Ground Beetles (Coleoptera: Carabidae) and Ecologically Sustainable Forest Management in the Northern Hardwood Forests of Central Ontario

Ben Angel

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

Faculty of Forestry University of Toronto

Ground beetles (Coleoptera: Carabidae) and ecologically sustainable forest management in the northern hardwood forests of Central Ontario

Ben Angel

Master of Science in Forestry

Faculty of Forestry University of Toronto

2019

Abstract

Ecological sustainability is a priority in the management of Ontario’s northern hardwood forests. It is important to assess harvesting impacts, as well as the likelihood of recovery prior to the next harvest. In this thesis, I use carabid communities as indicators of forest management in sugar maple (Acer saccharum Marsh.) dominated stands in Central

Ontario to compare different partial harvesting systems within a recently cut experimental forest, as well as between old-growth and mature logged stands. Greater basal area removal resulted in greater deviations in carabid communities from uncut conditions among experimental treatments, while communities were similar between old-growth and mature logged sites, despite strong difference in forest structure. These findings suggest that pre- harvest conditions may sufficiently recover between partial harvests to support ecological sustainability over time. However, the ability to recover may depend on initial logging intensity and harvesting rotation length.

Acknowledgements

I would like to thank my supervisor Dr. Jay R. Malcolm for the opportunity to take on this study and for helping me to grow as a scientist. I thank my advisory committee members Dr. Sandy Smith, Dr. John Caspersen, and Dr. Marie-Joseé Fortin for their guidance and valuable inputs.

For invaluable work in carabid identification, I would like to thank Nurul Islam.

Thanks to the summer 2017 field crew for the hard work, dedication, and laughs along the way.

I would like to thank my friends and colleagues at the Faculty of Forestry for their support and for generally making my experience more enjoyable and memorable.

This work was funded by grants from the Natural Sciences and Engineering Research Council of Canada and the Faculty of Forestry Haliburton Forest Brown Fund (to Jay R. Malcolm). I am indebted to the owners of Haliburton Forest and Wild Life Reserve Ltd. for permission to work on their property.

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Table of Contents Abstract ...... ii Acknowledgements...... iii List of Tables ...... vi List of Figures ...... vii List of Appendices ...... ix General Introduction ...... 1 Chapter 1: Ground beetle (Coleoptera: Carabidae) responses to experimental partial harvesting in the northern hardwood forests of Central Ontario ...... 6 Introduction ...... 6 Methods ...... 10 Study Area ...... 10 Study Site ...... 11 Beetle Sampling ...... 13 Habitat Measurements ...... 14 Statistical Methods ...... 16 Results ...... 17 Carabids ...... 17 Forest Structure ...... 18 Relationships between carabids and forest structure...... 20 Discussion ...... 21 Chapter 2: Forest management history and ground beetle (Coleoptera: Carabidae) communities in the northern hardwood forests of Central Ontario ...... 27 Introduction ...... 28 Methods ...... 31 Study Area ...... 31 Site selection and establishment ...... 31 Beetle Sampling ...... 32 Habitat Measurements ...... 33 Statistical Methods ...... 36 Results ...... 37 Carabids ...... 37 Forest structure and tree species composition ...... 38

Relationships between carabids and habitat variables ...... 38 Discussion ...... 39 General Conclusions ...... 44 Literature Cited ...... 50 Tables and Figures ...... 58 Appendices ...... 76

List of Tables

Table 1.1 Comparisons of carabid species abundances (captures per 100 pitfall bucket-nights) and total abundance among treatments in an experiment comparing different partial harvesting systems in Haliburton Forest, Ontario in July 2016, June 2017 and July 2017. Only species occurring in at least 30% of replicates in an individual month are shown and total abundance includes all captures. Either negative binomial regression or one-way analysis of variance was used to test for treatment effects.

Table 1.2. Comparison of forest structure variables among silvicultural treatments in an experiment examining different partial-harvesting methods in Haliburton Forest, Ontario. Letters in common indicate homogeneous groupings in Tukey's multiple-comparison test.

Table 1.3. Statistical significance of relationships between carabid abundance (captures per 100 bucket-nights) and environmental variables for species occurring in at least 30% of experimental units in an individual month. Variables excluded snags and size classes. Either negative binomial regression or simple linear regression were used. For significant relationships, the direction of the slope is indicated in parentheses.

Table 2.1. Carabid species abundances (captures per 100 pitfall bucket-nights) and total abundance compared between old-growth and mature logged stands in Haliburton Forest, Ontario. Only species occurring in at least 30% of transects in an individual month are shown; total abundance included all species.

Table 2.2. Comparison of forest structure and tree species composition variables between old-growth and mature logged stands in Haliburton Forest, Ontario.

Table 2.3. Variance explained and statistical significance in redundancy analysis (RDA) of carabid species abundances constrained by habitat variables in June and July at both the transect and station levels for sampling in Haliburton Forest, Ontario, in 2017. Permutation results (9999 iterations) include significance tests of all RDA axes as well as the top three most important variables in each analysis, each tested singly.

List of Figures

Figure 1.1. The Blue Heron Demonstration forest experimental design. Each treatment block is an 84 m square. The red line represents the path of the primary skid trail and M = ICR.

Figure 1.2. First two axes from a PCA ordination on the covariance matrix of standardized carabid abundances from sampling in July 2016, June 2017 and July 2017. Carabid species acronyms consist of the first four letters of the genus and the first four letters of the species. Total abundance and treatment centroids (CON, STS, DLC, FMS and ICR) were passive variables (red). Figure 1.3. Standardized carabid species abundances from the rarest species for each month on the right to the most common on the left for July 2016, June 2017 and July 2017. Carabid species acronyms consist of the first letter of the genus and the first letter of the species.

Figure 1.4. Carabid species richness plotted against abundance in July 2016, June 2017 and July 2017. Colours correspond to treatments: CON = dark green; STS = light green; FMS = yellow; DLC = orange; and ICR = red.

Figure 1.5. First two factors of a varimax rotated PCA ordination on the correlation matrix of forest structure variables: T_BS = small tree treatment block basal area; T_BL large tree treatment block basal area; T_HS = small tree basal area harvested; T_HL = large tree basal area harvested; T_LS = small tree local basal area; T_LL = large tree local basal area; D_ES = early- decay, small DWD density; D_EL = early-decay, large DWD density; D_LS = late-decay, small DWD density; D_LL = late-decay, large DWD density; S_T = total snag basal area. Colours correspond to treatments: CON = dark green; STS = light green; FMS = yellow; DLC = orange; and ICR = red. Treatment centroids (stars) were added passively. The control plot of CON1 is also indicated.

Figure 1.6. The first two axes of redundancy analysis on the covariance matrix constraining carabid abundances in July 2016, June 2017 and July 2017 with structure variables (red): T_B = Treatment block basal area; T_R = Treatment block basal area harvested; T_L = local tree basal area; D_E = Early-decay DWD abundance; D_E = Early-decay DWD abundance; D_L = Late- decay DWD abundance; F1 = First varimax factor, F2 = Second varimax factor. Treatment centroids and total abundance were added passively (green).

Figure 2.1. Sampling transect unit including the 420 m long main transect, 100 m long perpendicular intersecting “fishbone” transects, and primary and secondary stations represented by blue and green dots respectively. Figure 2.2. First two axes from a Principal Component Analysis on the covariance matrix of standardized carabid abundances from sampling in June 2017 and July 2017 in Haliburton Forest, Ontario. Carabid species acronyms consist of the first four letters of the genus and the

vii first four letters of the species. Total abundance was added as a passive variable. Colours correspond to treatments: old-growth (green); mature logged (red).

Figure 2.3. Standardized carabid species abundances for June 2017 and July 2017 sampling in Haliburton Forest, Ontario. Carabid species acronyms consist of the first letter of the genus and the first letter of the species.

Figure 2.4. First two axes of a PCA ordination on the correlation matrix of forest structure and trees species composition variables for mature logged (red) and old-growth (green) sites in Haliburton Forest, Ontario (BA = basal area; SBY_BA = small yellow birch basal area; scale = Weibull scale parameter ; shape = Weibull shape parameter; BA_SEMI = basal area semivariance; %MH = percent sugar maple composition; %MH_SEMI = semivariance of percent sugar maple composition; conifer = conifer score; D_ES = small, early-decay DWD volume; D_EL = large, early-decay DWD volume ; D_LS = small, late-decay DWD volume; D_LL = large, late-decay DWD volume; S_ES = small, early-decay snag basal area; S_EL = large, early-decay snag basal area ; S_LS = small, late-decay snag basal area; S_LL = large, late-decay snag basal area; LD = leaf litter depth).

Figure 3.1. Standardized abundances of Pterostichus coracinus across treatments of varying basal areas in Haliburton Forest sampled in June 2017. Blue indicates data from Blue Heron Demonstration Forest (Chapter 1) with the star representing a mean of all treatments. Green indicates data from mature logged and old-growth stands (Chapter 2). Treatment acronyms: BH = Blue Heron Demonstration Forest; CON = control; STS = single-tree selection; FMS = financial maturity selection; DLC = diameter-limit cutting; ICR = intensive crown release: ML = mature logged; OG =old-growth.

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List of Appendices

Appendix 1. Location of Blue Heron Demonstration Forest (BH) and transects within old-growth (OG) and mature logged (ML) stands in Haliburton Forest and Wild Life Reserve (HF), adjacent to the south-west side of Algonquin Provincial Park (APP). Appendix 2. Species list in descending rank of abundance, indicating total number of individuals captured in July 2016, June 2017 and July 2017 in the five silvicultural treatments in the Blue Heron Demonstration forest in Haliburton Forest, as well habitat categories. Effort is listed in parentheses (bucket-nights).

Appendix 3. Species list in descending rank of abundance, indicating total number of individuals captured in June 2017 and July 2017 in old-growth and mature logged stands in Haliburton Forest, as well as habitat categories. Effort is listed in parentheses (bucket-nights).

General Introduction

Since the time of European settlement, the northern hardwood forests of Central Ontario have been subject to widespread logging in addition to land clearing for agriculture and other human activities (Pinto et al., 2008). Research suggests that across hardwood forests of the Great

Lakes-St. Lawrence forest region, present-day conditions differ from the original forests in both structure and tree species composition. This conclusion is based on comparisons of managed stands to remaining old-growth forests (Angers et al., 2005; Goodburn & Lorimer, 1998) and on descriptions found in historical tree surveys (Leadbitter et al., 2002; Pinto et al., 2008; Schulte et al., 2007; White & Mladenoff, 1994). Harvesting has focused on the removal of large, mature trees, which in turn has led to reduced natural tree death, snag formation and downed woody debris (DWD) inputs (Angers et al., 2005; Vanderwel et al., 2006a). The most distinct change in tree composition has been a decline in the abundance of coniferous species, specifically those that were commercially important in the past, including white pine (Pinus strobus L.), red pine

(Pinus resinosa Ait.), and eastern hemlock (Tsuga canadensis (L.) Carrière) (Pinto et al., 2008).

With the decreases in softwoods has come increases in hardwood species, particularly sugar maple (Acer saccharum Marsh.) (Pinto et al., 2008; White & Mladenoff, 1994). The sugar maple dominated landscapes we see today are thought to be attributed to a combination of the removal of valued tree species, fire suppression, and harvesting that creates small gaps that favour sugar maple regeneration (Leadbitter et al., 2002). These changes in forest structure and tree species composition can be expected to broadly affect habitat conditions, with consequences for community assemblages of the resident biota (Leadbitter et al., 2002; Quinn, 2004).

Hardwood logging remains widespread throughout Central Ontario, but the value of ecological sustainability in forest management has gained importance over time and is reflected in modern silvicultural systems (Quinn, 2004). Single-tree selection (STS) is the most common system used today and promotes sustainability by limiting basal area removal, emulating natural disturbances through the maintenance of uneven-aged diameter distributions, and by retaining certain habitat features (OMNRF, 2015). Although an improvement over historical approaches in that ecological impacts are explicitly considered, STS nonetheless comes with potential drawbacks, including a possible shift in species composition towards sugar maple dominance

(Crow et al., 2002) and reduced quantities of snags and DWD (Vanderwel et al., 2008;

Vanderwel et al., 2006b). These concerns may pose a stronger threat to long-term ecological sustainability in managed stands depending on the extent to which stand conditions are able to recover before the next harvest. If the conditions typical of older forests fail to become reestablished prior to the next harvest, then there is potential for impacts to accumulate from one harvest cycle to the next.

Close monitoring and experimentation are essential in evaluating current conditions and working towards improvement (Tierney, 2009), however, ecosystems consist of sometimes complex interactions between species and their environment. Attempting to study them in their entirety is difficult and the use of bioindicators is one way of working with the complexity of ecology (Lindenmayer et al., 2000). Bioindicators are biological taxa whose status reflects or indicates conditions in their environment (Siddig et al., 2017). By focusing on a single component that can be used to make inferences about a more complex system, bioindicators provide a potentially efficient and effective tool for gathering meaningful information about broader ecological responses to changes in the environmental (Siddig et al., 2017). They have

3 been grouped into indicators of abiotic or biotic factors, and depending on the objective of the study, a good forest management bioindicator may span both of these categories (Rainio &

Niemela, 2003).

Ground beetles (carabids) are well recognized as bioindicators (Koivula, 2011) and have been used for a range or purposes, including ecosystem and land type classification (Bergeron et al., 2012; Rykken et al., 2018), urbanization (Kotze et al., 2012), climate change (Brandmayr &

Pizzolotto, 2016) and have been proven to be particularly valuable in understanding ecosystem responses and sustainability in forest management (Langor & Spence, 2006; Niemelä et al.,

2007; Pearce & Venier, 2006; Work et al., 2008). One key characteristic of carabids in this regard is that they are abundant and commonly occurring across a wide range of regions, habitats, and successional stages, enabling the capture of many individuals within a relatively small area and short amount of time (Kotze et al., 2011). Since they are so common, they are popular subjects of entomological study, resulting in a richer knowledge of taxonomy, ecology and behaviour compared to many other insect families (Kotze et al., 2011). Carabids have short generation times and their high diversity in behaviours and life strategies across species strengthens community-level responses to change. They also have many food web connections and are important for soil decomposition, meaning that changes in carabid communities may reflect or result in broader changes in the ecosystem (Kotze et al., 2011; Thiele, 1977).

A forestry-influenced habitat variable that carabids are especially sensitive to is canopy openness (Niemelä, 2007), which is inherently affected by harvesting. Canopy openness influences many forest habitat and microclimatic conditions, including light and temperature levels, moisture regimes, leaf litter development, and understory plant communities (Martel,

1991) all of which are important to carabids (Larochelle and Larivière 2003; Siitonen, 2001;

Thiele 1977). Carabid species have often been classified into different ecological groups based on habitat openness, which serves as a useful tool to more easily interpret carabid compositional changes, although these descriptions can differ by geographical region (Work, et al., 2008). Tree species composition may also be influenced by forest management and can have implications for carabids. The division between conifer and deciduous composition is a particularly strong determinant of carabid community composition across Canada (Work, et al., 2008). Direct relationships between carabids and tree species are not well understood, and likely involve a range of indirect factors relating to forest type and associated ecosystem components (Bergeron et al., 2011).

As a primarily ground-dwelling family, carabids are heavily dependent on forest floor features to meet habitat requirements (Thiele, 1997). DWD is used as a source of nutrition, shelter, moisture, as well as a site for reproduction and associations with other saproxylic species

(Siitonen, 2001). Beetles, and especially carabids, constitute a substantial component of saproxylic communities (Siitonen, 2001; Work et al., 2004). Large and highly-decayed pieces have been found to be of particular importance, which can be attributed to their provision of a cool moist habitat, as well as soft wood suitable for oviposition and aestivation (Siitonen, 2001).

This form of DWD is also known to decline under partial harvesting (Angers et al., 2005;

Goodburn & Lorimer, 1998). The impact of forestry on DWD availability thus introduces another potentially valuable characteristic of carabids as forest management bioindicators. Leaf litter is another forest floor element that can similarly act as an important habitat feature for carabids. Reduced forest cover can lead to weaker leaf litter inputs (Martel, 1991) and leaf litter depth has been found to be greater in old-growth over mature logged forests (Vance & Nol,

2003), making it another forestry-sensitive variable with implications for carabid communities.

In this thesis, my objective is to evaluate the ecological sustainability of forest management in the northern hardwood forests of Central Ontario, with emphasis on STS.

Carabid communities, along with associated habitat components, were compared between treatments of differing logging histories and silvicultural prescriptions to assess the ecological impacts of harvesting. In Chapter 1, carabid communities were compared among recently-cut silvicultural treatments in the Blue Heron Demonstration Forest in Haliburton Forest, with treatments of higher intensity expected to result in lower abundances of species common to control plots, and possible increased captures of rare open-habitat species. This provided insight into the initial effects of partial harvesting along with potential differences among silvicultural systems. In Chapter 2, carabid communities were compared between old-growth and mature

STS-managed stands, with communities expected to be similar between treatments, but with possible differences at the species level. Here I evaluated the ultimate sustainability of STS at the end of the harvesting cycle using old-growth conditions as a reference. These comparisons enabled assessment of ecological conditions at both early and late stages of the harvesting cycle, as well as potential effects of differences in long-term management history and the use of different partial-harvest silvicultural systems. Together, the two chapters provide a comprehensive look at the sustainability of STS from the perspective of an important indicator group.

Chapter 1: Ground beetle (Coleoptera: Carabidae) responses to experimental partial harvesting in the northern hardwood forests of Central Ontario

Introduction

Most northern hardwood forests in Ontario are managed through partial harvesting, with selective systems such as diameter-limit cutting being used historically, and single-tree selection

(STS) being most commonly used today (Quinn, 2004). In contrast to clear-cutting, partial harvesting retains much of the original forest cover, with STS normally removing less than one third of existing basal area (OMNR, 2004). However, despite relatively low harvesting intensity, such harvesting can significantly alter forest structure (OMNR, 2012), with implications for forest biodiversity (Vanderwel et al., 2009). Additionally, impacts on habitat structure and local communities may vary depending on the exact nature of silvicultural prescriptions among different partial harvesting systems (Werner & Raffa, 2000).

Certain direct effects of logging, such as basal area reductions and the resulting opening of the canopy, are expected to closely follow silvicultural prescriptions. Other effects, such as snag and DWD stock changes are indirect, but can also be influenced by the nature of the forest management (Hautala et al., 2004; Vanderwel et al., 2008). Tree harvesting removes stems which would eventually end up as snags and DWD (Angers et al., 2005; Goodburn & Lorimer,

1998); at the same time, it provides a significant input of relatively small DWD pieces coming from tree crowns upon harvest (Vanderwel et al., 2008). Through machine destruction of existing

DWD, harvesting can also result in reduced DWD stocks, with large, relatively decomposed wood being most susceptible to destruction (Hautala et al., 2004). Safety considerations can lead to additional snag reductions (OMNR, 2012) and efforts to improve stand condition could

7 contribute to reduced woody debris inputs over time. These changes in forest structure modify habitat conditions and hence have the potential to alter forest communities. Species that rely on mature forest conditions may no longer find the stand suitable, whereas other species favoring more open conditions may colonize the stand (Doyon et al., 2005; Vanderwel et al., 2009).. At the forest floor level, changes in dead wood resources may influence populations of saproxylic species (Siitonen, 2001). These consequences have implications for ecological sustainability, which is an important part of successful forest management (Attiwill, 1994; Lindenmayer et al.,

2000).

When examining the effects of logging on forest biodiversity, understanding the short- term effects of the management is crucial. Further ecological succession builds upon these early changes, and they deserve special consideration in the development of sustainable silviculture.

The early response of carabids after clear-cutting in the boreal forest has been well-investigated, and several studies have found that the complete opening of the canopy induces decreases in forest specialists, with some species disappearing entirely after harvest, as well as increases in open-habitat specialists, many of which were not present before harvest (Koivula, 2002a).

Generally, the influx of open-habitat specialists outweighs the reduction in forest specialists, meaning that total abundance experiences an increase (Niemelä et al., 2007). In contrast, very little research has focused on the early effects of partial harvesting on carabid communities in northern hardwood forests, with only one study examining carabid response to STS within three years of cutting. They found that total carabid abundance was 50% lower in recently-logged compared to mature logged stands, but did not report a surge in open-habitat specialists (Vance

& Nol, 2003). Communities from mature STS-managed stands and unharvested old-growth stands were comparable, although two species were more common in logged sites.

Several other studies in northern temperate forests have investigated carabid response to forest management, but at later stages following the harvest. In the hardwood forests of southern

Quebec, no significant difference was found between mature logged stands dominated by 85-130 year-old sugar maple and stands 6-8 years after a selection cut (Moore et al., 2004). Studies conducted in Wisconsin and Upper Peninsula Michigan found stronger distinctions between assemblages from old-growth forests and in uneven-aged stands cut between four and thirteen years prior to carabid sampling (Latty et al., 2006). The selection system used in this northern region of the Great Lakes States is more intensive than STS in Ontario, with an 8-15 year cutting cycle, a minimum 16.1 m2/ha residual basal area, and maximum tree diameter of 45 cm DBH

(Goodburn & Lorimer, 1999) compared to a 20 year cycle, 18-20 m2/ha basal area target, and 60 cm DBH maximum diameter for STS (OMNR, 2004). These findings highlight the possibility that variation among partial harvesting systems may have implications for carabid communities.

The importance of DWD to carabids in northern hardwoods has received very little attention, but has been found to be a correlate at the community level (Latty et al., 2006). Associations between carabid species and DWD has been described in the boreal forest of northern Ontario, including for several species that are common in temperate forests to the south (Pearce et al.,

2003; Piascik, 2013).

Partial harvesting systems used in the Great Lakes-St. Lawrence forests have changed considerably over time and continue to evolve today. Early logging occurred at a time where ideas of sustainable forest management were not prevalent and when high grading was a common mode of harvesting (Quinn, 2004). The earliest organized system of widespread use was diameter-limit cutting (DLC), a selective, even-aged system that is a historically important method in Ontario and remains the basis for some municipal tree bylaws (OMNR, 2012). DLC is

9 now considered poor practice and is no longer recognized as a silvicultural system in Ontario due to potential declines in the quantity and quality of growth stock in future generations (OMNR,

2012). DLC may also result in greater long-term reductions in snag and DWD abundances than in stands managed under selection silviculture (Angers et al., 2005; Goodburn & Lorimer, 1998).

Out of improvements to diameter-limit cutting came selection systems, with STS being the most widely used system in the northern hardwood forests of Ontario, particularly on public lands

(OMNR, 1998). It is used specifically in upland tolerant hardwood forests and aims to reduce environmental impacts by maintaining an uneven-aged diameter distribution, improving stand health, and retaining wildlife features (OMNR, 2012). One concern with STS is that it can result in a relatively closed canopy with only small gaps, leading to the dominance of shade tolerant trees such as sugar maple (Acer saccharum Marsh.) and reduced growth opportunity for mid- tolerant species such as yellow birch (Betula alleghaniensis Britt.) (Neuendorff et al, 2007;

Shields et al, 2007). The desire to resolve this problem has motivated ideas for alternatives to

STS that create larger gaps in the canopy, such as group selection (Halpin et al., 2017). Intensive crown release (ICR) is an example of a system with higher basal area removal and can act to simulate group selection conditions along with uncut control areas in this study. ICR is an approach originally practiced in Germany that focuses on the potential of a few high-quality stems by removing nearby competitors and can thus result in areas of low canopy cover. Another criticism of STS is that it results in stands of poor productivity, as well as the harvest of low quality stems (Miller & Smith, 1993). In response to these drawbacks, new approaches have modified STS to have greater basal area reduction, promoting greater regeneration and tree growth, and the prioritization of the removal and nurturing of high quality trees (Miller & Smith,

1993). Such systems aim to optimize economic return and have been termed Financial Maturity

Selection (FMS) systems. Financial maturity refers to high quality trees that have grown to be large in size, but have yet to experience depreciation in value due to defects that come with further aging and growth (Cecil-Cockwell & Caspersen, 2015). These various systems differ in respect to basal area retention and residual diameter distribution prescriptions, which could in turn result in differing impacts on biodiversity, and carabids in particular (Work et al., 2008).

The objectives of this chapter were to measure habitat structure and carabid communities in a recently created silviculture experiment established in a sugar maple-dominated site in

Central Ontario, which included control plots along with four silvicultural treatments that varied in their target basal area removal and nature of the residual forest, as well as their historical relevance and the eventual potential for economic returns. Specific objectives were to: (1) examine responses of carabid communities, with more intense systems (i.e., greater basal area removal) predicted to result in lower numbers of forest specialist and carabids in general, and with the potential for an increase in open-habitat specialists, (2) to compare the early structural effects of the various partial harvesting systems, with systems of higher harvesting intensity predicted to result in higher amounts of small, young DWD, as well as a possible decrease in snags and large decaying DWD, and (3) investigate relationships between forest structure variables and carabid communities, with variables related to harvesting expected to show the strongest relationships.

Methods

Study Area

The study was conducted at a site within Haliburton Forest and Wild Life Reserve, a privately-owned forest located in Central Ontario (45°13' N, 78°35' W) in the Great Lakes-St.

Lawrence forest region (Rowe, 1972). Forest cover is dominated by sugar maple (Acer saccharum Marsh.) in association with American beech (Fagus grandifolia Ehrh.), yellow birch

(Betula alleghaniensis Britt.), and eastern hemlock (Tsuga canadensis (L.) Carrière). The mean annual temperature in the region is 5.0 °C, with monthly means ranging from −9.9 °C in January to 18.7 °C in July, and an average of 1074 mm of annual precipitation (Environment Canada,

2018). Logging in the area dates back to the removal of white pine beginning in the mid-1800s, followed by the harvesting of other target species such as yellow birch in the 1940s. More organized forest management began in the 1960s with the use of selective systems and a shift to modern selection systems in the 1970s and 1980s (Vanderwel et al., 2008).

Study Site

The Blue Heron Demonstration Forest was established to compare the financial productivity of several partial harvesting systems and is located at the south end of Haliburton

Forest (Appendix 1). The 19.76 ha site was originally a relatively homogeneous stand of sugar maple forest, with secondary components of American beech, eastern hemlock and yellow birch.

Pre-harvest tree inventories indicated basal area proportions of 75% sugar maple, 9% beech 6% hemlock and 4% yellow birch. The experimental layout was an approximately rectangular grid of twenty-eight 84 by 84 m contiguous blocks of forest, with each block acting as a treatment replicate. Treatment placement followed an approximate Latin square design to minimize spatial bias (Fig. 1.1). Four of the treatments were replicated six times and one (ICR - see below) was replicated four times.

The treatments were diameter-limit cutting (DLC), single-tree selection (STS), financial maturity selection (FMS), intensive crown release (ICR), and control (CON). The lone objective of DLC was to remove all stems over 30 cm DBH up to a limit of 15 m2/ha residual basal area.

Under STS, the target basal area was 18-20 m2/ha, with size class stocking of: Pole 6, Small 6,

Medium 5, Large 3 (where pole is defined as 10-24 cm DBH, small as 26-36 cm DBH, medium as 38-48 cm DBH, and large as 50-60 cm DBH using 2-cm diameter classes) (OMNR, 2004).

There is also a focus on crown spacing and stand improvement through removal of unsuitable growing stock, as well as the retention of wildlife features (OMNR, 2012). Financial maturity selection focuses on the removal of stems at financial maturity, approximately 40 cm DBH. The target basal area was 16 m2/ha, with size class stocking of: Pole 5, Small 5, Medium 5, Large 1.

FMS follows similar crown spacing, stand improvement and wildlife retention values to STS, with an additional priority of retaining veteran trees (1 m2/ha) with non-infectious defects. The two objectives of ICR were to release small (DBH < 35 cm) crop trees by removing any competing stems within 12 m, and to retain some large (DBH > 35 cm) crop trees. Crop trees were defined as having no defects up to the height of approximately 5 m. New crop trees are assumed to grow in unmarked areas between crop trees (the “filler forest”), meaning that ICR can result in harvesting that is heavy in some places and light in others, depending on the density and distribution of crop trees. ICR does not specify a particular basal area target, but it is nonetheless expected to be much lower on average than the other treatments. The control was unharvested, but in some cases contained spur skid trails used to access other treatment blocks.

Establishment of the experimental site began in 2014, with harvesting taking place in 2014 and

2015. The entire site had been most recently logged under STS in 1993.

Beetle Sampling

During three distinct sampling periods (July 21-28, 2016, June 19-28, 2017, and July

19-28, 2017), carabids were trapped in the center of each block using a "drift fence" pitfall array

(Piascik, 2013). Each array was in the shape of a "Y", consisting of a central pitfall bucket and three buckets in the shape of an equilateral triangle, each positioned 1.2 m away from the central bucket. The center bucket was connected to each end bucket by a 30-cm high geotextile sheet buried 10 cm deep in the ground and held in place by wooden stakes flush with the bucket edges.

The geotextile acted as a "drift" fence by guiding beetles to fall into the buckets at either end.

The tops of buckets lay flush with the ground and had dimensions of 11 cm in diameter and 13.3 cm in depth. When not in use, buckets were covered by plastic lids and leaf litter.

Traps were activated by removing the bucket lids and filling each bucket one-third full of

5% salt solution including a small amount of dish soap. Traps were left active for six consecutive nights, then trap contents were collected and preserved in 70% ethanol. Each collection contained the contents of all four buckets of an array. Carabids were pinned and identified using keys from Bousquet (2010). Carabid species were assigned to ecological groupings based on habitat preference classification systems used in other studies (Latty et al., 2006; Vance & Nol,

2003) as well as species descriptions found in Larochelle & Larivière (2003).

To correct for missing effort, abundances were standardized to captures per 100 bucket- nights. Following Piascik (2013), an outer bucket was counted as one bucket and the center bucket counted as 2.236 buckets (2.236 times the capture rate of outer buckets). Each pitfall array trapped for six nights thus totaled for 31.4 bucket-nights. If more than one outer bucket

(i.e., six bucket-nights) was compromised, then the entire sample was excluded from analysis.

This occurred for one block in each of the DLC and CON treatments in July 2016 and one in the

FMS treatment in June 2017. Missing effort was otherwise rare (under 5% of bucket-nights).

Habitat Measurements

In the summer of 2017, BAF2 prisms sweeps were conducted at the center of each block. All live trees found to be "in" and of a diameter at breast height (DBH) of ≥10 cm were measured to DBH. I used prism sweeps rather than fixed area plots because they were less labor intensive and they provided a more comprehensive representation of large stems within a stand.

Prism sweeps were also performed by Haliburton Forest for tree marking during the establishment of the experiment. Because trees to be harvested were marked, these sweeps provided basal area measurements for both before and immediately after the harvest. Each treatment block was divided into quarters to form four equal square plots, and prism sweeps were situated at the center of each plot. This tree marking data provided more comprehensive measurements for each treatment block, whereas my prisms better described the local tree community in the vicinity of a trap. Accordingly, these two sets of measurements are here referred to as "block" and "local" measurements, respectively. There were no block prism sweeps in the control plots, so fixed-area data and local prism sweeps, weighted at 80% and 20% respectively, were used to fill this gap in the data. Here, 250 m2 circular fixed-area plots also used the four-plot system.

Data obtained from prism sweeps were then used to calculate total basal area and basal area of relatively small (<38 cm DBH) and large (≥38 cm DBH) stems for both the local and block data, and for the latter, for before, after and harvested basal area. This size classification

15 represents the distinction between small and medium saw logs in Ontario forest management

(OMNR, 2004) and hence was relevant to residual stocking level targets of uneven-aged systems

(FMS and STS). This size classification also enabled the alignment of size classes between local and block data, which assigned trees to saw log size classes without a precise DBH measurement.

DWD was sampled via the line-intersect method (Wagner, 1968) along 84-m-long transect lines running perpendicular to each other that were parallel to the block boundaries and intersected the block center (Fig. 1.1). Any piece of intersecting DWD ≥10 cm in diameter had its diameter measured and was categorized with respect to size (less than, or equal to or greater than median diameter [13.8 cm]) and decay class. Decay class was determined as either "early" or "late" decay depending on the condition of the outer wood layers below the bark. If they were spongy or easily damaged by kicking, they were classified as late-decay; otherwise they were early-decay. These assignments corresponded roughly to decay classes 3-5 and 1-2 from Maser et al. (1979), respectively. From these data, I derived four DWD variables: the number of pieces per 100 m in the small and late, small and early, large and late, and large and early classes. I used numbers of pieces rather than volumes because preliminary ordinations indicated that the number of pieces explained more variation in carabid communities.

Snags were measured in a belt prism transect: that is, any snags of ≥10 cm DBH found to be “in” from any point along the DWD transects were measured to DBH. Only the transects along the one DWD dimension (the y dimension in Fig. 1.1) were used in order to avoid overlapping effort. Snag occurrence was low, hence only total snag basal area was calculated without any size or decay classifications.

Statistical Methods

One-way analysis of variance (ANOVA) was used to compare carabid abundances and habitat features among the five treatments. In some cases, log transformations were applied to structure variables to normalize residuals. Negative binomial regression was used on carabid abundances in place of analysis of variance if the assumptions of normality or homogeneity were violated, as evidenced from examinations of residuals. A negative binomial regression was deemed appropriate if the deviance value was between 0.8 and 1.2. If the deviance value was below 0.8 then, then log transformations were used on carabid abundances to bring the value to within the required range before testing. Linear regression was used to explore relationships between carabid abundances and habitat variables. Negative binomial regression and log transformations were applied to carabid abundances using the same criteria as for ANOVAs.

Statistical tests were undertaken using SAS (v. 9.4).

Ordination techniques were used to investigate carabid community composition, habitat variation, and treatment effects. Principal components analysis (PCA) on the correlation matrix was used for habitat variables, and a varimax rotation was applied in order to obtain an axis that best described the harvest-intensity gradient (see below). A PCA was also used for carabid communities, but on the covariance matrix, which gave less weight to rare species. Redundancy analysis (RDA, 9999 permutations) was used to constrain carabid abundances using forest structure variables. All ordinations were performed using Canoco (v. 4.5).

Carabid abundance and species richness values were calculated using EstimateS (Version

9, R.K. Colwell, http://purl.oclc.org/estimates).

Results

Carabids

In total, 806 individuals were captured of 21 carabid species. Synuchus impunctatus (Say,

1823) (54%) and Pterostichus coracinus (Newman, 1838) (24%) made up the majority of July

2016 captures, while Pterostichus coracinus (31%), Agonum retractum (LeConte, 1848) (23%) and Pterostichus pensylvanicus (LeConte, 1873) (18%) were dominant in June 2017. July 2017 shared dominant species with both other months; Pterostichus coracinus (48%), Synuchus impunctatus (24%), Agonum retractum (16%). There were five species exclusive to either July

2016 or June 2017 and no species captured only in July 2017.

None of the carabid species abundances showed significant differences among the treatments (Table 1.1), however, ordinations provided evidence of a gradient in species composition according to harvest intensity. In both July 2016 and July 2017, treatment centroids were positioned along the total carabid abundance vector in approximate order according to harvesting intensity, with CON showing greatest abundance, STS, DLC, and FMS intermediate, and ICR showing the least (Fig. 1.2). Here, total abundance was dominated by Pterostichus coracinus and Synuchus impunctatus; other relatively long species vectors in the same direction included Agonum retractum, Sphaeroderus canadensis Darlington 1933, and Pterostichus pensylvanicus. In both months, the majority of uncommon or rare species vectors, including

Agonum fidele Casey, 1920 and Olisthopus parmatus (Say, 1823) were negatively related to total abundance (Fig. 1.2.).

In June 2017, the first axis, rather than the vector for total abundance, divided common and rare species, but was only somewhat aligned with treatment intensity (Fig. 1.2). Here,

Pterostichus pensylvanicus, a species less common in other months, was most important for axis

2. Unlike other common species, Pterostichus pensylvanicus was more associated with FMS and

DLC rather than CON and STS, which can explain why the patterns seen in July 2016 and June

2017 did not hold true in this month.

The pattern of contrast between the responses of relatively abundant species compared to rarer ones was evident in rank-abundance plots for the various treatments (Fig 1.3). With the progression from uncut controls to silvicultural systems of progressively lower basal area retention, dominant species common in control plots tended to decrease in abundance and uncommon species tended to experience either little effect, an increase in abundance, or they appeared for the first time (Fig. 1.3). The trend was less evident in June 2017 where dominant species Agonum retractum and Pterostichus pensylvanicus did not fall in abundance until ICR and were most common in FMS and DLC. Pterostichus coracinus was common across all months and followed the trend of decrease in abundance with higher harvesting intensity, except that it was consistently more common in STS than CON (Fig. 1.3).

In rarefaction analyses, low abundance and high species accumulation were consistently true in the ICR plots for the three sampling periods. Difference in positioning among the remaining treatment rarefaction curves was less pronounced, although FMS and DLC generally exhibited lower abundances and higher species accumulation than CON and STS (Fig. 1.4).

Forest Structure

Post-harvest basal area in the treatment blocks followed expectations from the prescriptions except that DLC was higher than FMS, and STS was slightly below the target basal area retention range of 18-20 m2/ha. For the block prism sweeps, all harvested basal area, and

19 post-harvest tree basal area variables differed significantly among the treatments, with the exception of small post-harvest trees (Table 1.2). At the local level, only large trees experienced a significant treatment effect. Total DWD pieces and all early DWD variables also differed among treatments and were found to be much higher in harvested treatment than in controls.

Late-decay DWD and snags did not significantly vary among the treatments, but were most common in CON, although not significantly so (Table 1.2). Generally, variables that experienced treatment effects showed CON to be significantly different from the harvest treatments, and with

ICR at the other extreme, although statistical significance among the logged treatments was infrequent (Table 1.2).

The varimax-rotated PCA formed a first axis strongly correlated with basal area and early-decay DWD variables, and thus acted as a succinct measure of harvesting intensity (Fig.

1.5). Along the first axis, CON showed the lowest score and exhibited strong separation from the logged treatments; ICR, in contrast, had the highest scores. Of the remaining treatments, STS was nearest to CON, and both FMS and DLC fell roughly between STS and ICR (Fig. 1.5). Snag basal area and late-decay DWD piece abundances were more-or-less orthogonal to the main harvesting intensity axis (in the positive direction) as was small tree basal area (negative direction). Small tree variables highlighted the contrast between FMS (low small tree basal area) and DLC (high small tree basal). The control block of CON1 was a structurally unusual block that may have been subject to low-impact horse logging in the mid-2000s, and was a strong contributor to the second varimax factor (Fig. 1.5). CON1 was characterized by extreme values for structural measurements, including high values for late-decay DWD and low values for early- decay DWD, as well low basal area, specifically for small trees.

Relationships between carabids and forest structure

Of the structure variables measured, those related to the first varimax factor proved to be the most important in describing variation in species composition. Four common species were found to be correlated with at least one tree variable, early-decay DWD, or factor one itself

(Table 1.3). Pterostichus coracinus was not significantly correlated to post-harvest basal area, but in July 2016 and June 2017 it was less abundant in blocks with greater harvested basal area and was negatively correlated to the first varimax factor. In July 2016, Synuchus impunctatus was highly correlated with both high block basal area and low harvested block basal area, as well as negatively correlated with the presence of early-decay DWD and the first varimax factor in

July 2016. Of all species, Sphaeroderus canadensis was most related to tree variables and was positively correlated to local tree basal area as well as block basal, and was negatively correlated to harvested basal area and the first varimax factor. These were also the common species that showed the strongest signs of treatment effects relating to harvest intensity (Table 1.1). In July

2016 and July 2017, total carabid abundance was associated with high block basal area, and both harvested basal area and values for the first varimax factor were negatively correlated to total carabid abundance in July 2016. Total abundance was also lower in blocks with more early- decay DWD pieces in both July 2016 and July 2017, but this effect is primarily due the dominance of Synuchus impunctatus. Pterostichus pensylvanicus was the only species to be significantly correlated to the second varimax factor and responded negatively to tree basal area and positively to late-decay DWD in June 2017 (Table 1.3).

The redundancy analyses showed similarities with PCAs in that rare and common species vectors were divided, and centroids were approximately distributed according to treatment intensity. This was especially true in July 2016, and to a lesser extent in July 2017 (Fig. 1.6). The

21 opposition of rare and common species and treatment effects were less clear in June 2017.

Environmental variable effects, however, looked to be relatively strong across the sampling periods, and the first varimax factor in particular appeared to have an important influence (Fig.

1.6). In June 2017, the second varimax factors was also important and was aligned with vectors of species found in CON1. Environmental variables best explained carabid variation in July 2016

(44%, p < 0.01), with the first varimax factors accounting for much of the percent explained

(26%, p < 0.01). July 2017 was similar to July 2016, but effects were much weaker and not significant. Structure variables including local basal area (9%, p = 0.05) and late-decay DWD abundance (9%, p = 0.04) were significant in explaining carabid variation in June 2017, and the first (8%, p = 0.07) and second varimax factors (8%, p = 0.09) were equally important. Overall, however, RDA axes were not significant in accounting for carabid variation in June 2017 (27%, p = 0.08). These relationships between carabid abundances and forest structure variables, along with multivariate analysis at the community level, outline strong treatment effects, despite insignificance in ANOVAs at the species level.

Discussion

Abundances of individual carabid species did not show significant differences across the various treatments, but regressions against forest structure variables and community-level analyses revealed highly significant responses to the harvesting. Together, these illustrated that systems of higher harvesting intensity tended to be associated with lower carabid abundances, but higher species richness. Species that were most common in control plots tended to decrease in abundance with higher intensity treatments, whereas less common and rare species became

22 more prominent. These patterns are similar to the responses of carabid communities to clear- cutting in boreal forests; namely, decreases in previously dominant forest dwelling carabids and increases or influxes of open-habitat species (Niemelä et al., 2007), but were manifested to a lesser degree. In clear-cut studies, open-habitat species increased in abundance to the point of causing increases in total abundance (Niemelä et al., 2007). In my study, common species remained the most common across treatments and increases in less common and rare species were relatively subtle and dependant to some extent on singletons.

Of dominant species experiencing decreases in response to increasing harvest intensity,

Synuchus impunctatus was the best example, and had previously been found to be more common in mature logged stands rather than recent STS-cut stands in Algonquin Provincial Park (Vance

& Nol, 2003). Synuchus impunctatus has been described as a generalist species and is common in lightly forested areas (Larochelle and Larivière, 2003). The species did not, however differ in abundance between mature logged and old-growth northern hardwood forests in either Central

Ontario (Vance & Nol, 2003) or northern Wisconsin and the Upper Peninsula of Michigan

(Werner & Raffa, 2000). I also found that Synuchus impunctatus was negatively correlated with the number of pieces of early-decay DWD. In a downed-wood removal experiment in boreal

Ontario, Piascik (2013) found that Synuchus impunctatus was negatively correlated with small early-decay DWD, suggesting it may be an unfavourable habitat feature for this species. Early- decay DWD, however, is a clear sign of harvesting intensity and may be acting as an indicator of forest disturbance and canopy openness, rather than a direct deterrent on the forest floor.

Pterostichus coracinus was the other dominant species that followed this trend and was also found to be more common in mature logged stands than in recent STS-cut stands in Algonquin

Provincial Park (Vance & Nol, 2003). It was, however, more common in mature logged than old-

23 growth stands. This species has been described as a forest generalist and is currently the best- known indicator for mature logged vs. old-growth northern hardwood forests based on findings in both Vance & Nol (2003) and Werner & Raffa (2000). In my study, Pterostichus coracinus was found to be more common in the STS rather than control plots, possibly reflecting an affinity for moderately disturbed forests. Among the remaining species, Sphaeroderus canadensis was the most affected by treatments of increasing intensity, which fits with its designation as a forest specialist (Larochelle and Larivière 2003). It was also the only species to respond negatively to local basal area, suggesting it could be sensitive to harvesting at a smaller scale. Three species in particular (Synuchus impunctatus, Pterostichus coracinus and

Sphaeroderus canadensis) showed strong potential as indicators of partial harvesting disturbances.

Two other common forest specialists, Pterostichus pensylvanicus and Agonum retractum, did not appear to act as indicators of harvesting intensity. Abundances of Agonum retractum were fairly constant between treatments and were only reduced in ICR. It was, however, negatively correlated to early-decay DWD abundance in July 2016, which was highest in ICR.

Pterostichus pensylvanicus also exhibited no significant treatment effect, but responded negatively to local basal area. There was no relationship to block basal area, perhaps indicating that lower basal area within close proximity was more important than conditions across the surrounding block. Both species were also found not to differ between mature logged stands and recent STS cut stands in Algonquin Provincial Park (Vance & Nol, 2003).

Pterostichus pensylvanicus was more common in blocks with more large, late-decay

DWD and expressed the only positive relationship found with DWD resources in this chapter, although the relationship was highly influenced by CON1. Findings on DWD associations of

Pterostichus pensylvanicus in the Northern Ontario boreal forest are conflicting (Piascik, 2013), but other species found in this study (Pterostichus coracinus, Platynus decentis (Say, 1823),

Pterostichus melanarius (Illiger, 1798), Loricera pilicornis (Fabricius, 1775), Agonum retractum, and Synuchus impunctatus) were all associated with DWD resources (Pearce et al.,

2003; Piascik, 2013). In the present study, harvesting resulted in increases in early-decay DWD; however, large, late-decay DWD supplies were low and varied little between treatments. In

Northern Ontario, carabid associations with DWD were found to be stronger in clear-cut rather than areas of mature forest (Pearce et al., 2003). Carabids rely on DWD as a source of shelter, moisture and coolness, which are more limiting in open areas and could explain this occurrence.

This suggests that DWD may be less important as source of shelter in partially harvested hardwood forest compared to more open clear-cut-managed forests.

Four other carabid species (Cymindis cribricollis Dejean, 1831, Pterostichus tristis

Dejean, 1828, Pterostichus mutus (Say, 1823) and Platynus decentis) were common enough for analysis and did not appear to be strongly influenced by harvesting intensity. The remaining 12 rare species combined for 30 individuals across the three trapping periods and accounted for under 4% of total captures. Half of these rare captures came from the four ICR treatment blocks; together FMS and DLC accounted for eight, CON1 accounted for five, and the remaining 11

CON and STS blocks accounted for only two. Of the 12 species, the majority are not forest specialists and instead are associated with open habitats or wetlands rather than forest habitats

(Appendix 2). The tendency of these rare species to be captured in blocks treated with high intensity systems, most notably ICR, supports the idea that rare species uncommon in forest habitats experience increased abundances after harvests, even partial harvests (Klimaszewski et al., 2005). Two species (Agonum fidele and Pterostichus luctuosus (Dejean, 1828)), were

25 captured only within ICR3, ICR4 and DLC24 at the southeastern edge of the site (Fig 1.1). This area was not far from a wetland, which was under 150 m away from each of the three traps, and both of these species are commonly found in wet areas. These species may therefore have been captured in these blocks due to wet rather than open conditions. It is also possible, however, that harvesting impacts could contribute to the wet conditions. Fewer trees means less evapotranspiration (Tromp-van Meerveld & McDonnell, 2006), and soil compaction can lead to poor drainage and puddling after spring thaw or rain events (Brais & Camiré, 1998).

Not surprisingly, the various logging treatments resulted in strong changes in tree basal area and distributions along the varimax logging axis that were aligned with retention targets.

Prescribed diameter distributions were also reflected through values for small and large tree basal area. Removal of small and large trees was more balanced for uneven-aged systems, although FMS focused more on large stems. The diameter-limit for DLC led to basal removal being heavily skewed towards larger stems, which resulted in high small tree retention. FMS and

DLC were very similar in total basal area, but had different diameter distributions. This demonstrates how total harvested basal area alone cannot fully describe the effects of a prescription. Large and small trees have different roles in ecosystems and may contribute differently to canopy openness (Goodburn & Lorimer, 1999; Vuidot et al., 2011). Early-decay

DWD followed the effects of basal area removal, which is expected given the pulse of fresh

DWD resulting from the tree crowns left on site. This DWD input could be beneficial to organisms that rely on DWD resources, but it is made of primarily small pieces, which appear to be of less importance to carabids than other classes of DWD (Piascik, 2013). Both late-decay

DWD and snags were most abundant in control plots (although not significantly so) and both were negatively correlated with the amount of early-decay DWD. Control plots, however, had

26 much lower DWD stocks than old-growth forests (Vanderwel et al., 2008), meaning that effects of the recent harvest could be diminished due to changes experienced during earlier rotations.

Silvicultural systems with lower residual basal area targets were found to drive forest structure and carabid community composition further from conditions found in uncut control plots. This idea places ICR, and to a lesser extent DLC and FMS, as systems with the highest risk for ecological damage. FMS and DLC resulted in stronger effects than STS but were far from what was seen in ICR. A key question concerning the net ecological effects of harvesting systems is the likelihood that populations will achieve pre-harvest conditions prior to the next harvest. If harvesting impacts alter habitat structure beyond the point of recovery, then continued application of the system would presumably lead to a shift in community structure from the dominance of forest dwelling species to stronger components of species common in open areas.

Widespread use of the harvesting system without appropriate reserves could thus draw the sustainability of the overall management system into question. The present chapter, however, looks exclusively at the immediate level of basal area removal and does not address the prospect for subsequent tree growth and regeneration. While greater basal area removal pushes forest structure further from pre-harvest conditions, it also increases light availability, creating opportunities for tree growth and regeneration (Miller & Smith, 1993). This higher regeneration potential could thus help to offset heavier losses in basal area and in turn increase the chances of a full recovery. Dead wood supplies are also of particular concern, although in this case the main impacts appear to have resulted from harvests prior to the experiment.

The Blue Heron Demonstration Forest treatment design provided a suitable site for this study, but the 84 m by 84 m treatment blocks were much smaller than entire stands that are the usual target of partial harvesting management. Forest stands can range in size depending on the

27 landscape, but the Ontario provincial tree marking guide includes guidelines for stands of over

20 hectares (OMNR, 2004). For responses of structural and habitat variables to the experimental manipulation, this difference in size should have little effect, but implications for carabid abundances are more complicated, and likely depend on both the habitat preferences and dispersal abilities of individual species. This complexity makes outcomes hard to predict, but mosaics of smaller harvested forest patches tend to be more favorable to forest species and support generally higher species diversity than larger harvested areas (Koivula et al., 2002b).

Differences in abundances of Pterostichus coracinus and Synuchus impunctatus between recently logged and unlogged stands in Vance & Nol (2003) were clearer than findings in my study, which supports this idea. If it is true that the Blue Heron Demonstration Forest experimental design is dampening the effects of partial harvesting on carabid abundances, then the findings of this study could be underestimated in comparison to stand level differences occurring in typical silvicultural applications.

In conclusion, carabid responses to partial harvesting observed in my study draw parallels to patterns described after clear-cutting, but were less pronounced. This effect was observed to be stronger under silvicultural systems that removed greater amounts of tree basal area. Findings highlight the role of the extent of basal area removal in the long-term ecological sustainability of partial harvesting silviculture in northern hardwood forests.

Chapter 2: Forest management history and ground beetle (Coleoptera: Carabidae) communities in the northern hardwood forests of Central Ontario

Introduction

The implementation of single-tree selection (STS) has become widespread across the managed northern hardwood forests of Ontario and is currently the principal silvicultural system in the region, particularly on public lands (OMNRF, 2015). With its beginnings dating back as early as the 1950s in areas including Algonquin Provincial Park (Quinn, 2004), many STS- managed stands have now experienced multiple harvesting rotations under the system. Since

STS is such a prevalent silvicultural system, any associated ecological effects could result in broad negative impacts across the managed hardwood forests of Central Ontario, making monitoring particularly important. The fact that STS has now been in use for multiple harvesting rotations allows for more effective evaluation of potential effects that could have implications for long-term ecology sustainability.

Continued harvesting can lead to continued structural simplification and diminished frequency and quality of DWD and snags, which is of particular concern (Angers et al., 2005;

Siitonen, 2001). Saproxylic species, which depend on DWD at some point in their life cycle, collectively make up an important component of forest biodiversity (Freedman et al., 1996;

Siitonen, 2001). DWD is also an important source of organic matter in forest ecosystems and contributes to soils upon decay (Harmon et al., 1986). Snags support saproxylic species through microhabitats found in loose bark and decaying wood, and provide a source of cavities and foraging opportunities for birds and mammals (Freedman et al., 1996).

Gap creation during STS may also be problematic in that it can result in frequent, regular small gaps, unlike the more sporadic and often larger gaps created by blowdowns and other natural disturbances in old-growth forests (Quinn, 2004). STS gap structure has been found to promote sugar maple dominance and limit regeneration opportunities for mid-tolerant species, such as yellow birch, that require larger gaps for adequate light conditions (Halpin et al., 2017).

DWD is also an important substrate of yellow birch establishment, posing another potential threat to their abundances in STS-managed stands (Shields et al., 2007). Yellow birch are of particular ecological value due to their provision of foraging habitat for birds and other species

(Schwartz et al., 2005; Shields et al., 2007).

Although not a specific concern to STS, which retains more basal area than many silvicultural systems, forest cover reductions also pose a potential threat to biodiversity. Species that rely on mature forest conditions may no longer find the stand suitable even at the time of the next rotation, whereas early-successional species and others favoring more open conditions may increase in abundance (Doyon et al., 2005; Vanderwel et al., 2009).. The loss of mature trees, tree species diversity, snags, and DWD could mean changes in communities and a reduced diversity of microhabitats (Vuidot et al., 2011), highlighting the importance of continued monitoring. Together, these logging effects have the potential to contribute to overall landscape homogeneity. Homogenization can also increase susceptibility to threats such as pests, disease, climate change and other forms of disturbance (Schulte et al., 2007). If the effects of STS are strong enough so that pre-harvest conditions are not recovered by next harvest, effects may accumulate, and sustainability would be compromised.

Studies contrasting ground beetle (Carabidae) communities between old-growth and mature regenerating clear-cut-managed stands in boreal forests have observed strong carabid

30 community differences and identified several old-growth specialists (Niemelä et al., 2007;

Spence et al., 1996). In northern hardwood forests, carabid community differences between old- growth and mature logged stands were limited to the abundances of relatively few species and provided little evidence of vulnerable old-growth specialists (Latty et al., 2006; Vance & Nol,

2003; Werner & Raffa, 2000). Higher abundances of one species (Pterostichus coracinus) in mature logged stands was common across northern hardwood studies and it was the key species in distinguishing community differences related to logging history. As outlined in Chapter 1, saproxylic behaviour in carabids in northern hardwood forests has received little attention, but has been documented in the boreal forest of northern Ontario (Pearce et al., 2003; Piascik, 2013).

This study aims to assess the extent to which logging history has influenced current forest communities, with a focus on carabids as a bioindicator group, as well as the likelihood of sustainability in current forest management. Specific objectives are three-fold. First, I compare carabid communities between mature logged and old-growth stands, with communities expected to be similar, but with the possibility of differences at the species level, particularly higher abundances of Pterostichus coracinus in mature logged stands. Secondly, I compare forests structure and tree species composition between mature logged and old-growth stands, with snags and DWD expected to be more abundant in old-growth forests, along with greater basal area, abundance of large trees, and greater leaf litter depth. Mature logged stands may also show greater sugar maple dominance, lower abundances of coniferous species and mid-tolerant hardwood species such as yellow birch, as well as overall stronger homogeneity in forest structure and tree species composition as compared to old-growth stands. Finally, I investigate relationships between forest structure variables and carabid communities, with variables related to harvesting expected to show the strongest relationships. Relationships with DWD are expected

31 depending on saproxylic behavior of species captured; basal area and species composition may also influence carabid communities.

Methods

Study Area

The study was conducted in Haliburton Forest and Wild Life Reserve, a privately-owned forest located in Central Ontario (45°13' N, 78°35' W) in the Great Lakes-St. Lawrence forest region (Rowe, 1972). Forest cover is dominated by sugar maple (Acer saccharum Marsh.) in association with American beech (Fagus grandifolia Ehrh.), yellow birch (Betula alleghaniensis

Britt.), and eastern hemlock (Tsuga canadensis (L.) Carrière). The mean annual temperature in the region is 5.0 °C, with monthly means ranging from −9.9 °C in January to 18.7 °C in July, and an average of 1074 mm of annual precipitation (Environment Canada, 2018). Logging in the area dates back to the removal of white pine beginning in the mid-1800s followed by the harvesting of other target species such as yellow birch in the first half of the 1900s. More organized forest management began in the 1960s with the use of selective systems and a shift to modern selection systems in the 1970s and 1980s. Old-growth stands remain in areas where terrain impeded access

(Vanderwel et al., 2008).

Site selection and establishment

Six mature logged and three old-growth stands were selected as study sites (Appendix 1).

Mature logged stands had an average age of 23 and a range of 20 to 28 years since previous

32 harvest. These stands had been selection-cut at least twice, with a prior history of selective harvesting. Old-growth forests had no record of management history and site inspections revealed no stumps created by harvesting, including no large, heavily-decayed stumps indicating past high-grading of conifers. All stands had high concentrations of sugar maple, were of relatively flat topography, and were not close to waterbodies, wetlands, and other study sites. A

420-m long transect (Fig. 2.1) was established at each site to be the basis of both carabid sampling and forest measurements. Transects were placed within 30 m of the stand boundaries

(FRI) and were as straight as possible, but in some cases they spanned two stands of equivalent composition. Six primary stations were located at 84-m intervals along the transect, with five secondary stations evenly spaced between them. Perpendicular “fishbone” transects of 100 m were established in intersection with the six primary stations and their lengths were split evenly on both sides of the main transect (Fig 2.1).

Beetle Sampling

During two distinct sampling periods (June 19-28, 2017, and July 19-28, 2017), carabids were trapped at primary stations using a "drift fence" pitfall array at each (Piascik,

2013). Each array was in the shape of a "Y", consisting of a central pitfall bucket and three buckets in the shape of an equilateral triangle, each positioned 1.2 m away from the central bucket. The center bucket was connected to each end bucket by a 30 cm high geotextile sheet buried 10 cm deep in the ground and held in place by wooden stakes flush with the bucket edges.

The geotextile acted as a "drift" fence by guiding beetles to fall into the buckets at either end.

The tops of buckets lay flush with the ground and had dimensions of 11 cm in diameter and 13.3 cm in depth. When not in use, buckets were covered by plastic lids and leaf litter.

Traps were set by removing the bucket lids and filling each bucket one-third full of 5% salt solution including a small amount of dish soap. Traps were left active for six consecutive nights, then trap contents were collected and preserved in 70% ethanol. Each collection contained the contents of all four buckets of an array. Traps were set during the two sampling periods: one in June 2017 and one in July 2017, and traps at the various sites were all set within three nights of each other. Carabids were pinned and identified using keys from Bousquet

(2010). Carabid species were assigned to ecological groupings based on habitat classification systems used in other studies (Latty et al., 2006; Vance & Nol, 2003) as well as species descriptions found in Larochelle & Larivière (2003).

To correct for missing effort, abundances were standardized to captures per 100 bucket- nights. Following Piascik (2013), an outer bucket was counted as one bucket and the center bucket counted as 2.236 buckets. Each pitfall array trapped for six nights thus totaled for 31.4 bucket-nights. If more than one outer bucket (i.e., six bucket-nights) was compromised, then the entire sample was excluded from analysis. This occurred in 5 out of 36 arrays in mature logged stands sampled in June 2017 and in 1 out of 18 arrays in old-growth stands sampled in July 2017.

Missing effort was otherwise rare (<5% of bucket-nights).

Habitat Measurements

In the summer of 2017, BAF2 prisms sweeps were conducted at both the primary and secondary stations. All live trees found to be "in" and of a diameter at breast height (DBH) of

≥10 cm were identified to species and measured to DBH. I used prism sweeps rather than fixed- area plots because they were less labor intensive and provided a more comprehensive sample of large stems within a stand. Data obtained from prism sweeps were then used to calculate total

34 basal area and percent sugar maple composition. Basal area of relatively small (<38 cm DBH) and large (≥38 cm DBH) yellow birch stems were also calculated. This size classification represents the distinction between small and medium saw logs in Ontario forest management

(OMNR, 2004) and was used to separate the basal area of large, mature stems from basal area of younger yellow birch trees whose growth would have more likely been affected by STS. Percent density was calculated in 3-cm size classes in order to characterize the diameter distribution of each stand, to which 2-parameter Weibull distributions were fit. The shape parameter describes the shape of the probability density function, with small values indicating shapes similar to a negative exponential function and larger values indicating a normal distribution. Scale represents the DBH class at which 63.2% of total tree density has accumulated. In the context of tree diameter distributions in this study, higher shape values corresponded to wider spread in density percentages across size classes, whereas higher scale values corresponded to a higher average

DBH. Stand heterogeneity was measured using within-transect semivariance calculations for both basal area and percent composition of sugar maple, where greater semivariance values indicated greater heterogeneity from station to station within a transect. It was calculated by taking the mean of absolute differences between adjacent stations along an individual transect.

The transect means were then partialled out using linear regression to derive a measure that was statistically independent of the mean, although variance was not removed because it contributes to the characterization of heterogeneity.

Spatial relationships between transect locations and surrounding conifer cover were quantified using GIS-based mapping techniques. Satellite photos of each stand (DigitalGobe images in Google Earth; dated March 4, 2015 and April 18, 2015) both taken in spring, but before leaf-out, were classified into areas of conifer or non-conifer using supervised

35 classification. Conifer cover scores for each station were then generated by using the heat conduction model of Malcolm et al. (2017) in which the number of interactions was chosen such that fit between model values and variation in species-specific beetle abundances was maximized. Here, conifer pixels were heat sources, and heat diffused into the surrounding non- conifer landscape. Modelled temperatures acted as proxies of local conifer abundances, weighted by distance.

DWD was sampled via the line-intersect method (Wagner, 1968) along both the main transect and fishbones, totaling 1020 m (Fig. 2.1). Any piece of intersecting DWD ≥10 cm in diameter had its diameter measured and was categorized with respect to size (less than, or equal to or greater than median diameter [18.7 cm]) and decay class. Decay class was determined as either "early" or "late" decay depending on the condition of the outer wood layers below the bark. If they were spongy or easily damaged by kicking, they were classified as late-decay; otherwise they were early-decay. These assignments corresponded roughly to decay classes 3-5 and 1-2 from Maser et al. (1979), respectively. DWD volumes were calculated for size and decay classes.

Snags were measured in a belt prism transect: that is, any snags of ≥10 cm DBH found to be “in” from any point along the 420-m long transect were measured to DBH. As previously, they were categorized with respect to size (less than, or equal to or greater than the median diameter [33.3 cm]) and decay class. Decay class was determined as either "early" or "late" decay depending on the extent to which the tree crown structure and bark of the trunk remained intact. An early-decay snag retained over 50% of coarse branches and bark may have been beginning to loosen, while a late-decay snag retained fewer than 50% of coarse branches, and bark was loose or completely sloughed off. These assignments corresponded roughly to decay

36 classes 1-2 and 3-5 from Anderson and Rice (1993), respectively. Snag basal area was calculated for size and decay classes.

One leaf litter depth sample was taken at each primary station. A small hole was dug at a location on the forest floor void of shallow roots and stones and within 1 m of the pitfall array.

The leaf litter depth layer was then measured to the nearest centimeter.

Statistical Methods

Carabid abundances and habitat features were compared between treatments using t-tests.

General and generalized linear regression was used to explore relationships between carabid abundances and habitat variables. Negative binomial regression was deemed appropriate if the deviance value was between 0.8 and 1.2. If the deviance value was below 0.8 then, then log transformations were used on carabid abundances to bring the value to within the required range before testing. Statistical tests were undertaken using SAS (v. 9.4).

Ordination techniques were used to investigate carabid community composition, habitat variation, and treatment effects. Principal component analysis (PCA) on the correlation matrix was used for habitat variables. A PCA was also used for carabid communities, but on the covariance matrix, which gave less weight to rare species. Redundancy analysis (RDA, 9999 permutations) was used to constrain carabid abundances using habitat variables at both transect and array levels. All ordinations were performed using Canoco (v. 4.5).

Results

Carabids

In total, 799 individuals were captured of 19 species. Agonum retractum (50%)

Pterostichus coracinus (13%) and Pterostichus pensylvanicus (12%) made up the majority of

June 2017 captures, whereas Synuchus impunctatus (41%), Pterostichus coracinus (23%) and

Agonum retractum (19%) were dominant in July 2017. There were nine species exclusive to June

2017 and four species captured only in July 2017.

Differences in abundances for each of the species were not significant between mature logged and old-growth stands (Table 2.1). In the June 2017 PCA, old-growth communities were similar amongst the three sites and were distinguished from four of the six logged sites, primarily along the second ordination axis (Fig. 2.2). Pterostichus coracinus was the most important species on the second axis and showed the strongest sign of a treatment effect of any individual species (Table 2.1), with highest abundances in mature logged stands. No sign of a treatment effect was apparent in the July 2017 PCA (Fig. 2.2). In June 2017, common species including

Agonum retractum, Pterostichus coracinus and Pterostichus pensylvanicus were slightly more abundant in mature logged stands and rare species captures were more frequent, with six species only found in that sampling period (Fig 2.3), although twice as many mature logged sites were sampled compared to old-growth. These tendencies, however, were not consistent in July 2017, where abundances of common species, including Synuchus impunctatus and Agonum retractum were higher in old-growth stands and rare species were not more common in one treatment (Fig

2.3).

Forest structure and tree species composition

There were strong differences in forest structure between mature logged and old-growth sites, and treatments were distinguished along the first ordination axis (Fig. 2.4). High snag basal area and DWD occurrences, particularly of large size classes (Table 2.2), were most important in defining differences between mature logged and old-growth treatments. Basal area was also significantly higher in old-growth stands, as well as conifer cover scores, which was the only tree species composition variable to exhibit a treatment effect. Remaining variables including

Weibull parameters, semivariance values, leaf litter depth, and percent sugar maple were not significant between treatments (Fig. 2.4). Percent sugar maple composition was the strongest variable (longest vector) along the second ordination axis and was therefore important in explaining variation independent of logging effects. Small yellow birch basal area was positively associated with mature logged stands and was nearly significant (Table 2.2).

Relationships between carabids and habitat variables

In redundancy analyses, habitat variables were only significant in explaining carabid variation at the station level in July 2017. In that ordination, small early-decay DWD volume, conifer cover score, and percent sugar maple composition were most important (Table 2.3).

These were also the strongest variables in describing variation at the transect level, but not significantly. Tree species composition variables were also important in June 2017, along with large, late-decay DWD and basal area semivariance at the transect level (Table 2.3).

Overall, tree species composition variables appeared to be most important in explaining carabid abundances, with the dominant species in each month (Agonum retractum in June 2017

39 and in Synuchus impunctatus in July 2017) showing negative associations with conifer cover, while Pterostichus coracinus was associated with low percent sugar maple composition in July

2017 (Table 2.4). DWD was also important in describing carabid variation, but was primarily negatively related to species abundances: Agonum retractum was less common in areas of higher volumes of large, late-decay DWD, while Synuchus impunctatus was negatively correlated to small, early-decay DWD volume and Pterostichus coracinus was negatively correlated to large, early-decay DWD volume. However, Sphaeroderus canadensis and Agonum retractum were both positively associated with volumes of late-decay DWD (Table 2.4). Sphaeroderus canadensis was the only species to have a significant relationship to basal area, being more common in stands of higher basal area (Table 2.4). Based on relationships with semivariance variables, heterogeneity also appeared to have implications for carabids communities, specifically in June 2017, when abundances of Pterostichus pensylvanicus and Pterostichus coracinus were negatively correlated to basal area semivariance and percent sugar maple composition semivariance respectively (Table 2.4). Pterostichus tristis abundances were higher in stands of shallow leaf litter depth in June 2017 and this was also true for Pterostichus coracinus in both months (Table 2.4). This was the only relationship to occur in both sampling periods.

Discussion

Carabid communities were similar between mature logged and old-growth stands, suggesting a recovery since the previous harvest. Communities were also similar between STS- managed and old-growth stands in Algonquin Provincial Park in Ontario (Vance & Nol, 2003),

40 although some species differed significantly between the treatments, the strongest being higher abundances of Pterostichus coracinus in mature logged compared to old-growth stands. Single- tree selection used in stands in Vance & Nol (2003) would have been very similar to silvicultural practices in Haliburton Forest, but stand age was slightly younger in Vance and Nol (2003) than in this study, with some stands as young as 15 years since harvest. In Michigan and Wisconsin, carabid communities were also similar between mature logged and old-growth stand (Latty et al.,

2006; Werner & Raffa, 2000), but with more apparent community level differences as described by a split between mature logged and old-growth communities along the first axis of an NMDS ordination. Similar to findings in Vance & Nol (2003), higher occurrences of Pterostichus coracinus in mature logged stands was the most important treatment effect found at the species level. In the studies in Wisconsin and Michigan, stands were even younger and cut at a minimum four, but up to 13 years prior to carabid sampling. The selection system used in this northern region of the Great Lakes States is also more intensive than STS in Ontario, with an 8-15 year cutting cycle, a minimum 16.1 m2/ha residual basal area, and maximum tree diameter of 45 cm

DBH (Goodburn & Lorimer, 1999), rather than a 20 year cycle, 18-20 m2/ha basal area target, and 60 cm DBH maximum diameter used as guidelines for STS (OMNR, 2004). Overall, findings across studies support high levels of recovery, but evidence for sustainability (i.e., a return to unlogged conditions) was strongest in the present study. This suggests potential for sustainability, but that success is sensitive to the nature of the silvicultural system and time between harvests. Specifically, it appears that systems with higher basal area retention, and longer time periods between repeated harvests, show the highest potential for sustainability.

Here, I use the term "sustainable" in the sense that if at least some of the stands in a harvested

41 landscape have communities that are characteristic of old-growth forests, then there are greater chances of the landscape maintaining old-growth communities over time.

Although I found no difference between the treatments, carabid species did respond to treatment-related variation in habitat variables. DWD volume was primary inversely correlated with carabid abundances, and low abundances of Synuchus impunctatus in areas of high volumes of small, early-decay DWD was the strongest of such relationships. This association was also made in Chapter 1 and was suggested to be related to tree removal and thus canopy openness. In this chapter, however, I found no relationship between this form of DWD and basal area, suggesting that it may have a direct influence on Synuchus impunctatus abundances.

Sphaeroderus canadensis was the only species correlated to basal area and the relationship was positive. This finding agrees with its description as a forest specialist, and a similar relationship to basal area was apparent in Chapter 1. Reduction in leaf litter depth is another harvesting- affected forest feature that was found to be related to carabid species abundances. Carabid relationships to both percent sugar maple composition and basal area semivariance values also suggest potential impact of forest homogenization, should it occur over time.

Tree species composition variables, particularly conifer cover scores, were strongest in describing carabid variation. Both Agonum retractum and Synuchus impunctatus were dominant species that were negatively associated with conifer cover and have previously been found to be more abundant in deciduous forests (Work et al., 2008). Pterostichus coracinus was less common in stands of high sugar maple composition in July 2017, and has been associated with mixed and coniferous forests (Work et al., 2008). Tree species composition differences between the treatments could be attributed to the long-term decline in the abundances of coniferous forests and increase in sugar maple abundance in the Great Lakes-St. Lawrence forests of

Ontario, but site-specific terrain differences could have been important in dictating surrounding tree species composition as well. Regardless of why old-growth stands tended to be associated with greater surrounding conifer cover scores in this study, findings of relationships between carabids, conifer cover and sugar maple composition are relevant to ecological changes that may have followed landscape level shifts in tree species composition associated with logging history in the region. Despite the possibility that tree species composition may be influenced by logging history, I found that structural differences remain the most important features in distinguishing mature logged and old-growth forests. The fact that tree species composition was generally more important than structural features in explaining variation in carabid species suggests that site level differences in terrain and forest type may be of greater influence on carabid communities than logging impacts in mature logged forests, and therefore can be interpreted as evidence for ecological recovery.

Low quantities of DWD and snags as well as lower basal area were consistent with mature logged forests in other studies (Angers et al., 2005; Goodburn & Lorimer, 1999; Hale et al, 1999; Vanderwel et al., 2006a). A difference in conifer cover scores was also significant, and as described above, may be due to site effects rather than logging history. Old-growth sites were found in areas of dissected terrain, where difficult access may have prevented logging. As a result, the maple stands that I sampled were in closer proximity to areas of steeper inclines and edges of waterbodies, which could explain greater conifer cover. It is therefore difficult to conclude that this compositional effect was truly a logging effect. Reduced opportunities for yellow birch establishment is a well described potential consequence associated with STS, but I found the opposite, although not significantly so. The role of site-specific forest type likely plays a role here, but perhaps silvicultural practices in Haliburton Forest tend to maintain a canopy gap

43 structure better suited to the recruitment of yellow birch than elsewhere across northern hardwood forests.

Old-growth stands were predicted to have higher heterogeneity in both basal area and percent sugar maple composition, and stem diameters were expected to be distributed more broadly across size classes, specifically in classes of larger diameter. High values for scale and low values for shape were thus as predicted. These predictions relating to diameter distribution and stand heterogeneity were not confirmed, but variables generally behaved in the predicted direction between treatments (Fig. 2.4). Diameter distribution can see considerable recovery between harvests, but long-term changes in structural and compositional heterogeneity may be more difficult to reverse.

In conclusion, I found that carabid communities were similar between old-growth and mature logged sites, despite strong structural differences between the two stand types. These findings provide evidence of ecological sustainability in selection logging in northern hardwood forests, but comparisons with other literature suggests that this sustainability may be sensitive to initial harvesting intensity and harvest rotation length.

General Conclusions

Since the movement away from selective harvesting in the mid to late 1900s, single-tree selection (STS) has been the principal silviculture system used in the management of northern hardwood forests in Central Ontario (Quinn, 2004). Minimizing ecological impacts is a priority in STS management, but the realization of these intentions must be monitored to assess sustainability. Ecological conditions are difficult to measure in their entirety, therefore I used carabids as bioindicators of forest management. The findings of this thesis highlight the role of basal area removal in initial ecological response to harvesting, but provides evidence for potential recovery before the next harvest, and therefore greater likelihood of sustainability over time. With increased intensity of basal area removal among recently-cut treatments, carabid communities displayed decreases in abundances of species common to uncut control plots and a shift toward the importance of uncommon and rare species. This effect was consistent with responses to clear-cutting in boreal forests, although less pronounced. Chapter 1 highlighted the implications of initial basal area reductions and gave insight into differences in initial impacts of past, current and potential future systems of use in the northern hardwoods of Central Ontario. In contrast, mature logged and old-growth stands were structurally distinct, but there were no significant differences in carabid communities between the treatments. The primary difference between treatments was higher quantities of snags and DWD in old-growth stands, but an overarching dependence on the resources by saproxylic carabid species was not supported, and carabid relationships with DWD were more often negative than positive.

Instead, the strongest variables in predicting carabid variation were related to tree species composition, specifically conifer cover. Although conifer cover was found to be higher in the

45 areas surrounding old-growth stands, this may be attributed to topography rather than logging history. Altogether, these findings suggest that at the stand level, and from the perspective of ground beetles, current mature logged northern hardwood forests provide similar conditions to old-growth forests. This was found despite differences in forest management history and structural components, and suggests that recovery since previous harvest was sufficient to support long-term sustainability.

Three species were found to have positive correlations to DWD volume: Pterostichus pensylvanicus was associated with large late-decay DWD in Chapter 1, although this relationship was driven primarily by the experimental block of CON1, Agonum retractum was associated with large late-decay DWD in Chapter 2 at both transect and station levels, and Sphaeroderus canadensis was associated with small late-decay DWD in Chapter 2 at the station level.

Pterostichus pensylvanicus and Agonum retractum were also found to be positively correlated to

DWD volumes in the boreal forests of Northern Ontario, along with other common species

(Synuchus impunctatus and Pterostichus coracinus), uncommon and rare species found in this study (Platynus decentis, Pterostichus melanarius and Loricera pilicornis) and several species that were not captured in this study (Pearce et al., 2003; Piascik, 2013). All correlations between carabids and DWD and other habitat variables, with exception to the relationship between

Pterostichus coracinus and leaf litter depth, were only found during one sampling period, suggesting that carabid response to habitat conditions may be temporally dependent. The species structure of carabid communities in the Great Lakes-St. Lawrence region is known to change significantly across the year (Adlam et al., 2017; Barlow, 1970; Epstein 1990; Levesque, 1994).

These changes are dependent on the combination of seasonal activity patterns and reproductive

46 periods of individual carabid species, which have been found to coincide with climate conditions, specifically temperature fluctuations (Barlow, 1970; Boer, 1981; Kotze et al., 2012).

Across northern hardwood studies, Pterostichus coracinus has been the strongest indicator species of forest management (Latty et al., 2006; Vance & Nol, 2003; Werner & Raffa,

2000) and the species is consistently abundant from early spring to late fall across Eastern

Canada (Barlow, 1970). Not only is it an indicator of mature logged over old-growth forests, but low abundances after recent harvest in comparison to mature logged forests were found in

Chapter 1 and described in Vance & Nol (2003). Higher abundances were found in forests of intermediate disturbance levels in Vance & Nol (2003) and this trend is clear across treatments of varying basal area from Chapters 1 and 2 of this thesis (Fig. 3.1). Higher abundances of

Pterostichus coracinus in control plots in Chapter 1 compared to mature logged transects in

Chapter 2 may be a result of control plots being situated amongst a mosaic of recently harvested forest blocks. Occurrences within the Blue Heron Demonstration Forest may thus be influenced by both the basal area within local treatment blocks and the broader fragmented site of varying basal area levels. The fact that Pterostichus coracinus was not significantly different between treatments in this chapter is a strong piece of evidence for sustainability, particularly in light of other studies where it was more abundant in mature logged stands (Latty et al., 2006; Vance &

Nol, 2003; Werner & Raffa, 2000).

Synuchus impunctatus is another species that has been found to respond to forest disturbance, but has not been found to differ in abundance between mature logged and old- growth stands. In Chapter 1, Synuchus impunctatus was less abundant in treatments of higher harvesting intensity and responded similarly to strong disturbances in Vance & Nol (2003), with lower abundances in recently logged compared to mature logged stands. Synuchus impunctatus

47 abundances were also found to be lower in affected areas within two years following the 1998 ice storm in southern Quebec (Saint-Germain & Mauffette, 2001). Contrary to these findings,

Synuchus impunctatus is known as a generalist, and was more common within 12 and 13 year old strip clear-cuts compared to adjacent mature STS-logged hardwoods in Moore et al., (2004), and was described as an open-habitat species in the Canadian boreal forest (Spence et al., 1996).

Responses to harvesting may depend on the region or forest type (Work et al., 2008), but these findings suggest that Synuchus impunctatus abundances may be highest in forests of intermediate disturbance, similar to Pterostichus coracinus. Affinity for forests of intermediate disturbance is also consistent with the designation of generalists for both of these species. Of all species,

Sphaeroderus canadensis displayed the strongest evidence of dependence on mature forests. It was less abundant in treatments of higher harvesting intensity in Chapter 1, and was positively correlated to old-growth related variables, including basal area and small late-decay DWD in

Chapter 2. Sphaeroderus canadensis was also more abundant in old-growth than mature logged stands in both months, but not significantly. It has been described as a forest specialist, but otherwise the ecology and behaviour Sphaeroderus canadensis has not been described in detail

(Larochelle and Larivière 2003). It has only been caught in relatively small numbers in studies in

Ontario and Quebec (Moore et al., 2004; Vance & Nol, 2003), and was not captured by Werner

& Raffa (2000) in Michigan and Wisconsin. This is consistent with its description as an eastern species (Work et al., 2008). Many of the uncommon or rare species in the Blue Heron

Demonstration Forest were not known forest specialists, but were rather generalists or associated with either open or wet habitats (Appendix 2 and 3). These species were not abundant enough for analysis, but collectively, these rare species contributed to logging intensity responses in Chapter

1. Examples of such rare species were present in Chapter 2 as well, but displayed no evident treatment effects.

A weakness of this study was small sample size due to both a low number of study sites and limited total carabid trapping effort. Only three old-growth stands were sampled due to their rarity in Haliburton Forest. Along with being uncommon, old-growth stands were isolated within the surrounding managed landscape, which may support different community assemblages than a greater area of continuous unlogged forest (Niemelä et al., 2007). Carabid trapping only took place over two six-night periods (one in June and one in July) in 2017, with the exception of the

July 2016 sample in the Blue Heron Demonstration Forest. This made for a relatively small sample of carabid individuals within a narrow timespan. Capture numbers, however, were sufficient to perform both community and species level analysis for each sampling period individually. Investigating samples separately provided snapshots of carabid communities, which change across the reproductive season. This approach therefore avoids the potential dilution of effects caused by pooling data from ever-changing communities across a season, although different samples could still be compared and analyzed as a whole. Vance & Nol (2003) is the previous study with the most similarities to my thesis in that they compared carabid communities between STS-managed and old-growth stands in Central Ontario. Overall, conclusions were similar to what was found in my thesis, but they found stronger carabid community differences between mature logged and old-growth stands, particularly higher abundances of Pterostichus coracinus in mature logged stands. This could be attributed to the slightly older mature logged stands in this thesis. Synuchus impunctatus and Pterostichus coracinus were caught in lower abundances in recently harvested rather than mature logged stands, although these effects were less evident in my study. The heterogeneous matrix of treatment blocks of varying basal area

49 within the Blue Heron Demonstration Forest may have resulted in weaker differences in conditions among treatments in comparison to using entire stands as experimental units as in

Vance & Nol (2003), and could potentially explain the difference in findings between the studies. Similarities to carabid response to clear-cutting, however, were more evident in my study than in Vance & Nol (2003), and were most common in silvicultural treatments of higher intensity that were unique to this study, including DLC and ICR. The comparison of carabid communities between treatments within the Blue Heron Demonstration Forest allowed for the evaluation of differences in ecological impacts among multiple silvicultural systems of relevance to forest management in the northern hardwood forests of Central Ontario within a small area of common environmental conditions. This enabled findings of a fine gradient in carabid response to multiple levels of basal area retention. The investigation of relationships between carabids and harvesting-related habitat variables of forest structure and tree species composition was another strength of this study and is an area of research that has not been thoroughly explored in the northern hardwoods of the Great Lakes-St. Lawrence region.

The addition of intermediately-aged logging treatments and sites differing in number of harvesting rotations under STS-management would help to better understand how carabid communities respond to harvesting impacts across the harvesting cycle, to better assess the potential for accumulation of harvesting impacts across rotations, and consequently the long- term sustainability of STS. This thesis focuses primarily on stand level effects, but sampling of carabids in close proximity to DWD would presumably be more effective in testing for saproxylic behavior.

In conclusion, this thesis provides evidence that ecological conditions may be sufficiently recovered to support sustainability across rotations of STS, despite differences in structure

50 elements associates with old-growth stands, such as snags and DWD. The likelihood of sustainability, however, may be sensitive to initial harvesting intensity and harvesting frequency.

Literature Cited

Adlam, C., Despland, E., & Beaulieu, F. (2017). Spatial and Seasonal Patterns of Ground Beetle (Coleoptera : Carabidae) Activity Across Habitats At the Mont St . Hilaire Biosphere Reserve, Quebec. Journal of the Entomological Society of Ontario, 148, 23–38.

Anderson, H.W., & Rice J.A. (1993). A tree marking guide for the tolerant hardwoods working group in Ontario. Ontario Ministry of Natural Resources Science & Technology Series, Volume 7. 227 p.

Angers, V. A., Messier, C., Beaudet, M., & Leduc, A. (2005). Comparing composition and structure in old-growth and harvested (selection and diameter-limit cuts) northern hardwood stands in Quebec. Forest Ecology and Management, 217(2–3), 275–293.

Attiwill, P. M. (1994). The disturbance of forest ecosystems: the ecological basis for conservative management. Forest Ecology and Management, 63(2–3), 247–300.

Barlow, C. A. (1970). Phenology and distribution of some Pterostichus (Coleoptera: Carabidae) of Eastern Canada. Journal of the New York Entomological Society, 78, 215-236.

Bergeron, J. A. C., Blanchet, F. G., Spence, J. R., & Volney, W. J. A. (2012). Ecosystem classification and inventory maps as surrogates for ground beetle assemblages in boreal forest. Journal of Plant Ecology, 5(1), 97–108.

Bergeron, J. A. C., Spence, J. R., & Volney, W. J. A. (2011). Landscape patterns of species-level association between ground-beetles and overstory trees in boreal forests of western Canada (Coleoptera, Carabidae). ZooKeys, 147, 577–600. den Boer, P. J. (1981). On the survival of populations in a heterogeneous and variable environment. Oecologia, 50, 39-53.

Bousquet, Y., (2010). Illustrated identification guide to adults and larvae of northeastern North American ground beetles (Coleoptera: Carabidae). Pensoft Series Faunistica No. 90, Pensoft Publishers, Sofia, Bulgaria.

Brais, S., & Camiré, C. (1998). Soil compaction induced by careful logging in the claybelt region of northwestern Quebec (Canada). Canadian Journal of Soil Science, 78(1), 197–206.

Brandmayr, P., & Pizzolotto, R. (2016). Climate change and its impact on epigean and hypogean carabid beetles. Periodicum Biologorum, 118(3), 147–162.

Cecil-Cockwell, M.J.L. & Caspersen, J.P. (2015) A simple system for classifying sugar maple vigour and quality. Canadian Journal of Forest Research, 45, 900–909.

Crow, T. R., Buckley, D. S., Nauertz, E. A, & Zasada, J. C. (2002). Effects of management on the composition and structure of northern hardwood forests in Upper Michigan. Forest

Science, 48(1), 129–145.

Doyon, F., Gagnon, D., & Giroux, J. F. (2005). Effects of strip and single-tree selection cutting on birds and their habitat in a southwestern Quebec northern hardwood forest. Forest Ecology and Management, 209(1–2), 101–116.

Environment Canada. (2018). Climate normals for Haliburton, Ontario. Available from http://www.climate.weatheroffice.ec.gc.ca/climate_normals/index_e.html [accessed 15 November 2018].

Epstein, M. E., & Kulman, H. M. (1990). Habitat Distribution and Seasonal Occurrence of Carabid Beetles in East-central Minnesota. The American Midland Naturalist, 123(2), 209– 225.

Freedman, B., Zelazny, V., Beaudette, D., Fleming, T., Johnson, G., Flemming, S., Forbes, G., Gerrrow, J.S., Johnson, G., & Woodley, S. (1996). Biodiversity implications of changes in the quantity of dead organic matter in managed forests. Environmental Reviews, 4(3), 238– 265.

Goodburn, J. M., & Lorimer, C. G. (1998). Cavity trees and coarse woody debris in old-growth and managed northern hardwood forests in Wisconsin and Michigan. Canadian Journal of Forest Research, 28(3), 427–438.

Goodburn, J. M., & Lorimer, C. G. (1999). Population structure in old-growth and managed northern hardwoods: an. Forest-Ecology-and-Management, 118(1/3): 11-29, 118.

Hale, C. M., Pastor, J., & Rusterholz, K. A. (1999). Comparison of structural and compositional characteristics in old-growth and mature, managed hardwood forests of Minnesota, U.S.A. Canadian Journal of Forest Research, 29(10), 1479–1489.

Halpin, C. R., Lorimer, C. G., Hanson, J. J., & Palik, B. J. (2017). Predicted long-term effects of group selection on species composition and stand structure in northern hardwood forests. Forest Ecology and Management, 400, 677–691.

Harmon, M.E., Franklin, J.F., Swanson, F.J., Sollins, P., Gregory, S.V., Lattin, J.D., Anderson, N.H., Cline, S.P., Aumen, N.G., Sedell, J.R., Lienkaemper, G.W., Cromack, K., Jr., and Cummins, K.W. (1986). Ecology of coarse woody debris in temperate ecosystems. Advances in Ecological Research, 15, 133-302.

Hautala, H., Jalonen, J., Laaka-Lindberg, S., & Vanha-Majamaa, I. (2004). Impacts of retention felling on coarse woody debris (CWD) in mature boreal spruce forests in Finland. Biodiversity and Conservation, 13(8), 1541–1554.

Klimaszewski, J., Langor, D. W., Work, T. T., Pelletier, G., Hammond, H. J., & Germain, C. (2005). The effects of patch harvesting and site preparation on ground beetles (Coleoptera, Carabidae) in yellow birch dominated forests of southeastern Quebec. Canadian Journal of

Forest Research, 35(11), 2616–2628.

Koivula, M. (2002). Alternative harvesting methods and boreal carabid beetles (Coleoptera, Carabidae). Forest Ecology and Management, 167(1–3), 103–121.

Koivula, M. J. (2011). Useful model organisms, indicators, or both? Ground beetles (Coleoptera, Carabidae) reflecting environmental conditions. ZooKeys, 100, 287–317.

Koivula, M., Kukkonen, J., & Niemelä, J. (2002). Boreal carabid-beetle (Coleoptera, Carabidae) assemblages along the clear-cut originated succession gradient. Biodiversity and Conservation, 11(7), 1269–1288.

Kotze, D. J., Lehvävirta, S., Koivula, M., O’Hara, R. B., & Spence, J. R. (2012). Effects of habitat edges and trampling on the distribution of ground beetles (Coleoptera, Carabidae) in urban forests. Journal of Insect Conservation, 16(6), 883–897.

Kotze, D., Brandmayr, P., Casale, A., Dauffy-Richard, E., Dekoninck, W., Koivula, M. J., Löver, G.L., Mossakowski, D., Noordijk, J., Paarmann, W., Pizzolotto, R., Saska, P., Schwerk, A., Serrano, J., Szyszko, J., Taboada, A., Turin, H., Venn, S., Vermeulen, R., Zetto, T. (2011). Forty years of carabid beetle research in Europe - from taxonomy, biology, ecology and population studies to bioindication, habitat assessment and conservation. ZooKeys, 100, 55– 148.

Langor, D. W., & Spence, J. R. (2006). Arthropods as ecological indicators of sustainability of Canadian Forest. The Forestry Chronicle, 82(3), 344–350.

Larochelle, A. and M.-C. Larivière. (2003). A natural history of the ground-beetles (Coleoptera: Carabidae) of America north of Mexico. Pensoft Series Faunistica (Sofia-Moscow) 27, 583 pp.

Latty, E. F., Werner, S. M., Mladenoff, D. J., Raffa, K. F., & Sickley, T. A. (2006). Response of ground beetle (Carabidae) assemblages to logging history in northern hardwood-hemlock forests. Forest Ecology and Management, 222(1–3), 335–347.

Leadbitter, P., Euler, D., & Naylor, B. (2002). A comparison of historical and current forest cover in selected area of the Great Lakes St. Lawrence Forest of central Ontario. Forestry Chronicle, 78(4), 522–529.

Levesque, C., & Levesque, G.-Y. (1994). Abundance and Seasonal Activity of Ground Beetles (Coleoptera : Carabidae) in a Raspberry Plantation and Adjacent Sites in Southern Québec (Canada). Journal of the Kansas Entomological Society, 67(1), 73–101.

Lindenmayer, D. B., Margules, C. R., & Botkin, D. B. (2000). Indicators of biodiversity for ecologically sustainable forest management. Conservation Biology, 14(4), 941–950.

Malcolm, J. R., Valenta, K., & Lehman S. M. (2017). Edge effects in tropical dry forests of

Madagascar: additivity or synergy? Landscape Ecology, 32, 327-341.

Martel, J., Mauffette, Y., & Tousignant, S. (1991). Secondary effects of canopy dieback: The epigeal carabid fauna in Québec Appalachian maple forests. Canadian Entomologist, 123, 851-859.

Maser, C., R. G. Anderson, K. Cromack Jr., J. T. Williams, and R. E. Martin. (1979). Dead and down woody material: In wildlife habitats in managed forests: The Blue Mountains of Oregon and Washington, ed. J. Thomas, 78-95. Agriculture Handbook #553. Washington, DC: USDA Forest Service.78-95.

Miller, G. W., & Smith, H. C. (1993). A Practical Alternative to Single-Tree Selection? Northern Journal of Applied Forestry, 10(1), 32–38.

Moore, J.-D., Ouimet, R., Houle, D., & Camiré, C. (2004). Effects of two silvicultural practices on ground beetles (Coleoptera: Carabidae) in a northern hardwood forest, Quebec, Canada. Canadian Journal of Forest Research, 34(4), 959–968.

Neuendorff, J. K., Nagel, L. M., Webster, C. R., & Janowiak, M. K. (2007). Stand structure and composition in a northern hardwood forest after 40 years of single-tree selection. Northern Journal of Applied Forestry, 24(3), 197–202.

Niemelä, J., Koivula, M., & Kotze, D. J. (2007). The effects of forestry on carabid beetles (Coleoptera: Carabidae) in boreal forests. Journal of Insect Conservation, 11(1), 5–18.

Ontario Ministry of Natural Resources. (2015). Forest Management Guide to Silviculture in the Great Lakes-St. Lawrence and Boreal Forests of Ontario. Queen’s Printer for Ontario, Toronto.

Ontario Ministry of Natural Resources. (2012). A Silvicultural Guide to Managing Southern Ontario Forests. Ontario Ministry of Natural Resources. Queen’s Printer for Ontario, Toronto.

Ontario Ministry of Natural Resources. (1998). A Silvicultural Guide for the Tolerant Hardwood Forest in Ontario. Ontario Ministry of Natural Resources. Queen’s Printer for Ontario, Toronto.

Ontario Ministry of Natural Resources. (2004). Ontario Tree Marking Guide. Ontario Ministry of Natural Resources. Queen’s Printer for Ontario, Toronto.

Pearce, J. L., & Venier, L. A. (2006). The use of ground beetles (Coleoptera: Carabidae) and spiders (Araneae) as bioindicators of sustainable forest management: A review. Ecological Indicators, 6(4), 780–793.

Pearce, J. L., Venier, L. A., McKee, J., Pedlar, J., & McKenney, D. (2003). Influence of habitat and microhabitat on carabid (Coleoptera: Carabidae) assemblages in four stand types. The

Canadian Entomologist, 135(03), 337–357.

Piascik, P. W. (2013). Responses of ground beetles ( Coleoptera : Carabidae ) to variation in woody debris supply in boreal northeastern Ontario. M.Sc. thesis. University of Toronto.

Pinto, F., Romaniuk, S., & Ferguson, M. (2008). Changes to preindustrial forest tree composition in central and northeastern Ontario, Canada. Canadian Journal of Forest Research, 38(7), 1842–1854.

Quinn, N. W. S. (2004). The presettlement hardwood forests and wildlife of Algonquin Provincial Park: A synthesis of historic evidence and recent research. Forestry Chronicle, 80(6), 705–717.

Rainio, J., & Niemela, J. (2003). Ground beetles ( Coleoptera : Carabidae ) as bioindicators. Biodiversity and Conservation, 12: 487–506.

Rowe, J.S., (1972). Forest Regions of Canada. Canadian Forest Service, Ottawa, Canada.

Rykken, J. J., Capen, D. E., Mahabir, S. P., Rykken, J., Capen, D. E., & Mahabirt, S. P. (2018). Ground Beetles as Indicators of Land Type Diversity in the Green Mountains of Vermont. Conservation Biology, 11(2), 522–530.

Saint-Germain, M., & Mauffette, Y. (2001). Reduced ground beetle activity following ice damage in maple stands of southwestern Quebec. Forestry Chronicle, 77(4), 651–656.

Schulte, L. A., Mladenoff, D. J., Crow, T. R., Merrick, L. C., & Cleland, D. T. (2007). Homogenization of northern U.S. Great Lakes forests due to land use. Landscape Ecology, 22(7), 1089–1103.

Schwartz, J. W., Nagel, L. M., & Webster, C. R. (2005). Effects of uneven-aged management on diameter distribution and species composition of northern hardwoods in Upper Michigan. Forest Ecology and Management, 211(3), 356–370.

Shields, J. M., Webster, C. R., & Nagel, L. M. (2007). Factors influencing tree species diversity and Betula alleghaniensis establishment in silvicultural openings. Forestry, 80(3), 293–307.

Siddig, A. A. H., Ellison, A. M., Ochs, A., Villar-Leeman, C., & Lau, M. K. (2017). How do ecologists select and use indicator species to monitor ecological change? Insights from 14 years of publication in Ecological Indicators. Ecological Indicators, 60, 223–230.

Siitonen, J. (2001). Forest Management, Coarse Woody Debris and Saproxylic Organisms : Fennoscandian Boreal Forests as an Example. Ecological Bulletins, 49, 11–41.

Spence, J. R., Langor, D. W., Niemelä, J. K., Carcamo, H. A., & Currie, C. R. (1996). Northern forestry and carabids: the case for concern about old-growth species. Annales Zoologici Fennici, 33, 173–184.

Thiele, H. U. (1977). Carabid beetles in their environments. A study on habitat selection and adaptations in physiology and behaviour. Springer-Verlag, Berlin, Germany.

Tierney, G. L., Faber-Langendon, D., Mitchell, B. R., Shriver, W. G., & Gibbs, P. (2009). Monitoring and evaluating the ecological integrity of forest ecosystems. Frontiers in Ecology and the Environment, 7(6), 308-316.

Tromp-van Meerveld, H. J., & McDonnell, J. J. (2006). On the interrelations between topography, soil depth, soil moisture, transpiration rates and species distribution at the hillslope scale. Advances in Water Resources, 29(2), 293–310.

Vance, C. C., & Nol, E. (2003). Temporal effects of selection logging on ground beetle communities in northern hardwood forests of eastern Canada. Ecoscience, 10(1), 49–56.

Vanderwel, M. C., Malcolm, J. R., & Smith, S. M. (2006). An integrated model for snag and downed woody debris decay class transitions. Forest Ecology and Management, 234(1–3), 48–59.

Vanderwel, M. C., Mills, S. C., & Malcolm, J. R. (2009). Effects of partial harvesting on vertebrate species associated with late-successional forests in Ontario’s boreal region. Forestry Chronicle, 85(1), 91–104.

Vanderwel, M. C., Thorpe, H. C., Shuter, J. L., Caspersen, J. P., & Thomas, S. C. (2008). Contrasting downed woody debris dynamics in managed and unmanaged northern hardwood stands. Canadian Journal of Forest Research, 38(11), 2850–2861.

Vanderwel, M.C., Caspersen, J.P., Woods, M.E. (2006) Snag dynamics in partially harvested and unmanaged northern hardwood forests. Canadian Journal of Forest Research, 36(11), 2769–79.

Vuidot, A., Paillet, Y., Archaux, F., & Gosselin, F. (2011). Influence of tree characteristics and forest management on tree microhabitats. Biological Conservation, 144(1), 441–450.

Van Wagner, C. E. (1968). The line intercept method in forest fuel sampling. Forest Science 14:20-26.

Werner, S. M., & Raffa, K. F. (2000). Effects of forest management practices on the diversity of ground-occurring beetles in mixed northern hardwood forests of the Great Lakes Region. Forest Ecology and Management, 139, 135–155.

White, M., & Mladenoff, D. J. (1994). Old-growth forest landscape transition from pre-European settlement to present. Landscape Ecology, 9(3), 191–205.

Work, T. T., Koivula, M., Klimaszewski, J., Langor, D. W., Spence, J. R., Sweeney, J., & Hebert, C. (2008). Evaluation of carabid beetles as indicators of forest change in Canada.

The Canadian Entomologist, 140, 393–414.

Work, T. T., Shorthouse, D. P., Spence, J. R., Volney, W. J. A., & Langor, D. (2004). Stand composition and structure of the boreal mixedwood and epigaeic arthropods of the Ecosystem Management Emulating Natural Disturbance (EMEND) landbase in northwestern Alberta. Canadian Journal of Forest Research, 34(2), 417–430.

Tables and Figures

1 Mean (±SE) Carabid species CON STS FMS DLC ICR F or χ2 value P value July 2016 (n = 5) (n = 5) (n = 6) (n = 6) (n = 4) Synuchus impunctatus2 46.47(8.98) 36.08(9.82) 24.83(7.69) 27.06(7.11) 11.94(5.26) 2.14 0.11 Pterostichus coracinus2 17.83(5.92) 20.69(8.96) 7(3.54) 6.9(4.46) 10.35(6.01) 1.08 0.39 Sphaeroderus canadensis3 10.19(5.36) 3.18(1.16) 1.91(1.27) 1.06(1.06) 0.8(0.8) 6.2 0.18 Cymindis cribricollis2 3.82(1.19) 2.65(0.98) 3.18(2.01) 3.71(2.52) 1.59(1.59) 0.22 0.92 Pterostichus pensylvanicus3 1.27(0.78) 2.65(1.28) 0.64(0.64) 0.53(0.53) 0.8(0.8) 2.08 0.72 Pterostichus tristis3 2.55(1.19) 0.53(0.53) 1.27(0.78) 2.12(1.06) 0 6.61 0.16 Agonum retractum3 1.27(0.78) 1.59(1.09) 2.55(1.19) 1.06(1.06) 0 4.16 0.38 Total abundance2 84.03(11.15) 68.44(19.18) 43.29(8.21) 42.97(7.11) 30.24(8.27) 2.84 0.05 June 2017 (n = 6) (n = 6) (n = 6) (n = 5) (n = 4) Pterostichus coracinus2 11.67(2.12) 12.39(3.6) 8.49(5.11) 3.82(1.86) 4.17(1.99) 1.3 0.3 Agonum retractum2 4.77(1.36) 6.99(2.96) 8.49(3.15) 7.79(1.52) 2.39(1.52) 0.94 0.46 Pterostichus pensylvanicus2 4.24(2.42) 4.09(1.12) 6.9(3.01) 7(3.82) 1.78(1.04) 0.64 0.64 Platynus decentis3 1.59(0.71) 1.84(1.31) 0.53(0.53) 0.64(0.64) 3.37(2.26) 2.11 0.72 Pterostichus mutus3 1.06(0.67) 1.19(0.76) 2.12(1.57) 1.27(0.78) 3.98(3.98) 1.04 0.9 Total abundance2 28.65(3.18) 27.56(7.57) 32.36(8.68) 25.13(5.4) 22.19(5.62) 2.04 0.12 July 2017 (n = 6) (n = 6) (n = 6) (n = 6) (n = 4) Pterostichus coracinus2 10.86(3.69) 13.76(4.32) 5.31(1.78) 8.49(4) 2.39(1.52) 1.5 0.23 Synuchus impunctatus2 8.33(3) 4.24(2.56) 6.37(2.32) 0.53(0.53) 0.8(0.8) 2.3 0.09 Agonum retractum3 3.56(1.77) 3.71(3.71) 2.12(1.57) 3.18(1.64) 0.8(0.8) 1.1 .89 Total abundance2 25(5.56) 23.31(7.2) 15.38(2.9) 15.92(3.77) 5.57(2.39) 0.3 0.87 1 CON = Control; STS = Single-tree selection; FMS = Financial maturity selection; DLC = Diameter-limit cutting; ICR = Intensive crown release. See text for details. 2 F and P values are from one-way analysis of variance. 3 χ2 and P values are from negative binomial regression.

Means(±SE)1 Variable CON STS FMS DLC ICR F value2 P value2 (n = 6) (n = 6) (n = 6) (n = 6) (n = 4) Trees (DBH ≥ 10 cm) Treatment block basal area (m²/ha) All trees Pre-harvest 22.49(1.21) 23.5(1.18) 22.08(0.47) 22.75(0.53) 20.5(1.02) 1.18 0.35 Harvested 0a 5.75(0.7)b 7.53(0.57)b 7.17(0.4)b 9(2.41)b 15.75 <0.01 Post-harvest 22.49(1.21)a 17.75(0.51)b 14.56(0.39)bc 15.58(0.4)bc 11.5(2.46)c 15.11 <0.01 Small tree (DBH < 38cm) Pre-harvest 12.04(0.98) 14.42(1.36) 13.5(0.72) 14.75(0.69) 12.88(0.72) 1.37 0.28 Harvested 0a 3.42(0.85)b 2.89(0.41)ab 1.83(0.36)ab 4.25(2.14)b 3.89 0.01 Post-harvest 12.04(0.98) 11.0(0.56) 10.61(0.48) 12.92(0.69) 8.63(2.19) 2.49 0.07 Large tree (DBH ≥ 38cm) Pre-harvest 10.45(1.57) 9.08(0.83) 8.58(0.96) 8(0.84) 7.63(1.16) 0.95 0.45 Harvested 0a 2.33(0.44)b 4.64(0.77)c 5.33(0.49)c 4.75(0.92)bc 16.12 <0.01 Post-harvest 10.45(1.57)a 6.75(0.44)b 3.94(0.31)bc 2.67(0.57)c 2.88(0.38)c 15.05 <0.01 Local basal area (m²/ha) All trees 24(3.31) 22(2.13) 17.33(1.23) 20(2.42) 16.5(3.86) 1.4 0.27 Small tree (DBH < 38cm) 13.67(1.96) 14.33(0.8) 12.33(1.67) 16.67(2.29) 12.5(2.06) 0.94 0.46 Large tree (DBH ≥ 38cm) 10.33(2.5)a 7.67(1.82)ab 5(1.13)ab 3.33(0.67)b 4(1.83)ab 2.93 0.04 DWD (width ≥ 10 cm) Total abundance (pieces/100m) 5.26(1.51)a 10.22(1.5)ab 8.63(0.82)ab 10.02(2.18)ab 12.65(0.45)b 2.92 0.04 Early-decay, small (width < 13.8 cm) 0.89(0.37)a 4.17(0.84)ab 4.27(0.85)ab 4.27(1.71)ab 6.4(0.66)b 3.32 0.03 Early-decay, large (width ≥ 13.8 cm) 0.79(0.37)a 3.77(0.81)b 2.58(0.65)ab 3.47(0.74)ab 4.61(0.66)b 4.51 0.01 Late-decay, small (width < 13.8 cm)3 0.99(0.66) 0.79(0.29) 0.5(0.32) 0.6(0.27) 0.74(0.28) 0.43 0.79 Late-decay, large (width ≥ 13.8 cm)3 2.58(1.09) 1.49(0.45) 1.29(0.85) 1.69(0.39) 0.89(0.38) 0.89 0.49 Snags (DBH ≥ 10 cm) Snag basal area (m²/ha) 1.02(0.32) 0.41(0.08) 0.48(0.09) 0.52(0.17) 0.5(0.31) 1.53 0.23 1 CON = Control; STS = Single-tree selection; FMS = Financial maturity selection; DLC = Diameter-limit cutting; ICR = Intensive crown release. See text for details. 2 F and P values are from one-way analysis of variance. 3 Variables were log transformed prior to analysis.

Forest structure variables1 Carabid species T_B T_H T_L D_E D_L F1 F2 Agonum retractum July 2016 3,4 0.47 0.47 0.79 0.04(-) 0.87 0.3 0.63 June 20172 0.99 0.98 0.63 0.19 0.88 1 0.99 July 20173 0.41 0.74 0.47 0.67 0.76 0.77 0.73 Cymindis cribricollis July 20163 0.86 0.74 0.87 0.44 0.7 0.72 0.46 Platynus decentis June 20173,4 0.24 0.17 0.07 0.22 0.21 0.23 0.17 Pterostichus coracinus July 20162 0.3 0.03(-) 0.24 0.16 0.79 0.07 0.97 June 20172 0.12 0.04(-) 0.18 0.24 0.41 0.03(-) 0.64 July 20172 0.18 0.14 0.99 0.13 0.69 0.2 0.74 Pterostichus mutus June 20173,4 0.48 0.32 0.92 0.94 0.9 0.67 0.77 Pterostichus pensylvanicus July 20163,4 0.43 0.61 0.36 0.48 0.46 0.66 0.74 June 20172 0.67 0.77 <0.01(-) 0.22 <0.01(+) 0.44 <0.01(+) Pterostichus tristis July 20163 0.44 0.3 0.74 0.3 0.62 0.27 0.67 Sphaeroderus canadensis July 20163 0.03(+) 0.02(-) 0.01(+) 0.13 0.83 <0.01(-) 0.98 Synuchus impunctatus July 20162 <0.01(+) <0.01(-) 0.42 <0.01(-) 0.64 <0.01(-) 0.42 July 20173 0.1 0.43 0.48 0.11 0.44 0.26 0.79 Total abundance July 20163 <0.01(+) <0.01(-) 0.08 <0.01(-) 0.88 <0.01(-) 0.48 June 20173 0.92 0.38 0.33 0.16 0.27 0.68 0.16 July 20172 0.04(+) 0.08 0.84 0.04(-) 0.32 0.12 0.4 1T_B = Block basal area; T_H= Block basal area harvested; T_L = Local basal area; D_E = Early-decay DWD abundance; D_L = Late-decay DWD abundance; F1 = First varimax factor, F2 = Second varimax factor. See text for details. 2P values are from linear regression. 3P values are from negative binomial regression. 4Variables were log transformed prior to analysis.

Mean (±SE)1 Carabid species Mature logged Old-growth F value1 P value1 June 2017 (n = 6) (n = 3) Agonum retractum 17.41(4.48) 12.81(3.04) 0.45 0.52 Pterostichus coracinus 5(1.2) 2.51(0.73) 1.89 0.21 Pterostichus pensylvanicus 4.18(1.25) 2.86(0.94) 0.46 0.52 Sphaeroderus canadensis 2.41(0.71) 3.21(1.05) 0.41 0.54 Pterostichus tristis 1.59(0.43) 1.6(0.3) <0.01 0.98 Total abundance 35.17(6.74) 25.12(3.94) 0.97 0.36 July 2017 (n = 6) (n = 3) Synuchus impunctatus 6.45(2.1) 8.19(4.71) 0.16 0.7 Pterostichus coracinus 4.07(1.04) 4.05(2.08) <0.01 0.99 Agonum retractum 2.83(0.75) 4.26(1.93) 0.73 0.42 Pterostichus tristis 1.33(0.38) 1.1(0.55) 0.11 0.75 Total abundance 16.09(2.87) 19.74(6.28) 0.38 0.56

1F and P values are from one-way analysis of variance.

Table 2.2. Comparison of forest structure and tree species composition variables between old-growth and mature logged stands in Haliburton Forest, Ontario.

Mean (±SE)1 Variable Mature logged Old-growth F value1 P value1 (n = 6) (n = 3) Trees (DBH ≥ 10 cm) Basal area (m²/ha) 23.58(0.55) 25.88(0.47) 7.15 0.03 Basal area semivariance2 -0.44(0.61) 0.88(0.21) 2.18 0.18 Weibull parameters Scale 26.88(0.49) 28.21(0.7) 2.43 0.16 Shape 2.07(0.03) 1.95(0.1) 2.3 0.17 Species composition %MH 0.73(0.07) 0.59(0.14) 0.97 0.36 %MH semivariance2 -0.02(0.02) 0.04(0.02) 2.66 0.15 Small BY (DBH < 38cm) basal area 0.76(0.17) 0.18(0.1) 3.94 0.09 Conifer cover score 12.83(2.49) 37.22(14.14) 6.01 0.04 DWD (width ≥ 10 cm) Total volume (m3/ha) 43.87(4.77) 86.6(8.48) 22.92 <0.01 Early-decay, small (width < 18.7 cm) 2.79(0.23) 4.34(1.68) 1.81 0.22 Early-decay, large (width ≥ 18.7 cm) 8.56(2.59) 30.52(1.39) 31.74 <0.01 Late-decay, small (width < 18.7 cm) 5.52(0.28) 6.97(0.94) 3.85 0.09 Late-decay, large (width ≥ 18.7 cm) 27(4.72) 44.77(9.09) 3.8 0.09 Snags (DBH ≥ 10 cm) Total basal area (m²/ha) 1.11(0.14) 4.18(0.39) 89.87 <0.01 Early-decay, small (DBH < 33.3 cm) 0.06(0.02) 0.14(0.08) 1.59 0.25 Early-decay, large (DBH ≥ 33.3 cm) 0.12(0.03) 0.23(0.02) 5 0.06 Late-decay, small (DBH < 33.3 cm) 0.4(0.08) 1.01(0.51) 3.05 0.12 Late-decay, large (DBH ≥ 33.3 cm) 0.54(0.11) 2.8(0.21) 110.06 <0.01 Leaf litter depth 1.83(0.16) 2(0.29) 0.32 0.59 1F and P values are from one-way analysis of variance (df = 1, 8). 2Semivariance values are residuals calculated upon partialling out means.

%Variation P value Habitat variable1 explained June 2017 Transect RDA axes 96 0.11 BA_SEMI 34 0.07 %MH 25 0.14 conifer 18 0.22 Station RDA axes 19 0.3 conifer 9 0.02 D_LL 5.3 0.07 %MH 5.2 0.07 July 2017 Transect RDA axes 95 0.26 D_ES 28 0.07 conifer 18.6 0.23 %MH 16.1 0.26 Station RDA axes 30 <0.01 D_ES 11 <0.01 MH% 7 0.02 conifer 7 0.02 1Refer to Figure 1.3 for habitat variable acronyms.

Table 2.4. Statistical significance of relationships between carabid abundances (captures per 100 bucket-nights) and forest structure and species composition variables for carabid species occurring in at least 30% of transects in an individual month from sampling in Haliburton Forest, Ontario, in 2017. For significant relationships, the direction of the slope is indicated in parentheses.

Habitat variables1 D_ES D_EL D_LS D_LL LD BA BA_SEMI shape scale %MH %MH_SEMI conifer Species Transect Station Transect Station Transect Station Transect Station Transect Station Transect Station Transect Transect Transect Transect Station Transect Transect Station Agonum retractum June 20172 0.62 0.54 0.48 0.94 0.72 0.35 0.37 0.07 0.99 0.91 1 0.85 0.09 0.44 0.48 0.14 0.07 0.9 0.2 0.01(-) July 20173 0.89 0.41 0.96 0.54 0.51 0.55 0.03(+) 0.05(+) 0.77 0.78 0.44 0.75 0.96 0.06 0.19 0.18 0.16 0.93 0.07 0.16 Pterostichus coracinus June 20173 0.3 0.06 0.04(-) 0.23 0.32 0.72 0.71 0.84 0.04(-) 0.06 0.08 0.99 0.93 0.75 0.55 0.35 0.31 0.04(-) 0.6 0.35 July 20173 0.34 0.48 0.86 0.49 0.56 0.74 0.23 0.56 0.03(-) 0.13 0.9 0.88 0.84 0.36 0.8 0.05 0.02(-) 0.48 0.65 0.9 Pterostichus pensylvanicus June 20172 0.57 0.47 0.38 0.57 0.67 0.72 0.73 0.52 0.88 0.81 0.79 0.51 0.01(-) 0.37 0.64 0.32 0.53 0.39 0.92 0.54 Pterostichus tristis June 20173 0.85 0.86 0.81 0.63 0.32 0.63 0.72 0.67 0.03(-) 0.3 0.59 0.8 0.81 0.25 0.2 0.61 0.71 0.08 0.73 0.46 July 20173,4 0.23 0.36 0.47 0.89 0.46 0.85 0.65 0.35 0.42 0.26 0.41 0.73 0.1 0.25 0.68 0.11 0.78 1 0.8 0.45 Sphaeroderus canadensis June 20173 0.53 0.62 0.63 0.45 0.22 0.04(+) 0.89 0.6 0.66 0.68 0.01(+) 0.78 0.39 0.97 0.86 0.33 0.78 0.05 0.85 0.66 Synuchus impunctatus July 20172 0.08(-) <0.01(-) 0.88 0.86 0.27 0.43 0.45 0.49 0.39 0.3 0.54 0.27 0.71 0.81 0.98 0.43 0.25 0.49 0.23 0.03(-) Total abundance June 20173 0.59 0.47 0.21 0.57 0.72 0.74 0.55 0.42 0.76 0.27 0.94 0.86 0.02(-) 0.26 0.67 0.14 0.03(-) 0.55 0.21 <0.01(-) July 20173 0.09 <0.01(-) 0.84 0.42 0.61 0.98 0.6 0.47 0.14 0.1 0.46 0.55 0.92 0.51 0.72 0.68 0.25 0.68 0.84 0.56 1Refer to Figure 1.3 for habitat variable acronyms 2 P values are from linear regression. 3 P values are from negative binomial regression. 4Variables were log transformed prior to analysis.

Figure 1.1. The Blue Heron Demonstration forest experimental design. Each treatment block is an 84 m square. The red line represents the path of the primary skid trail and M = ICR.

DLC FMS Cymicrib Ptertris SyncimpuSynuimpu Chlaaest Platdece Loripili CON Ptermela Agonretr Harpsolm Olisparm Ptermutu Pterpens Elapclai Sphacana ICRM Total abundance Agonfide STS

July 2016

Ptercora 100.4 -1.0 -1.0 1.5

Pterpens

FMS DLC Ptertris Gasthome Total abundance Pterrost Calofrig Agonretr Cymicrib Olisparm Sphacana Agontide Pterluct Ptercora Ptermutu Agoncorv SynuimpuSyncimpu CON STS Platdece

ICRM June 2017 101.0 -1.0 -1.0 1.5

SynuimpuSyncimpu

CON

Total abundance DLC

Agonretr

Sphacana

Agonfide Ptertris Ptercora Pterpens STS Cymicrib Platdece

ICRM FMS July 2017 101.5 -1.0 -1.0 1.5 Figure 1.2. First two axes from a PCA ordination on the covariance matrix of standardized carabid abundances from sampling in July 2016, June 2017 and July 2017. Carabid species acronyms consist of the first four letters of the genus and the first four letters of the species. Total abundance and treatment centroids (CON, STS, DLC, FMS and ICR) were passive variables (red).

50 CON50 STS50 FMS50 DLC50 ICR 40 40 40 40 40 July 2016 30 30 30 30 30

20 20 20 20 20

10 10 10 10 10

0 0 0 0 0 Si Pc Sc Cc Ar Pt Pp Pe Pd Pu Af Po Lp Hs El Ch Si Pc Sc Cc Ar Pt Pp Pe Pd Pu Af Po Lp Hs El Ch Si Pc Sc Cc Ar Pt Pp Pe Pd Pu Af Po Lp Hs El Ch Si Pc Sc Cc Ar Pt Pp Pe Pd Pu Af Po Lp Hs El Ch Si Pc Sc Cc Ar Pt Pp Pe Pd Pu Af Po Lp Hs El Ch 16 16 16 16 16 14 CON14 STS14 FMS14 DLC14 ICR 12 12 12 12 12 10 June 2017 10 10 10 10 8 8 8 8 8 6 6 6 6 6 4 4 4 4 4 2 2 2 2 2 0 0 0 0 0 Pc Ar Pp Pu Pd Sc Pt Cc OpGh Af Pl Pr Si Cf Ac Pc Ar Pp Pu Pd Sc Pt Cc OpGh Af Pl Pr Si Cf Ac Pc Ar Pp Pu Pd Sc Pt Cc OpGhAf Pl Pr Si Cf Ac Pc Ar Pp Pu Pd Sc Pt Cc OpGh Af Pl Pr Si Cf Ac Pc Ar Pp Pu Pd Sc Pt Cc OpGh Af Pl Pr Si Cf Ac

CarabidAbundance 15 CON15 STS15 FMS15 DLC15 ICR

July 2017 10 10 10 10 10

5 5 5 5 5

0 0 0 0 0 Pc Si Ar Cc Sc Pt Af Pp Pd Pc Si Ar Cc Sc Pt Af Pp Pd Pc Si Ar Cc Sc Pt Af Pp Pd Pc Si Ar Cc Sc Pt Af Pp Pd Pc Si Ar Cc Sc Pt Af Pp Pd

Species

Figure 1.3. Standardized carabid species abundances from the rarest species for each month on the right to the most common on the left for July 2016, June 2017 and July 2017. Carabid species acronyms consist of the first letter of the genus and the first letter of the species.

16 AugustJuly 20162016 14 12 10 8 6 4 2 0 0 20 40 60 80 100 120 140

16 June 2017 14 12 10 8 6 4 2

Species RichnessSpecies 0 0 10 20 30 40 50 60 70

12 July 2017 10

0 0 10 20 30 40 50

Carabid Abundance

4 CON1 D_LL D_LS S_T 2

T_BL T_HS 0

T_LL T_HL T_BS D_EL -2 D_ES T_LS

-3 -2 -1 0 1 2 3

July 2016

Synuimpu T_H

ICR

June 2017 July 2017

ICR T_H

Synuimpu

ICR T_H

Figure 1.6. The first two axes of redundancy analysis on the covariance matrix constraining carabid abundances in July 2016, June 2017 and July 2017 with structure variables (red): T_B = Treatment block basal area; T_H = Treatment block basal area harvested; T_L = local tree basal area; D_E = Early-decay DWD abundance; D_E = Early-decay DWD abundance; D_L = Late- decay DWD abundance; F1 = First varimax factor, F2 = Second varimax factor. Treatment centroids and total abundance were added passively (green).

June 2017

Calaadve

July 2017

Synuimpu

Figure 2.2. First two axes from a Principal Component Analysis on the covariance matrix of standardized carabid abundances from sampling in June 2017 and July 2017 in Haliburton Forest, Ontario. Carabid species acronyms consist of the first four letters of the genus and the first four letters of the species. Total abundance was added as a passive variable. Colours correspond to treatments: old-growth (green); mature logged (red).

20 ML OG

15 June 2017 10

0 Ar Pc PpSc Pd Pt OpCc Sl AfPmPaAaGhAp Ar Pc PpSc Pd Pt OpCc Sl AfPmPaAaGhAp

14 12 ML OG 10 8 July 2017 6 4 2 0 Si Pc Ar Pt Sc Cc PpPm Hr HiC Aaa Si Pc Ar Pt Sc Cc PpPm Hr HiC Aaa

Appendices

Appendix 1. Locations of Blue Heron Demonstration Forest (BH) and transects within old- growth (OG) and mature logged (ML) stands in Haliburton Forest and Wild Life Reserve (HF), adjacent to the south-west side of Algonquin Provincial Park (APP).

Appendix 2. Species list in descending rank of abundance, indicating total number of individuals captured in July 2016, June 2017 and July 2017 in the five silvicultural treatments in the Blue Heron Demonstration forest in Haliburton Forest, as well as habitat categories. Effort is listed in parentheses (bucket-nights).

July 2016 June 2017 July 2017 CON STS FMS DLC ICR CON STS FMS DLC ICR CON STS FMS DLC ICR Total Habitat Species (n=147.08) (n=188.4) (n=188.4) (n=147.1) (n=124.7) (n=188.4) (n=176.4) (n=141.1) (n=188.4) (n=119.7) (n=182.4) (n=182.4) (n=188.4) (n=188.4) (n=124.7) (n=2476) association Synuchus impunctatus (Say, 1823) 73 68 41 22 14 0 0 0 1 0 14 8 1 12 1 267 generalist Pterostichus coracinus (Newman, 1838) 28 39 13 8 13 22 21 6 16 4 20 24 16 10 3 244 generalist Agonum retractum LeConte, 1848 2 3 2 2 0 9 12 12 16 3 6 7 6 4 1 84 forest Pterostichus pensylvanicus LeConte, 1873 2 4 1 1 1 8 7 11 13 2 1 1 0 0 0 43 forest Sphaeroderus canadensis Darlington, 1933 16 6 2 3 1 2 0 3 3 0 1 0 3 0 0 40 forest Cymindis cribricollis Dejean, 1831 6 4 7 4 2 0 2 1 2 0 0 2 2 1 0 34 open Pterostichus tristis Dejean, 1828 4 1 4 1 0 2 0 1 3 0 2 0 1 0 0 19 forest Pterostichus mutus (Say, 1823) 0 1 0 0 0 2 2 2 4 4 0 0 0 0 0 16 generalist Platynus decentis (Say, 1823) 1 0 0 1 0 3 3 1 1 4 0 0 1 1 0 16 forest Agonum fidele Casey, 1920 0 0 0 0 1 0 0 0 0 3 0 0 0 1 2 7 swamp Olisthopus parmatus (Say, 1823) 0 0 0 0 1 2 0 1 0 1 0 0 0 0 0 4 forest Pterostichus melanarius (Illiger, 1798) 0 1 1 0 2 0 0 0 0 0 0 0 0 0 0 4 generalist Gastrellarius honestus (Say, 1823) 0 0 0 0 0 2 0 1 1 0 0 0 0 0 0 4 forest Pterostichus rostratus (Newman, 1838) 0 0 0 0 0 2 0 0 0 0 0 0 0 0 0 2 forest Pterostichus luctuosus (Dejean, 1828) 0 0 0 0 0 0 0 0 0 2 0 0 0 0 0 2 swamp Chlaenius aestivus Say, 1823 0 0 0 1 0 0 0 0 0 0 0 0 0 0 0 1 forest Harpalus somnulentus Dejean, 1829 0 0 0 0 1 0 0 0 0 0 0 0 0 0 0 1 open Elaphrus clairvillei Kirby, 1837 0 0 0 0 1 0 0 0 0 0 0 0 0 0 0 1 swamp Loricera pilicornis (Fabricius, 1775) 0 0 0 1 0 0 0 0 0 0 0 0 0 0 0 1 swamp Agonum corvus (LeConte, 1860) 0 0 0 0 0 0 0 0 0 1 0 0 0 0 0 1 open Calosoma frigidium Kirby, 1837 0 0 0 0 0 0 0 0 1 0 0 0 0 0 0 1 forest CON = Control; STS = Single-tree selection; FMS = Financial maturity selection; DLC = Diameter-limit cutting; ICR = Intensive crown release. See text for details. Habitat association designations are based off of classification systems from Latty et al. (2006) and Vance and Nol (2003), as well as carabid species descriptions by Larochelle and Larivière (2003).

June 2017 July 2017

Mature logged Old-growth Mature logged Old-growth Total Habitat Species (n = 985.9) (n = 559.5) (n = 1131) (n = 552.1) (n = 3228.4) association Agonum retractum forest LeConte, 1848 180 72 32 23 307 Pterostichus coracinus generalist (Newman, 1838) 52 14 46 22 134 Synuchus impunctatus generalist (Say, 1823) 0 0 73 46 119 Pterostichus pensylvanicus forest LeConte, 1873 45 16 1 1 63 Sphaeroderus canadensis forest Darlington, 1933 25 18 9 7 59 Pterostichus tristis forest Dejean, 1828 16 9 15 6 46 Platynus decentis forest (Say, 1823) 18 9 0 0 27 Olisthopus parmatus forest (Say, 1823) 15 1 0 0 16 Cymindis cribricollis open Dejean, 1831 5 1 4 2 12 Calathus advena (LeConte, 1848) 2 0 0 1 3 forest Agonum fidele swamp Casey, 1920 2 0 0 0 2 Pterostichus mutus generalist (Say, 1823) 2 0 0 0 2 Sphaeroderus lecontei Dejean, 1826 2 0 0 0 2 forest Pterostichus adstrictus generalist Eschscholtz, 1823 2 0 0 0 2 Pterostichus melanarius generalist (Illiger, 1798) 0 0 0 1 1 Gastrellarius honestus forest (Say, 1823) 0 1 0 0 1 Amara pennsylvanica Hayward, 1908 1 0 0 0 1 open Harpalus rubripes open (Duftschmid, 1812) 0 0 1 0 1 Harpalus indigens open Casey, 1924 0 0 1 0 1 Habitat association designations are based off of classification systems from Latty et al. (2006) and Vance and Nol (2003), as well as carabid species descriptions by Larochelle and Larivière (2003).