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OCR for page 73
4
THE STATUS AND FUNCT/ON/NG
OF PACIFIC NORTHWEST FORESTS
/NTRODUCT/ON
In this chapter, the status, condition, and sustainability of the forested
ecosystems and associated plant and animal species of the Pacific
Northwest are reviewed and assessed. The effects of forest-use patterns
and management practices on timber and nontimber species are
evaluated, and the discussion includes an analysis of species that are
threatened or endangered by forest cutting and habitat fragmentation.
The chapter ends by considering the long-term effects of the loss of
biological diversity on the stability and functioning of ecosystems In
general. The findings presented in this chapter provide importantbases
for the assessment of forest management practices in Chapters 7 and S.
FOREST COND/T/ON
Genera/ Criteria of Condition
A system that is in good condition is one that retains its basic structures
end processes (Rapportl989~. The condition of en ecosystem represents
more than the absence of disease. it is also the ability to resist or recover
quickly from environmental stressors. Because ecosystems seldom have
clear boundaries, ecological condition spans spatial scales. The structure
of landscapes, for example, shapes processes (e.g., hydrology and
propagation of disturbances) that influence the integrity of stands and
73
OCR for page 74
74
Pacific Northwest Forests
streams. The integrity of streams depends additionally on the integrity
of riparian forests and of upsiope forests that control sediment inputs to
the streams. individual species modify processes in many ways that
influence ecological health. One example is the regulatory role played
by birds and predatory insects in consuming tree-feeding insects (e.g.,
Torgersen et al. 1990; Marquis and Whelan 1994~. The natural enemies
of insect pests require habitats such as large dead wood.
The traditional approach to assessing forest condition on the basis of
appearance (an inventory) may not detect early changes in condition, an
issue that was addressed well by the National Research Council in 1994
(NRC 1994~. The committee recommenced that assessment of ecosystem
condition also should consider 1) the stability of soils and watersheds,
2) the integrity of nutrient cycles and energy flows, and 3) the function-
ing of ecological processes that facilitate recovery from damage. Those
same factors underlie the health of forests. The authoring committee
concluded that assessing rangeland condition by comparing the
abundance and kinds of plants growing in an area with a benchmark
plant community (a list of plants expected on rangeland in excellent
condition) was inadequate because it did not ensure protection of
processes critical to ecosystem sustainability.
Not all ecosystem types are represented extensively on public lands
in the Pacific Northwest-examples include lowland floodplain forests,
oak woodlands, and coastal tidal marshes. And checkerboard owner-
ships of public and private lancis hinder effective assessment and
management of forest ecosystem patterns and processes.
Measurement of key ecological and ecosystem processes is more able
Man an inventory to provide sensitive indicators of forest condition.
Such processes inclucLe I) rates of nutrient (especially nitrogen) capture
(from soil or atmosphere) and fixation in biotic tissue; 2) rates of water
flux needed to maintain cellular function and evapotranspiration rates
(to prevent cellular cavitation and wilting); 3) rates of nutrient cycling
and biotic processes that ensure adequate supplies of critical nutrients
(especially nitrogen) or maintain balance among critical nutrients (such
as C:N:P:K:Cai ratios) and minimize nutrient leaching from the system;
and 4) rate of development of soil and canopy characteristics that
maintain favorable temperature and humidity, atmospheric intercep-
lCarbon:nitrogen:phosphorus:potassium:calcium.
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Status and Functioning
75
lion, and control of water flow and erosion. Clearly, it is not practical
to measure these In all or even a large number of stands. However, such
measurements from a representative subset of stands could provide
critical insight into the {ong-term trends In forest condition that result
from alternative management practices. Maintenance of species
assemblages and ecosystem processes both are important in measuring
forest condition.
Ecosystem health problems in Pacific Northwest forests can be
grouper! into Free general, Interrelated categories: 1) increased
vulnerability to insects, pathogens, fire, and drought; 2) extreme
fragmentation or loss of habitat and of its biological diversity; and 3) soil
degradation. Outgrowth forests can be lost to fire, drought, or insects
just as they are lost to chainsaws. Disturbance regimes that are too
frequent and too severe degrade son's ant! increase sediment export to
streams. Degraded soils grow new forests slowly and in some cases, not
at all (Perry et al. 1989a).
The Role of Bio/ogica/ Diversity
Efforts to conserve biological diversity are based on the assumption that
biodiversity has value. That value has been documented In numerous
publications, including books by EhrTich and Ehrlich (1981), Wilson
(198S, 1992), and NRC (1999b). Some of the major points raised in those
treatments are reviewed below, as are more recent findings, especially
those that relate the functioning of ecosystems to their biodiversity
(Johnson et al. 1996~.
The number and genetic variability of plant, animal, and microbial
species that live in a given location is caned its biodiversity.
Biodiversity also encompasses biologically mediated processes in a
habitat. Forestry and other land-use practices are influencing the
bioctiversity of the entire Pacific Northwest region. However,
biodiversity is also important on small scales. Practices that reduce the
biodiversity of a I,000-acre forest stand, for instance, can greatly affect
its functioning even if those practices do not threaten any species with
extinction. The loss of local biodiversity—i.e., the loss of species from
a given habitat—canbe of great ecological importance. Factors Mat lead
to the lass of local biodiversity include the conversion of naturally
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76
Pacific Northwest Forests
regenerating old-growth forests into rapid-rotation forests dominated
by a single tree species; the fragmentation of habitats via road building,
agriculture, and clearcu~ing; changes in fire regimes; fertilization; and
application of pesticides.
The loss of local biocliversity is a concern because accumulating
evidence medicates that viable populations of indigenous species are
important to Me rates, seasonality, and direction of processes contribut-
ing to overall ecosystem functioning(Temple 1977; Bormann anct Likens
1979; Franklin et al. 1989; Schowalter and Filip 1993~. A few experi-
ments have addressed the functional importance of particular species or
species assemblages. Tilman and Downing (1994) manipulatecl plant
diversity ire a grassland ecosystem via nitrogen addition and found that
primary productivity during a drought was significantly related to plant
diversity. Productivity in the lowest diversity plots dropped to about
one-tenth of what it had been before the drought, but in the most
diverse plots it only dropped to one-half of its precErought level. The
productivity of the more diverse plots was stabilized by cErought-
tolerant species compensating via increased growth for reduced
productivity of drought-intolerant species. Because high diversity stands
are more likely to have disturbance-resistant species in them, on average
they should be more stable, i.e., more resistant, to disturbance.
Schowalter and Turchin (1993) measured the growth of experimentally
introduced southern pine beetle (Dendroctonusironfalis) populations in
pine/hardwooc} forests in which tree diversity had been manipulatecE
by reducing densities of pines or hardwoods. Only dense, single-species
plantations of pines were conducive to population growth of this pine-
killing insect; the presence of nonhost hardwoods or shrubby vegetation
apparently interfered with discovery of interspersed or scattered hosts.
In both of the studies above, primary productivity was more stable
when plant diversity was higher. These results are supported by much
nonexperimental research on factors contributing to pest outbreaks
(Kareiva 1983; Schowalter et al. 1986; Hunter anct Aarssen 1988~.
Viable populations of indigenous species might be critical for
maintenance of ecological processes in Pacific Northwest forests, but
few studies have addressed the contributions of particular species or
species assemblages to processes.
Many examples of the importance of biodiversity in ecosystem
functioning are obvious. A critical role of organisms is in decomposi-
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Status and Functioning
77
tion the breakdown of organic structures into their physical elements,
including energy. There is increasing evidence that ecosystems with
higher levels of organ~smal diversity are better at carrying out produc-
tivity (e.g., Risch 1980; Courtney 1985; Ewe] 1986; Ewel et al. 1991; Frank
and McNaughton 1991; Naeem et al. 1994; Titan et al. 1996) and
nutrient retention (Tilman et al. 1996~. Organisms create structures and
communities that interact with and alter the physical world—such as
forests ancl coral reefs—and that provide habitat for other organisms
Mat influence acEctitional processes. Organisms and biological structures
have important influences on the hydrologic cycle (e.g., through
condensation, interception, ancE evapotranspiration) anct on geomorphic
processes, such as erosion.
That biodiversity of ecosystems is linked to their functioning was first
proposed by Elton (1958), further developed by Odum (1969), and then
called into question by May (1973), Goodman (1975), and others.
However, May and Goodman were addressing a different aspect of
stability than Elton or Odum. Elton and Odum were referring to the
stability of an entire ecosystem, whereas May and Goodman were
referring to the stability of a single species within an ecosystem. May,
for instance, demonstrated theoretically that the population size of a
species is expected to be more stable (i.e., return to equilibrium more
rapidly after a perturbation) if the species lives by itself than if it
competes with many other species. His result is still considered robust.
However, May did not explore the effect of a perturbation on the
stability of total community biomass, ecosystem primary productivity,
soil nutrient conservation, or other such ecosystem characteristics.
Those characteristics, however, were the attributes of greater interest to
Elton and Odum. A recent field study (Tilman 1996) has supported both
Elton and May.
Tong-term acclimation of ecosystems to changes in climate and other
environmental variables is primarily dependent upon available
bioctiversity. Obviously, greater numbers of species and greater genetic
variability among species provides for a larger number of biological
building blocks for ecosystem adjustment end acclimation. Given ever-
changing environments, the capacity to acclimate is central to the iong-
term sustainability of ecosystem processes. Such changes are obvious
in the shifts of species' abundances documented in I,000- to 10,000-year
records obtained by studying pollen profiles in lake sediments.
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78
Pacific Northwest Forests
Relatively unimportant species restricted to particular microsites during
one climatic regime often become important and widespreact as climate
shifts (e.g., Anderson 1990) or as a disease pathogen invades a habitat
(Davis et al. 1997~. The reservoir of genetic diversity within individual
species and populations is central to their ability to adapt to environ-
mental change. In view of this, focus on genetically engineered
genotypes of crop plants and forest trees has raised concern regarding
the loss of genetic diversity that m~ghtbe important to future conditions.
Ewel and co-workers (1991) experimentally established tropical
successional sequences that differed in plant biodiversity. They found
that more diverse communities were more nutrient conserving and more
productive. Naeem et al. (1994) experimentally established British
grassland communities that differed in their plant, herbivore, and
decomposes diversity. They found that diversity led to significantly
increased primary productivity as measured by the rate of photosynthe-
sis. Diversity also significantly affected decomposition, nutrient
retention, anct vegetation structure. In reviewing these studies on the
effects of biodiversity on ecosystem productivity and stability, Kareiva
(1994) concludecE that the loss of biodiversity leads to "less productive
ecosystems, vulnerable to environmental perturbations, and plagued by
declining soil fertility." This conclusion was supported by a field study
in which the plant diversity of 147 prairie grasslanct plots was manipu-
latecl and found to affect directly total plant community productivity,
nutrient use, and nutrient leaching loss (Tilman et al. 1996~.
The redistribution of species across the globe is one of the most
significant effects that humans have on ecosystems. The negative
consequences of exotic species in natural and managed ecosystems
de~nonstrates that the contribution of biological diversity to ecosystem
functioning is not merely a function of the number of species present,
but of their identities and evolutionary interrelations.
Resistance and Resilience
Biodiversity provides stability (resistance) and recovery (resilience) in
the face of disturbances that disrupt important ecosystem processes.
Resistance often results from complex linkages among organisms, such
OCR for page 79
Status and Functioning
79
as food webs that provide alternate pathways for achieving particular
flows of energy anct nutrients. For example, the presence of numerous
fungal species capable of forming mycorrhizae in a terrestrial ecosystem
buffers the ecosystem against the loss of individual species ancl makes
total loss of mycorrhizal function unlikely.
McNaughton (1977) explored theoretically the possible effects of
bioctiversity on ecosystem stability. He found that increased functional
diversity within an ecosystem was expected to make an ecosystem more
stable. Numerous studies have demonstrates! that increased plant
diversity helps stabilize primary production in response to climatic
change. For instance, in 1977, a severe ctroughtin eastern Europe caused
greatly decreased plant growth and crop yields. Inept et al. (1982)
compared the effects of the drought on the productivity of two Czecho-
slovakian fields, one of which had low plant-species richness and the
other win high plant-species richness. The productivity of the species-
poor field fell to I/3 of its predrought level, whereas that of the species-
rich field only fell to 2/3 of its predrought level. Although any
comparison of two sites is open to alternative interpretations, many
subsequent, better-replicated studies have found this same effect. For
instance, Frank and McNaughton (1991) found that the most diverse
Yellowstone grasslands were most stable in response to drought.
Similarly, a [Long-term study of 207 plots in Minnesota grassland and
savanna also included a period during which there was a major
drought. Because the productivity and species composition of all plots
was annually measured before, cLuring, ancE after the 1987-1988 drought,
it was possible to include extensive statistical controls for numerous
potentially confounding variables when determining the effects of
biodiversity on the drought resistance and resilience of these plots
(Tilman and Downing 1994~. Primary productivity in more diverse
communities was most resistant to and recovered most rapidly from
drought. Indeed, the least-diverse plots suffered a 4- to S-fold greater
loss of productivity than the most diverse plots, ancE recovered much
more slowly from the drought than the most diverse plots (Figure 4-~.
Diversity increased ecosystem resistance and resilience, because more
diverse areas were more likely to contain some species that were
drought resistant. Those species increased growth in response to the
ctecreased abundances of their drought-sensitive competitors.
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80
Pacific Northwest Forests
0.0 -
-
.0 -0.5
9
-
· -
U)
~ -1 O
3
o
-1 5
T T _ ~ _ T T I
~1
; 1 · . · . I ~ 'T-- ~ . . ~ .
O ~ 10 15 20 25
Plant Species Richness Before Drought
1
-1/2
1f4 o
o
._
Ads ~
o
m
~f46
FIGURE ~1. Effects of plant biodiversity on drought resistance of
grasslands. Source: Adapted from Tilman and Downing 1994.
The significance of biodiversity in ecosystem resilience has been de-
bated for many years, but evidence of its importance is now emerging
from long-term research. Diversity-related resistance is particularly
relevant to the management of agricultural and forest ecosystems to
minimize the spread of species-specific pathogens and pestinsects. And
although single-species plantations result in high levels of production
of specific products or resources, they also have a much higher risk of
infestation than do more complex systems.
McNaughton (1985) presented data for the Serengeti savanna that
demonstrated that areas with greater plant diversity were more resilient
to natural grazing pressure, because ungulate grazers fed more
selectively in areas with greater plant diversity. Plant species not
consumed by grazers were able to increase rapidly in biomass once
freed from competition with the species preferred by the grazers. Those
compensatory increases tended to stabilize primary productivity in
areas with greater plant biodiversity.
OCR for page 81
Status and Functioning
Landscape Change and Threars to Biodiversity
87
The long-term effects of ecosystem destruction and fragmentation are
well known and well documented: habitat loss and fragmentation
inevitably ancE unavoidably lead to species extinctions (e.g., MacArthur
and Wilson 1967; Diamond 1972; Terborgh 1974; Ehrlich anct Ehrlich
1981; Wilson 198S, 1992~. Some rare species are extirpated when the
only areas in which they live are destroyed. However, many other
species are sufficiently harmed by habitat loss that their populations
begin a slow clecline that often encts in extinction (e.g., Diamond 1972;
Terborgh 1974; Wilson 1992~. Some species that survive the destruction
of their prime habitat are left living in marginal habitats in which they
cannot maintain viable populations. Such remnants are "sink"
populations that are maintained only via continued immigration from
neighboring viable habitats. The presence of transient species is thought
to be the primary reason mainland sites of a given size contain more
species than island habitats of the same size (MacArthur and Wilson
~967~. Second habitat fragmentation isolates populations and decreases
localpopulation size. Smallpopulation size increases the chance of local
extinction (May 1973), and isolation decreases the chance of
recolonization by members of the same species. The long-term net effect
is the eventual extinction of formerly abundant species (Tilman et al.
1994~. Recent research postulates that the species most harmed by
habitat fragmentation are the species that are best adapted (i.e., most
specialized for the characteristic conditions) to the region and that they
will undergo selective extinction in the remaining fragments of
protected, undisturbed habitat (Tilman et al. 1994~.
Island biogeographic theory (MacArthur and Wilson 1967; SimberIoff
1984) predicts that the species richness that can be maintained within a
particular region depends on the size of the region and on its degree of
isolation from other regions. The close relationship between species
richness and habitat area has been widely demonstrated (e.g., Tilman et
al. 1994~. Several factors contribute to species-area relationships.
Habitat area alone affects species diversity because smaller areas
support smaller populations that are more susceptible to extinction than
larger populations in larger areas. Area is also a convenient surrogate
variable for environmental characteristics correlated with it. For
instance, as habitat size increases, the range of environmental conditions
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82
Pacific Northwest Forests
includecl within the habitat also might increase. In spatially homoge-
neous and heterogeneous regions, species richness increases with area,
but the intercept and slope of that relationship depend on habitat
heterogeneity (e.g., Simpson 1964; Pianka 1967; Greenstone 1984; Miine
and Forman 1986~. For instance, Simpson (1964) showed that mammal
species per unit land area were greatest in areas of North America with
the greatest topographic relief. The degree of isolation of an area, such
as the distance of an island from the mainland or the fragmentation of
forest patches by intervening agricultural activity, influences the rate at
which species immigrate to that area. The dependence of species
richness on area has been strongly supported by several studies from a
wide range of ecosystems (e.g., Smith 1974; Power 1975; Rosenzweig
1975; SimberIoff 1976; Molles 1978; Nilsson and Nilsson 1978; Rey 1981;
Welis 1983; Brown anc! Gibson 1983; Malmquist 1985; Rydin and
Borgegarc! 1988; T~omolino et al. 1989~.
Over the past decade, the concept of metapopulation dynamics has
been developed (see Hanski and Gilpin (1997~. Various authors have
addressed the clevelopment of the metapopulation concept (Hanski and
SimberIoff 1997), effects of population fragmentation into more or less
isolated demes on dispersal dynamics (Harrison and Taylor 1997),
consequences for gene flow and genetic heterogeneity of the
metapopulation (Hedrick and Gilpin 1997), capacity for recolonization
following local extinction of isolateci demes (Foley 1997; Thomas and
Hanski 1997; Wiens 1997), ability of host-specialists to track their hosts
in the shifting mosaic (Frank 1997), and application for conservation
(Hanski and Simberioff 1997~. As we wouIct expect, populations
maintain themselves as long as recolonization and population growth
produce sufficient numbers of dispersing individuals to balance local
extinction. Any population can be reduced or fragmented to the point
where this balance no longer is maintained. Therefore, the most
vulnerable taxa are, of course, those species that have relatively low
reproductive rates and relatively low vagility, and thereby are less
capable of maintaining gene flow and of recolonizing habitat islands in
fragmented! environments (Samways 1995~. A number of studies from
other regions have demonstrated reduced species diversity, influx of
invasive species, and altered ecological function following fragmenta-
tion, e.g., Aizen and Feinsinger (1994), Bawa (1990), Klein (1989), Powell
and Powell (1987), Punttila et al. (1994), Steffan-Dewenter and
OCR for page 83
Status and Functioning
83
Tscharntke (1997~. Pollination and seed dispersal may be processes at
particular risk, especially for increasingly isolated understory plants that
depend on insect pollination and seed dispersal, as opposed to wind
pollination, for successful recruitment (Aizen and Feins~nger 1994; Bawa
1990; Powell and Powell 1987; Steffan-Dewenter and Tscharntke 1994),
a dependency shared by many understory plants in Pacific Northwest
forests.
Other studies describe how habitat loss relates to the loss of
biodiversity (Harris 1984; SimberIoff 1984; Wilson 198S, 1992~. The
relationship between the amount of habitat exploited and the rate of
extinction is not linear, nor are the extinctions resulting from habitat
destruction usually instantaneous.
A recent experimental study of habitat fragmentation provides
insights into some of its other effects (Kruess and Tscharntke 1994~.
Kruess and Tscharntke established local plant populations that differed
in their degree of isolation and fragmentation. They observed that the
plant-feeding insects that attack the plant species were equally good at
colonizing plants, independent of the degree of plant population
isolation. However, the predators and parasitoids of the insects were
much less abundant in more isolated, fragmented populations. Plant-
feeding insects in large, intact stands of this plant species were 2 to 5
times more likely to be attacked by predators and parasitoids than those
in fragmented, isolated patches. Thus, habitat fragmentation, by
preferentially harming predators and parasitoicEs, led to increased
damage to plants from herbivores. Those results are consistent with
Schowalter's (1995) results in Pacific Northwest forests.
Figure 4-2 illustrates the link between habitat loss and species
extinctions. it is assumed that the relationship between the number of
species (S) and area (A) can be approximated by the simple formula S =
cAZ, where c is equivalent to the average number of species encountered
in an area of unit size, and z is the fractional increase in species per
fractional increase in area. The effect of habitat loss on regional species
diversity is a sharply increasing function of Me proportion of the habitat
destroyed. The two curves shown in Figure 4-2—for z values of 0.1 and
0.35—span the known range of z values, and thus represent likely upper
and lower bounds on the effects of a given degree of habitat lass on the
eventual extinction of species. For instance, the destruction of 50% of a
habitat should lead to the eventual loss of 7-22% of the existing species;
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Status and Functioning
97
protrude above the general forest canopy, within a mile of open water.
On the Columbia River estuary of Oregon and Washington, bald eagles
selected remnant stands of old-growth forest near shoreline for nesting
habitat (Garrett et al. 1993~. Because bald eagles are sensitive to
disturbance, human disturbance needs to be minimized curing nest-site
occupation.
v
~ ^ it,
Flammulated owl, boreal owl, great gray owl (Strix nebulosa), white-
headed woodpecker, black-backed woodpecker (Picoidles arcticus), three-
toed woodpecker (Picoides tridaclyZus), several hawks and owls, and
many other bird species are dependent on old-growth interior forests.
Pileated woodpeckers (Dryocopus piZeatus) require forests with an old-
growth component (McClelland 1979~. Lewis' woodpecker is most
abundant in forested communities with large trees (Robbing et al. 1983),
which indicates that efforts to promote original gallery pine forests
might benefit that species. Goldeneyes (BucephaZa cZanguZa), kestrels
(FaZco sparverius), Williamson's sapsucker (Sphyrapicus thyroi~leus),
mountain chickadee (Paws gambeZi), yellow-belliec3 sapsucker
(Sphyrapicus Darius), mountain bluebird, and common flicker (CoZaptes
auratus) all prefer to use large-diameter trees for nesting (McClelland et
al. 1979~.
Large mammals, including the grizzly bear, black bear (`Ursus
americanus), American marten (Martes americana), fisher (Martes
pennant)), lynx (Lynx canadensis), and wolverine (GuZo guZo), inhabit
forests In the Pacific Northwest and can be affected by management
practices (Henjum et al. 1994; Ruggiero et al. 1994~. Human-caused
mortality is the major limiting factor for the grizzly across its range, and
bears are at heightened risk in areas with roads. Togging from occupied
grizzly habitat might improve forage conditions by creating herbaceous
and shrubby habitat, but it also reduces habitat suitability by elim~nat-
~ng overstories used for security and resting areas. The intensified
human activity that accompanies timber operations is an additional
threat to the grizzly.
Other mammals, including the gray wolf (Cants lupus) and moose
(AZces Dices) rely on Pacific Northwest forest habitat. Restoration of the
gray wolf to the northern Rockies presents management challenges
similar to the grizzly, and timber-management guidelines that benefit
the prey of wolf also should benefit the wolf and the species on which
it preys. Moose habitat in central Idaho forests consists of closed-
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98
Pacific Northwest Forests
canopy forests, often with Pacific yew in the understory (Peek et al.
1987~. Poaching is known to be an important limiting factor for moose
in Idaho (Pierce et al. 1985~.
The northern flying squirrel and Townsend's chipmunk are important
prey species in Pacific Northwest forests (Carey et al. 1999~. The flying
squirrel consumes ectomycorrhizal fungi most abundant in late-
successional forests. This squirrel is also a major prey species for the
northern spotted owl. The chipmunk also consumes fungi, as well as
seeds anti fruits. These species, along with the red-backed vole, may be
good ~ncticators of functioning in these forests (Carey et al. 1999~.
Recommendations to conserve biocliversity have included thinning
second-growth forests less than 50 years old at variable densities to
increase crown differentiation, canopy stratification, and understory
development. Retention of legacies by minimizing site preparation and
burning so that ectomycorrhizal fungi link early in the successional
process also have been recommenced. Various activities to retain coarse
woody debris, long, and snags in these forests could also be directed at
maintaining fungi and associated small mammal prey species.
Amphibians and reptiles have begun to receive attention as concerns
mount over their recent declines (Gibbons 1988; Raphael 1988~. Older
forests appear to support a greater diversity of species of herpetofauna
than younger forests (Raphael 1988; Welsh and Lind 1988), but Raphael
(1988) pointed out that structural characteristics, including coarse
woody debris, hardwooct understory, and abundance of insist sites are
the critical criteria that dictate presence or absence of amphibians and
reptiles. in addition, forest fragmentation might have an isolating effect
on herpetofauna, which disperse across unsuitable habitat with
difficulty. No Pacific Northwest forest amphibians are listed as
threatener! or endangered. In Idaho, the Coeur d'Alene salamancler,
which inhabits moist talus adjacent to forested areas, is classified as a
species of special concern (Groves and Meiquist 1990), ant] nine species
of herpetofauna are listed for monitoring status in Washington.
Salmon and Other Fisheries
Anadromous salmon (salmon that spawn in freshwaters and migrate to
the ocean) and steelhead trout in the Pacific Northwest are rapidly
declining in abundance. Substantial numbers of native genetically
OCR for page 99
Status and Functioning
99
distinct lines (stocks) have become extinct within the past century.
Although debates over causes for the declines are heated, the status of
many stocks is not in dispute. Most of the salmonid stocks of the
mainstem Columbia River have been declining in abundance through-
out the past century and have decreased sharply since 1970. Salmonids
of coastal streams have exhibited similar, although less consistent,
declines. Habitat and water-quality degradation, often associated with
foresttg and other land-use practices, are widespread regional problems.
Few areas of high-quality salmon habitat remain, and many of the best
remaining sites are in forested headwaters (NRC 1996~.
Declines in Pacific salmon have been noteci since the late IS00s. As
the nation's attention turned to the West Coast and the California gold
rush, logging and salmon fishing became major commercial industries
in the Pacific Northwest. The first commercial salmon harvest occurred
in the Columbia River in IS61. The first hatcheries were established on
the McCloucl River of California in 1872 and on the CIackamas River of
Oregon in 1876 in response to decreases in salmon caused by overfishing
and habitat degradation (Stone 1897~. By the turn of the century,
Washington, Oregon, Idaho, and California had passed laws to limit the
freshwater harvest of salmon, anct statutes were enacted to prevent
private damming of streams and dumping of material into surface
waters.
Resource-management agencies in the Pacific Northwest have raised
concerns over the continuing decline of salmon throughout the
twentieth century and developed numerous regulations to protect fish
and habitats. T~ocal areas or specific stocks of salmon have benefitted
from these actions, but because of geographic scope and cumulative
effects of human activities, Pacific salmon have continued to decline.
Indeed, in response to the continuing problem, Oregon banned
commercial and sport salmon fishing in 1994, but reopened it in 1999.
Recent analyses of available information on specific stocks of the five
species of Pacific salmon document the regional and pervasive extent of
the loss of salmon (NRC 1996~. The Northwest Power Planning Council
(1986) estimated that the numbers of salmon in the Columbia River
basin declined from 10-16 minion before the mid-nineteenth century to
fewer than 2.5 million fish in the 1970s. In 1987, the U.S. Fish and
Wildlife Service examined trends in salmon numbers for Alaska,
Washington, Oregon, Idaho, and California during 1968-1984 (Konkel
and McIntyre 1987~. Thirteen of the 657 salmon populations for which
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Pacific Northwest Forests
adequate clata were available became extinct during those 17 years.
Significant trencts, either increasing or decreasing, were observed for
30% of the populations—populations in Alaska tended to increase, but
24% of the populations in the Pacific Northwest declined significantly
during the study period.
In 1991, the American Fisheries Society examined available evidence
on the status of anadromous salmonid stocks in Washington, Oregon,
Idaho, and California and concluded that more than 106 stocks have
become extinct. Of the remaining stocks, 160 were classified as at
serious risk of extinction, and an additional 54 were considered of
special concern. Salmon of the Columbia River have undergone the
greatest proportional loss of stocks, reflecting the history of intense
cotnmercial fishing, loss of habitat, and mortality associated with dams.
Subsequent evaluations of the status of salmon in the Pacific
Northwest support the conclusions of the American Fisheries Society
and identify ac[ditional stocks or habitats that have been lost or face
serious risks. The Wilderness Society (1993) formed a pane} of regional
experts that concluded that salmon habitat on public and private lands
has been seriously impaired by land-use practices, including forestry.
All species of anactromous salmonicts in the Pacific Northwest have
undergone stock extinctions within the past century, and pink salmon
is the only species in which the majority of stocks are not known to be
declining.
Causes for the declines of anactromous salmonids are numerous and
include habitat loss and conversion of streamsicte forest to agricultural,
range, residential, or urban lands. Forest practices on existing forest
lands, agricultural practices, grazing, dams Mat block passage, excessive
commercial harvest, and sport harvest also contribute to declining
populations (NRC 1996~.
Over the past several years, scientists have established links between
ocean conditions and upwelling, which brings nutrients into offshore
environments. Production and survival of mature fish in the ocean is
tied to climatic cycles in the North Pacific and to fishing pressure.
Evidence of long-term cycles can be seen in the comparison of coho in
Oregon and Washington and pink salmon in Alaskan fisheries. Catches
of adult coho in Oregon and Washington are low when catches of pink
salmon are high in Alaskan waters, changing over cycles of 20-35 years.
But historic cycles are a small part of the current declines in salmon. The
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Status and Functioning
107
declines are relater! to offshore conditions as well as forest and other
land-use practices.
Long-term records of cutthroat trout in the Alsea watershed
Oregon detnonstrate that populations in watershed dominated by
clearcuts are still at less than one-third of historic levels. The decrease
is not attributable to ocean conditions, because cutthroat numbers in
adjacent basins with mature forests are still at levels observed before
harvest. Much of the population in the area is resident and not linked
to the ocean environment.
Numerous other factors have contributed to the decline in salmon
populations, including mar~ne-mammal predation. Marine mammal
protection has allowed populations of seals and sea lions to recover, and
they consume returning adult salmon in the estuaries and mouths of
coastal rivers. However, most analyses estimate that salmon make up
a minor portion of the diet of marine mammals, and the adctitional
mortality has a minor effect on salmon populations.
Invertebrates
Invertebrate species characterizing disturbecl or early successional
systems typically are adapted to wide variation in environmental
conditions. Rather than being characteristic of particular communities,
they often are present as parts of relatively nondistinct species assem-
blages. (That is, early successional assemblages tend to be made up of
widespreact, disturbance-adapted species that are not specific to a
particular location. They are, of course, distinct from later successional
assemblages, as demonstrated by Schowalter (1995~. For example, many
of the herb and shrub species characteristic of early successional
communities in Westside forests (e.g., snowbrush (Ceanothus veZutinus),
manzanita (ArctostaphyZos), fireweed (Epilobium angustifoZium), and
daisies) are present in various early successional or frequently disturbed
meadow and shrub communities from Alaska to northern California,
and the species of aphids, ants, and other invertebrate species associates]
with those communities are widely distributed in western North
America (Furniss and Carolin 1977~.
Species closely associated with late-successional forests typically
depend on the moderate conditions and resources provided by that
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Pacific Northwest Forests
forest structure. The organisms often cannot survive the more extreme
conditions of open or disturbed areas (Seastedt 1984; Schowalter et al.
1986; Schowalter 1989~. Many invertebrates, especially those that
disperse by flying or ballooning, are capable of moving extensively in
search of resources and can stray or be blown into unsuitable areas. But
efforts to survive and reproduce do not necessarily contribute to stable
species populations. For example, Schowalter (1989) found grass- and
crop-feecling insects at the tops of old-growth Douglas-fir, ancl Edwards
(1987) documented an extensive invertebrate deposition on glaciers on
Mt. Ranier. Schowalter (1995) found old-growth Douglas-fir canopies
had the highest invertebrate biodiversity, followed by mature Douglas-
fir, shelterwood Douglas-fir, and then regenerating Douglas-fir, which
hail substantially lower invertebrate diversity. A largely different suite
of invertebrate species lived in the canopies of old-growth hemlock
(Schowalter 1995~. Thus, the mix of Douglas-fir, hemlock, and other
plants characteristic of old-growth, but not younger, forests substan-
tially increases invertebrate biodiversity. Remnant mature trees in
thinned stands (shelterwoods) can provide important refuges for many
invertebrate species. Late-successional forest thus provides resources
critical to survival of many species that, in turn, contribute to productiv-
ity of those forests. Many arboreal and forest-floor arthropods are
abundant only in old-growth forests (e.g., Schowalter 1995; Winchester
1997), but most species remain poor known.
Fungi
Mutually beneficial associations between fungi and plant root systems
are frequent in plant communities around the world (Harley and Smith
1983; Brundrett 1991~. Fungi and plants are physiologically interdepen-
dent, with plants supplying the energy needs of fungi and fungi
providing the plants with nutrients taken from the soil that plants need
but are unable to access in sufficient quantities. Symbiotic fungi also can
defend plants against root pathogens, and they protect plants against
heavy-metal toxicity (Vogt et al. 1987; Vogt et al. 1991; Wilkins 1991 ).
Mushrooms, including those collected for human consumption, are
the reproductive structure of filamentous fungi that form "ectomycor-
rhizal type" associations with trees. The Pacific Northwest is one of the
prime areas for mushroom collecting in the world, in the dry Eastside
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Status and Functioning
703
Ponderosa pine forests and the wet Westside Douglas-fir and hemlock
forests (Molina et al. 1993; Pin and Molina 1996~.
In the dry and wet forests of the Pacific Northwest, mushrooms are
an important food source for small mammals, and some species, such as
the California red-backed vole (Clethrionomys californicus), are almost
totally dependent on mushrooms for their subsistence (Foge] and
Trappe 1978; Maser et al. 1978~. The California red-backed vole is found
in young, mature, and old-growth forests in the Pacific Northwest but
is common ~ old-growth (Maser et al. 1978~. In western Oregon
coniferous forests, some small mammals disappear from areas where
trees have been harvested and only reappear when the regenerating
forest reaches the pole-sapling stage, because they are dependent on the
coniferous forest canopy (Ure and Maser 1982~.
Mycorrhizal
fungi are crucial to
ecosystem processes
because they facili-
tate and accelerate
the rate at which
plants are able to re-
establish and recover after disturbances (Trappe and Luoma 1992~. in
Eastside and Westside forests, the maintenance of the symbiotic
associations in forests is correlated closely with the presence of large,
coarse, woody debris. Large, coarse wood is important In the ability of
symbionts to survive over the short term (within annual cycles of
drought) (Harvey et al. 1978) as wed as over the long-term (e.g., from
one disturbance to the next). The direct effects of the loss of mycorrhizal
fungi or their propagules associated win forest-management practices
on regeneration is dramatically shown in a study conducted by Perry et
al. (1992~. Forest-harvest practices (e.g., applying herbicides to
deciduous shrubs that maintained the inoculi) caused loss of fungal
inoculum, resulting in failure of woody revegetation at the site.
Lan~/-use ant/ forest-management
practices have greatly inf/uenced
popu/ations of numerous species, and
have placed some of these species in
danger of extinction.
V/ABLE POPULAT/ONS
AND THE CONSERVATION OF B/OD/VERS/TY
T~and-use and forest-management practices have greatly influenced
populations of numerous species, and have placed some of these species
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Pacific Northwest Forests
in danger of extinction. To prevent extinction, viable populations must
be managed. A viable population is the number of individuals "that will
insure (at some acceptable level of risk) that a population will exist in a
viable state for a given interval of time" (Gilpin and Soule 1986~. That
can refer either to local populations or metapopulations, depending on
the circumstances.2
The number of individuals that constitute a viable population is
difficult to determine but depends in part on
· The population structure, social dynamic, and breeding characteris-
tics of the species in question.
· Environmental fluctuations, particularly the possibility of cata-
strophic events that sharply reduce population size (Shaffer 1981~.
· Environmental stresses, such as pollution, that reduce the vigor of
individuals.
· Various aspects of habitat quality' including, in particular, the
manner in which habitats are arrayed across the landscape" e.g., in
large blocks, isolated fragments, or fragments connected by habitat
bridges.
· The period for which viability is being assessed.
Populations that drop below some minimum size are drawn into
what Gilpin and Soule (1986) call the extinction vortex—i.e.,they will
become extinct unless extraordinary measures are taken. Populations
that are losing individuals become at risk before that point, however,
and when the population reaches a size near We extinction vortex, a
random environmental event (such as drought, unusual weather, or
disease) may draw the population into the vortex. Most information
concerning minimum population size comes from models of vertebrate
populations. They have not been tested for the plants, insects, and
microbes that constitute the vast majority of Earth's biota.
Population size affects the survival chances of individuals in two
~eneralwavs {Soule 1983~1986: Lands 19881: one related to demo~ranhv
i, _ , _ ~~ ~ __ ~ -I ~ ~~r~
2In this discussion, "population" is used to cover both possibilities. For
example, saving enough old-growth Douglas-fir to support 100 spotted owls
will not save the owl, because it takes more than 100 individuals to maintain
a viable population, and probably more than 1,000.
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Status and Functioning
705
(the patterns of population growth and decline), the other to the ability
retain genetic variability and avoid inbreeding.
Demographic Factors
Populations of most species fluctuate depending on numerous factors
in their environment; the smaller a population is, the greater is the
chance that one of its down cycles will either lead to its extinction or
reduce it such that genetic or social cteterioration triggers a slow slide to
extinction. The population size at which a slide to extinction begins is
partly determined by population genetics—especially inbreeding and
loss of genetic diversity. Nongenetic factors also are important. In some
species, small populations have relatively high genetic diversity, yet still
are at risk because of random variation in population size.
The risk of extinction from demographic variations in population size
is particularly high where factors such as infectious disease or large-
scale natural events (e.g., hurricanes and wildfires) have the potential to
reduce population sizes sharply (Hanson and Tuckwell 1981; Lande
1993~. Small populations cto not contain sufficient individuals to be
buffered against such catastrophic losses. A particular threat is the
possibility that two or more events that reduce population size occur In
quick succession, not allowing the population time to recover from one
before it is further reduced by another. Such multiple threats are most
likely on lanctscapes where human activities are greatest.
Some animals depend on numbers for defense, foraging, or effective
breeding. Such species have a social threshoict population size below
which the group becomes dysfunctional and unable to persist. In many
cases, the social threshold is considerably higher than that determined
solely by demographics or genetics (Soule 1983~.
Catastrophic events of one kind or another occur in almost any
environment, and local populations of many species may disappear
periodically from a given region. A species is not threatened by this as
long as other healthy populations can provide a source of immigrants
to replace local losses. However, when those other populations are in
decline or highly fragmented and isolated, losses within local ecosys-
tems may not be replaced, and extinction may occur. It is estimated that
100,000 to 300,000 species worldwide are threatened with extinction via
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Pacific Northwest Forests
this mechanism (Culotta 1994~. The area required to buffer species
against demographic calamities can be quite large: 200,000 ha or more
for some species that live In forested habitats (Shugart and Seagle 1985~.
Generic Factors
Small populations sizes usually result in loss of genetic variation,
particularly if the populations remain small for any length of time.
Small, isolated populations can be at risk genetically for two reasons.
First, fewer individuals exchanging fewer genes leads to inbreeding,
which frequently results in loss of vigor anct inability to resist stress
(T~ancle 1988~. Second, small populations tend to lose genetic variability
through random processes called genetic drift.
An effective population size is the number of breeding adults that
would provide the rate of inbreeding observed in a population if mating
were random and the sexes were equal in number. Therefore, effective
population size is the number of males anct females that actually
contribute equally to the gene pool, on average, from generation to
generation In many species, many incLividuals clo not breed or produce
relatively few offspring; the effective population size is much less than
the actual population size. For most mammals and birds, effective
population sizes are no more than 25-50% of actual population size.
Animals that live in family groups, such as wolves, typically have only
one mating pair per group. Soule and Wilcox (1980) estimated that an
actual population size of 600 wolves would be required to maintain an
effective population of 50 that prevented inbreeding depression.
The genetic factors involved in population sizes for plants have
additional aspects not encountered in higher animals. Plants rely on
some intermediary—wind or animals—for pollen and seed dispersal.
Most pollen falls in the immediate neighborhood of the contributing
parent, and plants with seeds dispersed by animals depend totally on
the welfare of those animal vectors, including the area needed by the
vectors to maintain viable populations. In addition, trees, probably
because they are long-lived, accumulate more harmful genes than
animals, and hence are likely to require a high level of outbreeding to
prevent inbreeding depression (Iedig 1986~. Management-related
genetic selection for desirable growth traits has the potential to
artificially narrow the gene pool of a whole species—selected and wild
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Status and Functioning
707
trees—because selected individuals presumably contribute to the gene
pool in proportion to their numbers (Ledig 1986~.
Popu/ation Viabi/ityAna/ysis
Theoretical considerations and empirical observations suggest that
when genetic and demographic factors are accounted for, population
of animals and plants must contain at least several thousand inclivictuals
to be viable (Whitford 1983; Soule 1987; Thomas 1990~. However, no
single number is applicable to all species, nor does a single number
necessarily apply to any one species in all environmental situations
(Gilpin ancl Soule 1986; Thomas 1990~.
Population viability analysis (PVA) can be used to identify threats
faced by a species and evaluating the likelihood that it will persist for a
given time into the future. PVA models take into account the view of
habitats as landscape phenomena and can incorporate numerous
features, including demographic stochastisity, environmental uncer-
tainty, natural catastrophes, and genetic uncertainty. Conservation and
recovery planning often must account for those and other variables,
particularly because many endangered species exist in small popula-
tions, and without appropriate planning, a single event might destroy
an entire species (NRC 1995; Akcakaya et al. 1999~. Several PVAs have
been conducted in recent years, including ones for the marbled murrelet,
the northern spotted owl, ant] the red-cockaded woodpecker.
Representative terms from entire chapter:
population size
