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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

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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 biodiversityi.e., the loss of species from a given habitatcanbe 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 worldsuch as forests ancl coral reefsand 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

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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.

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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

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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 4-2for z values of 0.1 and 0.35span 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

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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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700 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 populationspopulations 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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702 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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704 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 vortexi.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 geneticsespecially 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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706 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 intermediarywind or animalsfor 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 speciesselected and wild

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Status and Functioning 707 treesbecause 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.