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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BASSMAN, JOHNSON, FINS AND DOBROWOLSKI
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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ROCKY MOUNTAIN ECOSYSTEMS
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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BASSMAN, JOHNSON, FINS AND DOBROWOLSKI
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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ROCKY MOUNTAIN ECOSYSTEMS
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.
1085
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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BASSMAN, JOHNSON, FINS AND DOBROWOLSKI
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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ROCKY MOUNTAIN ECOSYSTEMS
ways in which stand management can supplement natural processes and simultaneously defray the costs of rectifying past
management practices (Fulé et al. 2001).
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