WETLANDS, Vol. 28, No. 4, December 2008, pp. 883–895
’ 2008, The Society of Wetland Scientists
LONG-TERM VEGETATION DEVELOPMENT OF RESTORED PRAIRIE
POTHOLE WETLANDS
Myla F. J. Aronson and Susan Galatowitsch
University of Minnesota
305 Alderman Hall
St. Paul, Minnesota, USA 55018
E-mail: arons076@umn.edu
Abstract: Although wetland restoration has been a key part of U.S. environmental policy for 20 years
(i.e., ‘‘no net loss’’), there is little long-term data on restorations to guide planning and assessment.
Understanding how restored wetland communities deviate from natural conditions, and how long those
deviations persist, can provide important insights into the mechanisms of recovery and improve
restoration practice. This study reports the results from a 19-year survey of 37 restored prairie pothole
wetlands in northern Iowa, southern Minnesota, and southeastern South Dakota. Complete floristic
surveys were performed for each of the wetlands in 1989, 1990, 1991, 2000, and 2007. The accumulation
of wetland species across all sites was greatest during the first 12 years after reflooding (14.4 species/year),
after which the rate declined to 1.6 species/year. Proximity to natural wetlands and a semi-permanent
water regime favored species accumulations during the first 12 years, but changes since then are primarily
linked to water regime. Semi-permanent wetlands have experienced fewer major gains and losses in
species richness, whereas temporary and seasonal wetlands have been less stable. From 2000 to 2007,
extinctions exceeded colonizations in all wetlands, resulting in a convergence of beta diversity. Although
77% of the species considered common to natural wetlands in the region established in these restorations,
70% of those considered infrequent have not. The likelihood that these restorations will eventually
support many additional species appears low, given the presence of barriers to recovery, especially the
dominance of invasive perennials (e.g., Phalaris arundinacea and Typha angustifolia/x glauca) on all sites
and the low colonization efficiency of wet prairie, sedge meadow, and woody perennial species.
Management, such as active revegetation of these low efficiency species guilds, particularly sedge meadow
and wet prairie perennials, and invasive species control is needed to ensure that restored prairie wetlands
support the region’s biodiversity. The important barriers to the recovery of prairie pothole restoration:
isolation, infrequent flooding, and invasive species, are all factors that do not self-correct over time and
need to be addressed during planning by establishing sound practices for initial implementation and longterm vegetation management.
Key Words: beta diversity, dispersal limitation, homogenization, invasive species, Iowa, Minnesota,
priority effects, South Dakota, species richness, wetland restoration
INTRODUCTION
effort to track the recovery of restored wetlands.
Monitoring of mitigation restorations is typically
limited to 3-5 years under regulatory programs to
ensure compliance with permit conditions (Mitsch
and Wilson 1996) while other wetland restorations
are generally not monitored. Consequently, generalizations on restored wetland development are not
often possible due to the paucity of long-term data
and the lack of replication at the regional scale
(Zedler 2000).
Comprehensive, long-term monitoring of abiotic
and biotic ecosystem parameters provide the best
scientific basis for improving practice, but it is
generally cost and time prohibitive. Vegetation
responses to restoration actions, however, have been
effective for detecting lags in recovery (e.g., Gibson
et al. 1994, Galatowitsch and van der Valk 1996a).
Twenty years ago, the U.S. adopted a federal
policy of ‘‘no net loss’’ for wetlands, following
George H.W. Bush’s presidential campaign pledge
(1988). Under this policy, wetland losses that cannot
be avoided must be mitigated through restoration or
creation. The science and practice of wetland
restoration received limited attention before the era
of ‘‘no net loss’’ (Whigham 1999) and so implementation proceeded based on trial and error. Because
wetland ecosystems likely take decades or centuries
to recover (Joosten 1995, Mitsch and Wilson 1996),
long-term ecological studies are critical for advancing the scientific basis for wetland restoration
decision-making. Unfortunately, the ‘‘no net loss’’
policy experiment was not linked to any systematic
883
884
For example, the lower-than-expected productivity
of Spartina foliosa in Sweetwater Marsh reflected
problems in sediment nitrogen levels unlikely to selfcorrect over time because of site preparation and
hydrologic design of the project (Gibson et al. 1994).
Poor growth of Spartina limited the suitability of the
restored wetland for key species, such as the clapper
rail, which was a focus of the mitigation. Vegetation
composition and abundance will frequently reflect
the hydrology and sediment characteristics of
wetlands so these are typically important diagnostic
traits. Understanding how restored wetland communities deviate from natural conditions and how
long those abnormal deviations persist can provide
important insights into the mechanisms of recovery
and offer key guidance for assessing the status of
future restorations (Galatowitsch 2006).
Thousands of wetlands restored in the late 1980s
under the Conservation Reserve Program (CRP) of
the 1985 Food Security Act presented an important
opportunity to advance wetland science and practice. While these programs initially were motivated
by concerns over soil loss and water quality
impairment, they soon became recognized as the
first potential opportunity for landscape-scale habitat restoration in North America. Confidence was
so high at the start of CRP that federal agencies had
no provisions for ecological monitoring of completed projects. Insights from CRP have come primarily
from individual researchers who studied particular
aspects of the recovery of these ecosystems (e.g.,
LaGrange and Dinsmore 1989, Delphey and Dinsmore 1993). One such study assessed the vegetation
response to reflooding for 64 wetlands in the
southern prairie pothole region, the glaciated terrain
in Minnesota and adjacent Iowa and South Dakota
that once was a complex of tallgrass prairie and
freshwater wetlands and is now predominantly corn
(Galatowitsch and van der Valk 1995, 1996b).
Prairie pothole wetlands were widely considered
among the simplest to restore because hydrophytic
plants were expected to recolonize efficiently from
long-lived seedbanks and propagules dispersed by
waterfowl (Galatowitsch and van der Valk 1996a).
Therefore, restorations consisted of reflooding by
removal of ditches and tile lines, but did not include
planting or follow-up vegetation management.
Within three years of reflooding, Galatowitsch and
van der Valk (1996b) observed that some plant
guilds were able to recolonize newly flooded habitats
rapidly while others were not arriving and establishing. Remnant seedbanks were depleted after
decades of drainage and cultivation (Wienhold and
van der Valk 1989); colonization differences likely
reflected differential dispersal ability (Galatowitsch
WETLANDS, Volume 28, No. 4, 2008
and van der Valk 1996b, Kettenring 2006). Nine
years later, surveys of these same restorations
showed colonization lags persisted for some guilds,
notably sedge meadow and wet prairie species. As
importantly, they showed that the overall rate of
plant colonization corresponded to wetland isolation in the landscape (Mulhouse and Galatowitsch
2003). Insights from observing these wetlands
periodically during the first 12 years following
restoration suggested it may be necessary to plant
some species that have minimal capacity to disperse
in the now highly fragmented landscapes of this
region and that re-assembly of communities could
be accelerated by selecting restoration sites proximate to natural wetlands. After a decade, the
consequence of slow recolonization of the native
flora was clear: a few invasive species (especially
Phalaris arundinacea) dominated nearly all sites
(Mulhouse and Galatowitsch 2003). Longer-term
data were needed to determine if invasive species
have arrested succession in these restorations and if
further reassembly would occur despite isolation and
competition (Galatowitsch 2006).
In 2007, 37 of the remaining 39 restored wetlands
surveyed by Galatowitsch and van der Valk (1995)
and Mulhouse and Galatowitsch (2003) were
revisited. This study represents one of the longest
data sets on revegetation and community development of restored wetlands. Restored in 1988, these
37 wetlands have been surveyed five times in the last
19 years. This extensive, regional data set allows us
to examine the long-term community development
of restored prairie pothole wetlands. The restoration
of multiple, discrete prairie pothole wetlands such as
these allows us to examine community assembly
including the importance of stochastic processes and
priority effects on restoration projects (Young et al.
2005, Chase 2007). Using this data set, we have
examined 1) the development of plant diversity and
wetland plant guilds over 19 years, 2) species
turnover and changes in beta diversity over time,
3) the importance of dispersal limitation and priority
effects in structuring these communities, and 4) how
successful restored prairie pothole wetlands are in
supporting the plant diversity typical of remnant
natural wetlands.
METHODS
Study sites consisted of thirty-seven 19-year old
restored upland depressional wetlands located in
northern Iowa, southern Minnesota, and southeastern South Dakota (for site map see Mulhouse and
Galatowitsch 2003). These wetlands are all situated
on soils typical for poorly drained depressional
Aronson & Galatowitsch, DEVELOPMENT OF RESTORED PRAIRIE POTHOLES
features in the region, i.e., calcareous clay-loams and
silty clay loams (cumulic haplaquolls, typic haplaquolls) or highly decomposed organic deposits
(sapric and hemic) (Galatowitsch 1993). All of the
restorations were conducted in 1988 by breaking
tiles, plugging ditches, and retaining water with
dikes to achieve a semi-permanent or seasonal water
regime (Galatowitsch and van der Valk 1996b).
While most of these restorations were part of the
federal Conservation Reserve Program, a few in
Minnesota were restored under a similar state
program, Reinvest-In-Minnesota. All restorations
were in active cultivation immediately prior to
restoration (Galatowitsch and van der Valk 1995).
Wetland and aquatic plants have not been planted in
the restorations, so the flora reflects natural
colonizations from seed banks or dispersal.
The original study, performed between 1989 and
1991, consisted of 64 wetlands (Galatowitsch and
van der Valk 1995). In 2007, 39 wetlands remained.
The majority of sites were lost to reconversion to
agricultural production between 1991 and 2000
(Mulhouse and Galatowitsch 2003). Two additional
sites were lost between 2000 and 2007. In 2007, one
site greater than 9 ha was eliminated from the study
because it became impossible to ensure the reliability
of a comprehensive floristic survey with dense
vegetative cover and unconsolidated sediments over
large areas. One other site could not be sampled
because property access was not granted.
Vegetation Survey
Wetlands were surveyed during mid-summer visits
in five sample years: 1989, 1990, 1991, 2000, and
2007, repeating the approach described in Mulhouse
and Galatowitsch (2003). In every year, a complete
floristic list was compiled, and species cover was
estimated for each wetland zone present at a site.
The vegetation of typical prairie potholes develop
into zones corresponding with the depth and
duration of flooding (Galatowitsch and van der
Valk 1996b). These zones typically form in concentric rings of open water, mudflat, emergent, wet
meadow, and buffer (i.e., area between wet meadow
and cultivated land). In the initial surveys done from
1989 to 1991, some zones were not yet completely
developed. Therefore, surveys were done at the
vegetated margin at the high water line in addition
to the open water and mudflat to capture the
developing emergent and wet meadow zones. Zone
classifications were made on site according to
vegetation type and hydrology. Within each zone,
the cover of each species was estimated using a
seven-point scale (Mueller-Dombois and Ellenberg
885
1974) consisting of the following cover classes: r 5
one individual with insignificant cover, + 5 few
individuals with insignificant cover, 1 5 1%–5%
cover, 2 5 . 5%–25% cover, 3 5 . 25%–50%
cover, 4 5 . 50%–75% cover, and 5 5 . 75%
cover. Some species were indistinguishable from
each other in their vegetative states or due to
hybridization; these were categorized into the
following species groups: Amaranthus hybridus/retroflexus, Aster praealtus/simplex/lanceolatus, Bidens
frondosa/vulgata, Calamagrostis canadensis/stricta,
Echinochloa crusgalli/muricata, Potamogeton foliosus/pusillus, and Typha angustifolia/x glauca. Nomenclature follows Great Plain Flora Association
(1986).
Data Analyses
To examine the changes in the entire regional
restored flora, the total number of plant species
observed across all wetland restorations was calculated for 1989 (representing initial establishment),
1991, 2000, and 2007. Obligate upland plants
(USDA, NRCS 2007) were removed from all
analyses. Species were classified into species guilds
based on Galatowitsch and van der Valk (1994),
Galatowitsch et al. (2000), and life history and
wetland indicator status (USDA, NRCS 2007).
These guilds included aquatic species, emergent
perennials, sedge meadow perennials, wet prairie
perennials, woody plants, native annuals, introduced
annuals, and introduced perennials (Mulhouse and
Galatowitsch 2003).
To examine species turnover at the site-level, we
calculated the total number of plant species observed at each wetland restoration for sample years
1991, 2000, and 2007. The change in species richness
between 1991 and 2000 and between 2000 and 2007
was calculated as the difference between the two
time periods and analyzed using repeated measures
ANOVA (Proc MIXED in SAS version 9.1, SAS
Institute, Cary, NC, USA). The frequency of
colonizations and extinctions between 1991 and
2000 and between 2000 and 2007 was calculated
for each wetland. The difference in colonizations
and extinctions between the two time periods was
analyzed using repeated measures ANOVA (Proc
MIXED in SAS version 9.1).
The potential effects of wetland size, isolation,
and hydrology on the proportion of change in
species richness were also evaluated. Wetland size
was determined from field topographic surveys by
using tile or surface outlets as the maximum pool
(Galatowitsch 1993). The size of these 37 wetlands
ranged from 0.2–6.8 ha. Landscape isolation was
886
estimated to be the mean distance (m) to the four
nearest natural wetlands, as described in Mulhouse
and Galatowitsch (2003). Wetlands were classified
based on the permanency of water in the deepest
zone, following Stewart and Kantrud (1971). We
assigned each wetland to a class based on the
frequency of flooding into midsummer and on the
vegetation found in the central, deepest zone.
Frequency of flooding was determined from late
summer aerial photography for federal crop compliance (National Agriculture Imagery Program),
and from water depth records collected during
vegetation surveys and by landowners. Five wetlands that were rarely or never flooded to midsummer and that lacked emergent and open water zones
were classified as ephemeral or temporary (Class I/
II). Twelve wetlands that were typically saturated
but not flooded in midsummer and that had a
central emergent zone where classified as seasonal
(Class III). Twenty wetlands that were flooded
through the summer in most years and had a central
open water zone (with rooted submersed aquatics)
were classified as semi-permanent (Class IV).
The proportion of change in species richness was
calculated for two time periods, 1991 to 2000 and
2000 to 2007, as the proportion of change in species
richness between the sample years. Kruskal-Wallis
non-parametric analysis of variance with SiegelTukey’s test for differences in dispersion (Proc
NPAR1WAY in SAS version 9.1) was used to
determine the differences among the three wetland
classes in the proportion of the change in species
richness for both time periods. Zar (1999) recommends the Siegel-Tukey procedure for non-parametric data. To test if wetland class was confounded
with wetland size or landscape isolation, analysis of
variance (Proc GLM in SAS version 9.1) was used to
determine differences among wetland classes in
wetland size and isolation. Wetland class was
confounded with wetland size; there was a significant difference in size among wetland classes
(ANOVA F 5 4.83, P 5 0.0136). Ephemeral
wetlands on average are smaller (0.32 ha) than
seasonal (1.04 ha) and semi-permanent wetlands
(2.32 ha). Landscape isolation was not significantly
different among wetland classes (ANOVA F 5 0.04,
P 5 0.9596). We also explored the relationship
between the number of species at each site as a
function of landscape isolation, and wetland size/
flooding frequency for 2007 survey data using a
factor-ceiling distribution as had previously been
reported for 1991 and 2000 (Mulhouse and Galatowitsch 2003).
We used multi-response permutation procedures
(MRPP) on species composition data to test for
WETLANDS, Volume 28, No. 4, 2008
changes in beta diversity for each wetland zone (wet
meadow, emergent, and open water) among 1991,
2000, and 2007. The mudflat zone was combined
with the emergent zone for all sample years because
these zones lie at the same elevation. If significant
differences were found among the three sample
years, pairs of sample years were then tested for
differences in composition using MRPP with a
Bonferroni test of P 5 0.05 divided by the number
of tests for each zone (four). MRPP is a nonparametric method for testing differences among
pre-defined groups (McCune and Grace 2002). We
used cover data for this analysis using the midpoints
along the seven point scale (r 5 0.1%, + 5 0.5%, 1 5
3.5%, 2 5 15%, 3 5 37.5%, 4 5 62.5%, 5 5 87.5%).
We performed MRPP on rank-transformed distance
matrices with Sorensen’s distance measure. Multiresponse permutation procedures were performed
using PCORD version 4.1 (McCune and Mefford
1999).
Comparisons to natural wetland plant communities were done using a regional floristic list from
Galatowitsch and van der Valk (1994). Each species
was classified by the hydrologic zone were it was
best represented, its life history guild, and its
distribution (common, infrequent, rare). This list
and was compiled from published floras (primarily
Great Plains Flora Association 1986, Gleason and
Cronquist 1991), state-generated species lists from
Natural Heritage programs, and from herbarium
records at Iowa State University, Ames, Iowa. The
floristic list was reviewed by curators from the major
herbarium collections in the region and other
experienced wetland botanists (see Galatowitsch
and van der Valk 1994). The common, infrequent,
and rare species typical of natural wetlands were
examined for presence in the 37 restored wetlands at
any time period between 1989 and 2007. Common
species of natural wetlands that were present in , 4,
or , 10%, of the restored wetlands at any of the
four sampling periods were also calculated.
RESULTS
Long-term Patterns in the Regional Restored Flora
In 1989, the first year after restoration, the
number of species present across all sites (i.e.,
regional species richness) was 125 (Table 1). Two
years later, in 1991, the regional species richness
increased by 10% to 138 species (6.5 species/yr). The
majority of species accumulation occurred between
1991 and 2000. In these prairie wetlands, regional
species richness increased from 1991 by 94% to 268
species in 2000 (14.4 species/yr). From 2000 to 2007,
Aronson & Galatowitsch, DEVELOPMENT OF RESTORED PRAIRIE POTHOLES
887
Table 1. Species richness for 1989, 1991, 2000, 2007, and total richness of all years for 37 prairie pothole wetlands by
guild. Species new to the flora and lost are presented for two time periods, 1991–2000 and 2000–2007.
Species Richness
Guild
All
Years
1989
1991
2000
2007
New to
Flora 2000
New to
Flora 2007
Species
Lost 2000
Species
Lost 2007
Aquatic
Emergent
Introduced annual
Introduced perennial
Native annual
Sedge meadow
Wet prairie
Woody
Total Richness
17
21
43
35
36
76
71
39
338
7
12
25
21
26
17
11
6
125
13
18
19
20
20
28
15
5
138
15
19
36
26
32
67
46
27
268
15
19
29
27
22
62
66
39
279
3
2
18
8
13
38
31
22
133
1
1
3
6
1
8
24
12
56
1
1
1
2
1
0
0
0
6
2
2
11
6
12
13
4
0
50
the total number of species increased by less than 5%
to 279 species in 2007 (1.6 species/year).
Changes in species richness for the regional flora
differed by wetland plant guild. Within the first
three years after restoration, both aquatic and
emergent species rapidly established and species
richness reached saturation. These guilds did not
accumulate substantial species richness between
1991 and 2007 (Table 1). The aquatic species
richness in 1991 represented over 75% of the total
richness accumulated in the 19 years of this study.
Over 85% of the emergent species that accumulated
over the 19 years of this study were present by the
third year. Likewise, introduced perennial species
also rapidly established, with 60% of the total
species present the first year after restoration.
Species richness of introduced perennials gained
only seven species from 1991 to 2007 (Table 1).
In contrast, sedge meadow, wet prairie, and
woody species exhibited the greatest increase in
species colonization between the third and twelfth
years. In 1991, , 40% of the sedge meadow species
that accumulated over the 19-year study were
present, while by 2000, 88% had colonized. The
sedge meadow guild lost more species than it gained
between 2000 and 2007, resulting in a decline in
richness, from 67 to 62 species. Wet prairie and
woody species continued to accumulate species
between 2000 and 2007. In 2000, only 65% of wet
prairie species and 69% of woody species observed
in restorations had colonized. By 2007, 93% of wet
prairie species and 100% of the woody flora had
colonized (Table 1). The species richness of annual
species, both introduced and native, was variable
throughout the 19 years of this study, with natives
declining between 2000 and 2007 (Table 1).
Only six species were lost between 1991 and 2000
and included aquatic, emergent, invasive annual,
invasive perennial, and native annual species (Ta-
ble 1). Between 2000 and 2007, 51 species were lost.
Over 85% of the species lost were present in two or
fewer wetlands in 2000. Species lost between 2000
and 2007 were largely introduced and native
annuals. In addition, there was a loss of 13 sedge
meadow species between 2000 and 2007 (Table 1).
Long-term Patterns at the Site-scale
As in 2000, the factor-ceiling analysis in 2007
showed that as isolation increased, the maximum
species richness decreased (Mulhouse and Galatowitsch 2003). Departures from this maximum were
related to frequency of flooding and wetland size.
Because there was no change in this relationship in
2007 (i.e., the slopes and intercepts are the same), the
results are not shown.
Patterns in species changes at the site scale
corresponded to those at the regional scale. The
average species richness of restorations decreased
from 60.0 species in 2000 to 56.9 species in 2007.
This followed a gain in average change in species
richness between 1991 and 2000 of 38.0 species
(Figure 1). The average change in species richness
between the two time periods significantly differed
(Repeated measures ANOVA, F 5 89.42, df 5 1, P
, 0.0001). Between 1991 and 2000, all sites gained
species. This gain in species richness ranged from
nine to 97 species. Between 2000 and 2007, less than
one-third of sites gained species. The change in
species richness ranged from 222 to 23 species.
Twenty-five sites lost species between 2000 and 2007.
Change in species richness between 2000 and 2007
was related to wetland class, although this was not
the case between 1991 and 2000. Between 1991 and
2000, ephemeral wetlands experienced changes in
species richness ranging from 59% to 417%. The
proportion of species richness change of seasonal
wetlands ranged from 58% to 520% and for semi-
888
Figure 1. Average number of colonizations, extinctions,
and the average change in species richness (SR) for 37
restored prairie pothole wetlands. Gray bars represent the
time period between 1991–2000, white bars represent the
time period between 2000–2007. Bars with different
lowercase letters are significantly different within category; Colonizations: F 5 74.16, df 5 1, P , 0.0001;
extinctions: F 5 87.50, df 5 1, P , 0.0001; SR: F 5 89.42,
df 5 1, P , 0.0001.
permanent wetlands the proportion of change in
species richness ranged from 45% to 760%. There
were no significant differences in the mean (Kruskal
Wallis X2 5 1.23, df 5 2, P 5 0.5400) or dispersion
(Siegel-Tukey X2 5 0.81, df 5 2, P 5 0.6700) of the
proportion of change in species richness between
1991 and 2000 among wetland classes. The proportion of change in species richness between 2000 and
2007 varied greatly among wetland classes (Figure 2). All of the ephemeral wetlands experienced
changes in species richness of . 10% (6). This was
true for all but one of the seasonal wetlands as well.
In contrast, changes in species richness were greater
than 10% for only 30% of the semi-permanent
wetlands (Figure 2). While there were no significant
differences in the average proportion of species
change among wetland classes (Kruskal Wallis X2 5
1.89, df 5 2, P 5 0.3900), there was a significant
difference in dispersion among wetland classes
(Siegel-Tukey X2 5 12.11, df 5 2, P 5 0.0023).
Small, ephemeral wetlands experienced much greater variability in species richness change than did
larger, semi-permanent wetlands (Figure 2).
The overall rate of new species accumulation in
restorations decreased between 2000 and 2007. New
colonizations per site were significantly greater
between 1991 and 2000 than between 2000 and
2007 (Repeated measures ANOVA, F 5 74.16, df 5
1, P , 0.0001), indicating rapid community assembly between 1991 and 2000 (Figure 1). Between 1991
and 2000, the average number of colonizations per
WETLANDS, Volume 28, No. 4, 2008
Figure 2. Proportion of change in species richness
between 2000 and 2007 (number of species in 2007/
number of species in 2000) for 37 restored prairie wetlands
by wetland class. Black bars represent ephemeral wetlands, light grey bars represent seasonal wetlands, and
dark grey bars represent semi-permanent wetlands.
site was 45.0 (5 species/yr), approximately double
the average colonizations (22.7) between 2000 and
2007 (3.2 species/yr). Extinctions however, significantly increased between 2000 and 2007 when
compared to extinctions between 1991 and 2000
(Repeated measures ANOVA, F 5 87.50, df 5 1, P
, 0.0001). The average number of species extinctions in 2000 (6.8, 0.8/yr) increased by almost 300%
in 2007 (19.8, 2.8/yr) (Figure 1).
The vegetation of the wetland zones, open water,
emergent, and wet meadow, exhibited change over
time (Table 2, Figure 3). In the open water zone,
there was little change in the most frequently
observed species over time (Table 2). In all years,
Potamogeton foliosus/pusillus, P. pectinatus, and
Lemna minor were among the most frequently
observed species in the open water zone. In 1991,
P. nodosus and P. zosteriformis were also frequent.
In 2000 and 2007, Ceratophyllum demersum was a
common species. In all years, P. foliosus/pusillus was
the dominant taxa in terms of percent cover.
Statistically, the vegetation composition of the open
water zone did not significantly differ between 1991
and 2000 or between 2000 and 2007 (Table 3).
However, the open water vegetation of 1991 was
significantly different from the vegetation of 2007
(Table 3). The open water community was significantly more similar across all sites in 2007 than the
vegetation of 1991 (Table 3).
The vegetation of the emergent zone exhibited
greater differences in species composition across
time than the open water zone (Table 2). Multiresponse permutation procedure showed a signifi-
89.5
73.7
57.9
47.4
31.6
50.0
50.0
46.7
43.3
43.3
40.0
52.9
44.1
38.2
35.3
35.3
Potamogeton foliosus/pusillus
Potamogeton pectinatus
Lemna minor
Potamogeton nodosus
Potamogeton zosteriformis
Polygonum amphibium
Scirpus fluviatilis
Lemna minor
Amaranthus rudis
Typha angustifolia/x glauca
Echinochloa crusgalli/ muricata
Elytrigia repens
Cirsium arvense
Bromus inermis
Ambrosia artemisiifolia
Phalaris arundinacea
27.1
9.0
38.2
4.8
16.1
0.9
17.4
32.0
3.6
7.9
11.2
11.2
6.4
5.4
1.1
0.4
Cover
Freq
62.5
58.3
41.7
41.7
37.5
83.3
75.0
72.2
69.4
66.7
61.1
100
92.3
78.4
78.4
75.7
73.0
2000*
Open Water
Potamogeton foliosus/pusillus
Potamogeton pectinatus
Ceratophyllum demersum
Lemna minor
Najas flexilis
Emergent
Phalaris arundinacea
Scirpus fluviatilis
Scirpus validus
Eleocharis palustris
Typha angustifolia/x glauca
Polygonum amphibium
Wet Meadow
Phalaris arundinacea
Cirsium arvense
Asclepias syriaca
Polygonum amphibium
Aster praealtus/simplex/lanceolatus
Rumex crispus
44.6
8.1
1.2
0.7
2.8
0.7
13.8
22.3
4.7
10.1
43.4
1.0
32.3
24.5
0.8
9.1
13.8
Cover
Phalaris arundinacea
Polygonum amphibium
Scirpus fluviatilis
Aster praeltus/simplex/lancelolatus
Cirsium arvense
Asclepias incarnata
Solidago canadensis
Scirpus fluviatilis
Phalaris arundinacea
Polygonum amphibium
Typha angustifolia/x glauca
Scirpus validus
Lemna minor
Potamogeton pectinatus
Lemna minor
Ceratophyllum demersum
Potamogeton foliosus/pusillus
Lemna trisulca
Spirodela polyrrhiza
2007*
100
82.9
80.0
77.1
74.3
71.4
71.4
93.5
90.3
77.4
77.4
74.2
58.1
90.9
81.8
68.2
68.2
59.1
50.0
Freq
66.0
3.72
4.1
3.4
1.0
0.5
1.7
22.3
22.5
3.3
56.3
4.2
23.1
29.7
21.7
7.8
47.2
15.9
7.6
Cover
* The number of sites surveyed in 1991 was 19 in the open water zone, 30 in the emergent zone, and 34 in the wet meadow zone. The number of sites surveyed in 2000 was 24 in the
open water zone, 36 in the emergent zone, and 37 in the wet meadow zone. The number of sites surveyed in 2007 was 22 in the open water zone, 31 in the emergent zone, and 35 in the
wet meadow zone.
Freq
1991*
Zone
Table 2. Dominant species in each wetland zone represented by frequency and average percent cover for sample years 1991, 2000, and 2007. The six most
frequently occurring species observed in greater than one-third of sites surveyed in each zone in each year are listed with their percent occurrence across all sites.
Average percent cover where present of each species are also listed.
Aronson & Galatowitsch, DEVELOPMENT OF RESTORED PRAIRIE POTHOLES
889
890
WETLANDS, Volume 28, No. 4, 2008
Table 3. Ranked transformed MRPP using Sorensen’s
distance measure on cover data of three wetland zones in
37 restored prairie pothole wetlands. Results are for pairs
of sample years for each wetland zone.
Average Sorensen’s
Distance
1991
Figure 3. Average Sorensen’s dissimilarity measure by
sample year for each wetland zone. The vegetation of the
wet meadow zone ( ) was significantly different among
sample years (A 5 0.175, P , 0.0001). Emergent
vegetation (#) was also significantly different among
sample years (A 5 0.122, P , 0.0001). Open water (.)
vegetation was significantly different among sample years
(A 5 0.089, P , 0.0001).
N
cant difference in the species composition of the
emergent zone among all sample years (Table 3,
Figure 3). On average, Typha angustifolia/x glauca
had the greatest percent cover (43.4% and 56.3%) of
any species in 2000 and 2007 (Table 2). In 1991,
Typha angustifolia/x glauca only had an average of
7.9% cover. Phalaris arundinacea was observed in
the emergent zones of over 90% of the sites surveyed
in 2007 but only 20% of sites surveyed in 1991. In all
years, Polygonum amphibium, Scirpus fluviatilis, and
Typha angustifolia/x glauca were among the most
frequently observed species (Table 2). The beta
diversity of the emergent zone decreased among
sites over time, with 1991 having the greatest average
Sorensen’s dissimilarity and 2007 having the smallest
average Sorensen’s dissimilarity (Table 3, Figure 3).
In the wet meadow zone, Phalaris arundinacea
was observed in 100% of sites in 2000 and 2007, but
only 35% of 1991 sites. Of the most frequently
occurring species, P. arundinacea and Cirsium
arvense were the only wet meadow zone species
common in all years. In 2000 and 2007, P.
arundinacea was by far the most dominant species
with an average of 44.6% cover in 2000 and 66.0%
cover in 2007. Species composition over time was
significantly different (Table 3, Figure 3). However,
the vegetation assemblages of 2000 and 2007 were
more similar in composition to each other than to
1991 (Table 3). The beta diversity of the wet
meadow zone appears to be converging with time;
Sorensen’s dissimilarity significantly decreased
across time, with 1991 having the greatest average
2000
Wet Meadow
1991–2000 0.6655
2000–2007
1991–2007 0.6581
0.2183
0.5903
Emergent
1991–2000
2000–2007
1991–2007
0.3151
0.5673
Open Water
1991–2000
2000–2007
1991–2007
0.6035
0.6371
0.5243
0.5643
0.4244
0.5693
2007
A
P
0.3610
0.1678
0.135 ,0.0001
0.042 0.0005
0.181 ,0.0001
0.3821
0.2354
0.108 ,0.0001
0.037 0.002
0.134 ,0.0001
0.4127
0.3027
0.0176*
0.029 0.0335*
0.121 ,0.0001
* Not significant after Bonferroni correction of P , 0.0125.
distance (Table 3, Figure 3). The greatest decrease
in Sorensen’s dissimilarity occurred between 1991
and 2000 (Figure 3).
Comparison to the Flora of Natural Wetlands
Over 22% of species common to natural wetlands
have never colonized the restored wetlands surveyed
in this study (Table 4). The majority of common
species that never colonized were wet prairie (14)
and sedge meadow (10) species. An additional 25
species considered common to natural wetlands were
present in , 10% of restored wetlands over the 19
years of this study (Table 4). Of infrequent species
found in natural wetlands, 70%, or 90 species, have
never colonized the restored wetlands. Rare species
of natural wetlands are most poorly represented in
the restored wetland flora. More than 93% of rare
species found in natural wetlands have not colonized
the restored wetlands in this study. Only five out of
78 rare species listed have colonized the restored
wetlands.
DISCUSSION
Clear patterns emerge from this long-term study
of restored prairie pothole wetlands. First, regional
plant species richness of restored wetlands stabilized
within 19 years, with most reassembly occurring in
the first 12 years following reflooding. Second,
wetland plant guilds varied in their colonization
efficiencies, with some guilds stabilizing after 3 years
Aronson & Galatowitsch, DEVELOPMENT OF RESTORED PRAIRIE POTHOLES
891
Table 4. Species considered common to prairie potholes (Galatowitsch and van der Valk 1994) but missing from or
present in less than 10% of study sites in the 37 restored prairie pothole wetlands surveyed 19 years post-restoration.
Guild
Missing Species
Aquatic
Wolffia columbiana
Emergent
Glyceria striata
Iris shrevei
Sagittaria cuneata
, 10% Species
Ricciocarpus natans
Introduced Annual
Polygonum hydropiper*
Native Annual
Eragrostis hypnoides
Hordeum pusillus
Cyperus aristatus*
Eleocharis obtusa
Panicum capillare
Sedge Meadow
Amorpha fruticosa
Cardamine pensylvanica
Carex hystericina
Equisetum pratense
Lysimachia hybrida
Ranunculus cymbalaria
Senecio aureus
Stellaria crassifolia
Veronica anagallis-aquatica
Veronica peregrine
Aster novae-angliae*
Calamagrostis canadensis/stricta
Caltha palustris
Carex stipata
Carex stricta
Cicuta maculata
Lippia lanceolata*
Lycopus uniflorus
Lysimachia ciliata
Lysimachia quadrifolia
Stellaria longifolia
Wet Prairie
Agrostis hyemalis
Allium canadense
Anemone canadensis
Cirsium flodmanii
Galium obtusum
Heracleum sphondylium
Hypoxis hirsuta
Liatris ligulistylis
Liatris pycnostachya
Phlox pilosa
Pycnanthemum virginianum
Senecio pauperculus
Thalictrum dasycarpum
Veronicastrum virginicum
Cerastium nutans*
Cyperus odoratus
Desmodium canadense
Helenium autumnale*
Lathyrus palustris
Lobelia spicata
Silphium laciniatum
Silphium perfoliatum
Woody
Salix discolor
Salix petiolaris
Salix rigida
Salix amygdaloides
* Not observed at any sites in 2007.
while others are still accumulating species after 19
years. Third, the vegetation of the wetland zones
across all sites was converging with time, resulting in
a decrease of beta diversity and the biotic simplification of the regional restored wetland flora.
Fourth, larger, semi-permanent wetlands experienced less species change and therefore greater
community stability than smaller, drier wetlands.
Finally, restored wetlands, even after nearly 20 years
of community development, do not function to
support the plant diversity of natural wetlands. The
patterns found in this study indicate that the assembly
of communities in prairie pothole wetlands after
restoration is driven primarily by stochastic processes
(e.g., Chase 2007). These communities are primarily
dispersal assembled communities with strong priority
effects driven by invasive plant species.
Long-term Patterns in the Regional Restored Flora
It is clear from this study that the majority of the
flora present in restored prairie pothole wetlands
will colonize in the first 12 years after restoration.
The greatest increase in species richness occurred
between years three and 12 and we observed only
minor additions to the flora in the last seven years,
892
indicating that the first 12 years of a restoration is
the most important for community assembly. Now
that species accumulation has slowed and extinctions are higher than new colonizations in these
restorations, it is unlikely that significant species
accumulation will occur in the future. Species
richness has stabilized and new additions will be
minor. However, this pattern could not have been
detected if the surveys of these restored wetlands had
not extended over two decades.
According to the efficient community hypothesis,
vegetation should reassemble rapidly after the
abiotic conditions are restored (Galatowitsch and
van der Valk 1996a). Because these restorations took
10 to 20 years for the development of the vegetation
to stabilize after hydrology was restored, they are
much less efficient than was widely presumed by
practitioners. Analysis of species guilds shows,
however, that colonization efficiency varies greatly,
with three of seven guilds capable of rapid reassembly. This efficiency likely stems from differences in
propagule pressure, dispersal ability, and the availability of safe sites for establishment. With wetland
losses across this region exceeding 90% (Dahl and
Johnson 1991), most of the suitable habitat for
hydrophytic species occurs along roadside ditches,
which are nearly universally dominated by invasive
perennials, such as Phalaris arundinacea and Typha
angustifolia/x glauca. In a study of seed rain in
wetlands within this region, Kettenring (2006) found
that introduced/invasive species (including Typha
and Phalaris) accounted for . 90% of seed rain in
natural wetlands and 77% in restored wetlands. Not
surprisingly, introduced/invasive perennials are one
of the high-efficiency guilds in this region. Species in
the other two high efficiency guilds, submersed and
floating aquatics and emergent perennials, likely
benefited from dispersal by waterbirds (primarily
ducks) that make frequent trips between wetlands
with open water (LaGrange and Dinsmore 1989,
Delphey and Dinsmore 1993, Mueller and van der
Valk 2002). Waterfowl are estimated to disperse
seeds between 20 to 30 km and possibly up to
1,400 km (Mueller and van der Valk 2002). Therefore, although these restored wetlands are isolated
between 0.2 and 2.6 km from natural wetlands
(Mulhouse and Galatowitsch 2003), the aquatic
and emergent guilds are effective colonizers even in
this highly fragmented agricultural landscape.
Three guilds of native perennials that are low
efficiency colonizers, wet prairie, sedge meadow, and
woody species, lack the advantage of either high
propagule availability or a reliable dispersal vector
(Galatowitsch and van der Valk 1996a). Consequently, species in these guilds are potentially
WETLANDS, Volume 28, No. 4, 2008
limited by the lack of safe sites, pre-empted by
dense stands of perennial grasses (primarily Phalaris,
but also Bromus inermis) that form monotypic
stands around the perimeter of nearly all sites.
Phalaris has been shown to preclude simultaneously
establishing native species (Reinhardt Adams and
Galatowitsch 2006) by its ability to rapidly establish
and grow (Reinhardt Adams and Galatowitsch
2005), limiting light availability to slower growing
native perennials (Wetzel and van der Valk 1998).
Long-term Patterns at the Site-scale
Competition with invasive species is one explanation for the loss in sedge meadow species observed
between 2000 and 2007; these losses were anticipated
because most species observed in 2000 were represented only by small populations (Mulhouse and
Galatowitsch 2003). Over 85% of the extinctions
that occurred between 2000 and 2007 were species
with small populations (i.e., few individuals previously recorded) at only one or two of the 37
wetlands surveyed. In addition to competitive
exclusion, small populations are more likely to
suffer stochastic extinctions (Shaffer 1981, Soulé
1986), particularly in isolated habitats in fragmented
landscapes (Saunders et al. 1991) and due to typical
hydrologic fluctuations. Another explanation is that
small populations were not found again because of
low detectability. Finally, management activities at
some wetlands, particularly herbicide treatment of
Cirsium arvense, may have affected many forb
species that occupy the wetland perimeter (personal
observations in all study years).
The beta diversity of these wetlands decreased
over time showing biotic simplification of the
wetland flora; this pattern was most pronounced
for the emergent and wet meadow zones. Invasive
Typha angustifolia/x glauca and Phalaris arundinacea
became more frequent and increased in dominance
across the restored wetlands particularly between
2000 and 2007 and could be driving the biotic
simplification of the emergent and wet meadow
zones, respectively. In this study, it was not possible
to analyze the relationship between native species
diversity and invasive species due to the high
abundance of the invasive species in all years at all
sites. Both Typha angustifolia/x glauca and Phalaris
arundinacea form dense, monodominant stands,
with little available space or light for other species
(Werner and Zedler 2002, Maurer et al. 2003). In
Pennsylvania, low species richness of created wetlands was attributed to both dispersal limitations of
natives and the invasion of Typha spp. at the created
wetlands (Campbell et al. 2002).
Aronson & Galatowitsch, DEVELOPMENT OF RESTORED PRAIRIE POTHOLES
After almost 20 years, these restorations lack the
expansive sedge meadows typical of prairie pothole
wetlands in this region (Galatowitsch and van der
Valk 1994). Only four restorations had Carex
species cover over 30% by 2007. Phalaris quickly
established after restoration and was present at 35%
of the restored wetlands in 1991, with large
populations at some sites. In contrast, no restorations had Carex spp. cover of more than 15% in
1991. This lack of an early developing sedge
meadow zone on all but a few sites would have
allowed Phalaris to overtake available suitable
habitat and quickly dominate (Wetzel and van der
Valk 1998, Budelsky and Galatowitsch 2000, Green
and Galatowitsch 2001, Perry et al. 2004, Reinhardt
Adams and Galatowitsch 2006). In these restored
wetlands, Phalaris established early and the priority
effects of this species are still evident today.
We observed that smaller, drier wetlands are
floristically less stable in the long term than larger,
semi-permanent wetlands, exhibiting greater turnover in species richness between sampling intervals.
The hydrology of shallow, natural wetlands is more
sensitive to annual variability in precipitation
(Winter 2000), not only drying earlier each year,
but also experiencing more years with no standing
water. In a wet year, rapid colonizers, such as some
aquatic and emergent plants, may appear on a site
but then disappear during years of below-average
rainfall. In contrast, deeper, semipermanent wetlands will be more likely to have at least small areas
of saturated or flooded habitat in a given year,
serving as refugial sites in all but the most severe
droughts. Another cause of rapid species turnover in
small, temporarily flooded restored wetlands could
be that they are hydrologically anomalous compared
to natural, temporary wetlands or semipermanent
wetlands. Because of federal guidelines, restored
wetlands almost universally have a dike constructed
at the outlet, formed by excavating a deep depression within the basin (Galatowitsch and van der
Valk 1994). When the volume of water stored in a
wetland is low (as is the case in temporary wetlands),
it is confined to this small excavated area, rather
than saturating the soil across the wetland basin.
These restored wetlands are not developing a sedge
meadow flora, but instead have a small area that
alternates between open water and mudflat, with
most of the basin persistently dry.
Comparison to the Flora of Natural Wetlands
While common species appear to succeed, at least
in terms of species richness, in restored wetlands,
infrequent and rare species of natural wetlands do
893
not. Infrequent and rare species found in natural
wetlands are poorly represented in these restored
wetlands. The majority of species common to
natural wetlands that have never been found in
these 37 restored wetlands were sedge meadow and
wet prairie perennials. For example, 16 of the 40
common Carex species were not found in the
restorations. Low propagule pressure, poor dispersal
ability (Galatowitsch and van der Valk 1996c,
Seabloom and van der Valk 2003), and competition
with invasive plant species (Green and Galatowitsch
2002) are likely limiting to many species, including
these Carex spp., that are now restricted to only a
handful of small protected areas (Kettenring 2006).
In addition, these 19-year old restored wetlands
still have a higher abundance of annuals than do
natural wetlands, including Amaranthus rudis, Bidens cernua, Polygonum pensylvanicum, and Rorippa
palustris. These annuals are often highly associated
with restored prairie pothole wetlands (Seabloom
and van der Valk 2003, Galatowitsch and van der
Valk 1996a). In addition, Phalaris arundinacea, one
of the dominant species of all wetlands studied here,
although found in natural wetlands is not as
abundant or widespread as it is in restored wetlands
(Galatowitsch and van der Valk 1996a). In emergent
wetlands in Connecticut, 12-year old created wetlands were dominated by invasive species, mainly
Phragmites australis, while native species, Carex
stricta and Typha latifolia, dominated at natural
wetlands (Moore et al. 1999).
CONCLUSIONS
Over the past 20 years, implementation of the ‘‘no
net loss’’ wetland policy progressed from the simple
goal of avoiding losses in aerial extent, to maintaining the wetland base within a locale and the
distribution of habitat types (e.g., Zedler 1996).
Advancing restoration standards so these habitats
reliably support a comparable level of biodiversity to
natural wetlands remains a critical need (Zedler
2000). Restored wetlands may fail to provide
expected biodiversity support because there has
not been adequate time for recovery and reassembly
or because there are abiotic and/or biotic barriers
that are likely to persist. In the case of prairie
pothole restorations, results of our long-term study
suggests that the recovery/reassembly phase lasted
approximately a decade, and that the vegetation of
these sites will not eventually resemble that of
natural wetlands. Biotic barriers, chiefly the differential in colonization efficiency between invasive
and native perennials, limited the recovery of all
study wetlands, whereas abiotic barriers (i.e.,
894
unfavorable hydrology) may be important in seasonal and ephemeral wetlands. Our findings demonstrate that addressing both biotic and abiotic
barriers to recovery is necessary when planning,
designing, and implementing wetland restorations so
they support typical levels of biodiversity for that
habitat type.
To ensure a higher level of recovery of prairie
pothole restorations, we outline five guidelines based
on this long-term study. First, the selection of
restoration sites should be prioritized near remnant
natural wetlands in order to increase propagule
pressure and decrease dispersal limitation of native
perennials. Landscape isolation has been shown to
be an important factor in the dispersal limitation of
aquatic plants (Godwin 1923). With reduced isolation, recovery of species in the sedge meadow and
wet prairie guilds increases (Galatowitsch 2006).
Second, restoration of semi-permanently flooded
wetlands along with seasonal and ephemeral wetlands should be implemented. Historically, wetlands
in the prairie pothole region were part of complexes
of small and large wetlands. Semi-permanent
wetlands are key elements of wetland complexes as
they are floristically more stable than ephemeral and
seasonal wetlands. Third, less frequently flooded
wetlands (seasonal to ephemeral) should not be
constructed by excavating pits to make dikes, which
limits soil saturation across the basin and instead
creates hydrological ‘‘flashy’’ pits favoring shortlived annuals and aquatics rather than sedge
meadow and wet prairie perennials. Fourth, active
revegetation via planting of vegetative stock and
seeding of native sedge meadow and wet prairie
perennials is necessary to ensure restoration success.
Restored wetlands can efficiently resemble natural
wetlands with planting (Galatowitsch 2006). Planting is especially important for low efficiency guilds.
Planting native species early in the restoration
process to establish canopy cover will also reduce
the available habitat for Phalaris arundinacea
(Lindig-Cisneros and Zedler 2002, Maurer et al.
2003). Finally, control of invasive plants species,
such as Phalaris and Typha, is essential to maximize
biodiversity. Control of Phalaris must start early in
the restoration process so that native communities
are allowed to establish (Green and Galatowitsch
2001) and must be factored into long-term management plans.
For existing restored wetlands in the region,
facilitating additional species accumulation may
actually be more problematic because removing
biotic barriers is likely more difficult than allowing
them to arise in the first place. Because the cover of
invasive perennials is very high on all sites, eradica-
WETLANDS, Volume 28, No. 4, 2008
tion will require herbicides and will often result in the
loss of most of the species that have colonized over
the past 20 years. Since seedbanks of the invasives
are likely to be abundant, weed control would need
to be pursued for many years. Without such
measures, we anticipate that species turnover will
continue to be driven by local extinctions, rather
than colonizations, and will drive the vegetation
composition over time. In the past decade, we
observed that biotic homogenization of the flora
has begun. This homogenization will likely trigger
further declines in species richness (McKinney and
Lockwood 2001) if management practices are not
implemented. Following this loss of plant diversity
and the lack of expansive sedge meadows, it can be
expected that other taxa, such as the avian communities will also decrease (Delphey and Dinsmore
1993). The restoration of wetlands in the prairie
pothole region to support regional biodiversity is of
critical importance. Given the guidelines we have
outlined above, we believe restored prairie pothole
wetlands can support this regional biodiversity.
ACKNOWLEDGMENTS
We thank Paul Wetzel and Eric Seabloom who
provided helpful suggestions for data analysis. The
University of Minnesota’s Fesler-Lampert Endowment provided financial support.
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Manuscript received 10 June 2008; accepted 25 August 2008.