2

Forest Health

This chapter contains the committee’s definition of forest health, which includes ecological, economic, and sociocultural factors. It summarizes the threats facing North American forests from insect pests and pathogens and introduces, as examples, the cases of four tree species affected by one or more of these pressures. These case study species are referenced throughout this report. This chapter concludes by describing the effects these threats have on forest health and ecosystem services.

DEFINING FOREST HEALTH

The committee spent much of its early deliberations discussing the term forest health. It heard a number of presentations on the topic (see Meeting 2 in Appendix B) and consulted the scientific literature (e.g., Kolb et al., 1994; Helms, 1998; Raffa et al., 2009; USDA-FS, 2009; Trumbore et al., 2015). On the basis of its information-gathering efforts, the committee agreed on the definition of forest health for this analysis as

Forest structure is the horizontal and vertical distribution of plant material, including ground vegetation and dead or fallen woody material, shrubs, and understory, midstory, and overstory trees (Bennett, 2010). Structure also concerns the age distribution of the trees in the forest. Forest stands are considered even-aged if all of the trees are within the same age class. A forest with uneven-aged structure is a stand with three or more age classes (Bennett, 2010). In practice, size is often used as a proxy for age. Forest structure affects seedling growth, survival, and crown formation of trees as well as the formation of habitat niches (von Gadow et al., 2012).

Forest composition refers to the identity and frequency of plant species found in a stand or landscape, including grass, forbs, shrubs, and trees. In other words, it is the entire plant community of the forest (Moore, 2004). Forest composition, directly or indirectly, affects all other biota present.

Trees play an important role in ecological processes, that is, the cycling of water, nutrients, and energy through the ecosystem, as well as in the natural successional dynamics, that is, the changes in plant species composition and structure following a disturbance (Glitzenstein et al., 1986; Keeton and Franklin, 2005). Trees’ influence on plant species composition and structure affects in turn the other species present in the system.

Healthy forests support economic, ecological, and sociocultural functions. Economic functions relate to the quality and quantity of timber or other vegetation products and game extracted from a forest as well as revenues generated through recreational uses of the forest. Ecological functions include habitat for wildlife, maintenance of biodiversity, soil erosion control, climate regulation, flood control, and effective maintenance of water quality. Sociocultural functions concern aesthetic, spiritual, and cultural values (DeFries et al., 2005; Cooper et al., 2016).

Forest productivity refers to the net primary productivity of plants in the forest system (reflected by the difference between the carbon captured via photosynthesis and that lost via respiration) (Landsberg and Waring, 1997).

Resilience in a forest ecosystem describes its capacity to absorb a disturbance¹ without a significant long-term change to the forest community functions and processes that existed before the disturbance (Holling, 1973; Millar and Stephenson, 2015; Seidl et al., 2016). For this report, resilience is specifically defined as a forest’s ability to maintain its structure, processes, and functions in the long term; however, the committee was mindful of other aspects of resilience in response to disturbance (e.g., resistance, absorption, reorganization, and transformation; Fisichelli et al., 2016). In particular, transformative resilience, that is, the capacity to change into a new system when disturbance makes the existing system untenable (Walker et al., 2004), could be of great relevance in the context of using biotechnology in forest ecosystems.

Like forests themselves, the assessment of whether a forest is healthy is not static. The assessment of the health of a forest will change not only with the evaluation of its structure, composition, processes, function, productivity, and resilience, but also with the state of knowledge about these aspects of forest health. Increasing numbers of studies are also demonstrating that climate change is also altering various aspects of forest health (Boisvenue and Running, 2006; Reyer et al., 2017; Paquette et al., 2018).

THE VALUE OF HEALTHY FORESTS

A healthy forest can be valued for the benefits it provides to humans and also for its own sake. An instrumental view of forest health takes it as a means to an end: the betterment of human welfare. In contrast, the intrinsic value of a forest does not depend on its contribution to human society (NRC, 2005). While the instrumental valuation of the forest ecosystem is framed in terms of the services it provides to humans, intrinsic value concerns the value a forest may have in itself, independent of its usefulness to human beings. Here, both perspectives on valuation are introduced.

Maintaining forest health is essential for the conservation and sustainable management of the many ecosystem services provided to humans by forests. Ecosystem services are the goods and services that are of value to people, provided wholly or in part by ecosystems (Olander et al.,

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¹ Natural disturbance is part of the normal functioning of a forest. Forested systems undergo successional and cyclical changes in structure and composition, which help to maintain high levels of biodiversity (Perry, 1994; Barnes and Wagner, 2004). Healthy forests may withstand natural disturbances either by being able to maintain similar properties (i.e., showing resistance) or by being able to recover many of their original properties afterward (i.e., being resilient). Land management practices can influence forest function and productivity following disturbance (Millar and Stephenson, 2015).

2015). In 2005, the Millennium Ecosystem Assessment categorized these services as provisioning, regulating, supporting, and cultural (Shvidenko et al., 2005; see Box 2-1).

Many ecosystem services that are provisioning, regulating, or supporting are biologically mediated (Burkhard and Maes, 2017). Trees help form and retain soil, cycle nutrients, and store carbon (e.g., Seidl et al., 2016). They filter and regulate the flow of water, first by intercepting rainfall in the canopy. The reduced volume and speed of the rain allows more water to be absorbed into the ground and, combined with the roots’ soil retention properties, controls flooding and reduces erosion (Ellison et al., 2017). Second, roots take up nutrients and pollutants in the subsurface water, preventing these elements from filtering into the groundwater supply. Trees improve air quality by intercepting pollutant particles (Nowak et al., 2014). Water vapor cools the surrounding environment when it evaporates from leaves. Trees buffer the landscape from the heat of the sun and the force of winds, and forests provide food and habitat for pollinators, fish, wildlife, and other organisms, as well as food, fuel, and products for humans.

Cultural ecosystem services are diverse (Milcu et al., 2013). They vary according to the intended or desired use of an ecosystem, such as recreation or creation of traditional forest products. Additionally, forests provide substantial cultural heritage or identity and spiritual, educational, and aesthetic values (Cooper et al., 2016). The values at stake may vary by individual or group. For example, some people may value mountain bike trails through a forest, whereas others may value the same area for its wildlife viewing opportunities or for a spiritual connection felt to nature when in that space. People may also place existence or nonuse value on forests simply because they wish to preserve the ecosystem or species within it (NRC, 2005).

Alongside the services they provide to humans, ecosystems such as forests may also be thought to have intrinsic value, value for their own sake. Intrinsic value, however, can be understood in different ways. Subjective intrinsic value arises from human evaluative attitudes. In the context of forests, for instance, people might intrinsically value forest ecosystems or wild animals or the perceived state of wildness itself. Objective intrinsic value describes value that is believed to exist on the basis of certain properties or features, independent of anyone’s evaluative attitudes (Sandler, 2012, 2018). If someone argues that human lives are valuable on the basis of certain properties

humans have, whether or not anyone actually values human lives, then they are defending the objective intrinsic value of human life. If someone argues that a forest ecosystem is objectively intrinsically valuable, they are maintaining that it has intrinsic value whether or not any human actually values it. Although the existence of objective intrinsic value is disputed on the ground that values must be created by valuers, the existence of objective intrinsic value in species, ecosystems, individual organisms, or all three has often been assumed or defended in conservation and environmental ethics (e.g., Soulé, 1985; Taylor, 1986; Rolston, 1988).

The relationship between intrinsic value and existence value is complex. Because existence value is based on human preference, it is clearly distinct from objective intrinsic value. Existence value and subjective intrinsic value, however, are much closer in meaning, and some definitions take existence value to be synonymous with subjective intrinsic value (e.g., Aldred, 1994). However, Davidson (2013:175) interprets existence value “as the (willingness to pay for the) benefits one derives from something’s mere existence, although one has no current or future plans for its active use.” Existence value, on this account, entails some kind of benefit or satisfaction to the valuer. Intrinsic value, on the other hand, does not imply any benefit to the valuer; rather, the existence of something with intrinsic value “exerts a moral duty on us to take it into account.” Therefore, Davidson suggests something could have intrinsic value without existence value; for example, a rat in a kitchen has intrinsic value, in that the human in the kitchen has a duty not to harm it, but presumably that person would prefer for the rat not to exist at all. Given this understanding of intrinsic value, Davidson argues that intrinsic value, though not existence value, falls outside the scope of ecosystem services because it is not in any sense about nature’s services to humans.

In this report, the committee adopts ecosystem services as the basis for assessment of the instrumental impacts of introducing a biotech tree to counter a threat to forest health. Chapter 5 presents a specific framework for defining ecosystem services in impact assessment that is compatible with regulatory decision making (discussed in Chapter 6). The impact assessment considers the potential benefits, risks, and trade-offs of the introduction of a biotech tree by evaluating expected changes in forest ecosystem services. However, the committee also believes that consideration of the intrinsic values of a healthy forest could usefully broaden the scope of public deliberations about the use of biotechnology (discussed in Chapter 7). Chapter 4 considers some of these values and the ways in which they may be affected by the introduction of a biotech tree to a forest ecosystem.

A healthy forest—that is, one in a condition that sustains the components of an ecosystem over time and space—is more likely to sustain ecosystem services of value to individuals and society. When assessing the impact of a threat (such as an invasive insect) on forest health, evaluating the effect of that threat on the biologically mediated processes and the cultural and aesthetic values of the forest ecosystem provides the basis for assessing how the provision of ecosystem services may change. When adverse effects are experienced or anticipated, alternative means of returning the forest ecosystem to health are considered, including the introduction of a biotech tree that can resist the threat. The remainder of this chapter reviews the scope of the threat from insect pests and pathogens facing North American forests and the implications of that threat for the forest ecosystem and the ecosystem services it provides.

THREATS TO FOREST HEALTH FROM INSECT PESTS AND PATHOGENS

Despite being part of the forest natural disturbance regime, outbreaks of insects and pathogens have dramatically increased in number and impact since the mid-19th century (Aukema et al., 2010; Boyd et al., 2013). The most recent national insect and disease risk assessment, conducted in 2012 by the Forest Service of the U.S. Department of Agriculture (USDA), estimated that 32.9 million

hectares (81.3 million acres)—that is, almost 7 percent of all forested² or treed³ land in the United States—were at risk of losing at least 25 percent of tree vegetation between 2013 and 2027 due to insects and diseases (Krist et al., 2014; see Figure 2-1). That assessment placed 9.4 million more hectares (23.3 million acres) at risk than was estimated in 2006 (Krist et al., 2014).

Most of these outbreaks have been caused by introduced insects and pathogens or by native species within their natural range as well as those expanding their geographic ranges due to climate change (Liebhold et al., 1995; Lovett et al., 2006; Sambaraju et al., 2012; Weed et al., 2013). Climate change is further compounding the impact of insects and pathogens by increasing abiotic stresses on trees, which may result in reduced defenses and increased susceptibility (Breshears et al., 2005; Berg et al., 2006). As a result, the impacts of insects and pathogens are among the greatest threats to forest ecosystems in North America (Moser et al., 2009; Krist et al., 2014; Lovett et al., 2016).

As the frequency of insect and pathogen outbreaks increases, forest resilience and the ecosystem services associated with forests are threatened (Millar and Stephenson, 2015; Seidl et al., 2016). The next section describes general threats posed by insects and pathogens and their interaction with climate change.

Introduced Insect Pests and Pathogens

Since the 1600s, around 450 species of insects and at least 16 species of pathogens have been introduced and become established in continental U.S. forests. Of those, 14 percent of the insects (62 species) and all of the pathogens have been classified as high-impact species (Aukema et al., 2010); that is, they cause some combination of tree mortality, canopy thinning, growth loss, defoliation, and decreased reproduction or regeneration. At least 2.5 introduced, established insect

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² Forested land contains at least 10 percent tree canopy cover.

³ Treed land is an area with measurable tree presence, including urban areas and land in the Great Plains with trees that does not meet the definition of forested land.

species have been detected each year since 1860 (Aukema et al., 2010). Given their cryptic nature and difficulties in early detection, there is little information on the rate of pathogen introduction.

Increases in human mobility and trade are the major pathways of introductions (Pyšek et al., 2010; Brockerhoff et al., 2014; Early et al., 2016). Pathogens and insect defoliators have generally been introduced with live plants (Liebhold et al., 2012). The introduction of insect borers, the most damaging group (see Box 2-2), is usually associated with wood packaging material (Aukema et al., 2010, 2011). The number of introduced borer species (including bark and ambrosia beetles) has dramatically increased since the 1990s, averaging 1.6 new introductions per year, reflecting the increased use of wood packaging materials and the growth in global trade (Haack, 2006; Aukema et al., 2010; see Figure 2-2). These introductions continue despite proactive requirements for treatment of wood pallets and shipping containers (Haack et al., 2014).

Some of these introductions have had devastating consequences in North American forests; impacts have ranged from temporary declines in population productivity to the functional extirpation of an entire species (see case study of the American chestnut, below). In many instances, the introduced insect pests and pathogens lack natural competitors, predators, parasites, or pathogens to regulate their populations (i.e., enemy release; Keane and Crawley, 2002), giving them a temporary fitness advantage that could contribute to their virulence (Hajek et al., 2016). The damage these species cause can be linked to a lack of resistance in the host tree (Herms and McCullough, 2014). Table 2-1 summarizes many of the nonnative pests threatening North American tree species.

The majority of introduced insect pests and pathogens are found in the northeastern United States (Liebhold et al., 2013; see Figure 2-3). This geographic pattern likely reflects the number

TABLE 2-1 18 Nonnative Forest Insects and Pathogens in North America with Current or Potential Future High Impacts

SOURCE: Adapted from Lovett et al., 2016.

of introductions, the historically high propagule pressure, and the impact of anthropogenic disturbance on the ability of the pests to invade in this region (Liebhold et al., 2013). This distribution is also correlated with the diversity of tree species, which is higher in the eastern half of the country (Liebhold et al., 2013). Once established, the average radial rate of spread—5.2 km per year—seems to be similar for all groups of insect pests and pathogens (Liebhold et al., 2013).

Insect Pests and Pathogens Under Climate Change

Climate change is opening new opportunities for colonization by both native and introduced insect species (Harvell et al., 2002; Logan et al., 2003). Forecasted temperatures for the mid-21st century indicate decreases in the length of the cold season and the incidence of extreme cold spells (IPCC, 2013). Cold winter temperatures, cold snaps, and short growing seasons have kept many insect pest species in the United States from moving into higher elevations and more northern latitudes (Carroll et al., 2004; Esper et al., 2007; Dukes et al., 2009). However, with warmer conditions, many insects are colonizing regions that previously had been unsuitable (Williams and Liebhold, 1997; Battisti et al., 2005). In addition, changes in climate are affecting the frequency and magnitude of outbreaks of both native and introduced pests. Outbreaks are predicted to increase in frequency and magnitude in the future. In areas where cold has previously limited establishment, warmer temperatures will likely allow an increase in development and reproductive rates and survival of many insects and pathogens (Ayres and Lombardero, 2000; Bale et al., 2002). An example is the native mountain pine beetle (Dendroctonus ponderosae) outbreak in North America between 1990 and 2010, which killed millions of hectares of pines and has been estimated to be an order of magnitude larger than any previously recorded event (Meddens et al., 2012; Raffa et al., 2013). This outbreak was associated with a reduction in cold snaps (i.e., periods of four consecutive days with average temperature below −20°C (Sambaraju et al., 2012) and overall warmer summer and winter temperatures. Warmer temperatures have also allowed an expansion of the territory of the mountain pine beetle hundreds of kilometers farther north in British Columbia and movement across Alberta into jack pine forests (Pinus banksiana), where it threatens the boreal forest as an invader. Likewise, the southern pine beetle (Dendroctonus frontalis) is moving northward into new forests on the eastern coast of the United States. In Alaska, Canada, and Colorado, outbreaks of the spruce beetle (Dendroctonus rufipennis) have increased with warmer weather and drier summers (Berg et al., 2006), and the beetle’s spread has been predicted to increase as warmer conditions facilitate faster insect development (Bentz et al., 2010; see Figure 2-4).

Changes in temperature and precipitation associated with climate change may become the most influential driver of pathogen outbreaks, because these changes could simultaneously affect host susceptibility and pathogen growth, reproduction, and infection (Sturrock et al., 2012). Forecasts of future climate indicate likely changes in pathogen overwintering survival, changes in host susceptibility to pathogen attack due to other stressors (e.g., drought conditions, ozone, or damage from storms), or changes in life cycles of associated species such as insects that disperse pathogens (Dukes et al., 2009; Weed et al., 2013). However, the outcome of these changes—higher or lower virulence—will likely be site specific (Sturrock et al., 2012). For example, Phytophthora ramorum, an introduced oomycete that causes sudden oak death, may experience a decrease in favorable environmental conditions in the eastern United States, but an increase in favorable sites in the western United States (Venette and Cohen, 2006; Venette, 2009) and Europe in response to climate change (Bergot et al., 2004).

Given that some pathogen species rely on insects for their dispersal (Wingfield et al., 2016), effects of climate change on the insect populations would likely cause changes in pathogen dynamics. For example, the two fungi that cause beech bark disease (Neonectria farinata and N. ditissima) are spread by a scale insect, Cryptococcus fagisuga. The extent of the infestation had

been restricted by cold winter temperatures, but with the onset of mild winters and dry autumns associated with climate change, both the scale and the fungi will likely move to northern latitudes and affect beech trees that had previously been shielded from the pathogen (Houston and Valentine, 1988; Stephanson and Coe, 2017).

EFFECTS OF INSECT PESTS AND PATHOGENS ON TREES AND ECOSYSTEM SERVICES

Adverse effects on forest health caused by increases in the frequency and magnitude of insect and pathogen outbreaks are already being observed and are likely to continue. This section reviews the effects on some specific tree species and genera; the feasibility of using biotechnology to address threats to these species is discussed in subsequent chapters. This section also examines more broadly the effects of insect pests and pathogens on forest health and ecosystem services.

Case Study Trees

A variety of introduced insect pests and pathogens (many included in Table 2-1) and the exacerbated pressure of some native insects and diseases facilitated by climate change threaten the long-term survival of many forest tree species native to North America. Rather than elucidating all threats, the committee decided to focus on four cases chosen by consensus and based on the following criteria:

• The severity of the threat.
• The causative agent(s) (insect, pathogen, or complex systems involving insect vectors or obligate pathogens with alternate hosts).
• The origin of the insect or pathogen (native or nonnative).
• The impact of climate instability and fire on the severity and extent of the disease or infestation.
• The ecological, economic, and cultural values of the host tree species.
• The use or potential use of the host tree species for plantation forestry.
• The efficacy or feasibility of traditional strategies to protect forest health (biological control, pesticide use, containment strategies, and selective tree breeding).
• The efficacy of gene insertion or gene-editing strategies if already in place.
• The feasibility of gene insertion or gene-editing strategies if not yet attempted or tested.
• Geographical distribution and phylogenetic position of the host species.

The four selected case studies—American chestnut (Castanea dentata), whitebark pine (Pinus albicaulis), ash (Fraxinus spp.), and poplar (Populus spp.)—represent a wide range of forest health problems with different combinations of characteristics in terms of the above criteria (see Table 2-2). In two cases, the committee chose specific host trees that face more than one pest pressure (American chestnut and whitebark pine). In the other two cases (ash and poplar), the committee examined the implications of a specific pest for a genus of trees. The native ranges of the major host tree species vary considerably in extent but together cover much of the United States (see Figure 2-5). Forest ecosystems, rural and urban, have all experienced negative ecological and economic impacts from tree mortality caused by the insects and pathogens examined in these studies. All of the species have clear ecological and cultural value, and all but whitebark pine have economic value. Critical for this study, the species vary in development and feasibility of a biotech solution to reduce vulnerability to the insect pest or pathogen involved. The case studies are introduced here and referenced throughout the rest of the report.

American Chestnut (Castanea dentata)

In the 19th century, the range of American chestnut extended from Maine to Mississippi along the Appalachian Mountains (Little, 1977; see Figure 2-5). American chestnuts were fast growing, and trees could reach 37 meters in height and 5 meters in diameter on favorable sites (Buttrick, 1925; Wang et al., 2013). The number of mature trees prior to the introduction of chestnut blight was estimated to be 4 billion (Detwiler, 1915), representing a major fraction of the forest biomass in many eastern forests (Braun, 1950). At some locations in the Appalachian Mountains, the American chestnut was considered to be a foundation species because of its strong influence on ecosystem structure and function (Youngs, 2000; Ellison et al., 2005a). In some regions, one in four trees in the canopy was reported to be an American chestnut (Johnson, 2013).

In 1904, American chestnuts at the Bronx Zoo in New York City died from infection by a fungal pathogen initially identified as Diaporthe parasitica but later renamed Cryphonectria parasitica. The pathogen was likely introduced on Japanese chestnuts imported to the United States as early as 1876 (Anagnostakis, 1987; Anagnostakis and Hillman, 1992).

The disease spread more or less unchecked, extending over the entire range of the American chestnut by the 1950s (see Figure 2-6). Traditional control measures, such as chemical treatments or clearing and burning, were ineffective (Stoddard and Moss, 1913). The pathogen maintained virulence over time, and almost all mature chestnuts were killed (Hepting, 1974; Russell, 1987).

The pathogen causing chestnut blight is necrotrophic, entering through small wounds in the outer bark, killing the living vascular cambium, and then developing cankers on the dead tissues. In susceptible trees, the fungus eventually girdles the branches and main stem, blocking the transfer of nutrients and resulting in tree death (Anagnostakis, 2000). In blight-tolerant Asian chestnut trees, lignified callus may surround the wound and restrict the growth of cankers; in susceptible trees, the fungus is able to overcome this resistance, leading to mortality.

In 2018, surviving chestnut trees existed mainly in shrubby growth forms that result from the formation of sprouts from the root collar. The sprouts grow for several years until they are again infected by C. parasitica and die back. Each cycle—resprout followed by fungus infection and dieback—weakens the tree until it eventually dies (Griffin, 2000). Sprouts rarely reach reproductive maturity and seeds are seldom produced (Paillet, 2002). Thus, the American chestnut persists mainly as a multistemmed shrub with only a few large chestnut trees remaining, often at the periphery of the tree’s range, presumably as “escapes” (i.e., trees that have not yet been exposed to the pathogen).

The loss of the American chestnut was devastating for rural communities that depended on the tree for food, livestock feed, and timber (Youngs, 2000; Freinkel, 2009). Equally devastating were the changes to the forest ecosystem due to the loss of a foundational species (Freinkel, 2009).

TABLE 2-2 List of Variables Considered by the Committee When Selecting Case Studies

^aSee discussion in section “Social and Ethical Considerations” in Chapter 4.

^bSee Boxes 3-1 and 3-2 in Chapter 3.

^cSee Box 3-4 in Chapter 3.

Other nonnative Castanea species have been planted in urban environments or as orchard trees for commercial production of chestnuts, but they do not fill the same ecological niche as the American chestnut. Chinese chestnut (C. mollissima) and Japanese chestnut (C. crenata) are typically small trees, lacking the fast growth and tall form of American chestnut. The European chestnut (C. sativa) has a growth and form somewhat similar to American chestnut as compared to the Asian species, but the European chestnut trees growing in North America are susceptible to the same diseases as the American chestnut and are not as frost tolerant. The Asian species usually do not live as long as American chestnut. In a forest setting, the other Castanea species are not competitive; they do not grow tall enough or fast enough to compete for light against the native American chestnut or other native tree species (Wu and Raven, 1999; Fei et al., 2012). The American chestnut has lost the role it once had as a foundational species that influenced other species and ecosystem processes.

As with many trees, the American chestnut faces more than one threat. In southern Appalachia, the introduced oomycete Phytophthora cinnamomi causes black lesions on the roots, eventually killing the tree by killing the root system (Crandall et al., 1945). Trials of restoration plantings in this region reveal that P. cinnamomi persists in the soil long after the mature chestnuts die and kills the majority of planted chestnut seedlings within a few months (Rhoades et al., 2003). Asian chestnut gall wasp (Dryocosmus kuriphilus), accidentally imported on Asian chestnut cuttings in 1974 (Payne et al., 1976), attacks both Asian and American chestnuts. The galls suppress shoot growth and nut development.

American chestnut is the committee’s only case study of a species that has essentially been lost throughout its native range as of 2018. Oaks and maples have filled in for this species over much of the range and maintained some of the forest functions (Keever, 1953; Woods and Shanks, 1959; McCormick and Platt, 1980). Although acorns have replaced chestnuts as mast sources to some extent, oaks have episodic mast years, unlike the consistent, substantial annual mast produced by the American chestnut and chestnut’s relatives, the chinquapins (Castenea pumila and C. ozarkensis). Population dynamics of species dependent on the nuts were likely affected, with cascading food web impacts. At least five moth species obligate on chestnuts have gone extinct (Opler, 1978; Wagner and Van Driesche, 2010). Economies and cultures of human communities originally reliant on American chestnut products were also altered (Davis, 2006); chestnut has been identified as a cultural keystone species (sensu Garibaldi and Turner, 2004).

Whitebark Pine (Pinus albicaulis)

Whitebark pine is a high-elevation tree of the western United States and Canada (see Figure 2-7). It spans over 18^o latitude and 21^o longitude, but within that area it establishes only within a narrow elevational distribution extending from the subalpine to treeline (Tomback et al., 2016). The tree exhibits high phenotypic plasticity (i.e., an ability to grow in different forms in response to its environment). In open stands, it grows as a large wide-crowned tree, whereas in dense stands it takes a linear form similar to lodgepole pine. On harsh windswept ridges, it forms krummholz—dwarfed, gnarled trees that seldom reach more than 1–2 meters in height, even when hundreds of years old. In the subalpine, it sometimes grows in mixed stands, often with subalpine fir, Engelmann spruce, and lodgepole pine. In the upper extent of the subalpine and at treeline, whitebark pine is typically the only tree present (Tomback et al., 2016). It is a long-lived tree, sometimes reaching ages of 1,000 years or more (Perkins and Swetnam, 1996). It grows slowly and typically does not begin to reproduce until at least 20–30 years of age and not fully until 60 or more years (McCaughey and Tomback, 2001).

Whitebark pine is considered to be both a keystone and a foundational species. As a keystone, its presence sustains the biodiversity and function of the community of which it is part. As a foundational species, it is responsible for creating the conditions that allow the community to assemble

in the first place (Tomback et al., 2016). At the upper limits of its elevational range, whitebark pine establishes in areas too harsh to support other tree species (Weaver and Dale, 1974; Tomback and Linhart, 1990). In these places, whitebark pines provide shelter and contribute to soil development, allowing other plant species to establish (Arno and Hoff, 1990; Callaway, 1998). “Life islands” of shrubby vegetation often develop at the base of these trees, providing food and nesting habitat for birds and small mammals and stabilizing rocky slopes. Cover provided by the trees regulates snowmelt, retaining water in the subalpine for longer into the spring and supporting flows in mid and low elevations for an extended period into the summer (Farnes, 1990).

The tree is threatened by several factors including human-induced changes in fire regimes (suppression), an introduced fungal pathogen (Cronartium ribicola, the causal agent of a disease called white pine blister rust), a native bark beetle (the mountain pine beetle, Dendroctonus ponderosae), and climate change (increased drought). Individually, each threat is serious. These factors also interact, exacerbating the rate and degree of decline. Together, these threats pose an extremely complex problem for the conservation and restoration of this tree.

More than half of all whitebark pines in the northern United States and Canada are already dead. In some areas, only about 2 percent of mature (reproductive) trees remain (Kendall and Keane, 2001; Zeglen, 2002; Smith et al., 2008). Seeds are dispersed by birds in the jay family, specifically Clark’s nutcrackers (Nucifraga columbiana), that open the cones and cache the seeds for later use. Seeds in unretrieved caches germinate to produce new whitebark pines. In areas where few mature trees remain, foraging becomes inefficient and the nutcrackers reduce visitation to these sites, thus lowering the potential for regeneration (McKinney and Tomback, 2007; McKinney et al., 2009; Barringer et al., 2012).

Mortality has been most severe in the central and northern Rocky Mountains and in the coastal mountain ranges, whereas southern populations remain fairly robust primarily due to a lack of rust and beetle activity as of 2018. Canada listed whitebark pine as endangered in 2010 (COSEWIC, 2010). The tree’s status in the United States is “recommended for listing, but precluded” (USFWS, 2011). Preclusion, in this case, is based on a lack of funding and its lower priority for recovery relative to several other species. As of 2018, the tree’s status under the Endangered Species Act was under re-review, with a decision slated for 2019.

North American Ash (Fraxinus spp.)

There are 16 ash species native to North America, of which green ash (Fraxinus pennsylvanica) and white ash (F. americana) are the most widely distributed. The native range of green ash includes the Eastern Temperate, Great Plains, and Northern Forests ecoregions in North America (Omernik, 1995, 2004; CEC, 1997; Omernik and Griffith, 2014; see Figure 2-8). Although green ash grows abundantly in riparian zones in mesic temperate forests, it can persist in upland forests and seasonally dry urban environments throughout the eastern and central United States. In the Great Plains ecoregion in the western part of the range, green ash can be locally abundant in riparian zones or along ephemeral streams (Rumble and Gobeille, 1998; Lesica, 2009). Although this species occupies only 1–4 percent of the landscape in this region, green ash woodlands support a disproportionately large component of biological diversity, including migratory songbirds, gallinaceous birds, and native ungulates (Boldt et al., 1979; MacCracken and Uresk, 1984; Hodorff and Sieg, 1986; Rumble and Gobeille, 1998). Additionally, 43 native arthropod species are solely dependent on green and white ash during some part of their life cycle, and 30 additional species have only 2–3 known host plants, one of which is ash (Gandhi and Herms, 2010b).

First detected in Detroit, Michigan, and Windsor, Ontario, in 2002, the emerald ash borer (EAB, Agrilus planipennis Fairmaire [Coleoptera: Buprestidae]) poses an acute threat to all of the native ash species in North America (Herms and McCullough, 2014). The International Union for

Conservation of Nature Red List of Threatened Species lists five North American ash species—green ash, white ash, black ash (F. nigra), pumpkin ash (F. profunda), and blue ash (F. quadrangulata)—as critically endangered due to nearly 100 percent mortality following attack, limited ability to regenerate under repeated attack, and rapid spread of the insect, largely through unintentional human agency. EAB, native to Asia, had spread to 31 states and 3 Canadian provinces as of May 2018 (see Figure 2-8).

The insect kills 99–100 percent of green ash trees in forest stands within 7 years of first detection (see Figure 2-9a) and kills urban green ash plantings as fast or faster, due to the extensive use of grafted green ash cultivars (Rebek et al., 2008; Smitley et al., 2008; Knight et al., 2012). Females oviposit in bark cracks and crevices, laying 60–80 eggs. Larvae hatch in a few weeks, feed voraciously on the phloem and other living tissues under the bark and complete four instars before overwintering as prepupae (Cappaert et al., 2005). Pupation occurs in the spring, and adults emerge starting in mid-May and continuing throughout the summer (Poland et al., 2011). EAB feeding destroys the vasculature and the tissue that forms new vessels and bark, ultimately girdling the main stem and thus killing the host (see Figure 2-9b).

Green ash, as well as the other ash species listed as critically endangered, has some capacity to regenerate from root and stump sprouts even after EAB infestation (Kashian, 2016). However, EAB also kills these resprouts, removing any mechanism for regeneration via vegetative propagation. Ash seedlings may be initially abundant after extensive mortality among adult trees (Kashian and

Witter, 2011), giving the impression that ash will recover. However, when these seedlings reach 2–3 cm in stem diameter, EAB infestation again inflicts high mortality. Ash does not have a persistent seedbank, so once mature trees are killed, it is nearly impossible for the species to reestablish itself.

The near synchronous loss of green ash has had a cascade of negative impacts, including the rapid loss of naturally occurring riparian forests, which are composed mainly of green or black ash (Gandhi and Herms, 2010a,b; Hausman et al., 2010; Kovacs et al., 2010; Knight et al., 2013), billions of dollars in tree removal cost to local governments, and the loss of a valuable utility hardwood used for cabinets, furniture, tool handles, restoration of antique cars, wooden snowshoes, guitars, and baseball bats. Five or more hawk moth species that specialize on Fraxinus are hypothesized to be at risk from the loss of ash to EAB (Wagner and Van Driesche, 2010). Thus, without effective and timely intervention, the EAB invasion threatens two of the most widely distributed hardwood species in the riparian forests of eastern North America and the most extensively used group of tree species for soil conservation, rural water management, urban green spaces, and utility woodworking as well as the species that depend on Fraxinus. It also threatens to continue its spread west, where it will likely kill western species of ash that have so far been unaffected.

Poplar (Populus spp.)

This case study presents an example of an incipient invasion of a pathogen native to forest ecosystems in eastern North America that poses a threat to an ecologically important native tree group in western North America as well as to a sector of the forest products industry. There are eight native species of Populus in North America and multiple hybrids (Cooke and Rood, 2007), but the focus of the case study is on three species: black cottonwood (P. trichocarpa), the closely related balsam poplar (P. balsamifera), and widespread eastern cottonwood (P. deltoides) (see Figure 2-5). These species are model organisms for basic research, so in some ways this tree species may represent a best-case scenario for the potential of biotechnology to prevent or mitigate a forest health crisis.

In open environments, black cottonwood is a dominant native tree in lowland riparian ecosystems in Oregon, Washington, and British Columbia (Franklin and Dyrness, 1973), where it plays essential roles in stream ecology (Pastor et al., 2014) and as habitat for birds and mammals (Kauffman and Krueger, 1984; Isaacs et al., 1993, 1996; Bryce et al., 2002). Black cottonwood populations typically become established following deposition of sand and gravel following episodic floods, resulting in bands of even-aged cohorts that line river floodplains (Braatne et al., 1996). The species produces abundant seeds with cotton-like appendages that facilitate long-distance dispersal by wind and water (Slavov et al., 2010; DiFazio et al., 2012) and enable deposition on newly created substrates following recession of floodwaters. It also spreads vegetatively by root sprouts or abscised branches, leading to the development of large clonal stands in some locations (Gom and Rood, 1999; Slavov et al., 2010). As a result, this species is critical for floodplain soil stabilization and provides habitat for other species. Black cottonwood populations have shown evidence of decline in recent decades, in part because of a loss of establishment opportunities due to flood control (Dykaar and Wigington, 2000; Braatne et al., 2007). However, extensive gallery forests of this species are still a prominent and valued component of the landscape in the Pacific Northwest.

In research, the genus Populus is widely recognized as a model for woody tree biology (Taylor, 2002; Jansson and Douglas, 2007). The genus has several desirable experimental characteristics, including a small genome (Tuskan et al., 2006), easy vegetative propagation via stem cuttings and tissue culture, ability to hybridize (Induri et al., 2012), and short generation time (Stanton et al., 2010). These features have made Populus an attractive model for applied studies focused on enhancing productivity in intensive plantation settings for pulp, biofuel, and solid wood (Dickmann, and Kuzovkina, 2014). Populus spp. have also been a primary target of basic research in the areas of physiology, ecology, and evolutionary biology. Consequently, abundant genetic and genomic

resources are available for this genus (Tuskan et al., 2006; Evans et al., 2014; Zinkgraf et al., 2016; Fahrenkrog et al., 2017).

The fungal pathogen Sphaerulina musiva (synonym, Septoria musiva) is native to eastern North America, with a historical distribution that largely mirrors that of its primary natural host, eastern cottonwood. The pathogen causes blotches and stem cankers in P. deltoides, P. balsamifera, P. trichocarpa, and hybrid Populus cultivars in North America (see Figure 2-10). The disease initially occurred primarily in natural populations of P. deltoides in the east, where it was mostly manifested as leaf spots (Waterman, 1954). However, it has since spread from eastern forests to intensively cultivated eastern plantations of native and hybrid poplars, where it commonly causes stem and branch cankers, often leading to breakage of the main stem and death of the tree (Ostry and McNabb, 1985; Dunnell et al., 2016). In the most detailed published survey of a large-scale outbreak, Strobl and Fraser (1989) documented occurrence of S. musiva canker in intensively cultivated hybrid poplar in Ontario. Within 5 years of the establishment of susceptible hybrid clones in the region, more than 150 hectares (370 acres) of plantations were affected by the disease, and 79 percent of the area planted with susceptible clones had disease outbreaks (Strobl and Fraser, 1989). This disease can clearly have rapid and devastating impacts on intensive plantations of susceptible varieties (Feau et al., 2010).

Of even greater concern are reports of stem cankers caused by S. musiva in natural populations of black cottonwood in Pacific Northwest forests, where the disease is not native and was unknown until 2006 (Callan et al., 2007; Herath et al., 2016). Both P. trichocarpa and P. balsamifera show high susceptibility to this disease (LeBoldus et al., 2013; Herath et al., 2016), so the threat of a large-scale outbreak has caused substantial concern among scientists, members of the forest industry, land managers, and the public (Feau et al., 2010). Black cottonwood may be particularly

vulnerable to an outbreak of this disease. In the core of its range along rivers of northwestern North America, black cottonwood often occurs in dense, even-aged stands in climates and microsites that are characterized by abundant moisture (DiFazio et al., 2011), which could facilitate spread of the disease. Furthermore, P. trichocarpa populations are already in decline due to flood control and habitat loss (Rood and Mahoney, 1990; Dykaar and Wigington, 2000), so a disease outbreak could be particularly problematic for the long-term viability of the species.

Effects on Forest Health and Ecosystem Services

The case studies are not isolated examples of species in decline. Rather, given the rate of introductions of nonnative insect pests and pathogens and the effects of climate change on distribution and abundance of native insects and pathogens, their trajectory is likely to become the norm in North American forests. The frequency and magnitude of outbreaks and the rate of tree mortality are likely to increase. These impacts will have significant effects on forest health and ecosystem services (Dukes et al., 2009; Millar and Stephenson, 2015; Lovett et al., 2016; Liebhold et al., 2017). As outlined above, ecosystem services are generally defined as the direct and indirect contributions of ecosystems to human well-being (Braat and de Groot, 2012; see also the discussion in Chapter 5).

The most immediate effects of increased insect and pathogen activity (native and introduced) on forest health are reductions in productivity and alterations of nutrient, carbon, and water cycles (Lovett et al., 2006). In the case of extended or severe tree mortality, as in the American chestnut, substantial losses of other forest species and some ecosystem services can be expected.

The impact of increased insect pest and pathogen activity on ecosystem services is strongly linked to the proportion of the canopy affected. Increases in the effects of host-specific insects and pathogens that target dominant and keystone tree species will likely result in the most severe and long-term impacts (Ellison et al., 2005a). For example, eastern hemlock (Tsuga canadensis) dominates forest stands in its northern range and moist coves in the south. Loss of the hemlock due to the nonnative hemlock wooly adelgid (Adelges tsugae) has caused the loss of several wildlife species associated with hemlock (Tingley et al., 2002; Ellison et al., 2005b), affected soil processes (Jenkins et al., 1999), and changed local hydraulic flow (Ellison et al., 2005a). These impacts may occur even where other tree species rapidly colonize areas once occupied by hemlock (Orwig et al., 2002), as the ecosystem services provided by one species may differ from those provided by others. For example, in the Southern Appalachians, the effects of hemlock trees on stream flow and temperature sustain unique communities of salamanders, fish, and other stream invertebrate species that will be lost without hemlocks (Snyder et al., 2002).

In areas of low tree diversity, outbreaks of insect pests and pathogens can have devastating consequences for regulating and supporting services, as a large proportion of the canopy can be affected with no replacement species naturally recolonizing afterward. This is the case with whitebark pine. The ecological void created by the loss of whitebark pine (see case study above) will be vast because this species supplies numerous resources, including shelter and food to wildlife species, water regulation through snowpack retention, and soil development, which facilitates the establishment of other plant species (Arno and Hoff, 1990; Farnes, 1990; Callaway, 1998).

Intrinsic properties of the ecosystem may mediate the magnitude of the loss of ecosystem services. High-diversity forests are home to more introduced insect pests and pathogens (Liebhold et al., 2013), but the loss of one tree species in these areas may be compensated by other species. For example, even though white ash (Fraxinus americana) is a conspicuous species in eastern North American forests, it does not dominate these stands (Prasad et al., 2007-ongoing). As of 2018, EAB was causing the death of most adult white ash trees across large areas. However, the void left by the death of ash trees is rapidly being filled by other tree species, such as maples (Margulies et al., 2017). Maples likely supply some of the ecosystem services provided by ash but may not support

the biodiversity reflective of an uninvaded forest. The same was true for the eastern forest when the American chestnut declined; the replacement species do not produce the mast, timber, or stature and do not provide the cultural or spiritual values of the original forest (Davis, 2006). Additionally, while replacement species offer at least a temporary mitigation of some impacts, the continual influx of nonnative insects and pathogens could subject the replacement species themselves to impacts in the future, a factor to consider when deciding whether to try to restore species in jeopardy of extirpation.

However, even if impacts can be mostly mitigated by replacement tree species, the costs can still be substantial. Shortly after EAB was found in the United States, the U.S. Forest Service projected the lost timber value from ash trees in forested lands could be close to $280 billion (Nowak et al., 2003). Additionally, the anticipated cost of losing these species in urban settings was estimated to be between $20 billion and $60 billion (USDA-APHIS, 2003) due to loss of property value and cost of removal. Using this subset of ecosystem values, EAB is the most economically devastating invasive insect pest in North American history (Herms and McCullough, 2014).

The effects of insect pests and pathogens on individual trees have cascading impacts on populations, reducing reproduction and survival. In the most extreme cases, local extirpation of the tree species and extinction or extirpation of species dependent on the tree may result (e.g., the already mentioned extinction of five moth species with the loss of the American chestnut) (Opler, 1978; Wagner and Van Driesche, 2010). Such species-specific effects can then translate into changes in community assemblage and structure, and thus, ecosystem functionality. The loss of whitebark pine may reduce the complexity and function of high-elevation ecosystems in the west and contribute to the decline of grizzlies and other wildlife as well as ecosystem services related to water and sediment regulation. The loss of ash trees affects not only natural communities; loss of city trees has had a large effect on property values (Aukema et al., 2011). The decline of black cottonwood in the West would adversely affect riparian habitats.

CONCLUSIONS

Based on its evaluation of the scientific literature and the information it gathered from invited speakers, the committee defined forest health as a condition that sustains the structure, composition, processes, function, productivity, and resilience of forest ecosystems over time and space. An assessment of this condition is based on the current state of knowledge and can be influenced by human needs, cultural values, and land management objectives. North American forests are struggling to maintain healthy conditions because of increasing stresses, on to which outbreaks of introduced insects and pathogens and the geographic expansion of native pests due to climate change are layered. While impossible to fully isolate, the direct adverse effects of pests on forest health have significant impacts on the ecosystem services that forests provide.

Conclusion: Healthy forests provide valuable ecosystem services to humans.

The ecological processes performed by forests and the cultural and aesthetic values attached to forests are important to individuals and to society. Forests provide food and habitat for pollinators, fish, wildlife, and other organisms, as well as food, fuel, and products for humans.

Conclusion: The health of North American forests is threatened by the introduction and spread of nonnative insects and pathogens and the epidemics of native pests exacerbated by environmental stress due to climate change.

At least 62 insect species and 16 pathogens that cause tree mortality, canopy thinning, growth loss, defoliation, or decreased reproduction or regeneration have been introduced to North Amer-

ica. Some of these introductions have had devastating consequences in North American forests. Increases in human mobility and trade are likely to lead to more such introductions. Climate change is opening new opportunities for colonization by both native and introduced insect species and affecting the frequency and magnitude of outbreaks of both native and introduced pests. Outbreaks are predicted to increase in frequency and magnitude in the future.

Conclusion: Tree species in forest ecosystems, tree plantations, and urban landscapes across North America are threatened by insect pests and pathogens.

The four case study species selected by the committee—American chestnut (Castanea dentata), whitebark pine (Pinus albicaulis), ash (Fraxinus spp.), and poplar (Populus spp.)—serve as examples of diverse ecosystems and habitats that are experiencing adverse impacts from tree mortality caused by insect pests and pathogens. The American chestnut was a foundation species because of its strong influence on ecosystem structure and function and an economic resource for communities before its extirpation. Whitebark pine creates and sustains community biodiversity at high elevations. Ash woodlands support biodiversity and provide benefits to humans as a popular urban landscape tree. Black cottonwood stabilizes streambanks and provides habitat for birds and mammals; poplars are also model trees for research and an important resource for production of pulp, biofuel, and solid wood.

Conclusion: Many forest tree species are threatened by more than one insect pest or pathogen.

American chestnut, whitebark pine, ash, and poplar are just four examples of North American tree species that have been or are in danger of being extirpated. They are subject to one or more pest threats, and whitebark pine, in particular, is losing habitat to climate change. The number of (see Table 2-1) and trend in (see Figure 2-2) introduced threats and the geographic expanse of all pest threats represented by the four case study species (see Figure 2-5) suggest that native trees throughout North America are in danger of or may become subject to pest outbreaks that adversely affect forest health.

Conclusion: As the frequency of insect and pathogen outbreaks increases, many forest tree species are in jeopardy of being lost from the landscape, resulting in changes to ecosystem services.

The growth in global trade, the increase in human mobility, and the warming of the climate are all contributing to the increased pest pressure that forests now face. The magnitude of pest outbreaks may permanently change the structure, composition, processes, function, productivity, and resilience of forest ecosystems. As tree species are lost from the landscape, the species obligate to those trees will be lost as well.

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