Comparative Strategies for Accelerated Wetland Restoration on Agricultural Land

Final Report for GNC03-015

Project Type: Graduate Student
Funds awarded in 2003: $10,000.00
Projected End Date: 12/31/2006
Region: North Central
State: Ohio
Graduate Student:
Faculty Advisor:
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Project Information


Early comparisons of wetland restoration strategies showed supplemental planting in historically wet agricultural areas to be a more effective means of restoring native wetland plant diversity than other restoration strategies. Measures of plant species richness and diversity compared favorably to other, similarly aged restorations in other parts of Ohio. The persistence of wetland species in the remnant seedbank helped facilitate quicker establishment of vegetative cover in both planted and unplanted wetlands. However, unplanted wetlands had far fewer plant species than planted wetlands. Early hand-pulling successfully suppressed Typha spp. establishment, but was inadequate for controlling Phalaris arundinacea, even when coupled with herbicide applications.


Discharge from agricultural settings is the most pervasive non-point source of pollution affecting surface water quality in the United States (USEPA 1984). To protect surface water quality and to respond to more stringent non-point source standards, cost-effective methods to treat agricultural non-point source discharges are urgently needed (Petersen 1998).

As a result, the restoration of wetlands on former agricultural land has received increasing public support through USDA programs including the Conservation Reserve Enhancement Program (CREP) and the Wetland Reserve Program (WRP). To date, restoration of wetlands on active farms or recently fallow agricultural land is becoming an effective strategy for improving both habitat quality on farms and water quality off the farms. However, processes of natural succession into functional and sustainable wetlands are complicated by the presence of a variety of exotic and invasive species, distances from native seed sources, and the impacts of site history. This has been demonstrated in numerous studies of past projects where created and restored wetlands were found to be devoid of individual, and in some cases entire functional groupings of native plant species many years after restoration (Galatowitsch and van der Valk 1996a and 1996b, Moore et al. 1999, Mulhouse and Galatowitsch 2003, Seabloom and van der Valk 2003).

More projects are needed to improve our understanding of how to accelerate the process of wetland habitat restoration within agricultural settings and the degree to which early management of these restorations is needed in order to produce sustainable levels of biodiversity and water quality improvement. The results of these studies would be of interest to farmers, restoration practitioners, public managers, and resource management professionals across Ohio and the Midwest.

The Black River Watershed, including the area in which the project took place, was listed in 1991 as one of 43 toxic hotspots in the Great Lakes basin. While many industrial discharges in the northern part of the watershed have been addressed, agriculture in the upstream southern areas is the number one contributor to non-point source pollution. In addition, 90% of the original wetlands in the watershed have disappeared, in large part due to agriculture. It was (and is) the goal of this project to investigate what early restoration strategies will provide the best short- (and long-term) wetland plant diversity and water quality improvements, and to serve as an interactive example of the benefits of restoration for local farmers, students, resource managers, and restoration practitioners.

Project Objectives:


"The overall goal for the project was to install and develop an experimental design for six ½ acre, hydrologically isolated wetland cells that would facilitate short- and long-term studies of wetland restoration techniques on active or fallow agricultural land. It was, and will continue to be our goal that this study, and future studies, will advance local and general knowledge about restoring wetland ecosystems on active farms or recently cropped areas.

Short term outcomes targeting faculty at The Ohio State University and Oberlin College, graduate student Joshua Smith and undergraduate students include: a) development of experimental designs to determine optimal seeding combinations and the impacts of nutrient runoff; b) determination of indicators for ecosystem processes and results of treatments; and c) establishment of monitoring system that will allow for real-time collection of data and display of data on a publicly accessible website. For this study in particular, Joshua Smith compared the effects of seeding, planting, and invasive species management strategies on the initial establishment of wetland plant communities in a previously tiled agricultural field that had been fallow for two years. A future study will examine the effects of simulated nutrient and sediment runoff on the established plant community and the effects of each initial management strategy on water quality.

Intermediate outcomes are to target public school students, farmers, and city officials with a focus on development of an educational outreach program targeting public schools, public officials, farmers, and landowners. The long-term outcomes of the project will be to improve restoration ecology practices applied to wetlands on agricultural fields."


To be completed by the end of 2003:

1)Establish six, half-acre, hydrologically independent wetland cells for the research and continued monitoring of wetland restoration strategies on agricultural land (completed).

2)Complete baseline soil, water and seedbank sampling of the site (completed).

3)Decide and implement an agreed upon experimental design regarding planting, seeding, and invasive species management regimes for each wetland cell (completed).

4)Complete the purchasing and installation of sampling and monitoring equipment (completed).

To be completed by the end of 2004:

1)Create a centralized location for storage of sample vouchers and complied data (completed).

2)Complete 1st year plant community survey in each wetland cell (completed).

3)Reassess experimental design and provide adaptive management to address any unforeseen biological or environmental issues (completed).

To be completed by the end of 2005:

1)Complete 2nd year plant community survey of each wetland cell (completed).

2)Finish the creation of a website to display photographs, data, and project findings (in progress).

3)Successful defense of Master’s Thesis by Joshua Smith (not completed until December 2006)


Click linked name(s) to expand/collapse or show everyone's info
  • Jay Martin
  • Martin Quigley


Materials and methods:


The study site was located at 44270 Oberlin Elyria Road at the George Jones Memorial Farm outside of Oberlin, Ohio. The site is home to the Ecological Design and Innovation Center (E.D.I.C.), a non-profit organization that promotes sustainable land use planning and development through a number of on-site practices and local programs. Situated one mile east of town at 41° 17’ 38” N latitude and 82° 13’ 03” W longitude, the site is home to six experimental freshwater marshes designed to be identical in size, shape (perimeter) and basin topography. Marshes at the study site were named and referred to as “wetlands” and will be referred to as such to provide consistency with the previously established terminology. Individual wetlands were constructed by excavating shallow depressions and creating earthen levee walls in a reverted old-field on the south end of the property.

The site was historically tiled and cropped in a standard corn and soybean rotation until its donation to Oberlin College in 2000. It is located at an elevation of 236 meters above sea level in the Plum Creek Basin of the Black River Watershed—a northward flowing tributary to Lake Erie. Each wetland has a rectangular shape; 68.6 meters long by 35.1 meters wide (approximately .24 hectares each). The wetlands are set side-by-side in a single row from west (Wetland one or W1) to east (Wetland six or W6). The wetlands were bordered on three sides (northern, eastern and western) by eight meter wide earthen levee walls and consist of deepwater basins in the northern ends, with gently sloped (1:100) mudflat basins tapering towards unbounded southern ends.


Rainfall and sheet flow from the adjacent old-field to the south comprised the only hydrologic inputs to the wetlands. Control boxes with adjustable weirs were built into the northeastern corners of each wetland for the purpose of adjusting water levels. Based on weekly surface water measurements, adjustments were made to the water levels of three wetlands in the early spring of 2005 to increase the similarity of deepwater basin depths and the southern extent to which each mudflat gradient was inundated.


Surveying and soil identification efforts prior to construction showed the site to be a member of the Mahoning soil series. Two soil classifications in particular, Mahoning silt loam (MgA) and Mahoning-Tiro silt loams (MkA), dominated the site. These soils are both typified by shallow slopes (0-2 percent) and severe wetness because of perched water tables near the surface (Masi, 2000, unpublished data). They are on the wave-cut lake plain and the soil is predominantly gray clay and is considered to be moderately suited for wetlands and wetland plants.


An analysis of the local seedbank was performed between March and May of 2004. Soil samples for the seedbank survey were collected on March 29, 2004 along three transects: one through an undisturbed old-field area near the wetlands and two through the disturbed, graded area representative of the soils in the wetland basins (one along both the north and south edges of the wetlands). Transects were located immediately north and south of the wetlands for two reasons: 1) To avoid confounding of the results by disturbing the seeds already sown in wetlands receiving the planting treatment; and 2) To ensure the inclusion of disturbed soils representative of the newly graded earth that comprised the wetland basins. A third group of seedbank samples were taken along a transect in a nearby undisturbed part of the remaining old-field north of the wetlands. This was done in order to facilitate a comparison with the disturbed soil of the wetlands.

Soil samples were taken with a standard hand-corer. Cored samples were five centimeters in diameter and eight centimeters deep and were taken from two random points every 50 meters along each transect. Soil cores were then homogenized by transect and mixed with sterilized potting soil (30:60 mix) in a 1:2 ratio by volume (sample to potting soil). In accordance with van der Valk (1981), soil mixes from each transect were then divided into two subsamples and placed in a tray for germination. In a similar fashion to the seed bank studies done by Willis and Mitsch (1995), one subsample tray was then submersed under one centimeter of water to approximate flooded conditions while the other subsample tray was watered daily in an attempt to approximate a wetland mudflat during drawdown (van der Valk, 1981). Seeds in each tray were allowed to germinate and grow to an identifiable size before they were removed from the trays.


Rebar was used in each wetland to delineate a permanent grid across the entire basin of each wetland. This was done in order to provide a means for fixed, non-random sampling within each basin, and in larger part, to help facilitate future, repeatable comparisons of areas within and between the wetlands.

When describing the sampling grid, the wetlands are best thought of as existing on an x, y-axis with the y-axis along the longer, north-south alignment and the x-axis along the shorter east-west alignment. As such, seven evenly spaced transects were placed parallel to the x-axis. This was done in order to cover the entire elevation gradient of each wetland from the shallow south end to the deeper north end. Seven evenly spaced transects were also marked out parallel to the y-axis of each wetland basin. Because their rectangular shape, each wetland was nearly twice as long as it was wide. As such, horizontal transects were spaced 10 meters apart while vertical transects were spaced five meters apart. The ends and intersections of each transect were then marked with ½ inch (diameter) by 48 inch (length) rebar stakes to form a permanent 60m x 30m grid in each wetland basin.


Following construction in July of 2003, annual ryegrass was sown in and across the wetlands to delay colonization until the experiment could begin. A long-term experiment was then initiated on October 29, 2003, three months after the wetland basins were completed. Two wetlands were left unplanted and open to natural colonization while the other four were planted and seeded with twenty-two wetland species typical of restoration projects and common to Ohio Wetlands. Seed and plant propagules were collected from local populations and purchased from Spence Restoration nursery in Indiana. Five-hundred and seventy-two (572) plants and rhizomes representing 11 species were planted evenly across the four planted wetlands while 11 other species were introduced via broadcast seeding. For the planting effort, all wetlands were drawn down in order to provide better access to all submerged and saturated areas. All propagules were planted at and between rebar stakes along the edges of the wetlands and across the previously flooded mudflat gradient per recommendations by Dr. David Benzing of Oberlin College. Seeds of individual species were divided evenly among the wetlands and homogenized with a sand mixture to prevent settling by weight. This mixture was then broadcast by hand-cranked broadcasting tools throughout each wetland in the planting treatment.


During both years of the study, three common invasive wetland plants were periodically removed from all wetlands. The plants were Phalaris arundincacea, Phragmites australis, and any Typha species; all of which were present in dense, nearly monotypic stands, at various locations on and around the property. No distinction was made between the native species Typha latifolia, non-native Typha angustifolia and the more invasive hybrid Typha x glauca. This was done for two reasons. First, Typha latifolia and Typha angustifolia by themselves can spread rapidly, and when present together may also result in the production of the more invasive hybrid Typha x glauca. Secondly, Oberlin researchers felt that it may be unrealistic to expect land-owners to be able to consistently distinguish between native Typha latifolia, non-native Typha angustifolia and the hybrid Typha x glauca species when trying to apply similar management techniques on their own properties in the future. As such, initial management for any and all Typha propagules was performed in all wetlands.

Typha biomass was removed by hand while Phalaris arundinacea was removed through cutting of the foliage at ground level, followed by an immediate application of concentrated Roundup © to remnant foliage on dry land, so as not to contaminate water. In 2004, this was done three times for both Typha and Phalaris arundinacea: late May, late June and late July. In 2005, there were only two sessions of management for Typha (late June and early August) because of a visual observation of a decrease in Typha. Phragmites australis, despite being observed in local roadside ditches, was not observed in any of the wetlands in either year of the study.

A measure of biomass (dry weight) was calculated for the total amount of each species removed from each wetland during each year. This parameter was calculated by drying a subsample of known wet mass at 100 degrees Celsius for 24 hours. The ratio of subsample dry weight to wet weight was then multiplied by the total wet weight removed from each wetland for the year. Comparisons of Phalaris arundinacea occurrence after 2004 were performed by comparing the frequency (number of quadrats where observed / total quadrats in each wetland) and median cover observed in the mudflats of each wetland. This change was necessary because of the large amounts of biomass encountered in 2004. Phalaris arundinacea frequency and cover surveys were performed on July 9, 2005 and again on June 15, 2006.


The developing macrophyte community in each wetland was surveyed in August of 2004 and 2005. Survey methodologies were based upon those used by the Ohio EPA as adapted from the North Carolina Vegetation Survey (Peet et al., 1998). This plot-based method employs a set of 10m x 10m quadrats in a 20m x 50m layout totaling .1ha in area. Using the existing rebar grid, sample quadrats were set up in identical locations in each wetland. In order to include more edge habitat and to better observe planting effects, sampling quadrats were delineated in a checkerboard pattern rather than the 20m x 50m orientation described by Peet et al., 1998. In total, sampling efforts within each basin amounted to 50% of the area of each wetland basin.

In both years of the study, vegetation surveys were performed by me and Oberlin College professors Dr. David Benzing and Dr. John Peterson. All species within each quadrat were identified to species for each sampling period (2004 and 2005), with unknown specimens being vouchered for later identification to the lowest possible taxonomical level with the aid of Gleason and Cronquist (1991) and The Illustrated Guide to Gleason and Cronquists’ Manual (Holm, 1998).

Once all species in each quadrat had been identified and/or vouchered, cover classes were assigned for each species in each sampling quadrat as recommended by Peet et al. (1998). This method uses defined cover classes for observed aerial cover, based upon a doubling factor. This method was chosen for its reported robustness in the face of individual observer error and its repeatability between different observers. Because of the large number of quadrats (54) and a need to conserve sampling time, this method was also chosen for its relative speed and simplicity compared to other sampling techniques. To help reduce the amount of subjective error, all three researchers produced independent estimates of individual species and total emergent cover during sampling. A final cover class was then agreed upon by all three researchers. In this study, vegetative cover was described as the proportion of sample area covered by either emergent macrophyte foliage or the foliage of floating-leaved species such as Potomogeton nodosum or Nymphaea odorata. Submerged species observed during field surveys were given cover class designations for purposes of species diversity and Floristic Quality Assessment Index calculations, but were not included in the estimated vegetative cover recorded for each sample quadrat. Species richness was calculated as the number of macrophyte species observed in a given area. In this study, total species, native species, and non-native species richness were recorded in each of nine sample quadrats in each wetland. From these samples, total, native and non-native richness was also found for each wetland as a whole.


Shannon-Weiner species diversity scores (Krebs 1999) were calculated for each sample quadrat using the following equation:

H´ = Summation(pi)(loge pi) Eqn. 1

where H´ = index of species diversity (information content of sample), and pi = proportion of the total sample of the species i. In practice, H´ varies from 0 (a community with only one species) to no more than 5 for most biological communities (Krebs 1999).

Floristic Quality Assessment Indices (FQAIs) are derivations of species diversity scores that add a component of relative importance to each species within an observed community. A common criticism of diversity indices is that they give equal weighting to all species, without regard to rarity, environmental tolerances or specificity of habitat requirements (Andreas et al., 2004). In practice, this means that an evenly distributed community of ten highly ubiquitous or invasive species would produce the same species diversity score as an evenly distributed community of ten endangered or low tolerance species. FQAI’s then, address this problem by incorporating an additional element of species weighting, known as a coefficient of conservatism (C of C). Each C of C is “an expression of the taxon’s autecology as it relates to its fidelity to narrow or broad habitat requirements with respect to all other taxa in the flora (Andreas et al., 2004).” For this study, the Floristic Quality Assessment Index (FQAI) currently in use by the Ohio EPA was used. Non-native species and plants not identified to species are precluded from the calculation. The equation is as follows:

I = Summation (CCi)/sq.root(Nnative) Eqn. 2

where I = the FQAI score, CCi = the coefficient of conservatism of plant species i, and Nnative = the total number of native species occurring in the community being evaluated. This equation is the original FQAI equation as formulated by Swink and Wilhelm (1979). In practice, FQAI scores can give watershed or wetland managers a basis from which to prioritize wetland conservation efforts, and a means by which to assess the success of wetland restorations (Andreas et al., 2004). In an attempt to provide a better relative assessment of wetland restoration success between our treatments, FQAI scores were calculated for each wetland. These scores were then compared between wetlands, treatments and sampling seasons.


Species Richness and Shannon-Wiener diversity scores (Krebs, 1999) were calculated for each sampling quadrat using PCORD4 Statistical Software version 4.20 (McCune, B. and M.J. Mefford, 1999). Native and non-native species richness also was calculated for each quadrat. Analysis of Variance (ANOVA) was used in Systat, version 11, (Systat Software Inc., 2004) to test for statistically significant differences in vegetative cover, native species richness, non-native species richness and Shannon-Wiener diversity scores between planted and unplanted treatments and between individual wetlands. Tukey’s post hoc multiple comparisons procedure was used to test for statistically significant differences between multiple sample means at the 95% confidence levels. In instances where data could not be transformed to meet the assumptions of parametric tests, p-values were derived from non-normal Kruskal–Wallace tests.

Research results and discussion:


Twenty-two different macrophyte species were observed in the seedbank study of the wetlands. Of these, only five had wetland designations (OBL or FACW), with the majority of species having upland designations (UP or FACU). Four of the observed seedbank species, Lolium perenne, Poa annua, Poa pratensis, Digitaria ischaemum—all upland non-natives—were not found in any of the wetlands in either year of the study. Conversely, not all wetland species observed in the landscape were observed in the seedbank study. These species were, Scirpus cyperinus, Juncus effuses, and Cyperus strigosus, all of which were observed in at least one or more of the wetlands in one or both years of the study.


Vegetative cover varied within each wetland largely because of differences in water depths. While some sample quadrats were located in largely deepwater areas with sparse vegetative cover, other quadrats were at higher elevations on wet-meadow/old-field areas of the mudflat gradient, and as a result had much greater vegetative covers. This was illustrated by plotting average depth of each quadrat against the observed vegetative covers in each quadrat in both 2004 and 2005. Large variations in water depth also occurred within certain individual sampling quadrats, particularly around the edge of the deepwater basins in each wetland. The effects this had on vegetative cover were shown by plotting the coefficient of variation of the average depth of each sample quadrat against vegetative covers. Mudflat quadrats with low coefficients of variation were revealed to have high vegetative covers while deepwater and deepwater edge quadrats with high coefficients of variation were revealed to have low vegetative covers.

In regard to wetlands as a whole, an Analysis of Variance showed no significant difference in vegetative cover between planted and unplanted wetland treatments (df =1, p = .065*) or between individula wetlands (df=5, p = .422*) in 2004. This was again true in 2005, as no statistically significant differences between treatments (df=1, p=.169*) or between individual wetlands (df=5, p=.080*) was observed. However, in 2004 unplanted wetland 4 (W4) had the highest mean vegetative cover (68% ± 41%) of all the wetlands in the study, indicating that planting had no effect on the amount of initial vegetative cover.

Water levels were adjusted between 2004 and 2005 according to results from an elevation survey performed on July 19, 2004, which revealed the existence of morphological differences both across mudflat gradients and in relation to the depth of the deepwater basins at the north end of each wetland. As such, results from the second sampling season reflected these changes in water levels, indicating that initial vegetative cover was affected more by variations in water levels than by planting efforts. In 2005, the two wetlands that received water level increases (planted W3 and unplanted W4) both declined in mean vegetative cover. W4 in particular, decreased considerably between 2004 and 2005 from 68% ± 41% to 31% ± 32%. Conversely, the greatest increase in mean cover between 2004 and 2005 was observed in the two wetlands (W5 and W6) that received no adjustment and displayed steady declines in water levels throughout the growing season. These results indicate that the high mean vegetative cover observed in W4 was a result of its shallower deepwater basin providing a more suitable habitat for the establishment of emergent macrophytes during the first year of the study, and that the subsequent drop in vegetative cover observed in W4 in 2005 was a result of the anthropogenic increase in water level at the beginning of that year.

These findings regarding water level effects on wetland vegetation establishment are consistent with past studies (van der Valk and Davis 1978, van der Valk 1981, van der Valk 1994). Cassanova and Brock (2000) performed a study on the effects of wetland water depth and flooding regimes on the emergence and establishment of various wetland species. They found a positive correlation between duration of inundation and the observed levels of species richness and biomass. The shortest durations of flooding produced the greatest biomass and species richness, while the longest durations of flooding producing the least. Mitsch and Willis (1995) also observed similar results in a study on the effects of submerged and moist soils on the emergence of wetland plants from both natural and artificial seedbanks. These findings illustrate the importance of the role of hydrology in wetland restorations, particularly where goals include the timely establishment of plant communities. In this regard, too much water may be just as problematic to sediment and nutrient sequestration as too little water for farmers and land-owners aiming to restore wetlands with productive vegetative cover rather than expanses of unvegetated open water.


Seeded species success in the EDIC wetlands was somewhat poor compared to results from other studies. In a series of depressional wetlands in southeastern Wisconsin, Reinartz and Warne (1993) observed 17 of 22 seeded species to have established in one or more of their constructed wetlands, while on six of eleven seeded species were observed to have established during this study. Wetlands in the Wisconsin study were also small and depressional, however, no seedbank was detected at any of the 17 wetlands in their study. This may have removed a component of competition from their study that existed for species seeded at the EDIC wetlands, as 22 species were detected in the seedbanks at the EDIC site. Further contributing to the seeding success of the Wisconsin study may have been the inclusion of a greater number of wetlands, both of varied sizes and locations in the landscape relative to this study. Increased site diversity most likely resulted in more combinations of environmental conditions suitable for germination and establishment by the larger diversity of seeded species. By comparison, there were only four planted wetlands in this study, all of which were replicates in regard to size and shape.

The edge habitat around the periphery of the deepwater basins differed considerably between planted and unplanted wetlands. In addition to the Juncus effuses, Eleocharis obtusa and Alisma subcordatum that dominated the unplanted wetland edge, planted wetland edge cover also included a number of visible patches of introduced species. In particular, Sagittaria latifolia, Sparganium eurycarpum, Nymphaea odorata and Pontederia cordata had all established colonies by the first sampling season, and were spreading considerably and vegetatively from points where they were planted the fall before. This difference in edge cover species composition between treatments was even more apparent in 2005 as Sagittaria latifolia, Sparganium eurycarpum, Nymphaea odorata and Pontederia cordata had continued to spread along the wetland edges, and into the deeper water of the planted wetlands.

In 2004, for wetlands as a whole, or wetlands as a whole, all four planted wetlands had greater total species richness than either of the unplanted wetlands in 2004. Native species richness was the same or higher in planted wetlands, while non-native species richness was the highest in the two unplanted wetlands. Comparisons of sample means between treatments showed planted wetlands to have significantly higher total species richness and native species richness than unplanted wetlands. However, despite both unplanted wetlands having higher non-native species richnesses than their planted counterparts, mean non-native richness was found to not be significantly different between treatments in 2004 (df = 1, p = .408).

From 2004 to 2005, observations of wetlands as a whole showed that native species richness either stayed the same (W3 and W6) or increased (W2 and W5) in planted wetlands, while decreasing in both unplanted wetlands. The number of non-native species and total species also increased in all four planted wetlands while decreasing in both unplanted wetlands. A comparison of means for 2005 showed all three richnesses to be significantly higher in planted wetlands than in unplanted wetlands. These results indicate that planting in restorations can significantly increase levels of native species richness, but that consideration needs to be given to the source of the propagules, as non-native species may also be inadvertently spread or introduced.

Wetland flora in other constructed wetlands can also be compared to that of the EDIC wetlands to assess the effects of management (species introductions), age, and location on plant community development. In this regard, the EDIC wetlands appear to be developing at rates similar to wetlands in other studies in the Midwestern United States. Despite being considerably smaller than all of the other wetlands, the EDIC wetlands compare very favorably in regard to both total species richness and wetland species richness. This supports the suggestion by Semlitsch and Bodie (1998) that in addition to reptiles and amphibians, small wetlands can also be very important to sustaining higher levels of plant biodiversity and that wetlands as small as 0.2 ha should be protected until more data can be collected regarding their importance to habitat connectivity and regional metapopulation dynamics.

Metapopulation interactions and source-sink dynamics deem that where multiple suitable habitats exist it is possible for individual populations of the same species to have influences on each other and on other suitable and unoccupied habitats. In the context of this study, it would appear that larger amounts of small wetland restorations could have a considerable effect on regional wetland plant diversity if these restorations are supplemented with propagules and seeds of native species that might otherwise be slow or unable to reach the sites on their own due to dispersal limitations and habitat fragmentation.

To date, this research indicates that early planting and seeding efforts at such restorations contributes to the initial plant diversity of even small restorations. Planting and seeding wetland restorations throughout agricultural landscapes would result in shorter distances between individual populations of more species, thereby helping to sustain smaller individual populations through a more frequent exchange of propagules with other nearby populations. Galatowitsch and van der valk (1996a), for example, found that unplanted three-year-old restorations previously drained by ditches were rapidly colonized by emergent perennials found to be persisting as small remnant populations in other areas of the ditches. In the same study, unplanted wetlands restored to tiled areas were still largely lacking any emergent component of vegetation. In another study, Seabloom and van der Valk (2003) found wetland restorations in previously farmed areas of the prairie pothole region to have lower species richnesses, greater compositional variability and a lack of flora distinctive of native wetlands. These studies and the EDIC study all show that farmers and land-owners interested in wetland restoration need to consider the landscape context of these restorations, and the potential need and benefits of supplemental planting and seeding efforts.


Shannon-Wiener Diversity scores also were calculated for each sampling quadrat in both years of the study. In 2004, mean species diversity scores for sample quadrats were significantly higher in planted wetlands than in unplanted wetland treatments (df = 1, p = .002*). Diversity scores were also significantly different between individual wetlands (df = 5, p = .001). Tukey’s HSD multiple comparison tests for significance at the 95% confidence level were used to identify between which wetlands these significant differences occurred. Results showed unplanted W4 to have significantly lower mean species diversity than all planted wetlands except W3. Despite both being planted, W3 also had significantly lower means species diversity than W6. This, however, was probably due to the aforementioned water level manipulations imposed on W3, and the decreases in vegetative cover that were also observed.

In 2005, mean species diversity again significantly differed between sample quadrats in planted and unplanted treatments (df = 1, p = .001*) and between individual wetlands (df = 5, p = .003). Tukey’s tests showed W4 to still have significantly lower mean species diversity than planted W5 (p = .003) and planted W6 (p = .008), but no longer less than planted W2. Also in 2005, the two unplanted wetlands displayed the lowest mean species diversity of all six wetlands.


Floristic Quality Assessment Index (F.Q.A.I.) scores were also calculated for each wetland as another means of comparing individual wetland plant communities as a whole (Andreas et al., 2004). Results showed planted wetlands to have significantly higher FQAI scores (2004: df=1, p =.001* and 2005: df=1, p =.001*) than unplanted wetlands, with mean planted wetland scores (2004: 16.3 ± 1.5 and 2005: 16.3 ± 0.6) being almost twice as high as mean unplanted wetland scores (2004: 7.3 ± 1.4 and 2005: 8.4 ± 0.2) in both years. In planted wetlands, FQAI scores decreased slightly in all but one wetland (W3) as three seeded species were observed in 2005 that were not observed in 2004. In unplanted wetlands, W1 had nearly the same FQAI score in 2005 (8.3) as it did in 2004 (8.2), while W4 increased from 6.3 to 8.5 because of the observation of two introduced species that had migrated from adjacent planted wetlands most likely via muskrats or birds.

Both unplanted EDIC wetlands had lower FQAI scores than unplanted wetlands in other studies. While the EDIC wetland lacked connectivity with any upstream hydrological inputs, the unplanted wetlands at the Olentangy River Wetland Research Park (ORWRP) in Columbus, Ohio (Mitsch et al. 2005) and the unplanted Indian Lake Wetlands in central Ohio (Fink 2001) were both relatively close in proximity or directly connected to larger order streams. As such, unplanted wetlands in these studies were exposed to more frequent introductions of greater amounts of wetland plant propagules, either indirectly or as the result of episodic flood events. It should also be noted that the EDIC wetlands were five times smaller than the unplanted ORWRP wetland and six times smaller than the Indian Lake Wetland (Fink, 2001). These results further illustrate the need for supplemental planting and seeding efforts in areas with limited hydrologic connectivity and in fragmented and isolated habitats typical of agricultural landscape settings.

The potential benefits of planting and seeding efforts can also be seen by comparing the planted EDIC wetlands to planted wetlands from other studies. Despite observances of positive correlations between species richness and wetland size (Reinartz and Warne 1993, Galatowitsch and van der Valk 1996a), the planted EDIC wetlands had species richnesses similar to those of sites many times larger, indicating that planting at the isolated EDIC wetlands has had a considerable effect on the initial plant community compositions of the planted wetlands. The relatively lower FQAI scores observed in other, similarly aged, planted wetlands (Ross Labs and ORW1), further support the notion that planting and seeding considerably help to increase the species richness and diversity of restored wetlands, particularly where former or current agricultural practices have resulted in habitat fragmentation and seedbank degradation that might otherwise yield species poor wetland restorations and/or wetlands dominated by invasive or exotic species.


Observations of invasive species showed Typha propagules to be small and sporadically located in mudflat gradients and along wetland edges. Phalaris arundinacea was considerably more prevalent in the shallow upper-mudflat (southern end) areas of the wetland basins. Phragmites australis, though observed in nearby roadside drainage ditches, was not observed in any of the wetlands in either year of the study. Management efforts for Typha in all cells were successful in preventing the establishment of vegetatively expanding colonies.

The genus Typha is of particular interest to restoration managers because of the species’ prominence in the Midwest and ability to quickly establish and aggressively spread throughout a range of water depths. Typha populations are naturally suppressed by muskrat populations where localized pockets of deeper water can provide habitat for escaping predators. Where muskrat colonies persist, cycles of Typha expansion and defoliation provide an important amount of habitat heterogeneity and open space for the recruitment and re-establishment of many other plant species from the existing seed bank and adjacent landscape during water level drawdowns. In the absence of muskrats, Typha expansion can continue unchecked, potentially resulting in dominant stands of Typha species. The deep water areas were provided, the earthen dike walls separating individual wetland basins warranted muskrat trapping in order to maintain dike wall structural integrity. As a result, this study was able to look at the effects of intense initial management of Typha species by site managers, without a confounding component of potential muskrat herbivory. In this study, the amount of harvested Typha biomass was considerably less in 2005 than in 2004. These results appear to indicate that early Typha management can be a successful means of slowing or even preventing an initial domination of restorations by aggressively spreading Typha species. Furthermore, the “head-start”provided to the remaining planted or recruited species could have the potential to also reduce the rate of future Typha establishment as the other recruited or planted species continue to establish and spread across remaining habitable sites, all the while producing a more diverse wetland seedbank.

In 2005, no statistically significant differences in mean Phalaris arundinacea cover existed between planted and unplanted wetlands (df=1, p =.748). In 2006, however, planted wetlands had a statistically larger amount of Phalaris arundinacea cover than did unplanted wetlands (df=1, p=.001) due largely to the encroachment of Phalaris arundinacea from nearby established colonies. The significant increase in Phalaris arundinacea cover in W1, W5 and W6 was most likely attributable to the proximity of these wetlands to the existing stands of Phalaris arundinacea in the adjacent landscape. While W2, W3, W4, and W5 were insulated on two sides by other experimental wetlands, W1 and W6 were exposed to landscape communities on three sides.

Despite a twice-per-summer regiment of intensive weed-whacking followed by applications of concentrated roundup © to exposed plant tissue, the mean amount of cover of Phalaris arundinacea increased considerably in all four planted wetlands between 2005 and 2006. Kruskal Wallace non-parametric tests for significance showed statistically larger median cover values for Phalaris arundinacea in W1, W5 and W6 in 2006 than in 2005. Two wetlands in particular, W5 and W6, displayed enormous increases in both the number of sample quadrats where Phalaris arundinacea was observed (frequency) and the mean percent cover in which Phalaris arundinacea was observed. Even in wetlands where the number of sample quadrats containing Phalaris arundinacea decreased, the mean cover increased, indicating vegetative expansion and a failure to control for the species in the wetlands.

Although our management regime of cutting followed by herbicide applications proved not to be an effective means of diminishing the amount of Phalaris arundinacea at the EDIC site, management for invasive species at individual wetland sites is still important because of the effects that their high reproductive rates and quick vegetative expansion can have on other local species. By allowing invasive species to establish and persist, wetland managers may be facilitating the spread of invasive species to surrounding areas where they can potentially inhibit the establishment of other native wetland plant species (Budelsky and Galatowitsch 1999, Calloway et al. 2003, Mulhouse and Galatowitsch 2003). Because the effected native species may play an important role in the normal functioning of the local wetland ecosystem, it is important that wetland management practices include some form of invasive species control.

Results from past studies indicate that planting, even in the absence of management, can suppress the domination of wetlands by Typha species. At the ORWRP in Columbus, Ohio, numerous studies have documented the quick spread and domination of Typha in an unplanted wetland, while a second planted wetland has persisted for over 10 years with less Typha cover and a more diverse plant community (Mitsch 2005). In Southeastern Wisconsin, a study of created depressional wetlands also showed smaller average amounts of Typha in seeded (12.9% ± 7.36%) versus unseeded (17.2% ± 4.36%) wetlands after only two years (Reinartz and Warne, 1993). By comparison, Typha species comprised <1% of cover in all six EDIC wetlands during both years of the study, indicating that Typha species expansion was successfully being suppressed by hand-pulling management efforts comparison that did not exist in other studies (Reinartz and Warne 1993, Galatowitsch and van der Valk 1996a and 1996b).

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Participation Summary

Educational & Outreach Activities

Participation Summary:

Education/outreach description:

To date, this project has yielded a non-peer reviewed publication for Ohio State University graduate student Joshua Smith in the Proceedings from The Ohio Invasive Plant Research Conference (2005). The publication is a five page summary of the project design and future plans for the site and is included in an Appendix at the end of this report. This project also resulted in a successfully defended Master’s Thesis for Joshua Smith (2006), partially fulfilling his requirements for the attainment of the degree of Master of Science in Environmental Science. Joshua graduated from The Ohio State University in December of 2006 after successfully completing the remaining requirements of the degree. A copy of his Master’s Thesis is also included in an Appendix at the end of this report. A future outcome of this project could also include future publications in a number of peer-reviewed research journals.

A number of undergraduate students from Oberlin college have assisted at the EDIC wetland site in a number of different capacities, including assistance with planting and seeding the wetland, managing for invasive species, trapping muskrats, installation of sampling grids, annual plant surveys, and daily recordings of individual wetland water levels and climatological data. Environmental studies classes led by Dr. John Petersen also take numerous trips to the EDIC wetlands each semester as an opportunity for a first-hand view of the various types of colonizing wetland flora and fauna, and to observe the logistics and potential benefits of wetland restoration. Visiting farmers and local residents interested in purchasing organic produce produced at the George Jones Memorial Farm through other EDIC projects have also been given tours and been allowed to observed the wetlands at their leisure. A number of local residents have also taken to making the EDIC wetlands a featured stop on evening walks or afternoon drives in the summer months, as a foot path across the George Jones Memorial farm connects the various projects of EDIC to the outskirts of the town of Oberlin, Ohio.

Project Outcomes

Project outcomes:

Generally speaking, the EDIC wetlands have provided a place where the local public and interested land-owners can see the first-hand benefits and application of programs like the Conservation Reserve Enhancement Program (CREP) and the Wetland Reserve Program (WRP). Oberlin students can now view and participate in the monitoring and restoration process of a local wetland restoration. More specifically, Joshua Smith’s graduate research project yielded the following results:

1) Where remnant wetland seedbanks exist, planting efforts may not significantly increase the initial amount of vegetative cover in constructed wetlands. Marginally productive agricultural areas with a history of wet hydrologic conditions can have varying degrees of remnant seedbanks primed for wetland plant emergence. As such, our findings support the general preference for wetland construction and restoration in these areas. Farmers or land-owners in rural areas looking to restore wetlands on their property should focus on these marginally productive and already wet, depressional areas, as it is likely that remnant wetland species may still be persisting in the seedbank or in nearby refugal corridors like drainage ditches. By already having a pre-existing hydrology suitable for wetland development, restoration efforts can be far less expensive and labor intensive, and may be able to develop more quickly into functioning wetland ecosystems than in areas where wetlands may have never before existed.

2) Significant increases in native species richness and diversity can be initially achieved in small wetlands through planting and seeding. Site isolation and dispersal limitations may require that plant introductions take place in order for some native species to reach constructed wetlands in a timely manner. This is especially important in smaller, more isolated wetlands, for the effects that it may have on long-term metapopulation dynamics and plant community succession in fragmented landscapes typical of agricultural areas. Though small wetlands are often not protected by current legislation, they too can be important to maintaining regional biodiversity for the roles they play in source-sink dynamics and habitat connectivity. Wetlands as small as 0.2 acres can quickly develop vegetative cover and a diverse community of native wetland species, especially where remnant seedbanks exist and planting efforts take place.

3) Better invasive species management efforts need to be developed for control of invasive species both on-site and in the local landscape. The spread of colonizing Typha species can be slowed by early removal and planting efforts, but established invasive species such as Phalaris arundinacea may prove to be much more persistent and difficult to suppress. When existing in the local landscape, vegetative expansion can facilitate quick invasion, even in the face of management efforts. Though the long-term effects of planting are still unknown, strong measures should be taken to control for invasive species both on-site and in the local landscape in order to prevent established invasive species from spreading both within a site and to other local areas.

Economic Analysis

Basic costs for implementing wetland restoration projects in agricultural settings primarily include construction, and any loss of income from land that may have otherwise been used for crop production. To minimize the later and to increase the chances of restoring proper wetland hydrologies and higher levels of wetland plant biodiversity, it is recommended that such restorations be performed in historically wet, low-yield areas where tiling or ditching has been a historical necessity for crop production. By breaking drainage tiles are filling drainage ditches, much of the construction and grading costs that might otherwise be required in drier areas could be avoided.

Specifically, wetland construction at the EDIC site included the grading of the landscape to accommodate 6 wetland cells at a total cost of $15, 270.00. On a per wetland basis, this amounts to $2,545.00 per ½ acre wetland, or $5,090.00 per acre of wetland. Per the restoration strategy recommended through the findings of current studies at the EDIC site reported previously in this report—supplemental planting and seeding of native species—the additional cost for purchasing said propagules and seeds was $550.00 per ½ acre wetland, or $1100.00 per acre of wetland. Supplemental planting costs could also be offset (and were in this study) if restoration practitioners are willing, able, and allowed to collect native seeds and propagules from other local established restorations or native wetlands.

Much of the construction costs can be recouped through the various financial incentives offered in the aforementioned C.R.E.P. and W.R.P. programs, while annual rent payments will also provide additional income to land-owners for the acreage that is taken out of production for these restorations. Until studies pertaining to the water quality improvements achieved by each wetland strategy (planting and not-planting) are performed, no cost-benefit or larger economic analysis of nutrient (fertilizer) and sediment runoff sequestration can be calculated. However, these and other studies are in the planning stages at the EDIC site.

In summary, wetland restorations are a potential means for land-owners and farmers to convert low-yield, active, or formerly tiled or ditched agricultural areas back into wetland areas through programs like C.R.E.P. and W.R.P. Construction costs are offset by financial incentives provided by these programs, and annual rental payments are based on the amount of acreage enrolled. Results from this study show that with a small additional effort—either in the form of a monetary purchasing of propagules or in the form of the labor required for off-site propagule and seed collecting—supplemental planting and early management for invasive species can largely increase the amount of restored native plant biodiversity in a much shorter amount of time that what might be expected without such efforts.

Farmer Adoption

To date, there is no specific number of farmers that have “adopted” the ideas and recommendations presented in this study. However, Brad Masi, the farm manager at the George Jones Memorial Farm is very happy with the results and plans on continuing with the current management plans and research projects of local Oberlin professors. It is his hope that that many more short- and long-term studies will continue to be performed so that local residents, managers, and farmers might better understand the benefits of wetland restoration in agricultural areas, and the extent to which different wetland management strategies can provide improvements in terms of water quality and increases in local and regional wetland biodiversity.

Specific recommendations after two-years of wetland management studies include: 1) That farmers aim to restore seasonally wet, low productivity agricultural areas that have required ditching or drainage tile, as these areas may already have hydrological regimes approximating wetland conditions. As such, considerably less labor may be needed in order to restore wetland hydrology to these areas and there is a greater chance that local remnant populations of native wetland species have persisted nearby or in the seedbank. 2) To plant and seed locally or nursery obtained native wetland species in order to help establish plants that might not otherwise be able to establish on their own due to geographical or site isolation constraints. 3) That wetland restoration design should allow for the inclusion of muskrat colonies, where warranted, so as to provide natural suppression of typically dominant wetland flora. This could both increase diversity by increasing habitat heterogeneity, while also decreasing the amount of continued management that might be required to maintain an increased diversity. 4) To initially manage for aggressive invasive species such as Typha species and Phalaris arundincacea in order to prevent the future domination of the wetlands by these species, and consequently, a lower amount of native species diversity—particularly where muskrat activity and site isolation are concerns.


Areas needing additional study

Future studies at the EDIC wetlands should include the ability of previous restoration strategies (planting and not planting) to resist the establishment and spread of Typha and Phalaris arundinacea once initial management efforts for the exclusion of these species have ceased. Another study of the effects of artificially pulsed sediment and nutrient loads into the planted and unplanted wetlands should be performed in order to gauge the effects that various concentrations and sediment loading rates might have on the established plant communities, and what if any, water quality treatment functions become less efficient as a result. Studies assessing the ability of each restoration strategy (planted and unplanted) to improve water quality, and changes in each strategy's ability to do so in the presence of varied nutrient and sediment pulses could then be evaluated in relation to any persisting or changing measures in wetland vegetation or soil parameters. By drawing together the results from these and other studies, it may be possible to create documentation or resource guides for making better site-specific recommendations for land owners or wetland restoration managers undertaking restoration efforts in a variety of active and former agricultural landscapes.

Any opinions, findings, conclusions, or recommendations expressed in this publication are those of the author(s) and do not necessarily reflect the view of the U.S. Department of Agriculture or SARE.