Tree Physiology 23, 1081–1089 © 2003 Heron Publishing—Victoria, Canada

Rocky Mountain ecosystems: diversity, complexity and interactions† JOHN H. BASSMAN,1,2 JON D. JOHNSON,3 LAUREN FINS 4 and JAMES P. DOBROWOLSKI1 1

Department of Natural Resource Sciences, Washington State University, Pullman, WA 99164-6410, USA

2

Author to whom correspondence should be addressed ([email protected])

3

Department of Natural Resource Sciences, Washington State University, Puyallup Research and Extension Center, 7612 Pioneer Way, Puyallup, WA 98371-4998, USA

4

Department of Forest Resources, University of Idaho, Moscow, ID 83844-1133, USA

Received January 23, 2003; accepted May 4, 2003; published online October 1, 2003

Summary The interior west of North America provides many opportunities to study ecosystem responses to climate change, biological diversity and management of disturbance regimes. These ecosystem responses are not unique to the Rocky Mountains, but they epitomize similar scientific problems throughout North America. Better management of these ecosystems depends on a thorough understanding of the underlying biology and ecological interactions of the species that occupy the diverse habitats of this region. This review highlights progress in research to understand aspects of this complex ecosystem. Keywords: biological diversity, disturbance, ecology, ecophysiology, genetics, water.

Introduction The Rocky Mountain region of North America, extending from central New Mexico through western Canada to northern Alaska (Figure 1), is topographically diverse, with interspersed high plains, basins, valleys, canyons, alpine tundra and glaciers. This topographic mosaic is reflected in a wide range of environmental conditions of temperature, solar irradiance, wind, and water availability that change radically over short distances with shifts in slope, aspect and elevation. The indigenous forest trees and associated vegetation have evolved in response to the environmental diversity of the region, resulting in considerable genetic variation, especially for species whose ranges span large portions of the Rocky Mountain region (Rehfeldt 1994), e.g., Pinus ponderosa Dougl. ex Laws. (Oliver and Ryker 1990) and Pseudotsuga menziesii var. glauca (Biessn.) Franco (Hermann and Lavender 1990). Although vegetation in this landscape is a function of physiography, it also reflects historical and current disturbance caused by natural forces, such as fires, avalanches and insect outbreaks, as well as anthropogenic influences such as highgrading, fire suppression and domestic livestock grazing patterns (Veblen et al. 1994, Long and Smith 2000). Increasingly, anthropogenic influences are negatively impacting biological

diversity in Rocky Mountain ecosystems. Because water is scarce throughout much of the western Rocky Mountains, water use and impacts on riparian areas underlie many of the landscape-level changes (Rood et al. 2003). Impacts of other long-term management practices, such as fire exclusion, have affected species composition and vegetation dynamics, resulting in serious forest health problems and increasing the risk of large-scale wildfires. Better management of sensitive ecosystems depends on a thorough understanding of the underlying biology and ecological interactions of species with each other and their environments (Long 2003). Much has been learned about the physiology, genetics and ecology of forest trees in the Rocky Mountains, but we still do not fully understand the complex interactions and long-term impacts of altered disturbance regimes and concomitant or subsequent changes in insect populations and pathogen dynamics. The objective of this review is to highlight examples of these interactions.

Ecophysiological challenges to trees in Rocky Mountain ecosystems Rocky Mountain ecosystems are characterized by high elevations or persistent water deficits. Low soil and air temperatures, high incident radiation (including ultraviolet radiation and reflectance off snow) and winter desiccation injury to perennial plants characterize high-elevation environments, whereas lower elevations are characterized by high temperatures, high vapor density deficits and region-wide summer droughts, resulting in daily and seasonal water deficits. High-elevation environments Temperature appears to be a key factor limiting tree growth at high elevations (Smith et al. 2003). In extra-tropical forests of both the northern and southern hemispheres, summer temperatures (mean warmest month) apparently control the elevation at which tree lines develop, whereas winter temperatures (mean coldest month) affect the type of species found at the

† This paper was among those presented at the 17th North American Forest Biology Workshop “Rocky Mountain ecosystems: Diversity, complexity and interactions,” sponsored by the Tree Physiology and Forest Genetics working groups of the Society of American Foresters and held at Washington State University, Pullman, WA.

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Figure 1. Continent of North America showing location of Rocky Mountains with the five major sections identified.

tree line (Jobbagy and Jackson 2000). This finding is consistent with reports that low air temperatures during the growing season limit vegetative growth of trees, whereas photosynthesis is not temperature-limited (Kozlowski et al. 1991, Smith et al. 2003). Trees at the tree line typically assume a characteristic krummholz form that has been attributed to a carbon imbalance caused by an increase in non-assimilating tissue. In subalpine fir (Abies lasiocarpa (Hook.) Nutt.) krummholz, carbon balance is strongly influenced by temperature and winter injury, with no apparent reallocation of carbon into non-photosynthetic tissue (Cairns and Malanson 1998). In larch (Larix

spp.) and pines (Pinus spp.), carbon allocation to foliage is equal or higher, respectively, as elevation increases, whereas tree height decreases with elevation but total tree biomass does not (Bernoulli and Körner 1999). The development of krummholz tree islands that move across the landscape at high elevations has received recent attention (Smith et al. 2003). Soil phosphorus limitations have been implicated in determining both species composition and the direction in which the tree islands move. Shiels and Sanford (2001) found higher plant-available P under Engelmann spruce (Picea engelmannii Parry ex Engelm.) krummholz than

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under bristlecone pine (Pinus aristata Engelm.) krummholz, but no other soil chemical differences could be ascertained. Parker and Sanford (1999) found no difference in soil P between tree islands and adjacent tundra dominated by grasses. Karlsson and Weih (2001) tested the hypothesis that low soil temperature limits N uptake by mountain birch (Betula pubescens Ehrh.) at the tree line in Sweden. They found no difference in soil temperature between sites colonized by mountain birch and nearby non-forested habitats, and were unable to explain why mature trees could not survive on sites presently unoccupied by birch. The combination of low temperature and high irradiance at high elevations leads to low-temperature photoinhibition (LTP) of photosynthesis (Smith et al. 2003). The ability of a species to either avoid or tolerate LTP has been related to tree survival at high-elevation sites. Germino and Smith (1999, 2000) found that A. lasiocarpa exhibited greater tolerance to LTP than P. engelmannii, but the latter showed greater avoidance of LTP through increased needle inclination and clumping. Tolerance to LTP was related to survival at high elevations in Eucalyptus species exposed to freezing temperatures and high irradiances (Close et al. 2002) and in a subalpine Rhododendron species (Neuner et al. 1999). Close et al. (2000) also found a positive relationship between LTP and anthocyanin production in Eucalyptus. The epoxide cycle has been implicated in protection of chlorophyll from photooxidation (Devlin and Barker 1971), and foliage concentrations of epoxide cycle pigments have been positively correlated with elevation, incident solar irradiance and UV radiation (Robakowski and Laitat 1999, Tegischer et al. 2002). Another important ecophysiological trait of trees growing at high elevation is winter desiccation injury (Kramer and Kozlowski 1979) caused by low soil temperatures restricting (at or above freezing) or preventing (below freezing) water absorption (Wan et al. 2001). Shoot desiccation can result from increased transpiration as a result of high air temperatures or wind velocities, or from direct needle water loss when blowing snow abrades the cuticle. Cairns (2001) studied A. lasiocarpa krummholz in Glacier National Park and found that the incidence of winter injury increased with elevation and on southwest aspects. Within krummholz patches, injury incidence was found mostly on the windward edge. Xylem embolism during winter desiccation was studied in tree line Picea abies (L.) Karst. growing in the Central Alps (Mayr et al. 2002). Conductivity losses of up to 100% were observed at water potentials down to –4.0 MPa, and vulnerability thresholds (water potential at 50% loss of conductivity) decreased with increasing elevation. This decrease was attributed to smaller tracheid, pit and pit pore diameters in trees at the higher elevations. Water stress The Rocky Mountain region is characterized by extended drought periods during summer and early fall. High daytime air temperatures and incident radiation combine to drive high rates of transpiration (Kramer and Kozlowski 1979, Kramer 1983, Kozlowski et al. 1991). The severity of the summer

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drought can be exacerbated by coarse, shallow soils (low soil water reserves) and by slope aspect, with southwest exposures being more xeric and northeast exposures more mesic. However, seedlings and young trees are capable of adapting to water stress through osmotic adjustment and changes in tissue elasticity, lessening the effect of subsequent stress cycles (Seiler and Johnsen 1994, Kozlowski and Pallardy 2001). Long-term allometric adjustments occur when a tree establishes and develops in a given water regime. Cinnirella et al. (2002) reported that tree water transport homeostasis is achieved by a combination of short-term stomatal regulation and optimal allocation among foliage, conducting sapwood and absorbing root tissue. Sperry et al. (2002) suggested that tree water use is altered in response to drought through species-specific alterations in hydraulic conductance from the soil to the canopy. This suggestion was reinforced by a study of hydraulic properties of contrasting oak species (Cavender-Bares and Holbrook 2001). As soil water content increased, leaf area per shoot increased and the ratio of sapwood to leaf area decreased. Maherali and DeLucia (2001) compared ponderosa (P. ponderosa) pine growing on contrasting sites in Nevada and found that trees at the dry site allocated more biomass to sapwood relative to foliage, thereby preventing xylem water potentials from reaching the point of xylem embolism. Similarly, at high elevations in the northern Rocky Mountains, subalpine fir, a climax species, had higher daily tree water use and twice the leaf to sapwood area, but a lower leaf-area-based sap flow (Sala et al. 2001) compared with whitebark pine (Pinus albicaulis Engelm.), an early seral species. In a high-elevation meadow in Arizona, changes in leaf area relative to sapwood area controlled the responses of stomatal conductance and hydraulic conductance to water stress in ponderosa pine and limber pine (Pinus flexilis James) (Fischer et al. 2002). Tree water use did not differ between the wet and dry summer seasons because of tight stomatal control of water loss. As the size of a tree increases, tissue water storage becomes more important in meeting transpirational demand. Zweifel et al. (2001) concluded that the use of internal water reserves in the bark and foliage of transpiring Norway spruce (P. abies) saplings helped optimize water transport by buffering peaks of extreme water consumption (Zweifel and Häsler 2001). Such buffering may reduce the formation of emboli in the xylem, maintaining hydraulic continuity during dry periods (Meinzer et al. 2001). A linkage between hydraulic architecture and leaf physiology has been demonstrated (Meinzer et al. 2001), but the signaling mechanism remains controversial. Aasamaa et al. (2002) reported a strong correlation among foliar gas exchange, hydraulic characteristics and endogenous concentrations of leaf abscisic acid (ABA) in six tree species. Furthermore, exogenous application of ABA modified shoot hydraulic conductivity and stomatal closure, suggesting that ABA may be an important chemical signal of water stress. Passive hydraulic redistribution (also known as “hydraulic lift”) of water within the soil profile, usually from lower to upper soil horizons, has large implications for water relations of Rocky Mountain trees (Dawson 1996, Meinzer et al. 2001).

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Caldwell et al. (1998) suggested that large quantities of water are lifted at night to supply transpiration the following day. During this process, water leaks from tree roots, increasing the available water in the upper soil horizon. This water redistribution has been found to support smaller, less deeply rooted plants, including tree regeneration in the understory. With higher soil water content in the upper horizon, nutrient availability and uptake may be enhanced (Caldwell et al. 1998). Dawson (1996) used hydrogen stable isotope composition to distinguish soil water from groundwater and found that the proportion of transpired water from groundwater increased with tree size. Several studies have utilized carbon isotope discrimination (δ13C) as an integrative, long-term measure of water-use efficiency in trees (Schimel 1993) and a means to study stomatal limitation of photosynthesis (Warren et al. 2001). Common garden provenance trials of several Rocky Mountain tree species, including Douglas-fir (Pseudotsuga menziesii var. glauca), ponderosa pine and western larch (Larix occidentalis Nutt.), have shown high correlations between δ13C and wateruse efficiency (Zhang et al. 1993, 1994, 1996, Zhang and Marshall 1995). Because water availability is a common environmental constraint to tree growth in the Rocky Mountains, selection of forest tree genotypes for water-use efficiency could improve the productivity of planted forests.

Water in Rocky Mountain ecosystems Water inputs to Rocky Mountain ecosystems derive primarily from the Pacific Ocean and the Gulf of Mexico in at least three distinct precipitation patterns (Smith 1994). The Plains Type, extending from the crest of the Rockies eastward, is characterized by pronounced summer maximum rain from air masses moving from the Caribbean and the Gulf of Mexico. The Arizona Type, extending west from the Rockies’ crest in New Mexico and Arizona, is a monsoonal pattern. Winters have limited precipitation deriving from cyclonic storms in the Pacific Ocean with occasional heavy snows. There is a pronounced spring dry period, followed by showers and thunderstorms in mid-summer fueled by moisture from both the Gulf of California and the Gulf of Mexico. Precipitation diminishes again in the late fall. The sub-Pacific Type characterizes areas lying west of the crest and east of the Sierra Nevada, Cascades, and Coast Range of British Columbia. There is a winter maximum of precipitation derived primarily from storms moving eastward from the Pacific Ocean. Storms may influence watershed processes such as flooding, landslides, erosion and forest blowdown, and as they move inland and meet the Rocky Mountains, orographic effects increase the amount and alter the form of precipitation (Beschta 1998), producing diverse ecosystems ranging from arid and semiarid plateaus to alpine snowfields. Flooding, in particular, has created geomorphic diversity by sculpting the landscape surface, controlling plant and animal species distributions, and altering successional processes. Major storms continue to reinitiate riparian plant community succession, and they flush sediment, organic materials and nutrients from

headwater streams to downstream lowlands. These events alter watershed and stream form and function, with consequences for downstream aquatic life and water quality (Clifton et al. 1999). Periodic low-intensity flooding and changes in stream channels lead to braided riparian areas and are considered essential for healthy riparian ecosystems (Rood et al. 2003). Understanding this spatial and temporal variability in precipitation intensity, form and amount, and the hydrologic response of watersheds, is critical for conservation of water quality and quantity and for restoring watershed functions in Rocky Mountain ecosystems. Humans have altered hydrologic regimes throughout the Rocky Mountains. During the late 1800s, removal of beaver and the utilization of anadromous fish stocks along the Columbia River system impacted Rocky Mountain ecosystems (Beschta 2000). Subsequent logging, livestock grazing and agriculture have resulted in increased erosion and sedimentation (Beschta 2000, Thurow 2001). Excessive timber harvesting reduces evapotranspiration and interception, which increases soil water content, temporarily increasing surface and subsurface flows and destabilizing slopes. Other forest practices have caused changes in slope steepness, slope-water effects, soil strength and vegetation rooting strength, promoting higher peak flows, channel erosion, and greater landslide activity (Robison et al. 1999, Sidle 2000). Landslides are the principal erosion process on steep forested slopes throughout the Rocky Mountain region (Swanson et al. 1987) and occur most frequently after intense winter rains or rain-on-snow events. Roads disrupt surface and subsurface flow patterns, create a load on fill slopes and remove support of the cut slope by channeling water along the road surface, thus causing slides greater than those caused by vegetation removal alone (Covington et al. 1994). These landslides alter flora and fauna distribution patterns and reset successional trajectories. Agriculture throughout the lower elevations of the Rocky Mountains has further impacted watersheds through increased delivery of agrichemicals to streams, elimination of riparian buffer zones, and channelization of streams. Grazing pressure from domestic livestock and growing populations of deer, elk and bison continues to degrade meadow complexes and riparian habitats. Excessive grazing by domestic livestock has altered upland plant and litter cover, leading to reduced infiltration, greater runoff and accelerated erosion. Fire exclusion in Rocky Mountain ecosystems and the resulting increases in overstory trees and litter accumulation have increased evapotranspiration and interception, leading to reduced water availability in upland soils and surface, subsurface and instream flows. Fire exclusion with concomitant fuel accumulation has recently led to catastrophic wildfires that reduce rain-absorbing plant and litter cover and alter the physicochemical characteristics of surface soils (Clary et al. 2000). Wildfires are becoming more frequent, affecting riparian areas directly (stream temperature and chemistry changes) and indirectly (hydrologic regime, erosion, sediment and debris loading, cover reduction), and triggering changes in water quality, quantity and timing. Wildfires and their associated hydrological effects are typically characterized as pulsed distur-

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bances to which many aquatic organisms, including native salmonids, are adapted (Reiman 1997). However, greater fire frequency coupled with the loss of well-connected, spatially complex habitats due to chronic management effects (e.g., conventional road construction and maintenance and timber harvest) could lead to long-term damage to aquatic systems. Agriculture and urban growth throughout the Rocky Mountains have been accompanied by the redistribution of water across the landscape by dams and inter-basin transfers (e.g., most recipients of water from the Colorado River are outside of the basin). Diversion of streams for irrigation, and flood control through stream channelization and the removal of beavers and their dams, have contributed to reduced upstream storage and reduced low flows. Flood control actions have reduced the amount of sediment entering and passing through stream systems, thereby reducing the available fish rearing habitat (Wissmar et al. 1993). Hydroelectric dams, which obstruct river flow and inundate river valleys, also destroy fish habitat. By the early 1970s, 65 million acre-feet of water were impounded by dams, including 14 on the Columbia River and 13 on the Snake River (Beschta 2000). Today there are 1,025 dams obstructing the water flow in Washington State alone. Throughout the Rocky Mountain region, the quality and quantity of water flowing over and through a watershed is used as a measure of ecosystem health. Gregersen et al. (2000) believe that water will become the key land management issue in the 21st century as western populations continue to expand and human activities continue to deplete and pollute aquifers and degrade surface water quality, producing conflicts over water allocation.

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replaced western white pines in the moist mid-elevation forests, and in the cold high-elevation forests, spruce and subalpine fir have severely encroached into former whitebark pine forests (Hann et al. 1998). Past harvest practices, the introduction of white pine blister rust and mountain pine beetles, and the exclusion of stand-replacement fires have shifted regional ecosystems toward more mature stages of forest development and toward late-successional, shade-tolerant species, particularly in northern Idaho and western Montana (Byler and Hagle 2000, Fins et al. 2002a, 2002b). Fire exclusion is implicated in the decline and lack of recruitment of new stems of aspen, a species that propagates largely by clonal root sprouts (Romme et al. 1995). These widespread changes have been accompanied by dramatic increases in insect and pathogen populations and alarming declines in forest health (Atkins et al. 1999, Byler and Hagle 2000). It appears that some northern Rocky Mountain ecosystems may have surpassed their capacity to withstand environmental stresses such as repeated seasonal droughts. This shift has broad implications. The “overwhelming majority of plant diversity” is found among the understory species in western forests (Pfilf et al. 2002). Understory plant species provide habitat for a variety of other flora and fauna (fungi, soil microorganisms, insects, mammals and birds) and the understories of young and early successional forests differ from those of older late successional forest types. Loss of diversity in the overstory will be accompanied by loss of diversity in the understory (Pfilf et al. 2002).

Disturbance and interactions Biological diversity in Rocky Mountain ecosystems Tree species native to Rocky Mountain ecosystems vary widely in their adaptive responses to climatic and geomorphic variation (Rehfeldt 1994). Some species, such as western white pine (Pinus monticola Dougl. ex D. Don), may be considered broadly adapted generalists, whereas the adaptive responses of other species, such as Douglas-fir, more closely track environmental gradients (Rehfeldt 1994). Compared with forests in other regions of North America, the predominance of wide-ranging, coniferous, wind-pollinated tree species suggests relatively high genetic diversity in the forests of the Rocky Mountains (Hamrick and Godt 1989). Rocky Mountain ecosystems show broad-scale changes in plant communities. White pine blister rust (Cronartium ribicola) and fire exclusion have dramatically reduced whitebark pine populations in recent years and the species is described as “functionally extinct in more than a third of its range” (Kendall 2003). In Idaho, populations of western white pine, ponderosa pine, western larch, aspen (Populus tremuloides Michx.) and whitebark pine have decreased considerably compared with historic values (Atkins et al. 1999). Regionally, Douglas-firs and true firs occur in unprecedented numbers in the dry, low-elevation forests that were historically dominated by ponderosa pine. Douglas-firs, true firs and lodgepole pines (Pinus contorta Dougl. ex Loud.) have largely

Disturbance is ubiquitous to every ecosystem. The type of disturbance and its timing, frequency and severity affect the spatial heterogeneity or patchiness of landscapes, which in turn influences present and future patterns of disturbance (Veblen et al. 1994). Interactions among different types of disturbances (e.g., fire, windthrow, insects and pathogens) are common in forests of western North America; however, there is little quantitative data describing these interactions (Veblen et al. 1994). The effects of anthropogenic influences (past and current mining activities, atmospheric deposition and climate change, road construction and concomitant habitat fragmentation, increasing exurban and resort development, wildland recreation and clearcutting) are difficult to distinguish from those of natural disturbances (Savage 1994). In the Rocky Mountains, fire has historically been the most important form of natural disturbance (Veblen et al. 1994 and literature cited therein, Hann et al. 1998, Long 2003). In many mid- to low-elevation fire-adapted communities, a frequent fire return interval (30 to 50 years) resulted in predominantly low intensity surface fires that eliminated accumulated fuel and prohibited invasion by later successional species, resulting in open park-like stands (Agee 1997, Rollins et al. 2001). Stand-replacing fires were rare. By contrast, in subalpine and boreal forests, stand-replacing crown fires were the norm, with return intervals of 200 to 300 years (Veblen et al. 1994, Johnson et al. 2001). In the southern Rocky Mountains, fire occur-

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rence was strongly tied to interannual drought conditions and associated with alternating climatic cycles of El Niño-Southern Oscillation (ENSO) and La Niña prior to Euro-American settlement (Donnegan et al. 2001). The extent and rate of burning have decreased substantially since the late 19th century in both wilderness and non-wilderness areas of the Rocky Mountains (Agee 1997, Rollins et al. 2001). Successful firefighting and fire prevention programs reduced wildfire acreage to historic lows by the 1950s (Agee 1993). Fire exclusion during the 20th century has resulted in fuel accumulations and the development of multi-layered canopies composed of shade-tolerant but fire-intolerant species in areas where these were historically absent or occurred in only small numbers. Consequently, there are now more continuous, heavier, and three-dimensional fuel loads that result in more severe fires once ignited (Agee 1997). It is predicted that decreased fire frequencies and associated increases in fuel loading and homogeneity will result in larger, more severe fires than have occurred in previous centuries (Rollins et al. 2001). Indeed, heavier fuel accumulations are leading to an increase in wildfire acreage despite aggressive firefighting efforts in many areas (Agee 1997, Arno and Allison-Bunnell 2002). In 2000, 3.4 million ha burned in the United States (Hesseln 2001). In 2002, wildfires burned about 2.7 million ha (National Interagency Fire Center, Boise, ID). It is estimated that, as a result of fire exclusion and past management practices, approximately two-thirds of the 200 million acres of federally managed wildlands in the United States, which are adapted to frequent fire regimes, are in moderately to severely degraded condition (Fulé et al. 2001). The unnatural shift in species composition and fuel loading in fire-dependent ecosystems is now recognized and there has been a concerted effort during the past decade to restore the historical balance through prescribed burning (Agee 1997, Hesseln 2001). Nevertheless, many decades will be required to restore historical fuel conditions in the many areas requiring treatment. Restorative management scenarios that combine commercial thinning and prescribed burning offer significant promise of improvement in ecological parameters while simultaneously reducing fuel hazards (Lynch et al. 2000), but such studies are still in early post-treatment phases (Fulé et al. 2001, Fins et al. 2002b). There is considerable interaction between different sources of disturbance, but the importance of these interactions has only recently been recognized (Baker and Veblen 1990, Agee 1997). For example, the extent and shape of fires directly affect landscape heterogeneity and ecosystem diversity (Agee 1993). The sizes and shapes of fires are directly related to fire frequency and behavior, such that shifts in the size distribution of fires and rates of burning over long periods of time can significantly change the patch dynamics of landscapes (Rollins et al. 2001). Fire suppression activities during the 20th century resulted in reductions in average fire size without corresponding increases in fire frequency, leading to more homogeneous landscape structures even in wilderness areas (Rollins et al. 2001). A shift in composition and activity of forest tree insect

pests and pathogens has been associated with these landscape changes (Savage 1994, Veblen et al. 1994, Byler et al. 1997). In fire-adapted forest ecosystems, fire exclusion and concomitant shifts in species composition to more shade-tolerant species have resulted in forest communities with higher densities. Higher densities have rendered these forests more susceptible to environmental stress, which in turn has increased the incidence of insect and pathogen attack, resulting in further accumulations of fuel. In the Blue Mountains of Washington and Oregon, removal of fire has allowed defoliator-susceptible trees to increase, resulting in longer and more damaging insect epidemics (Agee 1997). In the northern Rocky Mountains, a distinct species shift has occurred since the early 1900s from a predominance of shade-intolerant to shade-tolerant species. The absence of fire caused the initial compositional change, but insects and pathogens drove later changes (Byler et al. 1997). Insect and pathogen composition shifted from predominantly white pine blister rust plus mountain pine and other bark beetles in the early 1930s to root diseases, stem decays, Douglas-fir beetle and spruce beetle (Dendroctonous rufipennis (Kirby)) by the mid-1970s (Byler et al. 1997). All indications are that, in the absence of fire, fungi and beetles are producing forests of low density, mature grand fir (Abies grandis (Dougl. ex D. Don) Lindl.) and subalpine fir mixed with stands of perpetually young, small Douglas-fir and true firs (Hagle et al. 1995). Forest insect pests and pathogens are also major driving forces of disturbance in Rocky Mountain ecosystems (Veblen et al. 1994, Wargo 1995, Long 2003). Although disturbance by these agents is more insidious than that of fire, their activities are responsible for considerable nutrient cycling and release of other resources from aging biomass. However, because insects and pathogens are selective, the areas affected by these disturbances differ from those affected by fire (Wargo 1995). In the southern Rocky Mountains, major outbreaks of spruce bark beetle have caused large-scale mortality in subalpine habitats, significantly influencing the structure of these forests (Baker and Veblen 1990, Eisenhart and Veblen 2000). It is commonly believed that extensive insect disturbances were precursors to the large fires of the 19th century and that the more recent scarcity of fires is partly a result of the absence of insect outbreaks (Baker and Veblen 1990). However, based on dendroecological records, these outbreaks have had complicated interactions with fire cycles (Eisenhart and Veblen 2000) and it is unclear whether beetles increase the susceptibility of forests to natural fire (Baker and Veblen 1990). Tree densities have increased during the 20th century on sites disturbed by spruce bark beetles during the 19th century, but it is unclear whether this is a result of subsequent fire suppression or natural recovery from beetle disturbance (Baker and Veblen 1990). Restoration of natural disturbance regimes is considered essential for maintaining biological diversity and has significant economic implications as the wildland–urban interface continues to expand (Everett et al. 2000, Hesseln 2001). However, rebalancing ecosystems requires a thorough understanding of both inherent disturbance regimes (Everett et al. 2000) and

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