(Title of the Thesis)*

SPRUCE BARK BEETLE DISTURBANCE IN THE FOREST-TUNDRA ECOTONES OF SOUTHWEST YUKON: IMPACTS AND PREDISPOSING FACTORS

Michelle Leigh Mazzocato

A thesis submitted to the School of Environmental Studies

In conformity with the requirements for

the degree of Master of Environmental Studies

Queen’s University

Kingston, Ontario, Canada

(May, 2015)

As climate warms, phytophagous forest insects are expected to advance into previously unoccupied regions and/or experience population eruptions within their extant ranges. Already stressed by abiotic factors, arctic and alpine treeline environments will therefore become increasingly vulnerable to insect disturbances. Using vegetation surveys and dendrochronological techniques, we examined factors that may have predisposed the forest-tundra ecotones of southwest Yukon to spruce bark beetle (Dendroctonus rufipennis) infestation, and investigated the subsequent effects of insect disturbance on the growth and establishment of white spruce (Picea glauca). Specific objectives were to evaluate (i) the relationships between stand structure and variance in beetle-induced mortality and (ii) growth patterns between stands affected and relatively unaffected by the outbreak. Information on spruce size, reproduction and health, shrub cover, seedling density and size, stand density, and basal area was collected from three elevations at six sites, divided equally into three mortality classes. A subset of mature spruce individuals – both living and deceased – was sampled from one low-mortality and one medium-mortality site. Results demonstrated that hosts selected by spruce bark beetle were typically large individuals growing in dense stands, with spruce density and basal area dictating the extent of stand mortality. The majority of mortality was concentrated at – though not exclusive to – forestline elevations, though incidence of attack was sporadic across the region. At forestline, deceased individuals shared a period of accelerated growth when trees were between 75 and 150 years old, not present in their surviving counterparts, and had stronger temperature-growth correlations in early summer of the growing season than for still-living trees. Post-outbreak growth releases were observed in several surviving forestline mature spruce, but this accelerated growth was restricted to stands that experienced high levels of spruce mortality. The results indicate that host selection mechanisms and competitive restrictions in forest-tundra ecotones are similar to those in low-elevation, closed-canopy

ii forests, but suggest that additional factors may increase susceptibility to attack in treeline environments. Insect disturbance has the capacity to significantly alter treeline positions and dynamics, and is therefore an important factor to consider when assessing the future of forest-tundra ecotones.

iii

Acknowledgements

I would like to thank my supervisor, Dr. Ryan Danby, for giving me the opportunity to conduct research in the beautiful southwest Yukon and for offering guidance throughout this thesis project. I’d also like to thank my field assistants, Lucas Brehaut and Diana Zeng, as well as my fellow researcher

Robyn Laing, for helping with data collection in spite of the shrubs, mosquitos and rain. A sincere thank you to everyone at the Arctic Institute of North America’s Kluane Lake Research Station for providing a home away from home, and to the Champagne and Aishihik First Nations for permitting research on their traditional lands. This research was funded by a Northern Scientific Training Program grant from

Aboriginal Affairs and Northern Development Canada, and a Discovery Grant from the Natural Sciences and Engineering Research Council of Canada (awarded to Dr. Ryan Danby).

To my friends and colleagues in Environmental Studies and Geography: I couldn’t have asked for a better group of people to have alongside me on this journey. Whether it was support in the lab, nerding out over mutual interests, or just hanging out doing whatever, you kept me going and helped me stay sane. You are a bunch of incredible, crazy and big-hearted individuals, and I’m lucky to have met you.

Special thanks to my roommates for making our house a home in every sense of the word, and for being such amazing friends.

Finally, thank you to my parents and friends back home for the unwavering encouragement and support via phone calls, text messages, and emails. To Joseph: thank you for always understanding, for always believing, and for always somehow knowing what I needed to keep me moving forward, even when I didn’t know it myself.

Table of Contents

Abstract ...... ii Acknowledgements ...... iv List of Figures ...... vii List of Tables ...... viii Chapter 1 Introduction and Literature Review ...... 1 1.1 Background ...... 1 1.2 Treeline ...... 2 1.2.1 Limitations to Growth at Treeline ...... 2 1.2.2 Climate Change and Treeline ...... 4 1.2.3 Global Implications of Treeline Advance ...... 7 1.3 Natural Disturbance ...... 8 1.3.1 Physical Disturbance ...... 10 1.3.2 Fire Disturbance ...... 11 1.3.3 Insect Disturbance ...... 14 1.3.4 Synergistic Effects ...... 18 1.4 Dendrochronology ...... 19 1.5 Research Objectives ...... 22 Chapter 2 Effects of Differential Spruce Bark Beetle Mortality across Forest-Tundra Ecotones in Southwest Yukon, Canada ...... 25 2.1 Introduction ...... 25 2.2 Methods ...... 28 2.2.1 Study Area ...... 28 2.2.2 Field and Laboratory Methods ...... 29 2.2.3 Data Analysis ...... 32 2.3 Results ...... 35 2.3.1 Ecotone Structure ...... 35 2.3.2 Spruce Reproduction ...... 36 2.3.3 Seedling Growth ...... 36 2.3.4 Susceptibility to Spruce Bark Beetle ...... 37 2.4 Discussion ...... 38

2.4.1 Extent of Spruce Bark Beetle Mortality ...... 38 2.4.2 Susceptibility to Spruce Bark Beetle Attack ...... 40 2.4.3 The Future Structure of the Forest-Tundra Ecotone ...... 42 Chapter 3 Growth Patterns of Subalpine White Spruce (Picea glauca) in Relation to Varied Levels of Insect Disturbance, Southwest Yukon, Canada ...... 55 3.1 Introduction ...... 55 3.2 Methods ...... 59 3.2.1 Study Area ...... 59 3.2.2 Dendrochronological Analysis ...... 60 3.3 Results ...... 65 3.3.1 Comparison of Standardized Chronologies...... 65 3.3.2 Climate-Growth Relationships ...... 68 3.3.3 Post-Outbreak Growth Releases ...... 68 3.4 Discussion ...... 69 3.4.1 Growth Trends and Climate-Growth Relationships ...... 69 3.4.2 Post-Outbreak Tree Growth ...... 72 Chapter 4 Summary and Conclusions ...... 85 References ...... 90 Appendix A Summaries of Site Characteristics ...... 105 Appendix B Comparing Ecotone Characteristics across Elevations and Mortality Levels ...... 111

List of Figures

Figure 1-1 Visual representation of an alpine forest-tundra ecotone ...... 24 Figure 2-1 Spruce bark beetle-affected trees within forest-tundra ecotones of southwest Yukon ...... 46 Figure 2-2 Location of Kluane region study sites, southwest Yukon ...... 47 Figure 2-3 Diagram of sampling technique ...... 48 Figure 2-4 Levels of female cone production in spruce trees (≥2m in height) at treeline and forestline elevations, separated by mortality class ...... 49 Figure 2-5 White spruce seedling size metrics and radial growth rates relative to age ...... 50 Figure 2-6 Box plots illustrating the size ranges of sampled white spruce that survived the recent spruce bark beetle outbreak (n=322) in comparison to those that were killed (n=101) ...... 51 Figure 2-7 Pre-outbreak spruce basal area and stand density relative to the extent of mortality induced by spruce bark beetle ...... 52 Figure 3-1 Location of Kluane region dendrochronological study sites, southwest Yukon ...... 76 Figure 3-2 Standardized chronologies for Mount Decoeli and Telluride Creek at treeline and forestline elevations ...... 77 Figure 3-3 Marker years (± 1.5 standard deviations of the ring-width mean) for each chronology ...... 78 Figure 3-4 Correlation between total monthly precipitation and the residual chronologies from Mount Decoeli and Telluride Creek for the 1968–1998 time period ...... 79 Figure 3-5 Correlation between mean monthly temperature and the residual chronologies from Mount Decoeli and Telluride Creek for the 1968–1998 time period ...... 80 Figure 3-6 Percentage change in radial growth (following Rubino and McCarthy (2004)) for mature spruce individuals growing at Mount Decoeli and Telluride Creek treeline and forestline elevations ...... 81

vii

List of Tables

Table 2-1 Post-outbreak attributes of forest-tundra ecotone sites with varied spruce bark beetle mortality across three elevations ...... 53 Table 2-2 Binary logistic analysis of trees surviving the recent spruce bark beetle attack and those that were killed ...... 54 Table 3-1 Attributes of the Mount Decoeli and Telluride Creek standardized ring-width chronologies ... 82 Table 3-2 Pearson’s product-moment correlation coefficients between standardized chronologies for the common period (1866 – 1998) ...... 83 Table 3-3 Growth trends from the Mount Decoeli and Telluride Creek chronologies from 1867–1998. .. 84 Table A-1 Description of Site 1 (Mount Decoeli – low mortality) ...... 105 Table A-2 Description of Site 2 (Boutellier Creek – high mortality) ...... 106 Table A-3 Description of Site 3 (Telluride Creek – medium mortality) ...... 107 Table A-4 Description of Site 4 (Mount Decoeli – low mortality) ...... 108 Table A-5 Description of Site 5 (Boutellier Creek – high mortality) ...... 109 Table A-6 Description of Site 6 (Telluride Creek – medium mortality) ...... 110 Table B-1 Results from Kruskal-Wallis tests to compare ecotone characteristics between elevations ... 111 Table B-2 Assessment of the relationship between percentage spruce bark beetle mortality and various ecotone characteristics, using Spearman’s rank correlation coefficient ...... 112 Table B-3 Results from statistical analyses comparing seedling attributes between elevations, mortality class, and transect-specific percentage spruce bark beetle kill ...... 113

viii

Chapter 1

Introduction and Literature Review

1.1 Background

Climate warming is expected to have pronounced effects on arctic and subarctic ecosystems, including forest-tundra ecotones (or, treelines). Sustained warm temperatures can lead to an advance of the altitudinal and latitudinal limits of trees and other plants (Lescop-Sinclair and Payette 1995,

Suarez et al. 1999, Sturm et al. 2001, Walther et al. 2005, Tape et al. 2006, Soja et al. 2007, Devi et al.

2008, Frei et al. 2010), alter growth forms (Lescop-Sinclair and Payette 1995, Gamache and Payette

2004, Danby and Hik 2007b, Devi et al. 2008), and affect the radial growth of woody species (Jacoby and

D’Arrigo 1995, Suarez et al. 1999, Barber et al. 2000, Vittoz et al. 2008, Salzer et al. 2009, Zhang et al.

2011). Widespread modifications to forest-tundra ecotones have the capacity to cause feedbacks to the climate system (Grace et al. 2002, Symon et al. 2005, Chapin et al. 2005) and lead to changes in the range distributions of, and relationships between, flora and fauna species (Scheller and Mladenoff 2005,

Parmesan 2006, Pauli et al. 2007, Frei et al. 2010, Ettinger et al. 2011, Myers-Smith et al. 2011a).

At the same time, warming temperatures may affect the distribution of phytophagous insects

(Logan and Powell 2001, Carroll et al. 2004, Bentz et al. 2010, Marini et al. 2012, Weed et al. 2013,

Murdock et al. 2013) and allow their expansion into previously unaffected high altitude and latitude regions. In addition, climate can trigger outbreaks of endemic insect populations (Schmid and Frye

1977, Werner and Holsten 1985, Ayres and Lombardero 2000, Berg et al. 2006, Werner et al. 2006, Raffa et al. 2008, Bentz et al. 2010, Weed et al. 2013). Understanding the characteristics and impacts of insect disturbance within forest-tundra ecotones will therefore become essential to predicting changes to the distributional limit of trees.

This thesis presents the results of research that examined subalpine forests affected by spruce bark beetle in southwest Yukon, Canada. The overarching goal of this research was to provide a better understanding of the impacts of insect disturbance on the growth and establishment of trees in a subarctic treeline environment, as well as the factors that may predispose forest-tundra ecotones to infestation. This chapter acts as an overarching introduction for the two empirical chapters of this thesis by providing an overview of the scientific literature on the characteristics of forest-tundra ecotones, the role of natural disturbances in treeline environments, and how climate change may influence both treeline, natural disturbance, and the interactions between them.

1.2 Treeline

In subarctic and alpine regions, forested areas become increasingly open and discontinuous as latitudes and elevations increase, until trees become prostrate and eventually stop growing altogether

(Körner 1998, Grace et al. 2002, Holtmeier and Broll 2005). The transition zone extending from closed- canopy forest to the latitudinal and altitudinal limits of tree distribution is referred to as the forest- tundra ecotone, or treeline (Figure 1-1). While it may manifest as an abrupt delineation at some locations, the more common treeline form is diffuse – a gradual and fragmented decline in tree size and density that only appears as a ‘line’ at a global scale (Payette et al. 2001, Callaghan et al. 2002, Grace et al. 2002, Körner and Paulsen 2004, Harsch and Bader 2011).

1.2.1 Limitations to Growth at Treeline

Numerous factors have been hypothesized to limit tree establishment and growth at treeline, thereby controlling the current position and spatial structure of the ecotone. These include, but are not restricted to: low air and soil temperatures, limited moisture availability, poor reproductive capacity and seedling establishment, seedling competition with field layer plants, insufficient carbon and/or nutrient uptake, restricted solar radiation, and repeated damage by frost, wind and herbivory (Tranquillini 1979,

Stevens and Fox 1991, Körner 1998, Sveinbjörnsson 2000, Grace et al. 2002, Holtmeier and Broll 2005,

Moen et al. 2008). However, the processes which govern treelines are highly complex and, despite decades of ecological research, the interactions between them are not well understood (Callaghan et al.

2002, Grace et al. 2002). Furthermore, the relative importance of each of these mechanisms appear to be spatially and temporally variable (Sveinbjörnsson 2000, Callaghan et al. 2002, Holtmeier and Broll

2005, Danby and Hik 2007b, Moen et al. 2008, Stueve et al. 2011).

That said, temperature is considered to be the most important factor regulating treeline growth and ecotone expansion; not only does it dictate the length of the growing season, but also the success of heat-dependent ecological processes (Tranquillini 1979, Körner 1998, Körner and Paulsen 2004). On a global scale, the 10°C July isotherm coincides reasonably well with the position of the forest-tundra ecotone in regions between 40° and 70° latitude (Körner 1998), which also encompasses the distribution of the boreal forest. At a physiological level, low temperatures have been proposed to limit apical meristem growth by inhibiting the production of new cells and restricting the rate at which photosynthetic materials can be used (Tranquillini 1979, Körner 1998). Low soil temperatures may also constrain both growth of roots and shoot elongation (Tranquillini 1979, Körner 1998). This impairs annual growth, as well as preventing individuals from adequately replacing biomass lost to mechanical or herbivorous damage. In addition, temperature plays an important role in seedling establishment and recruitment. Specifically, the production of viable seeds has been shown to be dependent on sufficiently high ambient summer temperatures (Sirois 2000, Holtmeier and Broll 2005). Once established, seedling survival is facilitated by the protection and warmer microclimates of surrounding dwarf shrubs and clusters of trees with stunted, prostrate forms known as krummholz mats. As they grow taller, these seedlings become increasingly decoupled from microsite temperatures and more vulnerable to damage or death from wind desiccation and/or herbivory (Grace et al. 2002, Smith et al.

2003). 3

1.2.2 Climate Change and Treeline

There is a global warming trend in response to anthropogenic increases in carbon dioxide and other greenhouse gases, with high elevations and latitudes experiencing climate warming the most acutely (Stocker et al. 2013). Regions north of 60° underwent substantial climate warming during the latter half of the 20th century, at a rate higher than at lower latitudes (Symon et al. 2005, Chapin et al.

2005, Prowse et al. 2009, Stocker et al. 2013). This trend is anticipated to continue, with temperatures predicted to increase by 2° to 7°C by 2100 (Symon et al. 2005). Similarly, mountain ecosystems are highly susceptible to climate warming and are predicted to experience two to three times more warming than experienced over the last one hundred years (Nogués-Bravo et al. 2007, Rebetez and

Reinhard 2007). This makes mountains at higher latitudes, such as those in southwest Yukon, particularly vulnerable to the effects of climate change.

As the most significant factor governing growth and establishment at treeline, warming temperatures are expected to have pronounced effects on forest-tundra ecotones. Evidence from both

North America and Eurasia suggests that some of these changes are already occurring, including alterations to vegetation distribution and structure, and the advancement of trees and shrubs into areas currently occupied by tundra. Prolonged periods of warm temperatures have led to increased vertical growth in seedlings and krummholz individuals such that they are able to overcome the negative effects of above-ground microclimates and achieve erect growth forms (Lescop-Sinclair and Payette 1995,

Gamache and Payette 2004, Danby and Hik 2007b, Devi et al. 2008). This has also facilitated increased seedling establishment, growth rates and survivorship (Szeicz and MacDonald 1995, Lloyd and Fastie

2003, Danby and Hik 2007b) in regions where warming is sufficient to promote sexual reproduction (cf.

Lescop-Sinclair and Payette 1995), as well as radial growth increases in mature individuals (Suarez et al.

1999, Vittoz et al. 2008, Salzer et al. 2009). In turn, increased tree establishment may lead to stand densification, as has been observed across northern Canada in response to an increase in temperature

following the end of the Little Ice Age (circa 1850) (Payette and Filion 1985, Scott et al. 1987, Szeicz and

MacDonald 1995, MacDonald et al. 1998). Increased treeline and tundra plant biomass (Devi et al.

2008, Rammig et al. 2010, Danby et al. 2011) and densification of shrub cover (Sturm et al. 2001, Tape et al. 2006, Myers-Smith et al. 2011b, Ropars and Boudreau 2012) in response to warming has also been reported in several regions. Furthermore, warming trends have been shown to promote the advance of arctic and alpine plants and woody species into areas which they did not previously occupy (Lescop-

Sinclair and Payette 1995, Suarez et al. 1999, Sturm et al. 2001, Walther et al. 2005, Tape et al. 2006,

Soja et al. 2007, Devi et al. 2008, Frei et al. 2010).

While common, the altitudinal and latitudinal advance of species is not universal. In northern

Canada and Alaska, for example, some studies have demonstrated an expansion of white spruce (Picea glauca (Moench) Voss) and black spruce (P. mariana (Mill.) BSP) at latitudinal treelines (Lescop-Sinclair and Payette 1995, Suarez et al. 1999), while others have documented significant infilling of the pre- existing stands but only minimal expansion (Payette and Filion 1985, Scott et al. 1987, Szeicz and

MacDonald 1995). According to Szeicz and MacDonald (1995), this limited advance of individuals despite recent warming is due to an inertia inherent in white spruce populations at the margins of their distribution. Another line of reasoning, proposed by Danby and Hik (2007a), is that warming-induced changes affecting white spruce within the forest-tundra ecotone are hierarchical in nature, beginning with an increase in individual growth, stature and reproductive capacity, then an increase in stand density, and finally an expansion in species distribution. As such, the advance, expansion and/or densification of treelines may lag behind climate warming by several decades (Paulsen et al. 2000,

Gamache and Payette 2004, Holtmeier and Broll 2005, Lloyd 2005).

Modifications to the forest-tundra ecotone will not necessarily be uniform across all locations, due to site-specific factors and climate interactions with other limitations to growth and/or advance.

For instance, aspect has been demonstrated to play an important role in modulating the type and rate 5

of change in white spruce at altitudinal treelines. In southwest Yukon, Danby and Hik (2007c) reported a consistent and rapid advance of treeline on south-facing slopes in response to mid-20th century warming, whereas north-facing slopes experienced significant increases in stand density. Similarly,

Stueve et al. (2011) found that recent upslope shifts and densification of central Alaskan treelines were considerably greater on south aspects than on north aspects. Seedlings on south-facing slopes have also been observed to experience greater radial growth than those on north aspects (Danby and Hik 2007b).

Likewise, species’ responses to climate warming are dependent on their relative position in the forest- tundra ecotone. In the European Alps, Frei et al. (2010) postulated that alpine species close to their upper range limits are highly responsive to warming temperatures. At their lower range margins, however, responsiveness is more idiosyncratic, with several species experiencing downslope shifts rather than the anticipated upwards shifts in distribution. In Alaska, Lloyd and Fastie (2002) found considerable variation in response to climate warming between white spruce populations at and below treeline. Specifically, trees located less than 75m downslope from treeline were far more likely to experience growth declines in response to increased temperature during the latter half of the 20th century than those at treeline itself.

Differential responses to warming temperatures at upper and lower range margins may be in part due to the current structure of treeline environments. In the northern forest-tundra ecotone of

Québec, for example, warming has triggered tree shoot elongation and a shift from krummholz to erect growth forms, whereas growth forms are already arborescent in the southern forest-tundra ecotone and warming has therefore instead altered reproductive patterns (Gamache and Payette 2004).

Concurrent biotic and abiotic influences may further complicate treeline responses to temperature increases, particularly at lower elevations and latitudes, where species are less temperature-limited than those at their range margins (Holtmeier and Broll 2005, Frei et al. 2010, Stueve et al. 2011, Ettinger et al. 2011, Ohse et al. 2012). Alpine plants at the lower limits of their range margins may demonstrate 6

a lack of upslope movement because they are in competition with species that have themselves moved upslope due to recent temperature changes (Frei et al. 2010).

In addition, while temperature increases are generally expected to affect trees at treeline positively, high temperatures may also increase evapotranspiration. This could subsequently lead to negative effects being experienced in areas where moisture is limited by temperature-induced drought or low precipitation levels (Jacoby and D’Arrigo 1995, Barber et al. 2000, Lloyd and Fastie 2002, Daniels and Veblen 2004, Wilmking et al. 2005, Lloyd and Bunn 2007, Chhin and Wang 2008, Zhang et al. 2011,

McLane et al. 2011). Findings by Miyamoto et al. (2010) suggest that growth of spruce, pine and fir in

British Columbia and Yukon will increase with higher growing season temperatures on cool sites (i.e. at higher elevations and/or on north-facing slopes) and decrease on warm sites (i.e. at lower elevations and/or on south aspects). Trees may also respond more positively to temperature during cool periods of climatic regimes, such as the Pacific Decadal Oscillation (PDO), than in warm periods (Ohse et al.

2012). In other words, positive growth at treeline appears to be at least partially dependent on the interaction between temperature increase and moisture availability. If temperature limitations on growth are lifted, tree growth may become decoupled from temperature and forest-tundra ecotones may instead become increasingly governed by factors such as natural disturbance (Ohse et al. 2012).

After analyzing a global dataset, Harsch et al. (2009) concluded that, given its influence on treeline structure and recruitment, disturbance may delay the initiation of treeline advance. In contrast, a study by Van Bogaert et al. (2011) demonstrated that disturbance may in fact prevent the occurrence of treeline advance altogether.

1.2.3 Global Implications of Treeline Advance

Several studies of alpine tundra vegetation have found increased species richness and diversity in response to recent warming trends, with long-term changes in community composition anticipated for

the future (Pauli et al. 2007, Rammig et al. 2010). That said, the advance of shrubs and trees is predicted to replace between 11 and 50% of tundra worldwide if temperatures continue to increase

(Barros et al. 2014). As such, the invasion of woody species into tundra ecosystems could have negative impacts on biodiversity. A review by Myers-Smith et al. (2011a) concluded that increases in shrub cover may result in the loss of shade-intolerant tundra plants. In addition, range shifts may lead to asynchrony between species’ predators, food, and habitat resources (Parmesan 2006), replacement by more competitive or warmth-tolerant species (Pauli et al. 2007, Frei et al. 2010), range contractions or extirpation (Scheller and Mladenoff 2005, Parmesan 2006, Ettinger et al. 2011). These changes will subsequently influence existing food webs, animal populations and human livelihoods. For instance, an increase in willow (Salix spp.) cover may simultaneously benefit forager populations of moose and ptarmigan and, by decreasing lichen availability, negatively impact caribou and reindeer populations as well as associated hunting and herding (Myers-Smith et al. 2011a).

Widespread changes to forest-tundra ecotones could also have significant global impacts through feedbacks to the climate system, although consensus on the nature of these feedbacks has not yet been reached. An increase in plant biomass production has a greater capacity to act as a carbon sink, meaning that expansion and/or densification of trees and shrubs could mitigate future warming through negative feedbacks (Grace et al. 2002). However, as woody species have a much lower albedo than the tundra (Foley et al. 1994, Betts 2000), tree and shrub infilling and/or advance could exacerbate climate warming through positive feedbacks (Symon et al. 2005, Chapin et al. 2005).

1.3 Natural Disturbance

A natural disturbance is a relatively discrete event that disrupts ecosystem, community or population structure and alters resource availability or the physical environment (White and Pickett

1985). Disturbance events can be abiotic (e.g. floods, volcanic eruptions, windstorms, landslides or fires)

or biotic (e.g. insect outbreaks or invasion of exotic species) in origin, and occur over relatively short periods of time. Disturbance regimes, on the other hand, describe the long-term spatial and temporal distribution of disturbances and their resultant impacts on organisms, communities or ecosystems

(Turner et al. 2001).

Because natural disturbances do not affect all areas to the same extent, they create complex mosaic-like heterogeneous patterns which vary across the landscape (Baker 1995, Turner et al. 2001).

These spatial patterns can dictate recovery from disturbance events, particularly in terms of community succession. Successional trajectories are highly dependent on both local environmental factors (e.g. soil moisture) and disturbance characteristics (e.g. intensity, size), and become more difficult to predict following large, intense disturbance events. An increase in disturbance frequency may also cause a qualitative shift in successional pathways (Turner et al. 2001).

Ecosystems are often well-adapted to, or even reliant on, natural disturbances for maintenance and success. For example, in the fire-prone North American boreal forest, black spruce (Picea mariana) are able to self-replace after fire disturbances due to their production of semi-serotinous cones (Rowe and Scotter 1973). Similarly, coevolved populations of plants and the insects which feed on them may form resilient systems (Murdock et al. 2013) where both host and herbivore will persist despite fluctuations in host abundance and the occurrence of insect outbreaks. Defined as the ability of an ecological system to retain its relationships after undergoing disturbance (Holling 1973), resilience is determined by a system’s evolutionary history and the range of disturbances it has experienced.

Alterations to disturbance regimes, such as those initiated by climate change, may therefore have significant impacts on the landscape (Symon et al. 2005, Stocker et al. 2013), though the response of the landscape structure will not be uniform (Johnstone et al. 2010a, 2010b) and may lag behind these changes (Baker 1995).

In the following subsections, an overview of the impacts of physical, fire, and insect disturbance is provided. Given the focus of this thesis, these subsections are not meant to review the extensive literature on natural disturbances. Rather, they are intended to emphasize the importance of natural disturbance at treeline in a changing climate.

1.3.1 Physical Disturbance

Abiotic physical disturbances, such as windstorms, landslides and snow avalanches, often have long-lasting impacts on the ecosystems they disrupt. Depending on their characteristics (e.g. type, size and intensity), physical disturbances can cause large-scale mortality or merely affect community structure and composition. For example, small windstorms can cause physical damage to forests by breaking branches and stems, or even uprooting individual trees (Mayer et al. 1989). Hurricanes and tornados, on the other hand, are capable of causing heavy tree mortality, reduced tree density, and openings in the canopy (Dale et al. 2001). If early-successional shade-intolerant species are more susceptible to windthrow, as suggested by Rich et al. (2007), windstorms will change community composition by favouring more late-successional species. Should wind disturbances become more frequent and intense, as projected by some climate scenarios, this could create increased landscape heterogeneity (Dale et al. 2001, Rich et al. 2007).

In alpine regions, geomorphological processes and wind disturbances exert influence on and, in some cases, control the elevation and dynamics of forest-tundra ecotones. Snow avalanches and mass- wasting events (e.g. rock falls, landslides and debris flows) have the ability to uproot or damage established trees at or below treeline (Butler et al. 2009). In addition, the substrates deposited downslope by these processes are often unfavourable for seedling establishment and growth, being unstable with a lack of soil moisture and nutrients (Holtmeier and Broll 2005, Butler et al. 2009). As such, these disturbances are able to create and maintain treelines at elevations well below favourable

climatic limits (Holtmeier and Broll 2005, Butler et al. 2009, Leonelli et al. 2011). In parts of Montana’s

Glacier National Park, for instance, snow avalanches are considered to be the primary controller of the forest-tundra ecotone, while debris flows frequently destroy trees attempting to colonize above treeline

(Butler et al. 2009). Since wind velocity tends to increase with elevation (Holtmeier and Broll 2010), alpine treelines are also more susceptible to wind disturbances than forests downslope (Cullen et al.

2001). However, windthrow seems relatively unimportant within alpine forest-tundra ecotones; rather, persistent strong winds at treeline, though not categorized as disturbance events, may cause physical damage (e.g. breakage of crowns and branches) and reduce the vertical growth of trees (Holtmeier and

Broll 2010).

With 21st century projections of increased temperatures and alterations to precipitation, many of these processes are subject to change. It is unclear whether the magnitude and/or frequency of snow avalanches will be affected by future climate change, given the inherent difficulties in predicting these events (Butler et al. 2009). However, permafrost degradation on steep alpine slopes may reduce slope stability and increase the probability of rock falls (Haeberli and Gruber 2009). Furthermore, because landslides can be triggered by intense rainfall and rapid snowmelt, alpine regions with increased precipitation or accelerated snowmelt will likely experience more frequent landslides (Dale et al. 2001).

Therefore, the upward migration of alpine treelines will likely continue to be inhibited by disturbances such as snow avalanches and mass-wasting, despite climatic conditions allowing for its advance (Butler et al. 2009, Holtmeier and Broll 2010, Leonelli et al. 2011).

1.3.2 Fire Disturbance

Wildfire is an important natural disturbance agent inherent to many forested ecosystems, and is widely considered to be a dominant driver of ecological processes in both the North American and

Eurasian boreal forest (McCullough et al. 1998, Flannigan et al. 2005, Soja et al. 2007, Payette et al.

2008). Fire disturbances can be categorized by their fuel types as ground fires (subsurface organic fuels), surface fires (needle and leaf litter, logs and ground vegetation), or crown fires (crowns of standing trees; typically ignited by surface fires) (McCullough et al. 1998). The frequency, intensity, areal extent, and type of fire disturbance is dependent on weather and climate, fuel availability and moisture content, ignition agents, and human influence (Rowe and Scotter 1973, McCullough et al.

1998, Dale et al. 2001, Flannigan et al. 2005, Soja et al. 2007). The ecological effects of fire disturbances are highly variable but include the mortality of individual trees and plants, the creation of seed beds favourable for seed germination, the loss of canopy and soil seed banks, and subsequent alterations to forest age structure, species composition, floristic diversity, and nutrient cycling (McCullough et al.

1998, Dale et al. 2001, Soja et al. 2007). Repeated episodic fire disturbances in the boreal forest have therefore created a complex mosaic of forest structure, assisting in the maintenance of heterogeneous and diverse species populations (Rowe and Scotter 1973, McCullough et al. 1998).

Fire disturbance has also been demonstrated to play an important role in modulating forest- tundra ecotone vegetation distribution and dynamics. For example, the mosaic-like structures of eastern Canadian subalpine and arctic treelines have been attributed to varied fire regimes throughout the late Holocene, with the length between fire disturbances in different zones contributing to the variance in dominant tree species (Payette et al. 2001, de Lafontaine and Payette 2010). However, these forest-tundra ecotones have experienced fire less frequently than the more continuous boreal forest, with a reduction in fire activity occurring with increasing proximity to tree species limit, due to fewer available fire fuels and cooler temperatures (Payette et al. 2008). While longer fire return intervals may allow trees to develop crowns that produce large crops of cones, those at higher latitudes have only limited reproductive potential (Sirois 2000). As such, fire disturbances at upper treelines have been found to cause a reduction in, or even failure of, tree regeneration (Payette and Filion 1985,

Payette et al. 2001, 2008, Coop et al. 2010). 12

As the climate warms, fire disturbance events across the boreal forest are anticipated to occur with more frequency and severity through increased fire season length, fire weather severity, flammability of individuals, and ignitions from lightning (Stocks et al. 1998, Dale et al. 2001, Symon et al.

2005, Flannigan et al. 2005, Soja et al. 2007). In addition to releasing carbon into the atmosphere during fire disturbances, the carbon storage capacity of the boreal forest is reduced post-fire (Harden et al.

2000). An increase in boreal fire activity will thus have positive feedbacks to the climate system, intensifying the climate change process and potentially leading to future fires.

Should climate-driven changes trigger a reduction in fire return intervals and/or an increase in area burned, this may lead to changes in species composition and stand dominance. The findings of

Brown and Johnstone (2012) suggest that serotinous communities are limited by seed availability and are therefore vulnerable to regeneration failure if repeat disturbances occur within a relatively short time frame (i.e. before the production of viable cones is possible). Likewise, seedling density tends to decrease with distance to burn interior (Coop et al. 2010). As such, seedling establishment following large, intense fires will be insufficient to recreate previous stand densities. Post-fire stand vegetation shifts from spruce-dominated forest to deciduous-dominated successional trajectories have been documented at the 1958 Takhini (Hogg and Wein 2005) and 1998 Fox Lake (Johnstone et al. 2010b) burn sites of southwest and southcentral Yukon, respectively, as well as various sites in interior Alaska burned in 2004 (Johnstone et al. 2010a). While site moisture was found to have a positive effect on all three tree species at these sites, warm and dry post-fire conditions may have favoured the recruitment of deciduous trees over spruce (Johnstone et al. 2010a), particularly on south-facing slopes (Johnstone et al. 2010b), and resulted in slower-than-expected post-fire succession (Hogg and Wein 2005). In treeline environments, where tree regeneration is already limited by seed availability and slow growth, fire disturbances may therefore lead to significant shifts in species composition. Consequently, changes to

fire disturbance regimes may result in a regression of treeline forests rather than the densification and/or advance of trees predicted with climate warming (Brown and Johnstone 2012).

1.3.3 Insect Disturbance

Phytophagous insects are prevalent disturbance agents in temperate forests, often affecting significantly more forest than wildfire (Malmström and Raffa 2000, Dale et al. 2001). Defoliating insects, such as the spruce budworm (Cloristoneura occidentalis Freeman, C. fumiferana Clemens) and the gypsy moth (Lymantria dispar L.), damage trees by eating leaves and needles, which can lead to reduced growth, increased susceptibility to other disturbances, and occasionally the death of the host tree

(Kulman 1971, Filion et al. 1998, McCullough et al. 1998, Wilmking et al. 2012). Other insects, including the mountain pine beetle (Dendroctonus ponderosae Hopkins), the North American spruce bark beetle

(D. rufipennis Kirby) and the European spruce bark beetle (Ips typographus L.), are phloem-feeding species that reduce plant vigour and generally cause the mortality of their hosts (Schmid and Frye 1977).

Damage to and mortality of trees by episodic insect outbreaks can increase understory plant diversity

(Stone and Wolfe 1996), accelerate the growth of remaining trees and saplings (Romme et al. 1986,

Veblen et al. 1991a, Parish et al. 1999, Zhang et al. 1999, Eisenhart and Veblen 2000, Kulakowski et al.

2003, Alfaro et al. 2004, Berg et al. 2006, Axelson et al. 2010, Sherriff et al. 2011, Smith et al. 2012,

Amoroso et al. 2013), and cause a shift in stand dominance (Veblen et al. 1991a, DeRose and Long

2010), thereby altering forest structure and composition. In addition, outbreaks may adversely affect nutrient cycling, carbon sequestration and biodiversity (Ayres and Lombardero 2000, Kurz et al. 2008).

However, they may still be crucial to maintaining long-term ecosystem integrity (Logan et al. 2003).

Insect infestations have been particularly acute in the North American cordillera in recent decades, and this is believed to be driven at least in part by climate change (Ayres and Lombardero

2000, Safranyik and Wilson 2006, Berg et al. 2006, Werner et al. 2006, Logan and Powell 2009, Sherriff

et al. 2011, Weed et al. 2013). For example, spruce bark beetle populations have affected over one million hectares of forest in Alaska (Werner et al. 2006) and approximately 400,000 hectares in southwest Yukon (Government of Yukon 2013) between 1990 and present day. In both cases, these outbreaks appear to have been triggered and sustained by prolonged warmer-than-average temperatures (Berg et al. 2006).

The spruce bark beetle (Dendroctonus rufipennis) is a coniferphagous species which attacks and reproduces in live trees, with white spruce (Picea glauca) and Engelmann spruce (P. engelmannii Parry ex. Engelm.) being the most abundant host species in its range (Furniss and Carolin 1977, Schmid and

Frye 1977, Maroja et al. 2007). Because of their ability to kill healthy trees and their capacity to cause extensive landscape-scale mortality during periodic population eruptions (Furniss and Carolin 1977,

Schmid and Frye 1977, Holsten et al. 1999), the spruce bark beetle is important both ecologically and economically (Veblen et al. 1991a, Ayres and Lombardero 2000, Maroja et al. 2007).

Selection and colonization of hosts is a complex process, of which some aspects are not yet fully understood. Spruce bark beetles are able to distinguish the potential defensive capacity of a host tree

(Raffa et al. 2008, Bentz et al. 2010). Whether these beetles initially select their hosts through the use of chemoreceptors to detect attractive host volatiles or by landing on them at random and sampling them for suitability has been much debated (Pureswaran and Borden 2005). Regardless, once a host tree has been identified, adult spruce bark beetles facilitate recruitment of other adults by emitting aggregation pheromones. Since this can attract several thousand beetles within a few days, colonizing beetles excrete repelling pheromones once tree resistance has been exhausted, so as to limit competition with other individuals. Adult spruce bark beetles bore into the phloem, copulate, and excavate galleries wherein they oviposit. The resultant larvae excavate additional feeding galleries. This process eventually results in the mortality of the host tree, so long as host defenses are unable to

overwhelm the beetle populations. The new generation of adult beetles then tunnel outward through the bark, emerge, and begin the search for a new host tree (Raffa et al. 2008, Bentz et al. 2010).

Spruce bark beetles typically target slow-growing, large-diameter trees (Schmid and Frye 1977,

Hard 1985, Doak 2004, Hart et al. 2014), although smaller trees may be attacked during severe epidemics (Veblen et al. 1991a, Werner et al. 2006, Hart et al. 2014). In addition, spruce bark beetle population size dictates the vigour of the host trees targeted. In response to attack, healthy conifers are able to pitch out adult beetles or flood brood galleries with resin (Raffa et al. 2005, Garbutt et al. 2006).

Healthy trees are therefore avoided unless beetle populations are large enough to overwhelm their defensive capacities (Wallin and Raffa 2004).

However, spruce bark beetle populations only erupt intermittently, remaining in an endemic state for long periods of time. A combination of host availability and suitability, beetle population density and synchrony, and climatic conditions are required to reach a landscape-level outbreak (Raffa et al. 2008).

In particular, summer temperatures may play an important role in driving outbreak intensity and frequency, through their influence on life history strategies. With increased temperatures, there are shifts in spruce bark beetle generation duration from semivoltine (2-year maturation) to univoltine (1- year maturation) broods (Schmid and Frye 1977, Werner and Holsten 1985, Berg et al. 2006), which in turn increases the probability of population outbreaks. Warmer phloem temperatures have a positive impact on beetle development (Werner and Holsten 1985). Also, milder winters may reduce cold- induced mortality (Schmid and Frye 1977, Ayres and Lombardero 2000, Bentz et al. 2010). These processes are believed to increase synchrony of adult emergence between populations (Logan et al.

2003, Raffa et al. 2008), thereby furthering outbreak intensity. Since moisture stress can reduce host vigour and defense mechanisms (Raffa et al. 2005, Berg et al. 2006, Kausrud et al. 2012), more frequent and/or severe temperature-induced drought may intensify outbreaks. A decrease in defensive capacity

will enhance host susceptibility to attack, as a reduced number of spruce bark beetles are required to overwhelm host defenses.

Warming temperatures may therefore facilitate epidemic-level outbreaks of endemic phytophagous insects, such as the spruce bark beetle, and/or favour an expansion of outbreak range by affecting beetle development rates, survival, emergence and voltinism (Schmid and Frye 1977, Werner and Holsten 1985, Ayres and Lombardero 2000, Logan and Powell 2001, Logan et al. 2003, Safranyik and

Wilson 2006, Berg et al. 2006, Bentz et al. 2010, Kausrud et al. 2012, Weed et al. 2013). Once the eruptive threshold has been reached, the initial factors which triggered it (such as high temperatures) may not be needed to sustain it (Raffa et al. 2008). Still, the beetle-temperature relationship may decouple once the supply of susceptible host trees is exhausted (Werner et al. 2006).

While insect disturbance can extend from closed-canopy forest into the forest-tundra ecotone, there have been relatively few studies documenting outbreaks at either latitudinal or altitudinal treeline. Those that do exist have focused on the historic occurrence of insect disturbance (e.g.

Caccianiga et al. 2008) or the effects of defoliating insects (e.g. Van Bogaert et al. 2011). From the limited information available, the main effect of insect disturbance at treeline appears to be mortality

(Caccianiga et al. 2008, Van Bogaert et al. 2011). Already stressed by abiotic factors, trees growing at the margins of their distribution may recover slowly from an insect outbreak (Raffa et al. 2008, Moen et al. 2008). As such, repeated or long-term infestations may have a lasting influence on treeline structures.

As climate continues to warm, trees at the altitudinal and latitudinal limits of their distribution may become more vulnerable to future insect disturbances. Individuals that experience positive growth trends with higher temperatures may fall within the desired size range for infestation, and those that undergo climate-associated drought stress will be more susceptible to attack. Temperature-induced population eruptions (Berg et al. 2006, Raffa et al. 2008) and geographic range expansions (Ayres and 17

Lombardero 2000, Logan and Powell 2001, Carroll et al. 2004, Bentz et al. 2010, Marini et al. 2012) of phytophagous insects may further threaten marginal trees. That said, the responses of these insect populations to climate change are complex and variable (Bentz et al. 2010, Marini et al. 2012).

Knowledge gaps regarding the effects of insect disturbance on forest-tundra ecotones in a warming climate must therefore be addressed to ensure the adequate management of and predictive capacity for future outbreaks.

1.3.4 Synergistic Effects

While major disturbances may appear to be isolated events, they are in fact often causally related. For example, fire and insect disturbances can interact synergistically to affect ecosystem succession, nutrient cycling, and species composition (McCullough et al. 1998). While fire disturbance is essential for stand renewal in many areas, it may limit recovery or expansion when paired with insect disturbance. With eastern spruce budworm (Choristoneura fumifera) outbreaks, the majority of black spruce cone biomass is lost; these seed-depleted stands are therefore unable to facilitate black spruce regeneration should fire occur, leading to stand opening and a shift in vegetation dominance and composition (Simard and Payette 2005). In addition, mortality caused by episodic insect outbreaks is believed to increase the likelihood of fire (McCullough et al. 1998, Jenkins et al. 2008), though this may need to coincide with appropriate climatic conditions and/or occur within a few years of the outbreak in order to be realized (Bebi et al. 2003, Kulakowski et al. 2003). Conversely, fire disturbances dictate stand susceptibility to future insect disturbance. Surviving trees that are scorched or otherwise wounded in low- to moderate-severity fires are less able to successfully resist insect attack (McCullough et al. 1998). Young post-fire stands, created by stand-replacing fires, tend to be less severely affected by subsequent insect outbreaks than those not burned due to the absence of preferentially-attacked larger hosts (Bebi et al. 2003, Kulakowski et al. 2003). Upon maturation, however, stands of homogeneous

species and age are more likely to promote a phytophagous insect population eruption than heterogeneous stands (Raffa et al. 2008).

These synergistic effects further limit the capacity to predict the effects of climate on treelines.

After all, forecasting the effects of large-scale, relatively infrequent disturbances can be difficult because of the complexity of factors which dictate subsequent ecological processes (Bebi et al. 2003).

Furthermore, the complex interactions between climate, disturbance and recruitment may not remain the same should temperature warming predictions hold true (Holtmeier and Broll 2005).

1.4 Dendrochronology

Dendrochronology, which studies the annual growth rings of trees, is a particularly useful method for evaluating tree response to climatic changes and disturbance events. As outlined by Stokes and

Smiley (1968), there are four conditions which must be met in order to successfully conduct dendrochronolgical analysis: (i) trees used must add only one ring per growth season, (ii) there is only one dominant environmental factor limiting growth, (iii) this factor must vary in intensity from year to year and cause subsequent variation in ring-widths, and (iv) the growth-limiting environmental factor must be dominant over a large geographical area. Still, even trees which normally produce annual rings may fail to do so in years of high stress. In some years, radial growth may be so limited that ring production is constrained to certain radii or does not occur at all (Stokes and Smiley 1968, Fritts 1976).

Should a period of stress occur within a normal growing season, a false ring may be produced through the formation of mid-ring cells which resemble those that indicate the termination of a growing season

(Stokes and Smiley 1968, Fritts 1976). For this reason, cross-dating is necessary to minimize the likelihood of ring-dating error due to the presence of missing, locally absent, or false rings (Speer 2010).

Once tree rings are assigned to a specific year, they can be calibrated with available climatic data.

To maximize the variability of annual ring-widths, trees experiencing greater sensitivity to climatic

factors are selected for this type of analysis (Fritts 1976), known as dendroclimatology. This makes trees at the margins of their range, such as those at latitudinal or altitudinal treelines, ideally suited for study given the temperature limitations on their growth (Speer 2010). By calibrating dated tree rings with climate data, it is possible to correlate the growth of trees to temperature and precipitation patterns.

For example, high-elevation and high-latitude white spruce in northwestern Canada and Alaska have been demonstrated via dendrochronological analysis to exhibit differential responses to increased summer temperatures. Inconsistencies in treeline response to warming have been linked to shifting phases of the Pacific Decadal Oscillation (PDO) (Ohse et al. 2012, Chavardès et al. 2013), moisture stress

(Jacoby and D’Arrigo 1995, Lloyd and Fastie 2002, Wilmking et al. 2005, Porter and Pisaric 2011), and tree age (Szeicz and MacDonald 1994, Mamet and Kershaw 2013), among others.

Given the assumption that past limiting conditions affected growth patterns in the same manner as in the present (Fritts 1976), tree rings can also be used to reconstruct past climate and extend the climatic record backwards in time beyond those of instrumental records. These reconstructions (e.g.

Jacoby and D’Arrigo 1995, Mann et al. 1999, Briffa et al. 2001, Jones et al. 2001) have been used to evaluate past climate variability and provide evidence that 20th century temperatures in the Northern

Hemisphere, including those of southwest Yukon (Youngblut and Luckman 2008), were warmer than at any point in the past millennium. Paleoclimatic records are particularly valuable in regions with few climate stations and/or a limited historical instrumental record, such as the Canadian arctic (Prowse et al. 2009).

Disturbance events, such as fire and insect infestation, also affect the growth of tree rings.

Dendrochronology can therefore be used to interpret these records and identify past disturbances. In the case of stand-replacing and other low-severity fires, fire-scarred trees are sampled to determine year of occurrence (Kulakowski et al. 2003, Drury and Grissom 2008, Axelson et al. 2010). An alternative method of establishing fire dates is to determine the death year of fire-killed trees (Drury and Grissom 20

2008). Past outbreaks of defoliating insects can be detected in ring-width chronologies as a period of growth suppression in the host tree (Kulman 1971). Bark beetle infestations, on the other hand, rarely result in growth reductions (cf. Rolland and Lempérière 2004) as host trees are typically killed during attack (Veblen et al. 1991b). Rather, past bark beetle outbreaks are most commonly detected by identifying sustained, accelerated radial growth in understory and canopy-dominant host and non-host tree species following the mortality of dominant overstory host trees (e.g. Romme et al. 1986, Veblen et al. 1991a, Parish et al. 1999, Zhang et al. 1999, Eisenhart and Veblen 2000, Kulakowski et al. 2003, Alfaro et al. 2004, Berg et al. 2006, Axelson et al. 2010, Sherriff et al. 2011, Smith et al. 2012, Amoroso et al.

2013). Veblen et al. (1991a, 1991b) observed that these growth releases can be sustained for more than

40 years. That said, releases triggered by a regional outbreak may not be precisely simultaneous due to slightly differential timing of attack and host mortality (Veblen et al. 1991b, Eisenhart and Veblen 2000).

Assigning a date of death to deceased host trees may further assist in the dating of insect outbreaks, and may be critical to identifying less intense outbreaks (Eisenhart and Veblen 2000). Another method of detecting historic bark beetle infestations is the recording of crescent-shaped resin accumulations

(Rolland and Lempérière 2004), although Caccianiga et al. (2008) caution that these may also be caused by other stressors. Finally, analyzing ring-width chronologies from non-host species and/or control areas can assist in determining whether a growth suppression or release was caused by climatic factors or disturbance events (Eckstein et al. 1991, Zhang et al. 1999, Alfaro et al. 2004, Sherriff et al. 2011).

Through the use of dendrochronology, the dating of past disturbances allows for the reconstruction of past disturbance regimes. In addition, age-structure analysis can be used to infer historical stand structure. Paired with growth and composition analysis of stands following a known disturbance, this provides insight into the impacts of different disturbances and how climate may alter the effects, intensities, and frequencies of disturbance events. For instance, Berg et al. (2006) demonstrated that spruce bark beetle outbreaks have occurred rarely in southwest Yukon over the last 21

250 years, whereas stands on the Kenai Peninsula of Alaska have experienced outbreaks at a mean return interval of 52 years. However, Sherriff et al. (2011) identified the most recent spruce bark beetle infestations on the Kenai Peninsula as being higher in severity than those of the past. As such, the recent climate-driven outbreaks in these regions point to a divergence away from historic disturbance patterns.

1.5 Research Objectives

Believed to have begun in 1990, spruce bark beetle (Dendroctonus rufipennis) populations reached epidemic levels throughout southwest Yukon at the turn of the last century (Government of

Yukon 2013), including the Kluane Lake region where our research was conducted. Affecting approximately 400,000 hectares to date, levels of spruce bark beetle kill peaked in 1998 and again in

2004, but have declined greatly since 2006 (Garbutt et al. 2006, Government of Yukon 2013). In several areas, white spruce (Picea glauca) mortality has extended as high as the altitudinal treeline. In addition, the region has undergone a significant increase of 2°C in mean annual temperature during the last 50 years (Prowse et al. 2009). The spruce bark beetle outbreak is believed to have been caused at least in part by prolonged warm temperatures, which have a positive influence on insect overwintering survival and voltinism, and often induce moisture stress in mature host trees (Berg et al. 2006).

The aim of this study was to assess the impact of insect disturbance on the growth and establishment of trees in a subarctic treeline environment, within the context of a warming climate, as well as to gain a better understanding of factors that may predispose treelines to insect disturbance.

Specific objectives were: (i) to determine whether stand structure differed with variation in spruce bark beetle mortality, and, (ii) through the use of dendrochronology, to assess and evaluate differential growth patterns of trees at sites affected and relatively unaffected by spruce bark beetle, and to relate these to available climate records. Each of these objectives is the focus of Chapter 2 and 3, respectively.

With high elevations and latitudes anticipated to continue to be affected by climate warming more than other regions (Symon et al. 2005, Nogués-Bravo et al. 2007, Rebetez and Reinhard 2007,

Stocker et al. 2013), the uncertainties regarding the impacts of increased temperatures on forest-tundra ecotones will become increasingly important to address. In addition, very little is known about how insect disturbance functions at treeline and what influence it has on this ecotone. An improved understanding of the interactions between climate-driven effects and insect disturbance on altitudinal forest-tundra ecotones is therefore essential for assisting in the prediction of future vegetation changes and disturbances, as well as informing the development of appropriate management strategies.

Figure 1-1 Visual representation of an alpine forest-tundra ecotone. Defined as the transition zone between closed-canopy forest and the limits of tree distribution, the forest-tundra ecotone is typically characterized by a gradual decline in tree size and density. At forestline (sometimes referred to as timberline), trees are still able to achieve large sizes but grow in open-canopy stands. Treeline is characterized by discontinuous groupings of trees, which are often unable to achieve the heights of lower elevation individuals; the term ‘treeline’ is also used by some authors to refer to the entire forest- tundra ecotone. At krummholz line, trees are rarely present and only occur in stunted, prostrate growth forms.

Chapter 2

Effects of Differential Spruce Bark Beetle Mortality across Forest-Tundra

Ecotones in Southwest Yukon, Canada

2.1 Introduction

During the latter half of the 20th century, regions north of 60° experienced substantially more climate warming than lower latitudes (Symon et al. 2005, Chapin et al. 2005, Prowse et al. 2009, Stocker et al. 2013). Alpine ecosystems are also acutely susceptible to climate warming (Stocker et al. 2013), and are anticipated to undergo two to three times more warming than experienced over the last one hundred years (Nogués-Bravo et al. 2007, Rebetez and Reinhard 2007). With temperatures in northern latitudes predicted to increase by 2° to 7°C by 2100 (Symon et al. 2005), this makes subarctic alpine regions particularly vulnerable to the effects of climate change.

Often referred to as treeline, the forest-tundra ecotone is the transition zone between closed- canopy forest and the upper altitudinal or latitudinal limits of tree growth. While numerous mechanisms have been hypothesized to limit tree establishment and growth at treeline (see Stevens and Fox 1991, Körner 1998, Sveinbjörnsson 2000, Grace et al. 2002, Holtmeier and Broll 2005, Moen et al. 2008), temperature is widely believed to be the most critical regulator given its role in dictating both the length of the growing season and the success of heat-dependent ecological processes (Tranquillini

1979, Körner 1998, Körner and Paulsen 2004). As such, warming temperatures are expected to have pronounced effects on forest-tundra ecotones.

North American and Eurasian treelines have been studied extensively to investigate vegetative responses to global temperature increases following the termination of the Little Ice Age (circa 1850).

Prolonged periods of warming have caused stunted, prostrate tree individuals (i.e. krummholz) to achieve erect growth forms (Lescop-Sinclair and Payette 1995, Gamache and Payette 2004, Devi et al. 25

2008), and have facilitated increased seedling establishment, growth rates and survivorship (Szeicz and

MacDonald 1995, Lloyd and Fastie 2003, Danby and Hik 2007b). Radial growth increases in mature individuals have also been observed (Suarez et al. 1999, Vittoz et al. 2008, Salzer et al. 2009). Some treelines have exhibited stand densification (Payette and Filion 1985, Scott et al. 1987, Szeicz and

MacDonald 1995, MacDonald et al. 1998), while others have experienced the altitudinal and latitudinal advance of trees into areas which they did not previously occupy (Lescop-Sinclair and Payette 1995,

Suarez et al. 1999, Soja et al. 2007, Devi et al. 2008). The variability of responses to warming among forest-tundra ecotones reflect site-specific conditions which may limit growth and/or advance despite increased temperatures, such as species physiognomy, local topography, and disturbance history

(Holtmeier and Broll 2005).

Many of the phytophagous insects which affect forests are endemic and evolved over millennia alongside their hosts (Furniss and Carolin 1977). However, changes to the Earth’s climate may modify these relationships not only by altering the distribution and composition of tree species, but also by influencing aspects of insect life history. In western North America, for example, the unprecedented magnitude and severity of recent bark beetle outbreaks in British Columbia (Dendroctonus ponderosae

Hopkins), Alaska and Yukon (D. rufipennis Kirby) are believed to have been driven at least in part by climate change (Ayres and Lombardero 2000, Safranyik and Wilson 2006, Berg et al. 2006, Werner et al.

2006, Logan and Powell 2009, Sherriff et al. 2011, Weed et al. 2013). While an abundance of susceptible hosts is a prerequisite for a bark beetle epidemic to occur (Raffa et al. 2008), suitable climate is just as important a condition. In particular, increased temperatures can trigger shifts in beetle generation from semivoltine (2-year maturation) to univoltine (1-year maturation) broods (Schmid and Frye 1977,

Werner and Holsten 1985, Berg et al. 2006) and reduce cold-induced winter mortality (Schmid and Frye

1977, Ayres and Lombardero 2000, Safranyik and Wilson 2006, Bentz et al. 2010). Temperature-induced

drought may further contribute to outbreak intensity by reducing host vigour and defense mechanisms

(Berg et al. 2006), allowing a smaller number of bark beetles to overwhelm host defenses.

In addition to promoting increased population levels within their historic ranges, temperature warming can facilitate an expansion of bark beetles into latitudinal and altitudinal areas previously considered climatically unfavourable (Logan and Powell 2001, Carroll et al. 2004, Marini et al. 2012,

Weed et al. 2013). Trees growing at the margins of their distribution, such as those at subarctic alpine treelines, may therefore become more vulnerable to insect disturbance as climate continues to warm

(Ayres and Lombardero 2000, Bentz et al. 2010, Marini et al. 2012, Murdock et al. 2013) and, being already stressed by abiotic factors, recover slowly from an outbreak (Raffa et al. 2008, Moen et al.

2008). However, while insect disturbance can extend into the forest-tundra ecotone, there have been relatively few studies documenting the effects of insect outbreaks on treelines. The principal outcome of insect disturbance at treeline appears to be tree mortality (Caccianiga et al. 2008, Van Bogaert et al.

2011), but little else is known. The influence of insect disturbance on treelines, within the context of a warming climate, must therefore be addressed to assist in the prediction of future vegetation changes in this ecotone.

The aim of this study was to examine the varied effects of a recent spruce bark beetle (D. rufipennis) infestation on forest-tundra ecotones in southwest Yukon, Canada, with a view to better understanding how insect disturbance might affect future forest-tundra dynamics. Specific research questions were: (i) what, if any, aspects of ecotone structure have been affected by spruce bark beetle mortality? (ii) are there characteristics of trees or tree stands that make them more susceptible to spruce bark beetle attack? and (iii) do these effects vary with elevational position within the forest- tundra ecotone? Vegetation surveys were used to collect information within stands of white spruce

(Picea glauca (Moench) Voss) across the forest-tundra ecotone, and the data were used to examine relationships between stand structure and beetle-caused mortality. 27

2.2 Methods

2.2.1 Study Area

Field work was conducted in the region south of Kluane Lake in southwest Yukon, during the summer of 2012. The study area runs northwest to southeast in parallel with the Shakwak Trench, a large valley that separates the Kluane Ranges of the St. Elias Mountains from the rounded mountain ranges of the Yukon Plateau. Six sites were selected for sampling, each located within the subalpine forest zone (1080-1370masl) (sensu Douglas 1974) of the Kluane Wildlife Sanctuary, adjacent to the

Kluane National Park and Reserve.

Located on the leeward side of the St. Elias Mountains, this region receives little of the precipitation carried inland on prevailing winds from the Pacific Ocean and is considered semiarid

(Scudder 1997, Krebs and Boonstra 2001). With alpine glaciers and summits rising to over 2130masl

(Scudder 1997), the steep environmental gradients of the region create localized climate conditions that cannot be accurately captured by any weather station. That said, the 30-year climate normals available for the nearest weather station at Burwash Landing (61°22’N 139°03’W, 806.2masl) still provide a reasonable approximation of climate for the study area. Mean annual temperature is -3.2°C, with mean monthly temperatures of -20.5°C and 13.1°C for January and July, respectively; mean annual precipitation is 274.7mm, with average annual snowfall totalling 105.5cm (Environment Canada 2014).

The dominant tree species in this region is white spruce (Picea glauca), which occupies the subalpine forest zone primarily as open-canopy stands. In a gradual transition, spruce canopy cover becomes more open and discontinuous with increasing elevation, finally terminating with the presence of very few spruce which grow in stunted, prostrate ‘krummholz’ formations. No secondary tree species were observed at these sites, although balsam poplar (Populus balsamifera L.) and trembling aspen

(Populus tremuloides Michx.) stands do occur in the lower elevation montane valley bottom forests of

the Shakwak Trench (Krebs and Boonstra 2001). American dwarf birch (Betula glandulosa Michx.) and various willow (Salix spp.) are the two dominant deciduous shrubs and grow throughout the ecotone.

Salix glauca (L.) is the predominant species of willow, but S. pulchra (Cham.), S. richardsonii (Hook.), and

S. arbusculoides (Andersson) are also present in various amounts. These shrubs often form dense layers over 1m in height, referred to here as shrub-fields, particularly at higher elevations.

Between 1990 and 2006, epidemic levels of spruce bark beetle (Dendroctonus rufipennis) affected approximately 400,000 hectares of forest in southwest Yukon (Garbutt et al. 2006, Government of

Yukon 2013), including the subalpine forests and forest-tundra ecotones of the Kluane Ranges (Figure

2-1). With the exception of a relatively small outbreak occurring between 1934 and 1942, the Kluane region is not believed to have experienced any prior extensive spruce bark beetle outbreaks (Berg et al.

2006). Prolonged periods of warmer temperatures in the region during the latter half of the 20th century (Prowse et al. 2009) are believed to have contributed to the recent epidemic by inducing moisture stress in mature host trees, increasing beetle overwintering survival, and doubling of the insects’ maturation rate, or voltinism (Berg et al. 2006).

2.2.2 Field and Laboratory Methods

Satellite imagery available on GoogleEarth was used to identify possible areas of study along the

Shakwak Trench between Kluane Lake and Haines Junction. Potential sampling sites were identified during a fixed-wing flyover, recorded on a map along with GPS coordinates, and considered for selection if they were adequately accessible by vehicle or on foot. Final site selection was refined in the field, avoiding geographical features such as creeks and valleys while being as representative of the surrounding area as possible. In this way, the influence of microsite variables was minimized. In total, six subalpine sites of varied spruce bark beetle mortality were selected (Figure 2-2), with two each of low (<20%), medium (20%–40%) and high (≥40%) tree mortality; hereafter, each of these site-level

mortality levels will be referred to as ‘mortality class’. All sites were at least 1km apart from each other, and were similar in terms of elevation, aspect and slope.

An elevational gradient extending from closed boreal forest to open tundra was identified at each site, with three 100m perpendicular transects established at: (i) ‘forestline’ (~1,100masl), defined as the transition between open-canopy forest and discontinuous forest with non-overlapping tree crowns; (ii)

‘treeline’ (~1,1175masl), defined as the transition between the aforementioned discontinuous forest and higher elevation shrub-fields; and, (iii) ‘krummholz line’ (~1,250masl), or tree species limit, where spruce is rarely present and only occurs in stunted or krummholz growth forms. Three sampling points, each 50m apart, were established along each transect (Figure 2-3).

Circular sampling plots of varying area were established at all three points along each transect.

Plots extended outwards to include the 12 spruce individuals greater than 0.5m in height nearest to the centre point. This sample size provided sufficient replication for the purpose of describing stand characteristics. A qualitative assessment of health, as well as distance from the centre point and respective orientation, was noted for each individual. Individuals were classified as saplings (0.5m–1.9m in height) or trees (≥2.0m in height) (Danby and Hik 2007c, Hofgaard and Rees 2008). A clinometer was used to determine the height of individuals that could not be measured with a tape measure. The basal diameter at root crown (DRC) was recorded for each individual, and the diameter at breast height (DBH)

(1.3m above ground) was measured when applicable. If an individual was part of a clonal colony, the number of stems (ramets) and their respective height classes were also recorded. The presence of male strobili was noted, and the abundance of female strobili was estimated using a four-point scale: low

(<10), medium (10–50), high (51–100), and very high (>100) (adapted from Hofgaard and Rees 2008).

Strobili were differentiated between those from the current year and those from prior years.

Data on shrub cover was obtained near the middle sampling point of each transect (Figure 2-3). A

10m x 10m point-intercept grid was established and the presence of live shrubs, as well as their respective species and height, was recorded at 1m intervals. These data were used to yield measures of shrub dominance (proportion of shrubs that were willow (Salix spp.)), percentage shrub cover, and mean shrub height for each elevation.

Information on spruce seedlings (individuals <0.5m in height) was also collected. A 100m x 2m belt transect was established between the two outermost sampling points of each transect (Figure 2-3).

When seedlings were present, their height, basal diameter, and position along the transect were recorded. To maximize the length of record available for age-growth analysis, thirteen of the tallest

(≥0.2m) seedlings were randomly selected for destructive sampling, and were uprooted and labeled for later analysis. A maximum of three seedlings were selected per transect to avoid compromising the regenerative potential of the post-outbreak landscape.

In the laboratory, each of the 13 seedlings collected in the field were cut at root crown to obtain a cross-sectional disk. Seedlings >0.25m in height (n=11) were also cross-sectioned at 0.25m above root crown; this falls within the range of heights above ground (0.20m–0.50m) at which mature trees are typically sampled to determine age (Gutsell and Johnson 2002). These cross-sections were immersed in water for a period of 48 hours and subsequently thin-sectioned (30µm) using a Leica SM2010R sliding microtome. Thin-sections were mounted on glass slides, and annual growth rings of each seedling were counted at both sampling heights (if applicable) using a microscope and transmitted light. Thin- sectioning has been demonstrated to provide more accurate seedling age estimates than the traditional method of using wood disks and reflected light, particularly in environments such as treeline where seedling growth may be supressed by climate and/or competition (Daniels et al. 2007, Cairns et al.

2012).

2.2.3 Data Analysis

Data analysis was grouped into the following themes: (i) ecotone structure; (ii) spruce reproduction; (iii) seedling growth; and (iv) susceptibility to spruce bark beetle. Non-parametric tests were used because data obtained from both the sampled seedlings and the vegetation surveys did not consistently conform to the assumptions of parametric statistics and data transformation could not correct this. Medians (with range) are therefore more representative descriptors of our data distributions (Sokal and Rohlf 2009) and are presented here rather than means, unless specified otherwise. Sample size was not large enough to test for significant interactions between elevation and mortality class, but data were graphed to visually inspect for possible interactions and we report these whenever present. All statistical analyses were performed using SPSS 22.0 (IBM, Chicago, Illinois, USA).

Ecotone Structure

Analysis of data was conducted with a view to determining whether ecotone structure was (i) differentially affected by varied levels of spruce bark beetle mortality, and (ii) if these effects were dependent on elevation within the forest-tundra ecotone. A Kruskal-Wallis test was used to determine if ecotone characteristics differed significantly between elevations (krummholz line, treeline and forestline). Sample size varied depending on the characteristic being analyzed, due to the intrinsic variance of the observational units: (i) seedling density, shrub cover, shrub dominance, and mean shrub height were calculated for each elevational transect; (ii) percentage saplings (proportion of individuals that were <2m in height), density of living spruce, and basal area of living spruce were calculated for each plot; and (iii) spruce height, basal diameter and diameter at breast height were analyzed with the individual tree as the unit of observation. In cases where significant differences were found, a subsequent pairwise comparison was conducted using a Mann-Whitney U test (with the Bonferroni correction) to determine which specific elevations differed from one another.

Because the spruce bark beetle mortality classes used in site selection were assigned categorically based on calculations of site-level spruce mortality, they did not necessarily capture the effects of finer variations in levels of spruce bark beetle kill. As such, we used Spearman’s rank correlation coefficient

(ρ) to test whether ecotone characteristics were influenced by the percentage of trees killed by spruce bark beetle. If an ecotone characteristic was found to exhibit differences with elevation (via the Kruskal-

Wallis tests), a Spearman’s rank correlation coefficient was also calculated for each respective elevation.

In this way, we were able to examine interactions between elevation and percentage spruce bark beetle mortality.

Finally, we wished to determine whether the overstory stand structure influenced the growth and distribution of shrubs and/or saplings. Spearman’s rank correlation coefficients were therefore calculated for the relationships between (i) spruce density, basal area and height (i.e. measures of canopy structure), (ii) shrub cover and mean shrub height, and (iii) seedling density.

Spruce Reproduction

We compared cone production of spruce trees (≥2m in height) among plots to determine whether spruce reproduction was affected by the recent spruce bark beetle outbreak or varied with elevation.

While the presence of old cones had been recorded, we could not differentiate between cones produced in 2011 and those of the 2010 mast year (Krebs et al. 2012, Archibald et al. 2012). As such, we only conducted analyses on current-year (2012) reproduction levels. Spearman’s rank correlation coefficient was used to determine whether the presence of male and female cones and/or levels of female cone production (low, medium, high, very high) were related to percentage mortality of a stand.

A Mann-Whitney U test was used to determine whether these variables were dependent on elevational placement within the forest-tundra ecotone.

Seedling Growth

Few studies have analyzed growth of white spruce seedlings in treeline environments, and none have examined it at treelines affected by bark beetle disturbance. Those studies that do exist indicate that small individuals at treeline – ostensibly seedlings – can actually be several decades old (e.g. Danby and Hik 2007b), meaning that the growth of these individuals is exceedingly slow. Spearman’s rank correlation coefficient was used to determine whether seedling age was related to basal diameter, height or mean radial growth rate. A Mann-Whitney U test was used to test the null hypothesis that seedling age, basal diameter, height and mean radial growth did not differ with elevation (i.e. treeline and forestline). Similarly, a Kruskal-Wallis test was used to test the null hypothesis that these aforementioned variables did not differ between mortality class (low, medium, high). Since mortality classes were assigned from percentage of susceptible trees killed by spruce bark beetle over the entire site (i.e. 3 elevations), we also used Spearman’s rank correlation coefficient to test whether any of the variables had relationships with transect-specific percentage kill.

Susceptibility to Spruce Bark Beetle

We conducted three separate statistical analyses to examine whether particular tree or plot-level characteristics increased susceptibility to spruce bark beetle attack. First, we used Spearman’s rank correlation coefficient to investigate whether plots with greater pre-outbreak stand density or basal area experienced higher levels of spruce bark beetle mortality. Secondly, size metrics from living

(n=322) and beetle-killed (n=101) spruce were compared using a Mann-Whitney U test to determine whether larger trees were more likely to be killed. Finally, we performed binomial logistic regressions to ascertain the effects of height, DRC and DBH on the likelihood an individual would succumb to spruce bark beetle attack. Separate regressions were run on each size metric to avoid violating test assumptions by including interrelated and highly correlated variables. To ensure that this test reflected 34

the metrics at time of attack as accurately as possible, the DRC and DBH measurements for each living individual were estimated for 1997, the median year of recorded mortality at these sites. This was achieved by using the 1997-2012 average basal and breast height radial growth rates of a subset of individuals which were sampled along each transect for analysis of annual growth rings (see Chapter 3).

Metrics for spruce killed by bark beetle were not adjusted, as they reflect the exact dimensions of the individuals at time of death. All analyses were conducted on the entirety of each sample set, regardless of elevational placement; results did not differ substantially when separated by elevation, and are therefore not presented here.

2.3 Results

2.3.1 Ecotone Structure

Stand characteristics varied significantly with elevation along the forest-tundra ecotone. Spruce were generally larger and grew more densely in lower elevation stands (Table 2-1). Individuals growing at treeline were significantly shorter and had smaller DBH than those at forestline (p<0.001); there was no significant difference in DRC between elevations. Density and basal area of living spruce trees and saplings were higher at forestline than at treeline (p=0.001 and p=0.009 for density and basal area, respectively). Seedling density also differed between elevations; post-hoc tests revealed that seedling density was significantly higher at forestline compared to krummholz line (p=0.004), but there was no significant difference between krummholz line and treeline (p=0.921) or between treeline and forestline

(p=0.081).

Analysis of the data using spruce bark beetle-induced mortality as a continuous variable revealed that all spruce size metrics were highly positively correlated with percentage of spruce bark beetle kill throughout the forest-tundra ecotone (p<0.001). Seedling density was also positively correlated with percentage mortality (p=0.045). Across the forest-tundra ecotone, stand density and basal area of living

spruce were negatively correlated with percentage mortality, albeit not significantly so (p=0.868 and p=0.343, respectively). That said, these correlations become statistically significant (p=0.011 for basal area), or nearly so (p=0.065 for stand density), when considering only forestline stands.

The mean height of shrubs also had a positive correlation with percentage mortality throughout the forest-tundra ecotone (p=0.043). However, we cannot attribute differences in shrub height to the spruce bark beetle outbreak because shrub height is also arguably related to elevation (p=0.071).

Similarly, the elevational differences of other stand characteristics (see above) prevented the analysis of the relationships between canopy dynamics, shrubs and seedlings.

A summary of site-specific ecotone characteristics can be found in Appendix A; more detailed results from the statistical analyses can found in Appendix B.

2.3.2 Spruce Reproduction

Of the 233 spruce individuals ≥2m in height that survived the recent spruce bark beetle outbreak,

34.3% produced female cones and 15.9% produced male cones in 2012. There was a significant negative correlation between stand mortality and percentage of spruce bark beetle-susceptible spruce that produced female cones in 2012 (p=0.048). The number of trees which produced cones did not depend on elevational placement within the forest-tundra ecotone (p=0.239 and p=0.308 for elevational analysis of female and male cone production, respectively; Figure 2-4). However, the proportion of trees with low female production (<10 cones) was significantly greater at treeline as opposed to forestline

(p=0.047; Figure 2-4). No other significant results were found.

2.3.3 Seedling Growth

Seedling ages ranged from 16 to 52 years with a median of 32 years (Figure 2-5). Basal diameters ranged from 0.3 to 1.3cm (median 0.7cm), heights ranged from 19.5 to 48.0cm, and mean radial growth rate ranged from 0.88 to 1.88mm·yr-1 (median 1.11mm·yr-1) (Figure 2-5). Seedlings took a median of 26

years to grow 25cm in height, though this varied between 11 and 33 years. Seedling age was positively correlated with both basal diameter (Figure 2-5b) and height (Figure 2-5a). However, mean radial growth rates were poorly correlated with seedling age (Figure 2-5c), which suggests that age is not a significant contributor to variation in annual ring widths of seedlings growing within the forest-tundra ecotone. Neither seedling age, height, basal diameter nor radial growth differed significantly between elevation, mortality class or percentage of spruce bark beetle kill.

2.3.4 Susceptibility to Spruce Bark Beetle

Individuals killed by spruce bark beetle were significantly larger than those that survived the outbreak in terms of height, basal diameter, and diameter at breast height (Figure 2-6). Results from our logistic regression models indicated that, individually, the predictors of spruce height, basal diameter, and diameter at breast height reliably distinguished between spruce individuals which succumbed to spruce bark beetle attack and those that did not (Table 2-2). The best predictive model was for DBH, which explained 63.5% (Nagelkerke’s R2) of the variance in spruce mortality, correctly predicting 86.3% of cases. The DRC model was able to explain 58.9% (Nagelkerke’s R2) of the variability in the dependent data and correctly classified 85.5% of cases. The height model was less accurate, explaining only 49.3% (Nagelkerke’s R2) of spruce mortality variance and correctly classifying 83.0% of cases. All three of these models correctly predicted 91.9% for survival. Correct predictions for mortality differed, however, with 71.3% for DBH, 65.3% for DRC, and 54.5% for height. High values of spruce height, DRC and DBH all indicated a higher probability of tree mortality. Our results indicated that a one-metre increase in spruce height will result in a 53.2% increase in the odds of succumbing to spruce bark beetle, with a one-centimetre increase in DBH and DRC resulting in a 19.4% and 13.3% increase, respectively (Table 2-2).

Certain tree stand characteristics were also found to increase susceptibility to spruce bark beetle attack. Specifically, pre-outbreak stand basal area and density were positively correlated with percentage mortality (Figure 2-7). Correlations were strongest for stand basal area, and were more significant for stand density when considering only the density of spruce individuals classified as trees

(ρ=0.396, p=0.017) as opposed to that of all spruce individuals within a stand (ρ=0.366, p=0.028).

2.4 Discussion

Following a major spruce bark beetle outbreak in southwest Yukon, we examined sites in forest- tundra ecotones that experienced low, medium and high mortality levels. As anticipated, given that we sampled across the entire ecotone, we found altitudinal differences in stand characteristics. However, we also detected differences in pre-outbreak stand characteristics between stands with minimal, moderate and extensive mortality. The impacts of insect disturbance were most clearly pronounced at, though not exclusive to, lower elevations. These findings demonstrate that spruce bark beetle is capable of causing the mortality of individuals within forest-tundra ecotones, including treeline elevations, and suggest that the host selection mechanisms are similar to those in low-elevation, closed- canopy stands. By dissimilarly affecting stands at different locations and different elevations, spruce bark beetle disturbance will likely contribute to and/or reinforce heterogeneity in forest stand structure and treeline elevation in the Kluane Ranges.

2.4.1 Extent of Spruce Bark Beetle Mortality

Intermittent spruce bark beetle outbreaks have the capacity to kill the majority of mature trees across extensive areas of low-elevation forest (Schmid and Frye 1977). We found this to generally hold true in the subarctic alpine forest-tundra ecotones of southwest Yukon, but also identified strong heterogeneity in beetle-induced mortality levels across the ecotone. Stands sampled at Mount Decoeli experienced minimal spruce bark beetle kill, with a maximum mortality of 16.7% at treeline and 40.0%

at forestline, and several stands being unaffected at both elevations. Treeline stands at Telluride Creek were unaffected by the outbreak, whereas forestline stands generally experienced high levels of mortality (between 20.0% and 91.7%). At Boutellier Creek, stands were greatly affected by spruce bark beetle at both treeline and forestline elevations, although these ranged from 0.0% to 77.8% at treeline and from 25.0% to 71.4% at forestline.

To compare levels of mortality within the forest-tundra ecotone with those in montane valley bottom forests, we examined data gathered from 27 forest locations in the Shakwak Trench and Alsek

River drainage that were affected by the same infestation (Garbutt et al. 2006). In their report, Garbutt et al. (2006) presented stand mortality in terms of volume (m3/ha) killed by spruce bark beetle. We used the whole stem cubic volume equation provided by Garbutt et al. (2006) to assess volume for each sampled spruce, then calculated the volume killed by spruce bark beetle as a percentage of all spruce for each of the two data sets. The median percentage of volume killed at the 27 low-elevation plots reported by Garbutt et al. (2006) was 52.1%, and ranged from 2.4% to 86.9%. Our study sites had a median of 71.7% volume killed across the forest-tundra ecotone, ranging from 0.0% to 85.7% at treeline elevations (median 8.8%) and from 14.8% to 93.7% at forestline elevations (median 72.9%).

The results of this comparison demonstrate that subalpine spruce bark beetle incidence was sporadic throughout southwest Yukon, regardless of elevation. Range of mortality extent was similar between low-elevation forests and sites within the forest-tundra ecotone, which suggests that higher elevations are equally as susceptible to spruce bark beetle attack as valley bottom forests. However, while capable of experiencing high mortality levels, treeline stands generally had fewer trees killed than forestline or valley bottom forests. It therefore stands to reason that the characteristics of treeline individuals and/or stands were less desirable to spruce bark beetle.

It is difficult to compare the southwest Yukon outbreak to spruce bark beetle events in other regions because of inconsistencies in measurement of outbreak intensity across studies. The majority of 39

authors have reported outbreak severity in terms of estimated tree mortality, as we have done. No standard exists, however, on the size of individuals to include in this tally. For example, post-outbreak mortality estimates were calculated for trees ≥20cm DBH in the subalpine forests of northwestern

Colorado (Hinds et al. 1965, Schmid and Hinds 1974). In contrast, the extent of spruce mortality in south-central and southwest Alaska was assessed for trees with breast height diameters of >12cm

(Sherriff et al. 2011) and >12.7cm (Boucher and Mead 2006), among others (Werner et al. 2006). Our usage of different mortality measurements, paired with our targeted rather than random sampling design, prevents the direct comparison of outbreaks. However, tree mortality generally exhibits variability across all studied landscapes, as well as a preference for trees and stands with particular characteristics.

2.4.2 Susceptibility to Spruce Bark Beetle Attack

Consistent with research conducted on spruce bark beetle outbreaks in other subalpine and montane valley bottom forests (Massey and Wygant 1954, Schmid and Frye 1977, Hard 1985, Doak

2004, Hart et al. 2014), we found that large spruce individuals were more likely to be killed during an infestation in forest-tundra ecotones. Our logistic regression models identified DBH as being the best predictor of tree susceptibility at both treeline and forestline, though all size metrics contributed to host selection. Individuals growing at treeline were typically smaller than those at forestline, which suggests that treeline was less favourable an elevation for spruce bark beetle populations. Still, several small- diameter trees were killed throughout the ecotone, a phenomenon known to occur during severe outbreaks (Veblen et al. 1991a, Werner et al. 2006, Hart et al. 2014). These individuals were likely attacked during the earlier stages of the outbreak, since host selection has been found to be temporally dynamic (DeRose and Long 2012, Hart et al. 2014). If host selection in the forest-tundra ecotone mirrors that of other locations, an abundance of stands with large trees at outbreak initiation may have

favoured the selection of hosts through other means until the supply declined and the remainder of spruce was necessarily selected based on size (DeRose and Long 2012).

Stands comprised of large, densely growing trees experienced high levels of mortality, congruent with the spruce bark beetle risk factors outlined for regions outside the forest-tundra ecotone (Schmid and Frye 1977, Reynolds and Holsten 1994). As such, most forestline stands were greatly affected by the recent outbreak and the majority of treeline stands were exempt. Deviations from this general trend can also be explained by stand characteristics. Forestline stands which experienced minimal mortality (Mount Decoeli) were comprised of smaller, open-grown trees, whereas treeline stands greatly affected by spruce bark beetle (Boutellier Creek) more closely resembled the densities and basal areas of forestline. These examples illustrate that stand characteristics, controlled in part by elevation, dictate susceptibility to spruce bark beetle within the forest-tundra ecotone rather than elevation itself.

Trees growing in dense stands are subject to greater competition for resources, which can decrease host vigour and enable fewer beetles to overcome tree defenses (Wallin and Raffa 2004, Raffa et al. 2005).

Competition may also reduce radial growth rate, which is an important determinant of host selection and survival (Hard 1985, Doak 2004). Furthermore, aggregation pheromones that are produced after successful colonization of a host tree attract additional beetles to adjacent trees (Hard 1989), thereby increasing the probability of infestation for trees in close proximity to already affected individuals (Hart et al. 2014). In this way, stand structure likely influences susceptibility to spruce bark beetle attack more than individual tree characteristics (Schmid and Frye 1977) in forest-tundra ecotones, and may be of particular importance during the beginning stages of an outbreak when beetle populations are still increasing.

While endemic populations of spruce bark beetle have been identified at other subarctic treelines

(Caccianiga et al. 2008), the widespread attack of trees in upper forest-tundra ecotones will only occur if an adequate supply of susceptible hosts is available. Increased stand density and synchronous 41

senescence of potential hosts in treeline environments may trigger spruce bark beetle population eruptions (Caccianiga et al. 2008). However, during large regional outbreaks, such as the disturbance that recently affected southwest Yukon, spruce bark beetles typically originate in valley bottom forests

(Schmid and Frye 1977, Holsten et al. 1999, Garbutt et al. 2006). It is therefore reasonable to assume that spruce bark beetle attack extends upslope as the supply of lower elevation mature spruce is exhausted. As a result, provided that a sufficient number of larger spruce are clustered together, treelines may be predisposed to infestation if they are located near forestline stands hosting large populations of spruce bark beetle.

2.4.3 The Future Structure of the Forest-Tundra Ecotone

We anticipate that stands affected by spruce bark beetle will take several decades to recover, though this progression is not likely to be ubiquitous across the ecotone. In the subarctic alpine treelines of northern Sweden, for example, extensive thinning of mountain birch (Betula pubescens

Ehrh. ssp. czerepanovii (Orlova) Hӓmet-Ahti) remained evident on the landscape over five decades after an outbreak of winter moth (Operophtera brumata L.) (Van Bogaert et al. 2011). Seed germination rates decrease with increasing elevation and latitude (Sirois 2000, Walker et al. 2012), which makes stands in forest-tundra ecotones particularly sensitive to the loss of cone-bearing individuals. In total, only 34.3% of surviving trees ≥2m tall produced female cones in 2012, with the most significant losses in stands that experienced high mortality. This loss of cone production is comparable to the effects of spruce bark beetle on the Kenai Peninsula of Alaska, where post-outbreak cone production was limited to 22% of living trees >1.5m in height (Boggs et al. 2008). Low-density white spruce stands are pollen-limited and have reduced reproductive success (O’Connell et al. 2006). As such, we expect that the combined effects of pollen-limitations and reduced cone production will cause the least successful post-outbreak germination to occur in high-mortality stands.

Contradicting the findings of Veblen et al. (1991a), who found no evidence that post-outbreak establishment was favoured in affected stands, we found seedling abundance to be greater in high- mortality stands. In addition, we were able to estimate that half of the seedlings growing across the forest-tundra ecotones of southwest Yukon had established after 1997 (the median year of recorded mortality), presumably in response to the outbreak. As beetle-killed snags decay and fall, seedling establishment should increase in affected stands. Post-outbreak seedlings in the forests of the Kenai

Peninsula favoured rooting on downed logs, especially those that were heavily decayed (Boggs et al.

2008); we observed a similar preference for decayed, downed timber in our forest-tundra ecotone stands. However, through dendrochronological analysis, we found that deceased trees at our sites can remain standing for over 60 years. Also, established seedlings in the forest-tundra ecotones of southwest Yukon are quite slow growing (median radial growth rate of 1.11 mm·yr-1). The slow progress of decay in standing deceased spruce (Hinds et al. 1965) could therefore limit recovery to stands in which snags have fallen.

Increases in understory vegetation cover following an outbreak may also have negative effects on spruce regeneration. Stands on the Kenai Peninsula, Alaska, have seen increased bluejoint

(Calamagrostis canadensis (Michx.) P. Beauv.) cover restrict spruce regeneration after epidemic levels of spruce bark beetle swept through the region in the 1990s (Boggs et al. 2008), although its impact may be neutral rather than negative in some areas (Boucher and Mead 2006). In southwest Yukon, similar suppressions of spruce regeneration could occur at forestline in response to competition with the taller shrubs that exist at this elevation.

Pre-existing saplings and understory trees will likely experience accelerated growth in response to spruce bark beetle disturbance, as has been nearly universally observed in beetle-affected forests outside of the forest-tundra ecotone (e.g. Romme et al. 1986, Veblen et al. 1991a, Parish et al. 1999,

Zhang et al. 1999, Eisenhart and Veblen 2000, Kulakowski et al. 2003, Alfaro et al. 2004, Berg et al. 2006, 43

Axelson et al. 2010, Sherriff et al. 2011, Smith et al. 2012, Amoroso et al. 2013), though this may be most pronounced in stands with high mortality (see Chapter 3). Because cone production tends to increase with tree size and age (Boggs et al. 2008, Gärtner et al. 2011), we expect that the rapid growth of immature trees will benefit the recovery rate of affected stands by increasing the percentage of cone- bearing trees. Until saplings are able to achieve overstory presence, the canopy of high-mortality locations will remain more closed than treeline elevations but more open than prior to the outbreak.

Stands with high mortality but low numbers of saplings will therefore take the longest to achieve pre- outbreak densities and basal areas, and may be more heterogeneous in terms of age and size than before. Since forestline stands in the region typically experienced the most beetle kill, they have greater risk of increased heterogeneity.

While endemic levels of spruce bark beetle could continue to thin forests as soon as trees reach attack-favoured sizes (Berg et al. 2006), the high levels of spruce mortality throughout southwest Yukon will prevent any major outbreaks from occurring in the near future since epidemic-level populations cannot subsist on an already exhausted supply of timber. However, this recent outbreak may have significantly increased the risk of future fire in southwest Yukon, a region where fire disturbance has historically been less significant than in other parts of the boreal forest (Garbutt et al. 2006) and in which no major fires have been recorded since the early 20th century (Francis 1996). Greater fire intensity and spread rate in post-epidemic stands is presumed to exist (McCullough et al. 1998, Jenkins et al. 2008), although some research has contradicted this (Bebi et al. 2003, Kulakowski et al. 2003).

Deceased, large-diameter trees are thought to greatly increase fuel loading, and an opening of the canopy can increase surface wind speeds and seasonal desiccation of understory fuels (Jenkins et al.

2008). The slow decay rate of deceased spruce in this region could cause fire hazard to remain high for many years (Garbutt et al. 2006). Should a stand-replacing fire occur, poor regeneration of white spruce may alter species dominance and composition (Hogg and Wein 2005), particularly on south-facing slopes 44

(Johnstone et al. 2010b). By homogenizing stand age, stand-replacing fires would favour future spruce bark beetle outbreaks (Raffa et al. 2005, Safranyik and Wilson 2006). Because fire frequency decreases with advancement towards treeline (Payette et al. 2001), stand-replacing fires are less likely to occur at this elevation. Still, major spruce bark beetle infestations may occur at treeline as trees in these stands mature and densify.

The varied extent of spruce bark beetle mortality across the forest-tundra ecotones of southwest

Yukon indicates that this recent disturbance has not caused a ubiquitous treeline recession. Rather, recession has been locally limited to high-mortality sites, such as Boutellier Creek, where substantial thinning occurred at both treeline and forestline; at sites with low and medium mortality, fragmentation was restricted to lower elevations. The treelines in this region are predominantly diffuse (i.e. tree height and density declining with increased elevation) and their positions are thought to be primarily controlled by growing season temperatures (Harsch and Bader 2011). As such, the warming temperatures predicted to occur in this high-latitude, mountainous region (Symon et al. 2005, Nogués-Bravo et al.

2007, Rebetez and Reinhard 2007) should lead to treeline advance (Harsch and Bader 2011). Since stand infilling could preclude upslope advancement (Danby and Hik 2007a), disturbances that thin forest-tundra ecotones could hinder treeline advance. In particular, the effects of insect disturbance would likely delay climate-induced treeline advance (Harsch et al. 2009) in locations where treeline mortality was minimal or non-existent, but may preclude it entirely (Van Bogaert et al. 2011) at high-kill sites where mortality was extensive at treeline. If temperature warming leads to the expansion of bark beetle disturbance into previously unaffected high latitudes and altitudes (Logan and Powell 2001,

Carroll et al. 2004, Marini et al. 2012, Weed et al. 2013), an increased number of forest-tundra ecotone stands could become fragmented and/or recessed.

(a)

(b)

Figure 2-1 Spruce bark beetle-affected trees within forest-tundra ecotones of southwest Yukon. Photos show mortality extending up to treeline (a) on the northeast-facing slopes of the Kluane Ranges at Boutellier Creek and (b) throughout the Clear Creek Valley, located just east of Kluane Lake. Photo (b) courtesy of Katherine Dearborn.

Figure 2-2 Location of Kluane region study sites, southwest Yukon. Study sites were categorized according to levels of spruce bark beetle mortality: the sites at Mount Decoeli (1 and 4), Telluride Creek (3 and 6) and Boutellier Creek (2 and 5) were classified as having low (<20%), medium (20%–40%) and high (≥40%) tree mortality, respectively. Sites 1 and 3 were used for subsequent dendrochronological analysis (Chapter 3).

Figure 2-3 Diagram of sampling technique. Information on spruce seedling height and DRC (light grey rectangles), as well as shrub dominance, height and cover (dark grey rectangles), were collected from each transect along the elevational gradient (krummholz line, treeline, forestline). The health, height, DRC, DBH, and reproductive capacity of 12 spruce individuals were recorded at three equally-spaced circular sampling plots of varying area (black circles) along the two lower-elevation transects.

Figure 2-4 Levels of female cone production in spruce trees (≥2m in height) at treeline and forestline elevations, separated by mortality class. Sites located at Mount Decoeli (MD), Telluride Creek (TC) and Boutellier Creek (BC) were classified as experiencing low, medium and high levels of beetle-induced mortality, respectively.

Figure 2-5 White spruce seedling size metrics and radial growth rates relative to age; correlations are indicated by Spearman’s rank correlation coefficient (ρ). Closed circles represent individuals growing at forestline, whereas those growing at treeline are indicated by open circles; grey circles represent data points shared by treeline and forestline seedlings. Circle size is dependent on the number of individuals with the same value. 50

Figure 2-6 Box plots illustrating the size ranges of sampled white spruce that survived the recent spruce bark beetle outbreak (n=322) in comparison to those that were killed (n=101). The boxes and whiskers denote the 25%–75% and 10%–90% quantiles, respectively. The black line within the boxes represents the median, and outliers are presented as dots.

Figure 2-7 Pre-outbreak spruce basal area and stand density relative to the extent of mortality induced by spruce bark beetle; correlations are indicated by Spearman’s rank correlation coefficient (ρ). Closed circles represent individuals growing at forestline, whereas those growing at treeline are indicated by open circles.

Table 2-1 Post-outbreak attributes (median and range) of forest-tundra ecotone sites with varied spruce bark beetle mortality (low, medium, high) across three elevations (krummholz line, treeline, forestline). In cases where spruce were absent, tree-related data were not collected (indicated by n/a in this table).

Krummholz line Treeline Forestline Low mortality – Mt. Decoeli Seedling density (per ha) 0.0 (0.0 – 0.0) 0.0 (0.0 – 0.0) 100.0 (100.0 – 100.0) Shrub cover (%) 76.4 (62.7 – 90.0) 83.6 (79.1 – 88.2) 73.2 (70.0 – 76.4) Willow dominance (%) 48.6 (41.8 – 55.5) 52.3 (30.9 – 73.6) 66.8 (64.6 – 69.1) Shrub height (m) 59.8 (58.3 – 61.4) 91.4 (55.6 – 127.2) 148.8 (125.8 – 171.9) Spruce height (m) n/a 3.94 (0.6 – 9.9) 3.7 (0.6 – 14.6) Spruce DRC (cm) n/a 14.6 (4.4 – 69.6) 9.8 (0.8 – 55.4) Spruce DBH (cm) n/a 8.6 (0.1 – 32.1) 9.9 (0.8 – 49.0) Saplings (%) n/a 12.5 (0.0 – 41.7) 16.7 (0.0 – 58.3) Stand density (per ha) n/a 77.4 (22.7 – 345.1) 670.1 (239.8 – 823.2) Basal area (m2 per ha) n/a 2.7 (0.4 – 10.8) 10.6 5.2 – 22.7) Stand-level mortality (%) n/a 0.0 (0.0 – 16.7) 7.1 (0.0 – 40.0)

Medium mortality – Telluride Creek Seedling density (per ha) 0.0 (0.0 – 0.0) 50.0 (0.0 – 100.0) 500.0 (350.0 – 650.0) Shrub cover (%) 31.4 (17.3 – 45.5) 64.5 (57.3 – 71.8) 70.0 (53.6 – 86.4) Willow dominance (%) 24.1 (2.7 – 45.5) 44.1 (40.0 – 48.2) 48.2 (42.7 – 53.6) Shrub height (m) 76.1 (51.5 – 100.8) 90.5 (80.3 – 100.7) 124.2 (112.4 – 136.0) Spruce height (m) n/a 2.9 (0.5 – 6.1) 9.2 (0.6 – 23.2) Spruce DRC (cm) n/a 15.0 (2.2 – 53.7) 32.8 (1.3 – 64.1) Spruce DBH (cm) n/a 7.4 (1.0 – 16.7) 25.0 (1.3 – 51.0) Saplings (%) n/a 37.5 (0.0 – 58.3) 4.2 (0.0 – 41.7) Stand density (per ha) n/a 138.5 (54.8 – 224.0) 263.5 (44.1 – 877.0) Basal area (m2 per ha) n/a 4.3 (1.4 – 6.0) 12.6 (0.5 – 77.0) Stand-level mortality (%) n/a 0.0 (0.0 – 0.0) 62.3 (20.0 – 91.7)

High mortality – Boutellier Creek Seedling density (per ha) 25.0 (0.0 – 50.0) 225.0 (50.0 – 400.0) 750.0 (450.0 – 1050.0) Shrub cover (%) 77.3 (74.6 – 80.0) 78.6 (67.3 – 90.0) 51.8 (40.9 – 62.7) Willow dominance (%) 43.2 (39.1 – 47.3) 58.6 (76.4 – 35.5) 41.8 (25.5 – 58.2) Shrub height (m) 114.3 (101.8 – 126.8) 122.2 (110.0 – 134.3) 122.3 (88.2 – 156.4) Spruce height (m) n/a 5.1 (0.5 – 16.9) 6.3 (0.5 – 33.2) Spruce DRC (cm) n/a 28.8 (1.3 – 120.0) 19.4 (1.0 – 59.9) Spruce DBH (cm) n/a 28.3 (0.2 – 65.5) 20.5 (0.1 – 42.7) Saplings (%) n/a 25.0 (16.7 – 50.0) 20.8 (0.0 – 41.7) Stand density (per ha) n/a 98.0 (51.6 – 485.8) 484.6 (159.3 – 1347.7) Basal area (m2 per ha) n/a 3.0 (0.6 – 16.7) 5.8 (0.8 – 9.0) Stand-level mortality (%) n/a 60.0 (0.0 – 77.8) 61.1(25.0 – 71.4)

Table 2-2 Binary logistic analysis of trees surviving the recent spruce bark beetle attack and those that were killed. The logistic regression coefficient is denoted as β, with the estimated odds ratio expressed as eβ. Significant values (p<0.05) are indicated in bold type.

Predictor β eβ p Spruce height 0.427 1.532 <0.001 Spruce DBH 0.124 1.113 <0.001 Spruce DRC 0.177 1.194 <0.001

Likelihood ratio test X2 df P Spruce height 168.820 1 <0.001 Spruce DBH 210.940 1 <0.001 Spruce DRC 214.784 1 <0.001

Hosmer & Lemeshow goodness-of-fit test X2 df p Spruce height 26.267 8 0.001 Spruce DBH 14.706 8 0.065 Spruce DRC 14.405 8 0.072

Chapter 3

Growth Patterns of Subalpine White Spruce (Picea glauca) in Relation to Varied

Levels of Insect Disturbance, Southwest Yukon, Canada

3.1 Introduction

In the boreal forests of North America, phytophagous insects are prevalent disturbance agents which often affect significantly more forest than wildlife (Malmström and Raffa 2000, Dale et al. 2001).

Since many of these insects have evolved alongside their forest hosts for millennia (Furniss and Carolin

1977), periodic outbreaks of endemic insect species may be integral to maintaining ecosystem integrity over the long term (Logan et al. 2003). However, alterations to Earth’s climate have the potential to modify these relationships.

Insect infestations have been particularly acute in the North American cordillera in recent decades, with millions of hectares of forest killed by endemic bark beetles (Coleoptera: Curculionidae,

Scolytinae). These outbreaks have been unprecedented in terms of magnitude and severity, and are widely believed to have been driven by climate change (Ayres and Lombardero 2000, Safranyik and

Wilson 2006, Berg et al. 2006, Werner et al. 2006, Logan and Powell 2009, Bentz et al. 2010, Sherriff et al. 2011, Weed et al. 2013). For instance, prolonged warmer-than-average temperatures in northern forests during the late 20th century (Symon et al. 2005, Prowse et al. 2009) led to the initiation and sustainment of epidemic-level populations of spruce bark beetle (Dendroctonus rufipennis Kirby) (Berg et al. 2006) which have affected over 1.1 million hectares of forest in south-central Alaska and southwest Yukon since 1990 (Werner et al. 2006, Government of Yukon 2013).

Endemic to all spruce forests in North America, spruce bark beetle populations largely subsist on weakened or recently killed white (Picea glauca (Moench) Voss), Engelmann (P. engelmannii Parry ex.

Engelm.), and Sitka (P. sitchensis (Bongard) Carr) spruce trees (Furniss and Carolin 1977, Schmid and 55

Frye 1977) and erupt only intermittently. Because healthy trees can successfully flood beetle galleries with resin, they are only attacked when beetle populations are large enough to overwhelm their defenses (Wallin and Raffa 2004, Garbutt et al. 2006, Raffa et al. 2008, Bentz et al. 2010). Attack by spruce bark beetles typically results in the mortality of their live hosts (Schmid and Frye 1977).

Consequent alterations to forest structure and composition may occur, including increased understory plant diversity (Stone and Wolfe 1996), accelerated growth of understory trees (Veblen et al. 1991a,

Eisenhart and Veblen 2000, Bebi et al. 2003, Berg et al. 2006, Axelson et al. 2010, Sherriff et al. 2011), and a shift in dominance towards unaffected tree species (Veblen et al. 1991a, DeRose and Long 2010).

Outbreaks may also adversely affect nutrient cycling, carbon sequestration and biodiversity (Ayres and

Lombardero 2000, Kurz et al. 2008).

When paired with an ample supply of susceptible host trees, milder winters and warmer growing seasons can facilitate epidemic-level spruce bark beetle outbreaks within their endemic geographical ranges. In particular, temperature increases can reduce cold-induced winter mortality (Schmid and Frye

1977, Ayres and Lombardero 2000, Bentz et al. 2010) and accelerate beetle development rates from a semivoltine (2-year) to an univoltine (1-year) life cycle (Schmid and Frye 1977, Werner and Holsten

1985, Berg et al. 2006). Because low beetle overwintering survival has historically resulted in infestations which rarely last more than three or four years (Garbutt et al. 2006, Werner et al. 2006), the aforementioned changes allow for outbreaks to last significantly longer, provided that there are sufficient numbers of mature spruce hosts (Berg et al. 2006). Temperature-induced drought may further intensify an outbreak by reducing host vigour and defense mechanisms (Werner et al. 2006,

Raffa et al. 2008), causing host defenses to be overwhelmed by fewer beetles.

In addition, warming temperatures can facilitate the expansion of bark beetles into previously unoccupied latitudes and altitudes. During the last three decades of the 20th century, mountain pine beetle (Dendroctonus ponderosae Hopkins) infestations within British Columbia increasingly affected 56

northern and high-elevation habitats that were formerly deemed climatically unsuitable (Carroll et al.

2004). Should the 2.5°C rise in global temperatures anticipated by the International Panel on Climate

Change (Stocker et al. 2013) come to pass, these populations are predicted to undergo a northern latitudinal range shift of up to 7° (Logan and Powell 2001). Since previous predictions of bark beetle range expansions have been exceeded (Weed et al. 2013), an increasing number of distributional limits will likely become affected by outbreaks. Trees growing at the margins of their distribution, such as those located at subarctic alpine treelines, may therefore become more vulnerable to insect disturbances as climate continues to warm (Ayres and Lombardero 2000, Bentz et al. 2010, Marini et al.

2012, Murdock et al. 2013). Already stressed by abiotic factors, these trees may recover slowly from an outbreak (Raffa et al. 2008, Moen et al. 2008). However, to date, relatively few studies have examined the effects of insect outbreaks within forest-tundra ecotones.

Evidence of endemic spruce bark beetle activity has been found to date back to the 18th century at the subarctic treelines of northern Québec by Caccianiga et al. (2008). No regional outbreaks were observed; instead, spruce bark beetles sporadically infested mature spruce individuals and, to a lesser extent, young spruce. Still, extensive mortality occurred at one of their sites during the 1990s, which the authors believed to be triggered by increased stand density and synchronous tree senescence.

Caccianiga et al. (2008) also conducted dendrochronological analysis to determine whether sustained increases in radial growth, otherwise known as growth releases, were triggered in surviving trees by the death of dominant canopy-level trees, as has been observed in many lower elevation stands (Veblen et al. 1991a, Eisenhart and Veblen 2000, Kulakowski et al. 2003, Berg et al. 2006, Sherriff et al. 2011). The selected method of growth release detection, however, prevented the detection of any growth responses in the last decade of a tree’s chronology, and therefore any effects of the late 20th century spruce bark beetle outbreak on subarctic treeline could not be determined.

Using repeat photography, Van Bogaert et al. (2011) was able to document the extensive thinning and fragmentation caused by an outbreak of winter moth (Operophtera brumata L.) in the subarctic alpine treelines of northern Sweden. Over five decades after the infestation, the large-scale mortality of mountain birch (Betula pubescens Ehrh. ssp. czerepanovii (Orlova) Hӓmet-Ahti) was still quite evident on the landscape. While Harsch et al. (2009) postulated that disturbance will delay the occurrence of climate-induced treeline advance, Van Bogaert et al. (2011) posited that severe disturbance – including biotic disturbance – could preclude it entirely. No dendrochronological analysis on affected or unaffected treeline individuals were conducted in their study.

Both Caccianiga et al. (2008) and Van Bogaert et al. (2011) have demonstrated the capacity for insect disturbance to occur in treeline environments. However, despite the novel information that these studies provide, neither offers information on the short- or long-term effects of insect disturbance beyond the principal outcome of host mortality, such as whether surviving individuals experience accelerated growth post-outbreak. Also, relatively little is known about the mechanisms used to select hosts at high altitudes and latitudes.

Through the use of dendrochronological techniques, this study aimed to assess and evaluate differential tree growth patterns between stands affected and those relatively unaffected by a spruce bark beetle outbreak that occurred in southwest Yukon, Canada from 1990 to 2006. Specifically, we investigated (i) whether there were identifiable differences in annual growth patterns between trees that were killed by spruce bark beetle and those that survived the infestation, and (ii) if a growth release was evident in spruce trees that were not killed during the outbreak. As very little is known of the nature of insect disturbance within forest-tundra ecotones, this study has the capacity to better inform predictions of future ecotone dynamics, particularly since these ecotones are temperature-limited and therefore expected to advance with continued climate warming.

3.2 Methods

3.2.1 Study Area

We conducted this study within the Kluane Wildlife Sanctuary in southwest Yukon. Located along the northwest-southeast path of the Shakwak Trench between Kluane Lake and the town of Haines

Junction, our sites were positioned within the subalpine forest zone of the Kluane Ranges (sensu

Douglas 1974). White spruce (Picea glauca) dominates the forests of the region, with canopy cover becoming gradually more open and discontinuous as elevation increases. No secondary tree species were observed at these sites, although balsam poplar (Populus balsamifera L.) and trembling aspen

(Populus tremuloides Michx.) stands can be found at lower elevation forests within the Shakwak Trench

(Krebs and Boonstra 2001).

Approximately 400,000 hectares of forest across southwest Yukon, including the subalpine forests of the Kluane Ranges, were affected by epidemic levels of spruce bark beetle (Dendroctonus rufipennis) from the early 1990s to the mid-2000s (Garbutt et al. 2006, Government of Yukon 2013). With the exception of a relatively small outbreak occurring between 1934 and 1942, the Kluane region is not believed to have experienced any prior major infestations (Berg et al. 2006), making this epidemic unprecedented in terms of magnitude, severity and duration. Prolonged periods of warmer temperatures during the latter half of the 20th century (Prowse et al. 2009) are believed to have contributed to the outbreak by inducing moisture stress in mature host trees, increasing beetle overwinter survival, and doubling of the insect’s maturation rate, otherwise known as voltinism (Berg et al. 2006). This infestation has extended as high as the altitudinal treeline and resulted in the death of more than half of all mature spruce in many parts of the region (Garbutt et al. 2006, Government of

Yukon 2013).

Despite its relative proximity to the Pacific Ocean, the Kluane region lies in the rain shadow of the

St. Elias Mountains and therefore experiences a semiarid, continental climate (Scudder 1997, Krebs and 59

Boonstra 2001). Instrumental records for 1981-2010 from the Burwash Landing climate station (61°22’N

139°03’W, 806.2masl) show a mean annual temperature of -3.2°C, with mean monthly temperatures of

-20.5°C for January and 13.1°C for July, and a mean annual precipitation of 274.7mm, with 105.5cm falling as snow (Environment Canada 2014).

3.2.2 Dendrochronological Analysis

We sampled growth rings from mature white spruce at two subalpine sites in the study region during the summer of 2012 (Figure 3-1). The site at Telluride Creek was moderately affected by the recent outbreak (medium levels of spruce mortality), while the site at Mount Decoeli was minimally affected (low levels of spruce mortality). Both sites were selected based on their accessibility, representativeness of the surrounding region, and similarity in terms of elevation, aspect and slope.

Transects were established at two elevations at each site: (i) ‘forestline’, the transition between open- canopy forest and discontinuous forest with non-overlapping tree crowns; and (ii) ‘treeline’, the transition between the aforementioned discontinuous forest and higher elevation shrub-fields populated with scattered, stunted krummholz-form trees. Neither site experienced spruce bark beetle attack at treeline. Mature forestline spruce killed during the outbreak was estimated to be 12.5% at

Mount Decoeli and 60.0% at Telluride Creek (see Chapter 2). A minimum of 12 trees was sampled along each treeline and forestline transect, with a maximum of 24 trees sampled if individuals killed by spruce bark beetle were present. Given that multiple samples were obtained from each tree, as described below, and that a strong common ring-width signal exists in southwest Yukon (Youngblut and Luckman

2008), this sample size was able to provide sufficient replication for our analysis.

The individuals selected for sampling were both representative of the area and as old as possible, to maximize the length of record (Speer 2010). Samples were obtained from two heights on each tree.

In order to obtain the best age estimate for each tree, a single increment core or cross-section was

taken as low to the ground as possible; in some cases, particularly in dead individuals, rot necessitated sampling from greater heights. In addition, either a disk or two increment cores perpendicular to each other were obtained from as close to breast height (1.3m above ground) as possible. The collection of a disk, as opposed to an increment core, was favoured when individuals had a small diameter (<10cm) at sampling height. We recorded the height from which each sample was taken, as well as the height, health, basal diameter (DRC), and diameter at breast height (DBH) of each sampled tree.

Handling, preparation and processing of spruce samples followed standard dendrochronological techniques (Stokes and Smiley 1968). Increment cores were mounted and glued to wooden mounts.

Cores and disks were then sanded with progressively finer sandpaper grades, up to 600 grit, to attain clearly visible rings. Ring-widths were measured to the nearest 0.001mm using a stereomicroscope and

Velmex sliding stage (Velmex Inc., Bloomfield, New York, USA) interfaced with MeasureJ2X software

(VoorTech Consulting, Holderness, New Hampshire, USA). Visual cross-dating was statistically validated using the program COFECHA (Holmes 1983). Of the deceased samples at forestline, six trees were removed from further analysis as they died prior to the recent spruce bark beetle outbreak. Of the five trees removed from the Telluride Creek chronology, year of death was identified as 1902 for two individuals, with the remainder dying in 1905, 1908 and 1949. The singular individual removed from the

Mount Decoeli chronology died in 1988.

Chronology Development

To avoid the use of partial rings, the final year of growth was discarded for both living and beetle- killed individuals. In order to preserve low-frequency variations, the raw ring-width data were standardized with the program ARSTAN (Cook and Krusic 2005) using either a modified negative exponential curve, a linear regression line with a negative slope, or a horizontal line passing through the mean (Fritts et al. 1969, Helama et al. 2004). For each site, ring-width measurements from all radii were 61

combined to create standardized and residual chronologies that represented the growth of (i) living trees at treeline, (ii) living trees at forestline, and (iii) trees at forestline killed by spruce bark beetle during the recent outbreak. Samples taken from breast height are generally considered to be best for deriving relationships with climate, as tree growth may not respond to climatic variables until it reaches a particular age and, by extension, height (Fritts 1976, Speer 2010). Samples obtained from root crown, however, can be incorporated into a chronology without compromising the detection of climatic signals

(Chhin and Wang 2005). By combining samples taken from the oldest part of a tree (root crown) with those from breast height, our chronologies were able to reflect the influence of climate and extend as far back in time as possible. Characteristics of these chronologies are summarized in Table 3-1.

Although the findings of St. George et al. (2013) suggest that Picea spp. only very rarely form absent rings, our samples had several incidences of locally absent (i.e. entirely missing from the sampled radius; n=15) and partially absent (i.e. present for only a small section of the sampled radius; n=33) rings.

Reasons for this relatively high number of absent rings are not clear and require further research.

Initial comparisons of tree growth patterns between chronologies were made by visually inspecting the standardized chronologies created by ARSTAN, as well as comparing marker years, which were identified as years in which ring-widths were ≥1.5 standard deviations above or below the mean

(as indicated by COFECHA). In addition, linear regression was used to estimate multi-decadal growth trends of each standardized ring-width chronology. Other studies, such as Lloyd and Fastie (2002) and

Mamet and Kershaw (2013), have favoured a 150-year time period divided into three intervals of equal length. However, because climate records from Burwash Landing date from late-1966 onwards, we chose to analyze our standardized chronologies using 33-year segments extending backwards from the termination of the common period in 1998 (1867–1899, 1900–1932, 1933–1965, and 1966–1998).

Finally, we compared Pearson’s product moment correlation coefficients between standardized chronologies for the common period 1866–1998. Evaluating similarities between chronologies via 62

correlation analysis has been used in other studies at or in close proximity to altitudinal treelines (e.g.

Lloyd and Fastie 2002, Youngblut and Luckman 2008, Trindade et al. 2011). We were particularly interested in comparing the growth patterns of trees at forestline which survived the recent spruce bark beetle attack with those that were beetle-killed, to assess the possibility that spruce bark beetle targeted trees with specific growth trends.

Climate-Growth Analysis

We used the program DendroClim2002 (Biondi and Waikul 2004) to assess the relationships between radial tree growth and climate. This program calculates Pearson’s product moment correlation coefficients between ring-widths and climate variables, with 95% confidence intervals derived from

1,000 bootstrapped samples (Biondi and Waikul 2004). Removal of temporal autocorrelation is necessary for determining climate-growth relationships (Speer 2010), therefore residual chronologies were used (e.g. Eisenhart and Veblen 2000, Vittoz et al. 2008, Mamet and Kershaw 2013, Chavardès et al. 2013). The climate signal strength of these residual chronologies was determined via the expressed population signal (EPS), and were found to meet the recommended threshold value of 0.85 (Wigley et al. 1984) during the period of climate analysis. Following Youngblut and Luckman (2008), who also worked in southwest Yukon, EPS values were calculated for each residual chronology using a 30-year moving window with a 29-year overlap. The residual chronologies were compared against monthly mean temperatures and precipitation totals from the Burwash Landing climate station, located approximately 58km and 77km from our Telluride Creek and Mount Decoeli sites, respectively.

Instrumental records at this climate station are relatively short, therefore we ran correlations on climate data from 1967 to 1998 (i.e. the first complete year of record to the last year of ring-widths shared by chronologies). In each analysis, a 21-month climate window extending from the January of the prior year to September of the growth year was assessed. This climate window was chosen so as to exclude 63

months from the growth year that do not contribute to ring formation (October – December) while accounting for potentially important conditions in the year previous.

Detection of Post-Outbreak Growth Releases

To detect the presence of growth releases in living spruce trees following the recent spruce bark beetle outbreak (approx. 1990 – present), we used the running median method outlined by Rubino and

McCarthy (2004), an adaptation of the radial growth averaging method developed by Nowacki and

Abrams (1997). The use of a running median as opposed to a running mean is arguably more applicable to the non-normality of biological trends typically found in radial tree growth, and therefore more statistically robust (Rubino and McCarthy 2004). Individual chronologies for each living forestline tree were created in ARSTAN (Cook and Krusic 2005) by averaging the raw ring-width measurements of all radii obtained from the respective spruce individual; standardization was not used. Percentage radial growth change (%GC) was calculated on a yearly basis for each tree from 1990 onwards as follows:

%GC = [(M2/M1)/M1] x 100, where M1 = the preceding 10-year median and M2 = the subsequent 10-year median; %GC was affixed to the last year of the preceding 10-year period (M1) (Nowacki and Abrams 1997). The 10-year window selected for creating the running median has been found to adequately smooth ring-width variability caused by short-term climatic variation (Berg et al. 2006) while capturing growth increases related to canopy disturbances (Nowacki and Abrams 1997). A release was defined as a minimum growth change of 50% sustained for a period of 5 years. Our release factor was more conservative than those chosen by others investigating historical growth releases in response to spruce bark beetle in closed subalpine spruce stands (e.g. Veblen et al. 1991a, Berg et al. 2006, Sherriff et al. 2011) in order to reflect the greater canopy openness of our study sites. It should be emphasized that, given the selection of a 10- year window, this radial growth averaging method excludes the initial and final 10 years of each ring- 64

width series from release detection due to an insufficient number of years from which to calculate %GC.

As such, 2001 was the last year during which a radial growth increase could be detected in our samples.

This led to some difficulties in light of the recent timing of the spruce bark beetle outbreak, as discussed in later sections of this chapter. However, the use of a shorter window would likely not have been long enough to remove the influence of climatic anomalies (Rubino and McCarthy 2004, Berg et al. 2006).

To verify that identified releases were not a response to climate, this analysis was also performed on samples from the treeline transects. Following Sherriff et al. (2011), we assumed that a release could be attributed to climate if identified in spruce from both the higher-elevation, open-canopy stands at treeline and the lower-elevation, less-open-canopy stands at forestline. However, if a release was only identified at forestline, we attributed the release to disturbance.

3.3 Results

3.3.1 Comparison of Standardized Chronologies

Series length at Mount Decoeli (site with low spruce bark beetle kill) ranged from 60 to 334 years, with a mean of 132.08 ± 59.63 years (SD), and ranged from 52 to 365 years (217.10 ± 115.75) at

Telluride Creek (site with moderate spruce bark beetle kill). The treeline chronology from Telluride

Creek was notably shorter than the lower elevation chronology from forestline, whereas the length of the Mount Decoeli chronologies did not differ substantially between elevations (Table 3-1, Figure 3-2).

That said, sample depth remained low until after 1900 for both treeline chronologies (Figure 3-2).

Spruce bark beetle-induced mortality occurred between 1994 and 1998 at Mount Decoeli and between 1996 and 2001 at Telluride Creek; the median year of overall recorded mortality was 1997.

Trees killed by spruce bark beetle ranged from 111 to 266 years of age at Mount Decoeli, with a mean of

158.36 ± 46.68 years. Deceased trees sampled at Telluride Creek tended to be older, ranging from 239 to 336 years of age (mean 297.29 ± 33.11 years). However, it must be emphasized that falling within

these ages did not necessarily predispose trees to attack. In fact, the oldest living tree sampled at forestline was three decades older than the oldest attacked tree.

Growth Trends at Treeline

The two treeline chronologies were significantly positively correlated with one another (Table 3-2) and exhibited similar growth trends. At Mount Decoeli, treeline tree growth exhibited significant trends throughout the period of analysis: growth trends were negative prior to 1900, positive from 1900 until

1930, negative throughout the 1930s, 1940s and 1950s, and positive from the 1960s to present day

(Table 3-3, Figure 3-2a). Treeline individuals at Telluride Creek followed these growth patterns closely

(Figure 3-2c), although no significant trends were observed prior to 1900 (Table 3-3), likely a function of low sample depth. During the last three decades of the 20th century, growth remained high but without any significant trend (Table 3-3), with more pronounced positive growth beginning in the mid-1990s and continuing to present day (Figure 3-2c). Interestingly, despite a significant positive correlation between the Mount Decoeli and Telluride Creek treeline chronologies (r=0.676), few marker years were shared between the two series (Figure 3-3).

Growth Trends at Forestline

The growth of beetle-killed and outbreak-surviving trees at Mount Decoeli had similar low- frequency growth trends throughout the length of record (Figure 3-2b). In the decades immediately prior to 1900, growth was positive in both beetle-killed trees and surviving individuals, although regression analysis identified this growth as insignificant for the latter (Table 3-3). Both chronologies exhibited steady growth from 1900 to the mid-1940s, then experienced decreased growth until approximately 1960, when growth began following a significant positive trend (Table 3-3). Year-to-year variations remained similar throughout the 20th century, though beetle-killed individuals generally 66

experienced higher rates of growth than those that survived the outbreak, especially between 1925 and

1975 (Figure 3-2b). Growth at forestline was weakly correlated with treeline growth at the same site

(Table 3-2).

The growth trends of surviving and beetle-killed trees were more strongly correlated at Telluride

Creek (Table 3-2), and there was reasonable agreement between the markers from the surviving and beetle-killed spruce chronologies (26.5% shared) at Telluride Creek’s forestline in contrast to the 13.7% agreement between the two forestline chronologies at Mount Decoeli (Figure 3-3). The Telluride Creek chronologies exhibited similar high-frequency trends from 1680 onwards, but displayed significant low- frequency differences between 1725 and 1900 (Figure 3-2d). During this period, particularly during the

1800s, the ring-widths of beetle-killed individuals were greater than their counterparts. The accelerated growth of these individuals began to decline in the late 1800s, which is reflected in the significant negative growth trend identified in our linear regressions. After 1900, however, the chronologies shared similar multi-decadal growth trends. Growth of both killed and non-killed spruce increased from 1900 until 1940, then decreased until the mid-1960s, and increased again after this point (Table 3-3). A period of accelerated growth has been observed in surviving individuals since 2000 (Figure 3-2d), possibly indicating the occurrence of post-outbreak growth releases (see 3.3.3). Growth was also correlated with treeline growth at Telluride Creek approximately twice as well as the equivalent chronologies at Mount Decoeli (Table 3-2).

Growth trends of beetle-killed trees at Mount Decoeli and Telluride Creek were strongly positively correlated, indicating similar year-to-year and multi-decadal variations; this was not the case with the surviving chronologies from these sites (Table 3-2). Despite the considerable variation in marker years at forestline (Figure 3-3), two marker years were exclusively shared by the beetle-killed chronologies

(1909 – narrow ring, 1942, wide ring). In addition, 1904 was identified as a narrow marker year in both of the surviving spruce chronologies but in neither of the beetle-killed chronologies. 67

3.3.2 Climate-Growth Relationships

Radial growth of white spruce growing at forestline was strongly negatively correlated with precipitation during February of the year prior to the growth season (Figure 3-4). All six residual chronologies exhibited negative correlations with spring/early summer precipitation of the growth year

(Figure 3-4), though stronger correlations were found at Mount Decoeli. Positive correlations between precipitation and tree growth were found in March of the growth year, as well as the end of the current growing season (Figure 3-4). At treeline, precipitation-growth relationships were strongest in August of the growth year (Figure 3-4a and d), while September of that year was typically favoured by forestline individuals (Figure 3-4b, c, e and f).

Growth at most sites was positively correlated with mean temperatures from the previous

January as well as during the later months of the growing season, including September (Figure 3-5). In particular, forestline individuals exhibited the strongest correlations with temperature in July (Figure

3-5b, c, e and f). Although not as strongly correlated, June was also statistically significant for beetle- killed trees (Figure 3-5c and f). Treeline individuals, on the other hand, displayed highest summer growth in June (Figure 3-5a and d). All six chronologies were most strongly positively correlated with the temperatures in March of the prior year (Figure 3-5). These chronologies also exhibited positive growth associations with temperature in the previous August, although these were not consistently significant.

3.3.3 Post-Outbreak Growth Releases

No releases were identified in any of the mature spruce trees sampled from Mount Decoeli following the initiation of the spruce bark beetle outbreak in the early 1990s, regardless of elevation

(Figure 3-6). While some radial growth increases were observed in these trees, none of these were sustained (≥5 years) and were therefore not classified as growth releases. Only one mature spruce individual sampled from treeline at Telluride Creek exhibited any sustained radial growth increases 68

(Figure 3-6); however, as no trees were affected by spruce bark beetle at this elevation, this increase could not be in response to beetle-induced mortality. At the forestline of Telluride Creek, a growth release was identified in six of the twelve (50.0%) mature spruce individuals sampled (Figure 3-6). A seventh tree demonstrated an increase in ring-widths from 1999 onwards but could not be classified as experiencing a growth release, given that 2011 was the last year included in our release analysis.

3.4 Discussion

3.4.1 Growth Trends and Climate-Growth Relationships

Our results indicate that trees which succumbed to the recent spruce bark beetle attack experienced prolonged periods of greater radial growth in the past relative to their surviving counterparts, despite similar high- and low-frequency trends. At Telluride Creek, the growth of beetle- killed individuals was significantly greater during much of the 1800s. Growth was greater in beetle-killed individuals at Mount Decoeli as well, although differences at this site were less acute and occurred a century later (between 1925 and 1975). The periods of time in which radial growth of beetle-killed trees surpassed those of surviving spruce are inconsistent. However, they correspond to time frames where the mean age of beetle-killed trees ranged between 77.8 (± 40.3) and 146.8 (± 50.2) years old. The timing also coincides with periods of regional ring-width incoherence as observed by Youngblut and

Luckman (2008).

The conditions which led to increased radial growth rates earlier in these beetle-killed trees’ lives may have ultimately contributed to their demise. These trees likely reached beetle-preferred sizes more quickly than other forestline trees, which may have led small, endemic populations of spruce bark beetle to infest – but not kill – these individuals in the past. Positive selection of previously attacked trees has been reported for both Dendroctonus micans (Gilbert et al. 2001) and D. rufipennis (Werner et al. 2006), and has been observed at altitudinal treelines (Caccianiga et al. 2008). If this premise is correct, long-

term growth patterns may be equally as important as radial growth rates immediately prior to a beetle outbreak (Hard 1985, Doak 2004) in determining tree survival, at least within forest-tundra ecotones.

The existence of marker years unique to either beetle-killed or surviving chronologies supports this hypothesis, although the importance of growth in these particular years is unknown.

As with other high-elevation stands in southwest Yukon (Youngblut and Luckman 2008) and several subarctic and subalpine tree populations (e.g. MacDonald et al. 1998, Cullen et al. 2001, Vittoz et al. 2008, Flower and Smith 2011, Zhang et al. 2011, Trindade et al. 2011, Porter and Pisaric 2011, Mamet and Kershaw 2013), ring-widths at both sites were well-correlated with growing year summer temperatures. However, our climate-growth analysis identified stronger early summer temperature- growth correlations for beetle-killed trees than for still-living forestline trees. This suggests that a differential response to climate may have predisposed particular individuals to attack and/or beetle- induced mortality, although the mechanisms behind this are not clear. Since microsite variables are able to alter tree response to climate (Wilmking et al. 2005), it is possible that shared microsite variables either enabled particular trees to overcome beetle attack or triggered the death of particular individuals. Because surviving and beetle-killed trees at forestline were comingled, however, the differences in climatic response between these chronologies cannot be easily attributed to a particular microsite variable, and may in fact be caused by other unidentified factors or a combination thereof.

Our results also identified differences in climate-growth relationships between white spruce growing at treeline and forestline elevations. Elevational placement has been reported to affect the influence of climate, with temperature-growth relationships increasing with proximity to treeline (Lloyd and Fastie 2002, Vittoz et al. 2008) and lower-elevation trees being more sensitive to precipitation

(Rolland and Lempérière 2004, Griesbauer and Green 2012, Ohse et al. 2012). Although we found that treeline chronologies had higher mean sensitivity values (t=7.22, p=0.002), a measure of year-to-year variation and therefore an indicator of climatic influence (Fritts 1976), we did not find that elevation 70

much altered the extent of climate-growth relationships. Rather, the majority of observed differences appear to be in timing. During the growth season, for example, treeline chronologies were best correlated with June mean temperatures whereas those at forestline were strongest in July. Similarly, rainfall during July and August of the growth season was well-correlated with individuals growing at treeline; forestline individuals, in contrast, were generally better correlated with September precipitation. We do not believe that these elevational differences are age-related (Szeicz and

MacDonald 1994) since the climate-growth relationships of young chronologies (treelines, Mount

Decoeli forestline) were not distinct from older chronologies (Telluride Creek forestline). Instead, our findings imply that even slight elevational differences in stands (<75masl, in this case) may dictate the temporal nature of climate-growth relationships within the forest-tundra ecotone. Because warm early summer temperatures allow for the earlier onset of snowmelt and soil thaw, thereby facilitating the absorption of moisture and nutrients required for growth (Fritts 1976, Kramer and Kozlowski 1979), it is plausible to assume that an earlier start to the growing season would disproportionately favour growth in the significantly slower growing trees (t=-13.03, p<0.001) at treeline.

All chronologies had significantly strong positive relationships with previous March temperatures, also reported by Youngblut and Luckman (2008), but the importance of other antecedent conditions to radial growth differed between sites. Mount Decoeli chronologies shared consistently strong, positive relationships with previous January temperatures. In addition, the strong negative relationships with spring precipitation present at Mount Decoeli were absent at Telluride Creek. This suggests better synchrony in climate-growth relationships between elevations at Mount Decoeli, and may serve as an indicator of the highly complex interactions with climate in southwest Yukon. In examining white spruce climate-growth relationships across Yukon, Griesbauer and Green (2012) noted that correlations were relatively low in the Kluane Lake region but believed this to be caused by complex interactions between climate and growth rather than a lack of climatic control on growth processes. Climate-growth 71

relationships in the region have also become less pronounced and increasingly variable during the most recent (1977-2007) positive phase of the Pacific Decadal Oscillation (PDO) (Chavardès et al. 2013).

Regional variation in response to climate may lead to increasingly divergent growth trends among trees, which may in turn contribute to a tree’s susceptibility to spruce bark beetle attack.

We believe that the implications of these aforementioned differences in radial growth and response to climate will be pertinent to stands throughout southwest Yukon, as well as subarctic and subalpine forest-tundra ecotones. Regardless of elevational placement or beetle-induced mortality, similar low-frequency trends were exhibited by our chronologies and closely paralleled those observed by Youngblut and Luckman (2008). Despite being situated in the montane valley bottom forests of the

Shakwak Trench, the chronologies created by Zalatan and Gajewski (2005) also exhibited generally similar growth trends. This is indicative of the strong ring-width signal of high-elevation stands in southwest Yukon for the last few hundred years (Youngblut and Luckman 2008), and implies that this signal may be shared by low-elevation stands as well. With future climate warming anticipated to continue affecting high elevations and latitudes more than other regions (Symon et al. 2005, Nogués-

Bravo et al. 2007, Rebetez and Reinhard 2007, Stocker et al. 2013), forest-tundra ecotones may become characterized by increasingly divergent radial growth and climate relationships (D’Arrigo et al. 2008) and insect disturbances, although additional research is required to identify the specific mechanisms, and their interactions, responsible for this divergence.

3.4.2 Post-Outbreak Tree Growth

Bark beetle infestations typically result in the mortality of living host trees (Schmid and Frye

1977), which allows for increased light penetration and reduced competition for resources. For this reason, beetle-induced mortality in closed-canopy forests often causes accelerated radial growth (i.e. growth releases) in neighbouring trees which can be sustained for decades (e.g. Romme et al. 1986,

Veblen et al. 1991a, Parish et al. 1999, Zhang et al. 1999, Alfaro et al. 2004, Berg et al. 2006, Axelson et al. 2010, Smith et al. 2012, Amoroso et al. 2013). Until now, however, it has remained unclear as to whether this effect holds true within forest-tundra ecotones. A recent study which examined the role of biotic disturbance on the subarctic treelines surrounding Hudson Bay attempted to detect the presence of coincident growth releases in surviving spruce, but was unsuccessful due to methodological constraints (Caccianiga et al. 2008). To our knowledge, this study is therefore the first to identify growth releases following insect disturbance in a forest-tundra ecotone.

Our comparison of treeline and forestline chronologies indicated that growth releases observed at forestline were due to insect disturbance, as treeline individuals lacked the coincident releases which would be expected if climate was the driver (Sherriff et al. 2011, Smith et al. 2012). We anticipated that growth releases would be evident in at least some of the remnant mature spruce individuals sampled from the recently beetle-affected subalpine stands of southwest Yukon, but only one of the two sites exhibited growth releases in the decade following epidemic initiation. This difference in post-outbreak growth between locations is likely a function of the differential levels of beetle-induced mortality.

Forestline trees sampled at Mount Decoeli, where mature spruce mortality averaged 12.5%, did not exhibit any sustained increases in radial growth. In contrast, forestline trees sampled at Telluride Creek

– where mortality averaged 60% – displayed evidence of a high-severity outbreak, defined by Sherriff et al. (2011) as one where growth releases are detected in ≥50% of surviving individuals. This implies that subalpine white spruce stands are affected by competition for resources, despite the relatively open nature of the canopy cover.

Our results echo that of Amoroso et al. (2013), who examined stand recovery of mixed-conifer forest in southeastern British Columbia following a severe mountain pine beetle (Dendroctonus ponderosae) epidemic. The post-outbreak growth rates of lodgepole pine (Pinus contorta var. latifolia), subalpine fir (Abies lasciocarpa) and interior spruce (Picea glauca x engelmanii) were highest in stands 73

which experienced high mortality rates, and were therefore concluded to be strongly affected by stand mortality levels. The findings of Veblen et al. (1991a) further support our hypothesis that greater stand- level mortality leads to greater post-outbreak growth rates. This suggests that trees in forest-tundra ecotones respond to insect disturbance in a similar manner. However, while Veblen et al. (1991a) identified increased radial growth in all stands affected by a major spruce bark beetle infestation in the subalpine forests of Colorado, we did not observe any growth releases in trees growing in less-affected stands. Although both Mount Decoeli (Yukon) and Walton Creek (Colorado) experienced minimal spruce mortality, the trees that we sampled within the forest-tundra ecotone grew in stands that were less dense (360 trees/ha) and had smaller basal areas (22.1 m2/ha) than those examined by Veblen et al.

(1991a) (density: 1542 trees/ha; basal area: 59.2 m2/ha). The occurrence and extent of post-outbreak growth releases is therefore likely determined by stand characteristics, which influence susceptibility to spruce bark beetle attack (see Chapter 2) and resource competition amongst trees, as well as mortality levels.

Our ability to predict the long-term implications of beetle-induced mortality on the trees of the forest-tundra ecotone is somewhat limited by the scope of our analysis. Our method of growth release detection prevented any observation of release initiation after 1996. Given that the majority of beetle- killed trees sampled at our sites were assigned a date of death after 1996 (72.7% and 85.7% at Mount

Decoeli and Telluride Creek, respectively) and that some studies (e.g. Parish et al. 1999, Alfaro et al.

2004, Sherriff et al. 2011, Smith et al. 2012) have suggested that trees may experience a delayed response to canopy thinning, it is very likely that additional growth releases have occurred in these stands over the last decade. The standardized ring-width chronologies indicated that accelerated growth was indeed sustained until 2012 (the year of sampling) at Telluride Creek, but we were unable to demonstrate this given the temporal window used in our analysis. Still, our results do illustrate that trees within forest-tundra ecotones benefit from the enhanced resource availability associated with 74

bark beetle outbreaks, and suggest that the extent of radial growth increase depends on overstory mortality levels and/or stand characteristics such as density and basal area. Our results also indicate that old trees (i.e. >250 years) are capable of responding substantially to canopy disturbance, contrary to the findings of Alfaro et al. (2004). While we did not examine subcanopy individuals, we presume that they have experienced equivalent, if not more acute, accelerated growth following the death of nearby overstory trees, as has been observed in closed-canopy stands (Veblen et al. 1991a, Smith et al.

2012). However, it remains to be seen whether the existent growth releases will be sustained for years to come, or if they will be shorter in duration than those in similarly affected closed-canopy stands. It is also unclear whether trees growing in minimally-affected stands have no response to beetle-induced mortality or merely a delayed response. We therefore recommend long-term monitoring of stands in the area to more accurately determine the duration of growth releases in surviving spruce individuals and its interactions with stand-level mortality, and to compare the long-term effects of insect disturbance on the forest-tundra ecotone with those that occur in closed-canopy forests.

Figure 3-1 Location of Kluane region dendrochronological study sites, southwest Yukon. Mount Decoeli (MD) was minimally affected by the recent spruce bark beetle outbreak, with 12.5% of mature spruce killed at forestline. Telluride Creek (TC) experienced much higher levels of beetle-caused mortality, with 60.0% of mature spruce killed at forestline. Neither site experienced spruce bark beetle attack at treeline.

Figure 3-2 Standardized chronologies for Mount Decoeli and Telluride Creek at treeline and forestline elevations. Solid lines represent ring-width indices; dashed lines represent sample depth. Black lines represent chronologies derived from living individuals; grey lines represent chronologies derived from individuals killed by spruce bark beetle.

Figure 3-3 Marker years (± 1.5 standard deviations of the ring-width mean) for each chronology. Bars above the horizontal line represent wide rings, whereas those below the line represent narrow rings. Of the 133 marker years identified, 63.9% were unique to only one chronology, 30.1% were shared by at least two chronologies, and 6.0% were opposing (i.e. narrow for one chronology and wide for another).

Figure 3-4 Correlation between total monthly precipitation and the residual chronologies from Mount Decoeli (Dec) and Telluride Creek (Tel) for the 1968–1998 time period. From top to bottom are the chronologies created from treeline (T), surviving forestline (F-a) and beetle-killed (F-s) individuals. Solid bars represent statistically significant (p<0.05) correlations, whereas open bars are non-significant values. Abbreviated months in lower case represent the year prior to growth; those in upper case represent the year of growth.

Figure 3-5 Correlation between mean monthly temperature and the residual chronologies from Mount Decoeli (Dec) and Telluride Creek (Tel) for the 1968–1998 time period. From top to bottom are the chronologies created from treeline (T), surviving forestline (F-a) and beetle-killed (F-s) individuals. Solid bars represent statistically significant (p<0.05) correlations, whereas open bars are non-significant values. Abbreviated months in lower case represent the year prior to growth; those in upper case represent the year of growth.

Figure 3-6 Percentage change in radial growth (following Rubino and McCarthy (2004)) for mature spruce individuals growing at Mount Decoeli and Telluride Creek treeline and forestline elevations. Solid line at the 50% mark denotes the threshold for classifying accelerated radial growth as a release.

Table 3-1 Attributes of the Mount Decoeli (low spruce bark beetle kill) and Telluride Creek (moderate spruce bark beetle kill) standardized ring- width chronologies. Values of pre-outbreak stand density and series length are means ± one standard deviation. Chronology time series were truncated to omit the final year of growth recorded by each radii. Mean sensitivity and series intercorrelation values were obtained from COFECHA output files.

Mount Decoeli Telluride Creek Treeline Forestline Treeline Forestline Parameter Living Beetle-killed Living Beetle-killed Pre-outbreak density (trees/ha) 51.46 ± 39.46 505.71 ± 402.35 116.88 ± 93.11 960.44 ± 840.80 Chronology length 1727-2011 1801-2011 1734-1998 1866-2011 1649-2011 1661-2000 Series length (years) 119.58 ± 77.03 120.77 ± 50.50 159.36 ± 46.68 80.92 ± 28.12 308.50 ± 40.19 298.29 ± 33.12 Number of trees 12 13 11 12 12 7 Number of radii 29 41 32 34 37 22 82 Mean ring-width (mm) 0.69 ± 0.34 0.96 ± 0.19 0.84 ± 0.29 0.70 ± 0.36 0.94 ± 0.45 0.98 ± 0.24

Mean sensitivity 0.26 0.21 0.21 0.26 0.21 0.22 Series intercorrelation 0.44 0.57 0.61 0.50 0.62 0.60

Table 3-2 Pearson’s product-moment correlation coefficients between standardized chronologies for the common period (1866 – 1998). Significant correlations (p<0.05) are indicated in bold type. Row headings are full chronology descriptors, column headings are abbreviations (Dec: Mount Decoeli, Tel: Telluride Creek, T: treeline, F: forestline, a: living, b: beetle-killed).

Series Dec-T Dec-F-a Dec-F-b Tel-T Tel-F-a Tel-F-b Mount Decoeli Treeline 1.000 Forestline – living 0.187 1.000 Forestline – beetle-killed 0.282 0.672 1.000

Telluride Creek Treeline 0.676 0.258 0.405 1.000 Forestline – living 0.400 0.378 0.675 0.562 1.000 Forestline – beetle-killed 0.524 0.450 0.611 0.425 0.822 1.000

Table 3-3 Growth trends from the Mount Decoeli and Telluride Creek chronologies from 1867 to 1998. Standardized regression coefficients (β) are shown for each series for the four time periods. Significant correlations (p<0.05) are indicated in bold type.

Series 1867 – 1899 1900 – 1932 1933 – 1965 1966 – 1998 Mount Decoeli Treeline -0.670 0.630 -0.747 0.463 Forestline – living 0.195 -0.020 -0.545 0.742 Forestline – beetle-killed 0.566 0.142 -0.489 0.539

Telluride Creek Treeline -0.084 0.745 -0.441 0.220 Forestline – living 0.247 0.778 -0.588 0.488 Forestline – beetle-killed -0.788 0.750 -0.665 0.542

Chapter 4

Summary and Conclusions

Although insect disturbance can occur in treeline environments (Caccianiga et al. 2008, Van

Bogaert et al. 2011), relatively little is known about the selection of hosts or the effects of insect disturbance in these high altitude and latitude environments. Our research investigated the effects of spruce bark beetle (Dendroctonus rufipennis Kirby) disturbance on the growth and establishment of trees in the subarctic alpine forest-tundra ecotones of southwest Yukon, and examined factors that may have predisposed treelines to infestation. By combining information collected from vegetation surveys and dendrochronological analysis, this research provides evidence that spruce bark beetle is capable of causing widespread – albeit sporadic – mortality within forest-tundra ecotones, including at treeline elevations, and that many of the host selection mechanisms at high elevations are similar to those in low-elevation, closed-canopy stands.

We identified strong heterogeneity in beetle-induced mortality levels across the forest-tundra ecotones of southwest Yukon. The number of trees killed by spruce bark beetle at our study sites ranged from 0.0% to 77.8% at treeline elevations (median 0.0%) and from 0.0% to 91.7% at forestline elevations (median 40.0%), with a median of 20.0% killed across the forest-tundra ecotone. Even when considering our six study sites in terms of mortality class (low, medium, high), stand-level mortality varied considerably between plots, elevations and sites. Despite the sporadic nature of beetle attack across the landscape, several trends emerged during our analysis in terms of host selection.

Consistent with research conducted at elevations below the forest-tundra ecotone, stands characterized by large, densely growing spruce trees experienced high mortality levels. Spruce size metrics, particularly diameter at breast height, were good predictors of the likelihood of an individual

succumbing to spruce bark beetle attack. However, our results indicated that stand structure likely had more influence on determining susceptibility to spruce bark beetle than individual tree characteristics.

Because treelines were generally comprised of smaller, open-grown trees, the impacts of insect disturbance were most clearly pronounced at – though not exclusive to – forestlines. Treelines that more closely resembled the densities and basal areas of forestline, and were located in close proximity to forestline stands hosting large populations of spruce bark beetle, also experienced high levels of mortality. This suggests that elevation itself did not restrict host selection. Rather, elevational influence on stand characteristics determined the upward extent of infestation.

During dendrochronological analysis, we found evidence that host selection mechanisms may also be influenced by long-term radial growth and/or climate relationships. Trees which succumbed to spruce bark beetle attack had experienced prolonged periods of greater radial growth than their surviving counterparts in the past. These periods of time were not consistent in the chronologies examined, but did correspond to when the mean age of beetle-killed trees ranged between 77.8 (± 40.3) and 146.8 (± 50.2) years old. In addition, beetle-killed trees had stronger temperature-growth correlations in the early summer of the growing season than for still-living forestline trees. Long-term growth patterns may therefore be important to determining tree survival, at least within the forest- tundra ecotone. However, the mechanisms behind these observations are not clear, and should be considered for future studies examining insect disturbance.

In examining the post-outbreak radial growth patterns of surviving mature spruce, we identified several instances of sustained radial growth (i.e. growth releases) in neighbouring trees as a response to beetle-induced mortality. Until now, this effect has only been observed in closed-canopy forests and this study is therefore, to our knowledge, the first to demonstrate that post-outbreak growth releases can occur within forest-tundra ecotones. These growth releases occurred in stands that experienced high levels of spruce mortality, but were absent in survivors sampled from low-mortality stands. This 86

implies that, despite the relatively open nature of the canopy cover, spruce growing in alpine forest- tundra ecotones are affected by competition for resources and that remnant trees will benefit from the removal of dominant overstory individuals. The occurrence and extent of growth releases following an outbreak appear to be determined by stand characteristics, which influence both the degree of competition present in a stand and susceptibility to spruce bark beetle infestations.

Although a high level of spruce bark beetle kill may positively impact the growth of remnant trees and saplings, excessive thinning of cone-producing trees in the forest-tundra ecotone will greatly delay regeneration after an outbreak. Until existing saplings are able to produce viable cones, the reproductive capacity of a stand will largely be determined by the number of surviving mature spruce.

High-mortality stands with low numbers of saplings are therefore anticipated to take the longest to achieve pre-outbreak densities and basal areas. Likewise, the generally poor reproductive success of low-density spruce (O’Connell et al. 2006) and low seed germination rates in the forest-tundra ecotone

(Sirois 2000, Walker et al. 2012) should cause the least successful post-outbreak germination to occur in high-mortality stands. Seedling germination may also be restricted by competition with tall shrubs (e.g.

Salix spp., Betula glandulosa) at forestline, and/or be limited to stands where favourable germination habitats are provided by downed beetle-killed snags. Still, seedling abundance was found to be greatest in affected stands, with an estimated 50% of seedlings establishing in response to the outbreak. This suggests that affected stands in the forest-tundra ecotone will eventually recover, though it may take decades to do so.

The recent spruce bark beetle outbreak has caused substantial spruce mortality across the forest- tundra ecotones of southwest Yukon, but has not caused a widespread treeline recession. Instead, recession has been limited to locations where extensive thinning occurred at both treeline and forestline elevations. As a consequence, despite climate warming, treeline advance may not occur at locations

with considerable treeline mortality, though advance may simply be delayed at locations where fragmentation was restricted to lower elevations.

We recommend that future studies on insect disturbance in forest-tundra ecotones increase site- level replication. Our selection of six study sites did not prevent inferences on impacts or predisposing factors of spruce bark beetle disturbance in treeline environments, but did prevent the examination of interactions between elevation and mortality. In addition, due to constraints associated with the analytical technique used, we were unable to detect growth releases in trees after 2001. As such, we recommend the establishment of long-term monitoring in the area to determine the duration of growth releases in surviving spruce and its interactions with stand-level mortality, and to compare these with insect disturbances in closed-canopy forests. Finally, climate-growth analysis was limited by the short duration of instrumental records in the study region. This limitation is common across the Canadian arctic and subarctic (Prowse et al. 2009), and should be taken into account when interpreting dendrochronological analysis.

The implications of our findings are applicable to stands throughout southwest Yukon, as well as subarctic and subalpine forest-tundra ecotones. If warming temperatures allow for the expansion of bark beetle disturbances into regions previously deemed climatically unfavourable (Logan and Powell

2001, Carroll et al. 2004, Marini et al. 2012, Weed et al. 2013), an increasing number of treeline environments could become infested and subsequently fragmented and/or recessed. This research demonstrates the validity of using dendrochronological analysis to identify shared pre-outbreak growth trends that could contribute to attack susceptibility, and the benefits of conducting post-outbreak growth release analysis. As such, we recommend that future studies of insect outbreaks in treeline environments utilize similar methodologies. In addition, to build on the findings of our research and create a considerable knowledge base for the characteristics and effects of insect disturbance within

forest-tundra ecotones, we recommend the establishment of long-term monitoring sites in beetle- affected forest-tundra ecotone stands across southwest Yukon.

References

Alfaro, R. I., R. Campbell, P. Vera, B. Hawkes, and T. L. Shore. 2004. Dendroecological reconstruction of mountain pine beetle outbreaks in the Chilcotin Plateau of British Columbia. Pages 245–256 in T. L. Shore, J. E. Brooks, and J. E. Stone, editors. Mountain Pine Beetle Symposium: Challenges and Solutions. Information Report BC-X-399. Natural Resources Canada, Canadian Forest Service, Pacific Forestry Centre, Victoria, British Columbia, Canada.

Amoroso, M. M., D. K. Coates, and R. Astrup. 2013. Stand recovery and self-organization following large- scale mountain pine beetle induced canopy mortality in northern forests. Forest Ecology and Management 310:300–311.

Archibald, D. W., A. G. McAdam, S. Boutin, Q. E. Fletcher, and M. M. Humphries. 2012. Within-season synchrony of a masting conifer enhances seed escape. The American Naturalist 179:536–44.

Axelson, J. N., R. I. Alfaro, and B. C. Hawkes. 2010. Changes in stand structure in uneven-aged lodgepole pine stands impacted by mountain pine beetle epidemics and fires in central British Columbia. The Forestry Chronicle 86:87–99.

Ayres, M. P., and M. J. Lombardero. 2000. Assessing the consequences of global change for forest disturbance from herbivores and pathogens. The Science of the Total Environment 262:263– 286.

Baker, W. L. 1995. Longterm response of disturbance landscapes to human intervention and global change. Landscape Ecology 10:143–159.

Barber, V. A., G. P. Juday, and B. P. Finney. 2000. Reduced growth of Alaskan white spruce in the twentieth century from temperature-induced drought stress. Nature 405:668–673.

Barros, V. R., C. B. Field, D. J. Dokken, M. D. Mastrandrea, K. J. Mach, T. E. Bilir, M. Chatterjee, K. L. Ebi, Y. O. Estrada, R. C. Genova, B. Girma, E. S. Kissel, A. N. Levy, S. MacCracken, P. R. Mastrandrea, and L. L. White, editors. 2014. Climate Change 2014: Impacts, Adaptation, and Vulnerability. Part B: Regional Aspects. Contribution of Working Group II to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge University Press, Cambridge, UK.

Bebi, P., D. Kulakowski, and T. T. Veblen. 2003. Interactions between fire and spruce beetles in a subalpine Rocky Mountain forest landscape. Ecology 84:362–371.

Bentz, B. J., J. Régnière, C. J. Fettig, E. M. Hansen, J. L. Hayes, J. A. Hicke, R. G. Kelsey, J. F. Negrón, and S. J. Seybold. 2010. Climate change and bark beetles of the western United States and Canada: direct and indirect effects. BioScience 60:602–613.

Berg, E. E., J. D. Henry, C. L. Fastie, A. D. De Volder, and S. M. Matsuoka. 2006. Spruce beetle outbreaks on the Kenai Peninsula, Alaska, and Kluane National Park and Reserve, Yukon Territory:

relationship to summer temperatures and regional differences in disturbance regimes. Forest Ecology and Management 227:219–232.

Betts, R. A. 2000. Offset of the potential carbon sink from boreal forestation by decreases in surface albedo. Nature 408:187–190.

Biondi, F., and K. Waikul. 2004. DENDROCLIM2002: A C++ program for statistical calibration of climate signals in tree-ring chronologies. Computers & Geosciences 30:303–311.

Boggs, K., M. Sturdy, D. J. Rinella, and M. J. Rinella. 2008. White spruce regeneration following a major spruce beetle outbreak in forests on the Kenai Peninsula, Alaska. Forest Ecology and Management 255:3571–3579.

Boucher, T. V., and B. R. Mead. 2006. Vegetation change and forest regeneration on the Kenai Peninsula, Alaska following a spruce beetle outbreak, 1987–2000. Forest Ecology and Management 227:233–246.

Briffa, K. R., T. J. Osborn, F. H. Schweingruber, I. C. Harris, P. D. Jones, S. G. Shiyatov, and E. A. Vaganov. 2001. Low frequency temperature variations from a northern tree ring density network. Journal of Geophysical Research 106:2929–2941.

Brown, C. D., and J. F. Johnstone. 2012. Once burned, twice shy: repeat fires reduce seed availability and alter substrate constraints on Picea mariana regeneration. Forest Ecology and Management 266:34–41.

Butler, D. R., G. P. Malanson, L. M. Resler, S. J. Walsh, F. D. Wilkerson, G. L. Schmid, and C. F. Sawyer. 2009. Geomorphic patterns and processes at alpine treeline. Pages 63–84 in D. R. Butler, G. P. Malanson, S. J. Walsh, and D. B. Fagre, editors. The Changing Alpine Treeline: the Example of Glacier National Park, MT, USA. Elsevier, Amsterdam, the Netherlands.

Caccianiga, M., S. Payette, and L. Filion. 2008. Biotic disturbance in expanding subarctic forests along the eastern coast of Hudson Bay. The New Phytologist 178:823–834.

Cairns, D. M., C. W. Lafon, M. F. Mouton, R. L. Stuteville, A. B. Young, and J. Moen. 2012. Comparing two methods for aging trees with suppressed, diffuse-porous rings (Betula pubescens ssp. czerepanovii). Dendrochronologia 30:252–256.

Callaghan, T. V., R. M. M. Crawford, M. Eronen, A. Hofgaard, S. Payette, W. G. Rees, O. Skre, B. Sveinbjörnsson, T. K. Vlassova, and B. R. Werkman. 2002. The dynamics of the tundra-taiga boundary: an overview and suggested coordinated and integrated approach to research. AMBIO Special Report 12:3–5.

Carroll, A. L., S. W. Taylor, J. Régnière, and L. Safranyik. 2004. Effects of climate change on range expansion by the mountain pine beetle in British Columbia. Pages 223–232 in T. L. Shore, J. E. Brooks, and J. E. Stone, editors. Mountain Pine Beetle Symposium: Challenges and Solutions. Information Report BC-X-399. Natural Resources Canada, Canadian Forest Service, Pacific Forestry Centre, Victoria, British Columbia, Canada. 91

Chapin, F. S., M. Sturm, M. C. Serreze, J. P. McFadden, J. R. Key, A. H. Lloyd, A. D. McGuire, T. S. Rupp, A. H. Lynch, J. P. Schimel, J. Beringer, W. L. Chapman, H. E. Epstein, E. S. Euskirchen, L. D. Hinzman, G. Jia, C.-L. Ping, K. D. Tape, C. D. C. Thompson, D. A. Walker, and J. M. Welker. 2005. Role of land-surface changes in arctic summer warming. Science 310:657–660.

Chavardès, R. D., L. D. Daniels, P. O. Waeber, J. L. Innes, and C. R. Nitschke. 2013. Unstable climate−growth relations for white spruce in southwest Yukon, Canada. Climatic Change 116:593–611.

Chhin, S., and G. G. Wang. 2005. The effect of sampling height on dendroclimatic analysis. Dendrochronologia 23:47–55.

Chhin, S., and G. G. Wang. 2008. Climatic response of Picea glauca seedlings in a forest-prairie ecotone of western Canada. Annals of Forest Science 65:1–8.

Cook, E. R., and P. J. Krusic. 2005. Program ARSTAN: A tree-ring standardization program based on detrending and autoregressive time series modelling, with interactive graphics. Tree-Ring Laboratory, Lamont Doherty Earth Observatory, Columbia University, Palisades, New York, USA.

Coop, J. D., R. T. Massatti, and A. W. Schoettle. 2010. Subalpine vegetation pattern three decades after stand-replacing fire: effects of landscape context and topography on plant community composition, tree regeneration, and diversity. Journal of Vegetation Science 21:472–487.

Cullen, L. E., J. G. Palmer, R. P. Duncan, and G. H. Stewart. 2001. Climate change and tree-ring relationships of Nothofagus menziesii tree-line forests. Canadian Journal of Forest Research 31:1981–1991.

D’Arrigo, R., R. Wilson, B. Liepert, and P. Cherubini. 2008. On the “divergence problem” in northern forests: a review of the tree-ring evidence and possible causes. Global and Planetary Change 60:289–305.

Dale, V. H., L. A. Joyce, S. McNulty, R. P. Neilson, M. P. Ayres, M. D. Flannigan, P. J. Hanson, L. C. Irland, A. E. Lugo, C. J. Peterson, D. Simberloff, F. J. Swanson, B. J. Stocks, and B. M. Wotton. 2001. Climate change and forest disturbances. BioScience 51:723–734.

Danby, R. K., and D. S. Hik. 2007a. Evidence of recent treeline dynamics in southwest Yukon from aerial photographs. Arctic 60:411–420.

Danby, R. K., and D. S. Hik. 2007b. Responses of white spruce (Picea glauca) to experimental warming at a subarctic alpine treeline. Global Change Biology 13:437–451.

Danby, R. K., and D. S. Hik. 2007c. Variability, contingency and rapid change in recent subarctic alpine tree line dynamics. Journal of Ecology 95:352–363.

Danby, R. K., S. Koh, D. S. Hik, and L. W. Price. 2011. Four decades of plant community change in the alpine tundra of southwest Yukon, Canada. AMBIO 40:660–671.

Daniels, L. D., and T. T. Veblen. 2004. Spatiotemporal influences of climate on altitudinal treeline in northern Patagonia. Ecology 85:1284–1296.

Daniels, L. D., T. T. Veblen, and R. Villalba. 2007. Use of thin sections to improve age estimates of Nothofagus pumilio seedlings. Ecoscience 14:17–22.

De Lafontaine, G., and S. Payette. 2010. The origin and dynamics of subalpine white spruce and balsam fir stands in boreal eastern North America. Ecosystems 13:932–947.

DeRose, R. J., and J. N. Long. 2010. Regeneration response and seedling bank dynamics on a Dendroctonus rufipennis–killed Picea engelmannii landscape. Journal of Vegetation Science 21:377–387.

DeRose, R. J., and J. N. Long. 2012. Factors influencing the spatial and temporal dynamics of Engelmann spruce mortality during a spruce beetle outbreak on the Markagunt Plateau, Utah. Forest Science 58:1–14.

Devi, N., F. Hagedorn, P. Moiseev, H. Bugmann, S. Shiyatov, V. Mazepa, and A. Rigling. 2008. Expanding forests and changing growth forms of Siberian larch at the Polar Urals treeline during the 20th century. Global Change Biology 14:1581–1591.

Doak, P. 2004. The impact of tree and stand characteristics on spruce beetle (Coleoptera: Scolytidae) induced mortality of white spruce in the Copper River Basin, Alaska. Canadian Journal of Forest Research 34:810–816.

Douglas, G. W. Montane zone vegetation of the Alsek River region, southwestern Yukon. Canadian Journal of Botany 52:2505–2535.

Drury, S. A., and P. J. Grissom. 2008. Fire history and fire management implications in the Yukon Flats National Wildlife Refuge, interior Alaska. Forest Ecology and Management 256:304–312.

Eckstein, D., J. Hoogesteger, and R. L. Holmes. 1991. Insect-related differences in growth of birch and pine at northern treeline in Swedish Lapland. Ecography 14:18–23.

Eisenhart, K. S., and T. T. Veblen. 2000. Dendroecological detection of spruce bark beetle outbreaks in northwestern Colorado. Canadian Journal of Forest Research 30:1788–1798.

Environment Canada. 2014. Canadian Climate Normals 1981-2010: Burwash Landing, Yukon. Available at http://climate.weather.gc.ca/climate_normals/index_e.html.

Ettinger, A. K., K. R. Ford, and J. HilleRisLambers. 2011. Climate determines upper, but not lower, altitudinal range limits of Pacific Northwest conifers. Ecology 92:1323–1331.

Filion, L., S. Payette, A. Delwaide, and N. Bhiry. 1998. Insect defoliators as major disturbance factors in the high-altitude balsam fir forest of Mount Mégantic, southern Quebec. Canadian Journal of Forest Research 28:1832-1842.

Flannigan, M. D., K. A. Logan, B. D. Amiro, W. R. Skinner, and B. J. Stocks. 2005. Future area burned in Canada. Climatic Change 72:1–16.

Flower, A., and D. J. Smith. 2011. A dendroclimatic reconstruction of June–July mean temperature in the northern Canadian Rocky Mountains. Dendrochronologia 29:55–63.

Foley, J. A., J. E. Kutzbach, M. T. Coe, and S. Levis. 1994. Feedbacks between climate and boreal forests during the Holocene epoch. Nature 371:52–54.

Francis, S. R. 1996. Linking landscape pattern and forest disturbance: fire history of the Shakwak Trench, southwest Yukon Territory. M.Sc. Thesis. University of Alberta, Edmonton, Canada.

Frei, E., J. Bodin, and G.-R. Walther. 2010. Plant species’ range shifts in mountainous areas—all uphill from here? Botanica Helvetica 120:117–128.

Fritts, H. C. 1976. Tree Rings and Climate. Academic Press, London, UK.

Fritts, H.C., J. Mosimann, and C. Bottorff. 1969. A revised computer program for standardizing tree-ring series. Tree-Ring Bulletin 29: 15-20.

Furniss, R. L., and V. M. Carolin. 1977. Western forest insects. Miscellaneous Publication No. 1339. U.S. Department of Agriculture, Forest Service, Washington DC, USA.

Gamache, I., and S. Payette. 2004. Height growth response of tree line black spruce to recent climate warming across the forest-tundra of eastern Canada. Journal of Ecology 92:835–845.

Garbutt, R., B. Hawkes, and E. Allen. 2006. Spruce beetle and the forests of the southwest Yukon. Information Report BC-X-406. Natural Resources Canada, Canadian Forest Service, Pacific Forestry Centre, Victoria, British Columbia, Canada.

Gärtner, S. M., V. J. Lieffers, and S. E. Macdonald. 2011. Ecology and management of natural regeneration of white spruce in the boreal forest. Environmental Reviews 478:461–478.

Gilbert, M., G. Vouland, and J.-C. Grégoire. 2001. Past attacks influence host selection by the solitary bark beetle Dendroctonus micans. Ecological Entomology 26:133–142.

Government of Yukon. 2013. Forest Health Report. Yukon Energy, Mines and Resources, Forest Management Branch, Whitehorse, Yukon, Canada. Available at http://www.emr.gov.yk.ca/forestry/pdf/forest_health_report_2013.pdf.

Grace, J., F. Berninger, and L. Nagy. 2002. Impacts of climate change on the tree line. Annals of Botany 90:537–544.

Griesbauer, H. P., and D. S. Green. 2012. Geographic and temporal patterns in white spruce climate– growth relationships in Yukon, Canada. Forest Ecology and Management 267:215–227.

Gutsell, S. L., and E. A. Johnson. 2002. Accurately ageing trees and examining their height-growth rates: implications for interpreting forest dynamics. Journal of Ecology 90:153–166.

Haeberli, W., and S. Gruber. 2009. Global warming and mountain permafrost. Pages 205–218 in R. Margesin, editor. Permafrost Soils. Springer-Verlag, Berlin, Germany.

Hard, J. S. 1985. Spruce beetles attack slowly growing spruce. Forest Science 31:839–850.

Hard, J. S. 1989. Sequence of trees attacked by spruce beetles in a mature even-aged spruce stand in south-central Alaska. Northwest Science 63:5–12.

Harden, J. W., S. E. Trumbore, B. J. Stocks, A. Hirsch, S. T. Gower, K. P. O’Neill, and E. S. Kasischke. 2000. The role of fire in the boreal carbon budget. Global Change Biology 6:174–184.

Harsch, M. A., and M. Y. Bader. 2011. Treeline form – a potential key to understanding treeline dynamics. Global Ecology and Biogeography 20:582–596.

Harsch, M. A., P. E. Hulme, M. S. McGlone, and R. P. Duncan. 2009. Are treelines advancing? A global meta-analysis of treeline response to climate warming. Ecology Letters 12:1040–1049.

Hart, S. J., T. T. Veblen, and D. Kulakowski. 2014. Do tree and stand-level attributes determine susceptibility of spruce-fir forests to spruce beetle outbreaks in the early 21st century? Forest Ecology and Management 318:44–53.

Helama, S., M. Lindholm, M. Timonen, and M. Eronen. 2004. Detection of climate signal in dendrochronological data analysis: a comparison of tree-ring standardization methods. Theoretical and Applied Climatology 79:239–254.

Hinds, T. E., F. G. Hawksworth, and R. W. Davidson. 1965. Beetle-killed Engelmann spruce: its deterioration in Colorado. Journal of Forestry 63:536–542.

Hofgaard, A., and G. Rees, editors. 2008. PPS Arctic Manual: Common protocols for field measurements and handling of collected material. Available at http://ppsarctic.nina.no/files/PPS%20Arctic%20Manual.pdf.

Hogg, E. H., and R. W. Wein. 2005. Impacts of drought on forest growth and regeneration following fire in southwestern Yukon, Canada. Canadian Journal of Forest Research 35:2141–2150.

Holling, C. S. 1973. Resilience and stability of ecological systems. Annual Review of Ecology and Systematics 3:1–23.

Holmes, R. L. 1983. Computer-assisted quality control in tree-ring dating and measurement. Tree-Ring Bulletin 43:69–78.

Holsten, E. H., R. W. Their, A. S. Munson, and K. E. Gibson. 1999. The spruce beetle. Forest Insect and Disease Leaflet No. 127. U.S. Department of Agriculture, Forest Service, Washington DC, USA.

Holtmeier, F.-K., and G. Broll. 2005. Sensitivity and response of northern hemisphere altitudinal and polar treelines to environmental change at landscape and local scales. Global Ecology and Biogeography 14:395–410.

Holtmeier, F.-K., and G. Broll. 2010. Wind as an ecological agent at treelines in North America, the Alps, and the European Subarctic. Physical Geography 31:203–233.

Jacoby, G. C., and R. D. D’Arrigo. 1995. Tree ring width and density evidence of climatic and potential forest change in Alaska. Global Biogeochemical Cycles 9:227–234.

Jenkins, M. J., E. Hebertson, W. Page, and C. A. Jorgensen. 2008. Bark beetles, fuels, fires and implications for forest management in the Intermountain West. Forest Ecology and Management 254:16–34.

Johnstone, J. F., T. N. Hollingsworth, F. S. Chapin, and M. C. Mack. 2010a. Changes in fire regime break the legacy lock on successional trajectories in Alaskan boreal forest. Global Change Biology 16:1281–1295.

Johnstone, J. F., E. J. B. McIntire, E. J. Pedersen, G. King, and M. F. J. Pisaric. 2010b. A sensitive slope: estimating landscape patterns of forest resilience in a changing climate. Ecosphere 1:167–180.

Jones, P. D., T. J. Osborn, and K. R. Briffa. 2001. The evolution of climate over the last millennium. Science 292:662–667.

Kausrud, K., B. Okland, O. Skarpaas, J.-C. Grégoire, N. Erbilgin, and N. C. Stenseth. 2012. Population dynamics in changing environments: the case of an eruptive forest pest species. Biological Reviews of the Cambridge Philosophical Society 87:34–51.

Körner, C. 1998. A re-assessment of high elevation treeline positions and their explanation. Oecologia 115:445–459.

Körner, C., and J. Paulsen. 2004. A world-wide study of high altitude treeline temperatures. Journal of Biogeography 31:713–732.

Kramer, P. J., and T. T. Kozlowski. 1979. Physiology of Woody Plants. Academic Press, San Diego, California, USA.

Krebs, C. J., and R. Boonstra. 2001. The Kluane Region. Pages 9–24 in C. J. Krebs, S. Boutin, and R. Boonstra, editors. Ecosystem Dynamics of the Boreal Forest. Oxford University Press, New York, New York, USA.

Krebs, C. J., J. M. LaMontagne, A. J. Kenney, and S. Boutin. 2012. Climatic determinants of white spruce cone crops in the boreal forest of southwestern Yukon. Botany 119:113–119.

Kulakowski, D., T. T. Veblen, and P. Bebi. 2003. Effects of fire and spruce beetle outbreak legacies on the disturbance regime of a subalpine forest in Colorado. Journal of Biogeography 30:1445–1456.

Kulman, H. M. 1971. Effects of insect defoliation on growth and mortality of trees. Annual Review of Entomology 16:289–324.

Kurz, W. A., C. C. Dymond, G. Stinson, G. J. Rampley, E. T. Neilson, A. L. Carroll, T. Ebata, and L. Safranyik. 2008. Mountain pine beetle and forest carbon feedback to climate change. Nature 452:987–990.

Leonelli, G., M. Pelfini, U. Morra di Cella, and V. Garavaglia. 2011. Climate warming and the recent treeline shift in the European Alps: the role of geomorphological factors in high-altitude sites. AMBIO 40:264–273.

Lescop-Sinclair, K., and S. Payette. 1995. Recent advance of the arctic treeline along the eastern coast of Hudson Bay. Journal of Ecology 83:929–936.

Lloyd, A. H. 2005. Ecological histories from Alaskan tree lines provide insight into future change. Ecology 86:1687–1695.

Lloyd, A. H., and A. G. Bunn. 2007. Responses of the circumpolar boreal forest to 20th century climate variability. Environmental Research Letters 2:045013.

Lloyd, A. H., and C. L. Fastie. 2002. Spatial and temporal variability in the growth and climate response of treeline trees in Alaska. Climatic Change 52:481–509.

Lloyd, A. H., and C. L. Fastie. 2003. Recent changes in treeline forest distribution and structure in interior Alaska. Ecoscience 10:176–185.

Logan, J. A., and J. A. Powell. 2001. Ghost forests, global warming, and the mountain pine beetle (Coleoptera: Scolytidae). American Entomologist 47:160–172.

Logan, J. A., and J. A. Powell. 2009. Ecological consequences of climate change altered forest insect disturbance regimes. Pages 98–109 in F. H. Wagner, editor. Climate Warming in Western North America: Evidence and Environmental Effects. University of Utah Press, Salt Lake City, Utah, USA.

Logan, J. A., J. Régnière, and J. A. Powell. 2003. Assessing the impacts of global warming on forest pest dynamics. Frontiers in Ecology and the Environment 1:130–137.

MacDonald, G. M., J. M. Szeicz, J. Claricoates, and K. A. Dale. 1998. Response of the central Canadian treeline to recent climatic changes. Annals of the Association of American Geographers 88:183– 208.

Malmström, C. M., and K. F. Raffa. 2000. Biotic disturbance agents in the boreal forest: considerations for vegetation change models. Global Change Biology 6:36–48.

Mamet, S. D., and G. P. Kershaw. 2013. Age-dependency, climate, and environmental controls of recent tree growth trends at subarctic and alpine treelines. Dendrochronologia 31:75–87.

Mann, M. E., R. S. Bradley, and M. K. Hughes. 1999. Northern hemisphere temperatures during the past millennium: inferences, uncertainties, and limitations. Geophysical Research Letters 26:759– 762.

Marini, L., M. P. Ayres, A. Battisti, and M. Faccoli. 2012. Climate affects severity and altitudinal distribution of outbreaks in an eruptive bark beetle. Climatic Change 115:327–341.

Maroja, L. S., S. M. Bogdanowicz, K. F. Wallin, K. F. Raffa, and R. G. Harrison. 2007. Phylogeography of spruce beetles (Dendroctonus rufipennis Kirby) (Curculionidae: Scolytinae) in North America. Molecular Ecology 16:2560–2573.

Massey, C., and N. Wygant. 1954. Biology and control of the Engelmann spruce beetle in Colorado. Circular No. 944. U.S. Department of Agriculture, Forest Service, Washington DC, USA.

Mayer, H., M. R. Raupach, W. Kohsiek, B. Gardiner, J. A. Clarke, R. Amtmann, J. M. Crowther, P. G. Jarvis, and R. Milne. 1989. Windthrow [and discussion]. Philosophical Transactions of the Royal Society B 324:267–281.

McCullough, D. G., R. A. Werner, and D. Neumann. 1998. Fire and insects in northern and boreal forest ecosystems of North America. Annual Review of Entomology 43:107–127.

McLane, S. C., L. D. Daniels, and S. N. Aitken. 2011. Climate impacts on lodgepole pine (Pinus contorta) radial growth in a provenance experiment. Forest Ecology and Management 262:115–123.

Miyamoto, Y., H. P. Griesbauer, and D. S. Green. 2010. Growth responses of three coexisting conifer species to climate across wide geographic and climate ranges in Yukon and British Columbia. Forest Ecology and Management 259:514–523.

Moen, J., D. M. Cairns, and C. W. Lafon. 2008. Factors structuring the treeline ecotone in Fennoscandia. Plant Ecology & Diversity 1:77–87.

Murdock, T. Q., S. W. Taylor, A. Flower, A. Mehlenbacher, A. Montenegro, F. W. Zwiers, R. Alfaro, and D. L. Spittlehouse. 2013. Pest outbreak distribution and forest management impacts in a changing climate in British Columbia. Environmental Science & Policy 26:75–89.

Myers-Smith, I. H., B. C. Forbes, M. Wilmking, M. Hallinger, T. Lantz, D. Blok, K. D. Tape, M. Macias- Fauria, U. Sass-Klaassen, E. Lévesque, S. Boudreau, P. Ropars, L. Hermanutz, A. Trant, L. S. Collier, S. Weijers, J. Rozema, S. A. Rayback, N. M. Schmidt, G. Schaepman-Strub, S. Wipf, C. Rixen, C. B. Ménard, S. Venn, S. Goetz, L. Andreu-Hayles, S. Elmendorf, V. Ravolainen, J. Welker, P. Grogan, H. E. Epstein, and D. S. Hik. 2011a. Shrub expansion in tundra ecosystems: dynamics, impacts and research priorities. Environmental Research Letters 6:045509.

Myers-Smith, I. H., D. S. Hik, C. Kennedy, D. Cooley, J. F. Johnstone, A. J. Kenney, and C. J. Krebs. 2011b. Expansion of canopy-forming willows over the twentieth century on Herschel Island, Yukon Territory, Canada. AMBIO 40:610–623.

Nogués-Bravo, D., M. B. Araújo, M. P. Errea, and J. P. Martínez-Rica. 2007. Exposure of global mountain systems to climate warming during the 21st Century. Global Environmental Change 17:420–428.

Nowacki, G. J., and M. D. Abrams. 1997. Radial-growth averaging criteria for reconstructing disturbance histories from presettlement-origin oaks. Ecological Monographs 67:225–249.

O’Connell, L. M., A. Mosseler, and O. P. Rajora. 2006. Impacts of forest fragmentation on the reproductive success of white spruce (Picea glauca). Canadian Journal of Botany 965:956–965.

Ohse, B., F. Jansen, and M. Wilmking. 2012. Do limiting factors at Alaskan treelines shift with climatic regimes? Environmental Research Letters 7:015505.

Parish, R., J. A. Antos, and M.-J. Fortin. 1999. Stand development in an old-growth subalpine forest in southern interior British Columbia. Canadian Journal of Forest Research 29:1347–1356.

Parmesan, C. 2006. Ecological and evolutionary responses to recent climate change. Annual Review of Ecology, Evolution, and Systematics 37:637–669.

Pauli, H., M. Gottfried, K. Reiter, C. Klettner, and G. Grabherr. 2007. Signals of range expansions and contractions of vascular plants in the high Alps: observations (1994-2004) at the GLORIA master site Schrankogel, Tyrol, Austria. Global Change Biology 13:147–156.

Paulsen, J., U. M. Weber, and C. Körner. 2000. Tree growth near treeline: Abrupt or gradual reduction with altitude? Arctic, Antarctic, and Alpine Research 32:14–20.

Payette, S., and L. Filion. 1985. White spruce expansion at the tree line and recent climate change. Canadian Journal of Forest Research 15:241–251.

Payette, S., L. Filion, and A. Delwaide. 2008. Spatially explicit fire-climate history of the boreal forest- tundra (Eastern Canada) over the last 2000 years. Philosophical Transactions of the Royal Society B 363:2301–2316.

Payette, S., M.-J. Fortin, and I. Gamache. 2001. The subarctic forest–tundra: the structure of a biome in a changing climate. BioScience 51:709–718.

Porter, T. J., and M. F. J. Pisaric. 2011. Temperature-growth divergence in white spruce forests of Old Crow Flats, Yukon Territory, and adjacent regions of northwestern North America. Global Change Biology 17:3418–3430.

Prowse, T. D., C. Furgal, B. R. Bonsal, and T. W. D. Edwards. 2009. Climatic conditions in northern Canada: past and future. AMBIO 38:257–265.

Pureswaran, D. S., and J. H. Borden. 2005. Primary attraction and kairomonal host discrimination in three species of Dendroctonus (Coleoptera: Scolytidae). Agricultural and Forest Entomology 7:219–230.

Raffa, K. F., B. H. Aukema, B. J. Bentz, A. L. Carroll, J. A. Hicke, M. G. Turner, and W. H. Romme. 2008. Cross-scale drivers of natural disturbances prone to anthropogenic amplification: the dynamics of bark beetle eruptions. BioScience 58:501–517.

Raffa, K. F., B. H. Aukema, N. Erbilgin, K. D. Klepzig, and K. F. Wallin. 2005. Interactions among conifer terpenoids and bark beetles across multiple levels of scale: an attempt to understand links between population patterns and physiological processes. Recent Advances in Phytochemistry 39:79–118.

Rammig, A., T. Jonas, N. E. Zimmermann, and C. Rixen. 2010. Changes in alpine plant growth under future climate conditions. Biogeosciences 7:2013–2024.

Rebetez, M., and M. Reinhard. 2007. Monthly air temperature trends in Switzerland 1901–2000 and 1975–2004. Theoretical and Applied Climatology 91:27–34.

Reynolds, K. M., and E. H. Holsten. 1994. Classification of spruce beetle hazard in Lutz spruce (Picea x lutzii) stands on the Kenai Peninsula, Alaska. Canadian Journal of Forest Research 24:1015–1021.

Rich, R. L., L. E. Frelich, and P. B. Reich. 2007. Wind-throw mortality in the southern boreal forest: effects of species, diameter and stand age. Journal of Ecology 95:1261–1273.

Rolland, C., and G. Lempérière. 2004. Effects of climate on radial growth of Norway spruce and interactions with attacks by the bark beetle Dendroctonus micans (Kug., Coleoptera: Scolytidae): a dendroecological study in the French Massif Central. Forest Ecology and Management 201:89– 104.

Romme, W. H., D. H. Knight, and J. B. Yavitt. 1986. Mountain pine beetle outbreaks in the Rocky Mountains: regulators of primary productivity? The American Naturalist 127:484–494.

Ropars, P., and S. Boudreau. 2012. Shrub expansion at the forest-tundra ecotone: spatial heterogeneity linked to local topography. Environmental Research Letters 7:015501.

Rowe, J. S., and G. W. Scotter. 1973. Fire in the boreal forest. Quaternary Research 3:444–464.

Rubino, D. L., and B. C. McCarthy. 2004. Comparative analysis of dendroecological methods used to assess disturbance events. Dendrochronologia 21:97–115.

Safranyik, L., and B. Wilson, editors. 2006. The Mountain Pine Beetle: A Synthesis of Biology, Management, and Impacts on Lodgepole Pine. Natural Resources Canada, Canadian Forest Service, Pacific Forestry Centre, Victoria, British Columbia, Canada.

Salzer, M. W., M. K. Hughes, A. G. Bunn, and K. F. Kipfmueller. 2009. Recent unprecedented tree-ring growth in bristlecone pine at the highest elevations and possible causes. Proceedings of the National Academy of Sciences of the United States of America 106:20348–20353.

100

Scheller, R. M., and D. J. Mladenoff. 2005. A spatially interactive simulation of climate change, harvesting, wind, and tree species migration and projected changes to forest composition and biomass in northern Wisconsin, USA. Global Change Biology 11:307–321.

Schmid, J. M., and R. H. Frye. 1977. Spruce beetle in the Rockies. General Technical Report RM-49. U.S. Department of Agriculture, Forest Service, Rocky Mountain Forest and Range Experiment Station, Fort Collins, Colorado, USA.

Schmid, J. M., and T. E. Hinds. 1974. Development of spruce-fir stands following spruce beetle outbreaks. Research Paper RM-131. U.S. Department of Agriculture, Forest Service, Rocky Mountain Forest and Range Experiment Station, Fort Collins, Colorado, USA.

Scott, P. A., R. I. C. Hansell, and D. C. F. Fayle. 1987. Establishment of white spruce populations and responses to climate change at the treeline, Churchill, Manitoba, Canada. Arctic and Alpine Research 19:45–51.

Scudder, G. G. E. 1997. Environment of the Yukon. Pages 13–57 in H. V. Danks and J. A. Downes, editors. Insects of the Yukon. Biological Survey of Canada (Terrestrial Arthropods), Ottawa, Ontario, Canada.

Sherriff, R. L., E. E. Berg, and A. E. Miller. 2011. Climate variability and spruce beetle (Dendroctonus rufipennis) outbreaks in south-central and southwest Alaska. Ecology 92:1459–1470.

Simard, M., and S. Payette. 2005. Reduction of black spruce seed bank by spruce budworm infestation compromises postfire stand regeneration. Canadian Journal of Forest Research 35:1686–1696.

Sirois, L. 2000. Spatiotemporal variation in black spruce cone and seed crops along a boreal forest – tree line transect. Canadian Journal of Forest Research 30:900–909.

Smith, J. M., S. J. Hart, T. B. Chapman, T. T. Veblen, and T. Schoennagel. 2012. Dendroecological reconstruction of 1980s mountain pine beetle outbreak in lodgepole pine forests in northwestern Colorado. Ecoscience 19:113–126.

Smith, W. K., M. J. Germino, T. E. Hancock, and D. M. Johnson. 2003. Another perspective on altitudinal limits of alpine timberlines. Tree Physiology 23:1101–1112.

Soja, A. J., N. M. Tchebakova, N. H. F. French, M. D. Flannigan, H. H. Shugart, B. J. Stocks, A. I. Sukhinin, E. I. Parfenova, F. S. Chapin, and P. W. Stackhouse. 2007. Climate-induced boreal forest change: predictions versus current observations. Global and Planetary Change 56:274–296.

Sokal, R. R., and F. J. Rohlf. 2009. Introduction to biostatistics. Dover Publications Inc., Mineola, New York, USA.

Speer, J. 2010. Fundamentals of Tree Ring Research. University of Arizona Press, Tucson, Arizona, USA.

101

St. George, S., T. R. Ault, and M. C. A. Torbenson. 2013. The rarity of absent growth rings in Northern Hemisphere forests outside the American Southwest. Geophysical Research Letters 40:3727– 3731.

Stevens, G. C., and J. F. Fox. 1991. The causes of treeline. Annual Review of Ecology and Systematics 22:177–191.

Stocker, T. F., D. Qin, G.-K. Plattner, M. Tignor, S. K. Allen, J. Boschung, A. Nauels, Y. Xia, V. Bex, and P. M. Midgley, editors. 2013. Climate Change 2013: The Physical Science Basis. Contribution of Working Group I to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge University Press, Cambridge, UK.

Stocks, B. J., M. A. Fosberg, T. J. Lynham, L. Mearns, B. M. Wotton, Q. Yang, J.-Z. Jin, K. Lawrence, G. R. Hartley, J. A. Mason, and D. W. McKenney. 1998. Climate change and forest fire potential in Russian and Canadian boreal forests. Climatic Change 38:1–13.

Stokes, M. A., and T. L. Smiley. 1968. An introduction to tree ring dating. University of Chicago Press, Chicago, Illinois, USA.

Stone, W. E., and M. L. Wolfe. 1996. Response of understory vegetation to variable tree mortality following a mountain pine beetle epidemic in lodgepole pine stands in northern Utah. Vegetatio 122:1–12.

Stueve, K. M., R. E. Isaacs, L. E. Tyrrell, and R. V. Densmore. 2011. Spatial variability of biotic and abiotic tree establishment constraints across a treeline ecotone in the Alaska Range. Ecology 92:496– 506.

Sturm, M., C. Racine, and K. Tape. 2001. Increasing shrub abundance in the Arctic. Nature 411:546–547.

Suarez, F., D. Binkley, M. W. Kaye, and R. Stottlemyer. 1999. Expansion of forest stands into tundra in the Noatak National Preserve, northwest Alaska. Ecoscience 6:465–470.

Sveinbjörnsson, B. 2000. North American and European treelines: external forces and internal processes controlling position. AMBIO 29:388–395.

Symon, C., L. Arris, and B. Heal, editors. 2005. Arctic Climate Impact Assessment. Cambridge University Press, New York, New York, USA.

Szeicz, J. M., and G. M. MacDonald. 1994. Age-dependent tree-ring growth responses of subarctic white spruce to climate. Canadian Journal of Forest Research 24:120–132.

Szeicz, J. M., and G. M. MacDonald. 1995. Recent white spruce dynamics at the subarctic alpine treeline of north-western Canada. Journal of Ecology 83:873–885.

Tape, K., M. Sturm, and C. Racine. 2006. The evidence for shrub expansion in Northern Alaska and the Pan-Arctic. Global Change Biology 12:686–702.

102

Tranquillini, W. 1979. Physiological Ecology of the Alpine Timberline. Springer-Verlag, Berlin, Germany.

Trindade, M., T. Bell, C. P. Laroque, J. D. Jacobs, and L. Hermanutz. 2011. Dendroclimatic response of a coastal alpine treeline ecotone: a multispecies perspective from Labrador. Canadian Journal of Forest Research 41:469–478.

Turner, M. G., R. H. Gardner, and R. V. O’Neill. 2001. Landscape Ecology in Theory and Practice: Pattern and Process. Springer, New York, New York, USA.

Van Bogaert, R., K. Haneca, J. Hoogesteger, C. Jonasson, M. De Dapper, and T. V. Callaghan. 2011. A century of tree line changes in sub-Arctic Sweden shows local and regional variability and only a minor influence of 20th century climate warming. Journal of Biogeography 38:907–921.

Veblen, T. T., K. S. Hadley, M. S. Reid, and A. J. Rebertus. 1991a. The response of subalpine forests to spruce beetle outbreak in Colorado. Ecology 72:213–231.

Veblen, T. T., K. S. Hadley, M. S. Reid, and A. J. Rebertus. 1991b. Methods of detecting past spruce beetle outbreaks in Rocky Mountain subalpine forests. Canadian Journal of Forest Research 21:242– 254.

Vittoz, P., B. Rulence, T. Largey, and F. Freléchoux. 2008. Effects of climate and land-use change on the establishment and growth of cembran pine (Pinus cembra L.) over the altitudinal treeline ecotone in the Central Swiss Alps. Arctic, Antarctic, and Alpine Research 40:225–232.

Walker, X., G. H. R. Henry, K. McLeod, and A. Hofgaard. 2012. Reproduction and seedling establishment of Picea glauca across the northernmost forest-tundra region in Canada. Global Change Biology 18:3202–3211.

Wallin, K. F., and K. F. Raffa. 2004. Feedback between individual host selection behavior and population dynamics in an eruptive herbivore. Ecological Monographs 74:101–116.

Walther, G.-R., S. Beißner, and C. A. Burga. 2005. Trends in the upward shift of alpine plants. Journal of Vegetation Science 16:541–548.

Weed, A. S., M. P. Ayres, and J. A. Hicke. 2013. Consequences of climate change for biotic disturbances in North American forests. Ecological Monographs 83:441–470.

Werner, R. A., and E. H. Holsten. 1985. Factors influencing generation times of spruce beetles in Alaska. Canadian Journal of Forest Research 15:438–443.

Werner, R. A., E. H. Holsten, S. M. Matsuoka, and R. E. Burnside. 2006. Spruce beetles and forest ecosystems in south-central Alaska: a review of 30 years of research. Forest Ecology and Management 227:195–206.

White, P. S., and S. T. A. Pickett. 1985. Natural disturbance and patch dynamics: an introduction. Pages 3–13 in S. T. A. Pickett and P. S. White, editors. The Ecology of Natural Disturbance and Patch Dynamics. Academic Press, San Diego, California, USA. 103

Wigley, T. M. L., K. R. Briffa, and P. D. Jones. 1984. On the average value of correlated time series, with applications in dendroclimatology and hydrometeorology. Journal of Climate and Applied Meteorology 23:201–213.

Wilmking, M., R. D’Arrigo, G. C. Jacoby, and G. P. Juday. 2005. Increased temperature sensitivity and divergent growth trends in circumpolar boreal forests. Geophysical Research Letters 32:L15715.

Wilmking, M., M. Hallinger, R. Van Bogaert, T. Kyncl, F. Babst, W. Hahne, G. P. Juday, M. de Luis, K. Novak, and C. Völlm. 2012. Continuously missing outer rings in woody plants at their distributional margins. Dendrochronologia 30:213–222.

Youngblut, D., and B. Luckman. 2008. Maximum June–July temperatures in the southwest Yukon over the last 300 years reconstructed from tree rings. Dendrochronologia 25:153–166.

Zalatan, R., and K. Gajewski. 2005. Tree-ring analysis of five Picea glauca–dominated sites from the interior boreal forest in the Shakwak Trench, Yukon Territory, Canada. Polar Geography 29:1–16.

Zhang, Q., R. I. Alfaro, and R. J. Hebda. 1999. Dendroecological studies of tree growth, climate and spruce beetle outbreaks in Central British Columbia, Canada. Forest Ecology and Management 121:215–225.

Zhang, Y., X. Shao, and M. Wilmking. 2011. Dynamic relationships between Picea crassifolia growth and climate at upper treeline in the Qilian Mts., Northeast Tibetan Plateau, China. Dendrochronologia 29:185–199.

104

Appendix A Summaries of Site Characteristics

Table A-1 Description of Site 1 (Mount Decoeli – low mortality).

Krummholz line Treeline Forestline Plot 1 Plot 2 Plot 3 Plot 1 Plot 2 Plot 3 Plot 1 Plot 2 Plot 3 Elevation (masl) 1209 1216 1213 1153 1149 1148 1087 1097 1100 Slope (°) 8 9 9 13 13 12 5 7 14 Aspect (°) 67 59 50 67 70 67 32 42 68 Northing (NAD83) 6749009 6749058 6749101 6749318 6749364 6749407 6749482 6749518 6749563 Easting (NAD83) 0345901 0345878 0345856 0346472 0346456 0346435 0346936 0346899 0346874

Shrubs Cover (%) 90.0 79.1 76.4

105 Dominance (% willow) 49.2 32.4 76.0 Mean height (m) 61.4 ± 34.8 55.5 ± 34.0 171.9 ± 90.3

Spruce Seedling density (/ ha) 0.0 0.0 100.0 Mean height (m) n/a 2.9 2.3 3.7 4.8 5.7 9.1 Height range (m) n/a 1.1 –4.1 0.6 – 4.4 1.8 – 5.9 1.5 – 12.3 0.6 – 9.6 5.1 – 13.4 Mean DRC (cm) n/a 10.8 12.1 21.8 14.4 21.7 32.8 DRC range (cm) n/a 4.9 – 2.2 4.4 – 28.8 7.6 – 67.5 2.0 – 54.5 0.8 – 51.5 17.3 – 55.4 Mean DBH (cm) n/a 6.5 4.6 9.0 10.9 18.4 25.5 DBH range (cm) n/a 2.2 – 14.0 0.1 – 12.2 2.3 – 18.3 0.8 – 49.0 7.4 – 38.1 10.8 – 42.1 Saplings (%) n/a 16.7 41.7 16.7 8.3 16.7 0.0 Density (/ ha): pre-outbreak n/a 35.2 96.5 22.7 309.5 968.5 260.9 Density (/ ha): post-outbreak n/a 35.2 96.5 22.7 309.5 645.7 239.2 Basal area (m2/ha): pre-outbreak n/a 0.4 1.5 1.2 11.6 52.8 22.7 Basal area (m2/ha): post-outbreak n/a 0.4 1.5 1.2 11.6 17.5 22.7 Stand-level mortality (%) n/a 0.0 0.0 0.0 0.0 40.0 0.0

Table A-2 Description of Site 2 (Boutellier Creek – high mortality).

Krummholz line Treeline Forestline Plot 1 Plot 2 Plot 3 Plot 1 Plot 2 Plot 3 Plot 1 Plot 2 Plot 3 Elevation (masl) 1272 1264 1265 1205 1195 1192 1129 1130 1125 Slope (°) 8 7 8 7 6 6 7 6 3 Aspect (°) 56 60 60 74 78 82 63 66 79 Northing (NAD83) 6760179 6760219 6760260 6761001 6761047 6761095 6761628 6761667 6761704 Easting (NAD83) 0648869 0648840 0648814 0649166 0649164 0649146 0649541 0649521 0649486

Shrubs Cover (%) 74.5 90.0 40.9 Dominance (% willow) 42.2 63.4 58.3 Mean height (m) 101.8 ± 49.1 134.3 ± 69.5 88.2 ± 51.8

Spruce Seedling density (/ ha) 0.0 50.0 1050.0

106 Mean height (m) n/a 6.3 6.0 6.2 5.3 4.8 8.6 Height range (m) n/a 0.6 – 16.9 1.1 – 13.3 0.5 – 13.9 0.9 – 11.9 1.4 – 10.3 1.0 – 14.0 Mean DRC (cm) n/a 40.5 40.3 34.9 19.9 18.3 35.5 DRC range (cm) n/a 1.3 – 120.0 2.4 – 80.7 1.8 – 83.5 1.2 – 43.0 3.7 – 45.0 2.0 – 58.2 Mean DBH (cm) n/a 33.9 35.0 26.4 15.2 12.3 26.1 DBH range (cm) n/a 2.2 – 54.0 1.5 – 65.5 10.0 – 42.7 1.5 – 31.4 0.1 – 32.8 10.0 – 42.7 Saplings (%) n/a 25.0 33.3 25.0 25.0 16.7 8.3 Density (/ ha): pre-outbreak n/a 175.6 103.2 120.8 1075.3 512.9 515.7 Density (/ ha): post-outbreak n/a 73.2 51.6 80.6 627.3 341.9 187.5 Basal area (m2/ha): pre-outbreak n/a 37.8 19.3 19.2 47.9 22.6 60.3 Basal area (m2/ha): post-outbreak n/a 3.2 2.7 3.7 9.0 2.2 5.3 Stand-level mortality (%) n/a 77.8 75.0 44.4 55.6 40.0 70.0

Table A-3 Description of Site 3 (Telluride Creek – medium mortality).

Krummholz line Treeline Forestline Plot 1 Plot 2 Plot 3 Plot 1 Plot 2 Plot 3 Plot 1 Plot 2 Plot 3 Elevation (masl) 1267 1263 1261 1170 1178 1175 1071 1059 1089 Slope (°) 15 17 17 6 7 8 11 17 6 Aspect (°) 79 75 71 112 103 96 72 92 26 Northing (NAD83) 6754218 6754263 6754312 6754685 6754731 6754775 6758522 6758567 6757562 Easting (NAD83) 0652815 0652803 0652793 0653436 0653415 0653399 0655232 0655241 0655248

Shrubs Cover (%) 17.3 71.8 53.6 Dominance (% willow) 15.8 63.9 100.0 Mean height (m) 51.5 ± 36.6 100.7 ± 54.7 136.0 ± 51.3

Spruce Seedling density (/ ha) 0.0 100.0 350.0

107 Mean height (m) n/a 3.9 4.6 2.6 10.2 10.3 10.3 Height range (m) n/a 1.9 – 6.1 2.4 – 6.0 0.5 – 4.8 2.7 – 16.1 4.7 – 16.9 3.5 – 23.2 Mean DRC (cm) n/a 18.4 21.5 14.9 31.7 40.3 40.0 DRC range (cm) n/a 6.2 – 34.6 9.8 – 53.7 4.9 – 47.0 17.5 – 45.5 27.8 – 60.1 11.4 – 64.1 Mean DBH (cm) n/a 8.9 10.6 6.0 19.9 27.0 27.9 DBH range (cm) n/a 3.4 – 13.1 4.7 – 16.7 1.7 – 9.2 10.3 – 27.8 16.8 – 39.4 8.8 – 51.0 Saplings (%) n/a 8.3 0.0 41.7 0.0 0.0 0.0 Density (/ ha): pre-outbreak n/a 54.8 71.9 223.9 1929.4 423.2 528.7 Density (/ ha): post-outbreak n/a 54.8 71.9 223.9 877.0 282.2 44.1 Basal area (m2/ha): pre-outbreak n/a 1.7 3.4 6.0 174.7 57.5 75.0 Basal area (m2/ha): post-outbreak n/a 1.7 3.4 6.0 77.0 33.1 0.4 Stand-level mortality (%) n/a 0.0 0.0 0.0 54.5 33.3 91.7

Table A-4 Description of Site 4 (Mount Decoeli – low mortality).

Krummholz line Treeline Forestline Plot 1 Plot 2 Plot 3 Plot 1 Plot 2 Plot 3 Plot 1 Plot 2 Plot 3 Elevation (masl) 1237 1240 1233 1158 1157 1159 1120 1116 1117 Slope (°) 8 10 8 8 6 9 6 6 8 Aspect (°) 38 60 54 64 59 62 64 62 60 Northing (NAD83) 6747562 6747607 6747651 6747643 6747690 6747735 6747659 6747705 6747752 Easting (NAD83) 0346272 0346254 0346231 0346802 0346783 0346762 0347136 0347129 0347110

Shrubs Cover (%) 62.7 88.2 70.0 Dominance (% willow) 59.0 68.1 78.9 Mean height (m) 58.3 ± 41.9 127.2 ± 53.4 125.8 ± 53.0

Spruce Seedling density (/ ha) 0.0 0.0 100.0

108 Mean height (m) n/a 5.6 6.0 3.7 3.2 4.5 3.4 Height range (m) n/a 1.8 – 9.9 3.0 – 9.0 2.0 – 5.5 1.1 – 9.0 1.4 – 14.6 0.6 – 11.0 Mean DRC (cm) n/a 17.9 31.5 15.1 7.3 12.9 10.4 DRC range (cm) n/a 7.5 – 31.2 10.7 – 69.6 7.4 – 20.9 1.2 – 22.0 2.3 – 51.5 1.0 – 42.0 Mean DBH (cm) n/a 11.7 19.4 8.0 4.7 9.2 10.2 DBH range (cm) n/a 2.6 – 21.2 7.5 – 32.1 2.8 – 13.0 1.5 – 16.2 0.9 – 43.0 1.1 – 32.1 Saplings (%) n/a 8.3 0.0 0.0 16.7 41.7 58.3 Density (/ ha): pre-outbreak n/a 367.9 70.1 194.6 800.0 898.5 757.7 Density (/ ha): post-outbreak n/a 367.9 58.4 194.6 800.0 823.7 694.6 Basal area (m2/ha): pre-outbreak n/a 10.8 6.8 3.9 5.2 25.3 15.4 Basal area (m2/ha): post-outbreak n/a 10.8 5.2 3.9 5.2 9.7 6.7 Stand-level mortality (%) n/a 0.0 16.7 0.0 0.0 14.3 20.0

Table A-5 Description of Site 5 (Boutellier Creek – high mortality).

Krummholz line Treeline Forestline Plot 1 Plot 2 Plot 3 Plot 1 Plot 2 Plot 3 Plot 1 Plot 2 Plot 3 Elevation (masl) 1238 1237 1234 1177 1165 1176 1115 1113 1102 Slope (°) 9 8 7 6 9 8 10 9 8 Aspect (°) 54 55 87 36 32 39 62 80 84 Northing (NAD83) 6761502 6761542 6761584 6762258 6762242 6762246 6762738 6762781 6762831 Easting (NAD83) 0647193 0647159 0647132 0647940 0647894 0647846 0648527 0648510 0648507

Shrubs Cover (%) 80.0 67.3 62.7 Dominance (% willow) 52.5 50.6 88.9 Mean height (m) 126.8 ± 64.4 110.0 ± 74.0 156.4 ± 76.8

Spruce Seedling density (/ ha) 50.0 400.0 450.0

109 Mean height (m) n/a 4.2 4.3 7.6 4.0 12.5 6.4 Height range (m) n/a 0.6 – 10.0 0.8 – 11.9 1.3 – 16.2 0.5 – 12.8 2.3 – 33.2 0.6 – 16.1 Mean DRC (cm) n/a 17.5 18.5 28.7 12.0 33.9 17.6 DRC range (cm) n/a 1.5 – 48.8 1.3 – 76.3 1.6 – 58.7 1.0 – 45.5 7.0 – 59.9 1.0 – 47.7 Mean DBH (cm) n/a 16.3 14.3 22.2 9.5 22.2 20.5 DBH range (cm) n/a 0.2 – 40.0 2.0 – 53.4 0.6 – 52.6 0.6 – 35.3 0.9 – 41.0 1.7 – 36.5 Saplings (%) n/a 50.0 16.7 16.7 33.3 0.0 41.7 Density (/ ha): pre-outbreak n/a 158.2 485.8 227.1 1617.2 477.9 1325.7 Density (/ ha): post-outbreak n/a 118.6 485.8 115.5 1347.7 159.3 662.9 Basal area (m2/ha): pre-outbreak n/a 7.8 16.7 27.0 38.4 52.2 79.6 Basal area (m2/ha): post-outbreak n/a 1.6 16.7 0.6 7.9 6.3 0.8 Stand-level mortality (%) n/a 50.0 0.0 70.0 25.0 66.7 71.4

Table A-6 Description of Site 6 (Telluride Creek – medium mortality).

Krummholz line Treeline Forestline Plot 1 Plot 2 Plot 3 Plot 1 Plot 2 Plot 3 Plot 1 Plot 2 Plot 3 Elevation (masl) 1300 1306 1293 1204 1204 1202 1077 1067 1063 Slope (°) 17 16 15 7 7 7 8 8 12 Aspect (°) 41 38 47 69 68 73 32 42 28 Northing (NAD83) 6755271 6755307 6755338 6756286 6756330 6756364 6759658 6759696 6759724 Easting (NAD83) 0651655 0651625 0651591 0652371 0652351 0652315 0653144 0653112 0653073

Shrubs Cover (%) 45.5 57.3 86.4 Dominance (% willow) 100.0 62.9 38.5 Mean height (m) 100.8 ± 39.2 80.3 ± 40.1 112.4 ± 57.5

Spruce Seedling density (/ ha) 0.0 0.0 650.0

110 Mean height (m) n/a 2.5 1.9 2.1 7.0 8.5 8.1 Height range (m) n/a 1.4 – 4.0 0.8 – 3.3 0.6 – 3.3 0.6 – 16.2 1.7 – 14.0 0.6 – 14.7 Mean DRC (cm) n/a 12.5 16.4 15.9 21.5 26.4 34.4 DRC range (cm) n/a 5.0 – 21.8 4.0 – 42.1 2.2 – 40.9 1.3 – 55.4 3.6 – 52.0 1.4 – 53.2 Mean DBH (cm) n/a 5.5 5.8 5.8 28.3 20.1 30.0 DBH range (cm) n/a 2.5 – 10.4 2.8 – 7.6 1.0 – 9.1 13.6 – 47.3 1.3 – 48.4 17.7 – 41.3 Saplings (%) n/a 41.7 58.3 33.3 41.7 8.3 8.3 Density (/ ha): pre-outbreak n/a 99.9 190.0 177.0 419.7 422.8 182.5 Density (/ ha): post-outbreak n/a 99.9 190.0 177.0 244.8 345.9 66.4 Basal area (m2/ha): pre-outbreak n/a 1.4 6.0 5.1 27.9 32.3 20.6 Basal area (m2/ha): post-outbreak n/a 1.4 6.0 5.1 2.9 21.1 4.0 Stand-level mortality (%) n/a 0.0 0.0 0.0 71.4 20.0 70.0

Appendix B Comparing Ecotone Characteristics across Elevations and Mortality Levels

Table B-1 Results from Kruskal-Wallis tests to compare ecotone characteristics between elevations (krummholz line, treeline, forestline). Significant values (p<0.05) are indicated in bold type.

n df H p Elevation Seedling density 18 2 10.934 0.004 Shrub cover 18 2 1.433 0.489 Shrub dominance 18 2 2.713 0.258 Mean shrub height 18 2 5.298 0.071 Spruce height 432 1 31.951 <0.001 Spruce DRC 431 1 2.045 0.153 Spruce DBH 383 1 14.228 <0.001 Percentage saplings 36 2 1.226 0.268 Post-outbreak density 36 1 12.112 0.001 Post-outbreak basal area 36 1 6.897 0.009

111

Table B-2 Assessment of the relationship between percentage spruce bark beetle mortality and various ecotone characteristics, using Spearman’s rank correlation coefficient. Results were calculated across the entire forest-tundra ecotone for all characteristics; ecotone characteristics which were previously identified as having significant elevational differences (Table B-1) had correlations assessed for each relevant elevation. Significant values (p<0.05) are indicated in bold type.

n ρ p Forest-tundra ecotone Seedling density 12 0.586 0.045 Shrub cover 12 -0.120 0.711 Shrub dominance 12 0.296 0.351 Mean shrub height 12 0.592 0.043 Spruce height 432 0.326 <0.001 Spruce DRC 431 0.271 <0.001 Spruce DBH 383 0.479 <0.001 Percentage saplings 36 0.014 0.936 Post-outbreak density 36 -0.029 0.868 Post-outbreak basal area 36 -0.163 0.343

Treeline Seedling density 6 0.419 0.408 Spruce height 216 0.314 <0.001 Spruce DRC 215 0.228 0.001 Spruce DBH 190 0.424 <0.001 Post-outbreak density 18 -0.251 0.316 Post-outbreak basal area 18 -0.202 0.422

Forestline Seedling density 6 0.464 0.354 Spruce height 216 0.208 0.002 Spruce DRC 216 0.257 <0.001 Spruce DBH 193 0.383 <0.001 Post-outbreak density 18 -0.445 0.065 Post-outbreak basal area 18 -0.584 0.011

112

Table B-3 Results from statistical analyses comparing seedling attributes between elevations (treeline, forestline), mortality class (low, medium, high), and transect-specific percentage spruce bark beetle kill.

Test values p Age Elevation U = 14.000 0.641 Mortality class H = 0.229 0.892 Percentage spruce bark beetle kill ρ = -0.091 0.768

Basal diameter Elevation U = 14.500 0.513 Mortality class H = 0.198 0.906 Percentage spruce bark beetle kill ρ = -0.178 0.560

Height Elevation U = 16.000 0.410 Mortality class H = 0.197 0.906 Percentage spruce bark beetle kill ρ = -0.115 0.709

Radial growth Elevation U = 13.000 0.769 Mortality class H = 0.195 0.907 Percentage spruce bark beetle kill ρ = -0.451 0.122

113