Urban Wetland Reconstruction Impacts on Vegetation: A Case Study
Megan A. Larson1,2*, Julien Shepherd1, John E. Titus1
1Department of Biological Sciences, PO Box 6000, State University of New York at Binghamton, Binghamton, NY 13902.2Center for Integrated Watershed Studies, Binghamton University, Binghamton, NY 13902. *Corresponding author.
Urban Naturalist, No. 45 (2021)
Abstract
This study documents the effects on the seed bank and standing vegetation of an urban wetland after its expansion to accommodate increased runoff. While total wetland area doubled, densities of seedlings emerging from soil cores declined by 66%. Significant species compositional changes occurred in the seed bank, as did changes among functional groups: non-natives unexpectedly declined, while annuals and graminoids increased. The standing vegetation consisted overwhelmingly of herbaceous taxa before and after the reconstruction, and plant cover, after complete denudation of the vegetation, recovered within 3 years. Compositional changes included unexpected declines in the relative cover of non-native species and increases in the relative cover of graminoids and obligate wetland species—the last in keeping with observed hydrologic change. Seeding and planting may not be necessary, even to reach short-term goals (i.e., less than 5 years) of high plant cover, in reconstructed wetlands.
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UUrban Naturalist
M. A. Larson, J. Shepherd, and J. E. Titus
2021 No. 45
1
2021 Urban Naturalist 45:1–16
Urban Wetland Reconstruction Impacts on Vegetation:
A Case Study
Megan A. Larson1,2,*, Julian Shepherd1, and John E. Titus1
Abstract - This study documents the effects on the seed bank and standing vegetation of an urban wetland
after its expansion to accommodate increased runoff. While total wetland area doubled, densities of seedlings
emerging from soil cores declined by 66%. Significant species compositional changes occurred in the seed
bank, as did changes among functional groups: non-natives unexpectedly declined, while annuals and
graminoids increased. The standing vegetation consisted overwhelmingly of herbaceous taxa before and
after the reconstruction, and plant cover, after complete denudation of the vegetation, recovered within 3
years. Compositional changes included unexpected declines in the relative cover of non-native species and
increases in the relative cover of graminoids and obligate wetland species—the last in keeping with observed
hydrologic change. Seeding and planting may not be necessary, even to reach short-term goals (i.e., less than
5 years) of high plant cover, in reconstructed wetlands.
Introduction
The projected global increase in urban areas (Nilon et al. 2017, Seto et al. 2012) will surely
encroach upon countless more wetlands. Furthermore, land use managers are likely to promote
the creation of new wetlands and the expansion of existing wetlands to accommodate increases in
urban stormwater runoff. The increased runoff that accompanies urbanization leads many cities
to look toward green infrastructure, such as retention ponds and treatment wetlands, to mitigate
runoff impacts (Balderas Guzman et al. 2018). In this case study, we take advantage of an unusual
opportunity to evaluate the impacts of a major reconstruction project on the vegetation of a wetland
that will receive increased runoff. This project, which completely regraded the entire surface of a
previously constructed wetland and doubled its size, serves as an example of what may be expected
in similar ecosystems in the future.
Vegetation is a key component of wetlands, which promote sedimentation, reduce flooding
downstream, and improve water quality through reductions in inorganic nitrogen and phosphorus
(Mitsch and Gosselink 2015). Different species of plants may vary in their ecological functions (e.g.,
Kao et al. 2003), and this potential variation motivates our focus on changes in plant community
composition subsequent to the reconstruction project.
Urban wetlands may experience frequent disturbances (Grayson et al. 1999) from erosion due
to altered hydrology (Ravit et al. 2017), removal of aboveground biomass for crop harvest (Vécrin
et al. 2007), invasive species management (Lawrence et al. 2016), and stormwater control (Blecken
et al. 2017). Because wetland construction projects commonly fail to monitor plant community
establishment, little is known about the response of urban wetland vegetation to disturbances (Zedler
2000). Emergence of seedlings from soil seed banks (Alderton et al. 2017, Kaplan et al. 2014,
Muller et al. 2013, Nishihiro et al. 2006) and regrowth from surviving plant parts (Combroux and
Bornette 2004, Combroux et al. 2002) have great potential to recolonize soil surfaces denuded of
vegetation. Disturbance may favor subsequent colonization by invasive species (D’Antonio and
Meyerson 2002, Matthews et al. 2009) as well as annuals (e.g., van der Valk 1981). The seed bank
1Department of Biological Sciences, PO Box 6000, State University of New York at Binghamton, Binghamton, NY
13902 USA2 Center for Integrated Watershed Studies, Binghamton University, Binghamton, NY 13902 USA
*Corresponding author: mlarson2@binghamton.edu.
Manuscript Editor: Paige Warren
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itself may also be altered by major disturbance (Neff et al. 2009, Osunkoya et al. 2014). The extensive
mixing and spreading of the sediment in our expanded wetland led to 3 hypotheses regarding the
influence of disturbance on the plant community: (1) the soil seed bank would be diluted, resulting
in reduced density of seedlings emerging from surface sediment samples, and there would be (2)
an increase in non-native plants and (3) an increase in annual plants in the standing vegetation
subsequent to the regrading.
In addition to disturbance per se, changes to wetland topography may alter the plant community
in terms of dominant wetland indicator status (WIS), as wetland plants are distributed within
wetlands by their responses to water level (e.g., Keddy and Ellis 1985, Roznere and Titus 2017,
Tiner 1991). In this case study, engineering plans called for greater topographic diversity in the
reconstructed wetland, both by creation of a deeper channel through the wetland and by raising
mounds elsewhere at the site. This planned topographic variation led to our fourth hypothesis: there
would be a broadening in the spectrum of plant species toward greater diversity of WIS categories
in the standing vegetation.
Furthermore, our data allowed us to compare the species composition and growth habits of
the seedlings that emerged from the seed bank before and after the regrade as well as examine
successional changes in the standing vegetation for 3 years after the construction project.
Study Site and Methods
Study Site
Lieberman is a small (0.15 ha) urban retention wetland located on the Binghamton University
campus in Vestal, NY, USA (Lat 42.087°, Long -75.962°; altitude 308 m). This site is referred to as
“Site 1” in Larson et al. (2016) and Larson and Titus (2018). The site is relatively isolated and does
not receive hydrological input from other wetlands, with the nearest wetland complex ~0.6 km away.
Lieberman receives runoff from 0.56 km2 of campus, including parking lots, paved roadways and
sidewalks, and buildings (Kearney et al. 2013). The main inlet, via culverts and drainage ditches, is
located in the southwest corner of the site (Fig. 1). Groundwater also seeps into the wetland, which
drains through a culvert into Fuller Hollow Creek, which in turn discharges into the Susquehanna
River (Zhu et al. 2008), the largest tributary of Chesapeake Bay. Binghamton lies in a deciduous
forest biome with a humid continental climate. Mean annual precipitation is 99.8 cm and mean
monthly low and high temperatures are 3.4 and 12.5°C, respectively (Wikipedia 2020). The campus
supports a relatively high volume of traffic: ~48% of the 13,000 full-time undergraduate students
commute to campus, in addition to administrators, faculty, staff, and ~4000 graduate students
(College Board 2016).
In 2004, the pre-existing pond was drained and a berm was built along the east side of the pond;
the resulting wetland supported wetland vegetation throughout and served as a retention pond to
accommodate campus runoff (Fig. 1A). This stormwater retention wetland included a small channel
near the inlet that opened to a larger, inundated marsh dominated by Sagittaria latifolia Willd.
(Broadleaf Arrowhead) and Alisma triviale Pursh (Northern Water Plantain). Myosotis scorpioides
L. (Water Forget-me-not), Typha x glauca Godr. (Hybrid Cattail), and Leersia oryzoides (L.) Sw.
(Rice Cutgrass) were common near the inlet and around the perimeter of the wetland (Larson et
al. 2016). There have been occasional sightings of herbivores, including Odocoileus virginianus
(Zimmerman) (White-tailed Deer), Ondatra zibethicus (L.) (Muskrats), and Branta canadensis
(L.) (Canada Geese), although the impacts of these species on the vegetation are unknown.
The wetland was regraded to accommodate increases in runoff from newly constructed
impervious surfaces on campus. Construction of the expanded Lieberman wetland with heavy
construction equipment began in July 2011 and resulted in haphazard earth moving and burial of
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Figure 1. Aerial photographs of Lieberman (A) before regrading and (B) after regrading. The white
dotted lines indicate the wetland border. Google Earth Pro 7.3.3.7786 (April 2006 and 13 May 2015,
respectively). Vestal, New York, USA. Google, New York GIS 2021 [28 March 2021].
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existing vegetation (Fig. 2) as well as the upheaval and spreading of wetland sediment over ~0.34
ha—more than doubling the wetland area. Sediment accumulation ponds, initially too deep to support
vegetation, were added near the main inlet and outlet and connected by a meandering channel (Fig.
1B). Landscape was created using sediment from the wetland. Regrading was completed in the early
spring of 2012, at which point topographic variation was not realized, because of the slumping of
relatively fluid sediments. No seeding occurred. Experimental plant propagules were added to the
wetland in July 2012, but all died within the first growing season. These experimental plants were
not planted in survey locations.
Seed Bank Collection
Sampling locations were selected randomly in April 2011 (prior to regrading) at 5 points
along each of 3 randomly selected transects perpendicular to a baseline bordering the east side of
the wetland. At each point, 2 sediment cores (15.2 cm diameter, 5 cm deep) were collected, for a
total of 30 cores. Standing water was present at only 1 point. Sediment was stored in a cold room
at 4.4 °C for 1 month. In early May 2014 (>2 years after regrading was completed), we collected
2 randomly selected cores from 6 random transects, for a total of 12 sediment cores.
Experimental setups for both seed banks were described in Larson and Titus (2018), based on
van der Valk and Davis (1978). These studies were conducted in temperature-controlled 1200-L
fiberglass tanks in the Research Greenhouse at Binghamton University. Large debris, including
rhizomes and tubers, was removed from the sediment. The 2 samples from each point were
combined, homogenized, spread in a 1-cm-thick layer over sand in germination trays, and subjected
to 2 treatments to maximize seedling yield: a permanent drawdown treatment with water levels 5
cm below the sediment surface and a permanent flooded treatment with water levels 5.5 cm above
the sediment surface. The germination trays were exposed to natural light for the duration of the
studies. Water temperatures were maintained at 23 °C by refrigerated circulators (CFF-500, Remcor,
Franklin Park, IL). Seedling data were collected weekly until seedling emergence ceased. Seedlings
that could not be identified in the germination trays were transplanted into separate pots until they
could be identified. Seedling densities for the 2 treatments combined were expressed as counts
per m2 of original sediment core surface. Data were also summarized as relative seedling density
for each species, i.e., the percent of all seedlings attributed to that species. For each seed bank,
a species was considered “common” if its relative seedling density exceeded 5% in either water
level treatment. Seedling counts were also summarized for the following functional groups: native
status (native vs. non-native), longevity (annual, biennial, or perennial), WIS (obligate wetland =
OBL; facultative wetland = FACW; facultative = FAC; facultative upland = FACU; and upland=
U), and growth habit (graminoid, forb, and vine).
Vegetation Sampling
For every sampling period, quadrats were chosen by randomly selecting transects perpendicular
to a baseline bordering the eastern edge of the wetland. Standing vegetation was sampled in June
2011 from 3 sampling points on each of 5 transects just prior to regrading (Larson et al. 2016). For
3 growing seasons after the regrade (2012–2014), we recorded the progression of revegetation by
sampling the standing vegetation in early July and early August each year. Absolute cover estimates
were recorded for each herbaceous species within 1-m2 quadrats, to the nearest 5% (Mueller-Dombois
and Ellenberg 1974). We recorded vegetation data from 4 quadrats along each of 11 transects in
the 3 growing seasons after regrading. Thus, 44 plots were sampled on each occasion, except in
2012 when only 39 plots were sampled because flooding limited access. For each species, cover
was also summarized as relative cover, i.e., the percent of the total cover recorded for all species,
and similarly summarized for each category within the functional groups. Species in the standing
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vegetation were considered “common” if the relative percent cover was greater than 5%. For some
analyses, data were summarized as mean percent cover, or the total cover of a species divided by the
number of sampling quadrats, and, for other analyses, data were converted to a presence–absence
basis to address the concern that vegetation samples were taken in different months. The presence
vs. absence of a species is likely to change much less on a seasonal basis than cover. Functional
group categories for species in the standing vegetation were the same as for seedlings, with two
exceptions: there were only four WIS categories because there were no upland (U) plants in the
standing vegetation, and the growth habit categories were graminoid, forb, and tree/shrub.
Plant Identification
Taxa were identified to the species level when possible using Gleason and Cronquist (1991),
with nomenclature updated according to the New York Flora Atlas (Weldy et al. 2017). We
identified 92.2% (2011) and 96.9% (2014) of seedlings at least to the genus level for the seed bank
assessments. The native status, longevity, WIS, and growth habit of each species were found using
the USDA plant database (USDA NRCS 2012) for the northeast region and the New York Flora
Atlas (Weldy et al. 2017). Cattails in the standing vegetation were identified as the invasive and
non-native Typha x glauca because of substantial variation in the gap size between male and female
flowers, as well as leaf width (Selbo and Snow 2004). Typha seedlings were identified as Typha sp.
because of a lack of these morphological traits. Five species of Juncus (Juncus acuminatus Michx.
[Tapertip Rush]; Juncus articulatus L. [Jointleaf Rush]; Juncus bufonius L. [Toad Rush]; Juncus
effusus L. [Common Rush]; and Juncus tenuis Willd. [Poverty Rush]) were identified but could
not be reliably identified to species. Therefore, for the purposes of our taxonomic, longevity, and
WIS analyses on seedlings, all Juncus species were combined into a single Juncus spp. category.
Data Analyses
To test our seed bank dilution hypothesis, we applied a Kruskal–Wallis test (Social Science
Statistics 2020) to compare mean seedling densities (sum of both water level treatments) from the
2011 and 2014 cores sampled. We used the same nonparametric test to compare species richness
of these 2 seedling communities. Compositional differences between the communities, on the basis
of species and functional groups, were assessed with chi-square tests based on seedling counts.
Figure 2. Photograph of the regrading
process in Lieberman,
depicting the complete denudation
of the wetland. Picture
taken by Megan Larson, used
with permission.
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For these chi-square tests, the null hypothesis tested was that the distribution of seedlings among
categories (species or each type of functional group) was the same for both years. Species and
categories were combined as necessary to ensure that all expected values were at least 5. We also
compared seed bank species compositions using Sørenson’s similarity index (Sørenson 1948) based
on seedling relative densities.
Changes in standing vegetation composition were summarized by calculating species richness
and Shannon–Wiener diversity (H’). To compare the pre-disturbance vegetation with its development
after regrading, the Mahalanobis distance (MD) between the 2011 vegetation sample and the 6 July
and August 2012–2014 samples was used to evaluate the distinctness of the 2011 pre-disturbance
vegetation for the 4 functional group categorizations. In each case, the data consisted of the number
of seedlings present in each category (2 categories each for native status and growth habit, 3 for
longevity, and 5 for WIS), and a variance–covariance matrix for 2012–2014 data was constructed in
Excel. The inverse of this matrix was determined manually following the procedure of van Biezen
(2013), and the MD was calculated in Excel according to McCaffrey (2017). Manual calculation of
the inverse matrix was feasible for these functional groups but not for the large number of species
in the dataset. The resulting MD values were compared to critical values for chi-square with the
degrees of freedom equaling the number of categories (McCaffrey 2017).
The related concept of outlier analysis was applied to evaluate differences in species composition,
expressed as presence-absence rather than cover data, between the 2011 sample and separate
samplings for July and August for 2012-2014. PC-ORD software (McCune and Mefford, 1999)
compared the Sørensen’s distances among the 7 samples, with an outlier defined as a sample for
which the mean distance to the other samples exceeded 2 standard deviations away from the overall
mean distances among the 7 samples.
Analyses of variances (ANOVAs) were performed to test for among- year differences in cover
and richness for all vegetation plots sampled in 2011-2014, and separately for 2012-2014. For each
year after the regrade, July and August data were averaged for each plot, and the same plots were
used throughout this period. Significant overall results were further analyzed at the P = 0.05 level
with Tukey’s honestly significant difference (HSD) tests.
Results
Seed Bank: Before vs. After the Regrade
We observed several substantial changes in the seed bank following the regrading project.
Seedling densities before regrading were nearly 3-fold greater than those in 2014 (Table 1; H =
6.01, df = 19, P = 0.014). Although we observed a decrease in the overall number of species (53
species before the regrade, 37 species afterwards), the decrease in species richness per sampling
point (2 cores) was not statistically significant (Table 1; H = 1.27, df = 19, P = 0.259).
The similarity index between the 2 seed bank surveys was 42% because of similar common
species, including Juncus spp. and L. oryzoides (Table 2). Alisma triviale was relatively more
abundant under flooded conditions before the regrade, while Schoenoplectus tabernaemontani
(C.C. Gmel.) Palla (Softstem Bulrush) was common after the regrade but not before. Despite some
similarities in common species, we observed an overall shift in species composition (c2 = 694.5, df
= 10, P < 0.001), likely due to the increase in S. tabernaemontani and Eleocharis sp. (Spikerush)
as well as a decrease in A. triviale, Lemna minor L. (Common Duckweed), and M. scorpioides in
2014. We did not observe Typha sp. seedlings before the regrade, and only 7 emerged afterwards.
Table 3 presents the results of the c2 tests, which showed a significant change in native status,
as the percent of non-natives declined from 14.1% of seedlings before the regrade to 7.8% after.
Proportions of seedlings in longevity categories also changed significantly, as the percent of
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annuals rose more than 6-fold from 0.7% to 4.7%. There was no significant change among WIS
categories, but growth habit profiles changed significantly, as graminoids rose from 58.9% to 86.0%
of seedlings, while forbs fell correspondingly.
Standing Vegetation: Before vs. After the Regrade
Mean absolute cover for all species per plot initially declined from 70.1% in 2011 to 50.6%
in 2012; thereafter, it increased to 88.8% in 2013 and 116.3% in 2014 (ANOVA, F3,138 = 19.7, P <
0.0001). Table 4 summarizes the species with the highest relative cover to compare the standing
vegetation prior to the regrade (2011) with the 3 subsequent years, for which relative cover was
averaged over the July and August samplings. The most notable changes from 2011 to 2012 were the
precipitous declines in relative cover for the non-natives M. scorpioides and Typha x glauca and the
sharp increases in Potamogeton sp. (Pondweed) and S. latifolia. During the 3 growing seasons after
the regrade, cover rose monotonically for L. oryzoides and S. tabernaemontani, remained high for S.
latifolia, and declined for Potamogeton sp. Outlier analysis showed that the species composition in
2011 differed substantially from the post-regrade sampling periods by virtue of its Sørensen’s distance
being 2.11 standard deviations from the mean of all 7 vegetation samplings. Species richness per
plot in the standing vegetation significantly decreased immediately after the regrade, then rebounded
(Fig. 3A; ANOVA, F3,138 = 11.2, P < 0.0001), while H’ for the vegetation overall sharply declined
immediately after regrading, then steadily increased by 2014 (Fig. 3b).
Changes in the distribution of plant cover among the 4 types of functional groups were substantial in
some cases (Table 5). The proportion of plant cover of non-native species first declined abruptly from
2011 to 2012 (Fig. 3C), due largely to the aforementioned decreases in relative cover of M. scorpioides
and Typha x glauca (Table 4), then remained low. For longevity, perennial species accounted for >97%
of the plant cover for all sampling years. The relative cover of graminoid species nearly doubled by 2014
and that of obligate wetland species increased from 89% prior to the regrade to >96% afterwards.
In contrast to our fourth hypothesis, the diversity of WIS categories actually declined.
Species 2011 2014
DD FL DD FLr
Alisma triviale Pursh (Northern Water Plantain) 5.5 57.4 2.2 16.7
Juncus spp. (Rushes) 42.3 0.1 57.0 34.9
Leersia oryzoides (L.) Sw. (Rice Cutgrass) 22.8 2.3 17.2 14.4
Lemna minor L. (Common Duckweed) 0.0 23.7 - -
Schoenoplectus tabernaemontani (C.C. Gmel.)
Palla (Softstem Bulrush)
- - 7.3 18.9
Veronica serpyllifolia L. (Thymeleaf Speedwell) 8.5 1.6 4.6 5.3
Table 2. Relative seedling density for species with values >5% in at least 1 treatment; drawdown (DD)
and flooded (FL).
Table 1. Mean density (seedlings/m2) and species richness for seedlings emergent from the soil seed bank before
(2011; 15 sampling points) and after (2014; 6 points) regrading. Means ± SE. Probability (P) is for t-tests.
Density Species
2011 9880 ± 1509 17.5 ± 1.7
2014 3367 ± 1225 13.3 ± 2.1
P 0.014 0.259
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Our findings via Mahalanobis distance values (Table 6) were based on presence–absence data,
and, thus, differ somewhat from the percent cover data. Of the 4 types of functional groups, only
WIS showed a significant change following the regrade, concomitant with the increase in obligate
wetland species.
2011 2014 Chi-square (df, P)
Native status Native 85.9 (4206) 92.2 (654) 21.8 (1, ***)
Non-native 14.1 (692) 7.8 (55)
Longevity Annual 0.7 (22) 4.7 (14) 40.5 (1, ***)
Biennial <0.1 (1) 0 (0)
Perennial 99.2 (3002) 95.3 (281)
WIS OBL 80.3 (2456) 82.3 (241) 7.5 (3, n.s.)
FACW 5.0 (153) 2.4 (7)
FAC 13.1 (981) 15.0 (44)
FACU 1.0 (32) 0.3 (1)
U 0.6 (17) 0 (0)
Growth habit Forb 41.1 (2046) 14.0 (101) 195.3 (1, ***)
Graminoid 58.9 (2937) 86.0 (618)
Vine <0.1 (2) -
Table 3. Percentages of seedlings in different functional group categories for native status, longevity, wetland
indicator status (WIS), and growth habit that emerged from the seed bank before (2011) and after (2014)
regrading. Seedling counts shown in parentheses. Chi-square values shown with degrees of freedom are all
significant at the 0.001 level (***), except for the non-significant (n.s.) value for WIS. OBL = obligate wetland;
FACW = facultative wetland; FAC = Facultative; FACU = facultative upland; U = upland
Species 2011 2012 2013 2014
Alisma triviale Pursh. (Northern Water Plantain) 3.7 1.3 2.4 7.1
Eleocharis palustris (L.) Roem. and Schult. (Common
Spikerush)
- 6.1 5.6 5.0
Leersia oryzoides (L.) Sw. (Rice Cutgrass) 13.6 2.0 9.7 10.9
Myosotis scorpioides L. (Water Forget-me-not) 18.0 0.7 - -
Potamogeton sp. (Pondweed) 10.0 55.8 23.2 19.0
Ranunculus sp. (Buttercup) 8.2 <0.1 - -
Sagittaria latifolia Willd. (Broadleaf Arrowhead) 7.9 28.3 44.1 35.3
Schoenoplectus tabernaemontani (C.C. Gmel.) Palla
(Softstem Bulrush)
- 2.1 7.8 12.4
Typha x glauca Godr. (Hybrid Cattail) 17.7 0.4 0.8 1.5
Table 4. Relative percent cover for species in the standing vegetation with cover >5% before (2011) and/
or after (2012–2014) the regrade. Non-native species are in bold.
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Figure 3. Values of (A) species richness,
(B) Shannon-Weiner diversity (H’),
and (C) relative % cover of non-native
species for the standing vegetation
before and after regrading. Data
for 2012–2014 are each the mean
of 2 samples. Means not sharing a
common letter differ significantly at P
= 0.05, according to Tukey’s honestly
significant differences tests.
A.
B.
C.
a
b
c
B
AB
A
A
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Despite few changes in general vegetation characteristics, in addition to the cover changes
cited above, we observed several shifts in the species composition of the standing vegetation from
2012 through 2014 (Table 4).
Discussion
Habitat Alteration Reflected in Vegetation
The regrading of Lieberman was a major habitat alteration that substantially impacted the
seed bank and the standing vegetation. As predicted in our first hypothesis, we observed a reduced
density of seedlings emerging from soil cores, indicating a dilution of the soil seed bank. The total
seedling density before the regrade is within the range of other urban wetlands in the Binghamton
area, but the seedling density in 2014 was significantly lower than other urban wetland seedling
communities (Larson and Titus 2018). The overall seed bank dilution is likely a consequence of
the dredging and leveling work needed to expand the wetland area. Other studies have shown
that seed banks can initially be negatively affected by major habitat alterations but then recover;
for example, an extreme flooding event increased seedling density but reduced species richness,
yet the riparian seed bank itself recovered quickly and was considered resilient (Osunkoya et al.
2014). Neff et al. (2009) reported that the density of seedlings emerging from sediment samples
collected from a recently restored tidal marsh significantly increased by more than 40-fold within
a year, and species richness was significantly higher than any reference site. The seedling density
and species richness of these studies are likely a result of prolific seed production, allowing the
seed banks to recover quickly. The lower seedling density from Lieberman sediment samples may
be a consequence of low seed dispersal into the wetland, possibly due to its isolated position in an
urban, fragmented landscape.
Although Juncus spp. and L. oryzoides remained important components of the seed bank
before and after the regrade, we observed an overall shift in species composition, including a
Type Category 2011 2012 2013 2014
Native status Native 72.2 95.2 97.0 94.4
Non-native 27.8 4.8 3.0 5.6
Longevity Annual 1.6 0.8 2.4 2.0
Biennial 0.5 0.5 - 0.2
Perennial 97.9 98.7 97.6 97.8
Growth habit Forb 77.1 73.0 64.9 60.1
Graminoid 21.1 27.0 35.1 39.9
Tree/shrub 1.8 - - -
WIS OBL 88.5 98.8 97.9 96.7
FACW 4.6 0.9 0.6 1.0
FAC 5.0 0.2 1.4 2.1
FACU 1.9 0.0 <0.1 0.2
WIS-H' 1.29 0.90 0.93 0.99
Table 5. Relative % cover of functional types by category for native status, longevity, growth habit, and
wetland indicator status (WIS) of standing vegetation before (2011) and after (2014) the regrade. WISH'
is diversity based on presence–absence data for WIS categories (as in Table 3).
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depletion of some species that were common in the seed bank before the regrade. Other species
emerged that were not observed in 2011. For example, S. tabernaemontani was present after the
regrade but not before; this species was unexpectedly common in the new standing vegetation.
The regrading project initially decimated the standing vegetation, but plant cover steadily
increased. Like the seed bank, we observed a shift in species composition after regrading. Counter
to our second hypothesis, the non-native M. scorpioides and Typha x glauca, 2 of the common
species in the 2011 standing vegetation, did not rapidly establish compared to native species, like S.
latifolia, L. oryzoides, and S. tabernaemontani. Although both seed banks contained few non-native
species, we were surprised that non-native species cover in the new standing vegetation was lower,
as many invasive species rapidly colonize disturbed sites (Bansal et al. 2019, Bowman-Cutway
and Ehrenfeld 2010, D’Antonio and Meyerson 2002, Ehrenfeld 2008, Meyer et al. 2013); however,
not all urban habitats have a high presence of exotics (e.g., Ehrenfeld 2005).
Although initially low, the presence of non-native species may increase, resulting in the need for
management of these species. We observed only 6 seedlings of the non-native Lythrum salicaria L.
(Purple Loosestrife) before the regrade and 2 after, perhaps signaling the beginning of L. salicaria’s
invasion into the site. Continued dispersal of L. salicaria from outside the wetland and seed rain
from established plants may increase the presence of L. salicaria. Although Typha had low seedling
densities in both seed bank surveys, we predict that Typha will continue to spread vegetatively in
the newly altered site, as Typha readily colonizes and forms monodominant stands in disturbed
habitats (Bansal et al. 2019). Future invasive species management may need to include L. salicaria
and Typha x glauca removal; for example, Ho and Richardson (2013) recommend the removal of
invasive species for 5 to 7 years to ensure native plant establishment and limit invasive species
dominance. Continued monitoring of the standing vegetation will provide important information
regarding invasive species management in urban wetlands.
After the regrade, we were surprised that S. latifolia and Potamogeton sp. were the 2 most
common species in the standing vegetation because neither taxon was common in the seed bank
surveys or in the 2011 standing vegetation (Table 4). We were surprised that S. latifolia and
Potamogeton sp. were the 2 most common species in the standing vegetation after regrading,
because neither taxon was common in the seed bank surveys or in the 2011 vegetation. The
establishment of these 2 species may have resulted from hydrologic change coupled with asexual
propagation, namely through the production of S. latifolia corms or tubers (Dorken and Barrett
2003, 2004; Van Drunen and Dorken 2012) and Potamogeton sp. rhizomes (Gleason and Cronquist
1991, Wiegleb and Brux 1991), respectively. Many Potamogeton species spread vegetatively from
turions and the fragmentation of stolons and rhizomes (Combroux and Bornette 2004, Kaplan et
al. 2014, Vári 2013). Asexual propagules of S. latifolia may be important for restoring vegetation,
perhaps because they increase the likelihood of survival in disturbed habitats (Dorken and Barrett
2003) and dispersal rates within sites (Dorken and Barrett 2004). Similarly, asexual reproduction
strategies of Potamogeton may be more successful in habitats with frequent disturbances (Wiegleb
Type # categories D P
Native status 2 1.90 >0.05
Longevity 3 2.19 >0.05
Wetland Indicator Status 4 17.42 0.0016
Growth habit 3 4.75 >0.05
Table 6. Mahalanobis D values based on presence–absence data for functional group types in the standing
vegetation before vs. after the regrade.
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and Brux 1991), and other studies have shown that Potamogeton species can readily germinate
under flooded conditions (Wang et al. 2016). Meyer et al. (2013) also observed that Potamogeton
can rapidly colonize newly restored side-channels along the Rhine River. The recovery of standing
vegetation in a riverine wetland after restoration was attributed to an increased recruitment from
rhizomes and other vegetative fragments, suggesting that bud banks can be important for wetland
recovery from major habitat alterations (Combroux and Bornette 2004, Combroux et al. 2002).
In contrast to habitats with established plant communities, heavily disturbed urban wetlands
have early successional, non-equilibrium communities that may exhibit substantial changes to the
standing vegetation composition. A Gleasonian approach would predict that we would see changes
in plant functional groups, with annual plant species rapidly establishing after a disturbance,
followed by the colonization of clonal perennial species (Odland and del Moral 2002, van der
Valk 1981). Although this approach suggests that clonal perennial species would eventually
establish, the standing vegetation at Lieberman was dominated by perennial species in less than
3 growing seasons; many of these species were clonal graminoids (e.g., Eleocharis palustris [L.]
Roem. and Schult. [Common Spikerush], L. oryzoides, and S. tabernaemontani). We also observed
an increase in graminoids emerging from the seed bank, suggesting that something is favoring
the establishment of graminoids in our system. We originally predicted that we would observe a
compositional shift toward an increase in annual species as well as a greater diversity among WIS
categories. Although we saw a slight increase in the relative percent cover of annuals, the vegetation
after the regrade remained largely composed of perennial species; thus, our third hypothesis was
incorrect. We also observed a greater percentage of OBL species (Table 5), rather than an increase
in the diversity of WIS. This trend was not observed in the seed bank, where the percentages of
OBL species remained constant. Changes in hydrology may have resulted in an environment that
favored obligate wetland species; the wetland is conspicuously more inundated since the regrade
(M Larson, pers. observ.), and much of the originally planned topography was not successfully
constructed. Landscape architects and engineers need to pay particular attention to rehabilitating
urban wetland hydrologies to favor wetland plant establishment of targeted functional groups
(Schwab and Kiehl 2017, Wang et al. 2016).
Management Implications
The complete regrading and expansion of this urban wetland effected a sharp decline in
the number of seeds in the seed bank. This dilution may have contributed to shifts in both the
standing vegetation and the seed bank communities. Although most of the species that were lost
after the regrading project were originally observed in small numbers, even common species
were drastically reduced in their seedling density. Despite having substantially lower seedling
densities in our seed bank study, the standing vegetation recovered after 3 years, indicating a
resilient urban wetland ecosystem.
While the current case study focuses on a single, relatively small wetland, we are not aware
of any other sites in our region subjected to the same degree of disturbance. Yet we expect that
created and constructed urban wetlands will become more common. The complete regrade and
expansion of our urban wetland provided a unique opportunity to document changes in species
composition of the seedlings that emerged from the seed bank as well as successional changes in
the standing vegetation, and it is encouraging that the urban wetland recovered within 3 growing
seasons without an increase in non-native species cover. It appears that seeding and planting may
not always be necessary to promote the re-establishment of vegetation after an existing wetland is
reconstructed. Allowing vegetation to reestablish from existing seed and bud banks may increase
the likelihood of rapid recolonization of a diverse aquatic plant community (Alderton et al. 2017).
While vegetation cover rapidly establishes in some systems (Meyer et al. 2013, Mitsch et al. 2005),
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other studies estimate that the time required for an ecosystem to recover after wetland restoration
or creation may be several decades, or even centuries (Curran et al. 2014; Johansen et al. 2017;
Jones and Schmitz 2009; Moreno-Mateos et al. 2012, 2015; Stefanik and Mitsch 2012). Moreover,
urban environmental conditions require modified restoration and rehabilitation designs to ensure
restoration project success (Ravit et al. 2017). Future case studies like ours will increase our
understanding of vegetation recovery of wetlands in urban landscapes.
Acknowledgements
The authors thank Craig Reynolds, Evan Schulz, Anthony Paolini, and James Liebner for their assistance
in the Research Greenhouse and field. We also thank Dr. Doug Wilcox for contributing to the design of the
germination trays, Dr. Mark Blumler for assistance in plant identification, and Dr. Weixing Zhu for advice on
statistical analyses. Anonymous reviewers provided constructive comments on earlier drafts of the manuscript. This
study was partially supported by the Center for Integrated Watershed Studies (CIWS) of Binghamton University,
State University of New York; a Wallace Research Foundation grant to Binghamton University; and a United
States Environmental Protection Agency Region 2 Wetland Development Grant (EPA-R2-09WPDG) awarded
to the Upper Susquehanna Coalition (USC), administrated by the Tioga County SWCD, and subcontracted to
Binghamton University. Although the information in this document has been funded in part by the United States
Environmental Protection Agency under assistance agreement CD-972253-09-0 to the Tioga County Soil and
Water Conservation District, it has not gone through the Agency’s publications review process and, therefore,
may not necessarily reflect the views of the Agency and no official endorsement should be inferred.
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