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2006 SOUTHEASTERN NATURALIST 5(1):113–126
Effects of Prescribed Fire on the Vegetation of a
Savanna-Glade Complex in Northern Arkansas
SEÁN E. JENKINS1,* AND MICHAEL A. JENKINS2
Abstract - In Spring of 1995 and 1997, 170 ha of a savanna-glade complex in the
Ozark Mountains of Arkansas were burned. These prescribed burns included 5 of 18
permanent plots established across the site in 1993. We surveyed the ground flora,
seedling, sapling, and overstory vegetation of these plots before and after burning.
The burns greatly impacted the sapling layer, where density decreased from 2540/ha
preburn to 610/ha after the second burn. Ground flora richness was unchanged
following burning, but evenness and diversity increased. Species richness, evenness,
and diversity also increased on the unburned plots. We observed large increases in
the cover of several glade and savanna species after burning. We observed similar
trends, but of lesser magnitude, on unburned plots.
Prescribed fire is an essential tool in the management of oak savanna,
glades, and other community types throughout eastern North America
(Abrams 1992, Cole et al. 1990, Rouse 1986, Stritch 1990). Oak savannas
and glades in the Ozark Highlands were historically maintained by anthropogenic
fire and had a diverse ground flora (Baskin and Baskin 2000, Guyette
and Cutter 1991, Jenkins 1997, Ladd 1991). However, the vast majority of
these communities have been lost since European settlement (Christensen et
al. 1996, Nuzzo 1986). Fire suppression has resulted in compositional and
structural changes in the remaining remnants, with a dramatic shift from
open landscape mosaics to closed-canopy, even-age stands of oaks and fireintolerant
species (Beilmann and Brenner 1951, Dey 2003, Ebinger 1997,
Guyette and McGinnes 1982, Kucera and Martin 1957, Lorimer 1993,
Robertson et al. 1997) and a corresponding loss of ground flora diversity
(McClain and Anderson 1990, Taft 1997). In the Ozark Highlands, savannas
have been overgrown with Quercus marilandica Muench. (blackjack oak),
Carya texana Buckl. (black hickory), and fire-sensitive woody species such
as Amelanchier arborea (Michx. f.) Fern. (shadbush). According to Reiter
(1991), glades have been invaded and fragmented by Juniperus viriginiana
L. (eastern red cedar), Q. marilandica, Ulmus alata Michx. (winged elm),
and Rhus copallinum L. (winged sumac). To maintain and restore remaining
remnants of savanna and glades, the National Park Service and other public
and private agencies have initiated prescribed burning programs. While
burning has been extensive, the impacts of prescribed fire have been difficult
1Department of Biological Sciences, Western Illinois University, Macomb, IL 61455.
2National Park Service, Inventory and Monitoring Program, Great Smoky Mountains
National Park, Gatlinburg, TN 37738. *Corresponding author - SEJenkins@
114 Southeastern Naturalist Vol. 5, No. 1
to assess due to the limited pre and post-burn data available from sites that
have had multiple burns.
On April 3–4, 1995, and again on March 19–20, 1997, the National Park
Service burned a 170-ha (420-acre) section of a large savanna-glade complex
at Cook Hollow, Buffalo National River, AR. Five permanent plots
from an 18-plot array established and sampled in 1993 as part of a long-term
study of landscape-scale fire effects (Jenkins and Rebertus 1994) were
burned both years. The original burn plan called for the area containing all
18 plots to be burned. However, due to staffing limitations, the burn unit was
reduced after the 18 plots were established. While this smaller burn reduced
the number of plots burned, it provided unburned reference plots for comparison.
Following the two burns, we resampled the woody vegetation
(trees, shrubs, and vines) on the five burned plots. During the summer of
1999, we resampled the ground flora on all 18 plots and the woody vegetation
on the 13 unburned plots.
The prescribed burns at Cook Hollow were conducted to accomplish two
management goals: (1) restore overstory structure by reducing woody stem
density and basal area, and (2) increase the cover of prairie and glade species
while increasing the overall diversity of ground flora species. In this paper,
we use data from the burned plots to determine whether the two burns met
these management objectives. In addition, we examine structural and compositional
changes on the burned and unburned plots to determine if factors
other than fire may be impacting the savanna-glade complex.
Cook Hollow is part of the Lower Buffalo Wilderness Area of the
Buffalo National River and is located at the confluence of the Buffalo
River with the White River in north-central Arkansas. The area is part of
the White River Hills Subsection of the Ozark Highlands Section within
the Eastern Broadleaf Forest Province (Keys et al. 1995) and is underlain
by strata of Ordovician limestone and sandstone, thus it has a rugged
topography and frequent rock outcroppings. Approximately 324 hectares
(800 acres) of oak woodland-savanna-glade complex cover the south aspect
of Turkey Mountain and the adjacent north aspect of Granite Mountain
(Fig. 1). The flora of the area is diverse. Jenkins and Rebertus (1994)
reported 240 species from the eighteen 500-m2 plots in Cook Hollow on
Turkey and Granite Mountain. In addition, Logan (1992) reported 193
species on 42 limestone glades located in the uplands along the Buffalo
National River, AR.
Carya texana, Quercus stellata Wangenh. (post oak), Quercus
marilandica, and Quercus prinoides Willd. (dwarf chinkapin oak) are dominant
overstory species across the glade complex at Cook Hollow (Jenkins
and Rebertus 1994), and they dominated the overstory on burned and unburned
plots sampled in this study. During the first survey in 1993, the
2006 S.E. Jenkins and M.A. Jenkins 115
combined importance value (IV; [relative basal area + relative density]/2) of
these 4 species was 69 on the burned and 55 on the unburned plots (Jenkins
and Jenkins 1999). Juniperus virginiana was a major overstory species on
the unburned plots (IV = 21.8), but it was mostly found in the pre-burn
sapling layer of the burned plots. Pinus echinata P. Mill. (shortleaf pine) was
a relatively important species on the burned plots (IV = 14.1), but it was a
minor component of the unburned plots (IV < 1). Carya texana and Q.
marilandica dominated the pre-burn sapling layers of both burned and
unburned plots (Jenkins and Jenkins 1999).
In 1993, eighteen permanent plots were established and sampled to
monitor the long-term effects of fire at Cook Hollow (Rebertus and Jenkins
1994; Fig. 1). In August of 1995 and 1998, following two prescribed burns,
the woody vegetation (trees, shrubs, and vines) was resurveyed on five of the
plots within the burn unit (1, 2, 3, 10, and 18; Fig. 1). During the summer of
1999, we sampled the ground flora (herbaceous plants and woody/semi
woody shrubs < 1 m tall) on all 18 plots and the woody vegetation on the 13
unburned plots. Because we had just sampled woody vegetation on the
burned plots in 1998, we did not resample these plots in 1999.
Figure 1. Locations of plots in Cook Hollow, Buffalo National River, AR. Plots 1, 2,
3, 10, and 18 are within the 170-ha (420-acre) unit burned in 1995 and 1997.
116 Southeastern Naturalist Vol. 5, No. 1
We sampled woody vegetation with a nested plot design. The dbh of all
trees ( ≥ 5 cm dbh) was measured within each 500-m2 plot. Saplings (2.5–
4.99 cm dbh) were tallied by species within two 100-m2 subplots nested
within each 500-m2 plot. During the 1993 survey, large seedlings (woody
stems > 0.5 m tall but < 2.5 cm dbh) were tallied in two 10-m2 circular
subplots, and small seedlings (woody stems < 0.5 m tall) were tallied within
a 1-m2 quadrat nested within each 10-m2 subplot. During the 1995 and
successive surveys, both large seedlings and small seedlings were tallied in
four 10-m2 subplots to reduce the large variability we observed in the
preliminary analysis of seedling data (Jenkins and Jenkins 1999).
We sampled the ground flora in June and in early September of 1993 and
1999. Some pre-vernal species may have already become senescent by June,
but most species were detected during sampling. The canopy cover of all
herbaceous, semi-woody, and woody ground flora species (< 1m tall) within
the plot was estimated in 20% cover-increment classes (1 = 0–20%, 2 = 21–
40%, 3 = 41–60%, 4 = 61–80%, and 5 = 80–100%) within each 500-m2 plot.
Species nomenclature follows Kartesz (1999).
We calculated overstory density (stems/ha), overstory basal area (m2/ha),
and sapling density (stems/ha) for each species on each plot for each
remeasurement. The percent of total sapling density comprised of dead stems
(hereafter “percent dead saplings”) was also calculated for each plot. We used
one-way repeated measures ANOVA to compare mean overstory density,
overstory basal area, sapling density, and percent dead saplings between years
(1993, 1995, and 1999) for the 5 burn plots. When ANOVA revealed a
significant effect, we used the Tukey multiple comparison test (α = 0.05) for
post-hoc comparisons between sample years. Paired t-tests were used to
compare overstory density, overstory basal area, sapling density, and percent
dead saplings between years (1993 and 1999) for the 13 unburned plots.
Percent dead sapling data were arcsine transformed prior to statistical analysis
to improve normality (Zar 1996). Even though the burned and unburned plots
were located on similar landscape positions and had similar species composition,
they are not true paired replicates. Therefore, we did not make direct
quantitative comparisons between burned and unburned plots.
Due to the large discrepancy in sampling area used for large seedlings
between the pre-burn and post-burn surveys, statistical comparisons were
only made between the post-burn surveys. Because of a severe growing
season drought, we observed no small seedlings during the 1998 resurvey of
the burned plots that could be identified to species. Consequently, only the
densities of large seedlings occurring in the two post-burn surveys (1995 and
1998) were compared on the 5 burned plots. All large-seedling densities
were converted to a per-ha basis prior to analysis with paired t-tests.
We calculated species richness (S), evenness (E), and Shannon-Weiner
diversity (H’) for ground flora species on each plot during each survey. We
2006 S.E. Jenkins and M.A. Jenkins 117
used paired t-tests to compare differences in ground flora species S, E, and H’
between the 1993 and 1999 surveys of burned and unburned plots. The scale
values of herbaceous vegetation were converted to class midpoints to calculate
S, E, and H’. Mid-point cover values were also used to examine changes in
the dominance of individual species. The covers of individual prairie grass
species, Andropogon gerardii Vitman. (big bluestem), Schizachyrium
scoparium (Michx.) Nash (little bluestem), Sorghastrum nutans (L.) Nash
(Indian grass), Sporobolus compositus (Poir.) Merr. var. compositus (composite
dropseed), Sporobolus clandestinus (Biehler) A.S. Hitchc. (rough
dropseed), and Bouteloua curtipendula (Michx.) Torr. (sideoats grama), were
combined for comparisons. Paired t-tests were used to compare herbaceous
cover mid-points between the 1993 and 1999 surveys of burned and unburned
plots. Prior to statistical testing, all cover data were arcsine transformed to
improve normality (Zar 1996). In comparisons of ground flora cover, individual
species or groups of species often have coverages that exceed 100%.
When this occurred, all cover values in comparison groups were mathematically
converted to a 100-point scale prior to transformation.
The mean overstory basal area on the burn plots did not differ significantly
between surveys (P = 0.759; 16.3 ± 5.7 m2/ha in 1993, 16.8 ± 6.0 m2/
ha in 1995, and 16.4 ± 5.5 m2/ha in 1999). Total basal area remained
relatively constant across the three surveys since most fire-killed trees were
of small diameter (data not shown). Mean overstory density was lower
following the burns, but these decreases were statistically insignificant (P =
0.358; 1164 ± 275 stems/ha in 1993, 992 ± 273 stems/ha in 1995, and 1048 ±
207 stems/ha in 1999).
On the unburned plots, basal area increased (P = 0.118) between 1993
(15.8 ± 2.1 m2/ha) and 1999 (16.5 ± 2.1 m2/ha). Tree density on the unburned
plots changed little between 1993 and 1999 (P = 0.554; 1246 ± 116 m2/ha in
1993 and 1274 ± 101 m2/ha in 1999). However, the basal area of three
species increased slightly; Carya texana (P = 0.056), Juniperus virginiana
(P = 0.049), and Quercus stellata (P = 0.044).
The results of the two post-burn surveys show that the fires had a
dramatic impact on the sapling layer (Fig. 2). The density of saplings
decreased after each successive burn from an initial density of 2540 stems/
ha in 1993 to 610 stems/ha after the second burn (Fig. 2A). We observed a
significantly greater percent of dead saplings following the two burns than
in the preburn survey (P < 0.05; Fig. 2B). The mean percent of dead
saplings increased from 3% in 1993 to 73% in 1995 after the first burn and
69% after the second burn. Carya texana, Quercus marilandica, Q.
stellata, and Sassafras albidum (Nutt.) Nees (sassafras) exhibited the
118 Southeastern Naturalist Vol. 5, No. 1
greatest overall reductions in density, although the reductions were not
statistically significant (p > 0.05; Fig. 2A). After the first fire, the percentage
of dead saplings for all these species was greatest on two upper slope
plots and a lower slope plot dominated by Pinus echinata (data not shown).
The mean density of saplings on the unburned plots did not change
significantly between surveys (P = 0.602; Fig. 3). In addition, we observed
no significant changes in the density of individual species between 1993 and
1999. The mean percent dead saplings did not differ between surveys for all
pooled species or any individual species (data not shown). On individual
plots located on more drought-prone glades, densities were reduced by up to
Figure 2. (A) Sapling density (mean stems/ha ± 1 SE) and (B) percent dead saplings
(mean percent ± 1 SE) on 5 burned plots sampled pre-burn in 1993 and post-burn in
1995 and 1998. Mean values from each sample year for each species were compared
using a one-way repeated measures ANOVA with Tukey multiple comparison tests.
Sample years within a species not superscripted with the same letter are significantly
different (P < 0.05).
2006 S.E. Jenkins and M.A. Jenkins 119
72% following three successive years of growing season drought (data not
shown). These glade plots have shallow soil over bedrock, and most of the
saplings that died were Quercus marilandica, Q. stellata, and Juniperus
virginiana. This loss of density was counterbalanced by density increases on
plots located in more mesic landscape positions.
Following the first burn in 1995, overall mean density of large tree
seedlings remained stable (4100 stems/ha in 1993 and 4050 stems/ha in
1995; Table 1). However, these results must be interpreted with caution
since sample area was increased in the post-burn surveys and statistical
analysis was only performed on post-burn data. The density of Sassafras
albidum was greater following the first fire. Of the oak species, Quercus
Figure 3. Sapling density (mean stems/ha ± 1 SE) on 13 unburned plots sampled in
1993 and 1999. Mean values from each sample year for each species were compared
using a one-way repeated measures ANOVA with Tukey multiple comparison tests.
Means did not differ significantly (P > 0.05) between years for any comparison group.
Table 1. Density per ha (mean ± 1 SE) of selected large seedling (stems > 0.5 m tall; < 2.5-cm dbh)
on the 5 burned plots. Data were collected pre-burn in 1993 and post-burn in 1995 and 1998. Mean
densities between 1995 and 1998 were compared with paired t-tests. Densities from 1993 are
presented for qualitative comparison; no statistical analyses were conducted on these data.
Species 1993 1995 1998 P-value
Carya texana 200 ± 100 500 ± 300 750 ± 100 0.326
Quercus marilandica 200 ± 200 550 ± 400 400 ± 200 0.727
Quercus stellata 1100 ± 1000 200 ± 100 250 ± 250 0.815
Sassafras albidum 1000 ± 800 2250 ± 1000 2050 ± 1000 0.528
Total Density 4000 ± 900 4050 ± 1400 4600 ± 900 0.504
120 Southeastern Naturalist Vol. 5, No. 1
stellata seedling density was lower after the first burn. The second burn did
not significantly increase total large seedling density or the density of any
individual species. Field observations showed that resprouts were less numerous
and not as vigorous after the second burn, especially those of woody
vine and shrub species including Vaccinium stamineum L. (deerberry), Acacia
angustissima (P. Mill.) Kuntze (prairie acacia), Passiflora lutea L.
(passion flower), Rhus aromatica Ait. (aromatic sumac), Rosa carolina L.
(prairie rose), Smilax bona-nox L. (green brier), Vaccinium pallidum Ait.
(lowbush blueberry), and Vitis spp. (grape vine). Although present in the
larger plots, these woody shrub and vine species were not encountered in the
permanent nested circular seedling plots due to patchy distributions and/or
lower abundances after the second burn.
Ground flora vegetation
The results of the pre- and post-burn surveys suggest that the prescribed
burns have influenced the diversity and composition of the ground flora
(Table 2, Fig. 4A). Species richness (S) did not differ significantly between
surveys (P = 0.840), but species evenness (E) and diversity (H’) increased
significantly (P = 0.02 and P = 0.04, respectively). However, we noted no
significant change in the cover of any individual mesic-site species after
burning (P > 0.1; data not shown). The combined cover of prairie grass
species increased greatly (171%) after the two burns (Fig. 4A). In addition,
the cover of Dichanthelium species (panic grasses) increased 82%, and the
cover of two glade species, Lespedeza virginica (L.) Britt.-slender bush
clover and Solidago radula Nutt. (western rough goldenrod), more than
doubled (Fig. 4). Triodanis perfoliata (L.) Nieuwl. (Venus looking glass—
a light-seeded annual) exhibited a large increase in cover due to the seedbed
created by the burns. Solidago nemoralis Ait. (gray goldenrod), a
common species in abandoned fields and open woods, disappeared from all
5 plots following the two burns.
We observed significant increases in the values of S, E, and H’ for the
unburned plots (Table 2). We observed a decrease in the cover of 7 mesic
and old-field species between the two surveys (Fig. 4B). These species
likely increased in cover as a result of the large rainfall in 1993, but
decreased in cover in response to the severe droughts of 1996, 1998, and
1999. Conversely, the cover of 5 dry-site and glade species increased between
surveys (Fig. 4B).
Table 2. Changes in ground flora species richness, evenness, and Shannon-Weiner diversity on
5 burned and 13 unburned plots in 1993 and 1999. Prescribed burns were conducted in 1995 and
1997. Mean values were compared between years with paired t-tests.
Richness Evenness Diversity
1993 1999 P-value 1993 1999 P-value 1993 1999 P-value
Burned 46.0 45.4 0.840 0.659 0.757 0.026 2.49 2.83 0.042
Unburned 50.5 55.4 0.035 0.689 0.784 < 0.001 2.69 3.13 < 0.001
2006 S.E. Jenkins and M.A. Jenkins 121
Burning was initiated at Cook Hollow to reduce overstory density and
basal area and increase ground flora diversity and extent. Although the
prescribed burns have not altered overstory basal area and density, the fires
have drastically reduced the density of both C. texana and Q. marilandica
Figure 4. Percent cover (mean ± 1 SE) of Prairie grasses, Dichanthelium species, and
selected individual species on (A) the 5 burned plots in 1993 (preburn) and 1999
(postburn) and (B) the 13 unburned plots in 1993 and 1999. Species on unburned plots
are grouped according to common habitat (mesic site and old-field species vs. dry site
and glade species). Mean species cover was compared in burned and unburned plots
between 1993 and 1999 with paired t-tests. *P < 0.1, **P < 0.05, ***P < 0.01. Species
that exhibited large or significant changes in cover are included.
122 Southeastern Naturalist Vol. 5, No. 1
saplings. Reduced recruitment of saplings and small trees into the larger size
classes as a result of repeated burning will lead to a gradual long-term
reduction in both tree basal area and density (Harmon 1984, Taft 1997).
After thirteen years of burning, White (1983) observed a reduction in the
density and basal area of stems less then 25 cm dbh in a Quercus
ellipsoidalis E.J. Hill (northern pin oak)-dominated former upland savanna
in Minnesota. Peterson and Reich (2001) found that burning every 2 to 3
years reduced sapling density as well as canopy ingrowth, while less frequent
fires lead to increased sapling density in savannas. In our study,
reductions in sapling density were not uniform across the burned plots. We
observed the greatest reductions on upper slopes, which experienced greater
preheating of fuels, and in pine stands, which have more pyrogenic liter.
Both of these factors led to greater fire intensity in these two areas (Jenkins
and Jenkins 1999, Jenkins et al. 1997).
On the unburned plots, we observed little change in either the overstory or
sapling layers between 1993 and 1999. As with the burned plots, our results
were not uniform across all unburned plots. We observed a decrease in the
density of Quercus marilandica, Q. stellata, and Juniperus virginiana stems
on unburned plots located on exposed glades. These glade sites have shallow
soils that amplify the impacts of drought on tree survival (Baskin and Baskin
2000). However, these decreases in stem density were counterbalanced by
density increases on plots located in more mesic landscape positions. Gradual
increases in the basal area and density of Carya texana, Juniperus virginiana,
and Quercus stellata on these sites may have a pronounced negative effect on
ground flora diversity due to increased shading by trees and saplings (Taft
1997). In particular, invasion by J. virginiana has been observed on other
glades (Erickson et.al. 1942, Heikens and Robertson 1995, Jeffries1985,
Kucera and Martin 1957, McClain and Ebinger 2002, Nelson and Ladd 1983)
and is often of concern to managers (Jenkins and Rebertus 1994, Reiter 1991).
While largely absent in the overstory of the burned plots, Juniperus virginiana
was a major component across the 13 unburned plots, where it is more of a
management concern. In 2000, the entire glade complex was burned and
current plans call for continued burning of this larger unit. This expanded
burning should eventually reduce the density of J. virginiana across the
savanna-glade complex. A combination of overstory thinning and prescribed
burning could be used to more rapidly decrease overstory shading and increase
ground flora cover and diversity (Neilson et al. 2003).
Our results show that the diversity of ground flora species and the cover
of prairie grasses and other savanna/glade species have increased on the
burned plots, meeting the management goals of the National Park Service.
Other studies have also shown increased cover of prairie grass species
following fire in oak savannas (DeSelm and Clebsch 1991, Heikens et al.
1994, Tester 1996). However, Taft (2003) observed a decrease in the cover
of C4 prairie grasses (e.g., Schizachyrium scoparium and Sorghastrum
nutans) following two burns of a dry sandstone barrens. The significant
2006 S.E. Jenkins and M.A. Jenkins 123
increases in E and H’ we observed may have resulted from post-fire
reduction in the cover of more mesic species that have invaded with fire
suppression. Similar decreases in the cover of mesic non-savanna species
have been observed in other oak savannas in the Midwest following both
single (Hruska and Ebinger 1995) and multiple burns (Tester 1996).
Since the ground flora was not sampled until two years after the second
burn, it is impossible to know the immediate effects of the fire on ground
flora diversity at the site. For instance, Nuzzo et al. (1996) observed an
increase in herbaceous species richness in an Illinois sand savanna after a
prescribed fire, but richness declined following a fire-free year. In a long
term study of post-fire effects on a southern Illinois barren, Taft (1997)
found that some annual and biennial species increased dramatically following
a dormant-season burn, but then declined to preburn abundances two
years after the fire. Heikens et al. (1994) found that the abundance of
herbaceous species did not significantly increase following a spring burn on
shale and chert savannas in southern Illinois.
The increase in species richness, evenness, and diversity on the unburned
plots was unexpected. However, since there were 6 years separating the
surveys, such fluctuations in diversity and composition may have occurred
as a result of factors other than fire including periodic droughts, spatiotemporal
variability in colluvial processes, and annual variability in
freeze-thaw cycles. Periodic disturbance, low soil fertility, and resource
heterogeneity have all been suggested as potential mechanisms that promote
diversity and prevent dominance by a single species or group of species
(Huston 1994, Tilman 1982). We conducted our first survey in 1993, one of
the wettest years on record in much of the central United States. Therefore,
an abundance of moisture may have given a short-lived competitive edge to
the more dominant species. These species likely increased in cover as a
result of the large rainfall in 1993, but decreased in cover in response to the
severe droughts of 1996, 1998, and 1999. As cover of these species decreased,
evenness may have increased, causing a corresponding increase in
species diversity. In addition, prairie grasses may have increased in cover
during droughts years as growing space was relinquished through the dieback
of more mesic woodland species. Cyclic changes in the cover of prairie
grasses, such as Schizachyrium scoparium, regardless of the burning regime,
have been noted (DeSelm and Clebsch 1991). However, total percent increases
in the cover of prairie grasses and several other glade/savanna
species were much greater on burned plots than on unburned plots, suggesting
that fire may have increased the cover of these species beyond the
baseline increase that occurred on the unburned plots.
While our results show that the two spring burns were effective in meeting
management goals, changing overstory structure will be a slow process under
this low-intensity burning regime. While spring has traditionally been the
dominant season for prescribed burning in the East, a more mixed seasonal
burning regime with more variable fire intensities may better reproduce past
124 Southeastern Naturalist Vol. 5, No. 1
conditions that maintained savannas and glades. Historically, burns in Midwest
savanna and glades occurred throughout the growing season with many
fires occurring during October and November (McClain and Elzinga 1994).
Varying the season, return interval, and size of burns will better maintain tree
recruitment and favor a variety of ground flora reproductive strategies
(Jenkins and Rebertus 1994). This will, in turn, enhance and maintain avian
and general faunal diversity (Brawn et al 2001, Madden et al. 1999).
The results of our study highlight the importance of including unburned
reference plots in fire effects monitoring. If data were not available from
unburned plots, observed increases in species diversity and cover would
have been attributed solely to fire. However, our results suggest that, in
addition to fire, changes in ground flora diversity and composition are
influenced by a combination of poorly understood factors including precipitation
cycles, light regimes, topography, edaphic characteristics, resource
heterogeneity, and disturbance. Understanding these factors may not be
critical in long-term monitoring studies that seek only to quantify change
across a given area. However, their understanding is critical to understanding
the related mechanisms responsible for this change.
We thank Robert Pulchow, Thara Baker-Alley, Joseph Alley, and Mike Morris for
their field assistance. We also thank Chris Webster and two anonymous reviewers for
helpful comments on an earlier version of this manuscript. This study was funded
through grants from the National Park Service, Buffalo National River, Harrison, AR.
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