Download Long-term dynamics of biotic and abiotic resistance to exotic species

Ecological Applications, 21(6), 2011, pp. 2105–2118
Ó 2011 by the Ecological Society of America
Long-term dynamics of biotic and abiotic resistance to exotic
species invasion in restored vernal pool plant communities
SHARON K. COLLINGE,1,2,4 CHRIS RAY,1
AND
FRITZ GERHARDT3
1
2
Department of Ecology and Evolutionary Biology, 334 UCB, N122 Ramaley,
University of Colorado, Boulder, Colorado 80309-0334 USA
Environmental Studies Program, University of Colorado, Boulder, Colorado 80309-0397 USA
3
Beck Pond LLC, 394 Beck Pond Road, Newark, Vermont 05871 USA
Abstract. Invasion of native ecosystems by exotic species can seriously threaten native
biodiversity, alter ecosystem function, and inhibit conservation. Moreover, restoration of
native plant communities is often impeded by competition from exotic species. Exotic species
invasion may be limited by unfavorable abiotic conditions and by competition with native
species, but the relative importance of biotic and abiotic factors remains controversial and
may vary during the invasion process. We used a long-term experiment involving restored
vernal pool plant communities to characterize the temporal dynamics of exotic species
invasion, and to evaluate the relative support for biotic and abiotic factors affecting invasion
resistance. Experimental pools (n ¼ 256) were divided among controls and several seeding
treatments. In most treatments, native vernal pool species were initially more abundant than
exotic species, and pools that initially received more native seeds exhibited lower frequencies of
exotic species over time. However, even densely seeded pools were eventually dominated by
exotic species, following extreme climatic events that reduced both native and exotic plant
densities across the study site. By the sixth year of the experiment, most pools supported more
exotics than native vernal pool species, regardless of seeding treatment or pool depth.
Although deeper pools were less invaded by exotic species, two exotics (Hordeum marinum and
Lolium multiflorum) were able to colonize deeper pools as soon as the cover of native species
was reduced by climatic extremes. Based on an information-theoretic analysis, the best model
of invasion resistance included a nonlinear effect of seeding treatment and both linear and
nonlinear effects of pool depth. Pool depth received more support as a predictor of invasion
resistance, but seeding intensity was also strongly supported in multivariate models of
invasion, and was the best predictor of resistance to invasion by H. marinum and L.
multiflorum. We conclude that extreme climatic events can facilitate exotic species invasions by
both reducing abiotic constraints and weakening biotic resistance to invasion.
Key words: abiotic constraints; biotic resistance; California; ephemeral wetlands; exotic species;
grasslands; Hordeum marinum; invasion; Lolium multiflorum; vernal pools.
INTRODUCTION
Invasion of native ecosystems by exotic species is a
leading cause of the loss of native biodiversity (Wilcove et
al. 1998, Lockwood et al. 2007). Moreover, the conservation and restoration of native plant communities is
often impeded by exotic species invasion (Suding et al.
2004). Ecological research regarding exotic species
invasion has focused primarily on the roles of resource
availability, productivity, disturbance, and native species
diversity in determining the vulnerability of native
communities (Fox and Fox 1986, Crawley 1987,
Rejmánek 1989, Ewel et al. 1999, Mack et al. 2000).
These factors may be categorized into two distinct classes
of mechanisms: abiotic constraints, which include physManuscript received 30 May 2010; revised 23 December
2010; accepted 19 January 2011. Corresponding Editor: R. A.
Hufbauer.
4 E-mail: [email protected]
ical conditions that limit performance of exotics; and
biotic resistance, whereby native communities compete
with exotic species for limiting resources. Abiotic
constraints are important when exotic species lack the
morphological and physiological traits required to
tolerate the abiotic conditions found in less invaded
communities (Fox and Fox 1986, Rejmánek 1989). Many
studies have shown that exotic species richness and
abundance generally decrease with increasing abiotic
stress (Forcella and Harvey 1983, D’Antonio 1993, King
and Grace 2000, Harrison et al. 2001, Holway et al. 2002,
Williamson and Harrison 2002, Gerhardt and Collinge
2003, Seabloom and van der Valk 2003, 2007).
Alternatively, exotic species may be excluded from less
invaded communities by competition from native species
assemblages. As reviewed in Mack et al. (2000), numerous
studies have addressed the roles of species and functional
diversity, species identity, and productivity in determining community invasibility. Despite many theoretical and
empirical studies of exotic species invasion, there is still no
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consensus regarding the relative importance of biotic and
abiotic constraints in determining the invasion resistance
of native communities.
Although abiotic constraints and biotic resistance
have typically been considered independently, they may
act synergistically and they may also be influential at
different points in time during the invasion process
(D’Antonio 1993, Dietz and Edwards 2006). Dietz and
Edwards (2006) suggested that invasions may occur in
distinct phases: a primary phase, in which exotics
increase rapidly in response to disturbances and/or
resource enrichment, and a secondary phase, in which
exotics respond to new ecological circumstances. In one
of the few long-term studies of invasion, Wiser et al.
(1998) found that the invasion process was controlled
initially by dispersal limitation and later by other biotic
and abiotic factors, especially those related to resource
abundance and species richness. Abiotic constraints and
biotic resistance may also interact in complex ways. For
example, the strength of competition may differ among
sites along abiotic gradients (Grime 1973, 1979; but see
Wiens 1977), or the relative impact of abiotic constraints
and biotic interactions may vary along an environmental
gradient (Connell 1961, Gurevitch 1986, Hacker and
Bertness 1999, Greiner La Peyre et al. 2001). The few
studies that have examined interactions between abiotic
and biotic controls on invasibility in plant communities
have shown that the outcomes of competitive interactions among native and exotic species depend on abiotic
conditions (e.g., David 1999), and that the relative
importance of abiotic constraints and biotic resistance
may vary among sites and among native communities
depending on the life stage of the invading species (e.g.,
D’Antonio 1993).
Our long-term study of restored vernal pool plant
communities in California, USA, has allowed us to
investigate the temporal dynamics of invasion and to
evaluate the relative importance of biotic resistance and
abiotic constraints governing invasion since 2002.
Vernal pools are spatially discrete, ephemeral wetlands
that occur in a variety of ecosystems. In Mediterranean
climates, distinct wet and dry seasons lead to winter and
spring inundation of pools, followed by complete drying
in summer (Keeley and Zedler 1998, Grillas et al. 2004,
Deil 2005). In California, these seasonal wetlands are
characteristically underlain by an impervious hardpan
or claypan (Holland and Jain 1988). Relative to other
habitats, vernal pools harbor a disproportionate number
of taxa characteristic of the California flora (Stebbins
1976, Zedler 1987, Keeley and Zedler 1998). The
extreme variation in physical conditions within these
pools, combined with the relative isolation of these
patchily distributed habitats, has resulted in the
development of unique communities of plants and
animals. For example, California’s vernal pools harbor
.60 endemic species of plants and invertebrates
(Holland 1976, Stone 1990, Witham 1998), representing
an important component of California’s native biolog-
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Vol. 21, No. 6
ical diversity (Jokerst 1990, Pollak and Kan 1998). Due
to destruction of vernal pools from agricultural and
residential development, many species restricted to
vernal pools in California are currently categorized as
rare or endangered (Witham 1998). Thus, protection
and restoration of native plant and animal populations
and communities are critical aspects of vernal pool
conservation within the region.
In previous research, we compared patterns and
mechanisms of plant community development and exotic
species invasions in naturally occurring and restored
vernal pools (Gerhardt and Collinge 2003, 2007, Collinge
and Ray 2009). Because most exotic and native species in
these pools are annuals, studying the year-to-year
variation in species abundance and distribution can
reveal the relative importance of factors that directly
influence population dynamics. We observed that, in
general, exotic species were strongly and negatively
affected by the relatively harsh abiotic conditions
associated with pool inundation during the winter wet
phase. Most exotic species were unable to tolerate
extensive periods of inundation, and therefore, their
abundances in the field were negatively correlated with
pool depth (Gerhardt and Collinge 2003). Greenhouse
experiments in which water depth was experimentally
controlled confirmed that inundation significantly affected most measures of exotic plant performance and played
the dominant role in determining invasibility in vernal
pools (Gerhardt and Collinge 2007). Both field and
greenhouse studies suggested that there were relatively
slight, though detectable, influences of native species
cover on exotic species occurrence. In field surveys of
naturally occurring and restored pools, the abundances of
individual exotic species were more often related to water
depth than to the cover of native species (Gerhardt and
Collinge 2003). These studies offered meaningful insights
into short-term patterns and processes of exotic species
invasion, as they focused predominantly on recently
reconstructed vernal pools (Gerhardt and Collinge 2003)
and community dynamics within a single growing season
(Gerhardt and Collinge 2007).
We now have data from a long-term experiment in
vernal pool restoration, including seven years of data on
plant community dynamics in pools that were initially
subject to different native plant seeding treatments
(Collinge and Ray 2009). This seven-year time series
allows us to more fully assess the temporal dynamics of
invasions under varying abiotic and biotic conditions,
during a critical phase in community development. In
established communities, there may be less opportunity
to observe the relative importance of biotic and abiotic
factors because some constraints have been alleviated by
the very changes (e.g., species turnover) that result in
community establishment. Our data address the need for
information on how exotic species invasion proceeds in
restored communities, for practical applications in
ecological restoration.
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EXOTIC SPECIES IN RESTORED VERNAL POOLS
In the present study, we addressed five questions
regarding exotic species invasions in restored vernal
pools: (1) Is there evidence for biotic resistance to exotic
species invasions? (2) Are restoration trajectories divergent for experimental pools that were initially seeded
with vernal pool species vs. those that were not seeded?
(3) Are exotic species invasions constrained by abiotic
factors, such as pool depth? (4) Do the two dominant
exotic species in this system, Lolium multiflorum and
Hordeum marinum, differ from other exotic species in
their responses to pool depth and seeding treatment? (5)
Are long-term dynamics well predicted by results from
short-term experiments? In addition to answering these
qualitative questions, we used information theory to
quantify the relative contribution of biotic and abiotic
factors affecting invasion. Previous studies of this
system suggest that biotic resistance to exotic species
should be related to S, the number of native seeds sown
in a pool, while abiotic constraints on invasion are
largely related to D, a pool’s maximum depth of
inundation (Gerhardt and Collinge 2003, 2007,
Collinge and Ray 2009). Using S and D as predictor
variables, we developed a set of candidate models to
explain invasion in each pool. We then used an
information criterion to calculate the weighted average
support for each predictor across models, providing a
quantitative result that can be used as prior information
in subsequent analyses of biotic vs. abiotic controls on
invasion in this system.
METHODS
Study site
We conducted this study at Travis Air Force Base
(AFB), Solano County, California, USA. Travis AFB is
located in the Sacramento Valley near Fairfield,
California (38815 0 0000 N, 122800 0 0000 W; 6 m elevation).
This area receives ;50 cm of rain per year, concentrated
in the winter wet season from December to April. There
is typically no measurable precipitation from May to
November.
Approximately 100 naturally occurring vernal pools
occur in the 15-ha study area, embedded in a grassland
matrix recently used for grazing and cultivation agriculture
(BioSystems Analysis 1994). The site was leveled over 50
years ago, using simple tools to eliminate the basin-andmound topography typical of vernal pool habitats. This
treatment did not destroy the hardpan clay layer that
supports vernal pools in this region. We have identified 76
plant species in vernal pools at this study site during the
past ten years (S. K. Collinge, unpublished manuscript).
Nearly half (35 of the 76 species, or 46%) of these species
are ‘‘exotics’’ not native to California, including Bromus
hordeaceus, Centaurea solstitialis, Cotula coronopifolia,
Erodium botrys, Hordeum marinum, and Lolium multiflorum (nomenclature follows Hickman [1993]). About
one-third (26 of the 76, or 34%) are ‘‘vernal pool’’ or
‘‘native’’ species, considered to be native indicators or
associates of vernal pools (Keeler-Wolf et al. 1998),
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including Deschampsia danthonioides, Downingia concolor,
Eryngium vaseyi, Lasthenia conjugens, Layia chrysanthemoides, Plagiobothrys stipitatus, Pleuropogon californicus,
and Veronica peregrina. The remaining 20% (15 of the 76)
are ‘‘other’’ species native to California, but not necessarily
restricted to vernal pool habitats; e.g., species described as
‘‘generalists’’ by Keeler-Wolf et al. (1998).
In early December 1999, we constructed 256 ‘‘experimental’’ pools in a regular grid bisected by an aircraft
landing strip. Many of these natural vernal pools line the
sides of this small landing strip. Experimental pools were
constructed on the same soils as natural pools in this
area (Antioch-San Ysidro complex and San Ysidro
sandy loam), were similar in elevation and topography
(0–2% slopes), and were located in close proximity
(150 m) to naturally occurring pools. Each experimental pool was located ;10 m from the nearest experimental pool, a typical distance between naturally
occurring pools at this study site. Experimental pools
were constructed in three size classes (25, 50, and 100
m2; n ¼ 85 each), to bracket the median area (78 m2) of
natural pools within the study site.
Using seeds collected from natural pools at the site,
each experimental pool received one of five seeding
treatments, as detailed in Collinge and Ray (2009). (1)
‘‘LACO-1x’’ pools received 100 seeds of L. conjugens in
1999. (2) ‘‘LACO-3x’’ pools received 100 seeds of L.
conjugens in each year 1999–2001 (300 seeds total). (3)
‘‘Group-A’’ pools received 300 seeds in 1999 (100 seeds
each of Group-A species L. conjugens, D. danthonioides,
and E. vaseyi ), followed by 300 seeds in 2000 (100 seeds
each of Group-B species L. conjugens, L. chrysanthemoides, and P. stipitatus). (4) ‘‘Group-B’’ pools received
Group-B species in 1999 followed by Group-A species in
2000. Within each pool, a single 0.50 3 0.50 m plot was
subjected to a seeding or control treatment, and each
treatment plot was permanently marked for long-term
observation. Seeds were mixed with a tablespoon of fine
sand and sown over lightly raked plots. (5) Control
pools were subject to a similar disturbance involving
raking and ‘‘sowing’’ of seed-free sand. Of the 256
constructed pools, 249 received the assigned treatment
and were analyzed for this study.
Field data collection
Because seasonal inundation is a defining property of
vernal pools, we measured pool flooding during three
wet seasons (limited financial resources precluded
measurements in additional years). At weekly intervals
during December–April, we measured water depth at the
center of each pool’s marked treatment plot during two
years early in the study (1999–2000 and 2001–2002) and
one year late in the study (2008–2009). From these data,
we calculated MiY, the maximum water depth recorded
within pool i during the wet season ending in year Y,
and Di, the mean maximum water depth in pool i. To
avoid a temporal bias in Di, we first averaged Mi2000 and
Mi2002 to obtain a single set of depths measured ‘‘early’’
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SHARON K. COLLINGE ET AL.
in the study that were then averaged with the single set
of depths measured many years later. In equation form,
Di ¼ 0.5(Mearly þ M2009), where Mearly ¼ 0.5(M2000 þ
M2002). Because each plot was positioned within a
relatively deep portion of its pool, plot and pool depth
have been closely related. For example, weekly measurements in 2001–2002 demonstrated no significant
difference between maximum water depth at the plot
center vs. at the pool center (paired t test, t ¼ 0.283, df ¼
236, P ¼ 0.778).
We measured plant occurrence in experimental pools
during the flowering phase (April–May) each year
during 2000–2008. We placed a square frame (0.50 3
0.50 m ¼ 0.25 m2) divided into 100 subquadrats (each 5
3 5 cm) over the treatment plot in each pool and
recorded the frequency (number of subquadrats out of
100 in which the species occurred) of each species
present in each year. Plants were identified to species
using Hickman (1993), and a voucher collection is
maintained in the Collinge Laboratory at the University
of Colorado, Boulder, Colorado, USA.
Data analyses
Most of our analyses focused on the relative cover of
two species classes: exotic species and species native to
California vernal pools. We characterized the ‘‘cover’’ of
each species class as the sum of the frequencies of all
species in that class. We compared the cover of each
species class across years and seeding treatments using
95% confidence intervals on mean cover, and using box
plots showing medians, quartiles, range, and outliers. In
box plots, notches on the sides of each box represent
bounds on the median, such that there is ‘‘strong
evidence’’ that two medians differ if their notches do
not overlap (Chambers et al. 1983:62). The initial (2002)
difference in exotic species cover between pools sown
with 0–100 seeds and those sown with 300–600 seeds
was determined using Welch’s two-sample t test.
Theories of resource competition predict that the
abundance of one species will decline with the abundance of a competitor, such that interspecific competition results in a negative correlation in the abundances
of species competing within replicate habitats. Similarly,
we sought evidence for interclass competition between
native and exotic species by analyzing the relative cover
of each class within pools. Because we did not expect
strong competition in sparsely vegetated pools, we used
quantile regression to detect and model a key aspect of
competition. We used the 95th regression quantile to
model an upper bound on the cover of native species in
our plots as a function of increasing exotic species cover,
interpreting a negative slope of regression as evidence
for competition. Interannual variation in this model was
investigated using 95% confidence bounds on each
annual slope and intercept of regression. Using the
same regression framework, we also investigated the
converse effect of native species cover on the upper
bound of exotic cover.
We quantified the relative support for biotic and
abiotic controls on invasion using an informationtheoretic approach (Burnham and Anderson 2002).
Our response variable (y) was the temporal (2002–
2008) average cover of exotic species (e), alone or in
combination with cover of native vernal pool species (v).
We considered three indices of invasion: y ¼ e, y ¼ e v,
and y ¼ e/(e þ v). The latter two indices include useful
but potentially confounding information on the relative
cover of native species. In each case, direct effects of
seeding on the cover of native species may vie with
indirect effects (like biotic resistance) to complicate the
interpretation of any relationship between invasion and
explanatory variables. Here, we focused on y ¼ e, while
also summarizing results for the other two indices.
Assuming that the success of exotic species in this
system should be affected by the number of native seeds
sown (S ) and the maximum depth of inundation (D)
within each pool i, we developed a set of candidate
models of the form yi ¼ f(Si, Di ). (We used a similar
approach to model the frequency of the dominant exotic
species.) Di was the temporally weighted average of
maximum water depths, and Si was coded as the number
of native seeds sown within the treatment plot of pool i,
divided by 100. Candidate models included linear and/or
nonlinear effects of either (or both) S and D. The
evidence in support of each fitted model was compared
using an information criterion (AICc). AICc discounts
model likelihood by the number of model parameters
and is appropriate for ranking non-nested models
(Burnham and Anderson 2002). The model with lowest
AICc is best supported by the data, and the positive
difference between AICc values (DAICc) was used as a
metric of relative support. Following Burnham and
Anderson (2002), we assumed that models with DAICc
, 2 were well supported by the data, while models with
DAICc . 4 were poorly supported. DAICc values were
used to generate an Akaike weight for each model, a
scaled metric of relative support that can be summed
across models to compare the relative support for
individual predictors (Burnham and Anderson 2002),
such as S and D. Analyses and plots were coded in R
version 2.6.2 (R Development Core Team 2008).
RESULTS
Although vernal pool species were initially more
abundant than exotic species in these restored pools,
exotics dominated this system by 2007 (Fig. 1). ‘‘Other’’
species comprised ,10% of plant cover in each year, and
were omitted from analyses.
Biotic effects
The initial dominance of vernal pool species was
driven by results from seeded pools. Pools sown with
more native seeds and more native species generally
exhibited lower frequencies of exotic species than pools
that were unseeded or received seeds of only one native
species (Fig. 2). For Group-A and Group-B pools, these
September 2011
EXOTIC SPECIES IN RESTORED VERNAL POOLS
FIG. 1. Temporal progression, 2002–2008, of vernal pool
species cover (dark gray) and exotic species cover (light gray) in
permanent study plots within a large number of experimental
vernal pools (n ¼ 119–249). The study area was on the Travis
Air Force Base, California, USA. The cover of each species
class within each pool was the percentage of sub-quadrats in
which the species occurred (out of 100 sub-quadrats in a 100cm2 sampling quadrat), summed over all species in the class.
Dark horizontal lines represent medians, boxes contain upper
(75%) and lower (25%) quartiles, and dashed whiskers contain
the range of the data excepting outliers. Outliers (lying outside
the box by more than 1.5 times the interquartile range) are
represented by open circles.
effects of seeding were significant in early years of the
experiment (Fig. 2a, b). In contrast, initial cover of
exotic species was higher in pools sown with few or no
native seeds (Fig. 2d, e). Specifically, mean cover of
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exotics in 2002 averaged 89.25 in control and LACO-1x
pools, and 70.90 in other pools (Welch’s two-sample t
test, t ¼ 2.199, df ¼ 224.583, P ¼ 0.014). However, even
densely seeded Group-A and Group-B pools were
dominated by exotic species in 2008, just seven years
after the final seeding treatment (Fig. 2a, b).
In both seeded and control pools, the combined cover
of exotic and vernal pool species was limited. The scatter
plots in Fig. 3 reveal the cover achieved by every
community, from those comprised of only vernal pool
species (points near the vertical axis in each graph),
through mixed communities (central points), to those
comprised of only exotic species (points near the
horizontal axis). The lack of points in the upper right
quadrant of each graph demonstrates a lack of
communities with high frequencies of both exotic and
vernal pool species. Seeded and control pools exhibited
similar temporal patterns in community structure.
Statistical evidence for this similarity derived from our
models of maximum cover. When fitted separately to
data from seeded vs. control pools, there was extensive
overlap in confidence intervals for model parameters;
e.g., the 95% CI on the slope (or intercept) of the
quantile regression for seeded pools overlapped the
fitted slope (or intercept) of the quantile regression for
control pools in almost every year. For this reason, we
report only on models of the combined data (dashed
lines in Fig. 3). In each year, the slope of the 95th
regression quantile relating the cover of native to exotic
species was significantly negative (Table 1). Slopes were
significantly more negative during 2004–2005, when
FIG. 2. Invasion trajectories, 2002–2008, following different seeding treatments in experimental vernal pools. The cover of each
species class within each pool was the percentage of sub-quadrats in which the species occurred (out of 100 sub-quadrats in a 100cm2 sampling quadrat), summed over all species in the class. Mean cover was calculated by averaging cover over all study pools
within a treatment. Vertical lines represent 95% confidence intervals on mean cover. See Methods: Study site for seeding treatment
descriptions.
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FIG. 3. Time series of community composition in experimental vernal pools. Scatter plots relate the total cover of native species
to the total cover of exotic species in each pool. The cover of each species class within each pool was the percentage of sub-quadrats
in which the species occurred (out of 100 sub-quadrats in a 100-cm2 sampling quadrat), summed over all species in the class. Total
species cover was as limited in pools seeded with native species (open circles) as in unseeded control pools (solid circles). Upper
bounds on total species cover were estimated using 95th regression quantiles (dashed lines), as summarized in Table 1.
TABLE 1. Results of regression for the 95th quantile of each annual relationship between the cover of vernal pool species and
exotic species (Fig. 3).
Intercept
Slope
Year
Fitted
value
Lower
95% CI
Upper
95% CI
Fitted
value
Lower
95% CI
Upper
95% CI
n
2002
2003
2004
2005
2006
2007
2008
287.85
318.25
260.45
250.00
198.00
225.42
142.04
275.05
309.42
248.84
232.95
184.06
202.32
135.77
339.01
331.60
277.73
260.83
204.94
239.04
173.66
0.8305
0.8375
1.3495
1.1905
0.9565
0.9015
0.6040
1.0394
0.9025
1.4527
1.2480
0.9800
1.0494
0.7673
0.6502
0.6547
1.2765
1.0951
0.9306
0.6450
0.5891
246
249
233
248
244
119
236
Note: Each regression models an upper bound on the summed frequencies of species in the community. Variation in sample size
(n) represents variation in sampling effort (2007 only) and occasional lack of access to certain sampling plots.
September 2011
EXOTIC SPECIES IN RESTORED VERNAL POOLS
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communities with relatively high cover of exotic species
were lost (Fig. 3c, d).
Abiotic effects
Results from quantile regression of native cover on
exotic cover included variation in intercepts among
years (Table 1). Significant reductions in intercept values
suggested three phases of decline in the resources
available to this community: 95% confidence intervals
did not overlap for sequential pairs of intercept values
estimated for 2003–2004, 2005–2006, and 2007–2008
(Table 1). This pattern of decline was robust to a
reversal of response and predictor values. Intercepts for
95th quantile regression models of maximum exotic
cover fell from 246 to 175 between 2002 and 2008 (95%
CI ¼ 219–271 for 2002, and 163–207 for 2008).
Results summarized in Fig. 3 and Table 1 indicate that
the upper bound on total species cover in these communities declined over time in the following sequence: In 2002,
maximum cover was similar in all communities, regardless
of whether native or exotic species dominated each
treatment plot (the slope of the 95th regression quantile
was not significantly different from1.0; Table 1). By 2004,
maximum cover had declined in all communities, but
especially in those dominated by exotic species (the
intercept had declined significantly and the slope was
significantly below 1.0). In 2006, 2007, and especially
2008, maximum cover declined mainly in pools dominated
by native species (intercepts declined significantly in 2006
and 2008, and slopes rose above 1.0). Between 2006 and
2008, many communities appeared to ‘‘leap’’ from the
vertical axis toward the horizontal axis in Fig. 3,
representing a shift from the dominance of native species
to the dominance of exotics.
The dramatic shift in community composition after
2005 corresponded with two extreme climatic events.
Extreme flooding occurred during the growing season of
2006 (Fig. 4a), when winter (October–April) precipitation in nearby Fairfield, California (;5 km from our
study site) was 104.32 cm, more than two standard
deviations above the long-term mean (for 1950–2009,
55.97 6 22.33 cm, mean 6 SD; available online).5 High
precipitation occurred both early (December) and late
(March) in the growing season of 2006 (Fig. 4b), and
was followed by declines in both native and (especially)
exotic species cover, as measured in April (Figs. 1 and
2). The density of points near the vertical axis in Fig. 3e
indicates that exotic species cover was very low in many
pools during 2006. Subsequent drought during the
growing season of 2007 (Fig. 4a) corresponded with
continued depression of native species cover, as the
performance of vernal pool species was hampered by a
total lack of precipitation in January and March of this
year (Fig. 4b). In contrast with native species, exotic
5
hhttp://www.wrcc.dri.edui
FIG. 4. Precipitation recorded at Fairfield, California (5
km from the study site). (a) Precipitation during the wet season
(1 October–30 April) of 2006 was much higher than the longterm (1950–2008) mean, while precipitation during the wet
season of 2007 was well below the mean. (b) Excess
precipitation during 2006 occurred first and foremost in
December, and the drought of 2007 began in January.
species cover began to rebound during the drought year
of 2007 (Figs. 1, 2, and 3f ).
Pool depth accounted for much of the variation
among pools in the average cover of exotic species, and
there was strong evidence for nonlinearity in this
relationship (Fig. 5a). The best model was quadratic,
but a saturating model garnered similar support (DAICc
¼ 0.60) and a logistic model was nearly supported
(DAICc ¼ 2.01 for a model with fitted upper and lower
asymptotes). A linear model was not supported (DAICc
¼ 16.28). Quadratic curves were also best for modeling
effects of pool depth on invasion data from each
individual year (not shown), and for modeling other
2112
SHARON K. COLLINGE ET AL.
Ecological Applications
Vol. 21, No. 6
FIG. 5. Effects of inundation and sowing native seed on the invasion of vernal pool plant communities by exotic species. The
mean cover of exotic species for each pool is the summed frequencies of exotic species within the study plot, averaged over the
period 2002–2008. Frequency was the percentage of sub-quadrats in which the species occurred, out of 100 sub-quadrats in a 100cm2 quadrat. (a) The best relationship between exotic species cover and water depth was quadratic (shown) or saturating (DAICc ¼
0.60). (b) Stronger biotic resistance was offered by pools that received more seeds of vernal pool species. This relationship appeared
quadratic, but a linear decline in invasion index was also well supported (DAICc ¼ 0.39 and 6.21 for linear and null models,
respectively). (c) The best model of invasion included a linear effect of pool depth and quadratic effects of pool depth and seeding
treatment, but a similar model with linear effect of seeding garnered similar support (see Table 2). The linear fit between observed
and modeled data (solid line) rises along the diagonal (dashed line).
indices of invasion. For both y ¼ e/(e þ v) and y ¼ e v,
quadratic models garnered support similar to the best
logistic models (DAICc ranges were 0.77–1.45 for
quadratic models, 4.74–5.92 for saturating models, and
22.41–24.53 for linear models). Across all indices of
invasion, quadratic models reached their minima (lowest
average indices of invasion) at annual maximum water
depths averaging 22–25 cm.
Relative support for biotic vs. abiotic resistance
Exotic species cover was lower in pools that received
more seeds of vernal pool species (Fig. 5b). The best
model of this treatment effect was quadratic with a peak
at zero (unseeded control pools), but a linear effect of
seeding was also supported (DAICc ¼ 0.39). Results were
similar for all three indices of invasion (quadratic
models garnered best support and DAICc ranged 0.39–
1.77 for linear models, 6.21–19.52 for null models).
Our model selection analysis suggested strong support
for effects of both pool depth (D) and number of native
seeds sown (S ) on invasion resistance in these vernal
pools. Again, results were qualitatively identical and
quantitatively similar among invasion indices. The best
model of exotic species cover (y ¼ e) included a negative
linear effect of pool depth and nonlinear effects of both
seeding treatment and pool depth (y ¼ 155.85 0.52S2 10.16D þ 0.21D2). This same model also received highest
support in analyses of y ¼ e v and y ¼ e/(e þ v), where
its relative support was even higher (w ranged from 0.55
to 0.62) than in the analysis reported in Table 2 (w ¼
0.46). A model with similar support included the same
predictors excepting a linear, rather than nonlinear,
September 2011
EXOTIC SPECIES IN RESTORED VERNAL POOLS
2113
TABLE 2. Relative support for alternative models of invasion resistance.
Model
y
y
y
y
y
y
y
y
y
;
;
;
;
;
;
;
;
;
b0
b0 þ b1S
b0 þ b1S2
b0 þ b1D
b0 þ b1S þ b2D
b0 þ b1D þ b2D2
b0 þ b1S þ b2D þ b3D2
b0 þ b1S2 þ b2D þ b3D2
b0 þ b1S þ b2S2 þ b3D þ b4D2
k
LL
AICc
DAICc
w
2
3
3
3
4
4
5
5
6
1317.10
1313.16
1312.97
1210.90
1202.28
1201.73
1190.96
1190.74
1190.73
2638.24
2632.42
2632.03
2427.90
2412.73
2411.62
2392.17
2391.73
2393.80
246.51
240.69
240.30
36.17
21.00
19.89
0.44
0.00
2.07
0.0000
0.0000
0.0000
0.0000
0.0000
0.0000
0.3719
0.4634
0.1646
Notes: The mean cover of exotic species, y ¼ e, is modeled as a linear or nonlinear function of seeding treatment (S ) and
maximum water depth (D) in each pool (see Methods: Study site for seeding treatment descriptions). Each bx represents a fitted
intercept or coefficient of regression. For each model, n ¼ 249. Relative support (Akaike weight, w) for each model is derived from
an information criterion (AICc), which balances a metric of model fit (LL ¼ log-likelihood) against the number of fitted parameters
(k). Better models have higher values of w and lower values of DAICc, the difference between AICc for a given model and the lowest
AICc. The best model is indicated in boldface.
effect of seeding (Table 2). The only other model with
some support was the full model; however, its added
parameter did little to improve model fit (cf. loglikelihood values in Table 2) and garnered a penalty
(DAICc ¼ 2.07 ’ 2) almost exactly as high as expected
for the addition of an uninformative parameter
(Burnham and Anderson 2002).
Using Akaike weights to compare support for
individual predictors, linear and nonlinear effects of
water depth were relatively strong, followed by nonlinear effects of seeding (Table 3). The order of relative
support for each predictor was identical across indices of
invasion, but the absolute support for S2 was elevated in
analyses of y ¼ e v and y ¼ e/(e þ v) due to lower model
uncertainty in those analyses. Our candidate models,
chosen a priori based on previously observed effects of S
and D, did not equalize the number of models (m)
containing each variable, and considering all possible
models is not sensible in this study. However, equal
representation is easily achieved by omitting the three
models in the middle of Table 2, and is nearly achieved
by omitting the fourth and seventh models. In the first
case, m ¼ 3 for each predictor and the support for each is
unchanged (to four decimal places) relative to wp values
in Table 3. In the second case, m ¼ 3 for all but D (m ¼
4), and the only appreciable change is a rise in relative
support for S2 (wp ¼ 0.16, 0.63, 0.63, and 0.63 for S, S2,
D, and D2, respectively). Only three of our candidate
models were strictly alternative (the second, third, and
fourth models in Table 2), with no overlap in predictor
variables. Among these univariate models, effects of
seeding were overwhelmed by effects of water depth (wp
’ 0.00, 0.00, and 1.00 for S, S2, and D, respectively).
Although pool depth reduced invasion by many exotic
species, two exotic species were able to invade pools of
any depth as soon as competition from vernal pool
species was alleviated. Hordeum marinum and Lolium
multiflorum were the dominant exotics within this
system, combining to account for over 29% of total
species cover across years, and over 51% of total species
cover in 2008 (Fig. 6a). The rise in frequency of these
two species followed the reduction in total species cover
after the extreme flooding in 2006, and the continued
reduction in vernal pool species cover after the extreme
drought in 2007. Both H. marinum and L. multiflorum
also exhibited relatively high tolerance for inundation in
this experiment. By 2008, both species were most
frequent in pools of intermediate depth (Fig. 6b, c), in
direct contrast with the general effect of pool depth on
exotic species invasion (Fig. 5a). Despite substantial
scatter in the data, there was strong support for a
quadratic effect of pool depth on the frequency of both
species (DAICc ¼ 18.34 for linear and 17.90 for null
models of L. multiflorum data; DAICc ¼ 3.21 for linear
and 9.46 for null models of H. marinum data).
Although we did not find support for typical effects of
pool depth on invasion by the dominant exotic species in
this system, we did find support for effects of biotic
resistance to invasion by these species. Mean frequencies
of L. multiflorum and H. marinum during 2002–2008
declined with the number of native seeds sown in a pool.
The best model of L. multiflorum frequency ( f ) was
quadratic ( f ¼ 40.10 0.19S2), with a peak at S ¼ 0.
There was also support for a simple linear decline in
frequency with number of seeds sown (DAICc ¼ 0.83),
and relatively low support for the null model (DAICc ¼
2.14). For H. marinum, the best model was linear ( f ¼
TABLE 3. Relative support for linear and nonlinear predictors
of exotic species invasion, given the set of invasion models
considered in Table 2.
Predictor
wp
m
S
S2
D
D2
0.54
0.63
1.00
1.00
4
3
6
4
Notes: Abbreviations are: S, seeding treatment; and D,
maximum water depth in pool. The support for a given
predictor (wp) is the sum of support (w in Table 2) for all models
that incorporate the given predictor. Relative values of wp were
insensitive to the apparent inequality in the number of models
(m) in which different predictors appear.
2114
Ecological Applications
Vol. 21, No. 6
SHARON K. COLLINGE ET AL.
FIG. 6. Trends in cover for the two dominant exotic species in this system. (a) For each year, the mean frequencies of Lolium
multiflorum (dark gray) and Hordeum marinum (light gray) are shown relative to the summed frequencies of all species occurring in
sampled pools. Frequency was the percentage of sub-quadrats in which the species occurred, out of 100 sub-quadrats in a 100-cm2
quadrat. Each vertical line represents one standard error. Both (b) L. multiflorum and (c) H. marinum achieved higher frequency in
pools of intermediate depth. The quadratic models (solid curves) shown garnered far more support than null models: DAICc values
for null models were 17.90 in panel (b) and 9.46 in panel (c).
12.21–1.00S ), but there was also support for a quadratic
decline in frequency with number of seeds sown (DAICc
¼ 0.86), and little support for the null model (DAICc ¼
9.67). For both species, similar treatment effects were
evident in every year of the study, but appeared
strongest in 2007. The frequency of L. multiflorum in
2007 averaged 36% lower in the more heavily seeded
Group-A and Group-B pools than in control pools. The
best model for L. multiflorum in 2007 was f ¼ 55.68–
0.59S2 (DAICc was 1.62 for a negative linear effect of
seeding, and 6.66 for the null model). The frequency of
H. marinum in 2007 averaged 78% lower in Group-A
and Group-B pools than in controls. The best model for
H. marinum in 2007 was f ¼ 32.45–4.10S (DAICc was
0.62 for a quadratic decline, and 12.29 for the null
model).
DISCUSSION
In summary, these long-term data from experimental
plant communities provide new insights into the roles of
biotic resistance, abiotic constraints, and disturbance
events in determining the long-term dynamics of
biological invasion. We found strong evidence for
pronounced effects of both biotic resistance and abiotic
constraints on exotic species invasions. Initially the
trajectories of native and exotic species in plots seeded
with native species differed from those in unseeded plots.
However, the occurrence of extreme flooding followed
by extreme drought dramatically impacted both native
and exotic species, ultimately allowing exotic species to
dominate nearly every plot. As detailed in the following
paragraphs, these results have important implications
for our understanding of biological invasions and also
suggest general guidelines for the conservation of native
September 2011
EXOTIC SPECIES IN RESTORED VERNAL POOLS
plant communities, including the biologically rich plant
communities associated with vernal pools.
Temporal dynamics of invasion
In control pools and pools seeded lightly with just one
native species, exotic species tended to dominate the
community throughout the experiment (Fig. 2d, e). In
pools seeded more heavily with native species, we
observed a temporal shift from dominance by native
species to dominance by exotic species (Fig. 2a–c). This
shift was facilitated by two years of extreme conditions
that appeared to release exotic species from competition
with native species in a two-step process. First, extreme
flooding reduced the frequency of all species, but
especially exotic species. Then, drought inhibited the
rebound of vernal pool species, which, unlike most
exotics, require substantial water resources for growth
and seed production (Bauder 1989, Zedler 1990, Collinge
et al. 2003). Thus, ironically, the importance of biotic
resistance and abiotic constraints became most evident
when each was relaxed. The rapid transition to communities dominated primarily by exotic species followed
immediately after a relaxation in biotic resistance caused
by flood-induced decline in both native and exotic species
in. The drought of 2007, representing a relaxation in
abiotic constraints on exotic species, allowed exotics to
rebound quickly from the floods of 2006, while native
species lagged or continued to decline (Fig. 1).
Competition was evident in every year of this
experiment, resulting in a strong inverse relationship
between native and exotic species cover in the most
productive pools (Table 1). We characterized the upper
bound on joint cover of native and exotic species using a
linear relationship (Fig. 3). A linear bound on joint
cover (above which resource competition precludes
growth) is appropriate if the strength of competition
between individuals is similar within and between native
and exotic species classes. In some years, however,
maximum cover appears nonlinear, with higher maxima
in pools dominated by either native or exotic species,
and lower maxima in pools that contain a mixture of
native and exotic species (see especially Fig. 3c–f ). This
nonlinearity may signify that competition is more
intense between native and exotic species classes than
within each class (Gilpin and Ayala 1973).
Heightened competition in pools with mixed assemblages, relative to pools dominated by either natives or
exotics, may have several important effects on community dynamics, as follows. Under typical abiotic
conditions, pools dominated by native species should
be relatively difficult for exotic species to invade, as
observed during the first five years of this experiment.
Despite a general decline in the cover of native species
during 2002–2006, exotics failed to replace natives prior
to major climatic disturbances. Conversely, it may be
difficult to restore native communities in pools overrun
by exotics. In addition, the rate of transition from native
to exotic species (or conversely) should be affected by
2115
the relative intensity of competition between species
classes. Heightened competition between native and
exotic species should hasten the outcome favored by
current abiotic conditions, whether the outcome is
invasion or restoration. Thus, even fleeting abiotic
conditions may facilitate transitions like invasion. This
prediction appears to be supported by the rapid invasion
of exotics observed in this system during 2007–2008.
Each of these predictions depends on the shape (linear
or curvilinear) but not the position of the isocline
describing maximum cover, so these predictions should
not be affected by the observed fluctuation in maxima
among years (Table 1, Fig. 3).
Fluctuations in maximum productivity can also
provide opportunities for invasion, and such fluctuations were clearly observed during the course of this
study. Maximum cover in pools dominated by native
species varied by more than a factor of two between
2003 and 2008, and evidenced a series of declines (Table
1). The most productive pools observed in this study
were those dominated by native species, suggesting that
native species can make more efficient use of abiotic
resources in this system, under certain conditions.
Maximum cover in pools dominated by exotic species
also declined by nearly a factor of two between 2003 and
2004, and did not recover during the remainder of this
study (Fig. 3). This pattern is most likely explained by
climatic events that reduced the productivity of exotic
species without eliminating their competitive advantage.
The importance of climatic and other abiotic conditions affecting the outcome of competition need not
overshadow the sometimes corresponding effects of
biotic conditions. We have demonstrated strong effects
of both abiotic ( pool depth) and biotic (seeding
treatment) conditions on time-averaged indices of
invasion and especially strong effects of biotic resistance
to invasion by the dominant exotics in this system. Our
temporal data reiterate the importance of biotic
resistance to invasion. The floods of 2006 and drought
of 2007 combined to produce a sequence of abiotic as
well as biotic effects that hastened exotic species
invasion, as follows: median (Fig. 1) and mean (Fig. 2)
exotic species cover fell significantly between 2005 and
2006 (95% confidence intervals overlapped in only one
treatment; Fig. 2a), and flooding is the only available
explanation for this dramatic decline. Exotic cover then
rebounded during the 2007 drought (Figs. 1 and 2), in a
pattern at first suggesting mainly abiotic effects on
invasion dynamics. In contrast, the average cover of
native species rose and fell in 2007–2008 in a pattern that
could be explained as lagged evidence of competition
with exotic species. (A one-year lag in native response
would be expected if competition affected seed production in these annual species.) Although reduced competition from exotic species may explain the slight rise in
native cover in all seeded pools during 2007 (Fig. 2a–d),
native cover was understandably outpaced during this
drought year by rising exotic cover. In every treatment,
2116
SHARON K. COLLINGE ET AL.
exotic cover in 2007 was significantly higher than in 2006
(95% confidence intervals did not overlap; Fig. 2), a
pattern not seen in native cover under any treatment.
The subsequent decline in native cover across all
treatments in 2008 (Figs. 1–3) cannot be explained by
the fairly typical amount of precipitation received in that
year (Fig. 4a). This decline did, however, coincide with
often unprecedented competition from exotic species
(Fig. 2c–e). Similarly, details of the trajectory followed
by exotic species suggest that invasion was facilitated by
a relaxation in biotic resistance. Notably, the 2007 rise in
exotic cover was increasingly dramatic in treatments
with lower seeding intensity and most dramatic in
control pools (Fig. 2), precisely where we would expect
weaker competition from native species. The rapid
replacement of native cover with exotic cover between
2006 and 2008, after four years of more gradual change,
also suggests that invasion was facilitated by a positive
feedback process or nonlinear, threshold dynamics.
There were no significant interannual changes in median
or mean native cover prior to the significant declines of
2007–2008, and no significant interannual increases in
exotic cover prior to those of 2006–2008 (Figs. 1 and 2).
Although many pools remained free of exotics during
the first five years of the experiment (points against the
vertical axis in Fig. 3a–e), exotic species were able to
invade almost every one of these during just the last two
years (Fig. 3g). Furthermore, although maximum
productivity was significantly lower in 2007–2008 than
earlier in this experiment (Fig. 3 and Table 1), exotics
were able to rebound from their nadir in 2006 to achieve
complete dominance by 2008.
The dynamics of invasion in this field experiment were
driven largely by two exotic species. Two exotic grasses,
Lolium multiflorum and Hordeum marinum, were relatively unaffected by flooding in 2006, and appeared to
thrive during the drought of 2007 (Fig. 6a). Previous
field observations showed that H. marinum tended to
occur in deeper pools and its abundance was positively
associated with pool depth (Gerhardt and Collinge
2003). In contrast, L. multiflorum occurred primarily at
pool edges and its abundance was negatively associated
with pool depth. In the same field study, the abundances
of both species were also negatively associated with
native plant cover. In a greenhouse study, survival of
both species was unaffected by inundation, but growth
and reproduction decreased with increasing water depth
(Gerhardt and Collinge 2007). However, both species
were also strongly influenced by interactions with
neighboring plants, as both growth and reproduction
decreased with increasing densities of native neighbors.
In agreement with these previous observations, both
species appeared to benefit dramatically from reduced
biotic resistance following the extreme climatic events in
our long-term field study. In addition, our large data set
allowed the detection of important nonlinearities in the
effects of seeding and inundation on both species. Both
species reached higher frequencies in pools of interme-
Ecological Applications
Vol. 21, No. 6
diate depth (Fig. 6b, c), allowing them to invade a broad
spectrum of pools in this system.
The pattern of exotic species invasion documented
here was presaged by a greenhouse experiment demonstrating the potential for L. multiflorum and H. marinum
to tolerate inundation (Gerhardt and Collinge 2007).
The same experiment demonstrated negative effects of
competition with native species on population growth in
these two exotics. One conclusion of that short-term
study was that abiotic resistance generally eclipsed biotic
resistance to invasion in this system, and that ‘‘biotic
resistance reduces invasibility only in more benign
environments’’ (Gerhardt and Collinge 2007:922). The
current results from our long-term field study suggest a
different conclusion: Biotic resistance was overcome
more easily in shallower pools, which presumably
represent more benign environments for exotic species.
The harshest environments (deepest pools) were invaded
only after biotic resistance was removed and a key
abiotic constraint (pool flooding) was relaxed by a
drought. For example, L. multiflorum had become well
established (frequency . 40%) in only eight pools deeper
than 15 cm by 2005, 0 by 2006 (the year of extreme
flooding), 20 by 2007 (the drought year), and 43 by 2008.
Thus, this long-term field study has revealed the
formerly cryptic potential for key effects of biotic
resistance, as well as revealing temporal features (the
sequence and confluence of events) that may allow
widespread invasion of vernal pool communities.
Lessons for conservation and restoration
The success of many ecosystem restoration efforts is
compromised by invasion of exotic species (D’Antonio
and Meyerson 2002, Suding et al. 2004, Funk et al.
2008), so our efforts to clarify factors affecting invasion
of restored vernal pools may be useful to future
conservation and restoration efforts. We observed that
exotic species cover varied inversely with the number of
native seeds initially sown, but this effect lasted only
while abiotic conditions were ‘‘typical.’’ Extreme flooding followed by extreme drought allowed invasion by
exotics regardless of initial seeding treatment. These
findings suggest that efforts to maintain and restore
native biodiversity are likely to be most successful under
a range of abiotic conditions typical for native
ecosystems, and management interventions may be
necessary following climatic extremes.
Our results may also appear to suggest that more
diverse communities offer more biotic resistance, in
agreement with many other theoretical and empirical
studies. However, density and diversity varied in tandem
across our seeding treatments, with one exception.
Although our study was not designed to evaluate the
relative effects of density and diversity on biotic
resistance, it did address effects of density in the case
of one species, Lasthenia conjugens. Pools sown with 300
seeds of this species offered more biotic resistance than
those sown with 100 seeds (Fig. 2c, d). This effect was
September 2011
EXOTIC SPECIES IN RESTORED VERNAL POOLS
likely driven by the initially higher abundance of L.
conjugens in more heavily seeded pools (Collinge and
Ray 2009).
Although vernal pools are unique in many respects,
the lessons learned from our long-term study of
invasibility in this system have important implications
for our understanding of the forces shaping biological
invasions in general. In particular, the unique characteristics of these systems allowed us to detect the
importance of climatic extremes in altering the abiotic
and biotic environments. Such disturbances have been
shown to play important roles in facilitating exotic
species invasion in a variety of ecosystems (Hobbs and
Huenneke 1992, Davis et al. 2000). In addition, our
study allowed us to detect both the cryptic effects of
biotic resistance and how the invasion process was
altered by changes in abiotic conditions and overall
ecosystem productivity. Both abiotic constraints and
productivity have been shown to be important determinants of the success of biological invasions (see reviews
in Ewel et al. 1999, Mack et al. 2000), but this long-term
study demonstrated one of the mechanisms whereby this
process occurs, principally by reducing the productivity
and presumably the competitive edge of the native plant
community.
Many environmental factors that act as abiotic
constraints on invasions, such as precipitation patterns
and the resulting levels of inundation, are not obviously
amenable to control by humans. In addition, climatic
variability is likely to increase as a consequence of global
climate change (IPCC 2007, Füssel 2009), leading a
number of authors to propose that biological invasions
will become more frequent (Mooney and Hobbs 2000 and
chapters therein, Hellmann et al. 2008, Dukes et al. 2009).
Our study contributes a relevant example of how extreme
climatic events can enhance exotic species invasions.
ACKNOWLEDGMENTS
We are indebted to two anonymous reviewers for many
helpful comments on a previous version of this manuscript. We
especially thank Jaymee Marty for assistance with all aspects of
this project, and we are especially grateful for the excellent field
assistance of C. Battaglia, C. Christian, J. Compton, W.
Johnson, T. Levy, and C. Wise. We thank M. Anderson, R.
Holmes, and K. Neurer for facilitating our research at Travis
AFB, and Y. Linhart, K. Rice, E. Warne, and D. Wright for
their insights into vernal pool ecology and restoration. This
project was funded by the Air Force Center for Environmental
Excellence, the U.S. Fish and Wildlife Service, CH2M Hill, and
a grant from the U.S. National Science Foundation (Long
Term Research in Environmental Biology, DEB-0744520).
Permission to collect seeds of L. conjugens was granted by the
U.S. Fish and Wildlife Service to S. K. Collinge (permit number
TE828382-3).
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