Invasion of Exotic Plant Species in Tallgrass
Prairie Fragments
ANNE C. CULLY,*§ JACK F. CULLY JR.,† AND RONALD D. HIEBERT‡
*Division of Biology, Kansas State University, Manhattan, KS, 66506, U.S.A.
†U.S. Geological Survey, Kansas Cooperative Fish and Wildlife Research Unit, Division of Biology,
Kansas State University, Manhattan, KS 66506, U.S.A.
‡Midwest Region National Park Service, Omaha, NE 68102, U.S.A.
Abstract: The tallgrass prairie is one of the most severely affected ecosystems in North America. As a result of
extensive conversion to agriculture during the last century, as little as 1% of the original tallgrass prairie remains. The remaining fragments of tallgrass prairie communities have conservation significance, but questions remain about their viability and importance to conservation. We investigated the effects of fragment
size, native plant species diversity, and location on invasion by exotic plant species at 25 tallgrass prairie sites
in central North America at various geographic scales. We used exotic species richness and relative cover as
measures of invasion. Exotic species richness and cover were not related to area for all sites considered together. There were no significant relationships between native species richness and exotic species richness at
the cluster and regional scale or for all sites considered together. At the local scale, exotic species richness was
positively related to native species richness at four sites and negatively related at one. The 10 most frequently
occurring and abundant exotic plant species in the prairie fragments were cool-season, or C3, species, in contrast to the native plant community, which was dominated by warm-season, or C4, species. This suggests that
timing is important to the success of exotic species in the tallgrass prairie. Our study indicates that some small
fragments of tallgrass prairie are relatively intact and should not be overlooked as long-term refuges for prairie species, sources of genetic variability, and material for restoration.
Invasión de Especies Exóticas en Fragmentos de Pastizal Alto
Resumen: El pastizal alto es uno de los ecosistemas más severamente afectados de Norte América. Sólo subsiste
el 1% de la extensión original de pastizales, debido a su conversión extensiva a terrenos agrícolas durante el
siglo pasado. Los remanentes de pastizal tienen significancia para la conservación, pero hay dudas sobre su
viabilidad e importancia para la conservación. Se investigaron los efectos del tamaño del fragmento, de la
diversidad de especies nativas y de la ubicación geográfica sobre la invasión por plantas exóticas en 25 sitios en
el centro de Norte América a distintas escalas geográficas. Utilizamos la riqueza de especies exóticas y la cobertura relativa como medidas de invasión. La riqueza de especies exóticas y la cobertura no se relacionaron con
el área al considerar todos los sitios en conjunto. No hubo correlación significativa entre la riqueza de especies
nativas y la de especies exóticas a escala de “cluster” o región ni al considerar todos los sitios en conjunto. A escala local, la riqueza de especies exóticas se correlacionó positivamente en cuatro sitios y negativamente en
uno. Las 10 especies exóticas más frecuentes y abundantes en los fragmentos de pastizal fueron especies C3 de
temporada fresca, mientras que la comunidad de plantas nativas se integraba principalmente por especies C4
de temporada cálida. Esto sugiere que la temporización juega un papel importante en el éxito de especies exóticas en el pastizal alto. Nuestro estudio indica que algunos fragmentos pequeños de pastizal alto están relativamente intactos y no deben dejar de tomarse en cuenta como refugios a largo plazo para especies de pastizal y
como fuentes de variabilidad genética y de material para restauraciones.
§Current address: Colorado Plateau Cooperative Ecosystem Studies Unit, P.O. Box 5765, Northern Arizona University, Flagstaff, AZ
86011–5765, U.S.A., email anne.cully@nau.edu
Paper submitted March 12, 2002; revised manuscript accepted October 27, 2002.
990
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Volume 17, No. 4, August 2003
Cully et al.
Introduction
Prior to European settlement in the region, the North
American tallgrass prairie extended from south-central
Canada to southern Texas (Fig. 1). Europeans arrived in
the prairie by the 1600s, but most of the conversion to
agriculture took place in the last 100 years. Today, as little as 1% of the original tallgrass prairie remains. This is
the largest reduction in size for any major ecosystem in
North America (Samson & Knopf 1994). The remaining
tallgrass prairie occurs primarily at sites where rocky
outcrops and shallow soils make it impossible to plow
and plant crops. The largest remaining continuous area
of tallgrass prairie extends from the Flint Hills of northcentral Kansas to north-central Oklahoma. After settlement by Europeans, the Flint Hills region provided forage for cattle, and livestock grazing continues to be the
predominant use of the land today. The tallgrass prairie
is bordered on the east by deciduous woodlands. On
this dynamic border, prairie patches occur within a
woodland-grassland mosaic. Away from the woodlands,
prairie patches have been isolated by human activities
and are surrounded primarily by agricultural lands. In
the tallgrass prairie area, annual precipitation averages
60–80 cm. To the west, mid- and short-grass prairies border the tallgrass prairie (Risser et al. 1981).
Fragments of tallgrass prairie have survived in scattered locations. Some are protected to conserve remnant plant and animal communities that were characteristic of the once-extensive ecosystem. Some are preserved
Figure 1. Tallgrass, mid-grass, and short-grass prairies
(Archibold 1995), and locations of four clusters of
tallgrass prairie sites.
Exotic Species Invasion of Tallgrass Prairie
991
in areas that have been set aside for reasons other than conservation (e.g., for their historical and cultural significance),
such as Pipestone National Monument in southwestern
Minnesota and Wilson’s Creek National Battlefield in
Missouri. Most are unprotected. Considering the drastic
reduction of the tallgrass prairie throughout North
America and the loss of natural ecosystem processes such
as fire and grazing, it is reasonable to question the conservation value of the remaining fragments.
Fragmentation combined with the loss of area may
cause species loss directly by eliminating habitats. Likewise, species numbers or richness may decline in fragments as a result of isolation and lack of immigration of
propagules from neighboring communities. Specific examples of species decline in the tallgrass prairie over the last
32–52 years are reported by Leach and Givnish (1996).
The tallgrass prairie is relatively homogeneous in species
composition, with few endemics. Many prairie species occur throughout the central part of the continent (Great
Plains Flora Association 1986). However, several tallgrassprairie plant species are listed by the U.S. Fish and Wildlife Service as threatened (likely to become endangered
in the foreseeable future), including prairie bush clover
(Lespedeza leptostachya), Meade’s milkweed (Asclepias
meadii), and eastern and western fringed prairie orchid
(Platanthera leucophaea and P. praeclara respectively).
Running buffalo clover (Trifolium stoloniferum) is listed
as endangered (U.S. Code of Federal Regulations 50, 17.11
and 17.12, 31 October 1996 ). In addition, such a dramatic loss of area may have led to the loss of ecotypes and
associated genetic diversity of more common species.
Fragmentation may also increase vulnerability to invasion
by exotic species because small areas have a higher ratio
of edge to area and thus more points of entry for exotics
as well as increased likelihood of disturbance (Hobbs &
Huenneke 1992). Here, we define exotic species as those
from other continents arriving in North America after
the time of Columbus, following the definition used by
the Great Plains Flora Association ( 1986 ). Exotic plant
species may directly compete with native species and
may cause changes in ecosystem processes that have
profound effects on native species (Mack 1989; Howe &
Knopf 1991; Pimm 1991; Vitousek 1992; Christian & Wilson 1999). The ecological consequences of most invasions
are undocumented, however, and their effects on ecosystem function are poorly understood (Pimm 1991; Parker
et al. 1999).
There is conflicting evidence from theoretical and empirical studies for the role of species diversity in the success or failure of invasions (Levine & D’Antonio 1999).
Both negative (Rejmánek 1989; Pysek & Pysek 1995; Tilman 1997) and positive (Robinson et al. 1995; Smith &
Knapp 1999) relationships between native species richness and invasion by exotics have been found. Areas
rich in species are thought by some to be resistant to invasion because resources are being fully used by species
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Exotic Species Invasion of Tallgrass Prairie
already present and new species are unable to find unexploited resources ( MacArthur & Wilson 1967; MacArthur 1972; Pimm 1991 ). Species-poor areas, on the
other hand, may have unexploited resources that are
available to new species.
Species richness can be viewed in terms of numbers
of species per unit area or species density (Glass 1981;
Eyster-Smith 1984; Holt et al. 1995; Kellman 1996; Harrison 1999). Using a species-density approach in community studies of invasive species has the advantage of providing information on the extent of invasion by exotic
species. If exotics are frequent and abundant, they may
be considered successful invaders and the ones with the
potential to have a large ecological impact.
Variables that may influence invasion success include
similarities in latitude, life history, climatic regime, and
seasonality between the home of the exotic species and
the invaded area. Invasions may be limited to certain
times of the year (Crawley 1989; Rejmánek 1989; Hobbs &
Huenneke 1992; Levine & D’Antonio 1999; Stohlgren et
al. 1999). Surrounding land use influences invasion; natural areas surrounded by cultivated land, tame pasture
( native grasslands converted to exotic species ), or urban landscapes, are more prone to invasion ( Smith &
Knapp 2001; Hobbs & Huenneke 1992) than natural areas surrounded by native cover. Another influence on invasion by exotic species includes management of native
plant communities.
We investigated fragmentation of the tallgrass prairie
and how the location of the remaining fragments may
have affected their invasion by exotic species. Specifically, we asked the following questions: (1) Does exotic
species richness increase as prairie fragments decrease
in area? ( 2 ) Is exotic species richness in prairie fragments related to their native species richness? ( 3 ) Do
prairie fragments surrounded by woodlands have lower
exotic species richness than those in agricultural settings? (4) What ecological characteristics are associated
with the tallgrass prairie invaders? We asked these questions at the local, cluster, and regional scales, and for all
the sites together.
Methods
Site Selection
The sites were grouped in four clusters; each cluster included a national monument and five other sites in the
area. We selected the first two clusters in the woodlandgrassland mosaic region, where small patches of prairie
vegetation occur within a matrix of woodlands and forests. The first cluster was from sites in and around Effigy
Mounds National Monument, Iowa; sites ranged in size
from 0.1 to 16.0 ha. The shortest distance between
sites was 5 km and the longest was 40 km. The second
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Cully et al.
cluster was selected from sites near Wilson’s Creek National Monument, Missouri, where the sites ranged in
size from 2.4 to 130 ha. The shortest distance between
sites was 30 km and the longest about 135 km. The third
and fourth clusters were located in the Great Plains region, where prairie fragments have been isolated by human activities, primarily farming and urbanization. The
third cluster was near Pipestone National Monument,
Minnesota; sites ranged in size from 0.5 to 821 ha, with
the minimum distance between sites about 10 km and
the maximum about 50 km. The fourth cluster was chosen from sites around Homestead National Monument,
Nebraska, where sites ranged in size from 0.5 ha to
12,000 ha. The minimum distance between sites was 0.5
km and the maximum distance about 270 km. Sites in
the Effigy Mounds and Wilson’s Creek clusters occurred
on thin soils and rocky outcrops of limestone. All sites
but two of the Homestead cluster fragments also had
similar soils and substrate. In the Pipestone area, shallow
soils rest on underlying formations of quartzite. Three
clusters had six sites and the fourth cluster had seven,
including a restoration site. All sites were ungrazed, and
most were on a 3- or 4-year burning cycle.
Data Collection
In 1996 and 1997, we established six transects at each of
25 sites. Transects were located in grassland with few or
no trees, although some of the small sites were surrounded by trees and shrubs. Within each prairie, the origins of the six transects were randomly determined. At
the larger-sized sites, all transects were located away
from the boundary, to avoid edge effects. We were not
interested in exotic species at the edge so much as their
incursion into the interior of patches of prairie. At the
very small sites, transects were unavoidably located near
the edges. Some sites were heterogeneous in their physical setting and included upland, slopes, and valley bottoms. Other sites were entirely on uplands or slopes.
Where topography allowed at a single site, we placed
two transects on upland locations, two on slopes, and
two in bottomlands. If a site did not have the range of topographic features, the six transects were then located
based on the topography (e.g., entirely on a slope). For
sites that were large enough, a grid was overlaid on a
map of the site, grid squares numbered, and numbers selected randomly for the transect origins. We determined
transect direction randomly. At four small sites, the location and direction of transects was constrained by the
necessity of fitting them all in without overlap. Most
transects were 50 m long, except at one very small site.
Five 10-m2 circular sampling plots were located at 10-m
intervals along each transect, for a total of 30 plots per
site. At one small site, sampling plots were set out along
shorter transect lines at the same intervals of 10 m.
More than five of the shorter transects were required to
Cully et al.
Exotic Species Invasion of Tallgrass Prairie
achieve a total of 30 sampling plots. No plot overlapped
another.
In 1997 and 1998, we recorded all vascular plant species and their abundance ( measured as relative canopy
cover) at all the sites. We visited all the plots twice each
year during the growing season, once in the early summer and again in the late summer, to capture the two
major peaks of growth and reproduction of the cool- and
warm-season species. We recorded the canopy coverage
of each species using coverage classes similar to those of
Daubenmire (1959).
At all sites, we sampled equal numbers of plots to
avoid bias due to unequal site size and sampling effort
( Simpson 1964; Rosenzweig 1995 ). The primary focus
of analysis was the number of native and exotic species
per unit area, or species richness.
Analysis
Data from the early and late-period sampling for both
1997 and 1998 were combined, and total numbers of
identified species were calculated. The cover values for
each species and the total cover values for all species at
each site were calculated, and the percentage of the total cover for each species was then determined. We
compared native and exotic plant species at several
scales: local (within sites), clusters (sites within clusters),
regional (clusters within regions), and for all sites combined. We compared species richness and cover values
following the methods of Ott ( 1993 ). We used PROC
Univariate and Levene’s test in PROC Glm (SAS Institute
1996) to determine that the data were normally distributed and that variances within clusters of sites were
equal. We used one-way analysis of variance to determine whether clusters differed from one another in exotic species numbers and cover. When statistically significant differences occurred, we used the Waller
Duncan K-ratio t test to determine which means were
different from the others. We also looked for relationships among area, species richness, and cover values of
993
native and exotic species by using Pearson correlation
coefficients, PROC Corr, and regression analysis ( SPSS
1997).
Results
Site Size and Exotic Species Numbers
We identified 573 plant taxa; 59 (10%) were exotic. (A
list of the exotic plant species can be obtained from
A.C.C.) The mean number of exotic species for all sites
was 9 (range of 2–18) (Table 1). For all sites considered
together, exotic species richness was not significantly
related to area ( p 0.85 ) ( Fig. 2 ). The two clusters
within the woodland-grassland mosaic region were significantly different from each other in exotic species
richness ( p 0.05) (Table 1). The clusters in the Great
Plains region were also significantly different from each
other in richness of exotic species. Two clusters, one
from the Great Plains region and one from the woodland-grassland mosaic region, had significantly higher
numbers of exotic species than the others ( p 0.05 )
( Table 1 ), suggesting that factors other than regional
variables ( i.e., surrounding land uses, climate ) influenced invasion.
Exotic Species Cover
For all sites, about 16% (15.7 SD15.45) of total cover
was from exotic species ( Table 1 ), with a minimum of
0.15% and a maximum of 56%. As with species richness,
there was no significant relationship between area and
exotic species cover for sites considered overall ( p 0.81) (Fig. 3). However, exotic species cover decreased
significantly with increase in area ( p 0.02) within the
Table 1. Mean number of exotic plant species and mean
percentage of exotic plant species cover (SD) ftom tallgrass
prairie fragments.*
Region and cluster
Woodland-grassland
Effigy Mounds
Wilson’s Creek
Great Plains
Pipestone
Homestead
All sites
Mean no. exotic
species (SD)
Mean cover exotic
species (%)
6.1b (2.7)
12.8a (4.9)
3.9c (3.7)
17.6b (18.5)
12.1a (2.2)
6.7b (4.8)
9.3 (4.8)
32.8a (13.3)
9.5b (5.3)
15.7 (15.5)
*The letters a, b, and c demote significant differences between means
within columns ( p 0.05) (e.g., in column of mean number exotic
species, 6.1b is significantly different from 12.8a, but not different
than 6.7b) (ANOVA, SAS 6.1 1996).
Figure 2. Exotic plant species richness by area of
tallgrass prairie fragments.
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Exotic Species Invasion of Tallgrass Prairie
Cully et al.
four sites: two in the Wilson’s Creek cluster ( r 0.453,
p 0.012; r 0.363, p 0.049) and two in the Effigy
Mounds cluster ( r 0.498, p 0.005; r 0.527, p 0.003 ). We found a significant negative relationship at
one site in the Pipestone cluster ( r 0.521, p 0.003).
Characteristics of Exotic Species
Figure 3. Exotic plant species cover by area of
tallgrass prairie fragments.
Wilson’s Creek cluster. The Pipestone cluster had the
highest exotic species cover at 33% ( p 0.05), and the
Effigy Mounds cluster had the lowest exotic species
cover at 4% ( p 0.05) (Table 1).
Numbers of Native and Exotic Species
For all sites considered together, there was no significant relationship between the numbers of native and exotic plant species ( p 0.93 ) ( Fig. 4 ). For the sites of
the Effigy Mounds cluster, exotic species increased significantly ( p 0.04) with increased numbers of native
species (Fig. 4), but this result was strongly affected by
an outlier.
At the local scale, we found a significant positive relationship between native and exotic species richness at
Figure 4. Exotic plant species richness by native plant
species richness in tallgrass prairie fragments.
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Volume 17, No. 4, August 2003
Twenty-seven percent of the exotic species were perennial; 13% (8 species) were perennial, rhizomatous grasses.
In spite of their relatively low numbers, perennial, rhizomatous grasses made up 68% of the total exotic species
cover (Table 2). Two of the grasses, Kentucky bluegrass
( Poa pratensis) and smooth brome ( Bromus inermis ),
together comprised 62% of the total exotic cover (Table
2) and 10% of total cover for all plant species. Kentucky
bluegrass was found in almost half of the 750 plots;
smooth brome was the second-most frequently occurring species (Table 2). Annual species comprised a high
percentage (63%) of the total number of exotic species
but occurred less frequently and were less abundant
than perennial, rhizomatous grasses (Table 2). Biennial
species made up the remaining cover for the top 10 species.
Smooth brome, tall fescue ( Festuca arundinacea ), Kentucky bluegrass, and Canada bluegrass ( Poa compressa ),
4 of the 10 most frequently occurring and abundant exotic species in our study, were introduced into North
America as forage crops and for landscaping. These species have been extensively planted throughout the eastern and midwestern regions of the United States for
many years and have escaped into native plant communities and become naturalized. They are considered aggressive weeds in settings where they are not purposefully planted ( Hitchcock 1971; Great Plains Flora
Association 1986). Annual brome grasses were accidentally introduced to North America and are now agricultural and range pests (Knapp 1996; Whitson et al. 1999).
Low hopclover ( Trifolium campestre ) is an herbaceous
annual species that has become naturalized in parts of
the Great Plains and is often seen in disturbed areas.
One of two biennial species to make it into the list of
top 10 exotic species in our study, wild carrot ( Daucus
carota ), is found throughout the United States in old
pastures, meadows, and grasslands. Western salsify ( Tragopogon dubius ) has likewise become naturalized in
North America (Great Plains Flora Association 1986;
Whitson et al. 1999). These species are considered
weeds in the tallgrass prairie in the sense that they are
growing in places where they are not wanted ( Randall
1997); their effects on the tallgrass prairie ecosystem are
not fully understood.
The 10 most frequently occurring and abundant exotic species ( Table 2 ) were all characterized by the C3
metabolic pathway (Downton 1975). The C3 plant species are adapted to growth and reproduction during
Cully et al.
Exotic Species Invasion of Tallgrass Prairie
995
Table 2. Frequency of occurrence, percentage of cover, and ecological and phenological characteristics of the 10 most frequently occurring
exotic plant taxa in tallgrass prairie fragments.
Plant taxa
Poa pratensis
Bromus inermis
Trifolium campestre
Festuca arundinacea
Tragopogon dubius
Bromus tectorum
Bromus japonicus
Poa compressa
Daucus carota
Bromus sp.
Occurrence of
exotics (%)
Cover of
exotics (%)
C3 /C4a
Life
formb
Phenologyc
49
16
16
7
5
5
4
4
4
3
39
23
10
2
1
1
4
4
1
7
C3
C3
C3
C3
C3?
C3
C3
C3
C3?
C3
G, P, Rh
G, P, Rh
H, A
G, P, Rh
H, B
G, A
G, A
G, P, Rh
H, B
G, A
May–August
May–July
May–September
May–October
May–July
May–June
May–July
June–August
April–June
May–?
a
Taken from Downton (1975) and Sage et al. (1999); ?, species not listed; C3 or C4 determined by phylogenetic relationships and phenology.
B, biennial; G, grass; H, herbaceous; P, perennial; Rh, rhizomatous.
c
Taken from Great Plains Flora Association (1986).
b
cool, moist conditions and become dormant or at least
less productive during hot, dry conditions. These 10 species exhibit similar phenological traits, in that all are capable of beginning their reproductive cycles early in the
growing season ( Table 2 ). The annual brome grasses
and the biennial species had the briefest period for completing their phenology.
Discussion
Small natural areas are expected to have both fewer species overall and a higher proportion of exotic species
than larger areas ( MacArthur & Wilson 1967; Hobbs &
Huenneke 1992 ). Contrary to expectations, our study
showed no significant relationship between area and exotic species richness or cover. This is in contrast to the
results of a study of serpentine grasslands of California,
where small patches of grassland had higher numbers of
exotic plant species than larger, contiguous areas of habitat (Harrison 1999). The differences in the two outcomes
may be due to differing ecological factors that govern
the potential for invasion in the two grassland systems
and to human factors. The serpentine grassland patches
are primarily a function of the presence of serpentine
outcrops; these outcrops are characterized by high ratios of magnesium to calcium in the soil, which most
plant species cannot tolerate. Tallgrass prairie fragments
are primarily the result of breaking up of the sod on formerly expansive and relatively contiguous grasslands.
In our study, we also found no significant relationship
between area and total species richness per sampling
unit for all sites considered together. Similar patterns of
equivalent native species richness in both small and
large areas have been reported in serpentine grasslands
(Harrison 1999), other studies of tallgrass prairie (Glass
1981; Eyster-Smith 1984; and Holt et al. 1995), and isolated gallery forests in the tropics (Kellman 1996).
Because the number of exotic species alone may not
give the full picture of the potential impact of invasion
( Parker et al. 1999), we used cover as a rough estimate
of the impact of exotic species at tallgrass prairie sites.
Results for exotic species cover were similar to those for
exotic species richness. For the all-site analysis, there
was no relationship between area and exotic species
cover. However, the exotic plant cover at sites in the
Wilson’s Creek cluster was significantly negatively related to site size. Those sites with the highest number of
exotic species also had the highest percent cover of
exotics.
Exotic species richness in the tallgrass prairie did not
appear related to regions in which the sites were located
( woodland-grassland mosaic or Great Plains regions ),
and the presence of a native woodland community surrounding the fragmented prairies did not necessarily diminish the effects of invasion. The smallest sites of the
Wilson’s Creek cluster were imbedded in a woodlandgrassland matrix, yet they have been invaded by many
exotic species. On the other hand, sites in the Effigy
Mounds cluster, also surrounded by woodlands, had few
exotic species. In addition to numbers of species, exotic
plant cover was far less extensive in the Effigy Mounds
cluster than the Wilson’s Creek cluster. The Pipestone
cluster in the Great Plains region included sites surrounded by croplands. These sites had high levels of exotic species richness and cover. The sites of the Homestead cluster, also in the Great Plains region, were
surrounded by a mixture of land uses: grazing (including
on native prairie ), croplands, and urban areas. These
sites had fewer exotic species and lower exotic cover
values than sites in their “sister” Pipestone cluster.
Communities with high species richness may be more
resistant to invaders than species-poor communities (Elton 1958; MacArthur & Wilson 1967; Pimm 1991; Robinson et al. 1995). In our study, there was no correlation
between native species richness and exotic species
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Exotic Species Invasion of Tallgrass Prairie
numbers when all sites were analyzed together. We
found significant positive correlations between native
and exotic species richness at the local scale at 4 of the
25 sites, and a significant negative correlation at 1 site.
Smith and Knapp (1999) reported a significant positive
relationship between native and exotic plant species
numbers at the local scale at the Konza Prairie Biological
Station. Stohlgren et al. ( 1999 ) found that, as in our
study, the relationship between the number of native
and exotic species varies depending on the scale of sampling: exotic species diversity is negatively related to native diversity at the subplot scale (samples within larger
plots) but positively related at the site scale. Because we
looked at invasion after the fact, we cannot be sure
whether or not species diversity played a role in the
past. We found little evidence, however, that native species diversity influences or determines exotic species
success in tallgrass prairie.
The ecological and growth-form characteristics of the
exotic species in our study suggest that the tallgrass prairie is vulnerable to invasive plants that have certain characteristics. All of the 10 most frequent and abundant
plant taxa were cool-season, or C3, species. This fits the
pattern reported by Smith and Knapp ( 1999 ) for the
Konza Prairie Research Natural Area in Kansas, where
90% of the exotic flora is C3 species. The dominance of
cool-season exotic invaders is in distinct contrast to the
dominance of native C4 grasses in the prairie vegetation.
It isn’t clear why C3 invaders are successful, but it may
be related to the fact that many of the invasive species
are capable of beginning their reproductive cycle early
in the year and can flower, produce fruit, and disperse
seed before the peak of growth occurs in the warm-season species. Those species that can use available space
and light before the warm-season species’ canopy reaches
its full extent may be able to maintain themselves in the
plant community. Later in the growing season, warm-season species reach their full abundance, obscuring and
overshadowing the cool-season species, many of which
have completed their reproduction and photosynthesis.
Perhaps during the spring and early summer more open
conditions provide the best opportunity for new species
to enter the plant community. Fire and grazing patterns of
the past probably also played a role in invasions by exotic
species such as Kentucky bluegrass, which are now widespread and abundant.
The effect of exotic species on ecosystem processes
was not the focus of this study, but their pervasive presence and abundance suggests that invasions may have
had a significant but undocumented effect on species diversity ( and perhaps ecosystem processes ) in the remnants of tallgrass prairie. Exotic perennial, rhizhomatous
grass invaders may compete for nutrients and moisture
with species of similar life form or phenology (Wedin &
Tilman 1990; Nernberg & Dale 1997; van der Kamp et
al. 1999 ). Early-season native species that may be dis-
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Cully et al.
placed by exotics include leguminous forbs (i.e., species
in the genera Psoralea and Dalea) and perennial grasses
(i.e., Koeleria cristata and Dichanthelium spp.). Forbs
are the source of much of the native species diversity in
the tallgrass prairie, and the effects of invasion on this
component of the community could be severe. Dominant native species in our study included big bluestem
( Andropogon gerardii ), Indian grass ( Sorghastrum
nutans), and switchgrass (Panicum virgatum), which
are also perennial rhizomatous grasses. Annual brome
grasses have altered ecosystems in the western United
States by changing the fire regime ( Mack 1989; Knapp
1996) and have altered nutrient cycling and soil biota in
western semiarid communities (Belnap & Phillips 2001;
Evans et al. 2001 ). The presence of exotic annual species in the tallgrass prairie may displace native annual
species such as six-weeks fescue ( Festuca octoflora )
and daisy fleabane (Erigeron strigosus), as well as seedlings of native perennials.
Area, region, and native species richness provided little explanation for invasion of tallgrass prairie fragments
by exotic species. Consequently, we must look at other
factors for more complete understanding of invasion in
prairie fragments (Howe 1993; Kowarik 1999). Our results suggest that cool-season species have a high probability of success in the tallgrass prairie. This information
may be useful in predicting which new exotic species
are likely to become invaders; limited resources can
then be applied most effectively to future invasions.
Other factors involved in successful invasions of the
tallgrass plant communities may include many we did
not measure or analyze. Soil characteristics, moisture
availability, and other ecological factors may be relevant,
but management and use of land surrounding the sites
are probably also key to explaining invasions. Additional
studies, addressing adjacent land use, species composition of surrounding areas, management regimes, and
ecological characteristics will be important in assessing
the potential for future invasions of tallgrass prairie.
In the past, it was assumed that fragmentation of and
exotic invasion into prairie fragments had caused loss of
native species. Our results show that some small patches
of tallgrass prairie are relatively intact and can support a
near-full complement of native species. Although these
tallgrass prairie units are not a substitute for large protected areas, small fragments of prairie should receive attention from conservationists and land managers as longterm refuges for prairie species and sources of genetic
variability and material for restorations.
Acknowledgments
The U.S. Geological Survey provided funding for this
study through the Natural Resources Preservation Program. The Division of Biology at Kansas State University
Cully et al.
also provided support. The Iowa Department of Natural
Resources, Konza Prairie Biological Station, the Minnesota Department of Conservation, the Minnesota Department of Natural Resources, the Minnesota Historical Society, the Missouri Department of Conservation, the
Missouri Department of Natural Resources, the National
Park Service, The Nature Conservancy, and the Wisconsin Department of Natural Resources gave us permission
to work on prairies under their management. R. Rovang,
L. Thomas, G. Willson, and M. Skinner helped locate
prairies for the study. R. Hamilton helped locate study
areas and provided logistical support. I. Barnard, G.
Towne, and A. Paulin helped with plant identification.
Our excellent field technicians included A. Ricehill, A.
Paulin, M. Wong, C. Perchellet, and T. Swanson. T. M.
Barkley, A. Knapp, and J. Blair reviewed the manuscript
at an early stage. Comments by S. Timming and S. Wilson greatly improved the paper.
Literature Cited
Archibold, O. W. 1995. Ecology of world vegetation. Chapman and
Hall, New York.
Belnap, J., and S. L. Phillips. 2001. Soil biota in an ungrazed grassland:
response to annual grass (Bromus tectorum) invasion. Ecological
Applications 11:1261–1275.
Christian, J. M., and S. D. Wilson. 1999. Long-term ecosystem impacts
of an introduced grass in the northern Great Plains. Ecology 80:
2397–2407.
Crawley, M. J. 1989. Chance and timing in biological invasions. Pages
407–424 in J. A. Drake, H. A. Mooney, F. di Castri, R. H.Groves, F. J.
Kruger, M. Rejmánek, and M. Williamson, editors. Biological invasions: a global perspective. Scientific committee on problems of
the environment, scope 37. Wiley, New York.
Daubenmire, R. 1959. A canopy coverage method of vegetational analysis. Northwest Science 33:43–64.
Downton, W. J. S. 1975. The occurrence of C4 photosynthesis among
plants. Photosynthetica 9:96–105.
Elton, C. S. 1958. The ecology of invasions by animals and plants.
Metheun, London.
Evans, R. D., R. Rimer, L. Sperry, and J. Belnap. 2001. Exotic plant invasion alters nitrogen dynamic in an arid grassland. Ecological Applications 11:1301–1310.
Eyster-Smith, N. M. 1984. Tallgrass prairies: an ecological analysis of 77
remnants. Ph.D. dissertation. University of Arkansas. Fayetteville.
Glass, W. D. 1981. The importance of refuge size in preserving species
of prairie legumes, goldenrods, and milkweeds. M.S. thesis. University of Illinois, Chicago.
Great Plains Flora Association. 1986. Flora of the Great Plains. University of Kansas Press, Lawrence.
Harrison, S. 1999. Local and regional diversity in a patchy landscape:
native, alien, and endemic herbs on serpentine. Ecology 80:70–80.
Hitchcock, A. S. 1971. Manual of the grasses of the United States. Dover Publications, New York.
Hobbs, R. J., and L. F. Huenneke. 1992. Disturbance, diversity, and invasion: implications for conservation. Conservation Biology 6:324–337.
Holt, R. D., G. R. Robinson, and M. S. Gaines. 1995. Vegetation dynamics in an experimentally fragmented landscape. Ecology 76:1610–
1624.
Howe, H. F. 1993. Managing species diversity in tallgrass prairie: assumptions and implications. Conservation Biology 8:691–704.
Howe, W. H., and F. L. Knopf. 1991. On the imminent decline of Rio
Exotic Species Invasion of Tallgrass Prairie
997
Grande cottonwoods in central New Mexico. Southwestern Naturalist 36:218–224.
Kellman, M. 1996. Redefining roles: plant community reorganization
and species preservation in fragmented ecosystems. Global Ecology and Biogeography Letters 5:111–116.
Knapp, P. A. 1996. Cheatgrass (Bromus tectorum) dominance in the
Great Basin desert: history, persistence, and influences to human
activities. Global Environmental Change 6:37–52.
Kowarik, I. 1999. Invasion potential and invasion success: on the relevance of man-made interactions. Proceedings of the fifth international conference on the ecology of invasive alien plants. La Maddalena, Sardinia, Italy.
Leach, M. K., and T. J. Givnish. 1996. Ecological determinants of species loss in remnant prairies. Science 273:1555–1558.
Levine, J. M., and C. M. D’Antonio. 1999. Elton revisited: a review of
evidence linking diversity and invasibility. Oikos 87:15–26.
MacArthur, R. H. 1972. Geographical ecology. Harper and Row, New
York.
MacArthur, R. H., and E. O. Wilson. 1967. The theory of island biogeography. Princeton University Press, Princeton, New Jersey.
Mack, R. N. 1989. Temperate grasslands vulnerable to plant invasions:
characteristics and consequences. Pages 155–180 in J. A. Drake,
H. A. Mooney, F. di Castri, R. H.Groves, F. J. Kruger, M. Rejmánek,
and M. Williamson, editors. Biological invasions: a global perspective. Scientific committee on problems of the environment, scope
37. Wiley, New York.
Nernberg, D., and M. R. T. Dale. 1997. Competition of five native prairie grasses with Bromus inermis under three moisture regimes.
Canadian Journal of Botany 75:2140–2145.
Ott, R. L. 1993. An introduction to statistical methods and data analysis. Duxbury Press, Belmont, California.
Parker, I. M., D. Simberloff, W. M. Lonsdale, K. Goodel, M. Wonham,
P. M. Kareiva, M. H. Williamson, B. Von Holle, P. B. Moyle, J. E. Byers, and L. Goldwasser. 1999. Impact: toward a framework for understanding the ecological effects of invaders. Biological Invasions
1:3–19.
Pimm, S. L. 1991. The balance of nature? The University of Chicago
Press, Chicago.
Pysek, P., and A. Pysek. 1995. Invasion by Heracleum mantegazzianum in different habitats in the Czech Republic. Journal of Vegetation Science 6:711–718.
Randall, J. M. 1997. Defining weeds of natural areas. Pages 18–25 in
J. O. Luken and J. W. Thieret, editors. Assessment and management
of plant invasions. Springer-Verlag, New York.
Rejmánek, M. 1989. Invasibility of plant communities. Pages 369–388
in J. A. Drake, H. A. Mooney, F. di Castri, R. H. Groves, F. J. Kruger,
M. Rejmánek, and M. Williamson, editors. Biological invasions: a
global perspective. Scientific committee on problems of the environment, scope 37. Wiley, New York.
Risser, P. G., E. C. Birney, H. D. Blocker, S. W. May, W. J. Parton, and
J. A. Wiens. 1981. The true prairie ecosystem. US/IBP synthesis series 16. Hutchinson Ross, Stroudsburg, Pennsylvania.
Robinson, G. R., J. F. Quinn, and M. L. Stanton. 1995. Invasibility of experimental habitat islands in a California winter annual grassland.
Ecology 76:786–794.
Rosenzweig, M. L. 1995. Species diversity in time and space. Cambridge University Press, New York.
Sage, R. F., M. Li, and R. K. Monson. 1999. Taxonomic distribution of
C4 photosynthesis. Pages 551–584 in R. F. Sage, and R. K. Monson,
editors. C4 plant biology. Academic Press, New York.
Samson, F., and F. Knopf. 1994. Prairie conservation in North America.
BioScience 44:418–442.
SAS Institute. 1996. SAS 6.1. SAS Institute, Cary, North Carolina.
SPSS. 1997. Sigma plot 4.1. SPSS, Chicago.
Simpson, G. L. 1964. Species density of North American recent mammals. Systematic Zoology 13:57–73.
Smith, M. D., and A. K. Knapp. 1999. Exotic plant species in a C4 dom-
Conservation Biology
Volume 17, No. 4, August 2003
998
Exotic Species Invasion of Tallgrass Prairie
inated grassland: invasibility, disturbance, and community structure. Oecologia 120:605–612.
Smith, M. D., and A. K. Knapp. 2001. Size of the local species pool determines invasibility of a C4 – dominated grassland. Oikos 92:55–61.
Stohlgren, T. J., D. Binkley, G. W. Chong, M. A. Kalkhan, L. D. Schell,
K. A. Bull, Y. Otsuki, G. Newman, M. Bashkin, and Y. Son. 1999.
Exotic plant species invade hot spots of native plant diversity. Ecological Monographs 69:25–46.
Tilman, D. 1997. Community invasibility, recruitment limitation, and
grassland biodiversity. Ecology 78:81–92.
van der Camp, G., W. J. Stolte, and R. G. Clark. 1999. Drying out of
small prairie wetlands after conversion of their catchments from
Conservation Biology
Volume 17, No. 4, August 2003
Cully et al.
cultivation to permanent brome grass. Hydrological Sciences Journal 44:387–397.
Vitousek, P. M. 1992. Effects of alien plants on native ecosystems.
Pages 29–41 in: C.P. Stone, C.W. Smith, and J.T. Tunison. Alien
plant species in native ecosystems of Hawai’i: management and research. University of Hawai’i Cooperative National Park Resources
Studies Unit, Honolulu.
Wedin, D. A., and D. Tilman. 1990. Species effects on nitrogen cycling:
a test with perennial grasses. Oecologia 84:433–441.
Whitson, T. D., L. C. Burrill, S. A. Dewey, D. W. Cudney, B. E. Nelson,
R. D. Lee, and R. Parker. 1999. Weeds of the west. University of
Wyoming, Laramie.