2167 0.6324 0.0613 0.6520 Pterostichus punctatissimus2 0.2985 0.4877 0.3598 0.9357 0.8097 0.7344 0.6165 0.9870 0.1340 Scaphinotus bilobus1 0.0028 (+) 0.2304 0.0181 (+) 0.0259 (+) 0.0003 (+) 0.0080 (+) <0.0001 (+) 0.7589 0.0803 Sphaeroderus nitidicollis3 0.4210 0.1446 0.1870 0.1662 0.2418 0.3667 0.8886 0.0741 0.7822 Sphaeroderus stenostomus1 0.8486 0.0685 0.5398 0.1912 0.9028 0.7593 0.5391 0.0275 (- ) 0.2744 Synuchus impunctatus3 0.8124 0.0400 (- ) 0.5829 0.0267 (- ) 0.3996 0.9119 0.6132 0.1885 0.6464 Trechus apicalis1 0.3890 0.4465 0.2149 0.0954 0.0314 (+) 0.3349 0.2177 0.2485 0.9440 a Superscripts indicate the trapping session(s) that was analyzed (1 = June and August, 2 = June, 3 = August). b For significant relationships (p < 0.05), signs of the regression slope are indicated in parenthesis to the right of the p value.

In terms of the number of significant relationships with the various DWD variables, each variable had two or more significant relationships with one or another of the twelve carabid species, with the exception of large-diameter early-decay conifer, which had one. Deciduous and conifer DWD had approximately equal numbers of significant relationships (12 and 13, respectively); however, deciduous DWD had more highly significant relationships than conifer

DWD (6 and 4, respectively). Concerning the signs of the significant relationships, eighteen were positive and eleven were negative. Of these, seven of the positive relationships were highly significant and four of the negative relationships were highly significant. Of the most significant

DWD relationships found for each carabid species, eight of twelve were with large-diameter logs, and ten of twelve were with late-decay logs. Of these, six of twelve were with large- diameter, late-decay logs of which five were conifers.

When the relationship with the most significant DWD variables was plotted, ten of the twelve abundant carabid species showed linear responses (Fig. 1.1 and Fig. 1.2). Only Agonum sordens and Pterostichus melanarius displayed a significant binomial relationship, in these cases with small-diameter, late-decay deciduous DWD and large-diameter, late-decay coniferous

DWD, respectively (Fig. 1.1).

Species-specific tests on rare species revealed four species with significant (p < 0.05) relationships with variation in the nine DWD variables, three of these with multiple variables. In all cases, the sign of the relationships for a given species were either always positive or always negative. Amara lunicollis was positively associated with small-diameter, late-decay conifer (p =

0.0407). Bembidion wingatei was negatively associated with both early and late-decay small-diameter conifer (ps = 0.0183). Carabus maeander was positively associated with small

Total carabid abundance Calathus ingratus Platynus decens Pterostichus coracinus Pterostichus melanarius 5.4 2.5 6 5 4 5.1 4 2.0 3 4 4.8 1.5 3 2 4.5 1.0 2 2 2 1 4.2 p = 0.0004 r = 0.3575 0.5 p = 0.0344 r2 = 0.3062 p = 0.0002 r2 = 0.4249 1 p = 0.0465 r2 = 0.1480 p = < 0.0001 r2 = 0.6528

ln(abundance + 1)+ ln(abundance y = 0.0336x + 4.3465 y = 0.0603x + 0.9648 y = 0.0709x + 3.3662 y = 0.0557x + 2.2651 y = 0.1108x + 0.1786 3.9 0.0 0 0 0 0 5 10 15 20 0 5 10 15 20 0 5 10 15 20 0 5 10 15 20 0 5 10 15 20 Large-diameter, late-decay conifer (m3) Large-diameter, late-decay conifer (m3) Large-diameter, late-decay conifer (m3) Large-diameter, late-decay conifer (m3) Large-diameter, late-decay conifer (m3)

Scaphinotus bilobus Agonum sordens Loricera pilicornis Trechus apicalis 3.5 1.6 0.9 2.0 1.2 2.5 1.5 0.8 0.6 1.5 1.0 0.4 0.3 2 2 0.5 0.0 2 0.5 2 p = 0.0016 r = 0.3431 p = 0.0093 r = 0.4118 p = 0.0314 r = 0.3322 p < 0.0001 r = 0.6402 ln(abundance + 1) + ln(abundance y = 0.2172x + 0.5858 y = 0.1277x + 0.1264 y = 0.0637x + 0.0844 y = 0.0345x + 0.4754 -0.5 -0.4 0.0 0.0 0 2 4 6 8 10 0 2 4 6 8 10 0 2 4 6 8 10 0 10 20 30 40 Small-diameter, late-decay deciduous (m3) Small-diameter, late-decay deciduous (m3) Small-diameter, late-decay deciduous (m3) Large-diameter, early-decay deciduous (m3)

Pterostichus adstrictus Bradycellus lugubris 2 p = 0.0048 r = 0.3808 p = 0.0021 r2 = 0.5849 4 y = -0.0371x + 2.7761 3 y = -0.0312x + 1.4289 3 2

2 1

1 0

ln(abundance + 1) + ln(abundance 1) + ln(abundance 0 -1 0 10 20 30 40 0 20 40 60 80 Large-diameter, early-decay deciduous (m3) Large-diameter, late-decay deciduous (m3)

Synuchus impunctatus Sphaeroderus stenostomus 2 p = 0.0267 r = 0.2162 p = 0.0275 r2 = 0.2028 3 y = -0.1246x + 1.7267 1.8 y = -0.0246x + 1.0484 1.5 2 1.2 0.9 1 0.6

ln(abundance + 1)+ ln(abundance

ln(abundance + 1) + ln(abundance 0 0.3 0 2 4 6 8 10 0 5 10 15 20 Small-diameter, late-decay conifer (m3) Large-diameter, late-decay conifer (m3)

and large-diameter early-decay conifer as well as small-diameter late-decay conifer (ps =

0.0159). Pterostichus luctuosus was negatively associated with total DWD volume, small-diameter early-decay deciduous DWD, and large-diameter late-decay conifer (ps =

0.0407).

Total carabid abundance showed a highly significant (p < 0.001) positive relationship with volume of large-diameter, late-decay conifer (Table 1.1, Fig. 1.1), but no significant

24 relationships with any of the other wood variables. Species richness was not significant for any of the nine wood volume variables (Table 1.1).

The asymptotic nature of the individual-based rarefaction curves indicated that the carabid communities were relatively well sampled in that most of the species in the sampling area were captured (Fig. 1.3). Rarefaction revealed no overarching pattern of variation in species richness among the three classes of wood volumes. Generally, considerable overlap of standard error bars were shown among the three DWD classes, with the possible exceptions of total DWD volume and large-diameter, early-decay deciduous DWD, where sites with the lowest volumes had the highest relative species richness. There was little consistency in the patterns shown among the nine DWD variables, however. Sites with the greatest wood volume had the highest relative species richness for only 1 of the 9 variables, whereas sites with low or medium wood volumes had the highest relative species richness in 8 of 18 cases. This difference in proportions was not significant, however (Fisher's Exact Test, p = 0.19).

Total woody debris volume Small-diameter, early-decay conifer Small-diameter, early-decay deciduous 40

Number of species of Number 0

Small-diameter, late-decay conifer Small-diameter, late-decay deciduous Large-diameter, early-decay conifer 40

Number of species of Number 0

Large-diameter, early-decay deciduous Large-diameter, late-decay conifer Large-diameter, late-decay deciduous 40

Number of species of Number 0 0 1000 2000 3000 4000 5000 0 1000 2000 3000 4000 5000 0 1000 2000 3000 4000 5000 Number of individuals Number of individuals Number of individuals

Discussion

Annual and seasonal population fluctuations among carabids similar to those found in this study have been reported elsewhere (Barlow 1970, Jones 1979, Willand et al. 2011). These fluctuations have been suggested to be associated with variation in temperature (den Boer 1981), seasonal activity periods coinciding with reproductive period (typically either spring or autumn)

(Lövei and Sunderland 1996, Langor et al. 2008), and changes in prey abundance (Symondson et al. 2002). In the current study I did not measure factors that might be affecting carabid population fluctuations; however, there was no significant difference in mean temperature across trapping months between years (p > 0.15; paired t-test with day as the replicate; Environment

Canada 2012) and so the higher carabid captures in June 2011 are not likely an effect of variation in temperature. The period of peak abundance for the 20 abundant species I analyzed coincided relatively well with their respective breeding periods as suggested by Lövei and Sunderland

(1996) and Langor et al. (2008). Of the thirteen species with peak abundances in June samples, ten are considered spring breeders and the two species with peak abundances in August are considered autumn breeders. Of the five species that did not show seasonal variation, four are considered autumn breeders.

Of the twelve abundant carabid species correlated with DWD, eight displayed positive relationships (Agonum sordens, Calathus ingratus, Loricera pilicornis, Platynus decens,

Pterostichus coracinus, Pterostichus melanarius, and Scaphinotus bilobus) and four negative

(Bradycellus lugubris, Pterostichus adstrictus, Sphaeroderus stenostomus, and Synuchus impunctatus). Within each species this relationship was consistent for all DWD variables.

Agonum sordens, Loricera pilicornis, and Trechus apicalis displayed positive relationships to the volume of small-diameter, late-decay, deciduous DWD. In the case of A. sordens this

27 relationship was curvilinear suggesting a threshold response. These largely nocturnal species are known to utilize DWD for diurnal shelter, and in the case of A. sordens and L. pilicornis overwinter in DWD (Larochelle and Lariviere 2003), and are often associated with wet habitats

(Larson et al. 1999, Bousquet 2010). Downed woody debris may, to some extent, provide the necessary wet or humid conditions in these mixedwood stands, which are otherwise a drier habitat than typically associated with these species. Calathus ingratus, Platynus decens,

Pterostichus coracinus, and Pterostichus melanarius were all positively correlated with large- diameter, late-decay coniferous DWD. The former three species are considered to be forest habitat generalists and are endemic to North America, while P. melanarius is an introduced species that is now widespread and sometimes associated with moist habitats (Niemelä and

Spence 1991, Niemelä and Spence 1999, Bourassa et al. 2011). These species are known to utilize woody debris for shelter and overwintering (Larochelle and Lariviere 2003). C. ingratus and P. decens are often found under logs, loose bark of DWD, and generally in moist areas

(Larochelle and Lariviere 2003). However, Pearce et al. (2003) found contradictory results in that P. decens was negatively associated with woody debris in mature (>100 years old) deciduous boreal forests, although they did not undertake experimental manipulations.

Gastropods, which are an important prey for P. melanarius (Symondson et al. 1996, Symondson et al. 2002, Oberholzer and Frank 2003), are often found in the moist environments under logs and bark of woody debris (Savely 1939, Kappes et al. 2006). Scaphinotus bilobus was positively correlated with six of the nine DWD variables and is usually found on moist and shady soils and is known to utilize DWD for shelter (Larochelle and Lariviere 2003). S. bilobus has also been associated with old growth forests containing high volumes of woody debris (Bertrand 2005,

Janssen et al. 2009). As a specialist gastropod predator, S. bilobus may also be drawn to DWD as

28 a substrate often associated with its prey. The results of this study are consistent with the existing literature and provide additional evidence of the importance of DWD as a habitat feature for S. bilobus.

The species and state of decay of DWD have been suggested to be important factors structuring saproxylic beetle assemblages (Siitonen 2001, Grove 2002, Jacobs et al. 2007) and associations with either early-decay or late-decay wood have been noted for insect communities

(Vanderwel et al. 2006). Different tree species have varying decay rates and chemical composition and therefore will provide variable microhabitats for insects (Harmon et al. 1986,

Siitonen 2001). Based on a combination of rearing from bolts and window traps on snags of

Populus in boreal aspen stands, Hammond et al. (2004) found that beetle species richness increased with wood decay whereas abundance was higher in early stages of decay. The microclimatic conditions provided by DWD will vary with the state of decay of the wood, with increasing humidity and cool conditions in later states of decay. Such variation in temperature and humidity are known to influence the distribution of carabids within a stand (Thiele 1977).

Late-decay wood was the best predictor variable for total carabid abundance and seven of the twelve abundant carabid species that were positively correlated with DWD, Agonum sordens,

Calathus ingratus, Loricera pilicornis, Platynus decens, Pterostichus coracinus, Pterostichus melanarius, and Trechus apicalis; each of these species, with the exception of T. apicalis, is known to utilize wood as an overwintering substrate (Larochelle and Lariviere 2003). Wood in an advanced state of decay is softer and more accessible than early-decay wood for individuals requiring access for overwintering or oviposition. For example, Pterostichus adstrictus individuals are often found ovipositing under the bark of moist decayed wood, a substrate also used by its three larval instars and pupae (Goulet 1974). These results suggest that wood in

29 advanced stages of decay are especially important for some carabid species. The range of microhabitats available in DWD will increase with the diameter and therefore it is likely that the size distribution of DWD is important to many saproxylic organisms (Martikainen et al. 2000). A

Canadian boreal forest study suggests that many beetle species depend directly on large-diameter woody debris (Hammond et al. 2004). In my study, total carabid abundance and four of the eight species positively correlated with DWD (Calathus ingratus, Platynus decens, Pterostichus coracinus, and Pterostichus melanarius) were associated with only large-diameter, late-decay coniferous DWD.

Interestingly, four of the abundant species in this study were negatively correlated with

DWD (most significant relationships Fig. 1.2). In other studies, three of these species have been positively associated with DWD in several habitats including mature and post-fire forests and clearcuts. Bradycellus lugubris, a species favouring wet environments, has been found to be positively associated with high volumes of early-decay DWD in clearcut habitats (Pearce et al.

2003), but similar correlations have not been reported in closed-canopy forests. Pterostichus adstrictus, a species known to utilize woody debris for egg and larval development, overwintering, and shelter (Goulet 1974, Larochelle and Lariviere 2003), has been positively associated with high volumes of woody debris in post-fire stands (Cobb et al. 2007) and woody debris of moderate-decay in clearcuts (Pearce et al. 2003). However, P. adstrictus is a habitat generalist with regionally varying habitat preferences and flexibility in oviposition substrates and has displayed a fast rate of development in a range of temperatures and moisture levels (Goulet

1974). Pearce et al. (2003) also found P. adstrictus to be less abundant in wet microhabitats in mixedwood boreal stands, which may explain the negative relationship with DWD, a feature associated with humid microclimate (Harmon et al. 1986). Synuchus impunctatus, also a forest

30 generalist that utilizes woody debris for shelter (Larochelle and Lariviere 2003), has been found to be positively associated with moderate to well decayed woody debris in clearcut habitats

(Pearce et al. 2003), however such results are not necessarily transferable to closed canopy forests. It is likely that DWD may be important for some carabid species only when other habitat features are limiting. For example, Pearce et al. (2003) found that woody debris was more important for carabids in clearcuts than in closed forest. DWD may be less important in the structurally complex habitats of mature forests. Indeed, my own work suggests this; for example, I found that DWD, in the form of slash piles, strongly increased local abundances of carabids in clearcuts (by 167% on average; Chapter 2), but only relatively weakly in the present study (for example, as seen in Fig. 1.1, an increase in the volume of large-diameter, late-decay conifer from 0 to 19 m3/ha increased total carabid abundance on average by 91%). Sphaeroderus stenostomus, a species considered to be a forest specialist and is often found sheltering under logs and overwintering in wood (Larochelle and Lariviere 2003, Bousquet 2010) was found to be negatively correlated with large-diameter, late-decay coniferous DWD. This is a curious result given that this substrate apparently provides ideal conditions for S. stenostomus and indeed was found to be a positive correlate for three other species that in the literature show associations with DWD similar to those reported for S. stenostomus. Furthermore, as described above, it is likely that DWD may be more important for some carabid species only when other habitat features are limiting. Species responding negatively to increasing volumes of DWD may also be influenced by the presence and distribution of their competitors that could be associated with

DWD. For example, carabids were found to be negatively correlated with the presence of

Formica ants in two studies in southern boreal forests in Finland (Niemelä et al. 1992a, Heliola

31 et al. 2001). Until additional manipulative experiments are undertaken, it will be difficult to tease apart such relationships.

Of the four rare species significantly correlated with DWD, two were positively associated and two were negatively associated, and for each species this relationship was consistent for all DWD variables. Each of these four species has been noted to utilize DWD for shelter, and in the case of Carabus maeander and Pterostichus luctuosus, also as an overwintering substrate (Larochelle and Lariviere 2003). Interestingly, C. maeander and P. luctuosus displayed opposite associations (positive and negative, respectively) with three different DWD variables each despite having very similar moist habitat preferences and suggested affinities with DWD, however these species are typically associated with wet environments (e.g., marshes and swamps) and are likely not in their preferred habitat in the relatively drier mixedwood forests (Bousquet 2010). The positive association of C. maeander with both small and large-diameter early-decay conifer and small-diameter, late-decay conifer is consistent with a species utilizing DWD for shelter and overwintering. Amara lunicollis is considered a generalist species that is typically associated with open habitats; however this species positive association with small-diameter, late-decay, conifer DWD is consistent with its reported DWD affinity (Larochelle and Lariviere 2003). Bembidion wingatei, a forest specialist species typically associated with both conifer and mixedwood forests was negatively associated with both early and late-decay, small-diameter conifer DWD. This species is generally only collected in low abundances and no specific affinities with DWD have been reported aside from observational data therefore it is difficult to speculate on this this result. In fact, with the exception of P. luctuosus, all of these species are often reported in relatively low abundances in a variety of forest types in North America (e.g., Werner and Raffa 2000, Pearce et al. 2003,

Klimaszewski et al. 2005, Pearce et al. 2005, Gandhi et al. 2008). In this study these relationships are relatively weak as a result of the rarity of these species (standardized abundances; A. lunicollis is represented by 9.07 individuals [present in 18.5% of sample plots],

B. wingatei is represented by 15.49 individuals [present in 22.2% of sample plots], C. maeander is represented by 12.09 individuals [present in 18.5% of sample plots], and P. luctuosus is represented by 11.83 individuals [present in 14.8 % of sample plots]).

Previous studies have found that carabid species richness and diversity increase with structural complexity such as dead wood (e.g., Fuller et al. 2008). In a woody debris manipulation study in approximately 50-year-old loblolly pine stands, for example, Ulyshen and

Hanula (2009) reported higher carabid species richness and diversity in sites with higher volumes of woody debris. In the present study, I found little evidence of an increase in carabid species richness with increasing volumes of DWD. Abundances, however, were consistently higher in sites with high volumes of coniferous DWD both small and large-diameter and early and late-decay, and were generally higher in sites with late-decay wood of various sizes and species. Large-diameter, late-decay conifer was a particularly good predictor of total carabid abundance. This is consistent with the species-level analysis in this study as seven of the twelve abundant species positively correlated with DWD were found to be specifically correlated with late-decay DWD and four of those seven species were specifically correlated with large- diameter, late-decay coniferous DWD. The presence of coniferous DWD, especially well- decayed large-diameter logs, appears to be an important substrate for carabids.

Unfortunately, not enough is known about individual carabid species life histories and microhabitat requirements to fully understand habitat affinities with DWD. Detailed studies on specific associations of carabids (autecology) with DWD would be very beneficial for further

33 interpretation of these results and in determining how critical DWD is for the survival of carabid species. Study of carabid egg, pupal, and larval stages, the most vulnerable phases due to limited mobility and sensitivity to extremes (Lövei and Sunderland 1996), could provide insight in to the importance of DWD in maintaining viable populations.

The results of this study provide evidence that several species are affected by limited availability of DWD, particularly large-diameter, late-decay logs. These results should be interpreted with caution, however, as it is difficult to isolate the importance of a single habitat feature and there may be other habitat factors that could explain variation in carabid abundances which have not been measured in this study such as soil moisture and pH, vertical stand stratification, litter composition and their potential interconnectivity (e.g., Paje and

Mossakowski 1984, Niemelä et al. 1992b, Antvogel and Bonn 2001). However this study was designed to limit confounding variables by sampling in stands of similar tree species composition and management history (the fact that they are all relatively mature post-logged sites) and by manipulation of DWD. Standing dead wood and stumps were not considered in this study because these forms of dead wood are unlikely to be important for the ground dwelling carabids; at the same time, the latter do provide some woody debris, and hence may have made the DWD relationships that I found weaker.

The results of this study generally support the increasingly accepted recognition of the importance of woody debris in forest ecosystems as limitations in the availability of DWD evidently have strong implications for some carabids in boreal mixedwood forests. Positive correlations of carabid species with various DWD types, particularly large-diameter logs and those in advanced stages of decay, indicate the need to maintain a diverse supply of DWD within managed forests in order to ensure the conservation of these species. Interestingly, over a wide

34 range of volumes, I found that total abundance and abundances of most species increased more- or-less linearly, indicating that management-associated reductions in DWD volumes can be expected to decrease populations. In the context of clearcut harvesting, where the majority of late-decay DWD is destroyed by machinery or exposed to drying conditions (Hautala et al.

2004), leaving aggregated retention patches that are representative of the given forest may serve to protect some existing DWD and provide a source of new DWD through the early stages of stand development. Maintaining larger patches will increase the long-term stability of residual trees and provide larger source habitats. Clearly, the more intensive harvest operations become with respect to features such as decreased rotation lengths and higher fibre removals, the greater will be the reductions in the abundances of DWD-loving species, making the value of residual leave areas, reserves, and protected areas even more important.

Chapter 2: The importance of slash for ground beetles (Coleoptera: Carabidae) in a biomass clearcut

Introduction

Clearcutting is the dominant forest management strategy in Canadian boreal forests and is largely aimed at mimicking natural disturbance (OMNR 2001). One key difference separating clearcutting from natural disturbance, however, is the reduced amount of woody debris that remains after the disturbance (Fridman and Walheim 2000, Siitonen 2001, Pedlar et al. 2002).

Growing concerns over climate change and diminishing petroleum supplies have led to an increased interest in the development of renewable energy sources, including the use of logging residues and stems of low commercial value as fuel. The development of this resource and the associated intensification of biomass harvest may lead to a further reduction in woody debris supplies and a loss of structural complexity in clearcuts. In parts of Europe, such intensive management has resulted in reductions of woody debris by as much as 90 – 98 % relative to levels found in unmanaged stands (Siitonen 2001).

Reductions in woody debris in intensively managed forests have been directly tied to negative impacts on saproxylic flora and fauna (Siitonen and Martikainen 1994, Siitonen 2001).

In Sweden, slash harvesting has been found to result in lower species richness of ground-active beetles in recently clearcut (< 1 year post-harvest) stands (Gunnarsson et al. 2004) and decreases in arthropod abundances over the longer term (15 – 18 years post harvest; Bengtsson et al. 1997).

This is especially concerning in light of the high numbers of red-listed invertebrates (referring to species considered extirpated, endangered, or threatened in an area) that are saproxylic, including beetles, of which an estimated 85 % are red-listed in Sweden (Jonsell et al. 1998). The harvesting of logging residues is a relatively new practice in Canada and therefore it may be

36 possible to minimize negative consequences similar to those observed in Europe. In order to assess potential impacts associated with slash harvesting, it is important to understand its ecological role in clearcuts.

Numerous studies have utilized carabid beetles (Coleoptera: Carabidae) to assess impacts associated with forest management. As a group, they are highly sensitive to habitat conditions and respond rapidly to habitat change (e.g., Niemelä et al. 1993a, Koivula et al. 2002). Structural features, such as leaf litter and logging residues have been found to have a significant influence on carabid abundances (Koivula et al. 1999, Koivula and Niemelä 2003). Slash is a dominant structural feature within clearcuts and, because it is largely composed of woody debris, it is likely an important habitat feature for many carabid species for which woody debris is an important substrate (e.g., Pearce et al. 2003, Latty et al. 2006, Cobb et al. 2007). Pearce et al.

(2003) found strong associations between several carabid species and woody debris in clearcuts of Ontario’s northwestern boreal region, although they did not examine biomass harvest-related slash reductions specifically. Studies in Europe have shown that harvesting slash after clearcutting results in changes in microclimatic conditions and vegetation (Proe et al. 2001,

Astrom et al. 2005, Dynesius et al. 2008) and microhabitat is considered to be among the most important factors in determining carabid assemblages (Thiele 1977). A number of European studies have examined associations of carabids with slash in clearcuts. In a slash-manipulation experiment in boreal Sweden, carabids were found to be significantly more abundant near slash than on bare ground (Nittérus and Gunnarsson 2006). In a comparison of Balsam fir (Picea abies) -dominated boreal clearcuts with and without slash harvest, Nittérus et al. (2007) found that sites with slash harvest had significantly fewer forest carabid species, a guild that is particularly sensitive to clearcutting (Szyszko 1990, Niemelä et al. 1993a, Niemelä et al. 1993b,

Haila et al. 1994, Duchesne et al. 1999, Heliola et al. 2001). In a Finnish study, Koivula and

Niemelä (2003) found that the amount of slash on the ground significantly explained variation in carabid communities. Finally, in a laboratory habitat-choice experiment in which total time spent in either slash or bare ground was measured, the two carabid species used in the experiment

(Pterostichus oblongopunctatus and Carabus hortensis) spent significantly more time in slash than away from slash (Nittérus et al. 2008).

Although carabids evidently respond positively to slash, current knowledge in the context of biomass harvesting is limited to European studies in largely conifer-dominated stands that contained very little post-harvest woody debris. Responses have not been examined for Canadian carabid assemblages or from harvests of mixedwood stands with substantial deciduous components. Clearcuts of mixedwoods in Canada contain comparatively higher volumes of post-harvest woody debris, in part because of lower commercial values of deciduous trees and in part because such cuts typically represent a first rotation of harvesting. For example, c.

5-year-old conifer-dominated clearcuts in Finland were found to contain 15.6 ± 25.1 m3/ha of woody debris > 10 cm in diameter and >130 cm in length (32.7 ± 27.9 m3/ha when all woody debris > 10 cm including stumps was included; Eräjää et al. 2010), whereas mixedwood clearcuts of the same age in Canada were found to contain 111.97 ± 35.14 m3/ha of woody debris

> 10 cm in diameter and > 50 cm in length (including stumps, however stumps comprise a very small portion of the volume; Pedlar et al. 2002). According to island biogeography theory

(MacArthur and Wilson 1967), associations of carabids with slash would presumably be weaker in clearcuts where more woody debris is available compared to clearcuts where such resources are in short supply.

The spatial distribution of habitat patches is considered to be of central importance in the conservation of biodiversity after disturbance (Taylor et al. 1993, Fahrig and Merriam 1994).

Understanding the distribution of carabids in a clearcut could provide insight into better emulating natural disturbance by identifying minimum distances to source habitats required to promote recolonization. For example, there is some evidence that carabid assemblages vary with distance from forest edge in clearcuts. In Finland, Koivula (2002) found that the abundance of forest species in clearcuts increased closer to the mature forest edge; in Canada, Pearce et al.

(2005) noted differences in carabid species composition between 10 and 100 meters away from forest edge, although overall abundance and species richness was similar between the two.

Changes in carabid composition with increasing distance from forest edge have not been examined in the context of biomass harvests.

In this study, I investigate whether slash contributes to maintaining carabid populations in a mixedwood boreal clearcut in northeastern Ontario 2.5 - 3.5 years after a biomass harvest. My specific objectives were to examine differences in local abundances and species richness of carabids near and far from slash piles and as a function of distance from forest edge.

Methods

Study Site

The study took place in a 207-ha clearcut harvested in the winter of 2007/2008 in the

Gorden Cosens Forest Management Unit in the vicinity of Kapuskasing in northeastern Ontario,

Canada (Fig. 2.1). This site represents a model cut implemented by the Forestry Research

Partnership as part of the Enhanced Forest Productivity project that is aimed at developing forestry practices required to maintain and enhance an economically viable forestry industry (Pitt et al. 2008 in litt.). In addition to a conventional full-tree harvest, some of the slash and non-

commercial tree species (combined representing 54 % of the total volume harvested) were chipped and removed from the site for subsequent use as fuel. Clearcut logging in the region is required to leave some live trees after the harvest; in this cut, they were aggregated into small forest islands dispersed throughout the cut of approximately 100 trees for every 4 ha of harvested

40 area (McPherson et al. 2008). To estimate the volume of woody debris ≥ 7 cm in diameter in the site, I superimposed a grid with 300-m spacing onto the cut and undertook line-intercept sampling (Van Wagner 1968) along the grid lines (5,348 m of sampling in total, residual forest islands and stumps were excluded; see Fig. 2.1). During the sampling, woody debris volume was quantified for several different types of ground cover: skid trails, fine-slash piles, coarse slash piles that could be sampled exhaustively at the time, coarse slash piles that required subsequent subsampling, and other. For piles that required subsequent sampling, I returned later and estimated the pile height profile along the transect line and, to estimate volume for the whole profile, exhaustively measured woody debris volume for a subset of the pile profile. Wood volumes for the various cover types (with number of m sampled) were: skid trails 1 m3/ha (126 m), fine-slash piles 75 m3/ha (630 m), originally-sampled slash piles 151 m3/ha (74 m), subsequently-sampled slash piles 308 m3/ha (42 m), and other 62 m3/ha (4476 m). The overall average was 65 m3/ha.

Prior to harvest activities, the site was an upland mixedwood forest of approximately 70 years of age that, according to a pre-harvest cruise by Tembec (Pitt et al. 2009 in litt.), consisted of spruce-pine-fir (84.8 m3/ha), aspen (30.5 m3/ha), white birch-balsam poplar (11.1 m3/ha), and cedar (5.5 m3/ha). Planting took place the summer following the cut across approximately 170 ha with an equal mix of white and black spruce over 134 ha and a further 36.3 ha planted with only black spruce to an average density of 2400 stems per ha. The remaining 36.5 ha, consisting of roads and small patches, was not planted to allow for natural regeneration. Surveys in 2009 indicate that total regeneration (including natural regeneration) consisted of 2879 stems per ha of conifers (mean height 34 cm planted, 53 cm natural) and 4834 stems per ha of hardwood (74% poplar; mean height 65 cm). Vegetation cover was estimated at 46% consisting of moss (25%;

41 mean height 1 cm), tall shrubs (7%; mean height 40 cm), low shrubs (5%; mean height 13 cm), and other vegetation types (combined 9%; mean height 19.25 cm, range 11-25). Aerial spraying of glyphosate-based herbicide FORZA (1.99 kg active ingredient/ha) occurred in August of 2009 and 2011 as a conifer-release strategy (Pitt et al. 2009 in litt.).

Experimental Design

I used a factorial design to investigate two treatments: proximity to slash piles (<5 m or

>83 m away) and proximity to forest edge (34-40, 66-84, or 181-268 m away). I first mapped all of the relatively large slash piles in the cut and then picked two at random at each of the three distances from the forest edge. Similarly, I picked at random six sites that were far from slash piles and met the same distance-from-forest criteria. Sites far from slash piles were on average

126 m from the large slash piles (range 84-160 m). All sites were relatively far from residual islands of live trees (mean = 79 m; range 42 - 163 m). I measured the approximate dimensions of the six slash piles sampled by superimposing an ellipse onto each and measuring the ellipse's major and minor axes and, based on the axes lengths, the ellipse area. Ellipse axes averaged 15.0

(range 11 - 20) and 12.8 (range 9 - 14) m, respectively, with slash pile area averaging 129 (range

86 - 198) m2. Estimated slash pile height averaged 1.6 (range 0.40 - 2.25) m.

Carabid Sampling

Carabids were sampled by use of pitfall buckets in combination with drift-fences formed into “Y”-shaped arrays. Drift fences were built of polypropylene geotextile ("green line") with one pitfall bucket placed at the end of each of the three arms and one in the centre of the Y (i.e.,

4 buckets per array). Each arm consisted of a piece of geotextile that was 0.5 m high and 3.65 m long, with 10 cm of the geotextile buried in the soil. When arrays were placed near slash piles, they were oriented such that two of the arms were close to the slash pile (1-3 m distant). Buckets

42 were 15 cm in diameter by 15 cm deep and were buried flush with the ground surface and the end(s) of the pitfall arms. During trapping, the buckets were filled with 5 cm of a 5 % saline solution to which a small amount of soap had been added to break surface tension. Carabids were sampled three times: 26 – 31 August, 2010 and 15 – 20 June and 18 – 23 August, 2011. During each sampling session, traps were active for five consecutive nights; buckets were covered with lids and soil at other times. Specimens were collected at the end of each five-night session and preserved in 70 % ethanol. Carabids were identified to species using keys from Lindroth (1961,

1963, 1966, 1968, 1969a, 1969b) and Bousquet (2010). Nomenclature follows Bousquet (2010).

A voucher collection was authenticated by H. Goulet at the Canadian National Collection,

Ottawa, and is located at the Faculty of Forestry, University of Toronto, Ontario.

Prior to statistical analysis, carabid abundances were standardized to account for missing effort by calculating the number of individuals per 100 bucket-nights. Missing effort (which was due to bears and other vagaries of field sampling) represented 11 % of end buckets and 17 % of center buckets in August 2010, 3 % of end buckets and 0 % of center buckets in June 2011, and no missing effort in August 2011. In order to calculate effort, I assumed that a centre bucket was

2.236 times as effective as an end bucket (see Chapter 1).

Statistical Analyses

Analyses were conducted based on means across the three sampling sessions and means were ln(x + 1) transformed to better meet the assumptions of normality and homogeneity, which were evaluated from plots of residuals in species-specific tests (see below). A principal component analysis (PCA) was performed on the covariance matrix of ln-transformed carabid abundances to examine overall patterns of community variation; in addition, I used redundancy analyses to test for effects due to slash and forest proximity, and their interaction.

Univariate tests (Analysis of Covariance) were undertaken for relatively abundant species

(those occurring in ≥ 50 % of samples) and for total carabid abundance and species richness. For rarer species, I classified abundances as above or below the species-specific median, and undertook contingency table analyses on counts in the two categories. In these analyses, distances from forest edge were classified into three classes (near, medium, and far).

To compare species richness while controlling for abundances, I undertook rarefaction analyses (before standardization for sampling effort). Rarefaction estimates were calculated using the formula for individual-based rarefaction (mean and variance) from Coleman et al.

(1982). Multivariate procedures were undertaken in Canoco for Windows (v. 4.5); other analyses were in SAS (v. 9.2).

Results

A total of 432 individuals from 18 genera and 41 species were collected (Table 2.1; Rank abundance curve: Appendix 6). Of these, 29 species are endemic to North America, 10 are

Holarctic, and 2 are introduced with European origin (Clivina fossor and Pterostichus melanarius; Bousquet 2010). The most abundant species (which occurred in ≥ 50% of samples) were: Agonum gratiosum, Agonum retractum, Bradycellus lugubris, Bradycellus nigrinus,

Platynus decens, Pterostichus coracinus, and Scaphinotus bilobus. The sample was dominated by open habitat and generalist species, with only four forest specialists (as described by Lindroth

1966, Larochelle and Lariviere 2003): Agonum retractum, Pterostichus pensylvanicus,

Sphaeroderus nitidicollis, and Sphaeroderus stenostomus.

Standardized mean abundance Near slash Away from slash

Total mean % of total Species Near Medium Far Near Medium Far abundance captures Agonum cupripenne (Say, 1823) 2.55 0 2.55 0 0 0 5.09 0.89 Agonum gratiosum (Mannerheim, 1853) 16.55 12.73 33.10 0 6.37 3.15 71.90 12.51 Agonum retractum LeConte, 1846 8.91 12.73 15.28 1.27 3.82 3.15 45.16 7.86 Agonum sordens Kirby, 1837 0 0 0 0 0 1.27 1.27 0.22 Agonum thoreyi Dejean, 1828 0 0 1.27 0 0 0 1.27 0.22 Amara erratica (Duftschmid, 1812) 0 1.27 1.27 0 1.27 0 3.82 0.66 Amara lunicollis Schiødte, 1837 0 2.55 1.27 0 0 1.57 5.39 0.94 Amara otiosa Casey, 1918 0 0 0 0 1.27 0 1.27 0.22 Amara patruelis Dejean, 1831 0 0 2.55 0 1.27 0 3.82 0.66 Anisodactylus harrisii LeConte, 1863 0 0 1.27 0 0 0 1.27 0.22 Badister obtusus LeConte, 1878 0 0 1.27 0 0 0 1.27 0.22 Bembidion fortestriatum (Motschulsky, 1845) 0 0 1.27 0 0 0 1.27 0.22 Bembidion grapii Gyllenhal, 1827 0 1.27 0 0 0 0 1.27 0.22 Bembidion mutatum Gemminger & Harold, 1868 1.27 0 0 0 0 0 1.27 0.22 Bembidion occultator Notman, 1919 0 0 1.27 0 0 0 1.27 0.22 Bradycellus lecontei Csiki, 1932 0 0 0 0 0 1.27 1.27 0.22 Bradycellus lugubris (LeConte, 1847) 8.91 3.82 2.55 6.37 3.82 12.29 37.75 6.57 Bradycellus nigrinus (Dejean, 1829) 5.09 3.82 0 7.64 11.46 4.12 32.13 5.59 Calathus ingratus Dejean, 1828 0 1.27 0 0 1.27 0 2.55 0.44 Chlaenius lithophilus Say, 1823 3.82 0 0 0 0 0 3.82 0.66 Clivina fossor (Linné, 1758) 0 0 0 0 1.27 0 1.27 0.22

Table 2.1. – continued

Harpalus fulvilabris Mannerheim, 1853 0 0 1.27 0 0 0 1.27 0.22 Harpalus laevipes Zetterstedt, 1828 3.82 1.27 2.55 0 0 0 7.64 1.33 Harpalus somnulentus Dejean, 1829 2.55 0 0 0 2.55 1.27 6.37 1.11 Notiophilus aquaticus (Linné, 1758) 0 0 0 1.27 0 0 1.27 0.22 Patrobus foveocollis (Eschscholtz, 1823) 1.27 0 0 0 0 1.57 2.85 0.50 Patrobus stygicus Chaudoir, 1871 0 0 0 0 1.27 0 1.27 0.22 Platynus decens (Say, 1823) 2.55 2.55 8.91 0 1.27 1.57 16.85 2.93 Platynus mannerheimii (Dejean, 1828) 0 1.27 0 0 0 0 1.27 0.22 Poecilus lucublandus (Say, 1823) 0 2.55 3.82 0 0 0 6.37 1.11 Pterostichus adstrictus Eschscholtz, 1823 11.46 2.55 11.46 0 0 0 25.46 4.43 Pterostichus commutabilis (Motschulsky, 1866) 0 0 1.27 0 2.55 0 3.82 0.66 Pterostichus coracinus (Newman, 1838) 19.10 73.85 38.01 15.31 28.01 8.59 182.87 31.81 Pterostichus luctuosus (Dejean, 1828) 0 11.46 6.37 0 0 0 17.83 3.10 Pterostichus melanarius (Illiger, 1798) 0 2.55 6.37 0 0 0 8.91 1.55 Pterostichus pensylvanicus LeConte, 1873 5.09 5.09 16.07 1.27 1.27 0 28.80 5.01 Pterostichus punctatissimus (Randall, 1838) 0 0 5.88 0 0 0 5.88 1.02 Scaphinotus bilobus (Say, 1823) 3.82 0 8.43 2.55 2.55 4.77 22.11 3.85 Sphaeroderus nitidicollis Guérin-Méneville, 1829 0 0 0 0 1.27 1.27 2.55 0.44 Sphaeroderus stenostomus Dejean, 1826 0 0 2.55 0 0 2.22 4.77 0.83 Trechus apicalis Motschulsky, 1845 0 0 1.27 0 0 0 1.27 0.22 Total abundance 96.77 142.60 179.15 35.68 72.57 48.10 574.88 100.00 Species richness 15 17 26 7 17 14

More than 73 % of all individuals were captured at traps near slash piles (Table 2.1). The first axis from the principal component analysis revealed clear separation between traps near slash piles (mostly to the right of the origin) and those away from slash piles (to the left of the origin), with the first two axes explaining respectively 30 and 24.5% of the total variance (Fig.

2.2). The great majority of species vectors also pointed to the right, indicating relatively greater abundances of most species near slash piles. In fact, eighteen carabid species were captured exclusively at slash piles, although ten of these were represented by only a single individual.

Conversely, seven species were captured exclusively far from slash piles, six of which were represented by only a single individual. In the redundancy analysis examining the effects of proximity to slash piles and proximity to forest, the former was highly significant (p = 0.0016), but the latter was not (p = 0.94). I found little evidence of interaction between the two (p > 0.45).

ANCOVAs on total abundance and species richness also found little evidence of interaction effects (p = 0.89 and 0.92, respectively), so I dropped the interaction terms and tested only main effects. As suggested by the PCA, total abundance was significantly higher near than far from slash piles (p = 0.0251) however species richness was not (p = 0.0592). I found little evidence that distance-from-forest-edge had an effect on carabid abundance or species richness when all samples were included in analysis (ANCOVA ps > 0.1). However, when only samples located at slash piles were analyzed (n = 6) using regression analysis, there was some indication of a positive distance effect on total carabid abundance (p = 0.0264; Fig. 2.3), but not when only samples far from a slash piles were tested (p = 0.64). Distance-from-forest effects on species richness were not significant (p = 0.11 and 0.45, respectively) in the same tests.

At the species level, ANCOVA revealed that of the seven abundant species only the forest specialist, Agonum retractum, was significantly more abundant near than far from slash

Agongrat 1.0 Platdece Agonthor Bembfort Trecapic Pterluct Bemboccu Agonretr Poeclucu Ptercomm Bradlugu Ptermela totabun Amarluni Calaingr Platmann Richness Sphaniti Bembgrap Ptercora Amarotio Bradnigr Sphasten Patrstyg Patrfove Chlalith Agonsord Clivfoss Pteradst Bradleco Notiaqua Amarerra Bembmuta Agoncupr Anisharr Harpsomn Amarpatr Pterpunc Badiobtu Harplaev Harpfulv

Scapbilo Pterpens -1.0 -0.6 1.0

A: y = 0.0037x + 3.8517 p = 0.0264 5 4 3

2 B: y = 0.0027x + 2.6431 p = 0.6387 1

ln(abundance + 1) + ln(abundance 0 0 100 200 300

Distance from forest edge (m)

piles (p = 0.0342). Fisher’s exact tests revealed that slash pile affinity was significant only for

Pterostichus adstrictus (p = 0.0152). Three other species (Poecilus lucublandus, Pterostichus luctuosus, and Pterostichus melanarius) were captured exclusively near slash piles and were close to significant (ps = 0.0606). Interestingly, 19 of the 21 individuals of P. pensylvanicus were captured near slash piles, although the species was recorded in only 5 traps. I found little evidence of distance-from-forest effects at the species level (ANCOVA ps > 0.13; Fisher’s exact tests ps > 0.05).

Rarefaction indicated that the species accumulation rate was nearly identical for the two treatments (Fig. 2.4 A). Rarefaction examining the three distances from forest edge indicated a general trend of samples located closest to the forest edge having fewer species and individuals than the samples located farther into the clearcut (Fig. 2.4 B). This trend was also evident when samples from the near-slash and far-from-slash were analyzed separately (Fig. 2.4 C).

40 40 30 A. B. C. 25 30 30 20 20 20 15 10 10 10 5

Number of species of Number 0 0 0 0 100 200 300 0 60 120 180 0 30 60 90 120 150

Number of individuals

Discussion

The significantly greater abundance of carabids close to slash piles suggest that slash is an important structural feature supporting carabid populations in clearcuts, even in cuts with relatively high volumes of woody debris (65 m3/ha in this case). This supports results from a

Swedish slash manipulation study that found significantly higher abundances of carabids at slash compared to away from slash in a relatively young (< 1 year post-harvest) conifer dominant clearcut (Nittérus and Gunnarsson 2006). This positive association with slash is not surprising given that slash piles are a dominant structural feature in clearcuts providing shelter and sources of prey for carabids, whose small-scale distribution has been suggested to result from active microhabitat selection (Niemelä et al. 1992). Slash presumably provides relatively cooler and moister conditions compared to the surroundings, and may act as a buffer during periods of extreme microclimatic conditions (Dynesius et al. 2008).

Gunnarsson et al. (2004) suggest that local accumulations of woody debris resulting from natural disturbances such as fire and insect outbreaks may have been an important resource for beetles even before the time of forest management. Many carabid species have reported associations with woody debris as oviposition and overwintering habitats (e.g., Goulet 1974,

Bousquet 2010) and mounds of woody debris, as found in slash piles, likely provides important habitat in a landscape largely denuded of this feature. My findings also corroborate those of

Pearce et al. (2003), who reported strong associations between many carabid species and woody debris in northwestern Ontario boreal clearcuts.

I found little evidence that slash supported higher carabid species richness than the surroundings when species richness was controlled for abundance via rarefaction. Rarefaction curves based on samples near slash and away from slash depicted almost identical species

51 accumulation rates (Fig. 2.4 A). This suggests that the higher abundances near slash piles were not the result of beta diversification (i.e., species specialized to certain microhabitats associated with woody debris), but instead reflected a general amelioration of harsh conditions that was valuable for virtually all species.

The most sensitive species to disturbance are likely those with specific habitat requirements and limited dispersal capabilities (den Boer 1981) such as forest specialist species.

Species-level analysis revealed that Agonum retractum and Pterostichus adstrictus were significantly more abundant near slash piles than away from slash. Both of these species have been found to occur in lower abundances in clearcuts than in adjacent unharvested stands 7-9 years after harvest in northwestern Ontario’s boreal mixedwood region (Duchesne et al. 1999).

Agonum retractum is considered a forest specialist (Lindroth 1966); the shady and humid microclimate found in slash piles may have more closely resembled conditions found in closed forests. These results provide some indication that slash may play a role in sustaining populations of forest species in clearcut habitats 2.5 – 3.5 years after harvest. Interestingly, there were no significant differences in abundances of A. retractum at the three distances from forest edge and it is unlikely that the individuals caught were wanderers from the intact forests bordering the site given that A. retractum likely has limited dispersal capabilities as it is wing-dimorphic and even long-winged individuals may not readily fly (Carter 1976).

Furthermore, forest species have been found to rarely enter open habitat and if they did, they were restricted to 20 m away from the forest edge (Koivula et al. 2004). Since carabids typically live 1 - 2 years (Lövei and Sunderland 1996) it is possible that these individuals are from new generations and not remaining individuals from before disturbance. This suggests that slash may

52 support forest species in early stand development stages, which is consistent with the findings of

Nittérus et al. (2007) which suggest that this may be the case even 5 -7 years after harvest.

It is not surprising that the abundance of Pterostichus adstrictus, considered a generalist species that is often found to prefer open habitat (Lindroth 1966, Larochelle and Lariviere 2003), was higher near slash due to this species affinity with dead wood. Adults are often found ovipositing under the bark of wood and this substrate is also utilized by its three larval instars and pupae for protection from temperature and desiccation (Goulet 1974). P. adstrictus abundances were also found to be highly correlated with logging residues in a Finnish boreal clearcut 1.5 years after harvest (Heliola et al. 2001). I found little evidence that the abundance of

P. adstrictus varied with distance from forest edge; however, Heliola et al. (2001) noted a decrease in abundance from the center of a boreal clearcut to forest edge and found that this species did not enter the adjacent forest at all. Interestingly, P. adstrictus has sometimes been classified as a forest specialist (e.g., Duchesne et al. 1999, Pearce et al. 2003) and its abundance was found to be much lower in clearcuts than in adjacent intact forests 7 – 10 years after harvest in northwestern Ontario (Duchesne et al. 1999). Both Agonum retractum and Pterostichus adstrictus have been positively associated with woody debris in clearcuts of northwestern

Ontario (Pearce et al. 2003). Pearce et al. (2003) suggest that because these species are spring to early summer breeders, their larval stage occurs in the summer when they are at most risk of desiccation, and hence that they may seek structural features, such as dead wood, that provide moister conditions.

The typical response of carabids to clearcutting is an increase in species richness driven by the colonization by open-habitat species (e.g., Niemelä et al. 1993a, Duchesne et al. 1999,

Heliola et al. 2001, Koivula et al. 2002). The high amount of light and higher temperatures in

53 clearcuts relative to closed forests provides suitable conditions for these species (Thiele 1977) that arrive within months after logging (Haila et al. 1994, Koivula and Niemelä 2003). Forest species are generally negatively affected by forest management. In Finland, managed boreal forests contained fewer forest specialist saproxylic beetle species compared to unmanaged forests (Koivula et al. 2002) and reductions in carabid forest species abundances have been observed after clearcutting (Szyszko 1990, Haila et al. 1994, Duchesne et al. 1999, Heliola et al.

2001). Losses of some forest specialists have been reported within two years after logging in western Canada (Niemelä et al. 1993a, Niemelä et al. 1993b) and several forest species did not recolonize regenerating stands even more than 20 years after harvest (Niemelä et al. 1993b).

Clearcuts with biomass harvest may have further negative effects; for example, Nittérus et al.

(2007) found that 5 - 7 years after harvest, clearcuts with slash removal had fewer forest carabid species than clearcuts without slash removal and suggested that slash harvesting may have long-term effects on carabid compositions. Here, I obtained evidence that this relationship also holds for open-loving species as total carabid abundance was significantly higher at slash piles.

This is consistent with the suggestion of Gunnarsson et al. (2004) that retaining some slash may provide a degree of buffering for sustaining individuals and species of ground active beetles in clearcuts that have been subjected to slash harvest. Slash may be particularly important in maintaining carabid species in relatively young clearcuts before other structural elements such as vegetation develop and may play a role in maintaining populations during stand development.

The response of A. retractum (and potentially P. adstrictus depending on which classification is followed) in the present study is in line with these observations suggesting that retaining slash may serve to sustain forest species at least 2.5 – 3.5 years after harvest. It should be noted that generalizations cannot necessarily be made for all forest species as some have been found in 5 –

7 year old clearcut sites where slash had been intensively harvested indicating that slash is not the only influencing factor (Nittérus et al. 2007). However as seen in the present study, even clearcuts with slash harvest have some residual slash that may support forest species.

I found little evidence that carabid species richness or total abundance varied with distance from forest edge when all samples (i.e., near and away from slash) were included.

Similarly, I found no significant relationship between the abundances of individual species and distances from forest edge. These results generally agree with Pearce et al. (2005) who, in a northwestern Ontario clearcut 5 - 10 years post-harvest, found that carabid species richness and overall abundance was similar 10 and 100 meters away from forest edge, although they found some differences in carabid species composition. However, I found that carabid abundance and species richness was generally higher in samples further away from forest edge. This trend can be interpreted in two ways. Presumably, structural complexity such as that provided by slash piles, would become increasingly important with increasing distance from forest edge; especially in supporting species with limited dispersal capabilities and that require cool and moist microhabitats. I found that when samples near slash piles were analyzed alone, there was a significant increase in total carabid abundance with distance from edge. Although the sample size was small, this result may indicate that slash is more important to carabids further from forest edge where conditions may be harsher than nearer the relative shelter of forest edge. It has been suggested that the shading provided by adjacent forest would make clearcut edges a less severe environment when compared with areas farther into the cut (Pearce et al. 2005) and the shelter, in the form of large trees, provided by adjacent stands may help forest species survive in small clearcuts (Koivula 2002). On the other hand, one would expect a decrease in species abundances with increasing distance to source habitats. There is some evidence of this in the

55 literature; for example, Koivula (2002) found that the abundance and distribution of carabids was negatively affected with increasing distance to the nearest source habitat. However this would likely be a more important factor for forest species as open habitat species and generalists are presumably colonizing from nearby fields etc. Interestingly, Koivula et al. (2004) found that forest species did not venture more than 20 m into open habitats, which is somewhat closer to the forest than the traps that I placed "close" to the forest (34-40 m). Distance from forest edge effects may also be weak in my study because the sample was dominated by open habitat and generalist species which would presumably be less sensitive to distance from forest edge.

The diverging rarefaction curves of samples at medium and far distances from forest edge

(Fig. 2.4 B) indicate the possibility that further sampling may reveal a more distinct difference in species richness. However, because the curve for samples near forest edge is asymptotic it may indicate that fewer carabid species prefer proximity to forest edge in clearcuts. It has been noted that there do not appear to be any edge specialists among boreal carabids in both Finland

(Heliola et al. 2001) and Canada (Spence et al. 1996) and open habitat species very rarely penetrated forest interior from clearcuts (Heliola et al. 2001).

Although slash pile size was not controlled for in this study, I found little evidence of slash pile area or height effects. Specifically, when I included these two variables in a multiple regression model of ln-transformed total abundance against distance from forest edge, neither was significant (stepwise test; ps > 0.15). Interestingly, Gunnarsson et al. (2004) found a positive correlation between ground-dwelling beetle abundance and slash pile height in cuts with slash harvest, but not in those without slash harvest. Nittérus and Gunnarsson (2006) also found a significant positive relationship (for carabid abundance specifically, in this case) in reference plots of a slash manipulation experiment. Both of these studies were based on relatively small

56 slash piles (maximum height 0.35 m and 0.42 m, respectively) compared to mine (height range

0.40 – 2.25m ). The positive correlation with height found for small slash piles, but not large piles, may indicate a threshold response. Microclimate amelioration effects may increase with slash pile size up to a certain threshold, but not above. Thus, a positive relationship may be found among piles of various sizes below this threshold, but not above it. Unfortunately, the two studies that found positive correlations with slash pile height also included samples with no slash cover (indicated as 0 cm height), which, given the positive associations of carabids with slash, may have confounded their results.

As a conifer-release strategy, the site was aerially sprayed with herbicide less than a month before the sampling in August 2011. It is not known if this affected carabid captures; however, other evidence suggests that it did not. In northwestern Ontario’s boreal mixedwood region, Duchesne et al. (1999) did not find differences in either total carabid abundance or species-specific abundances when comparing sites treated with two different herbicides and untreated cuts. No visible effects of the herbicide on vegetation were evident during the time of trapping (pers. obs.).

These results have important implications for forest management as it is evident that slash is an important structural feature in maintaining abundances of carabids in clearcuts where this resource is limiting. My results suggest that slash may serve as refugia for some forest species which have a higher conservation value because they comprise a small portion of carabid species in boreal landscapes (Niemelä 1993). Slash may be filling the role of structure and woody debris associated with natural disturbances (Bengtsson et al. 2000, Gunnarsson et al.

2004). In order to better meet the goals of mimicking natural disturbance and to promote the recovery of carabids, my results suggest that it is necessary to maintain sufficient quantities of

57 slash in clearcut landscapes. Interestingly, even though the clearcut that I studied was the site of biomass harvesting, the volume of woody debris that I found was quite high (woody debris ≥ 7 cm = 65 m3/ha); however, this amount of post-harvest woody debris is consistent with a biomass cut in that Pedlar et al. (2002) reported volumes of 111.97 ± 35.14 m3/ha (woody debris > 10 cm in diameter and > 50 cm in length) in traditionally-harvested mixedwood clearcuts of northwestern Ontario. Presumably, the value of slash to carabids would be even stronger in cuts from which even more woody debris had been removed, provided that such cuts still offer habitat for the full complement of species. Because slash piles were widely dispersed in the cut that I studied, and hence did not likely provide source habitats for the areas sampled far from slash, strikingly lower abundance and richness of carabids in the sites far from slash suggests that the woody debris volumes in these areas was a limiting feature for the populations of many species.

Further study will be necessary in order to determine critical amounts of slash and its spatial distribution required to maintain carabid populations over the long term.

General Conclusions

Woody debris is recognized as an important habitat feature for numerous species and plays a role in supporting biodiversity by maintaining the structural heterogeneity found in unmanaged forests (Hansen et al. 1991). It is increasingly accepted that policies aimed at maintaining woody debris supplies over time must be incorporated into forest management as a principle strategy to sustain ecological function (Graham et al. 1994, Lee et al. 1997, Sturtevant et al. 1997, Hagan and Grove 1999, Work et al. 2004, Langor et al. 2008). The findings of my thesis highlight the importance of woody debris in managed forests by demonstrating that management-related reductions in woody debris can have negative consequences for forest biota.

Carabids proved to be a good study organism by exhibiting many qualities of an effective indicator group. In particular, they demonstrated sensitivity to woody debris availability in both closed canopy forests and recent cutovers, and provided evidence of niche partitioning among various types of wood. As an abundant and largely predatory group, it is likely that impacts to carabids will resonate through the ecosystem (Lang et al. 1999).

In closed-canopy forests, carabids displayed significant relationships with volumes of a variety of woody debris types, ranging in decay stages, species, and size classes, and in particular with the volume of large-diameter, late-decay coniferous wood. Abundances of numerous species, as well as total carabid abundance, displayed positive linear relationships with manipulated woody debris volumes indicating that reductions in woody debris availability have the potential to result in reduced population sizes and cause regional extirpations. The importance of large-diameter, late-decay coniferous woody debris to carabids is of particular concern given that this resource is most likely to become limiting in managed forests of boreal

Ontario. Large conifers are the primary target of harvest operations and pre-existing late-decay

59 wood is often destroyed by machinery or exposed to drying conditions following logging operations (Hautala et al. 2004). From a management perspective, these results indicate the need to maintain a diverse supply of woody debris over time. Various management strategies have the potential to achieve these goals; for example, rotation ages that are sufficiently long to allow for the development of the diverse supply of woody debris characteristic of unmanaged stands

(Hansen et al. 1991, Rouvinen et al. 2002); retention of patches of representative forest within a cut to provide new sources of woody debris and to protect existing late-decay wood; and multi- cohort management (Bergeron et al. 2007).

Carabids also displayed a strong association with woody debris in the form of slash piles in clearcuts. Slash may be functioning as a substitute for both structural and woody-debris specific features associated with natural disturbances that carabids are adapted to (Bengtsson et al. 2000, Gunnarsson et al. 2004). Clearcuts, at least in early development stages, contain a relatively diverse array of carabid species with varying ecological associations (Koivula and

Niemelä 2003). The strong affinity of most species with slash piles, including open-habitat species, suggests that excessive harvesting of logging residue could have widespread, negative ecological consequences for virtually the entire community. These potential impacts in the early stages of forest development may determine subsequent responses as the forest matures. My results also provided some indication that slash piles were especially important in supporting high carabid abundances far from forest edges; however, the sample size was very small. Further study of the effects of slash pile distributions within clearcuts may yield important management guidelines given that the spatial distribution of habitat patches is considered to be of central importance in the conservation of biodiversity after disturbances (Taylor et al. 1993, Fahrig and

Merriam 1994).

Of the 20 abundant carabid species, four showed negative relationships with woody debris volumes. Three of these appear to be more typical of open habitats, and their negative relationships may reflect the increasingly mesic conditions associated with high wood volumes.

For example, Pterostichus adstrictus was less abundant in wet than dry microhabitats in mixedwood boreal stands of northwestern Ontario (Pearce et al. 2003). Another explanation is that these species are responding negatively to the presence of competitors. For example, carabids have been negatively correlated with the presence of Formica ants in two separate studies (Niemelä et al. 1992, Heliola et al. 2001). I also found that the abundance of

Sphaeroderus stenostomus, a forest species negatively associated with large-diameter, late-decay conifer wood was negatively correlated with the abundance of Pterostichus melanarius, a widespread introduced species that was positively associated with the same woody debris type

(see below). Gastropods, which are often found in the moist environments under logs and bark of woody debris (Savely 1939, Bohan et al. 2000, Kappes et al. 2006), are an important prey for both species. A possibility is that P. melanarius is responding positively to increased prey abundances associated with high woody debris volumes and outcompeting S. stenostomus in the process. Unfortunately, it is notoriously difficult to tease apart such relationships without further study, especially manipulative experiments. It is also possible that woody debris may be a more important substrate where other structural features are limiting. For example, although negatively correlated with woody debris volumes in closed-canopy sites, P. adstrictus was significantly more abundant near slash piles in the clearcut (Chapter 2). Similarly, although Agonum retractum showed little evidence of an association with woody debris in closed-canopy forests, it was also significantly more abundant near slash piles in the clearcut. Unfortunately, not enough is known about carabid life histories and microhabitat requirements to fully understand affinities

61 with woody debris. Studies of egg, larval, and pupal stages may be especially beneficial given the limited mobility of these life stages and their sensitivity to extremes in microclimatic conditions and to desiccation (Lövei and Sunderland 1996).

In total, 60 species of carabids were collected, with 24 being in common to both studies.

Of the 60, 42 are endemic to North America, 13 are Holarctic, and 5 are non-native European species (Amara eurynota, Carabus granulatus, Carabus nemoralis, Clivina fossor, and

Pterostichus melanarius). Of these non-native species, only P. melanarius was abundant. For four of the five species, all individuals captured were from four sites which are located within 20 km of the town of Kapuskasing; the exception was P. melanarius for which 87% of captures were within this distance. Although these species are capable of invading relatively undisturbed habitats such as these closed-canopy stands, this finding is consistent with expectations that introduced species will more often be found colonizing disturbed areas, especially those close to human habitations (Spence and Spence 1988, Niemelä and Spence 1991, Niemelä 1997, Niemelä and Spence 1999, Niemelä et al. 2002, Werner and Raffa 2003). P. melanarius has been found to be associated with human habitation both in Canada and its native range in Europe (Spence

1990, Niemelä and Spence 1991).

Perhaps equally interestingly, four of the five non-native carabids in the closed-canopy study were highly associated with only one of the nine sites. Respectively, 92 % and 87 % of C. fossor and P. melanarius were captured at this site and C. granulatus and C. nemoralis were exclusively captured at this site (although they were represented by only one and two individuals, respectively). Unfortunately, it is unclear what makes this site a non-native carabid hot-spot. The only obviously unique characteristic of the site was a small stream located between two of the site’s plots; however, this stream was > 200 m away from the closest plot. The habitat

62 preferences of three of these four species (C. granulatus, C. fossor, and P. melanarius) includes moist areas such as banks of water bodies, which may provide some explanation as none of the other study sites were as close to a water body (Niemelä and Spence 1991, Bousquet 2010).

Another possible explanation could be the sites proximity to the Kapuskasing dump (c. 5 km), that may have received waste nursery stock containing non-native carabid species (see below).

P. melanarius is thought to have arrived in Canada from Europe in the ballast of commercial ships (Lindroth 1957) and in nursery stock (Spence and Spence 1988). The first

Canadian record of P. melanarius was in Nova Scotia in 1926 (Majka et al. 2007, Bousquet

2010). As an efficient disperser, largely through flight, this species is now widespread in Canada in a variety of habitats ranging from urban areas to intact forests (Niemelä and Spence 1991,

Niemelä and Spence 1999, Bourassa et al. 2011). Some evidence suggests that P. melanarius can have negative impacts on native carabids. Spence and Spence (1988) observed a strong negative correlation between the abundances of P. melanarius and the native Pterostichus adstrictus in a variety of disturbed urban habitats in western Canada. Predation of P. adstrictus larvae by both adult and larval P. melanarius has also been observed in laboratory experiments (Currie and

Digweed 1996). Similarly, in a laboratory experiment of predator-prey and competitive interactions, the survival of adult P. adstrictus was found to be 35 % lower in the presence of P. melanarius, which was attributed to predation as opposed to competition for food (Currie et al.

1996). Niemelä and Spence (1991) did not find similar negative associations of P. melanarius with the native carabid fauna in riparian deciduous forests around Edmonton, Alberta, where, at the time of trapping, P. melanarius was the second most abundant carabid species. They attributed this lack of negative impacts to three possible explanations: 1) P. melanarius may have been colonizing an empty niche in these riparian forests, 2) food was not limiting and therefore

63 there was no resource competition, and 3) such negative interactions may be more common in the anthropogenic than forested habitats for reasons not understood. I undertook Spearman’s correlation analyses using total carabid abundance (excluding abundance of P. melanarius), species richness (excluding P. melanarius), and the standardized total captures of each of the 20 abundant species (represented in at least 50% of samples; using the sessions representing each species peak abundance; see Chapter 1) for the 27 closed-canopy plots. Total carabid abundance was actually positively correlated with the abundance of P. melanarius (p < 0.001) and I found little evidence of correlation with species richness (p > 0.5). Species-level analyses revealed that abundances of Platynus decens, Pterostichus adstrictus, and Pterostichus coracinus were positively correlated with the abundance of P. melanarius (p = 0.0052, 0.0279, and 0.0048, respectively) and abundances of Sphaeroderus nitidicollis and Sphaeroderus stenostomus were negatively correlated with that of P. melanarius (p = 0.0097 and 0.0349, respectively).

Interestingly, the positive correlation of P. adstrictus with P. melanarius is contrary to the above-mentioned literature. The negative correlations of S. nitidicollis and S. stenostomus (and

Scaphinotus bilobus, although not quite significant [p = 0.0894]) with P. melanarius are also an interesting result. These two (three including S. bilobus) species are gastropod feeding specialists and may be suffering from competition from P. melanarius, an important predator of slugs

(Symondson et al. 1996, Symondson et al. 2002, Oberholzer and Frank 2003). Slugs have been found in lower abundances in the presence of P. melanarius and, at least in Europe, display predator avoidance behaviour in responses to chemical cues from this carabid (Bohan et al. 2000,

Oberholzer and Frank 2003, Armsworth et al. 2005). My study was not designed to look at the effects of P. melanarius on native carabid fauna and is likely biased by the woody debris manipulation (for example, abundances of P. decens, P. coracinus, and P. melanarius were all

64 positively correlated with volumes of the same woody debris type while abundances of P. melanarius and S. stenostomus were oppositely correlated with volumes of the same woody debris type). It would be interesting to undertake future study to examine impacts resulting from the expansion of this species in boreal mixedwoods.

In conclusion, my research affirms the need to ensure a diverse and abundant supply of woody debris in order to maintain carabid populations in managed forests. Results also highlight the importance of retaining sufficient woody debris after harvest. Clearly, the more intensive harvest operations become (such as through decreased rotation lengths, more intensive fiber removal, and less residual leave) the greater are the expected reductions in abundances. Such considerations highlight the value of management regimes that better emulate natural disturbances and of protected areas.

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Appendices

Appendix 1. Effective length of drift fence arms for carabid captures in a three-arm drift fence pitfall array.

Presumably, if the location where a beetle encounters a drift fence is too far from the associated bucket, then the animal is unlikely to follow the fence all the way to the bucket. Thus, one can imagine that a drift fence arm has some effective length beyond which increasing the length of the arm does not result in greater capture rates in a bucket. Calling the radius of a bucket r and the effective length of the arm x, then a bucket with one arm would have a "capture perimeter" of (2πr + 2x); namely, the bucket perimeter plus two times x (i.e., both sides of the arm). Similarly, a bucket with three array arms would have a capture perimeter of (2πr + 6x). Knowing that a centre bucket captured 2.236 times as many individuals as an end bucket and in my case that r = 7.5 cm, we have 2.236 · (2π · 7.5 + 2x) = (2π · 7.5 + 6x). Solving for x, we find that x = 0.38 m. This number would suggest that from a carabid perspective, a three-arm pitfall array with each arm a length of 2 · 0.38 m = 0.76 m would be as effective as one with arm lengths of 3.65 m, as used in the present experiment.

Session % of June August June August total Species 2010 2010 2011 2011 Total captures Agonum fidele Casey, 1920 0 0 1 0 1 0.01 Agonum gratiosum (Mannerheim, 1853) 50 7 208 7 272 2.34 Agonum melanarium Dejean, 1828 0 0 1 0 1 0.01 Agonum mutatum (Gemminger & Harold, 1868) 1 0 0 0 1 0.01 Agonum retractum LeConte, 1846 339 16 546 3 904 7.79 Agonum sordens Kirby, 1837 53 1 217 8 279 2.40 Agonum superioris Lindroth, 1966 1 0 3 0 4 0.03 Amara eurynota (Panzer, 1797) 0 0 1 0 1 0.01 Amara lunicollis Schiødte, 1837 4 0 3 0 7 0.06 Bembidion fortestriatum (Motschulsky, 1845) 4 1 10 5 20 0.17 Bembidion wingatei Bland, 1864 8 5 4 5 22 0.19 Bradycellus lugubris (LeConte, 1847) 27 1 77 0 105 0.90 Calathus ingratus Dejean, 1828 87 8 65 14 174 1.50 Calosoma frigidum Kirby, 1837 2 0 15 0 17 0.15 Carabus granulatus Linné, 1758 0 0 1 0 1 0.01 Carabus maeander Fischer von Waldheim, 1820 1 0 8 4 13 0.11 Carabus nemoralis O.F. Müller, 1764 0 0 2 0 2 0.02 Clivina fossor (Linné, 1758) 16 2 33 10 61 0.53 Cymindis cribricollis Dejean, 1831 0 0 2 0 2 0.02 Diplocheila obtusa (LeConte, 1847) 0 0 0 1 1 0.01 Elaphrus clairvillei Kirby, 1837 1 0 7 1 9 0.08 Harpalus fulvilabris Mannerheim, 1853 36 21 130 21 208 1.79 Harpalus laevipes Zetterstedt, 1828 0 0 2 0 2 0.02 Harpalus somnulentus Dejean, 1829 1 0 1 0 2 0.02 Loricera pilicornis (Fabricius, 1775) 18 1 108 5 132 1.14 Patrobus foveocollis (Eschscholtz, 1823) 34 2 65 0 101 0.87 Patrobus longicornis (Say, 1823) 0 0 0 2 2 0.02 Platynus decens Say, 1823 553 56 2114 65 2788 24.03 Platynus mannerheimii (Dejean, 1828) 49 5 326 12 392 3.38 Pterostichus adstrictus Eschscholtz, 1823 214 51 760 15 1040 8.96 Pterostichus coracinus (Newman, 1838) 230 305 904 359 1798 15.49 Pterostichus luctuosus (Dejean, 1828) 0 0 9 0 9 0.08 Pterostichus melanarius (Illiger, 1798) 163 118 370 270 921 7.94 Pterostichus patruelis (Dejean, 1831) 0 0 1 0 1 0.01 Pterostichus pensylvanicus LeConte, 1873 276 44 668 7 995 8.57 Pterostichus punctatissimus (Randall, 1838) 12 3 252 4 271 2.34 Scaphinotus bilobus (Say, 1823) 28 18 42 52 140 1.21 Sphaeroderus nitidicollis Guérin-Méneville, 1829 14 173 37 181 405 3.49 Sphaeroderus stenostomus Dejean, 1826 33 74 57 33 197 1.70 Stereocerus haematopus (Dejean, 1831) 1 0 0 0 1 0.01

Appendix 4. - continued

Synuchus impunctatus (Say, 1823) 0 35 0 226 261 2.25 Trechus apicalis Motschulsky, 1845 14 3 14 9 40 0.34 Trichotichnus autumnalis (Say, 1823) 0 0 0 1 1 0.01 Total 2270 950 7064 1320 11604

Appendix 5. Rank abundance curve for carabid species collected in closed-canopy boreal mixedwood stands in northeastern Ontario.

ln(abundance + 1) + ln(abundance 0 0 10 20 30 40

Species rank

Appendix 6. Rank abundance curve for carabid species collected in a biomass clearcut that was formerly a boreal mixedwood stand in northeastern Ontario.

5 4 3 2 1

ln(abundance + 1) + ln(abundance 0 0 10 20 30 40

Species rank