Biol Invasions
DOI 10.1007/s10530-008-9241-2
ORIGINAL PAPER
Effects of experimental manipulation of light and nutrients
on establishment of seedlings of native and invasive woody
species in Long Island, NY forests
Jessica Gurevitch Æ Timothy G. Howard Æ Isabel W. Ashton Æ
Elizabeth A. Leger Æ Katherine M. Howe Æ Eliza Woo Æ
Manuel Lerdau
Received: 9 June 2007 / Accepted: 20 November 2007
Ó Springer Science+Business Media B.V. 2008
Abstract While earlier studies on the process of
invasion often focused on single factors or on the
general explanation of ‘disturbance,’ recent work has
Electronic supplementary material The online version of
this article (doi:10.1007/s10530-008-9241-2) contains
supplementary material, which is available to authorized users.
J. Gurevitch (&) T. G. Howard I. W. Ashton E. A. Leger K. M. Howe E. Woo M. Lerdau
Department of Ecology and Evolution, Stony Brook
University, Stony Brook, NY 11794-5245, USA
e-mail: [email protected]
T. G. Howard
New York Natural Heritage Program, 625 Broadway,
Albany, NY 12233-4757, USA
I. W. Ashton
Department of Ecology and Evolutionary Biology,
University of California Irvine, Irvine, CA 92697, USA
E. A. Leger
Department of Natural Resources and Environmental
Science, University of Nevada, Reno, 1000 Valley Road,
Reno, NV 89512, USA
K. M. Howe
Midwest Invasive Plant Network, The Nature
Conservancy, 1505 N. Delaware St., Suite 200,
Indianapolis, IN 46202, USA
M. Lerdau
Blandy Experimental Farm and Department
of Environmental Sciences, University of Virginia,
Clark Hall, 291 McCormick Rd, PO Box 400123,
Charlottesville, VA 22904-4123, USA
attempted to move towards a more mechanistic
understanding of the factors that promote plant
community invasion. Manipulative experiments provide a means for discerning causal relationships and
interactive effects of environmental factors in promoting invasion; such experiments have been
conducted in a number of grassland and shrub
ecosystems. This study extends multifactor manipulative experiments into forest communities to
compare factors influencing early seedling establishment for native and invasive woody plants. In Long
Island, NY, invasion patterns are correlated with
forest community type (pine barrens or hardwood),
light availability, and soil N and Ca. We conducted
manipulative field experiments in two different years
to determine the relative importance and interaction
of experimental gaps and N and Ca addition in pine
barrens and hardwood forests in promoting invasion.
We used seedlings of seven common native and
invasive species in the first experiment, and 16 native
and invasive species paired phylogenetically in the
second experiment. Light had the strongest effect on
plant growth; all plants grew more in gaps. We found
no difference in the average growth rates of native
and invasive species. Invasives responded more to
high resources than did natives, with highest relative
growth rates in gaps in the more fertile soils of the
hardwood forests. Opportunities for invasion may
differ from year to year, with differential success of
invaders only in some years and under some
environmental conditions. Clearly, to understand the
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J. Gurevitch et al.
complex interactions between resources and invasion
in forests will require many manipulative experiments across a range of environments and using
suites of invasive and native species.
Keywords Pine barrens Exotic species Phylogenetically paired experiment Multifactor
experiment Community invasibility Forest
invasion Nitrophilic plants Experimental gaps Northeastern US hardwood forests
Introduction
The spread of invasive, exotic species has received an
enormous amount of attention from ecologists and
others concerned with conservation of biodiversity.
Invasive organisms pose serious threats both to
humans and for the preservation of natural systems
(Lodge 1993). Disturbance has long been considered
a major factor facilitating invasion by non-native
species (e.g., Elton 1958; Crawley 1987; Lodge
1993), and more recent theory has proposed that the
effects of disturbance on invasion can be explained as
the result of a temporary increase in the amount of
unused resources (Davis et al. 2000). However, the
term disturbance encompasses a broad range of the
processes that occur during invasions, and even the
explanation that disturbance promotes invasion by
providing a pulse of resources covers a wide range of
factors that may affect organisms in very different
ways. For instance, disturbances in plant communities can separately or simultaneously affect soil
resource availability, plant competition, herbivory,
light availability, and seed bank dynamics (e.g.,
Orians 1986). Recently, attention has been turning
toward an analysis of more-specific factors promoting
or inhibiting plant community invasibility. We
believe that testing the relative effects of specific
mechanistic factors and their interactions (e.g., light
and N) in promoting or inhibiting invasion in a wide
range of plant communities will lead to the most
rapid advances in understanding the invasion process.
Such an understanding may also prove to be useful in
managing invasions.
The goal of our study was to experimentally test the
role and importance of multiple hypothesized aboveand belowground factors on early seedling establishment in determining forest community resistance or
123
susceptibility to invasion. To obtain a better understanding of invasions it is important to examine the
differential success of invasive and native species in
different environments in response to hypothesized
causal factors. We chose to use early seedling establishment for these experiments because this life-history
stage is critical in determining whether an invasion
succeeds or fails. Studies in grassland systems have
increasingly used factorial experiments to advance our
understanding of mechanisms promoting invasion of
non-native species (e.g., Malmstrom et al. 2006;
Chambers et al. 2007), although relatively few such
studies to date have considered the responses of more
than a small number of invasive species (generally one
species; but see, e.g., Foster and Dickson 2004). Very
few experimental studies in forests have considered the
effects of more than one factor in promoting invasion
of non-natives (but see Von Holle 2005). More
experimental studies are needed that test the effects
on multiple species of multiple hypothesized factors
and their interactions in inhibiting or promoting plant
community invasion.
Many forests of the eastern United States are at
particular risk for invasions by alien species, both
because these communities are exposed to a large
number of exotic species growing within close proximity in heavily perturbed conditions, and because of
the large anthropogenic influences such as N deposition and fragmentation that are currently affecting
these systems. Eastern forests today exist within an
urban/suburban matrix that provides a large potential
donor pool of exotic species poised to become
invasive. In contrast to more extensive or remote
ecosystems where dispersal of exotics to the ecosystem
may often limit invasions, invasions of forests in the
eastern USA are likely to be more influenced by factors
other than dispersal.
N deposition in forests of the eastern USA has
been identified as one of the critical environmental
problems facing these ecosystems, and many invasive
plants are known to be nitrophilic (Aber et al. 1989;
Scherer-Lorenzen et al. 2000; Gilliam 2006). Studies
have demonstrated that increases in N deposition
cause soil Ca to move into solution and leach from
the terrestrial ecosystem into stream water (Vitousek
et al. 1997; Likens 1998; Lawrence et al. 1999); as
these ecosystems become enriched in N from anthropogenic deposition, they can become depleted in Ca
(Aber et al. 1998). The potential role of resource
Effects of experimental manipulation of light and nutrients
availability in facilitating invasion (e.g., Levine and
D’Antonio 1999; Stohlgren et al. 1999) may be
particularly important in the eastern USA because
of the large influences that human activities have had
on the availabilities of both N and Ca over the last
50 years (Aber et al. 1998; Likens 1998; Lawrence
et al. 1999). These changes in resource availability
could profoundly alter the susceptibility of these
ecosystems to invasions (e.g., Dukes and Mooney
1999).
Previous research (Howard et al. 2004) in this
system revealed broad patterns of susceptibility to
invasion in Long Island forest communities that were
closely related to site quality factors, particularly soil
fertility. Thus, at landscape scales pine barrens sites
tend to be essentially uninvaded, while mesic deciduous hardwood sites often have higher rates of
invasion. Both descriptive and experimental evidence
confirmed that light, N, and Ca can be critical in
influencing susceptibility of forest communities to
invasion by exotic plants. We hypothesize that it is
the combination of more fertile soils with high light
levels that facilitates invasion in hardwood forest
sites in which the canopy has been disturbed, but that
pine barrens sites are resistant to invasion (at least by
the common invasive species currently present) due
to poor soils, regardless of light levels. Howard et al.
(2004) reported that invasion was strongly correlated
with both soil Ca and N. Pine barrens soils had
greater litter depth, lower total and nitrate N mineralization rates, and lower Ca (as well as Mg, P, and
CEC; Howard et al. 2004) compared to hardwood
forests soils. Soil N and Ca are correlated, however,
so it is impossible to determine from observational
data whether N, Ca, or both nutrients are causally
linked to invasion status. They found no relationship
between invasion and soil texture or pH, suggesting
that only some of the soil characteristics that
differentiate sites are important in determining invasion. Decomposition rates, foliar N concentration,
and N loss rates for experimentally manipulated litter
were increased in hardwood forests compared to pine
barrens sites and to mesic uninvaded sites (Ashton
et al. 2005; I. Ashton, unpublished data), suggesting
that feedbacks exist between N availability and
invasion success.
While documenting patterns of invasion is a
critical step toward understanding the underlying
processes governing invasions, alternative hypotheses
may explain the same patterns found in an observational study equally well, and it is usually impossible
to separate the effects of co-occurring factors.
Manipulative field experiments provide the most
robust tests of the factors determining community
invasibility, and offer the potential for testing which
factors are of greatest importance in promoting or
inhibiting the invasion of natural communities. In
addition, they offer the opportunity to examine the
role of environmental factors when propagule availability is eliminated as a factor determining the
degree of invasion. We conducted manipulative
factorial split-plot experiments to examine the
responses of a suite of native and invasive species
to experimentally created gaps and to the addition of
soil nutrients in two forest types, hardwood forests
and pine barrens, in two different years. Because the
experiments were somewhat different in the two years, their results also invite consideration of how to
compare the results of ecological field experiments to
reach more general conclusions, and how to evaluate
the agreement (or lack thereof) between different
experiments.
Methods
To examine the differential effects of nutrient additions,
light availability, and forest type on the establishment
and growth of woody native and invasive seedlings,
we conducted two similar manipulative experiments in
two years, 2000 and 2003. We used similar approaches
in the experiments in the two years, but the experimental designs differed in some details. In both years,
we manipulated N, Ca, and canopy opening (experimental gaps) using replicated, split-block designs.
In 2000, we chose six sites: three pine barrens sites
with lower background soil nutrient levels, and
three hardwood forest sites with higher soil nutrients,
based on the results of previous research (Howard
et al. 2004). The pine barrens sites were Sarnoff
Preserve, N 40°53.2120 W 72°38.6940 , Topping Path,
N 40°51.1360 W 72°43.9290 , and Hot Water Street,
N 40°52.4410 , W 72°42.8680 and the hardwood forest
sites were East Farm, N 40°54.3310 W 73°09.1040 ,
Weld Preserve, N 40°54.5100 W 73°12.6310 , and
Rocky Point, N 40°56.5740 W 72°56.9470 . At each
site, we identified two separate localities at least 20 m
apart, and chose one at random to create a gap by
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removing all trees using a chainsaw. In each plot, all
shrubs and herbaceous understory plants were cut at
ground level within a circular area of 6 m radius. We
did this to maximize the chances for survival of the
experimental seedlings, and to enhance the consistency
of the experimental treatments and our ability to detect
responses to them (e.g., a seedling under a shrub would
respond differently to an experimental gap than one that
was not).
In 2003, we used four of these sites (two pine
barrens sites, Topping Path and Sarnoff Preserve, and
two hardwood forest sites, East Farm and Weld
Preserve), each with a gap and undisturbed canopy
(non-gap) treatment, treated as above. We used fewer
sites in 2003 and greatly expanded the number of
species used, as described below.
In both years, three 2.5 9 2.5 m square blocks
were placed at each gap/non-gap locality, with each
block consisting of four 1 9 1 m plots separated by
0.5 m. Blocks were arranged radiating out from a
circle with their inner edges 1.5 m from the center
of the circle, which was at the center of the gap or
non-gap area. We chose the first block location in
each spot using a random compass direction from
the center of the circle. Blocks were equidistant
from each other, separated by [1 m at their inner
corners, and encompassed within a 4 m radius
circle. We randomly assigned treatments consisting
of one of four nutrient levels (+N, +Ca, +N and
Ca) or a control (no nutrients added) to each plot
within each block. In each 1 9 1 m plot, we
planted 14 seedlings in a 4 9 4 grid (14 seedlings
and two blank spots) with 20 cm separating each
individual centered in the middle of each plot. In
2003, 15 or 16 seedlings were planted in a 4 9 4
grid in each plot with 20 cm separating each
individual.
In both years, seeds were stratified for 9–16 weeks
(depending on species) over the winter in a laboratory
refrigerator at 2°C. Seedlings were planted in a
greenhouse on the Stony Brook University campus in
tubular Cone-tainers (2.5 cm diameter, 12 cm depth,
66 ml volume, RLC4 Pine Cell, Steuwe & Sons, Inc.,
Corvallis, OR) on 10–27 March 2000, and 19
February–4 March 2003 in standard potting mix.
Germination took approximately 3–5 weeks, depending on species, after which the seedlings were
maintained in the greenhouse until transplanting to
the field sites.
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Before planting in both experiments, we recorded
the height, stem diameter, and number of leaves (where
possible) of every individual to estimate initial mass.
We then measured another set of individuals, sacrificed,
dried, and weighed them, and used regressions to
estimate initial mass of those planted in the field.
In 2000 we planted seedlings of four invasive
species and three natives (Table 1), chosen to represent
dominant or locally common species and a range of
functional groups (trees, vines and shrubs). Seedlings
(2 individuals 9 7 species 9 4 nutrient treatments 9
3 blocks 9 2 light treatments 9 6 sites = 2016 individuals) were randomly assigned to treatment and
location and planted in the ground from 30 May to 16
June. On 20 June 2000, we replaced dead seedlings (55
individuals, or 2.7%).
In 2003, we chose eight native and eight invasive
species (Table 1), paired phylogenetically as closely as
possible, representing a range of functional types and
found regionally (in several cases phylogenetic triplets
were used). Because it was not possible in every case to
find congeneric or confamilial pairs meeting all of
these criteria and for which seeds were available, in
two cases species were paired by order (Ceanothus/
Elaeagnus, both N-fixing shrubs, and Lonicera/
Viburnum, both shrubs). Seeds were germinated as in
2000 in a greenhouse. Seedlings (1,464 individuals;
because numbers of individuals for some species were
limited due to poor germination, the experimental
design was unbalanced) were planted on 29 and 30
May 2003. Plants were individually protected from
deer and rabbits with plastic mesh cylinders (45 cm
high, 10 cm diameter) because an experiment in 2002
was largely destroyed by deer herbivory (J. Gurevitch,
unpublished data). On 23 June 2003, we replaced dead
seedlings (99 individuals, or 6.8%).
In 2000, Ca addition plots received 30 g of Ca in
the form of CaSO4 (101.90 g ± 0.09 g of CaSO4),
raked into the soil surface on 26 and 27 June. Because
of the lack of growth responses to this treatment in
2000, Ca was greatly increased in 2003. In 2003, Ca
addition plots received 200 g of Ca in the form of
CaSO4 (679.38 g of CaSO4), raked into the soil on 12
and 14 of May. In 2000, we added N in a solution of
(NH4)2SO4 and NaNO3 in three applications of 1 l
each (26/27 June, 2 August and 29 August) via
backpack sprayer with a total addition of 20 g N
m-2year-1. In 2003, we added N in the same total
amounts, manner, and form in two applications on
Effects of experimental manipulation of light and nutrients
Table 1 Species planted in the 2000 and 2003 experiments, with their functional groups, families and origins (exotic, E, or native, N)
Species planted 2000
Origin,
functional
group
Family
Species planted 2003
Origin,
functional
group
Family
Acer platanoides
E, tree
Aceraceae
Acer platanoides
E, tree
Aceraceae
Acer rubrum
N, tree
Aceraceae
Acer rubrum
N, tree
Aceraceae
Pinus rigida
N, conifer tree
Pinaceae
Ampelopsis brevipedunculata
E, vine
Vitaceae
Parthenocissus quinquefolia
N, vine
Vitaceae
Vitis novae-angliae
N, vine
Vitaceae
Berberidaceae
Elaeagnus umbellata
E, shrub
Elaeag-naceae
Berberis canadensis
N, shrub
Rosa multiflora
E, shrub
Rosaceae
Berberis thunbergii
E, shrub
Berberidaceae
Vitis novae-anglea
N, vine
Vitaceae
Celastrus orbiculata
E, vine
Celastraceae
Ampelopsis
brevipundiculata
E, vine
Vitaceae
Celastrus scandens
N, vine
Celastraceae
Ceanothus americanus
N, shrub
Rhamnaceae
Elaeagnus umbellata
E, shrub
Elaeagnaceae
Prunus serotina
N, tree
Rosaceae
Rosa caroliniense
N, shrub
Rosaceae
Rosa multiflora
E, shrub
Rosaceae
Viburnum acerifolium
N, shrub
Adoxaceae
Lonicera mackii
E, shrub
Caprifoliaceae
12–14 May and 9–18 July 2003. In both experiments,
pH was checked before and after nutrient additions
for each plot, but was unchanged, so no adjustments
for pH were necessary. All plots not receiving water
from the N application were given 1 l of water.
We harvested all the individuals between 21–28
September, 2000 and 26–27 August, 2003. We
excavated root systems for each seedling, and placed
the entire plant in a labeled paper bag. After drying to
constant weight in a drying oven, we separated roots
from shoots and weighed root, shoot, and total mass
separately to the nearest 0.001 g.
different transformations, suggesting that the analyses
of variance (ANOVAs) were relatively robust to the
violations of these assumptions. Akike information
criteria values were also best (smallest, indicating
better fit of the statistical model) using the natural log
of the relative growth rate, compared to other measures
of growth or final size (not shown). PROC MIXED
uses restricted maximum-likelihood estimates of the
variance parameters and Satterthwaite estimates of the
degrees of freedom (resulting in fractional degrees of
freedom); we used the ‘random’ statement to specify
denominators for the F tests.
Statistical methods
Results
We used PROC MIXED in SAS 9.1 to analyze the data.
Growth was calculated as the natural log of the relative
growth rate (RGR) (final estimated initial dry mass
divided by estimated initial dry mass, plus 1.0), and
back-transformed to create the figures. Residuals of
untransformed data and various transformations were
examined for fit to normal distributions and homoscedasticity; while the natural log transformed data met
the assumptions better than did other transformations,
it did not satisfy these assumptions. Nevertheless,
results did not differ strongly among the analyses using
Effects of invasive status, family, and species on
growth and allocation
Native and invasive species did not differ in growth
rates in either the 2000 or 2003 experiments (Fig. 1;
invasive status was not significant in 2000: F = 0.53,
df = 1, 5.08, P = 0.50 or in 2003: F = 1.13, df = 1,
8.38, P = 0.32). There were no significant interactions
between invasive status and any other factor for RGR
in 2000. Invasive plants allocated less tissue to roots
relative to shoots on average compared to native plants
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J. Gurevitch et al.
A
0.8
0.7
0.6
RGR
0.5
0.4
0.3
0.2
0.1
0.0
B
2000
2003
1.6
1.5
Root to shoot ratio
1.4
1.3
1.2
1.1
1.0
0.9
0.8
Invasive
Native
Fig. 1 (a). Mean relative growth rates [RGR, (final dry
weight - estimated initial dry weight)/estimated initial dry
weight] in 2000 and 2003 for native (triangles) and invasive
(circles) species ± standard errors; N = 516 and 845 native
and invasive plants in 2000, and 712 and 634 for natives and
invasives in 2003. Means were calculated on transformed data
(natural log + 1) and then back-transformed for graphing in
this and all subsequent figures. Invasive status was not
significant in either year (2000: F = 0.53, df = 1, 5.08,
P = 0.50, or 2003: F = 1.13, df = 1, 8.38, P = 0.32). (b).
Mean root-to-shoot ratios (root dry weight divided by shoot dry
weight) in 2003 for native (triangle) and invasive (circle)
species ± standard errors; N = 708 and 632 for native and
invasive plants. Means were calculated on transformed data
(natural log + 1) and then back-transformed for graphing.
Natives had significantly higher root-to-shoot ratios than
invasives (F = 14.29, df = 1, 11, P = 0.003). There were no
significant differences on average between native and invasive
plants in root-to-shoot ratios in 2000
in 2003 (Fig. 1b, F = 14.29, df = 1, 11, P = 0.003).
There were no significant differences on average
between native and invasive plants in root-to-shoot
ratios in 2000. While individual species varied widely
in root-to-shoot ratios, these ratios were remarkably
consistent within species across the two years (see the
electronic supplementary material); the differences
between the two years in allocation were largely due to
123
the differences between the species used in the two
experiments. Neither gaps nor the addition of Ca or N
affected the allocation to above- versus belowground
tissue in either year. Families differed in their root to
shoot ratios in 2003 (F = 5.92, df = 6, 10.1,
P = 0.007); there were no other significant factors or
interactions affecting root-to-shoot ratios.
Invasive species responded to hardwood forest
gaps with higher growth rates in comparison to native
species in 2003, while invasives and natives had
similar growth rates in intact hardwood forests, and
both responded similarly to gaps in the pine barrens
(Fig. 2a; significant three-way interaction between
status, community, and gap on RGR; F = 11.14,
df = 1, 46, P = 0.002). Invasive plants responded
more to added Ca than did natives in the pine barrens
but not in hardwood forest sites in 2003 (significant
three-way interaction between status, community,
and Ca addition, Fig. 2b; F = 5.11, df = 1, 1134,
P = 0.024). In 2003 there were also marginally
significant two-way interactions on growth between
invasive/native status and N addition (F = 4.14,
df = 1, 46.6, P = 0.048) and a significant interaction
between invasive/native status and gaps (F = 8.56,
df = 1, 6.71, P = 0.023).
Although there was no overall difference between
native and invasive plants, families in 2003 differed
strongly in growth rates (F = 8.43, df = 6, 8.28,
P = 0.004). Invasive and native species from different families responded differently to N addition
(Fig. 3; there was a significant three-way interaction
between status, family and N; F = 2.52, df = 12,
1143, P = 0.003). Other than that, families did not
respond differently to resources or community type
(there were no other significant interactions between
family and other factors). To examine whether the
2003 results might have been an artifact of the
species paired by order rather than family, we
eliminated those pairs from the data and carried out
a re-analysis; results were qualitatively the same for
family and all interactions with family (i.e., factors
and interactions remained significant or not
significant).
Family was not tested as a factor in 2000 because
the design of that experiment was not set up to test it,
but we were curious about possible differences due to
species, since invasive status did not have a significant effect on growth. To examine differences
among species we analyzed the 2000 experiment
Effects of experimental manipulation of light and nutrients
Fig. 2 (a). Mean RGRs in 2003 for native and invasive plants
in hardwood forest and pine barrens experimental sites in
experimental gaps and under intact canopies. Means were
calculated on transformed data (natural log + 1) and then
back-transformed for graphing. The standard error for the
community 9 status 9 gap interaction was 0.18 on the transformed data. There was a significant three-way interaction
between status, community, and gap on RGR (F = 11.14,
df = 1, 46, P = 0.002). (b). Mean RGRs in 2003 for invasive
and native experimental plants in hardwood forest and pine
barrens experimental sites with and without added calcium.
Means were calculated on transformed data (natural log + 1)
and then back-transformed for graphing. The standard error for
the community 9 status 9 Ca interaction was 0.20 on the
transformed data. There was a significant three-way interaction
between status, community, and Ca addition (F = 5.11,
df = 1, 1134, P = 0.024)
with a second model, ignoring status and using
species as a factor. Due to the reduction in actual a
when re-analyzing data a second time, we used a
conservative cutoff (0.05/2, or P = 0.025, for significance; we could not adjust P values directly as in a
Bonferonni correction because the exact values for
the results where P was less than 0.0001 were not
available). This analysis showed strong differences
among species overall in growth rates, and large
differences among species in their responses to
resources (community, N, and light) which were
independent of their status as natives or invasives
(e.g., Fig. 4). There were highly significant differences
among species (F = 35.84, df = 6, 1219, P \
0.0001), and large differences among species in their
responses to N (species 9 N: F = 8.95, df = 6, 1219,
P \ 0.0001), community (species 9 community,
F = 6.26, df = 6, 1221, P \ 0.0001), gap (F =
9.72, df = 6, 1217, P \ 0.0001) and in the three-way
interactions between species, gap, and N (F = 3.27,
df = 6, 1217, P = 0.003), species, community, and N
(Fig. 4; F = 7.88, df = 6, 1220, P \ 0.0001) and
species, community, and gap (F = 8.83, df = 6, 1214,
P \ 0.0001).
Community, soil nutrients, and gaps
In 2000, growth did not differ overall between pine
barrens and hardwood forest communities (F = 0.15,
df = 1, 4.52, P = 0.72), nor was there any main effect
of either N or Ca addition. However, growth responses
to N differed between pine barrens and hardwood forests
(community type 9 N addition interaction, F = 5.39,
df = 1, 17.1, P = 0.033): N decreased growth in the
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J. Gurevitch et al.
Native
2.5
Invasive
Capri/Adox
2.0
RGR
1.5
Berberidaceae
1.0
Rham/Eleag
Rham/Eleag
Vitaceae
0.5
Capri/Adox
Rosaceae
{ Rosaceae
Aceraceae
Aceraceae
Berberidaceae
Vitaceae
0.0
Celastraceae
-0.5
Celastraceae
-1.0
Control
+N
Control
+N
Experimental treatment
Fig. 3 Mean RGRs in 2003 for native and invasive experimental plants belonging to different taxonomic groups with
and without added nitrogen. Full family names and species
used are indicated in Table 1; symbols for families or family
combinations are the same for both panels of the figure. Means
were calculated on transformed data (natural log + 1) and then
back-transformed for graphing. The standard error for the
status 9 family 9 N interaction ranged from 0.18 to 0.23 on
the transformed data. There was a significant three-way
interaction between status, family, and N (F = 2.52,
df = 12, 1143, P = 0.003)
Hardwood
3.0
Pine Barrens
2.5
V. novae-anglea
2.0
RGR
A. brevipedunculata
1.5
P. rigida
P. rigida
1.0
A. rubrum
E. umbellata
R. multiflora
A. platanoides
0.5
V. novae-anglea
A. brevipundiculata
R. multiflora
E. umbellata
A. rubrum
0.0
A. platanoides
-0.5
Control
+N
Control
+N
Experimental treatment
Fig. 4 Mean RGRs in 2000 for experimental plants belonging
to different species with and without added nitrogen in
hardwood forest and pine barrens experimental sites. Native
species are indicated by closed circles and names in normal
font, non-native invasive species by triangles and species
names in bold font. Species functional types, invasive/native
status, and families are listed in Table 1. There was a highly
significant interaction between species, community, and N
(F = 7.88, df = 6, 1220, P \ 0.0001)
hardwood forest sites and increased growth in the pine
barrens sites (Fig. 5a). Gaps had a strong effect on
growth in 2000 (F = 11.59, df = 1, 24.4, P = 0.002),
and there were larger responses to gaps in the denser
hardwood forest sites than in the more open pine barrens
(community type 9 gap interaction, Fig. 5b, F = 6.53,
df = 1, 20.8, P = 0.02).
In 2003, growth was greater in the pine barrens
than in the hardwood forest sites (community,
F = 31.88, df = 1, 6.37, P = 0.001) and better
overall in gaps than under intact canopies (gap,
F = 33.29, df = 1, 43.9, P \ 0.0001). Heavy rainfall
during extended periods damaged seedlings in the
hardwood forest sites, but had a lesser effect in the
pine barrens sites. As in 2000, there was a significant
community by gap interaction, with larger responses
to gaps in the hardwood forest areas than in the pine
barrens (F = 8.57, df = 1, 43.9, P = 0.005, means
123
Effects of experimental manipulation of light and nutrients
A
1.5
RGR
1.0
pine barrens
0.5
hardwood forest
0.0
-0.5
no N
+N
Fertilizer treatment
B
1.5
hardwood forest
1.0
RGR
pine barrens
0.5
0.0
-0.5
intact
gap
Experimental gap treatment
Fig. 5 (a). Mean RGRs in 2000 for experimental plants in
hardwood forest and pine barrens experimental sites with and
without added nitrogen. Means were calculated on transformed
data (natural log + 1) and then back-transformed for graphing.
The standard error for the community 9 N interaction was
0.22 on the transformed data. There was a significant
community type 9 N addition interaction (F = 5.39, df = 1,
17.1, P = 0.033). (b). Mean RGRs in 2000 for experimental
plants in hardwood forest and pine barrens experimental sites
in experimental gaps and under intact canopies. There was a
significant community type 9 gap interaction (F = 6.53,
df = 1, 20.8, P = 0.02)
not shown). No other main effects or interactions
were significant besides those already discussed for
the two experiments. The results for neither year
depended strongly on the initial estimates of size, as
results were qualitatively similar (i.e., same patterns
and same factors significant) when final plant size
alone was analyzed rather than growth.
Discussion
Light and soil nutrients are two fundamental factors
often considered to mediate the success of biological
invasions. By conducting experiments on the contributions and interactions of these factors to the early
establishment of seedlings of native and invasive
species, we hoped to better understand the relative
importance of light, nitrogen, and calcium in mediating the differential success of invasive plants in two
different forest community types. A larger goal was
to develop one approach for discerning general
patterns governing invasion in eastern forests that is
not strictly limited to conditions in any one year,
community, or for any one species or even pair of
species. Balancing the need for broader generalizations with species-specific and context-dependent
results is one of the most difficult and longstanding
challenges in ecological research, and this is especially true for interpreting the outcome of field
experiments. Distinguishing truly general results
from those that vary depending on temporal, ecosystem, or phylogenetic context requires both individual
multifactor experiments, and even more importantly,
a substantial body of experimental data. The results
presented here offer the beginning of a contribution
to that necessary data on plant invasions in forests.
Our results suggest that strong and consistent
relationships are likely between light availability and
the success of invasive taxa for the deciduous forests
of the eastern USA. While it has long been felt that
disturbance mediates the success of invasives in these
forests, it may be useful to move beyond this
generalization to recognize that factors that increase
light availability (e.g., canopy removal, fragmentation due to development) promote invasion in these
mesic hardwood forests. Why might invasives be
better able to respond to increases in light availability
in comparison to native plants? One possibility is
suggested by the differences in root-to-shoot ratios
that we found in one of the years between natives and
invasives, with invasives showing much greater
allocation to shoots on average. We hypothesize that
greater leaf area for a given amount of biomass, i.e., a
greater allocation to energy and carbon acquisition,
may allow greater responsiveness to increased light
for invasives in some years. In contrast to the context
independence of the results with light, the N and Ca
addition responses showed a great deal of context
dependence. Under certain conditions, increases in
nutrients favored invasives over natives, while under
others there were no discernible effects of nutrient
enrichment. Fertilization in different years and
123
J. Gurevitch et al.
different communities had different impacts, again
suggesting that the environmental background against
which the experiment was conducted contributed to
the results obtained. For example, N additions had the
counterintuitive effect of decreasing plant growth
rates in hardwood forests, while increasing growth
rates in pine barrens, a result that may be due to high
rates of herbivory on N-fertilized seedlings in hardwood forest sites (Leger et al. 2007). Most nutrient
addition studies are conducted in one or a small
number of environmental backgrounds and so do not
encounter this real-life complexity.
While it is often assumed that invasive species
have higher growth rates than natives (e.g., Rejmanek
and Richardson 1996; Grotkopp et al. 2002), we did
not find this to be the case for experimental seedlings
in either of the two years of field experiments in two
different forest community types; overall differences
between natives and invasives were inconsistent and
not statistically significant. One of the most striking
differences between them was the stronger response
of invasives to light gaps in the richer soil of the
hardwood forests in 2003 (Fig. 2a). This experimental outcome meshes very well with the patterns of
plant invasion in northeastern forests: pine barrens
communities have very low levels of invasion, while
hardwood forests are often heavily invaded, particularly in gaps (e.g., Howard et al. 2004). One possible
explanation for this discrepancy between our results
and the assumption that invasive plants grow more
rapidly is that, rather than having consistently greater
growth rates in the field, perhaps invasive species
have higher maximum growth rates under optimal
conditions.
It seems very likely that opportunities for invasion
may differ from year to year, with strong responses of
invaders possible only in some years and under some
environmental conditions (Johnstone 1986; Davis
et al. 2000). Conducting experiments in two different
years suggests the reality of ephemeral opportunities
for invasion that would not be apparent in an
experiment conducted only once, in a single year.
There were consistently strong differences in average
growth rates among individual taxa: families (in
2003) and species (in 2000) differed greatly in RGRs.
There were also striking differences among taxa in
their responses to environmental conditions, including both experimentally manipulated resources and
habitat differences. Because of these differences
123
among species and families, opportunities for invasion may not be consistent across all invasive species
in any given year: conditions in a particular year and
with a particular set of environmental conditions that
favor the establishment of Celastrus orbiculata may
be neutral or unfavorable for Berberis thunbergii in
these forests, for instance.
No other experiments on woody plant invasions in
forests that we are aware of have been conducted in
which both above and belowground environmental
factors were manipulated and a suite of native and
invasive species’ responses tested. Mechanistic studies on invasions of herbaceous species in grasslands
have been conducted more frequently (e.g., Knops
et al. 1999; Naeem et al. 2000; Chambers et al.
2007), in contrast to the paucity of such experiments
on woody species in forests. Soil nutrient levels have
been found to be important in facilitating grassland
invasions in a number of systems (Daehler 2003;
Prober et al. 2005). In contrast, our results show that
invasives in forests appear to respond more strongly
and more uniformly to light than to any other
environmental factor, including soil nutrient levels.
Sanford et al. (2003) reported no differences in the
responses of invasive and natives in response to light
availability in woody species in a northeastern forest;
all species in that study responded strongly to light. A
future challenge for studies in forest ecosystems will
be to try and elucidate the relationships between light
and nutrients in mediating invasions in other forest
systems (e.g., Burke and Grime 1996; Davis et al.
2005). At a larger scale, it will be critical to try and
understand how global-scale environmental changes,
such as those in temperature and carbon dioxide
concentrations, interact with changes in resources
such as light and nutrients in facilitating biological
invasions.
Acknowledgements Funding from the US Environmental
Protection Agency Star Grant number EPA828900010 and the
U.S. Fish and Wildlife Service for an Upton Preserve Research
Grant is gratefully acknowledged. We thank the University of
Delhi and Professor Inderjit for sponsoring a symposium in
Delhi, India which stimulated the synthesis needed to complete
this paper. We appreciate the hard work of Laura Hyatt, Kerry
Brown, Andrea Green, Jonathan Hickman, Susan Natali, Sarah
Uihlein, Leslie Gonzalez, Annie Hauff, Tag Engstrom, Dustin
Brisson, J. Matthew Hoch, Windsor Aguirre, Dan Stoebel, Bengt
Allen, Catherine McGlynn, Christopher Noto, Erin Cooper, Paul
Bourdeau, Wei Fang, and Mike Axelrod in the greenhouse,
laboratory, and field. We thank the Nature Conservancy, Suffolk
County Parks Department, and the New York State Department
Effects of experimental manipulation of light and nutrients
of Environmental Conservation for permission and access to
field sites. Comments of an anonymous reviewer substantially
improved the clarity of the paper.
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