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2013 SOUTHEASTERN NATURALIST 12(2):353–366
Effects of Repeated-stand Entries on Terrestrial
Salamanders and their Habitat
Jessica A. Homyack1,2,* and Carola A. Haas1
Abstract - In recent years, silivicultural methods have shifted away from clearcut harvesting
towards greater retention of overstory trees through part or all of a rotation. However,
little is known about the effects of partial harvesting on wildlife populations. Thus, we
examined effects of high-leave shelterwood management on terrestrial salamanders prior
to and after an initial harvest and a subsequent overstory removal harvest (ORH) 13 years
later. On an experimental research site in southwestern Virginia, we compared changes in
salamander captures in this plot to a clearcut and control plot 1994–1996 and 2007–2009.
Compared to contemporaneous estimates from an unharvested control, salamander captures
were lower on shelterwood and clearcut plots 2-years after the initial harvest (1996)
and lower on the shelterwood plot 1- and 2-years after the ORH (2008, 2009). Captures
of the most common species, Plethodon cinereus (Eastern Red-backed Salamanders), followed
similar trends with fewer captures in both harvested plots 2-years after the initial
harvest (1996), but only the ORH differed from the control 2-years after the second partial
harvest (2009). Abundance of woody debris was greater in the shelterwood following
the ORH but was more decayed in the control plot. The regenerating clearcut (14 years
post-harvest) had deeper leaf litter and denser understory vegetation than the ORH. These
data are some of the first available describing effects of multiple harvest entries on terrestrial
salamanders and suggest cumulative negative impacts on salamanders may occur
from partial harvesting systems. More long-term monitoring of salamander populations
is justified in silvicultural systems with multiple entries with in a rotation.
Within many jurisdictions, forest harvesting has trended away from intensive
management (e.g., clearcutting) and towards greater retention of overstory trees
through all or part of the rotation (e.g., partial harvesting) (Fuller et al. 2004,
McWilliams et al. 2005, Siry 2002). Partial harvesting is a general term that
refers to forest stands in which multiple harvests are made during a single rotation
and some canopy trees remain for at least a portion of the rotation. From the
mid-1980s to the mid-1990s, clearcutting made up 45% and partial harvesting
accounted for 55% of the average annual harvest in the southern US (US Department
of Agriculture, Forest Servive 2010). The use of partial harvesting on such
a significant land area may have occurred in part because of the emphasis for land
managers to retain forest structure and biodiversity (Brunson and Shelby 1992,
Gillis 1990) and the more positive public perception of multi-canopied forests
compared to clearcuts (Bliss 2000, Brunson and Shelby 1992, Sedjo 1999). Thus,
1Department of Fish and Wildlife Conservation, Virginia Tech, Blacksburg, VA 24061.
2Current address - 1785 Weyerhaeuser Road, Vanceboro, NC 28586. *Corresponding
author - firstname.lastname@example.org.
354 Southeastern Naturalist Vol. 12, No. 2
silvicultural alternatives to clearcut harvesting are common across the southeastern
United States, yet knowledge regarding the influence of partial cutting, and
especially the cumulative effects of multiple stand-entries on wildlife populations
and their habitats is limited (Fuller et al. 2004, Homyack and Haas 2009,
McComb et al. 1993, Reichenbach and Sattler 2007).
At a landscape scale, partial harvests must be extended over a larger area to
produce the same amount of wood fiber per unit area as clearcut stands (Gillis
1990, Hagan 1996, Knapp et al. 2003), which may increase the effects of isolation
and fragmentation on wildlife populations (Morrison et al. 1992, Saunders
et al. 1991). At a stand scale, populations of species sensitive to disturbance may
not have had adequate time to return to preharvest abundances prior to subsequent
harvests in the rotation, and thus may face cumulative effects of multiple
harvest events. Alternatively, by retaining some characteristics of more mature
forest, partial harvests may have weaker effects on wildlife populations than
clearcutting (Fuller et al. 2004, McComb et al.1993, Semlitsch et al. 2009), but
these hypotheses have not been well studied. Finally, numerous structural habitat
elements important for wildlife at a sub-stand scale, such as amounts of coarse
woody debris and tree density and composition are altered by multiple stand entries
(Barg and Edmonds 1999, Lilieholm et al. 1990).
Terrestrial salamanders may be model organisms for examining effects of forest
harvesting on wildlife occurring at the soil-litter interface because they are physiologically
linked to microhabitat and microclimate features by their requirements
of cool and moist conditions for cutaneous respiration (Welsh and Droege 2001).
Additionally, terrestrial salamanders are long-lived, display low inter-annual
variation in abundances, are an apex predator in the leaf litter, are one of the most
abundant vertebrates in forested systems, and reach their highest levels of species
diversity in the central and southern Appalachians (Burton and Likens 1975,
Petranka 1998, Walton 2005, Walton and Steckler 2005, Welsh and Droege 2001,
Wyman 1998). Thus, salamanders have ecological characteristics that make them
useful for understanding broader effects of silvicultural practices on wildlife.
Across forested systems in North America, research has consistently demonstrated
that forest harvesting can have persistent negative effects on abundances
of terrestrial salamanders (deMaynadier and Hunter 1995, Dupuis et al. 1995,
Homyack and Haas 2009). In a meta-analysis of 14 studies, untreated control
stands had 5 times greater abundances of plethodontid salamanders than clearcut
forest stands (deMaynadier and Hunter 1995). Although a greater focus has been
on short-term (≤2 years) effects of clearcut harvesting on salamanders, available
information indicates that a wide range of forest practices that remove canopy
trees can have negative and lasting effects on abundances (Ash 1997, deMaynadier
and Hunter 1995, Homyack and Haas 2009, Petranka et al. 1993, Pough et
al. 1987, Tilghman et al. 2012). Despite that multiple harvests in a stand during
a single rotation are common for many silvicultural systems, effects from >1
harvest entry on terrestrial salamanders or their specific habitat components have
not been well-documented.
2013 J.A. Homyack and C.A. Haas 355
The goal of this investigation was to evaluate whether multiple harvest
entries in one type of partial harvest, a high-leave shelterwood system, had
cumulative negative effects on terrestrial salamanders in a central Appalachian
hardwood forest. We used a case-study approach to examine the effects
both pre- and post-treatment at an experimental research site in southwestern
Virginia. We predicted that captures of terrestrial salamanders would not have
recovered to pre-harvest numbers prior to the second stand entry and would
decline further after a second harvest. Specifically, we quantified (1) captures
of terrestrial salamanders 1-year prior to and 2-years following an initial
partial harvest and 1-year prior to and 2-years following a second overstory
removal harvest and within a similar clearcut harvest and untreated control
stand during the same time periods, and (2) within-stand habitat characteristics
important for terrestrial salamanders after the second entry. Although inferences
are limited by our case-study approach, this research expands the limited
knowledge of cumulative effects of multiple shelterwood entries on terrestrial
salamanders and their habitat.
Field Site Description
We compared the effects of several silvicultural treatments on terrestrial salamanders
on the Jefferson National Forest, Montgomery County, VA. The study
site (Blacksburg 1 [BB1]) is part of a long-term investigation of the effects of oak
regeneration methods on biodiversity, the Southern Appalachian Silviculture and
Biodiversity Project (SASAB), where terrestrial salamanders have been sampled
yearly since 1994 (Atwood et al. 2009, Belote et al. 2008, Homyack and Haas
2009). The study site was south-facing (153°), had a moderate slope (16%), had
no recent history of stand disturbance, and had uniform stocking of merchantable
trees (Wender 2000). Dominant overstory trees included Quercus spp. (oaks),
Liriodendron tulipifera L. (Yellow Poplar), Acer rubrum L. (Red Maple), and
Oxydendron arboretum L. (Sourwood) as well as small components of other
hardwoods and Pinus strobus L. (White Pine). Within the site, a single 2-ha plot
of each silvicultural treatment (control, shelterwood, and silvicultural clearcut)
were assigned randomly.
The initial treatments occurred during winter 1994–1995. For the silvicultural
clearcut (hereafter clearcut) plot, all stems >5 cm diameter at breast height
(DBH) were cut (pre-treatment basal area = 24 m2/ha; 1-year post-treatment basal
area = 4 m2/ha, 10-year post-treatment basal area = 7 m2/ha). Merchantable trees
were skidded and removed from the site. In the shelterwood plot, the overstory
was harvested in two entries to facilitate a cohort of advanced regeneration under
the partial canopy (Smith et al. 1997). Following the first harvest in 1994–1995,
17 m2/ha of the pre-treatment basal area of 27 m2/ha was retained. Residual stems
were dominant or co-dominant trees of preferred species (primarily oaks) with
DBH of 25–40 cm. During winter 2007–2008, residual overstory stems were harvested
with chainsaws and skidders, retaining 6 m2/ha of overstory basal area. No
356 Southeastern Naturalist Vol. 12, No. 2
treatments were applied to the control or clearcut plot during this time. During
the pre-, 1-year post-, and 10-year post-treatment time periods, the control plot
had 21, 21, and 22 m2/ha of overstory basal area, respectively.
We sampled terrestrial salamanders using night-time area-constrained
searches on rainy nights April–early June and September–October during
1994–1996 and 2007–2009. For the first sampling years, 1994 was the pretreatment
estimate and 1995 and 1996 were 1- and 2-years post-treatment.
During the 2007–2009 sampling window, 2007 represented the pre-ORH
sample and 2008 and 2009 were 1- and 2-years post-ORH, respectively. Prior
to the initial harvest, we established a 3 × 3 grid of sampling transects (n = 9
transects/plot) in each 2-ha plot, with each transect measuring 2 m × 15 m.
Transects were >30 m from plot edges and >30 m from each other. Each warm
rainy night, we randomly selected one transect from each plot, and 2–3 observers
hand-captured terrestrial salamanders active on the surface of transects. We
standardized our sampling to only occur nights (>1 hour after sunset) during or
after rain events when temperatures were >7 °C and the leaf litter was wet. We
rotated the order that treatment plots were sampled among nights, and transects
were only sampled once per sampling year. Salamanders were housed overnight
in a laboratory environment, and the next morning we confirmed species identification
and recorded morphological and reproductive data. We released all
salamanders at the point of capture within 24 hours. We did not individually
mark salamanders because toe-clipping may violate assumptions of tag loss
from regeneration of digits (Heatwole 1961) and may decrease survival (Davis
and Ovaska 2001, McCarthy and Paris 2004). Further, current marking technology
(e.g., Visible Implant Elastomer, Northwest Technologies, Shaw Island,
WA) was not readily available during the beginning of the study. We assumed
that counts of individuals were positively and linearly related to the true population
size and that detection did not differ across treatments, years, or species
(Mazerolle et al. 2007, Pollock et al. 2002, Reichenbach and Sattler 2007,
Welsh and Droege 2001). Knapp et al. (2003) and Homyack and Haas (2009)
provide additional details on salamander sampling methods.
Second, we a priori selected habitat characteristics that would be associated
with mediating microclimate, providing foraging habitat, or providing habitat
for brooding eggs for terrestrial salamanders. Habitat characteristics were measured
on all salamander sampling transects on each of the three treatment plots
in 2009, the second growing season after the ORH. To quantify coarse woody
debris (CWD), we counted root masses (≥7.6 cm diameter at the base), stumps
(<2 m height, ≥7.6 cm diameter), and logs (≥7.6 cm diameter, in contact with
ground) within each transect. We measured log diameters at both ends and the
length and converted it into a volume by using the formula for the volume of
the frustum of a cone (Volume = [1/3]π[r2 + rR + R2]h), where r = small radius,
2013 J.A. Homyack and C.A. Haas 357
R = large radius, and h = height. We calculated stump and root mass volume as
a cylinder with height and mid-point diameter. We evaluated decomposition
class of each piece of CWD and assigned it a value from 1–5 based on Maser et
al. (1979). We quantified density of trees (≥7.6 cm DBH, ≥1.5 m height, standing
at >45º from the ground) that occurred within each transect and density of
understory woody vegetation by counting the number of woody stems within
a plot that were >0.5 m tall, but <7.6 cm DBH. We measured leaf-litter depth
at six locations in a transect (centered at 2.0 m, 3.0 m, 7.0 m, 8.0 m, 12.0 m,
13.0 m) with a ruler held perpendicularly to the ground surface and averaged
them into a single value.
We determined whether captures of all salamanders or just Plethodon cinereus
Green (Eastern Red-backed Salamander) differed among treatments
after both an initial and a second harvest in the shelterwood plot with separate
Kruskal-Wallis tests comparing the control, shelterwood (or ORH), and clearcut
treatment plot for each year of sampling (Conover 1999). When tests were
significant (P < 0.05), we used a Bonferroni multiple comparison procedure to
evaluate which treatments differed. Secondly, we evaluated whether six a priori
selected habitat characteristics, including the volume, density, and decay class
of CWD, density of overstory trees, density of understory trees, and leaf-litter
depth differed among silvicultural treatments two-years after the ORH with
Kruskal-Wallis tests (Conover 1999). We used SAS 9.3 (SAS Institute, Cary,
NC) for analyses. Our analyses and thus the inferences made from our results
are limited by our case-study approach without site replication and use of transects
as the experimental unit. However, the random assignment of treatments
and long-term collection of data in our study are improvements over the traditional
before-after-control-impact (BACI) design, which often uses inferential
statistics such as t-tests and analysis of variance to describe the effects of a
stressor (Smith 2002).
We recorded 672 total terrestrial salamander captures across 33 sampling
nights. We sampled 4–7 sets of transects/year. Most (95%) salamanders
were Plethodon cinereus, but captures also included P. cylindraceus Harlan
(White-spotted Slimy Salamanders; 4% of captures) and <1% each of
Desmognathus fuscus Rafinesque (Northern Dusky Salamander), Eurycea
cirrigera Green (Southern Two-lined Salamander), and Gyrinophilus porphyriticus
Green (Spring Salamander) (Table 1).
Effects of shelterwood harvesting differed across years of post-harvest sampling
for total salamander captures. Total captures of terrestrial salamanders in
the control plot was similar to both the shelterwood and clearcut prior to the
initial harvest (1994: χ2 = 1.10, df = 2, P = 0.58) and the growing season after
358 Southeastern Naturalist Vol. 12, No. 2
treatment (1995: χ2 = 2.24, df = 2, P = 0.33). However, 2 years after the initial
harvest, there were 70–74% fewer salamander captures in the shelterwood and
clearcut plots compared to the control (1996: χ2 = 11.60, df = 2, P = 0.004)
Table 1. Number of salamander captures by species and year in a mature oak (control), silvicultural
clearcut, and shelterwood-harvested forest in southwestern Virginia, 1994–1996 and 2007–2009.
Plots were harvested in 1994–1995, and residual trees were removed in an overstory removal harvest
(ORH) in the shelterwood plot during winter 2007–2008.
Year Control Clearcut Shelterwood
Pre-treatment year 43 54 39
1-year post 59 52 45
2-years post 53 16 19
1-year pre-ORH 44 29 24
1-year post-ORH 55 21 13
2-years post-ORH 56 15 4
Total 310 187 144
Pre-treatment year 1 1 5
1-year post 7 2 3
2-years post 1 0 0
1-year pre-ORH 4 0 0
1-year post-ORH 1 0 0
2-years post-ORH 0 0 0
Total 14 3 8
Pre-treatment year 1 0 0
1-year post 0 0 0
2-years post 1 0 0
1-year pre-ORH 0 0 0
1-year post-ORH 0 0 0
2-years post-ORH 0 0 0
Total 2 0 0
Pre-treatment year 0 0 0
1-year post 1 0 0
2-years post 0 0 0
1-year pre-ORH 1 0 0
1-year post-ORH 1 0 0
2-years post-ORH 0 0 0
Total 3 0 0
Pre-treatment year 0 0 0
1-year post 1 0 0
2-years post 0 0 0
1-year pre-ORH 0 0 0
1-year post-ORH 0 0 0
2-years post-ORH 0 0 0
Total 1 0 0
Grand total 330 190 152
2013 J.A. Homyack and C.A. Haas 359
(Fig. 1). During the second sampling period in 2007–2009, salamander captures
did not differ significantly among treatments prior to the ORH (2007: χ2 = 2.86,
df = 2, P = 0.24), but were different both 1 year (2008: χ2 = 7.185, df = 2, P =
0.03) and 2 years (2009: χ2 = 9.10, df = 2, P = 0.01) after the second harvest of
the shelterwood plot (Fig. 1). The ORH had 58–76% fewer salamander captures
than either clearcut or control plots 1 year after treatment (2008: P < 0.05), but
at 2 years after treatment, was only significantly lower from the control (2009:
P < 0.05).
Mean captures of P. cinereus responded comparably, with similar number of
mean captures across treatments in the pre-treatment year (1994: χ2 = 1.23, df = 2,
P = 0.54) and 1 year after the initial harvest (1995: χ2 = 1.86, df = 2, P = 0.39).
The second year after harvest, the control plot had 2.8 and 3.3 times more captures
of P. cinereus than the shelterwood and clearcut plot, respectively (1996: χ2
= 11.07, df = 2, P = 0.004). During 2007–2009, captures of P. cinereus did not
differ the year prior to (2007: χ2 = 1.73, df = 2, P = 0.42) or 1 year after the ORH
(2008: χ2 = 5.31, df = 2, P = 0.07), but were different 2 years after the ORH
(2009: χ2 = 9.10, df = 2, P = 0.01). In 2009, captures of P. cinereus in the control
were greater than in the ORH (12.5 times greater, P < 0.05) but not the clearcut
plot (3.3 times greater, P > 0.05).
Structural habitat characteristics relevant to salamander life histories differed
among treatments after the ORH. The density (χ2 = 8.50, df = 2, P = 0.01) and
mean decay class (χ2 = 15.80, df = 2, P = 0.004), but not volume (χ2 = 4.73, df = 2,
Figure 1. Mean (SE) number of terrestrial salamander captures/30-m2 transect during
area-constrained night-time searches across three silvicultural treatments in southwestern
Virginia, 1994–2009. Different letters indicate among-treatment differences (P < 0.05).
360 Southeastern Naturalist Vol. 12, No. 2
P = 0.09), of CWD differed among treatments (Fig. 2). Although CWD density
in the control was only lower compared to the ORH, CWD in the control was
significantly more decomposed than either of the harvested treatments (P < 0.05)
(Fig. 2). Density of overstory trees was 1.5–2.8 times greater in the control than
in the harvested treatments (Fig. 2), but understory trees (<7.6 cm DBH) were
at a lower density (χ2 = 15.70, df = 2, P < 0.001; Fig. 2). Lastly, litter depth was
lowest in the ORH, but differed significantly only from the clearcut (χ2 = 8.86,
df = 2, P = 0.01; Fig. 2).
Figure 2. Forest structural characteristics in an untreated control, shelterwood, and
regenerating clearcut plot in southwestern Virginia. Forest structure was measured on
nine 2- × 15-m transects/plot in summer 2009, the second growing season following an
overstory removal harvest in the shelterwood plot and 14 years after the initial treatment
to the clearcut and shelterwood. Different letters indicate among-treatment differences
(P < 0.05).
2013 J.A. Homyack and C.A. Haas 361
In our study, both total captures of terrestrial salamanders and captures
of the most commonly encountered species responded negatively to both the
initial harvest and the ORH 13 years later. Salamander captures on the shelterwood
plot relative to the control declined in the second year after the initial
treatment, and declined again 1- and 2-years after the ORH. Other studies have
reported that partial harvesting has either negative effects (Brooks 2001, Grialou
et al. 2000, Homyack and Haas 2009, Knapp et al. 2003) or little effect
(Brooks 1999, Duguay and Wood 2002, Mitchell et al. 1996, Pough et al. 1987,
Reichenbach and Sattler 2007) on terrestrial salamanders. However, most prior
research has focused only on effects to salamanders from initial harvest entries.
In contrast, we quantified captures of salamanders through both the initial and
second harvest, and provided evidence that repeated stand entries can negatively
affect salamanders. At our experimentally manipulated study site, a second
harvest caused an additional decline in the numbers of salamanders captured
in the shelterwood as compared to the control and the 14–15 year-old clearcut.
The result that salamander captures after the ORH declined to levels as low as
those soon after the initial shelterwood is striking. Whether this negative effect
will have cumulative long-term impacts to salamander populations is currently
unknown as this study only examined salamander captures to two years
after the ORH. Unless salamander populations can recover more quickly after
the second harvest, partial harvest methods that require multiple stand entries
within a rotation, such as a group selection regime, could possibly permanently
depress salamander numbers (Homyack and Haas 2009).
Discrepancies among studies on the effects of partial harvesting on salamander
abundances may have resulted from variation in the type of partial harvest or
in basal-area retention (but see Tilghman et al. 2012). For example, after the
initial harvest in this study, the shelterwood plot retained 17 m2/ha of overstory
basal area. Other investigations of effects of partial harvesting on terrestrial
salamanders reported average residual basal areas of 9.1–18.3 m2/ha for a shelterwood
harvest (Reichenbach and Sattler 2007), 4–15 m2/ha for a range of
shelterwood management options (Harpole and Haas 1999), 6–14 m2/ha for a
range of partial harvesting options (Knapp et al. 2003), and 48 m2/ha for a light
thinning (Grialou et al. 2000), but many other researchers failed to report basal
area or specific harvesting type, hindering the interpretation of results across investigations
Besides residual basal area of harvested stands, other aspects of forest structure
likely influenced salamander numbers and communities after treatments. In
our case study, a priori selected habitat characteristics associated with forestdwelling
salamanders differed among an untreated control, regenerating clearcut,
and a recently harvested shelterwood plot. Abundance of CWD and leaf-litter
depth differed among control, clearcut, and post-ORH shelterwood plots. Not
surprisingly, the ORH tended to have more abundant woody debris than control
362 Southeastern Naturalist Vol. 12, No. 2
plots, likely due to addition of logging residues (Fig. 2b). Coarse woody debris
was decomposed further in the control compared to either harvesting treatment
(Fig. 2c), further indicating that logging slash was a primary source of downed
wood in treatment plots. Additionally, leaf-litter depth in the ORH was 55–68%
of that in control or clearcut plots, but understory tree density was greatest on
regenerating clearcut plots (Fig. 2e, 2f).
Changing the structure of mature forest from harvesting generally is perceived
to have negative effects on salamander populations. After harvesting, salamanders
are hypothesized to emigrate from disturbed areas (evacuation hypothesis),
retreat underground until conditions are more amenable (retreat hypothesis), or
die, either directly from harvesting equipment or indirectly from changes to habitat
(mortality hypothesis) (deMaynadier and Hunter 1995). Removal of overstory and
understory trees leads to less leaf litter on the ground, thus reducing available foraging
substrate and mediation of microclimate for salamanders. Further, opening
of the canopy layer from harvesting can cause increased ground temperatures and
decreased soil moisture, which may restrict movements, foraging opportunities,
and cutaneous respiration of terrestrial salamanders (Chen et al. 1999, Harpole
and Haas 1999, Jaeger 1980, Liechty et al. 1992) and increase energetic costs
(Homyack et al. 2011). Lastly, terrestrial salamanders rely on CWD for several
life-history requirements, including maintaining moisture and thermal balances,
access to mates and foraging opportunities, and substrates for brooding eggs (de-
Maynadier and Hunter 1995). Although harvesting produces large inputs of small
diameter logging slash, this small woody debris often decomposes and does not
persist through a rotation (Fraver et al. 2002, Spies et al. 1988).
In our study, woody debris from the clearcut harvest persisted through 14
years, so that density of CWD was similar to the control during this period.
The second growing season after the ORH, there were more individual pieces
of CWD in the shelterwood plot. However, because this CWD was not welldecayed
and total volumes were not increased significantly, slash may not have
been used by salamanders for foraging or brooding eggs. Thus, the large inputs
of logging slash after the shelterwood ORH may not have been sufficient to
overcome negative changes to microclimate or other life-history requirements
of terrestrial salamanders.
Although conclusions drawn from this case-study approach are limited to
our study area due to lack of replication, this experimental design included both
pre-treatment estimates and randomly applied treatments, both of which are uncommon
in investigations of forest harvesting on salamanders (deMaynadier and
Hunter 1995, Perkins and Hunter 2006, Reichenbach and Sattler 2007). Our exploratory
results indicate that silvicultural regimes that employ multiple entries
within a rotation have the potential to negatively affect terrestrial salamanders, at
least at our mixed hardwood sites and for the salamander community we examined.
Given that >60 years is expected to be required for salamander populations
in Appalachian oak forest to recover to pre-harvest levels of abundance from only
one stand entry (Homyack and Haas 2009), it is likely that silvicultural regimes
2013 J.A. Homyack and C.A. Haas 363
such as shelterwood systems that repeatedly reduce salamander populations will
require a longer period for population recovery, or may permanently suppress
populations (see discussion in Knapp et al. 2003). Forest managers will need to
weigh the consequences of partial harvests on biodiversity along with the higher
costs of harvesting, potentially negative effects on soil erosion due to multiple
stand entries within a rotation (Hood et al. 2002), and implications for landscapescale
effects when applying forest plans. Additional research should consider the
long-term effects of multiple harvest entries on relative abundances and demographics
of terrestrial salamanders on replicated study sites where both pre- and
post-harvest data are quantified.
This research was supported by United States Department of Agriculture-National
Research Initiative Grants to Haas et al. (9503196 and 2005-35101-15363) and an AdvanceVT
Doctoral Fellowship (SBE-0244916) provided to J. Homyack. We thank the
George Washington and Jefferson National Forests for logistical support, the numerous
field assistants for data collection, and the reviewers for impr oving this manuscript.
Ash, A.N. 1997. Disappearance and return of plethodontid salamanders to clearcut plots
in the southern Blue Ridge Mountains. Conservation Biology 11:983–989.
Atwood, C.J., T.R. Fox, and D.L. Loftis. 2009. Effects of alternative silviculture on
stump sprouting in the southern Appalachians. Forest Ecology and Management
Barg, A.K., and R.L. Edmonds. 1999. Influence of partial cutting on site microclimate,
soil nitrogen dynamics, and microbial biomass in Douglas-fir stands in western Washington.
Canadian Journal of Forest Research 29:705–713.
Belote, R.T., R.H. Jones, S.M. Hood, and B.W. Wender. 2008. Diversity-invasibility
across an experimental disturbance gradient in Appalachian forests. Ecology
Bliss, J.C. 2000. Public perceptions of clearcutting. Journal of Forestry 98:4–9.
Brooks, R.T. 1999. Residual effects of thinning and high white-tailed deer densities on
Northern Redback Salamanders in southern New England oak forests. Journal of
Wildlife Management 63:1172–1180.
Brooks, R.T. 2001. Effects of the removal of overstory hemlock from hemlock-dominated
forests on Eastern Redback Salamanders. Forest Ecology and Management
Brunson, M., and B. Shelby. 1992. Assessing recreational and scenic quality: How does
new forestry rate? Journal of Forestry 90: 37–41.
Burton, T.M., and G.E. Likens. 1975. Energy flow and nutrient cycling in salamander
poplations in the Hubbard Brook Experimental Forest, New Hampshire. Ecology
Chen, J., S.C. Saunders, T.R. Crow, R.J. Naiman, K.D. Brosofske, G.D. Mroz, B.L.
Brookshire, and J.F. Franklin. 1999. Microclimate in forest ecosystem and landscape
ecology. Bioscience 49:288–297.
Conover, W.J. 1999. Practical Nonparametric Statistics. John Wiley and Sons, New York,
NY. 584 pp.
364 Southeastern Naturalist Vol. 12, No. 2
Davis, T.M., and K. Ovaska. 2001. Individual recognition of amphibians: Effects of toe
clipping and fluorescent tagging on the salamander Plethodon vehiculum. Journal of
deMaynadier, P.G., and M.L. Hunter. 1995. The relationship between forest management
and amphibian ecology: A review of the North American literature. Environmental
Duguay, J.P., and P.B. Wood. 2002. Salamander abundance in regenerating forest stands
on the Monongahela National Forest, West Virginia. Forest Science 48:331–335.
Dupuis, L.A., J.N.M. Smith, and F.L. Bunnell. 1995. Relation of terrestrial-breeding
amphibian abundance to tree-stand age. Conservation Biology 9:645–653.
Fraver, S., R. Wagner, and M. Day. 2002. Dynamics of coarse woody debris following
gap harvesting in the Acadian forest of central Maine, USA. Canadian Journal of Forest
Fuller, A.K., D.J. Harrison, and H.J. Lachowski. 2004. Stand-scale effects of partial harvesting
and clearcutting on small mammals and forest structure. Forest Ecology and
Gillis, A.M. 1990. The new forestry. BioScience 40:558–562.
Grialou, J.A., S.D. West, and R.N. Wilkins. 2000. The effects of forest clearcut harvesting
and thinning on terrestrial salamanders. Journal of Wildlife Management 64:105–113.
Hagan, J.C. 1996. Clearcutting in Maine: Would somebody please ask the right question.
Maine Policy Review: July 1996.
Harpole, D.N., and C.A. Haas. 1999. Effects of seven silvicultural treatments on terrestrial
salamanders. Forest Ecology and Management 114: 349–356.
Heatwole, H. 1961. Inhibition of digital regeneration in salamanders and its use in marking
individuals for field studies. Ecology 42:593–594.
Homyack, J.A., and C.A. Haas. 2009. Long-term effects of experimental forest harvesting
on abundance and reproductive demography of terrestrial salamanders. Biological
Homyack, J.A., C.A. Haas, and W.A. Hopkins. 2011. Energetics of surface-active terrestrial
salamanders in experimentally harvested forest. Journal of Wildlife Management
Hood, S.M., S.M. Zedaker, W.M. Aust, and D.M. Smith. 2002. Predicted soil loss for
harvesting regimes in Appalachian hardwoods. Northern Journal of Applied Forestry
Jaeger, R.G. 1980. Microhabitats of a terrestrial forest salamander. Copeia 1980:
Knapp, S.M., C.A. Haas, D.N. Harpole, and R.L. Kirkpatrick. 2003. Initial effects of
clearcutting and alternative silvicultural practices on terrestrial salamander abundance.
Conservation Biology 17:752–762.
Liechty, H.O., M.J. Helmes, D.D. Reed, and G.D. Mroz. 1992. Changes in microclimate
after stand conversion in two northern hardwood stands. Forest Ecology and Management
Lilieholm, R.J., L.S. Davis, R.C. Heald, and S.P. Holmen. 1990. Effects of single-tree
selection harvests on stand structure, species composition, and understory tree growth
in a Sierra mixed conifer forest. Western Journal of Applied Forestry 52:43–47.
Maser, C., R.G. Anderson, K. Cromack, Jr, J.T. Williams, and R.E.Martin. 1979. Dead
and down woody material. Pp. 79–85, In J.W. Thomas (Ed.). Wildlife Habitats in
Managed Forests: The Blue Mountains of Oregon and Washington. US Department
of Agriculture, Forest Service, Washington DC. 512 pp.
2013 J.A. Homyack and C.A. Haas 365
Mazerolle, M.J., L.L. Bailey, W.L. Kendall, J.A. Royle, S.J. Converse, and J.D. Nichols.
2007. Making great leaps forward: Accounting for detectability in herpetological field
studies. Journal of Herpetology 41:672–689.
McCarthy, M.A., and K.M. Parris. 2004. Clarifying the effect of toe clipping on frogs
with Bayesian statistics. Journal of Applied Ecology 41:780–786.
McComb, W.C., T.A. Spies, and W.H. Emmingham. 1993. Douglas-fir forests. Managing
for timber and mature-forest habitat. Journal of Forestry 91:31–42.
McWilliams, W.H., B.J. Butler, L.E. Caldwell, D.M. Griffith, M.L. Hoppus, K.M.
Lausten, A.J. Lister, T.W. Lister, J.W. Metzler, R.S. Morin, S.A. Sader, L.B. Stewart,
J.R. Steinman, J.A. Westfall, D.A. Williams, A. Whitman, and C.W. Woodall. 2005.
The forest of Maine: 2003. Resource Bulletin NE-164. US Department of Agriculture,
Forest Service, Northeastern Research Station, Newtown Square, PA. 188 pp.
Mitchell, J.C., J.A. Wicknick, and C.D. Anthony. 1996. Effects of timber harvesting practices
on peaks of Otter Salamanders (Plethodon hubrichti) populations. Amphibian
and Reptile Conservation 1:15–19.
Morrison, M.L., B.G. Marcot, and R.W. Mannan. 1992. Wildlife-Habitat Relationships.
University of Wisconsin Press, Madison, WI.
Perkins, D.W., and M.L. Hunter. 2006. Effects of riparian timber management on amphibians
in Maine. Journal of Wildlife Management 70:657–670.
Petranka, J.W. 1998. Salamanders of the United States and Canada. Smithsonian Institution,
Petranka, J.W., M.E. Eldridge, and K.E. Healy. 1993. Effects of timber harvesting on
Southern Appalachain salamanders. Conservation Biology 7:363–370.
Pollock, K.H., J.D. Nichols, T.R. Simons, G.L. Farnsworth, L.L. Bailey, and J.R. Sauer.
2002. Large-scale wildlife monitoring studies: Statistical methods for design and
analysis. Environmetrics 13:105–109.
Pough, F.H., E.M. Smith, D.H. Rhodes, and A. Collazo. 1987. The abundance of salamanders
in forest stands with different histories of disturbance. Forest Ecology and
Reichenbach, N., and P. Sattler. 2007. Effects of timbering on Plethodon hubrichti over
twelve years. Journal of Herpetology 41:622–629.
Saunders, D.A., R.J. Hobbs, and C.R. Margules. 1991. Biological consequences of ecosystem
fragmentation: A review. Conservation Biology 5:18–32.
Sedjo, R.A. 1999. The potential of high-yield plantation forestry for meeting timber
needs. New Forests 17:339–359.
Semlitsch, R.D., B.D. Todd, S.M. Blomquist, A.J.K. Calhoun, J.W. Gibbons, J.P. Gibbs,
G.J. Graeter, E.B. Harper, D.J. Hocking, M.L. Hunter, D.A. Patrick, T.A.G. Rittenhouse,
and B.B. Rothermel. 2009. Effects of timber harvest on amphibian populations:
Understanding mechanisms from forest experiments. Bioscience 59:853–862.
Siry, J.P. 2002. Intensive timber management practices. Pp 327–340, In D.N. Wear and
J.G. Greis (Eds.). Southern Forest Resource Assessment. US Department of Agriculture,
Forest Service, Southern Research Station, Asheville, NC. 635 pp.
Smith, D.M., B.C. Larson, M.J. Kelty, and P.M.S. Ashton. 1997. The Practice of Silviculture:
Applied Forest Ecology. 9th Edition. John Wiley and Sons, New York, NY.
Smith, E.P. 2002. BACI Design. Pp. 141–148, In A.H. El-Shaarawi and W.W. Piergorsch
(Eds.). Encyclopedia of Environments. John Wiley and Sons. Chichester, UK.
366 Southeastern Naturalist Vol. 12, No. 2
Spies, T.A., J.F. Franklin, and T.B. Thomas. 1988. Coarse woody debris in Douglas-fir
forests of western Oregon and Washington. Ecology 69:1689–1702.
Tilghman, J.M., S.W. Ramee, and D.M. Marsh. 2012. Meta-analysis of the effects of
canopy removal on terrestrial salamander populations in North America. Biological
US Department of Agriculture, Forest Service. 2010. Major trend data. Avaialable online
at http://www.fia.fs.fed.us/slides/major-trends.pdf . Accessed 30 October 2012.
Walton, B.M. 2005. Salamanders in forest-floor food webs: Environmental heterogeneity
affects the strength of top-down effects. Pedobiologia 49:381–393.
Walton, B.M., and S. Steckler. 2005. Contrasting effects of salamanders on forest-floor
macro- and mesofauna in laboratory microcosms. Pedobiologia 49:51–60.
Welsh, H.H., Jr., and S. Droege. 2001. A case for using plethodontid salamanders for
monitoring biodiversity and ecosystem integrity of North American forests. Conservation
Wender, B.W. 2000. Impacts of seven silvicultural alternatives on vascular plant community
composition, structure, and diversity in the southern Appalachians. M.Sc. Thesis.
Virginia Tech, Blacksburg, VA.
Wyman, R.L. 1998. Experimental assessment of salamanders as predators of detrital food
webs: Effects on invertebrates, decomposition, and the carbon cycle. Biodiversity and