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Environmental Issues in Pacific Northwest Forest Management (2000)

Chapter: 4 The Status and Functioning of Pacific Northwest Forests

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Suggested Citation:"4 The Status and Functioning of Pacific Northwest Forests." National Research Council. 2000. Environmental Issues in Pacific Northwest Forest Management. Washington, DC: The National Academies Press. doi: 10.17226/4983.
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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

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.

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

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-

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.

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

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.

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.

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

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

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;

84 Pacific Northwest Forests Li}<ely UPON and Lower Bounds far Ext~n~ons O.S £ - __ Q 0~4 Q ~2 1,0 Z ~ O.35 / 0.0 0~2 0.4 o.6 0~8 4.o Proportion of Habitat Lost FIGURE 4-2. Effect of habitat loss on the extinction of species, derived from the species-area relationship (see Tilman et al. 1994~. destruction of 95% should lead to the extinction of 26-65% of the species. A srnaD increase in the proportion of habitat destroyed wouict lead to many more extinctions in an already fragmented landscape than would a comparable increase in destruction of an intact landscape. For a virgin mainland habitat, z should be about 0.15, but should increase to as much as 0.35 as more of the habitat is destroyed and made more isIandlike. Much of the previous discussion is based on theory and data from ecosystems outside the Pacific Northwest. Based on these consider- ations and given that about 80-90% of the Pacific Northwest has already been subject to forest cutting, one might suggest that additional loss of old-growth forest to agriculture, sprang] or short-rotation silviculture might threaten an ever-larger number of species. The steeply rising curves in Figure 4-2 show that loss of an additional ~ % of the original habitat when 90% has already been destroyed is expected to lead to the extinction of more than 4 times as many species as would a 1% loss in habitat when only ~ 0% has been destroyed. Certainly, no one would argue that habitat fragmentation or Toss have

Status and Functioning 85 anything but negative effects on the viability of species populations. Furthermore, those effects are likely to increase disproportionately as total habitat area decreases and fragmentation increases. But several factors make it very difficult to make quantitative predictions of the extent of actual loss we might expect in the Pacific Northwest for some additional increment of old-growth forest loss (Rochelle et al. 1999~. First, the analogy between a fragmented landscape ancE islands in an ocean is imperfect. - _ The matrix of cLis- turbed lances in the Pacific Northwest represents a com- plex collection of forested and nonforestecE lands that vary in their value as habitat and dispersal corridors. Second, relatively few species are entirely restricted to o1~-growth forests, making generalizations about the specific number of species jeopardized by an increment of old- growth loss impossible. Third, forests are constantly changing; cut-over forests do regrow ant! eventually reacquire oIci-growth features. These caveats do not change the fact that continuer! cubing of old- growth forests and conversion of naturally regenerating forests to other conditions poses greater risks for the biodiversity of the Pacific Northwest now than it ever has. This is a matter of particular concern for those species that depend most on the habitatieatures (e.g., stanching and downed woody debris) of old-growth forests. Pest outbreaks are nor random events that threaten forest resources but rather, at /east in part, are predictab/e responses to changes in forest condition; thus, they are va/uab/e ina/icators of changes in forest condition. Diseases and Pests Pest outbreaks are not random events that threaten forest resources but rather, at least in part, are predictable responses to changes in forest condition; thus, they are valuable indicators of changes in forest condition. Outbreaks of insect and pathogen populations are affected bv three general factors: 1) host condition: 21 host abundance ancE 3) host ~ ~ , - , , apparency, i.e., the ease with which an insect or pathogen can find a suitable tree species.

86 Pacific Northwest Forests Host condition. Host condition indicates plant responses to environ- mental factors. Plants respond to stresses such as extreme temperatures, flooding, and moisture or nutrient limitation by adjusting their physiological processes to increase evapotranspiration or to conserve water and nutrients (Mattson 1980; Mattson and Haack 1987; Waring and Cobb 1992; Lorio 1993; Koricheva et al. 1998; Schowalter et al. 1999~. Defensive chemicals, such as phenolics and terpenoids, which normally protect healthy plants from insects and pathogens, consume energy ant] nutrients that otherwise conic] be used for growth or to meet more immediate metabolic requirements associated with plant survival (Perry and Pitman 1983; Rhoades 1983; Strong et al. 1984; Becerra 1994; Harborne 1994~. Stressed plants typically reduce production of such defensive chemicals (Rhoades 1983; boric 1993~. Plant-feed~ng insects and pathogens have a wide range of abilities to colonize host plants, depending on their tolerances or ability to detoxify plant defensive chemicals (Harborne 1994~. Some insects and pathogens respond rapidly to small changes in concentrations of defensive chemicals in milcIly stressed plants, whereas others can reach outbreak levels only if there has been a considerable reduction in plant clefensive- chem~cai production (folio 1993~. Some of the more virulent early colonists feeding on mainly stressed plants can exacerbate stress and predispose plants to other insects and pathogens (Schowalter 1985; Paine et al. 1993), leading to rapid decline in plant condition. For example, tree defoliation after infection by root diseases often weakens trees sufficiently to permit colonization by tree-kiHing bark insects (Goheen and Hansen 1993~. Hostabundiance. Host abundance strongly effects insect end pathogen populations. Insects and pathogens usually occur in scattered stressed or injured plants, where their effects are primarily to thin stancis and accelerate successional transitions from shade-intolerant to shade- tolerant vegetation (Schowalter 1981~. Insect and pathogen populations typically remain small overall because individuals dispersing in search of new hosts are vuInerabie to a variety of mortality agents, including adverse weather conditions and predators. Most major outbreaks occur when insect and pathogen populations can grow on closely spaced and suitable hosts, a condition found most often in forest stands dominated by a single species, such as in forest plantations. Survival increases

Status and Functioning 87 because more suitable hosts are available end easily reached (Risch1980, 1981; Strong et al. 1984; Courtney 1935; Schowalter and Turchin 1993, Schowalter et al. 1986~. Host apparency. This term refers to the ease with which an insect or pathogen can find individuals of a suitable tree species. Insects adapted to feeding on particular plant species often use the defensive chemicals produced by their host plants and carried in the forest vapor stream (especially if plant injury increases chemical release to the airstream) as chemical beacons to locate host plants (Stanton 1983; Schowalter et al. 1986; Visser 1986~. Chemolocation can be disrupted by nonhost plants that produce and emit other chemicals that do not attract or even repel insects or pathogens (Visser 1986~. Thus, in vegetation containing a mixture of many plant species, host-finding by insects is difficult. In contrast, the strongly attractive vapor stream of single-species stands facilitate host-finding (Risch 1980,1981; Stanton 1983; Strong ~t ~1 1984 Courtney 1985; Schowalter 1986; Visser 1986). . ~ _ ~ · ~ , As a result of host condition, abundance, and apparency operating interactively, single-species plantations are particularly vulnerable to pest outbreaks. Commercial tree species typically are rapidly growing species that easily become stressed by water or nutrient limitation (Mario 1993~. Stress, combined with abundance and apparency, greatly increase the likelihood of pest outbreaks in single-species plantations, especially in forests where these conditions persist for long periods (Kareiva 1983; Schowalter et al. 1986; Schowalter ant] Turchin 1993~. The trend ~ silviculture toward planting genetically improved stocks might increase such risks by diminishing genetic variability in disease resistance. incidence of Pest and Disease Ourbreaks Susceptibility to pests and disease in the Pacific Northwest has increased primarily in Eastside forests. The change in condition has been created by altered tree species compositions since the late I8OO's (Carlson and I=otan 1988; Hagle and Goheen 1988), aggravated by drought (Patterson 1992~. Selective logging, especially removal of shade-intolerant species (e.g., western larch and ponderosa pine) has increased the abundance

88 Pacific Northwest Forests and distribution of drought-intolerant Douglas-fir and true firs, which have become susceptible targets of defoliators, bark beetles, and root pathogens (HaglLe and Schmitz 1993~. Various agents in western forests—tree-feeding insects, pathogens, and fire—have positive and negative roles. For example, stem decay fungi are particularly important to cavity-nesting birds, because they soften heartwood end facilitate excavation of cavities in living trees (Bull and Partridge 1986~. "Witches-brooms" formed by dwarf mistletoe (Arceuthobiun~ spp.) provide shelter for owls and porcupines. Root rots create patchiness that diversifies plant species composition. At low levels, spruce budworm (`Choristoneura occidentaZis) kills weaker trees, thereby providing more resources for the ~nore-tolerant survivors. Although some native pests occasionally kill trees over large areas, the structure of ecosystems and landscapes, along with the ecological relationships mediated by that structure, often dampens and absorbs the spread of disturbances, diseases, and pests (Perry et al. 1989b, 1994 ). Some management practices for Westside forests were developed to minimize real or perceived pest problems. For instance, Port Orford cedar root rot (Phytophthora ZateraZis) has been limited by sanitation of logging equipment and forest closure. (The rot was introduced into high-elevation areas by vehicle movement from infected valley areas and subsequently transmitted by down-slope water flow.) Slash burning and other residue treatments have been prescribed to prevent spread of bark beedes anct pathogenic fungi into crop trees. However, few of those insects or fungi were problems in old-growth forests, because species diversity and diversity of natural enemies limitecI pest abundance and spread (Schowalter 1989; Goheen and Hansen 1993; Schowalter 1995~. Lower tree species diversity in managed forests on the Eastside has increased the degree of disease development and severity on many sites and contributed to the higher mortality rates of trees than are evident on the Westside. For example, current annual mortality of Eastside Douglas-fir is estimated to be 4.~%, which is 4 times higher than its natural mortality rate; it is mainly due to fungal diseases (Hagle and Goheen 1988; Spies et al. 1990~. Natural mortality is caused primarily by the thinning associated with stand development as trees increasingly compete for space and also to periodic beetle outbreaks (Rebertus et al. 1992~. Hagle and Goheen (1988) suggested that Douglas-fir anct the true

Status and Functioning 89 firs are most susceptible to damage by root diseases in western Montana, northern Idaho, eastern Oregon, and eastern Washington and predicted chronic mortality of these species from root diseases would occur throughout the lives of the stands in which they are present. Nutrient status of a site and disease also are causally associated. For example, ArmiZlaria root rot problems are more severe in productive sites than in harsh, iess-productive sites because of more intense logging on productive sites and the resultant increased true fir component (Hagle and Schmitz 1993~. Some insect outbreaks result from changing climatic conditions, such as droughts, and are especially severe on sites where dominance has shifted to tree species susceptible to insects and pests (Mattson and Haack 1987~. For example, in west-central Idaho in the Cascade Ranger District, precipitation decreased 25% from 1985-1990. Mature Dougias- fir trees became stressed and were attacked by the Douglas-fir bark beetle (Dendroctonus pseudotsugae) (Patterson 1992~. As a result, annual Douglas-fir beetle-related mortality increased from fewer than 1,000 trees killed to more than 15,000 in 1989 (Patterson 1992~. Two major factors underlie current disease and pest problems in forests east of the Cascades crest. First is the large-scale conversion of Eastside forests from a landscape dominated by a mosaic of old-growth pine to one dominated by younger forests with a high component of true firs and Douglas-fir. That has effectively converted the regional landscape from one that dampens and absorbs disturbances, such as fire and pest outbreaks, into one that magnifies them (Perry 198Sb, 1994 ). This is a major reason for proposals to move landscapes of the Eastside back toward Me historic dominance by large ponderosa pine (Agee 1993~. The second factor is the introduction of foreign disturbances, i.e., stresses that have not been present historically, and against which the system has no effective defenses. Exotic insects and pathogens are obvious introductions, exemplified by the presence of nonindigenous white pine blister rust. And the potentialfor introducing and dispersing new exotic pest organisms could be exacerbated by U.S. importation of raw logs. Foreign stresses also include clearcutting and removing coarse woody debris (which reduce habitat for natural predators and parasites of pest insects and pathogens (Campbell and Torgersen 1982; Schowalter 1995), logging with heavy equipment (which compacts soils and diminishes ecosystem resilience), and disrupting Me historic regime

go Pacific Northwest Forests of frequent, gentle fires. Fire reduction anct logging have been the major contributors to change In forest structure and landscape patterns (Agee 1993; Hagle and Schmitz 1993~. Considerable evidence from throughout the interior west indicates that spruce budworm outbreaks cturing this century have been more frequent, more widespread, of longer duration, and more lethal to trees than previously (Anderson et al. 1987; Swetnam ant! Lynch 1989; Wickman et al. 1992~. Increasing populations of defoliating insects during the 20th century have resulted in large part from increases in true fir and Douglas-fir ingrowth in stands previously dominatect by nonhost ponderosa pine and larch (Hagle and Schmitz 1993~. Incidence of the root rot ArmiZZaria might also have been increased by the spread of these tree species. The basic ecological principle of host abundance discussed above is at work on the Eastside. True firs and Douglas-fir have become more abundant during the 20th century because early loggers took only ponclerosa pine, leaving behind the less-valuable firs. Furthermore, the spread of firs was facilitated by fire suppression (Agee 1993~. The resulting shift in tree species composition was dramatic. Eastside commercial forest land dominated by ponderosa pine has shrunk cturing the 20th century 75%-90%, depending on locale. Most of those forests are now dominated by Douglas-fir and true firs that once occurred primarily at the higher elevations. Larvae of spruce budworm and tussock moth (Orgl/ia pseu(lotsugata), the principal outbreak defoliators in Eastside forests, are dispersed largely by winct currents. Larval survival, hence the rate at which an infestation spreads and the area it covers, depends (among other things) on the proportion of dispersing larvae that lands on a suitable host. That depends, in turn, on the relative proportion of host and nonhost tree species across the landscape (Schowalter et al. 1986; Perry 198Sa; Schowalter and Turchin 1993~. When Douglas-fir and true firs existed as forested islands on mountains surrounded by nonhost ponderosa pine at mid and low elevations, dispersal of defoliators was limited. The much greater expanse of firs today facilitates the spread of insects, a phenomenon that has been exacerbated by several additional factors. Very high stocking density in some stands (due to fire exclusion) stresses trees, reducing their ability to resist insect and pathogen attacks. Firs, less-drought tolerant than ponderosa pine, are particularly

Status and Functioning 97 susceptible to drought stress In the lower elevation areas to which they have denigrated since fire exclusion was adopted as a management strategy. Abnormally low precipitation during the 1980s probably contributed to the most recent spruce budworm infestation. Recluced habitat (especially coarse woocly debris) for birds, ants, ancE other natural enemies might have also contributed to increased population growth and spread of ctefoliators (Torgersen et al. 1990~. In 1989, bark beetles and spruce budworm infestation was evident over 3.1 million acres in eastern Oregon. By 1991, more than 4.7 million acres was infested (Oregon Dept. of Forestry 1999~. The area of active infestation has since cteclined' comprising 2 million acres in 1992, and measurements in spring of 1993 indicated that budworm populations were very low. Nonetheless, the potential for widespread infestations of budworm, tussock moth, and ArmiZZaria will remain as long as current forest conditions persist. Comparative studies wouIct help to eluciciate the numerous factors that relate to pathogen and insect outbreaks in different forests and to understanct pathogen and insect prevalence. STATUS OF OTHER PLANT SPECIES In addition to their effects on timber species, land practices in the Pacific Northwest have influenced the stability and abundances of many other plant species of the region. The species that have been most immedi- ately and greatly affected by past lance use in the region are those species most closely associated with late successional and old-growth forests. Because most plant species abundances do not vary greatly among younger, mature, and old-growth forests in western Oregon and Washington (Spies 1991), few plant species have been directly endan- geredby the forest harvesting that has occurred. But that does not mean that their future is ensured. Indeed, habitat destruction and fragmenta- tion might eventually threaten many common plant species of the region (Tilman et al. 1994~. However, Pacific yew (`Taxus brevifolia) and the lichen isobaric spp., which have strong preferences for old-growth forests, are strongly affected by decreases in the spatial extent of old- growth forests (Spies 1991). In Idaho, Henderson et al. (1977) proposed 17 forest plants for federal listing under the Endangered Species Act

92 Pacific Northwest Forests (ESA), including I) for threatened status and 6 as enclangered. In contrast, many more animal species have been adversely affected by lane! use and deforestation. STATUS OF W/LDL/FE Habitat fragmentation, increased abundances of Douglas-fir and true fir, clecreasec] fire frequency, the much greater abundances of early successional forests, soil erosion into streams, anct construction of dams have affected many bird, mammal, fish, and amphibian species of the Pacific Northwest. Species that rely on old growth have frequently been greatly affected, and concern is growing for amphibians and reptiles, neotropical migrant passerine birds, endemic species, and other species that appear to be scarce but for which little is known. And although the ESA has focused attention on species such as the northern spotted owl anct the grizzly bear (Ursus arctos horribiZis), mandates predating the Multiple Use-Sustainect Yield Act of 1960 pertain to national forests and address the need to maintain key ecological processes and bioctiversity. Paulson (1992) summarized distribution and status of the avifauna in the Pacific Northwest, including Idaho, western Montana, Alaska south of the tundra, and British Columbia. Of 380 species that occur regularly in the region, 327 use the area for breeding. An additional Il0 species occur irregularly in the region and were not considerecl. The region has no endemic species, although seven species (trumpeter swan (Cygnus buccinator), Barrow's goldeneye (Bucephala isZandica), red-breasted sapsucker (Sphyrapicus rugar), white-headed woodpecker (`Pacoides aZboZarvatus), chestnut-backed chickadee (Paws ruiescens), varied thrush (Ixoreus naevius), and golden-crowned sparrow (Zonotrichia atricapilla)) are largely restricted to it. Vaux's swift (Chaetura vauxi), rufous hummingbird (SeZasphorus rufus), Pacific-siope flycatcher (Empidonox difficiZis), flammulated owl (Otusflammeolus), and Townsenc['s warbler (Dendroica townsendi) and hermit warbler (Dendroica occidentaZis) breed primarily within the region and winter elsewhere. Breeding species richness (the number of species) is high in the Rocky Mountain portion of this region because of the variety of habitats and climates, with a peak in richness located in northwestern Montana (Cook 1969; Paulson 1992~. Species that have shown sharp declines in the Rocky Mountain region

Status and Functioning 93 include the osprey (Pandion haZiaetus), bald eagle (Haliaeetus ZeucocephaZus), peregrine falcon (Falco peregrinus), willow flycatcher (Empidonax traiZZii), common nighthawk (ChordeiZes minor), Lewis' woodpecker (MeZanerpes Zewis), olive-sided flycatcher (Contopus cooper)), western wood pewee (Contopus sordiduZus), western bluebird (SiaZia mexicana), mountain bluebirct (SiaZia currucoides), loggerhead shrike (Lanius Zudovicianus), and red-eyed vireo ( Vireo oZivaceus) (Paulson 1992~. Habitat loss on southern winter range and on breeding range, habitat fragmentation Mat predisposes interior forest species to increased brood parasitismirombrown-headect cowbirds (MoZothrus ater), and predation from other species are probable causes (Paulson 1992~. Carter and Barker (1993) evaluated the status and reliability of data for these species from breeding bird surveys. Finch (1991) pointed out that status of neotropical migrants in western North America needs more evaluation. For western Oregon (Mesiow and Wight 1975) and the Rockies (Tobalske et al. 1991), logging changes species composition of forest passerines. Ground-nesting species and species that use openings and low-growing vegetation are found in recent clearcuts and early stages of regeneration. Foliage-foraging species and tree gleaners are less abundant in those areas. Conifer-tree-nesting species are least abun- dant, including golden-crowned kinglet (Regulus satrapa), Swainson's thrush (Catharus ustuZatus), varied thrush, and Townsend's warbler. McGarigal and McComb (1995) reported that 10 of 15 passerines were negatively affected by change in habitat area. Only 2 of the 15 species examined were affected by habitat configuration, anct 4 species were positively associated with more fragmented habitats. Seven species were associated with landscapes containing more fragmented late seral forests than expected. These authors conclucled that the relationship between late-seral forest area and bird abundance varied dramatically among species. Raptors (hawks, owls, and eagles) pose special problems for habitat retention in managed coniferous forests in the region. McClelland et al. (1979) recommendect that fores/-management practices should minimize harassment to prevent nest sites and territories from being abandoned, prevent direct destruction of habitat, and provide selective cutting systems that remove only specific trees and ensure retention of snags and potential snags. Uncommon raptor species of Pacific Northwest coniferous forests include the flammulated owl and the boreal owl

94 Pacific Northwest Forests (Aegolius funereus), which are associated with old-growth ponderosa pine forests and spruce-fir forests, respectively (Holt and Hillis 1987; Hayward et al. 1993~0 Goshawks (Accipitergentilis) require protection of riparian habitat and protection of known active nest sites, which usually are found only in mature forest with canopy closure (Thomas et al. 1993~. Because much of the controversy has centered on spotted owl populations, much is known about that species. Noon and Biles (1990) examined population attributes of the species to see what factors are most responsible for population fluctuation. The spotted owlis a typical K-selectec3 species (i.e., has a long generation time and small clutch sizes) with age at first reproduction being 2 years and with 2 eggs produced annually thereafter. Probabilities of survival to age ~ are low, but year-to-year probabilities become increasingly higher thereafter; maximum life is approximately 16 years. Noon ant! Biles (1990) demonstratect that spotted-owl populations are not sensitive to a wicie range of changes in juvenile survival; the factor that most affects population dynamics is aclult survival. Species that exhibit high adult survival and low juvenile survival are thought to evolve in habitat that favors adult longevity. Forests west of the Cascades crest and through- out much of eastern Oregon and Washington are generally more long lived and stable in composition than forests of the northern Rockies, which encourages adult owl survival. Relatively low fecundity precludes rapid repopulation after a ctecline, further suggesting adaptation to long-lived forests. Isolated habitats caused by logging, tree falls, or fire might not be recolonized rapidly, because survival of potential dispersers, which are typically young owls, is low. Table 4-] demonstrates that populations of spotted owls decreased from 1985-1993, and the rate of decrease accelerated in the I] areas studied, consistent with earlier analyses of life-table data by Noon and Bites (1990), which showed a declining population level. The decrease was attributed (Lance 1988; Doak 1989; Thomas et al. 1990) to the loss of mature anct old-growth forest habitat. Thomas et al. (1993) projected habitat declines at rates of 1.~-5.4% per year on intensive study areas. Spotted-ow] home ranges in Westside forests are typically 1,200-2,000 ha, of which 20-50% is old-growth forest (Thomas et al. 1990~. FEMAT (1993) evaluated numerous species in the range of the spotted owl with regard to old-growth habitat. They concluded "312 plants, IJ2 stocks of

Status and Functioning 95 TABLE 4-1. Summary of vital statistics of spotted owls from eleven ntensively studied areas west of the Cascades, 1985-1993 Mean Standard error Range Probability of survival to aged Probability of adult survival Number of female young fledged per breeding female Finite rate of increase .925 .015 .258 .844 .311 .036 .005 .125 .000-.418 .821 -.868 .073-.524 .830-~.019 aIncludes second-year birds. Source: Forsman et al. 1996 anadromous sahnonicts, 4 species of resident fish, and 90 terrestrial vertebrates were found to be closely associated with oict-growth forest conditions." On the Olympic peninsula, spotted-owl home ranges are estimated at 1,680-2,800 ha (Lehmkuh] and Raphael 1993), with vertical canopy layering and large snag diameters being best predictors of owl presence in oid-growth forest (Milis et al. 1993~. The large sizes of home ranges suggest that habitat management strategies designed to maintain spotted-owl populations should also maintain habitat for other species dependent upon old-growth forest and the associated ecosystem processes within those forests. However, that owl populations are declining across the Westside range implies that sufficient habitat might not be available currently to perpetuate populations. Strategies for reversing that trend are to protect current oIc3-growth forest used by owls and to identify and take advantage of opportunities to create suitable habitat in seconcI-growth forest. On BUM lands in western Oregon, spotted owl nest sites contained more old-growth forest and were in larger old-growth patches than randomly located sites in managed forests (Meyer et al. 1998~. With the exception of oict-growth patch size, no forest fragmentation indices were related to nest site selection. These forests consisted of approximately 25% old-growth and mature forest, as contrasted with estimates of approximately 60% of the Oregon Coast Range forests being in old growth before IS40 (Ripple 1994~. Habitat occupancy may be relater! to environmental variables such as density of snags and down logs that are habitats for prey.

96 Pacific Northwest Forests The situation is different in spotted-owl habitat in Eastsicle forests. Buchanan et al. (1995) reported that 60 of 83 nest sites examiner! In the Wenatchee, Wash.0 area were in forest in intermediate stages of succession, and the remaining 22 were in old-growth forests. Nesting was found in a wide range of stand ages, from54 to 700 years. For 23% of the nest sites, the area had been partially logged 40 or more years ago. Specific characteristics of nest sites in this area were trees with 35-60 cm Ebb with branches further above the ground than other trees within the area. Because food supplies and foraging efficiency were consicterect to have little influence on nest site selection, canopy height was consiclerect important in increasing the ability of the owl to detect avian predators, such as the great horned owl. Management efforts to minimize fire In this area might have favored creation of suitable habitat in the short term but also might have increased the risk of habitat loss due to severe fire. In addition, fecundity rates of spotted owls on Eastsicle forests appear to be higher than on Westside forests. Over 5 years, Eastside rates averaged 0.49 female fledglings per female (range 0.~-0.74) (Irwin ant] Fleming 1991), compared with .31 female fledglings per female on the Westside (Foreman et al. 1996~. The main conclusions from these comparisons is that the population dynamics and habitat relationships of spotted owls vary across the owes range, that populations are declining west of the Cascade crest, and that Insufficient habitat is now available on Westside forests to sustain owl populations. The marbled murrelet is another threatened bird species of the Pacific Northwest. It is a seabird that nests in late successional forests along the coast and adjacent Coast range (Nelson et al. 1992~. Gill-net fishing, of} pollution, and harvesting of old-growth forests within the range of the murrelet are considered threats. As part of an adaptive management approach, the Pacific Seabird Group (1993) recommended restriction of logging in known nest sites and similar habitats and restrictions in use of gill nets. Bald eagles, listed as endangered in this region, are increasing after losses primarily attributable to pesticide contamination (Montana Bald Eagle Working Group 1991~. However, pesticide residues in the lower Columbia River basin are still high and are affecting eagle productivity (Anthony et al. 1993~. Nesting habitat in this region is primarily mixed- species, multistoried forest with snags and large trees or snags that

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-

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

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

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 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

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

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

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

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 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.

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

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 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

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.

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People are demanding more of the goods, services, and amenities provided by the forests of the Pacific Northwest, but the finiteness of the supply has become clear. This issue involves complex questions of biology, economics, social values, community life, and federal intervention.

Forests of the Pacific Northwest explains that economic and aesthetic benefits can be sustained through new approaches to management, proposes general goals for forest management, and discusses strategies for achieving them. Recommendations address restoration of damaged areas, management for multiple uses, dispute resolution, and federal authority.

The volume explores the market role of Pacific Northwest wood products and looks at the implications if other regions should be expected to make up for reduced timber harvests.

The book also reviews the health of the forested ecosystems of the region, evaluating the effects of past forest use patterns and management practices. It discusses the biological importance, social significance, and management of old-growth as well as late-succession forests.

This volume will be of interest to public officials, policymakers, the forest products industry, environmental advocates, researchers, and concerned residents.

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