Vol. 13, Special Issue 6
Community Responses to Eastern Hemlock Loss Across a
Relena R. Ribbons*
Abstract - Tsuga canadensis (Eastern Hemlock) forests are experiencing widespread
mortality due to the invasive insect Adelges tsugae (Hemlock Woolly Adelgid). This study
sought to document the community response to HWA across latitude. I selected two sensitive
response groups—plants and ants— to use as biological indicators of ecosystem change
to monitor differences along a natural gradient of Eastern Hemlock mortality among 3
forest types: relatively healthy Eastern Hemlocks, dead or dying Eastern Hemlocks, and
hardwood stands. I sampled understory vegetation, ants, and soils at each site and compared
sites using a linear mixed-model to discern the best predictors of species density. I
also compared analyses of variances across forest types among response variables. There
was an average two-fold increase in understory vegetation species density between Eastern
Hemlock and hardwood forests; ant species density was not influenced by forest type.
Analysis of variance comparisons for understory vegetation showed that forest type affects
understory vegetation, a result which was attributable to differences in a few dominant
plant species. The linear mixed-model showed that Eastern Hemlock density and latitude
were important predictors for both ant and vegetation species densities; soil pH and stand
density were predictors for vegetation species density, and litter depth was a predictor for
ant species density. My findings show that large structural changes in Eastern Hemlock forest
communities (induced by the effects HWA) alter a foundation ecosystem by shifting the
composition of understory plant communities, but not ant communities.
There is evidence that global changes, such as the spread of invasive species,
are leading to large transitions in both structure and function in ecosystems (Chapin
et al. 2000, Ribbons 2014). A major question in ecology is how these changes will
influence community structure and ecosystem function. Forests serve as important
buffers for climate change; however, loss of dominant species can significantly impact
their buffering capacity (Whitehead 2011). Ecosystem transitions are increasing
under global pressures such as climate change and the spread of non-native
invasive pests (Estes et al. 2011), and there is uncertainty about the factors that
promote ecosystem resilience to repeated disturbances (Loreau et al. 2001). Studying
forests in transition can provide insight into how altered overstory vegetation
influences understory arthropod communities, including forest-floor ants, which
are known to be ecosystem engineers (Del Toro et al. 2012).
*University of Tennessee - Ecology and Evolutionary Biology Department, Knoxville,
Knoxville, TN 37996. Current address - School of Environment, Natural Resources, and
Geography, Bangor University, Bangor, Gwynedd, Wales LL57 2UW, UK; rribbons@
Manuscript Editor: JoVonn Hill
Forest Impacts and Ecosystem Effects of the Hemlock Woolly Adelgid in the Eastern US
2014 Southeastern Naturalist 13(Special Issue 6):88–103
2014 Vol. 13, Special Issue 6
Foundation species are those that define community structure by creating locally
stable conditions exploited by other species; they modulate and stabilize
fundamental ecosystem processes (Ellison et al. 2005). In the northern part of
its range, Tsuga canadensis (L.) Carrière (Eastern Hemlock, hereafter, Hemlock)
tends to be a dominant forest species. In the southern portion of its range,
Hemlocks occur most frequently in mixed hardwood stands, and are especially
important in riparian environments (Krapfl et al. 2011, Nuckolls et al. 2009).
Infestation by an introduced invasive pest, Adelges tsugae Annand (Hemlock
Woolly Adelgid [HWA]), has decimated southern hemlock forests (Kincaid 2007,
Kincaid and Parker 2008, Nuckolls et al. 2009) and has reached as far north as
Massachusetts (Orwig and Foster 1998, Orwig et al. 2012). Increases in soil temperature,
chemistry, and nutrient fluxes have been documented following Hemlock
decline (Cobb et al. 2006, Jenkins et al. 1999, Kizlinski et al. 2002, Orwig
et al. 2008, Stadler et al. 2006). The loss of Hemlocks has already led to changes
in forest hydrology by increasing stream temperature and soil pH (Evans et al.
2011, Ford and Vose 2007, Knoepp et al. 2011, Martin and Goebel 2013), and
other functional and structural changes, such as shifts in microbial communities,
are likely to follow as forests transition to hardwood-dominated stands. Although
the direct effects of dying Hemlock forests are well documented for plant species
associated with the forests (Orwig et al. 2012), the cascading effects of Hemlock
decline, including changes in ant communities and soil physical properties, have
not been explored in a single study; some effects have been explored independently.
For example, changes in ant, beetle, and spider community compositions have
been documented in Hemlock forests infested with HWA compared with logged
Hemlock and reference Hemlock forests (Sackett et al. 2011).
Terrestrial invertebrates, including ants, play important roles in mediating ecosystem
processes such as decomposition, and they directly influence soil microbial
and plant communities and the nutrient processes they regulate. Ants are seed
dispersers and soil bioturbators (Del Toro et al. 2012, Folgarait 1998) and affect
terrestrial vegetation structure and arthropod communities (Holldobler and Wilson
1990, Zelikova et al. 2008), but their influence on ecosystem functions has not been
well-studied (Sackett et al. 2011). The effects of dying Hemlock forests on invertebrates
are well documented at the guild or generic level on a local scale (Dilling
et al. 2007, Ingwell et al. 2012, Rohr et al. 2009); however, the cascading effects of
Hemlock decline have not been explored across a regional latitudinal gradient.
Linking the large-scale disturbances of Hemlock mortality with understory
plant and ant communities may provide vital information on the future trajectory
of these forests, with insights into forest-floor regeneration and ecosystem health.
One group of forest ants, the Aphaenogaster rudis complex (thread-waisted ants),
are common in Hemlock forests, and are known to disperse up to 77% of all forest
plant seeds (Sackett et al. 2011); as such, they directly influence plant community
dynamics and are organisms worthy of study. This study sought to quantify the
abundance and distribution of understory plants and leaf-litter ants in low-mortality
Hemlock, high-mortality Hemlock, and hardwood forests. I used a space-for-time
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substitution of Hemlock loss by comparing the three forest types, as designed by Ellison
et al. (2010) for the Hemlock Removal Experiment. I also aimed to determine
compositional differences across a latitudinal and mortality gradient.
Field Site Description
To examine patterns in community responses to the loss of Hemlocks, I sampled
9 sites throughout the species’ range, and across the HWA invasion range (Fig. 1;
Supplemental Table 1 in Supplemental File 1, aavailable online at http://www.
eaglehill.us/SENAonline/suppl-files/s13-sp6-2001m-Ribbons-s1, and, for BioOne
subscribers, at http://dx.doi.org/10.1656/HA2001M.s1). The nine sites were (in
order from south to north): Smoky Mountains (SM), Frozen Head (FH), Fall
Creek (FC), Gauley River (GR), Rothrock (RR), Willington Hill (WH), MacLeish
(ML), Black Rock (BR), and Finger Lakes (FL). At each of these 9 sites, I established
three 100-m2 (0.01-ha) plots, 1 in each of 3 forest-cover types: uninfested
or low-mortality Hemlock, heavily HWA-infested or dead Hemlock, and mixedhardwoods,
henceforth referred to as Hemlock, dead Hemlock, and hardwood,
respectively; no stands had a completely open canopy. I identified to species all
overstory trees within the plots, recorded diameter at breast height (DBH) (trees
with DBH > 8 cm) for each, and counted and identified to species all tree saplings
Figure 1. Map of the distribution of counties with HWA as of 2011, with study locations
overlaid in black circles (courtesy of the US Forest Service).
2014 Vol. 13, Special Issue 6
(tree height < 1 m, DBH < 8 cm) within the plot. I compared basal area, density, and
relative importance values among forest types and across latitude using two-way
analysis of variance (ANOVA). Within each plot, I collected 9 soil samples for pH
analysis and recorded soil moisture and temperature. To examine understory diversity
across gradient, I collected data on plant and ant communities using 9 randomly
stratified 1-m2 subplots (n = 27 subplots/site; 243 subplots in total) evenly spaced
in checkerboard patter within a grid .
I tried to establish plots within a site at similar topographic position, elevation,
and aspect to reduce potential confounding effects between microhabitat features
and forest types (see Supplemental Table 1 in Supplemental File 1, available online
and, for BioOne subscribers, at http://dx.doi.org/10.1656/HA2001M.s1). All field
sampling occurred during daylight hours (8:00 AM until 6:00 PM). I sampled each
site once between May and July in 2012, starting in the south and traveling northward
to track phenology over the season.
To determine soil physical properties, I extracted 9 soil cores using a checkerboard
pattern of sampling locations within each plot using an AMS soil corer
(15-cm depth, 5-cm diameter; AMS Soil, Inc., American Falls, ID). I sifted soils
using a 2-mm-mesh sieve to homogenize the soil for pH analysis, following the
soil sampling protocol of Carter and Gregorich (2008). Within each of these plots,
I collected 2 soil-moisture measurements using a HydroSense monitor, and 2 temperature-
point measurements using a standard thermometer.
For canopy tree species, I calculated basal area and density of the overstory
and understory (saplings and seedlings) for each plot. I then calculated relative
importance values for each species at each site, as the sum of relative basal area
and relative density for each species. All herbaceous and woody vegetation within
the subplots was identified to species and assigned a percent groundcover using
a modified Braun-Blanquet scale ranging from 1–100% in 5% increment classes.
Nomenclature follows Gleason and Cronquist (1991). In addition to assigning a
percent-cover estimate, I identified and counted all tree seedlings located within
each subplot. I listed all species encountered within the 100-m2 plot to account
for species not contained within the understory vegetation subplots (see Supplemental
Table 2 in Supplemental File 1, available online at https://www.eaglehill.us/
SENAonline/suppl-files/s13-sp6-2001m-Ribbons-s1, and, for BioOne subscribers,
I selected 9 stratified 1-m2 subplots at random and collected leaf-litter from
them for Winkler extraction following the protocol of Agosti and Alonso (2000).
After collection, I took the samples to the lab and left them to extract for 72 h until
the litter was dry (Ivanov and Keiper 2009, Ivanov et al. 2010). After emptying
litter from the Winkler, I added any remaining ants to the extracted ant samples,
placed the ants into 70% ethanol, and sorted them to species-level. I deposited ant
voucher specimens at the Harvard Forest Long-Term Ecological Research Site
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I determined species density, richness, percent groundcover for vegetation, and
the incidence of occurrence (presence/absence) for ants across the latitudinal gradient.
To test for differences among forest types and across latitude, I compared
richness estimates and species density for understory vegetation and ants using
two-way ANOVAs of forest type, latitude, and richness. Not all plant or ant species
ranges extended throughout the entire gradient, and I expected a decline in species
richness as latitude increased. To account for this potential effect, I compared intrasite
species density among forest types, and constructed a linear mixed model to
test which variables were the best predictors of vegetation and ant species density.
I tested for interactions between ants and vegetation, local site factors, soil properties,
and temperature to determine the best predictors of species density.
To test for differences in community composition and dissimilarity among
forest types across the latitudinal gradient, I used non-metric multidimensional
scaling (NMDS). The EstimateS program (Colwell 1997) was used to estimate
ant abundance across the sites; this program also showed that although ants were
under-sampled at all sites, they were similarly under-sampled across all sites. Raw
abundance data for both vegetation and ants were fourth-root transformed prior to
multivariate analyses, and all distance-based metrics used a Bray-Curtis dissimilarity
matrix. I visually analyzed vegetation and ant communities at each site using
NMDS plots to determine initial clustering and dispersal among forest types at a site
(n = 9 sites), with 95% confidence intervals around clouds of points to determine
differences among forest types. For clarity of results I used binary-transformed data
(presence/absence) for final NMDS graphs. NMDS were constructed in Primer 6
version 6.1.13 (Primer-E, Ltd.; see Anderson 2001, McArdle and Andersen 2001).
To test which variables were the best predictors of vegetation richness and ant
species density, I used a linear mixed model approach in the lme4 package (Bates
et al. 2012) in R statistical program (version 2.15.0; R Core Team 2013).
Total basal area and stand density did not differ significantly across latitude or
among forest types (Table 1); results of ANOVAs showed mean Hemlock basal
area and density differed significantly among forest types but not across latitude
(Table 2). Mean soil temperature, soil volumetric water content (VWC), litter
depth, and soil pH did not differ among forest types (Table 3). Vegetation percent
Table 1. Stand structure mean (standard deviation) values for basal area (m2/ha, BA), density (stems/
ha, Density) of all overstory tree species for each forest type, and Hemlock basal area (BA), relative
importance value (IV), and density (Density) for each forest ty pe.
Canopy BA Density BA IV Density
Dead hemlock 205.33 (98.08) 244.44 (48.76) 120.56 (68.83) 61.22 (24.60) 142.67 (36.40)
Hardwood 188.22 (80.27) 296.67 (109.10) 8.33 (16.99) 5.78 (12.22) 16.44 (27.54)
Hemlock 211.22 (86.82) 271.89 (94.18) 177.67 (91.42) 83.22 (13.80) 192.22 (66.10)
2014 Vol. 13, Special Issue 6
groundcover (similar to abundance) generally decreased as latitude increased along
the gradient, with the exception of hardwood sites, which showed increases in
groundcover (Fig. 2). Ant abundance remained fairly consistent along the gradient
except for the northernmost site which had lower ant abundance, with no significant
differences observed between forest types (Fig. 2). Vegetation and ant species richness
trends were similar to percent groundcover and abundance t rends (Fig. 3).
A total of 87 plant species and 29 ant species were collected across the gradient
(see Supplemental Tables 2, 3 in Supplemental File 1, available online at http://
www.eaglehill.us/SENAonline/suppl-files/s13-sp6-2001m-Ribbons-s1, and, for
BioOne subscribers, at http://dx.doi.org/10.1656/HA2001M.s1). Species density of
the plant communities differed significantly across latitudes (F = 7.87, P = 0.009)
and among forest types (F = 3.97, P = 0.03) (Table 2), and a post-hoc Tukey’s honest
significant differences test showed the greatest difference of all pairs between
Hemlock and hardwood plots (M = 7.33, P = 0.02). Vegetation species density
Table 3. Mean soil physical and chemical properties for each forest type, volumetric water content
(VWC), temperature, litter depth, and soil pH.
Canopy Soil VWC Soil temp (C) Litter depth (cm) Soil pH
Dead hemlock 15.81 (9.37) 16.43 (9.37) 1.77 (0.63) 4.41 (0.56)
Hardwood 13.11 (8.14) 16.73 (8.14) 1.79 (0.50) 4.52 (0.59)
Hemlock 10.15 (2.56) 15.86 (2.56) 1.89 (0.69) 4.45 (0.68)
Table 2. Analysis of variances table including degrees of freedom (df), F-values, and associated Pvalues
(α = 0.05, with * indicating a significant P-value) for forest community stand dynamics including
total basal area, stand density, Hemlock basal area, Hemlock density, vegetation species density,
and ant species density across latitude and among forest types.
Source DF F P
Total basal area
Latitude 1 0.350 0.56
Forest types 2 0.159 0.85
Hemlock basal area
Latitude 1 0.022 0.82
Forest types 2 15.110 0.001*
Total stand density
Latitude 1 0.890 0.35
Forest types 2 0.790 0.46
Latitude 1 0.744 0.39
Forest types 2 34.000 0.001*
Vegetation species density
Latitude 1 7.870 0.009*
Forest types 2 3.970 0.03*
Ant species density
Latitude 1 12.862 0.001*
Forest types 2 0.853 0.44
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generally was highest in hardwood forests in the far northern sites (WH, ML and
BR) and far southern sites (SM, FH, and FC), followed by dead Hemlock and Hemlock
forests; however the response was more mixed at the intermediate sites (GR,
RR and FL). A total of 3648 ants were collected using Winkler extractors. Ant species
density differed significantly across latitude (F = 12.862, P = 0.001), but not
among forest types (F = 0.853, P = 0.44) (Table 2).
The following predictor variables from each plot were put into the linear mixed
model: temperature (mean annual temperature as recorded by the closest weather
station), canopy (forest type), total basal area, total stand density, Hemlock density,
soil VWC, soil temperature, litter depth, and soil pH. Additional variables including
elevation, latitude, longitude, Hemlock basal area, and importance values were
measured at the sites, but not included in the model due to a high correlation with
at least one of the variables already in the model. To determine the best model for
predicting richness, I used AIC stepwise model selection in the MASS package
(Venables and Ripley 2002). For vegetation richness, the best model was predicted
by stand density + Hemlock density + soil temperature + soil pH + canopy type.
For ant richness, the best model was predicted by Hemlock density + temperature
+ litter + canopy type.
Figure 2. Percent groundcover of understory vegetation (a) and Mean abundance of ants (b)
across the 3 forest-cover types (black = dead hemlock, white = hardwood, and blue = healthy
Hemlock), with sites arranged from south to north (left to right) along the x-axis.
Figure 3. Mean richness of understory vegetation (a) and ants (b) across the 3 forest-cover
types (black = dead hemlock, white = hardwood, and blue = healthy Hemlock), with sites
arranged from south to north (left to right) along the x-axis.
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Differences in forest community structure due to hemlock loss
The loss of Hemlock in forests leads to changes in vegetation and some ant
communities (Figs. 4, 5); however, the effects are reflected associated with Hemlock
loss differ among forest types representing a gradient of Hemlock mortality.
I documented an overall decline in basal area, density and relative importance of
Hemlock across the latitudinal gradient, with highest losses at the southern sites,
consistent with previous studies (Kincaid 2007, Kincaid and Parker 2008). At many
sites, Hemlock remained within the understory; however, similar to the findings of
Orwig et al. (2008), regeneration was low at my sites, as represented by seedlings
Figure 4. Non-metric multidimensional scaling plots (NMDS) for ants across all 9 sites,
with clusters of points from the same forest type indicating distinct communities compared
with dispersal of points across all forest types (which indicate no relationship between ants
and forest types). Data were binary transformed, and ellipses around forest treatments indicate
a 95% confidence interval.
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within hardwood and dead Hemlock stands. Across the gradient, most sites contained
abundant Acer rubrum L. (Red Maple) seedlings, which were especially
prominent in dead Hemlock stands, suggesting that this species will largely replace
Hemlocks. This finding is consistent with reports for other mixed hardwood forests
across the eastern US (Abrams 1998). Typically, structural changes in overstory
vegetation and a shift to hardwood-dominated forests is first reflected in the understory
(Mahan et al. 2004, Orwig et al. 2002).
Individual species responses vary by forest type for vegetation, but not ants
Similar to other studies between forest types (Hill et al. 2008), species density
and plant-community composition varied among forest types across the gradient,
but I observed less variation in ant communities, which suggests that ants may
Figure 5. Non-metric multidimensional scaling analysis for understory vegetation across
all 9 sites, with clusters of points from the same forest type indicating distinct communities
compared with dispersal of points across all forest types. Data were binary transformed, and
ellipses around forest treatments indicate a 95% confidence inte rval.
2014 Vol. 13, Special Issue 6
respond less to forest type and are more sensitive to open versus closed habitats
(Del Toro 2013). Red Maple was present at all sites, but I found the greatest number
of Red Maple seedlings and the highest percent groundcover in the dead Hemlock
plots. A few plant species including Aralia nudicaulis L. (Wild Sarsaparilla), Hammamelis
virginiana L. (Witch Hazel), and Lindera benzoin L. (Blume) (Spicebush)
were only detected in the Hemlock and dead Hemlock forests, suggesting that these
plants will decrease in abundance in the future. Some sites had vegetation species
gains or losses characterized by forest types. For example, at site SM, Mitchella
repens L. (Partridgeberry) was absent in hardwoods, but was present with high
abundance in Hemlocks. Conversely, at site FH, Hexastylis arifolia (Michx.) (Little
Brown Jug; an ant-dispersed plant) was absent in Hemlocks, and highest in abundance
in hardwood stands. Witch Hazel and Partridgeberry were fairly widespread
across numerous habitats in eastern North American forests, including those not
sampled in this study, and as such, are not species of concern (NRCS 2014).
Disturbed environments are ideal settings for new invasions of non-native plant
species (Wardle et al. 2011), and increased species richness following Hemlock
mortality does not necessarily indicate increased native species diversity or serve
as an indicator of increased ecosystem health. This pattern of change has implications
for the long-term resilience and stability for forests with declining Hemlock
because the replacing stand may contain more non-native or invasive plant species.
I observed 4 non-native invasive species across the gradient of sites, including Berberis
thunbergii DC (Japanese Barberry), Microstegium vimineum (Trin.) A. Camus
(Japanese Stiltgrass), Lonicera morrowii A. Gray (Morrow’s Honeysuckle), and
Euonymus alata (Thun.) Siebold (Burning Bush). Studies in the Delaware Water
Gap National Recreation Area (Eschtruth et al. 2006) have documented the influx
of invasive non-native plant species that were not found in initial 1993 plot surveys,
but which occured in 35% of permanent plots 10 years later, including Ailanthus
altissima (Mill.) Swingle (Tree-of-Heaven), Alliaria petiolata (M. Bieb.) Cavara &
Grande (Garlic Mustard), Japanese Barberry, Japanese Stiltgrass, and Rosa multiflora
Thunb. (Multiflora Rose). As the spread of invasive plant species continues,
highly disturbed sites such as declining Hemlock stands appear to provide ideal
habitat for new populations to establish. Long-term studies and repeated sampling
efforts are needed to determine if invasive plant species colonize sites with high
Hemlock mortality. It has been shown that invasive pests that co-occur with HWA,
such as Fiorinia externa Ferris (Elongate Hemlock Scale) pose a threat to Hemlocks
(Preisser et al. 2008) by altering regeneration (Preisser et al. 2011), growth,
and foliar chemistry (Miller-Pierce et al. 2010).
Examination of ants provides insight into how structural changes in forest
vegetation influence insects which mediate important ecosystem services such as
nutrient cycling, seed dispersal, and decomposition (Del Toro et al. 2012, Folgarait
1998). Similar to vegetation, ant community composition and species density
respond strongly to hemlock loss, with some species increasing in frequency and
species density in declining hemlock forests; however, this correlation was not
consistent across the gradient of sites. This result suggests the relationship between
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changes in vegetation and ants is not a direct linear relationship, but a more
complicated interaction. For example, some ants are either favored by or prefer
the conditions of disturbed hemlock forests, such as Prenolepis imparis (Say),
Stigmatomma pallipes (Haldeman), and Temnothorax longispinosus (Roger);
however, with a more thorough ant sampling, this trend may not hold true as some
of these species typically nest in trees (T. longispinosus). Some ant species, such as
Aphaenogaster picea (Wheeler) and Stenamma schmittii (Wheeler) are resilient to
changes in forest ground cover or overstory and persist in similar abundance in both
healthy and declining hemlock forests. Several species preferred the hardwood
forests, including Lasius claviger (Roger), L. umbratus (Nylander), L. nearcticus
(Wheeler), and Ponera pennsylvanica (Buckley) (although P. pennsylvanica is
found in a wide variety of habitats throughout the region). One species was equally
common across hardwood and hemlock forest types but much less abundant in the
dead hemlock forests, Myrmica punctiventris (Roger), which I suggest is due to
either a difference in environmental conditions or competitive exclusion by moredominant
ant species once the forest habitat is disturbed (see Del Toro et al. 2013).
Although I observed marked declines in the abundance of Aphaenogaster rudis, as
hemlocks are removed from the overstory, this trend was not significant; however,
more sampling might have yielded a stronger trend. I postulate this phenomenon
might be caused by competitive exclusion of behaviorally sub-dominant ants in the
A. rudis complex when large-scale disturbances occur (Del Toro et al. 2013), since
their abundance decreases in the disturbed (dead hemlock) sites.
In this study, neither plant community composition nor species density can be
used as surrogates to consistently predict ant community composition or species
density, which is in agreement with Hill et al.’s conclusions (2008). This finding
contrasts with a study documenting vegetation as the best indicator for non-ant arthropod
richness (Schaffers et al. 2008), suggesting the vegetation–ant relationship
is decoupled by large-scale, stand-replacing Hemlock decline. Thus, the observed
differences between ant and plant species richness patterns among forest types
across the latitudinal gradient may be due to the disturbed nature of the forest
as Hemlocks declined, rather than by underlying differences in forest types. It is
possible that as canopy closure increases over time, that vegetation may become a
better predictor of arthropod richness.
Community composition differences vary among forest type and across latitude
Research has documented community and ecosystem responses to HWA at local
and regional scales (Cobb et al. 2006, Evans et al. 2011, Nuckolls et al. 2009, Orwig
and Foster 1998, Orwig et al. 2013, Stadler et al. 2005); however, a regional-scale
examination of the interactions between community responses to Hemlock loss has
been lacking. Although species diversity changes in vegetation communities across
the gradient I studied, I observed no statistical differences in ant communities
composition across the gradient. This finding suggests that ants are more resilient
to the physical changes in forest structure as Hemlock declines, and that Hemlock
species do not explicitly control ant fauna. Based on the ANOVA results, latitude,
2014 Vol. 13, Special Issue 6
a surrogate for the effects of climate, was the strongest predictor of differences
for ant species density. Forest type was important at local scales at some sites, but
not others. Both forest type and latitude were strong predictors for differences in
vegetation species density. In both models for ant and vegetation species density,
latitude, canopy type, and hemlock density were important parameters. For plants,
total stand density and soil pH were also important predictors of species density,
whereas litter depth was an important additional predictor for ants. These results
suggest that the loss of acidic needle inputs into the litter layer and increased standstocking
levels by newly recruited tree saplings as overstory Hemlocks die, are both
important factors that influence understory plant communities.
While patterns in species density were consistent with my predictions and
those observed in other studies (D’Amato et al. 2009), these trends varied across
latitude suggesting a non-uniform response to the loss of Hemlocks on a regional
scale. Contrary to my predictions, ants demonstrated variable responses in species
density, with no consistent trend among forest type, but a direct relationship with
latitude. My results suggest that plant species are more sensitive to changes in overstory
vegetation than ant communities, which may require more extreme habitat
alterations before a response is observed, e.g., a shift from a closed-canopy forest
to an open-canopy habitat. There are several caveats when considering this result
in the larger context of community and invasion ecology: 1) dispersal limitation
could be an important factor (Caspersen and Saprunoff 2005, Clark et al. 1999); 2)
Hemlock forests are species-poor with strong top-down controls over soil-chemical
properties such as pH, compared with other temperate forest ecosystems; and 3) the
invasive HWA is a host-specific pest, so although response patterns were similar
among Hemlock stands across the gradient, more generalist invasive pests may
elicit different responses in forest communities.
The loss of Hemlocks will alter the forest landscape in the eastern US and
change the trajectory of these forests in the future. Many forested ecosystems maintain
a legacy effect of Hemlocks, as noted by sharp contrasts between highly acidic
soil and greater litter depth at formerly Hemlock-dominated sites and the higher
pH soils and shallower litter layer in hardwood stands without previous hemlock
composition. In this study, I found that vegetation within the Hemlock community
was immediately influenced by changes in canopy light and the influx of novel organisms
at sites with Hemlock mortality. Ants appeared to be resilient to changes in
overstory and understory vegetation, a finding consistent with previous studies (Del
Toro 2013); however there were marked declines in important seed-dispersing species
(A. rudis complex) in hardwood and transitioning Hemlock forests compared
with intact Hemlock forests. While these changes may not have immediate effects
on ecosystem health, the long-term trajectory of these forest communities may be
significantly altered by the loss of ecosystem engineers such as the seed-dispersing
ant species. I suggest that managers consider local site factors and land-use history
at specific forests, and the results of regional and local research to inform their
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management decisions. In this study, I demonstrated that invasive HWA-induced
structural changes in Hemlock forest communities led to consistent compositional
shifts in understory plant communities but not in ant communities, suggesting that
ants are not good bio-indicators in these highly disturbed Hemlock forests.
Thank you to Aimee Classen, Nathan Sanders, Aaron Ellison, and Israel Del Toro for
discussions on study design and analysis. I also thank Israel Del Toro, Kelsey Carter, Yvan
Delgado de la Flor, Alex Pfenningwerth, and Joan Ribbons for field assistance. Thank you
to Israel Del Toro and Aaron Ellison for ant-identification assistance and comments on earlier
versions of this manuscript. Work was undertaken with permission from the following
agencies: National Park Service (permit #GRSM-2012-SCI-1108, #GARI-2012-SCI-0002),
Black Rock Forest Consortium, Harvard Forest, MacLeish Field Station, Tennessee State
Parks, Pennsylvania Department of Conservation and Natural Resources, and the United
States Forest Service Finger Lakes National Forest. The author was supported by a UTKEEB
summer research grant and the Bredesen CIRE-ESE fellowship.
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