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Impacts and Management of Hemlock Woolly Adelgid in
National Parks of the Eastern United States
Scott R. Abella*
Abstract - Introduced forest pests and pathogens are a major threat to national parks. This
paper reviews existing impacts, projected impacts, and management options for Tsuga canadensis
(Eastern Hemlock) forests in US national parks threatened by the introduced insect
Adelges tsugae (Hemlock Woolly Adelgid [HWA]). Eighty-five national parks, or 21% of all
parks in the US national park system, are encompassed within the range of Eastern Hemlock.
These 85 parks include iconic areas such as national battlefields and the Appalachian National
Scenic Trail. Four focal parks of this study—Delaware Water Gap National Recreation Area,
Great Smoky Mountains National Park, Shenandoah National Park, and New River Gorge
National River—are collectively visited by 16 million people annually and Eastern Hemlock
is a forest component in 2–26% of their area. Results of research on HWA impacts to forest
species composition, soil nutrient cycling, watersheds and fisheries, wildlife habitat, and visitor
experiences and safety have already been reported from these parks. A general principle
is that after Eastern Hemlock forest decline, some species (e.g., some avian species favoring
other tree species) benefit, while those associated with Eastern Hemlock must adapt or
decline. Forecasting future forest-tree composition is complicated by the fact that: (i) many
possible replacement tree species are themselves threatened by introduced damaging agents,
(ii) changes hinge upon understory dynamics such as invading exotic plants or expansion of
native shrubs, and (iii) this die-off event is occurring within a context of multiple interacting
factors such as elevated herbivory, climate change, and atmospheric pollution. Some
management strategies for parks include: intensive HWA chemical treatment at priority sites,
biocontrol, genetic manipulation for HWA resistance in Eastern Hemlock, exotic plant treatment,
facilitated establishment of native vegetation, or doing nothing, the last of which also is
likely to result in appreciable forest change. Threats to US national parks posed by introduced
forest pests and pathogens warrant heightened attention.
Introduction
The potential for introduced forest pests and pathogens to degrade resources of
national parks is formidable. The 1916 Organic Act (39 Stat. F35) created the National
Park Service (NPS) with the long-term mission of conserving, unimpaired,
significant US natural and cultural features for future generations. This mission
was reaffirmed by the 2006 NPS management policy that defined impairment as a
direct impact to a resource or value that is: (i) necessary to fulfilling specific purposes
identified in a park’s establishing legislation, (ii) key to a park’s natural or
cultural integrity, or (iii) identified in a park’s management plan as being of critical
*National Park Service, Washington Office, Natural Resource Stewardship and Science Directorate,
Biological Resource Management Division, 1201 Oakridge Drive, Fort Collins,
CO 80525. Current address - Natural Resource Conservation LLC, 1400 Colorado Street,
Boulder City, NV 89005; abellaNRC@gmail.com.
Manuscript Editor: Justin Hart
Forest Impacts and Ecosystem Effects of the Hemlock Woolly Adelgid in the Eastern US
2014 Southeastern Naturalist 13(Special Issue 6):16–45
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significance (NPS 2006). Forest die-offs caused by introduced pests and pathogens
could meet criteria for impairment because indigenous forests represent inherent
resource values, forests influence many other natural and cultural resource values,
and multiple human generations are required for old forests to develop.
Reducing or forestalling degradation from introduced forest-damaging agents
is a major challenge for the NPS and one that is expected to increase in difficulty
(Jenkins 2007). Numerous examples of major alterations to forest species composition
caused by introduced organisms already exist (Liebhold et al. 1995). A
classic example in eastern North America is Castanea dentata (Marshall) Borkh.
(American Chestnut). This native tree was eliminated as a dominant canopy species
by the 1950s, including in national parks, because of the introduced fungus
Cryphonectria parasitica (Murrill) Barr that caused Chestnut Blight (Ellison et al.
2005). Thus, this damaging agent had already dramatically impacted parks long before
the past few decades of recent climatic change (Lovett et al. 2006). In addition
to introduced fungi, over 450 species of exotic forest insects have been introduced
to North America (Aukema et al. 2009). Many of these introduced organisms are
still expanding their ranges, further underscoring their potential to impact national
parks. The introduced insect Adelges tsugae Annand (Hemlock Woolly Adelgid
[HWA]) that is killing Tsuga canadensis (L.) Carrière (Eastern Hemlock) trees in
eastern North America exemplifies the type of recent impact already incurred and
suggests the potential for future impacts to parks (Eschtruth et al. 2006). Moreover,
responses to this introduced pest illustrate issues in contemporary management that
seek to reduce impacts or facilitate forest adaptation to damaging agents in parks
(Johnson et al. 2008).
This review has three focus areas. First, it summarizes Eastern Hemlock ecology,
HWA ecology, and the value of Eastern Hemlock forests to parks threatened
by HWA. Second, it synthesizes data regarding existing and potential impacts of
HWA to four large HWA-infested parks where research has been conducted, and
then broadens discussion of potential impacts by including lands surrounding parks.
Third, it discusses potential management strategies for Eastern Hemlock forests in
a national park context and presents ideas for integration of forest health with NPS
programs. This emphasis on NPS lands for the HWA threat is timely, because while
a large body of literature is accumulating on HWA, national parks have some unique
considerations that differ from other lands. Moreover, several aspects of HWA management,
such as release of biocontrols or long-term genetic manipulation for HWA
resistance in Eastern Hemlock, illuminate challenges faced by managers who must
comply with existing NPS policy.
Ecology and Values of Eastern Hemlock Forests
Eastern Hemlock autecology
Eastern Hemlock is an evergreen, coniferous tree in the Pinaceae family. Godman
and Lancaster (1990) reported that ages over 400 years, diameters greater than 100
cm, and heights of 30 m or more are common for Hemlock trees in mature stands.
Eastern Hemlock is ranked as the most shade-tolerant of eastern tree species and it
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can survive with only 5% full sunlight (Goerlich and Nyland 2000). The slow-growing
species can remain as a suppressed seedling or sapling for centuries, but given a
canopy opening at any point after seedling establishment, it can grow ≈0.6 cm/yr in
diameter (Godman and Lancaster 1990). Eastern Hemlock relies on a seedling bank
for recruitment, because while seed production is plentiful, seeds stored in the soil
rarely retain viability more than one year (Hille Ris Lambers et al. 2005). Cones and
the seeds inside them are small, and the seeds are wind dispersed starting in October
and continuing through winter, depending on geographic location (Goerlich and
Nyland 2000). The species does not resprout if cut (Orwig and Foster 1998). Eastern
Hemlock is susceptible to several native and exotic fungi (e.g., root rots) and insects
that can damage or kill trees, but none are presently as devastating as HWA. Hemlock’s
unique combination of large size, longevity, shade tolerance, and evergreen
growth form is not readily replaced by other eastern tree species.
Eastern Hemlock has high moisture requirements, is susceptible to drought, and
might have reduced resistance to damaging agents on dry sites (Orwig et al. 2002).
In the southern part of its range, Eastern Hemlock typically is most dominant on
moist sites near streams, in drainages, or on topographically protected sites or north
aspects with damp, cool microclimates (Abella 2003, Hart and Shankman 2005,
Kincaid 2007). In northern areas, Eastern Hemlock occupies sites similar to those
in the South, as well as margins of swamps and lakes, and moist upland areas (Orwig
et al. 2012). Soil properties are variable on sites occupied by Eastern Hemlock
(Abella et al. 2003). Hemlock can form pure stands or grow in mixtures with numerous
other tree species varying within landscapes and among regions of its range
(Evans et al. 2011). Thus, extirpation of Eastern Hemlock would result in complete
stand replacement in pure Eastern Hemlock stands or alteration of mixed-species
stands (Fig. 1).
Eastern Hemlock’s recent biogeographic history reveals that a major decline
occurred in the past that is thought to have correlated with an extended dry period
(Hessl and Pederson 2013). Based on a study of buried pollen in Ontario, Canada
Eastern Hemlock expanded its importance in the region beginning ca. 9000 years
ago (Haas and McAndrews 2000). However, starting ca. 5800 years ago, Eastern
Hemlock declined about 6-fold during the next 1000 years when the climate warmed
and precipitation decreased from 83 to 70 cm/yr. Pollen of Pinus spp. (pine), Betula
spp. (birch), Fagus spp. (beech), Ulmus spp. (elm), Acer spp. (maple), and Quercus
spp. (oak) increased or was maintained during a 2000-yr period from the start of
Eastern Hemlock’s decline 5800 until 3800 years ago when Eastern Hemlock began
increasing. Thus, while Eastern Hemlock suffered a fairly recent major decline, it
is not thought to be as severe as today’s HWA-induced collapse. Moreover, new
stressors are prevalent (e.g., introduced organisms affecting tree species that increased
during Hemlock’s past decline) that could influence forest changes (Ford et
al. 2012, Small et al. 2005, Snyder et al. 201 1).
Value of Eastern Hemlock forests
Eastern Hemlock is considered a foundation species because it performs a unique
set of ecological functions not necessarily provided by any other tree species in
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eastern North American forests (Ellison et al. 2005, Vose et al. 2013). This species
plays a role in microclimate amelioration, watershed stabilization, soil ecology, and
nutrient cycling, as well as affecting plant species composition in forest communities
and wildlife habitat. Eastern Hemlock also has aesthetic and economic value
for humans (Table 1).
Because of its evergreen foliage and high crown-bulk density, Eastern Hemlock
creates a unique microclimate that moderates conditions throughout the year.
In summer, its foliage creates a dark, cool, moist microclimate below its canopy
that has important implications for habitat and watershed functions. In the southern
Appalachian Mountains of North Carolina for example, Webster et al. (2012)
found that a doubling of per hectare Eastern Hemlock foliage mass resulted in a
15% decrease in mean summer temperature. During winter, Eastern Hemlock can
ameliorate extreme temperatures and snowfall. In Vermont, Lishawa et al. (2007)
Figure. 1. a, b. View of mixed-species hillslope forest from afar and closer up with dead
Eastern Hemlock (the gray dead trees) in Great Smoky Mountains National Park. Photo by
S.R. Abella, 22 August 2013, along Highway 441 at 1200 m elevation near 35°35'44.9"N,
83°24'46.6"W (World Geodetic System 1984). c. Example of a typical site treated for Hemlock
Woolly Adelgid within Great Smoky Mountains National Park to maintain Eastern
Hemlock along riparian areas (photo by S.R. Abella, 21 August 2013). d. Trail damage from
the death and toppling of an Eastern Hemlock, now lying across a stream in the bottom right
of the photograph. Photo by S.R. Abella, 22 August 2013, near the Big East Fork Trailhead
(Shining Rock Wilderness, Pisgah National Forest, NC), 1200 m elevation near 35°21'32"N,
82°49'43"W.
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found that snow depth was 50% lower and minimum temperature averaged 2 °C
warmer in Eastern Hemlock forests than hardwood forests. Eastern Hemlock also
influences stream flow, stream water temperature, water chemistry, and light availability
for aquatic ecosystems differently than hardwood forests that also occupy
riparian areas (Ford and Vose 2007, Snyder et al. 2002). Comparative research to
clarify Eastern Hemlock’s functional role in watersheds, and whether these functions
vary among regions, is accumulating (e.g., Jenkins et al. 1999, Siderhurst et
al. 2010).
Several studies suggest that soil properties below Eastern Hemlock differ from
those below eastern hardwoods. Compared to five hardwood tree species in Connecticut,
for example, Finzi et al. (1998) reported that soil below individual Eastern
Hemlock trees exhibited elevated O horizon mass, C content, and N content, while
pH was lower. Similarly, at a more southerly location in Kentucky, Boettcher and
Kalisz (1990) reported that O horizon mass was greatest below Eastern Hemlock
while the pH of the O horizon and 0–5 cm mineral soil layer was low. Likely due
in part to acidic litter with a high C:N ratio, decomposition and N mineralization
below Eastern Hemlock can be slow (Boettcher and Kalisz 1990, Finzi et al. 1998).
Unique soil properties below Eastern Hemlock are thought to influence understory
plant species composition (Beatty 1984).
Eastern Hemlock forests support a unique assemblage of understory plants,
and total understory cover is low unless dominant, shade-tolerant shrubs such as
Rhododendron spp. (azalea, laurel) are present. In the southern Appalachians, for
example, species such as Hexastylis shuttleworthii (Britten and Baker f.) Small
(Largeflower Heartleaf), Mitchella repens L. (Partridgeberry), and Leucothoe
fontanesiana (Steud.) Sleumer (Highland Doghobble) were more important in
Table 1. Functions of Eastern Hemlock within ecosystems and for humans.
Function Examples Example reference
Microclimate creation All-year low-light environment; moderate Webster et al. 2012
temperature extremes
Soil protection Thick O horizons overlay mineral soil Beatty 1984
Nutrient retention Acidic, decomposition-resistant, high C:N litter Orwig et al. 2008
Carbon sequestration High biomass; high O horizon C content Finzi et al. 1998
Watershed value Shaded stream environment; coarse woody Siderhurst et al. 2010
debris production
Fish habitat Unique in-stream insect and fish composition Ross et al. 2003
Invertebrate habitat Unique communities; some species prefer Ingwell et al. 2012
Eastern Hemlock
Wildlife habitat 96 bird, 47 mammal species associates of Yamasaki et al. 2000
Eastern Hemlock in Northeast
Plant habitat Unique composition often with evergreen plants Abella and Shelburne 2004
Community diversity Increased within-landscape richness of Jenkins 2007
biological communities
Aesthetics Evergreen trees valued for their appearance Ellison et al. 2005
Recreation Hemlock location near water; ameliorated Quimby 1996
microclimate
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Eastern Hemlock forests than anywhere else on the landscape (Abella and Shelburne
2004). Rhododendron maximum L. (Great Rhododendron), a large evergreen
shrub, often dominates understories of Eastern Hemlock forests in the South. In
northern forests, Rogers (1980) concluded that understories of Eastern Hemlock
forests exhibited significant regional variation and were structurally unique within
landscapes. Eastern Hemlock forest understories contained abundant woody plants
(especially tree seedlings), evergreen plants, and ferns (Rogers 1980).
Eastern Hemlock contributes to landscape-scale wildlife diversity in at least
three ways by supporting: (i) some species largely restricted to Eastern Hemlock
forests, (ii) some species associated with other habitats but present in Eastern
Hemlock forests, and (iii) unique wildlife assemblages contributing to overall diversity
within landscapes (Ross et al. 2004, Yamasaki et al. 2000). In Wisconsin, for
instance, Howe and Mossman (1995) found that the bird species Dendroica fusca
Müller (Blackburnian Warbler) and Troglodytes troglodytes Vieillot (Winter Wren)
were three times as frequent in Eastern Hemlock as in hardwood forests. Yamasaki
et al. (2000) noted that 96 bird and 47 mammal species were associated with Eastern
Hemlock forests in northeastern states.
Eastern Hemlock provides numerous benefits to humans both inside and outside
park settings. Many ecological functions provided by Eastern Hemlock, such as
watershed functions, are also important to humans. Although timber production
is not a goal in parks, Eastern Hemlock was a source of timber and tannins in the
past, and pre-emptive logging of Eastern Hemlock on non-park lands before HWAinduced
mortality is an important contemporary issue (Kizlinski et al. 2002). While
ecological tradeoffs characterize pre-emptive logging, Eastern Hemlock is at least
temporarily providing lumber on some lands (Orwig et al. 2012). Partly because of
their frequent occurrence near streams, Eastern Hemlock forests are areas of human
recreation and valued for aesthetics (Quimby 1996).
HWA Autecology and Dynamics
HWA is a small (<1.5 mm long), aphid-like insect that is reddish-brown to
purplish-black in color. Its name comes from its woolly white appearance, which develops
as it matures and produces a covering of wool-like wax filaments to protect its
eggs. HWA is considered native to Japan where it inhabits forests containing Tsuga
sieboldii Carr. (Southern Japanese Hemlock; Havill et al. 2011). The insect is considered
a relatively innocuous component in its native range where, apparently, its
population is kept low by host resistance and natural enemies. In Connecticut, Mc-
Clure (1996) described three annual HWA generations: an overwintering generation
and two types of spring generations. All generations have six stages of development:
egg, four nymphal instars, and adult. These generations have different developmental
timing within a year (Havill et al. 2011). The largest adults average 1.4 mm long and
1.1 mm wide (McClure 1996). HWA kills Eastern Hemlock by sucking sap and depleting
the tree of starch reserves (Havill and Foottit 2007, Jonas et al. 2012).
Souto et al. (1996) noted that HWA was first reported in the eastern US in
the early 1950s, near Richmond, VA. From the mid-1950s to mid-1980s, HWA
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spread slowly through the mid-Atlantic states and was considered a pest only to
ornamental trees in urban environments and readily controllable using insecticides
(Souto et al. 1996). However, potentially aided by spread from a 1985 hurricane,
HWA continued dispersing to reach extensive natural stands of Eastern Hemlock,
and damage to Eastern Hemlock forests began increasing exponentially. As Havill
et al. (2011) summarized, HWA has not co-evolved with eastern North American
forest ecosystems, and as a result, Eastern Hemlock cannot resist or tolerate HWA
feeding, and the native community of natural enemies cannot keep HWA populations
below damaging levels.
Temperature appears important in HWA population dynamics and spread (Paradis
et al. 2008). Major declines in HWA populations in eastern North America have
followed cold years, but populations subsequently rebounded. HWA’s northern
spread also seems temperature-limited. McClure (1996) noted that HWAs introduced
to eastern North America were apparently from a non-cold-hardy strain,
because Japan has some other HWA populations that persist at low temperatures.
Climatic warming could accelerate HWA’s movement northward, but it should not
be assumed that warming is needed for this to occur because of potential adaptive
ability of HWA in the current climate (Butin et al. 2005). Morin et al. (2009) reported
that HWA has spread at an average rate of 9–20 km/year, and they further
noted that faster spread can occur through long-distance transport such as by humans
inadvertently moving infested material.
National Parks Impacted by HWA
Eighty-five parks, or 21% of the 401 parks in the US NPS system, are within
the range of Eastern Hemlock (Fig. 2). Parks within Eastern Hemlock’s range include
significant American natural and cultural features such as in Valley Forge
National Historical Park, King of Prussia, PA; Antietam National Battlefield,
Sharpsburg, MD; Gettysburg National Military Park, Gettysburg, PA; Hopewell
Culture National Historical Park, Chillicothe, OH; Appalachian National Scenic
Trail (GA–ME); and in the largest eastern parks including Great Smoky
Mountains National Park (GSMNP), Gatlinburg, TN; Shenandoah National
Park, Luray, VA; Sleeping Bear Dunes National Lakeshore, Empire, MI; and
Acadia National Park, Bar Harbor, ME. Based on 2011–2012 HWA distribution
maps, HWA had not spread into northern Maine and was not reported in Acadia
National Park (Harris et al. 2012), was absent from the upper Great Lakes including
parks in Michigan and Wisconsin, and had not reached disjunct Eastern
Hemlock populations in western Kentucky in Mammoth Cave National Park,
Mammoth Cave, KY (US Forest Service, Forest Health Technology Enterprise
Team, http://foresthealth.fs.usda.gov/portal).
Some isolated populations of Eastern Hemlock exist in western Georgia, Kentucky,
and Indiana, which include some national park units such as Mammoth Cave
National Park (Fig. 2). The disjunct populations might afford some opportunity to
attempt limiting establishment of HWA within these isolated areas through prevention,
early detection, and treatment of any incipient HWA populations (Hart 2008,
Koch et al. 2006).
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Table 2 lists 16 parks that exceed 2000 ha and contain Eastern Hemlock, with
10 of these parks—all but those in the Great Lakes, northern Maine, and western
Kentucky— reporting presence of HWA. Excluding Appalachian National Scenic
Trail for which records were not available, the 15 other parks have received a
combined visitation of 45 million people annually from 2008–2012 (Public Use
Statistics Office, NPS, Denver, CO). Based on vegetation maps available for 10 of
the parks (NPS, Vegetation Inventory Program, Fort Collins, CO) and a fire vegetation
map for Big South Fork National River and Recreation Area, Oneida, TN, pure
Eastern Hemlock forests or mixed-species forests with Eastern Hemlock occupy
about 2–26% of park area (Table 2). Thus, there are parks such as Mammoth Cave
National Park in which Eastern Hemlock forests are locally rare, unique communities
and parks such as GSMNP in which Eastern Hemlock forms a dominant part
of total forest cover. Overall, for the 11 parks with vegetation maps, the combined
Eastern Hemlock forest resource exceeds 60,000 ha.
Impacts in Parks
Eastern Hemlock mortality after HWA infestation can be rapid and pervasive
(Ford et al. 2012). There is some encouraging recent information, however, that
Eastern Hemlock mortality can be a slower process in some situations, at least in
Figure 2. Range of Eastern Hemlock corresponding with distribut ion of National Park Service
units (shown in black). Sixteen parks exceeding 2000 ha in size and within Eastern
Hemlock’s range are numbered according to the legend.
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northern areas, than was previously thought (Eschtruth et al. 2013, Orwig et al.
2012). Long-term monitoring in Delaware Water Gap National Recreation Area
(DWGNRA) in NJ/PA revealed that 65% of infested Eastern Hemlock trees originally
monitored in 1993 remained alive 20 years later in 2012 (Eschtruth et al.
2013). Mortality in both northern (Orwig et al. 2002) and southern regions (Martin
and Goebel 2012) is thought to occur most quickly and be most extensive on dry
sites where Eastern Hemlock trees are stressed. Duration of HWA infestation, however,
can overwhelm importance of other site factors that influence the health of
Eastern Hemlock forests (Martin and Goebel 2012, Rentch et al. 2009). Droughts
seem to be associated with rapid HWA-induced Eastern Hemlock mortality, e.g.,
the rapid mortality (often <4 years) recorded in GSMNP during a mid-2000s dry
period (J. Webster, GSMNP, Gatlinburg, TN, pers. comm.). Our current understanding
suggests that Eastern Hemlock mortality rates partly depend on site and
climatic conditions and can be rapid (within a few years) or slow over a period
of decades. Recent findings also suggest that further research to clarify site and
climatic factors affecting Eastern Hemlock mortality patterns is warranted to help
prioritize HWA management treatments (Eschtruth et al. 2013, Martin and Goebel
2012, Rentch et al. 2009). It should be noted that these recent studies indicate that
while slowed mortality at least in some cases is encouraging, the total mortality
Table 2. National Park Service units exceeding 2000 ha containing Eastern Hemlock forests. NP =
National Park, NL = National Lakeshore, NR = National River, NHP = National Historic Park, and
NRA = National Recreation Area.
States Size (ha) Hemlock ha (%)A Polygons (%)A AdelgidB
Acadia NP ME 14,473 2332 (6.8) 302 (6.2)
Apostle Islands NL WI 17,062 1382 (4.8) 733 (15.1)
Appalachian National ME to GA 42,853 – – ×
Scenic Trail
Big South Fork NR and NRA KY/TN 47,070 11,942 (24.1) – ×
Blue Ridge Parkway NC/VA 33,272 – – ×
Chesapeake and Ohio MD/DC/WV 5325 – – ×
Canal NHP
Catoctin Mountain Park MD 2384 – – ×
Cumberland Gap NHP KY/TN/VA 9852 166 (2.0) 38 (1.7) ×
Cuyahoga Valley NP OH 7697 – –
Delaware Water Gap NRA NJ/PA 22,761 3233 (11.6) 316 (6.7) ×
Great Smoky Mountains NP TN/NC 211,219 57,106 (26.0)C 11,604 (23.2)C ×
Mammoth Cave NP KY 21,045 262 (1.3) 73 (0.9)
New River Gorge NR WV 21,687 1339 (4.6) 855 (10.3) ×
Pictured Rocks NL MI 14,434 2157 (6.4) 516 (10.7)
Shenandoah NP VA 79,904 1903 (2.4) – ×
Sleeping Bear Dunes NL MI 22,953 687 (2.0) 443 (4.8)
AVegetation mapping could include a buffer area surrounding parks, and the percentage of Eastern
Hemlock area and number of polygons was calculated based on the total area of mapped vegetation.
(–) indicates mapping is not complete.
BSymbols (×) indicate Hemlock Woolly Adelgid is present within a park as of 2012.
CAmount of pure Eastern Hemlock forest approximates 6381 ha, or 2.9% of the park.
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is frequently extensive, and that trees weakened by HWA are susceptible to other
pests, windthrow, and other mortality-causing agents. Additionally, dieback of
Eastern Hemlock tree crowns, even when trees have not died, can have ecological
effects due to changes in light availability and other factors (Eschtruth et al. 2013).
Research regarding existing and projected HWA impacts has been published
for four NPS units currently infested by HWA (Table 3). Available research on
impacts to vegetation, fungi, watersheds, fisheries, wildlife, and invertebrates, is
summarized in the following sections for DWGNRA (NJ/PA), GSMNP (NC/TN),
Shenandoah National Park (VA), and New River Gorge National River (WV) to illustrate
types of impacts to parks and their broad geographic extent.
Delaware Water Gap National Recreation Area
HWA was detected in DWGNRA in 1989, and the park has the longest-term, most
extensively reported monitoring of HWA infestation, Eastern Hemlock decline, and
Table 3. Examples of impacts or potential impacts to Eastern Hemlock ecosystems from Hemlock
Woolly Adelgid reported from National Park Service units.
National Park Service unit/HWA-impact reports Reference
Delaware Water Gap National Recreation Area
Four bird species at risk strongly associated with Eastern He mlock Ross et al. (2004)
Fewer breeding Acadian Flycatchers in HWA-infested sites Allen et al. (2009)
Bryophyte cover doubled after HWA infestation Cleavitt et al. (2008)
Understory plant cover doubled; tree seedling changes variabl e Eschtruth et al. (2006)
Interaction between HWA and deer influence tree changes Eschtruth and Battles (2008)
Increased invasion of exotic plants in Adelgid sites Eschtruth and Battles (2009)
Overview of numerous past and present ecosystem changes Evans (2010)
Native Brook Trout had 3× abundance in Eastern Hemlock than Ross et al. (2003)
hardwood streams
Aquatic invertebrates differed between Eastern Hemlock and Snyder et al. (2002)
hardwood streams
Great Smoky Mountains National Park
Unique fungal communities associated with Eastern Hemlock, Baird et al. (2007, 2009)
including species new to science
Both Adelgid and chemical treatments for HWA can affect Falcone and DeWald (2010)
insects and birds
Evergreen shrub Great Rhododendron might affect tree and Krapfl et al. (2011, 2012)
understory plant changes
In-stream light increased overall, but Great Rhododendron cou ld Roberts et al. (2009)
mollify aquatic changes
Shenandoah National Park
Overall arthropod abundance could increase but at expense of Rohr et al. (2009)
Eastern Hemlock-associated species
Eastern Hemlock crown condition declines from 80% healthy tre es Willeford Bair (2002)
in 1990 to near 0% in 2000
New River Gorge National River
Uncertain future forest composition because of disease afflict ing Martin and Goebel (2012)
other tree species
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consequent ecological impacts in the national park system (Evans 2010). Evans et
al. (1996) provided an early projection of anticipated HWA impacts: gap infilling
by tree species such as Betula alleghaniensis Britton (Yellow Birch) and Acer rubrum
L. (Red Maple), reduced bryophyte abundance, increased overall abundance
of understory plants including more exotic plants, increased soil N availability,
increased understory and stream light levels and temperatures, reduced populations
of native Salvelinus fontinalis Mitchill (Brook Trout), and reduced populations of
birds strongly associated with Eastern Hemlock. Subsequent studies allow evaluation
of some of these projected impacts.
Effects of HWA on vegetation in DWGNRA have been significant. Eschtruth et
al. (2006) reported on changes from 1994 (pre-HWA) to 2003, with major Eastern
Hemlock mortality beginning before 1999. In 1994, over 95% of Eastern Hemlock
trees at Adams Creek within the park had healthy crown-vigor ratings, but only 9
years later in 2003, <35% of trees were healthy, and the rest were either declining
(45%) or already dead (20%). Light transmission to the understory more than
doubled from 1994–2003, supporting the prediction that forest light-levels would
increase. Note that while light likely decreased as other trees filled in the gaps, this
reduction would only be appreciable during the leaf-on growing season because
these tree species were deciduous. Understory-tree relative density changes were
difficult to discern, as Fagus grandifolia Ehrh. (American Beech) increased by
four-fold at one site but not another, and relative density of other species such as
Red Maple and Betula lenta L. (Sweet Birch) changed little. As Evans et al. (1996)
predicted, overall understory plant cover more than doubled from 1994–2003, but
remained below 17%. Future tree species composition of former Eastern Hemlock
forests is difficult to predict because of the numerous factors that can affect tree
establishment. For example, Odocoileus virginianus Zimmermann (White-Tailed
Deer) density exceeds 20 animals/km2 in DWGNRA, and these animals can dramatically
influence forest change following Eastern Hemlock decline (Eschtruth
and Battles 2008).
Some studies have not supported the prediction that native bryophyte
abundance would decline but have supported the prediction that exotic plant
cover would increase post-HWA. Cleavitt et al. (2008) reported that bryophyte
cover doubled from 1994 (pre-HWA) to 2006 (about 7 years post-HWA),
with increases in species such as Dicranum montanum Hedw. (Montane Dicranum
Moss) and Brachythecium rutabulum (Hedw.) Schimp. (Brachythecium
Moss), as well as species associated with coarse woody debris. Cleavitt et al.
(2008) further noted uncertainty about how long bryophytes might continue to
increase, and discussed factors such as establishment of hardwood trees and
shrubs that could affect future bryophyte composition. Eschtruth and Battles
(2009) found that exotic plants increased with light availability following Eastern
Hemlock mortality. Moreover, White-Tailed Deer herbivory interacted with
Eastern Hemlock mortality to further increase exotic plants while decreasing
native plants. The priority exotic plants Alliaria petiolata (M. Bieb.) Cavara
and Grande (Garlic Mustard), Berberis thunbergii DC. (Japanese Barberry), and
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Microstegium vimineum (Trin.) A. Camus (Nepalese Browntop) all increased outside
of exclosures open to deer herbivory in HWA-affected forests.
Snyder et al. (2002) and Ross et al. (2003) highlighted potential HWA impacts
to watershed functions and aquatic ecosystems in the park. Snyder et al. (2002)
found that aquatic invertebrate composition was strongly correlated with forest
composition (Eastern Hemlock versus hardwood forests). Streams draining Eastern
Hemlock forests had 8% of their aquatic taxa strongly associated with Eastern
Hemlock, and species evenness (one component of diversity) was higher in streams
running through Eastern Hemlock than hardwood forests. The authors concluded
that declines in Eastern Hemlock indicated potential for reductions in within-stream
and park-wide benthic diversity. Ross et al. (2003) found that native Brook Trout
were three times as prevalent in Eastern Hemlock as in hardwood forest streams.
Fish species richness did not differ between Eastern Hemlock and hardwood forest
streams, suggesting that changes might be mostly compositional.
Two studies in the park have examined relationships of bird species with Eastern
Hemlock forests. Ross et al. (2004) reported that bird species richness was greater
in hardwoods than in Eastern Hemlock forests, but four bird species strongly associated
with Eastern Hemlock were at risk. These four were insectivorous species including
Empidonax virescens Vieillot (Acadian Flycatcher), Vireo solitarius Wilson
(Blue-Headed Vireo), Dendroica virens Gmelin (Black-Throated Green Warbler),
and Dendroica fusca Müller (Blackburnian Warbler). Allen et al. (2009) further
compared abundance of Acadian Flycatcher across sites representing a gradient of
Eastern Hemlock mortality. Eastern Hemlock supported 90% of nests of this species,
and 70% fewer breeding pairs occupied HWA-infested sites. These studies,
combined with those on other lands (Becker et al. 2008, Tingley et al. 2002), suggest
that while some bird species might benefit from Eastern Hemlock mortality,
Hemlock-associated species such as the Acadian Flycatcher have declined. These
species will likely decline further unless they adapt to changing habitat conditions.
Great Smoky Mountains National Park
HWA was first identified in GSMNP in 2002 (Lambdin et al. 2006), and potentially
because of the park’s southerly location and subsequent drought, Eastern
Hemlock trees died rapidly, often within a few years of HWA infestation (Ford et
al. 2012). Krapfl et al. (2011, 2012) measured 32 plots established in the park in
2003 before full HWA infestation and resampled the plots in 2008–2009 when all
plots contained HWA. By the 5-year post-HWA measurement, Eastern Hemlock
overstory and understory tree density declined. Similar to results from DWGNRA
(Eschtruth et al. 2006), only 20% of overstory Eastern Hemlock had good crowncondition
ratings, with the remaining trees dead or severely declining. Overstory
tree species composition (44 species, mostly hardwoods, on all plots) and understory
tree and tall-shrub composition (79 species) were species-rich, but no difference
existed in relative stem density (excluding Eastern Hemlock) of species between
2003 and 2008/2009. The authors hypothesized that in many areas, future species
composition might hinge upon responses of the dominant native, evergreen shrub
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Great Rhododendron, which might partly buffer changes in light transmission to
the forest floor from Eastern Hemlock die of f (Krapfl et al. 2012).
Fungi were compared among tree species and between HWA-damaged and
healthy bark of Eastern Hemlock trees by Baird et al. (2007, 2009). Microfungal
assemblages did not differ between damaged and healthy bark within ≈1–2 years
of HWA infestation. A rare fungus, Thysanophora canadensis Stolk and Hennebert,
occurred only on Eastern Hemlock, raising the possibility that this fungal species
is at risk from Eastern Hemlock mortality (Baird et al. 2007). With over 55 fungal
species new to the park discovered in 2005–2006 during an All Taxa Biodiversity
Inventory, the possibility exists that numerous little-known organisms such as fungi
might be impacted by HWA before we even know the species exist in parks (Baird
et al. 2009).
Falcone and DeWald (2010) compared arthropods and insectivorous birds
between HWA-infested Eastern Hemlock stands treated with insecticide for two
years and untreated HWA-infested stands. The study did not detect differences
in total arthropod abundance or density of insectivorous birds between ongoing
treatment areas and untreated stands. However, herbivorous Hemiptera and larval
Lepidoptera were significantly lower in treated stands. This illustrated the potential
for non-target treatment effects (Falcone and DeWald 2010) and a difficulty in
evaluating impacts of both HWA and counter-treatments: lack of unaffected Eastern
Hemlock reference sites for comparison.
Roberts et al. (2009) compared characteristics of streams in the park flowing
through paired Eastern Hemlock and hardwood forests to project potential changes
if hardwoods replaced Eastern Hemlock. Unlike studies elsewhere, little relationship
between forest type and stream characteristics of discharge, water temperature,
pH, and NO3-N concentration was evident over a one-year period. There were some
differences in light levels (mainly during the leaf-off period of deciduous trees), but
presence of Great Rhododendron moderated differences. Light levels in hardwood
forests with Great Rhododendron understories resembled light levels of streams
in Eastern Hemlock forests. Their study, combined with Krapfl et al. (2012), suggested
that Great Rhododendron dynamics might influence numerous changes after
Eastern Hemlock trees die within parks.
Shenandoah National Park
Willeford Bair (2002) described networks of plots established in 1990 and 1999
to monitor the condition of Eastern Hemlock trees. From 1990–2000, the percentage
of Eastern Hemlock trees in the excellent crown-condition class dropped
sharply from 80% to near zero, with 49% mortality. While Eastern Hemlock trees
at higher elevations were still impacted by HWA, crown condition on average was
twice as good at elevations above 750 m than below. This finding was consistent
with observations that HWA populations currently seem limited by cold temperatures
(Morin et al. 2009). Willeford Bair (2002) made the important point that other
factors being equal, high-elevation Eastern Hemlock populations may warrant special
attention for preventive HWA management.
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Another study in the park compared arthropod assemblages in Eastern Hemlock
stands with those in hardwood forests that might replace the former. Rohr et al.
(2009) concluded that Eastern Hemlock loss might increase overall arthropod abundance,
including increased abundance for 23 arthropod taxa, while overall species
richness, including 7 taxa strongly associated with Eastern Hemlock, could decline.
New River Gorge National River
Martin and Goebel (2012) used a 9- to 32-year chronosequence of time-sinceinvasion
of HWA to evaluate dynamics in Eastern Hemlock forests in West Virginia
and Virginia riparian areas, including along the New River. As Willeford Bair
(2002) found, Eastern Hemlock decline was moderated at the highest elevations
and in topographic positions with the coolest temperatures, but time since invasion
of HWA was the primary factor in Hemlock decline. The authors concluded
that relatively slow decline of Eastern Hemlock stands over a period of years to
decades characterizes this park, similar to DWGNRA, and that while elevation or
topography may initially slow the decline, duration of HWA infestation eventually
predominates the dynamics of infestation. Based on the finding that saplings and
seedlings were at least as rapidly affected as canopy trees, the authors projected
essentially complete mortality of all Eastern Hemlock size classes (Martin and
Goebel 2012). The authors also expressed uncertainty about future forest composition
because composition could be contingent upon numerous factors, including
presence of the exotic, competitive tree Ailanthus altissima (Mill.) Swingle (Tree
of Heaven), which comprised over 50% relative basal area at some sites. Moreover,
American Beech, a shade-tolerant potential replacement tree, also is under stress
from infection with the introduced Beech Bark Disease.
Synthesis of impacts
Understanding HWA impacts within national parks and on other lands surrounding
parks is important for at least two reasons: connectedness of park ecosystems
with those of adjacent land (e.g., HWA spread, dispersion of biocontrol species into
and out of parks, and watershed connectivity), and the potential to extrapolate what
is learned on other lands and in parks, and apply it in the other. Research regarding
HWA impacts in national parks is generally consistent with research conducted on
other lands, and the collective literature has suggested several general principles
related to HWA impacts: (1) Cold temperature appears to limit HWA distribution,
and while warming temperatures might expedite HWA’s spread, it should not be
assumed that warming is needed for the continued spread of the infestation given
HWA’s apparent adaptive capacity (Butin et al. 2005, Paradis et al. 2008); (2) All size
classes of Eastern Hemlock trees have been killed by HWA, and while searching for
potentially resistant trees is warranted, resistance appears low to non-existent (Vose
et al. 2013); (3) Eastern Hemlock mortality might be fastest in warmer regions, but in
non-drought periods, mortality at a stand scale might still occur over a period of years
to decades (Ford et al. 2012); (4) HWA increases abundance of dead wood, which
can have numerous effects on ecosystems (e.g., soil nutrient cycling, input of wood
to streams) and create hazard trees at human-use sites (Evans and Shreiner 2008);
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(5) There likely will be short-term impacts immediately following Eastern Hemlock
mortality (e.g., increased light) and longer-term impacts likely contingent upon postinfestation
vegetation dynamics (Small et al. 2005); (6) Many impacts of Eastern
Hemlock mortality to ecosystem functions likely depend on post-infestation forest
structure and which resulting forest communities become established (Farnsworth
et al. 2012); (7) Tree species and vegetation that may replace Eastern Hemlock forests
are difficult to predict (Orwig et al. 2012); (8) Colonizing vegetation will likely
be controlled by numerous factors, such as damaging agents (e.g., exotic pests) affecting
potential tree species replacements, White-Tailed Deer herbivory, invasion
of exotic plants, and dynamics of competitive native plant species such as Great
Rhododendron (Eschtruth et al. 2006); (9) Impacts to ecosystem functions and variability
in colonizing vegetation within landscapes and among regions will likely be
substantial (Evans et al. 2011); (10) As with many ecological events, some species
will benefit from Eastern Hemlock mortality. Species associated with Eastern Hemlock
must adapt or decline, and Hemlock loss will reduce species diversity unless it
triggers long-term speciation (Adkins and Rieske 2013 , Becker et al. 2008, Yorks et
al. 2003); and (11) Further research to identify organisms strongly associated with
Eastern Hemlock could help determine which species are most at risk from Hemlock
loss (Baird et al. 2009).
Previous research has also further described specific ecological impacts of
Eastern Hemlock mortality: (1) Microclimate changes are projected and there will
be greater variation in temperature and light with loss of Hemlock’s ameliorating
influence, and more light will reach the forest floor, although this effect could be
partly offset if other evergreen species increase (Webster et al. 2012); (2) Nutrient
cycling will probably change, likely with altered soil organic matter pools
and increased N mineralization and nitrification (Stadler et al. 2006); (3) Several
comparisons exist of stream and fishery characteristics between Eastern Hemlock
forests and potential replacement hardwood forests (e.g., Ross et al. 2003), but
actual impacts likely will remain uncertain until longer-term post-Hemlock assessments
are available, including those that have encompassed the effects of major
storm events (Siderhurst et al. 2010). Based on existing work, impacts to post-infestation
stream characteristics might not be dramatic at every site, especially when
other evergreen plants colonize those formerly covered by Eastern Hemlock (Roberts
et al. 2009); (4) Understory plant cover and species richness have increased
after Eastern Hemlock mortality. However, exotic plant invasion also increases on
sites that formerly supported Hemlock stands (Eschtruth and Battles 2009); and
(5) Changes in wildlife and invertebrate species composition largely appear to have
hinged upon interaction between species favoring Eastern Hemlock and those that
do not favor habitat conditions created by Hemlock (Ingwell et al. 2012).
National Park Management
Results from studies of HWA invasion highlight challenges that the NPS will
increasingly face from introduced forest pests, including in key policy areas such
as biocontrol and facilitating park adaptation to ecological change. Moreover, Vose
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et al.’s (2013) review of potential HWA management actions, largely geared to
non-park lands, is useful for illuminating some aspects unique to managing a forest
die-off event in a national park context. A range of potential strategies to reduce
undesired impacts of HWA to parks warrants consideration (Table 4).
Collect genetic material
Resistance of Eastern Hemlock trees to HWA seems minimal (Orwig et al. 2002).
Nevertheless, surveys for potentially resistant trees are warranted within parks,
especially considering that pre-emptive salvage logging is not a goal in parks. Preemptive
logging has occurred on private lands, and while this can generate revenue
for landowners, it reduces the possibility of identifying resistant trees if any population
or individual-tree resistance happens to occur (Kizlinski et al. 2002). Ingwell
and Preisser (2011) described a citizen-science program in which volunteers survey
for potentially resistant Eastern Hemlock trees, and the authors suggested that some
trees showed resistance to adult HWA. This type of volunteer program can be highly
applicable in a national park context and is usually a low-cost tool if the only park
commitment is to provide coordination and training (Ingwell and Preisser 2011).
Regardless of whether there are HWA-resistant Eastern Hemlocks, collection of
genetic material from parks is warranted so that material is available for resistance
breeding programs and for Eastern Hemlock restoration if HWA can be controlled
(Jetton et al. 2010). American Chestnut represents an example of an ongoing program
to breed resistance to an introduced damaging agent and illustrates advantages
and disadvantages of such programs. Trees with genotypes similar to the original
wild type are available for test field-plantings, but their genetics have been altered
from the original species, and reintroduction of these trees might displace native
tree species that have partly replaced American Chestnut (Jacobs 2007). With due
consideration to these tradeoffs, current NPS policy allows flexibility in supporting
breeding programs to introduce resistant trees (NPS 2006).
Chemical treatment
Chemical treatments have served to keep individual trees or small stands of
Eastern Hemlock alive, and their use requires balancing negative, non-target effects
Table 4. Summary of preventive, adaptive, and restorative strategies for Eastern Hemlock forests
impacted by Hemlock Woolly Adelgid in a National Park Service context.
Activity Example reference
Constrain HWA spread to disjuncts Hart (2008)
Biocontrol Havill et al. (2012)
Chemical control Cowles (2009)
Manage deer density Eschtruth and Battles (2009)
Control exotic plants Evans and Shreiner (2008)
Reduce other stressors Ellison et al. (2005)
Promote replacement vegetation Vose et al. (2013)
Limit replacement tree die off Small et al. (2005)
Collect genetic material Jetton et al. (2010)
Protect public from hazard trees Johnson et al. (2008)
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of the treatment and its financial cost versus consequences of losing the trees (Knoepp
et al. 2012). The insecticide imidacloprid is a recommended chemical for HWA
control. It is either applied to soil, injected into tree trunks, or applied as a foliar
spray. Cowles et al. (2006) tested several trunk and soil applications on Eastern
Hemlock trees in Connecticut and concluded that soil applications reduced HWA
populations by 50–100% for >2 years, but no trunk injection-treatment significantly
reduced HWA. Further research determined the minimum doses required and the
optimal form (tablets versus powder) of soil-applied imidacloprid that would maximize
efficacy for HWA control and minimize the potential for undesired transport
to aquatic ecosystems (Cowles 2009). In the southern Appalachian Mountains,
Knoepp et al. (2012) reported that horizontal movement of imidacloprid in soil was
limited. They found that imidacloprid concentration at canopy drip lines was low
compared to the concentration near trunks. Their study illustrated that elevation and
soil organic matter (and potentially texture) can influence imidacloprid retention in
soil among sites.
GSMNP is an example of a park that has aggressively treated HWA (Johnson
et al. 2008, Webster 2010). The park has used imidacloprid to treat over 250,000
Eastern Hemlock trees since the early 2000s in prioritized areas such as along
streams, trails, roads, and specially identified Eastern Hemlock conservation areas
(K. Johnson and J. Webster, GSMNP, Gatlinburg, TN, pers. comm.). The park’s
treatment protocol has evolved over time, and the current preferred method is an
aqueous soil drench of imidacloprid applied to the soil surface around trunks of
trees (J. Webster, pers. comm.). For trees <63 cm diameter at breast height (DBH),
current chemical costs are $0.09/2.5 cm DBH at a dose rate of 0.7 g active ingredient/
2.5 cm DBH. Trees ≥63 cm DBH are dosed at 1.4 g active ingredient/2.5 cm
DBH. A range of tree sizes (including small trees <20 cm DBH) have been treated
by soil drench to conserve forest structure. Effectiveness of an individual treatment
now appears to last longer than previously thought. The park believes that treatments
effectively control HWA for 5 years and possibly as long as 8 years before
re-treatment is necessary to keep HWA populations low (J. Webster, pers. comm.).
If water is unavailable at treatment sites, the park uses imidacloprid tablets placed
in soil around tree trunks.
Countless untreated Eastern Hemlock trees have died throughout the GSMNP,
and given limited resources for treatment, it is not feasible to treat all trees within
the park. Park staff view chemical treatments as a way to maintain some priority
Eastern Hemlock stands and as a strategy to buy time until other chemical or biocontrol
treatments are developed, or other processes limit the spread and impact of
HWA (K. Johnson and J. Webster, pers. comm.). In addition to continuing treatment
of currently treated stands, the park is interested in assessing and treating some
high-elevation Eastern Hemlock stands that occur above 1400 m (Webster 2010).
HWA populations might be limited by cold temperatures at these sites, and Eastern
Hemlock conservation zones could be strategically expanded there.
DWGNRA provides another context in which active HWA control is ongoing
(Evans and Schreiner 2008). Since 2003, the park has treated about 25,000 Eastern
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Hemlock trees with insecticide (R.A. Evans, DWGNRA, Milford, PA, pers. comm.).
Developed visitor-use areas have been prioritized for treatment to prevent Eastern
Hemlock trees from declining and becoming hazard trees. The park’s second priority
has been treating riparian Eastern Hemlock forests to protect streams from
loss of fisheries habitat, prevent stream-bank erosion, and maintain water quality.
In upland areas, the preferred treatment method is imidacloprid tablet application
to soil around Hemlock trunks (R.A. Evans, pers. comm.). The number of tablets
applied increases with the square of the DBH of the tree (in inches), such that the
number of tablets equates to DBH2/10 for trees >15 cm (6 in.) in DBH. Thus, a
25-cm (10-in) DBH tree requires 10 tablets, costing $7–10 total for the tablets.
These chemical costs are greater than for a soil drench, but tablet applications do
not require mixing with water and are faster, safer, and cheaper to apply compared
to a soil drench at this park (R.A. Evans, pers. comm.). Trees in riparian areas and
near surface water are treated with imidacloprid stem injections, which cost about
10 times as much as tablet application for the same size tree. As at GSMNP, the
park is concerned about both non-target effects of the treatment and consequences
of leaving Eastern Hemlock trees untreated (Evans 2010). Key challenges facing
park managers regarding use of chemical treatments include: identifying priority
treatment sites to maximize benefit from limited resources; identifying non-target
effects and weighing them against the effects of Eastern Hemlock loss; and balancing
time and effort for chemical treatment with development of other treatments or
longer-term management interventions (Evans and Schreiner 2008).
Biocontrol
Biocontrol is the release of predators or damaging agents, usually from the
homeland of the organism targeted for treatment, to reduce a target organism in its
introduced habitat (van Lenteren et al. 2006). In the US, the process of preparing
biocontrols occurs under the Department of Agriculture’s Animal and Plant Health
Inspection Service (APHIS; Montgomery 2011). Potential biocontrol agents are
usually first screened by examining organisms associated with a target organism in
its native habitat. Next, feeding trials to determine food preferences of a potential
biocontrol for the target organism and some non-target organisms are conducted
in a controlled environment. If a biocontrol agent is authorized for release, the
organism is released into the wild and researchers evaluate its ability to become
established (van Lenteren et al. 2006, Montgomery 2011). Evaluations of full ecosystem
effects of a biocontrol organism are rarely available, but there are numerous
examples of biocontrols that have reduced the populations of their target organism
(Van Driesche et al. 2010). A biocontrol program is considered successful if the
cost and time (often decades) required for development and potential negative
non-target effects are offset by reductions in the target organism compared to doing
nothing or implementing other types of treatments.
Because, by definition, biocontrol involves release of one or more non-native
organisms, numerous authors have noted the care with which biocontrol programs
must be developed and the fact that ecological effects of a biocontrol in wildland
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ecosystems cannot be fully known until they are actually occurring (Simberloff
and Stiling 1996, van Lenteren et al. 2006). The likelihood that a biocontrol
organism will have immediate undesired effects, longer-term undesired effects
(e.g., switching food sources to native species if the target organism is reduced
by the biocontrol), or evolve to become invasive organisms themselves cannot be
completely dismissed a priori, but the possibilities can be reduced through careful
pre-release testing (van Lenteren et al. 2006).
Biocontrol currently represents a conundrum within NPS management policy
(NPS 2006). Integrated Pest Management Coordinators in the NPS have discretion
to authorize potential release of biocontrols approved for release in the US, which is
then contingent upon development and approval of plans at individual parks (NPS
2006). The conundrum is that promoting establishment of exotic species (through
release of biocontrols) within parks is generally inconsistent with NPS policy, but
protecting native species is consistent with policy, as is limited use of exotic species
to achieve resource objectives when viable alternatives are not available (NPS
2006). Two other important issues are that evaluations of non-target effects may or
may not need to be more conservative for parks than for initial APHIS approval (but
ability of the NPS to conduct these evaluations is limited), and park boundaries are
porous to biocontrol spread such that biocontrols may inhabit parks whether or not
released in parks.
Numerous biocontrol species have been released in North America as attempts
to control HWA (Montgomery 2011), and it should be recognized that this has
resulted in further exotic species introductions. Exotic species were needed as
biocontrols because predatory insects native to eastern North America have not coevolved
with HWA, are not specialists on HWA, and have not kept HWA densities
sufficiently low to prevent death of Eastern Hemlock (Havill et al. 2011). A major
current challenge is allowing sufficient time for careful field trials but still moving
biocontrol programs forward while some living Eastern Hemlock trees remain
(Vose et al. 2013).
Havill et al. (2012) reported on an unintended outcome that is already evident
following the release of two beetle biocontrols for HWA. Laricobius nigrinus
Fender (Black Beetle), native to western North America but not to eastern North
America, has been released at numerous field sites since 2003 within Eastern Hemlock’s
range. This beetle has reduced HWA density, but Eastern Hemlock stands
have not fully recovered from HWA infestation. Laricobius rubidus LeConte
(Tooth-necked Fungus Beetle) is the only member of the genus endemic to eastern
North America, and it feeds on HWA but prefers feeding on Pineus strobi Hartig
(Pine Bark Adelgid) that in turn feeds on the native tree Pinus strobus L. (Eastern
White Pine). The introduced biocontrol beetle has been hybridizing with the native
beetle, and effects of this hybridization are unclear (Havill et al. 2012). Given that
Eastern White Pine distribution overlaps with Eastern Hemlock and in many areas
is a potential replacement tree when Eastern Hemlock is lost, the potential for adding
another stressor to Eastern White Pine through release of an HWA biocontrol is
disconcerting (Vose et al. 2013).
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Laricobius nigrinus, Laricobius osakensis Montgomery, Sasajiscymnus tsugae
Sasaji McClure, Scymnus coniferarum Crotch, Scymnus sinuanodulus Yu and Yao,
and Scymnus ningshanensis Yu and Yao are biocontrol insects released in GSMNP
and DWGNRA (Evans and Schreiner 2008, Johnson et al. 2008). As of 2013, over
500,000 biocontrol beetles had been released in GSMNP (K. Johnson, pers. comm.)
and over 80,000 in DWGNRA (Evans and Schreiner 2008). The NPS faces difficult
choices regarding the use of biocontrol agents in parks because other treatments
such as insecticides are not feasible to conduct over large areas, and when they are
used, they can induce non-target effects; doing nothing (i.e. likely allowing hemlocks
to die) would also have major effects. Because Eastern Hemlock-dependent
organisms have not been fully identified within parks, there is no way to accurately
predict potential non-target impacts of biocontrols, nor, conversely, the full impacts
of Eastern Hemlock extirpation. Moreover, park boundaries are porous to biocontrol
organisms and parks are likely to contain the organisms whether or not they
are released directly within parks. This fact has led parks to release biocontrols for
HWA: parks would receive any non-target impacts of biocontrols regardless, but by
releasing biocontrols in parks, the parks could receive benefits faster (R.A. Evans,
pers. comm.). As an overall strategy for biocontrol and forest health, careful analysis
and monitoring of tradeoffs of biocontrols versus other treatments (including
doing nothing), while also ensuring that potential replacement tree species (e.g.,
Eastern White Pine) are not harmed, seems prudent.
Facilitate adaptation
Eastern Hemlock has already been lost from appreciable areas of parks, and
several strategies could help ecosystems adapt to current and potential Eastern
Hemlock losses. A major principle in conservation biology is that minimizing
other stressors can increase an indigenous ecosystem’s ability to adapt to a
primary stressor (Lovett et al. 2006). Actions that may promote adaptability of
ecosystems formerly dominated by Eastern Hemlock could include, but are not
limited to, reducing invasion of exotic plant species, maintaining soil health (e.g.,
limiting effects of pollution deposition), limiting unnatural levels of herbivory
such as those caused by overabundance of White-Tailed Deer, and protecting
and promoting colonizing native plant species (Webster et al. 2005). With opencanopy
conditions after Eastern Hemlock mortality, invasion by exotic plants has
already been reported in DWGNRA (Eschtruth and Battles 2009), and it is unclear
if invasive species will persist and for how long. Management treatments have
generally reduced target exotic plant species in localized projects on NPS lands,
but park-wide effectiveness remains uncertain as does long-term effectiveness
at controlling multiple invasions (Abella 2014). Eschtruth and Battles (2008) reported
that intensive herbivory by White-Tailed Deer has significantly influenced
transitions in post-HWA forests in DWGNRA, where Deer densities exceed 20
animals/km2 . Reduced White-Tailed Deer herbivory may correspond with increased
opportunity for tree recruitment and establishment of understory plants,
especially of plant species most susceptible to herbivory.
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Many of the tree species colonizing former Eastern Hemlock sites are deciduous
trees not thought to provide the same functions as Eastern Hemlock, such as yearround
amelioration of microclimates (Evans et al. 2011). In some areas such as the
southern Appalachians, the conifer Eastern White Pine might be the closest match
to Eastern Hemlock in terms of distribution and ecological traits (Vose et al. 2013).
Consideration could be given to promoting Eastern White Pine or other native trees
through planting or other means. This approach, however, poses a dilemma, because
a first step in dealing with HWA-impacted areas might ideally be to monitor
what plant species colonize former hemlock sites to evaluate whether management
intervention is needed or desirable. However, taking the time to monitor sites prior
to implementing managment could result in lost opportunities as the initial years
during and after hemlock mortality might afford the most amenable window to
establish desired vegetation. Experimentation and adaptive management would
likely be needed to identify optimal techniques and species for assisting tree establishment.
Eastern Hemlock habitats are specialized, and sites formerly vegetated
by Hemlock will have unique soil properties such as thick, acidic O horizons that
could affect plant recruitment (Finzi et al. 1998).
Some adaptation strategies discussed in Vose et al. (2013), primarily for non-
NPS lands, illustrate some potential similarities and differences for application
on NPS lands. For example, Vose et al. (2013) discussed introducing exotic tree
species, such as those from HWA’s native range that are resistant to HWA, at least
within their native range, to attempt to mimic functions provided by Eastern Hemlock.
The authors acknowledged that public acceptance of intentionally introducing
(more) exotic species to public lands is uncertain, and it should also be noted that
introducing exotic trees might interfere with colonization by native trees and may
not be legally authorized. While a range of novel management options may warrant
consideration (Vose et al. 2013), NPS (2006) management policy limits intentional
introduction of exotic species to NPS lands and recommends promoting native species
whenever possible. Other suggestions of Vose et al. (2013), such as collecting
Eastern Hemlock genetic material for potential resistance breeding programs, or
conserving remaining native species, would have greater congruence with NPS
policy than introducing exotic trees (NPS 2006). Another management option discussed
by Vose et al. (2013) was controlling aggressive native colonizers such as
Great Rhododendron in post-HWA sites. NPS (2006) policy allows control of native
species when natural processes have been interrupted. Such control measures
are being taken in some parks where native trees have increased in density during
a period in which natural fires have been suppressed. These woody species are being
mechanically thinned so surface fire can be safely reintroduced (e.g., Teraoka
2012). Although Great Rhododendron can suppress tree-seedling establishment in
the southern Appalachians (Ford et al. 2012), and thinning it may enhance recruitment
of other species, the benefit of controlling Great Rhododendron on NPS lands
is debatable because this species could perform some of the functions of Eastern
Hemlock such as providing evergreen cover and creating a shaded, sparsely vegetated
forest floor (Roberts et al. 2009).
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Leverage existing National Park Service programs
The NPS has several programs that could provide support to parks managing
HWA and post-infestation conditions. The Exotic Plant Management Team
program helps parks conduct treatments for exotic plants (Fraley et al. 2007) and
may be a valuable resource for reducing exotic plant invasions already noted as
a concern in post-HWA forests (Eschtruth and Battles 2009). The NPS also has
a network of native-plant nurseries, housed at some parks or in collaboration
with organizations such as Natural Resources Conservation Service plant materials
centers, which could be sites for collecting and processing Eastern Hemlock
genetic material and growing native plants for facilitated establishment on HWA
sites. The Integrated Pest Management Program could provide numerous services,
such as supporting scientific evaluations of specific biocontrol agents that
would inform development of park management policies (e.g., Havill et al. 2011)
and help parks balance tradeoffs resulting from a range of management alternatives
representing difficult decisions (Vose et al. 2013). The Vegetation Inventory
Program establishes vegetation-sampling plots across parks to produce maps of
current vegetation and could represent a major source of reference information
on distribution and composition of Eastern Hemlock Forest communities (Jenkins
2007). Moreover, the inventory plots could become long-term monitoring plots.
To enhance utility, it is important to follow key vegetation-sampling principles
such as using consistent sampling-unit sizes across sites to allow comparisons
and employing unbiased sampling, which can be done in a stratified framework
to maximize use of limited resources (Peet and Roberts 2013). Much has been
learned from independent, park-created networks of monitoring plots such as at
DWGNRA where plots have been re-measured since 1993, providing models for
other parks and NPS-wide programs (Evans 2010). The Inventory and Monitoring
Program (Fancy et al. 2009) could address specific questions regarding vegetation
transitions in the aftermath of HWA and monitor effectiveness of management
treatments such as facilitated colonization of native species. Moreover, a key
priority as a foundation for management is inventorying and understanding organisms
that are strongly associated with Eastern Hemlock. The Climate Change
Program could predict potential impacts of introduced forest pe sts and future forest
composition in specific parks (Iverson et al. 2008). To account for influences
on forest composition, these scenarios must include multiple factors (e.g., other
introduced forest pests, herbivory, disruption of natural disturbance regimes such
as fire) that interact with, and sometimes overwhelm, climatic influences (Lovett
et al. 2006).
Conclusion
Management of HWA and its impacts on parks would likely benefit from increased
understanding of baselines for comparison, both for identication of Eastern
Hemlock-dependent species and for evaluating management alternatives. Baird
et al. (2007, 2009) identified Eastern Hemlock-associated arthropod species not
only unknown in GSMNP, but never described to science. This knowledge gap is
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sobering in that even parks such as GSMNP visited by over 9 million people annually
could lose Hemlock-associated species before the species have even been
discovered. It may be difficult to develop adaptation strategies without understanding
the degree of dependence of organisms on Eastern Hemlock, their capacity to
adapt to using other habitat types when Hemlock dies, or their ability to persist
with chronic exposure to treatment regimes aimed at maintaining Hemlock. When
evaluating management options, it is useful to use potential replacement forests
(often deciduous forests) as a likely outcome of some management strategies and
assess that outcome as a basis for comparision with management actions that seek
to preserve intact Eastern Hemlock forests, which are unlikely to be maintained
in many areas (Roberts et al. 2009). For example, potential non-target effects of
chemical treatments for HWA should probably be compared against a baseline that
includes a forest without Eastern Hemlock (Falcone and DeWald 2010).
Potential for introduced forest-damaging agents to singly influence and interact
with other factors to affect future forest composition of parks warrants increased
attention. Lovett et al. (2006) concluded that forest destruction by introduced forest
pests might overwhelm influences of climate change over coming decades. It is
important to recognize that introduced pests had already dramatically impacted forests
in national parks well before recent decades of contemporary climate change,
most notably, the iconic extirpation of American Chestnut from introduced blight,
which was complete in most areas by the 1950s (Ellison et al. 2005). As a result of
American Chestnut loss, dramatically different forests exist to potentially adapt to
future climates. American Chestnut is no longer present as a dominant overstory
tree to experience today’s climate or a future climate, a situation that will likely be
repeated with the loss of Eastern Hemlock. Climate change is recognized as a major
factor with the potential to shape ecosystems over the next several decades (Shafer
2012), but current science suggests that rather than being a single, causative agent,
it is one of numerous, interacting factors that have shaped and will likely continue
shaping park ecosystems (Eschtruth and Battles 2009, Lovett et al. 2006, Orwig
et al. 2012). The primary current threat to Eastern Hemlock resources in national
parks is an introduced insect species, which is shaping ecosystem trajectories perhaps
for several future centuries.
Acknowledgments
This paper benefitted from discussions with NPS managers leading or involved with
efforts to manage HWA: Richard Evans of DWGNRA, and Kristine Johnson and Jesse
Webster of GSMNP. Richard Baird (Mississippi State University, Mississippi State, MS)
shared ideas regarding identifying Eastern Hemlock-associated species. I also thank Sharon
Altman (University of Nevada-Las Vegas, Las Vegas, NV) for formatting figures; and
Sharon Altman, Peter Budde (NPS), and two anonymous reviewers for providing helpful
comments on the manuscript. This manuscript is a contribution of the Biological Resource
Management Division, NPS, of the US Government. Any use of trade names is for descriptive
purposes only and does not imply endorsement by the US Government.
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2014 Vol. 13, Special Issue 6
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