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Creating riverine wetlands: Ecological succession, nutrient retention, and pulsing effects
Ecological Engineering 25 (2005) 510–527
Creating riverine wetlands: Ecological succession, nutrient retention, and pulsing effects William J. Mitsch ∗ , Li Zhang, Christopher J. Anderson, Anne E. Altor, Maria E. Hern´andez Olentangy River Wetland Research Park, School of Natural Resources, The Ohio State University, 352 W. Dodridge Street, Columbus, OH 43202, USA Received 17 February 2005; accepted 17 April 2005
Abstract Successional patterns, water quality changes, and effects of hydrologic pulsing are documented for a whole-ecosystem experiment involving two created wetlands that have been subjected to continuous inflow of pumped river water for more than 10 years. At the beginning of the growing season in the first year of the experiment (1994), 2400 individuals representing 13 macrophyte species were introduced to one of the wetland basins. The other basin was an unplanted control. Patterns of succession are illustrated by macrophyte community diversity and net aboveground primary productivity, soil development, water quality changes, and nutrient retention for the two basins. The planted wetland continued to be more diverse in plant cover 10 years after planting and the unplanted wetland appeared to be more productive but more susceptible to stress. Soil color and organic content continued to change after wetland creation and wetlands had robust features of hydric soils within a few years of flooding. Organic matter content in surface soils in the wetlands increased by approximately 1% per 3-year period. Plant diversity and species differences led to some differences in the basins in macrophyte productivity, carbon sequestration, water quality changes and nutrient retention. The wetlands continued to retain nitrate–nitrogen and soluble reactive phosphorus 10 years after their creation. There are some signs that sediment and total phosphorus retention are diminishing after 10 years of river flow. Preliminary results from the beginnings of a flood pulsing experiment in the two basins in 2003–2004 are described for water quality, nutrient retention, aboveground productivity, and methane and nitrous oxide gaseous fluxes. © 2005 Elsevier B.V. All rights reserved. Keywords: Created wetlands; Wetland restoration; Floodplain wetlands; Wetland succession; Olentangy River Wetland Research Park
1. Introduction
∗ Corresponding author. Tel.: +1 614 292 9774; fax: +1 614 292 9773. E-mail address: [email protected] (W.J. Mitsch).
0925-8574/$ – see front matter © 2005 Elsevier B.V. All rights reserved. doi:10.1016/j.ecoleng.2005.04.014
Wetlands are being created and restored at great frequency around the world both as “mitigation” wetlands that are meant to replace or compensate for wetland habitat loss and as wetland treatment systems for
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improving water quality. As important as the functions of wetlands are for providing values such as habitat structure and water quality improvement, there are scarcely any long-term data from thousands of created and restored non-tidal freshwater wetlands. In contrast, there are some results of long-term (>10-year) in the literature for created salt marshes and tidal freshwater marshes (Craft et al., 2002, 2003; Leck, 2003). Further, few studies have investigated how macrophyte diversity and cover affect ecosystem function in created and restored wetlands, despite the frequent use of macrophyte cover and species requirements as determinants of legal and ecological success of these wetlands in mitigating wetland loss (Mitsch et al., 1998; NRC, 2001). This paper presents the results of 10 years of study of two created riverine wetlands developed “from scratch” on non-wetland soils and maintained for those 10 years by pumped river water. These experimental wetlands allowed the simultaneous long-term study of three different aspects of newly created wetlands: (1) measuring the importance of wetland plant introduction on ecosystem function; (2) investigating the time it takes for the development of hydric soils at a site where no hydric soils previously existed; and (3) determining the long-term patterns of water quality changes of a flow-through wetland as it develops from open ponds of water to vegetated, hydric-soil, marshes. For the plant introduction study, our hypothesis in this study was that planted and unplanted wetlands would initially diverge in structure but eventually converge in structure. Three major questions in self-design were also part of this study: 1. Does human introduction of propagules have any measurable effect on ecosystem function? 2. At what rate and to what degree will a diverse biological community develop in newly created hydrologically open wetland ecosystems in which little biological life initially existed? 3. Does biodiversity affect ecosystem function? We believe that introducing macrophytes in created and restored wetlands could be regarded as a humaninduced switch that changes the resiliency of the wetlands and that this resiliency or lack thereof may be an important factor in allowing systems to cross thresholds that alter ecosystem function, recovery trajectory, and successional patterns. Because created and restored wetlands have rarely been monitored beyond the 5-year
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Fig. 1. Four possible outcomes of propagule introduction (e.g., planting of macrophytes) in newly created wetlands. Pathways 1 and 2 lead to healthy wetlands; Pathways 3 and 4 lead to a failed ecosystem state.
periods, the importance of species introduction in general and macrophyte planting in particular on ecosystem function remains poorly understood. Changes that are long-term in nature, e.g. woody plant invasion or soil carbon increases, can have significant effects on shorter-term dynamics of wetlands, e.g. algal and macrophyte vegetation dynamics. For example, Fig. 1 illustrates four possible ecosystem states possible for wetlands with two different initial conditions. In one case, a wetland is planted (high propagule introduction) and in the second case, a wetland is left to natural colonization (low propagule introduction). Achieving a healthy ecosystem state for shorter-term variables is possible in both situations (Pathways 1 and 2) but unexpected threshold shifts can occur (Pathways 3 and 4); these short-term dynamics couple with long-term ecosystem processes (Carpenter and Turner, 2001). We have 10 years detailed data on ecosystem structure and function in two 1-ha wetland basins maintained under strict and well documented hydrologic conditions that, when analyzed in a retrospective analysis, will reveal: (1) the importance of macrophyte introduction and subsequent macrophyte community diversity on ecosystem function and (2) the effects of this diversity on ecosystem resiliency to stressors and coupling of fast and slow turnover dynamics. Our hypothesis is that in wetland ecosystems with essentially identical forcing functions, the introduction of propagules can lead to alternative ecosystem states. These alternate states may or may not provide the functions necessary to meet restoration and regulatory needs for wetlands or provide the right conditions for water quality improvement.
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1.1. A whole-ecosystem approach Studies that attempt to link wetland function with structure are often done at inappropriate spatial and temporal scales to assist those responsible for the management of wetland landscapes. While there is no single optimum scale for ecosystem experimentation, it is well know that it is easier to apply statistical methods successfully when many small replicated systems or plots are used (Carpenter, 1998; Carpenter et al., 1998b). As a result, these scales are often chosen to make inferences about wetland management. Unfortunately, when results of small-scale, short-term studies are applied to full-scale conditions, conclusions are questionable at best. For example, Engelhardt and Ritchie (2001) manipulated seventy 1.5-m diameter wading pools with one, two, and three species of the submersed pondweed (Potamogeton spp.) over one growing season and found that higher algal biomass and higher phosphorus uptake occurred in the pools with highest macrophyte species richness. They concluded that higher species richness created up to 25% higher algal biomass and caused 30% more phosphorus uptake and thus would support more wildlife and fish. They further concluded that a wetland with high richness or diversity due to disturbance might better “sustain ecosystem functioning and promote the services of those wetlands to humans.” We consider this extrapolation of short-term results from 1.5-m pools to national wetland policy in a prestigious scientific journal to be infelicitous. Extrapolation of short-term microcosm experiments across temporal and spatial scales for adaptive management of ecosystems is questionable at best. Alternatives to the replicated small-scale mesocosms for wetland study are large-scale, long-term whole ecosystem studies that include more components of the ecosystem. Examples of such studies are described by Mitsch and Day (2004) and include systems such studies of tropical rain forests (Odum and Pigeon, 1970); forested watersheds (Likens et al., 1977); lakes (Schindler, 1977; Schindler et al., 1997; Carpenter et al., 1996,1998a), and wetlands (Odum et al., 1977; Mitsch et al., 1995). Whole-ecosystem studies are often criticized because the size, cost, and logistics do not allow for much if any replication. Yet the lack of replication can be compensated for by the decrease in variances that has been shown to result
when large-scale experiments are used. Results from large-scale experiments are less stochastic and thus more homeostatic, and often allow for the demonstration of ecosystem properties that otherwise would not appear in smaller scale experiments (Pomeroy et al., 1988; Odum, 1990; Carpenter et al., 1995; Carpenter, 1998). Kemp et al. (2001) illustrate in plankton experiments that as the scale of the experiment increases, the relative variance decreases. This may be a general principle and if it is, there is less need for great numbers of replications with increasing size of the experiment. There needs to be more understanding and acceptance of large-scale experiments and observations in the literature, if for no other reason than that they serve as a check on theories and management recommendations being published from smaller-scale studies chosen primarily because of elegant replications and statistics. 2. Methods 2.1. Experimental design Two 1-ha experimental wetlands and a river water delivery system were constructed in 1993–1994 at The Olentangy River Wetland Research Park, a 12ha wetland research site on the campus of The Ohio State University in Columbus (Fig. 2). Over 2400 plant propagules (mostly root stock and rhizomes) representing 13 species typical of Midwestern USA marshes were planted in one wetland (Wetland 1, W1) in May 1994 (Fig. 3). Wetland 2 (W2) remained unplanted. Both wetlands have received the same amount and quality of pumped river water and remained essentially identical hydroperiods over the 10-year period 1994–2003 (Fig. 4). River water is pumped into the wetlands continuously, day and night, except for occasional short-term unscheduled electrical failures. After start-up trials in 1994, a pumping protocol was developed that involves changing the pumping rate two or three times per week based on a formula that relates pumping rate to river discharge. In 2003, extramural funding from Ohio and Federal agencies was obtained for a pulsing study whereby artificial “floods” were introduced to the wetland basins; each wetland was administered with the same hydrologic conditions, so the “planting experiment” was not violated. Pumped inflow to each wetland has averaged 20–30 m/year. Water depths in the major portions of the wetland are
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Fig. 2. Olentangy River Wetland Research Park at Ohio State University, showing two experimental wetlands used for long-term whole-ecosystem study described in this proposal.
generally 20–40 cm in the shallow areas where most of the emergent macrophytes grow and 50–80 cm in the deepwater areas that were constructed in the wetland to allow overwintering of fish (for mosquito control) and long-term sediment storage. Early results illustrated system divergence and convergence (Mitsch et al., 1998) and water quality changes caused by the wetlands (Nairn and Mitsch, 2000; Spieles and Mitsch, 2000a) in the first 3 years of this study. There was a clear pattern of ecosystem divergence 6 years after planting (Mitsch et al., 2005a). Other studies published on the ecosystemscale experimental wetlands include those on aquatic system modeling (Metzker and Mitsch, 1997), algal
dynamics (Wu and Mitsch, 1998), hydrology (Koreny et al., 1999; Zhang and Mitsch, 2005), water quality (Kang et al., 1998), benthic invertebrates (Spieles and Mitsch, 2000b), sedimentation patterns (Harter and Mitsch, 2003), and Typha hybridization (Selbo and Snow, 2004). Methods for some of the prominent data collected annually at the experimental wetlands are described here. 2.2. Macrophyte coverage and community diversity Macrophyte coverage by dominant community has been estimated each year from aerial color photogra-
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to the same size basin map utilizing geographic information system software. 2.3. Macrophyte productivity Aboveground peak biomass has been used as an estimate of net aboveground primary productivity in the experimental wetlands since 1997 by direct aboveground harvesting of random 1-m2 plots in 16 general locations in each wetland along sampling boardwalks.
Fig. 3. Macrophyte species planted in experimental Wetland 1 (W1) in May 1994. Mudflat was designation used for shallow slightly sloping convex portion of experimental wetlands where the most diversity of plants were introduced. Graphic shows survival of individual plants, as marked by flagging, in June and August 1994 and August 1995.
phy taken at the period of peak biomass (late August), coupled with ground truth surveys. Ground surveys involved mapping plant communities along 500 m of seven transects in each wetland. Transects are on permanent walkways that are about 1.5 m above the wetland, thus giving a good perspective even with 3-m tall plants. A 10 m × 10 m grid system marked with permanent numbered PVC poles is used to identify the location of plant communities in each wetland. Vegetation community maps for each year are normalized
2.4. Soil sampling Soils have been comprehensively analyzed for color and organic content on three occasions: 1993 (prior to flooding); 1995 (17 months after flooding began), and in 2003 (approximately 10 years after construction). Surface soils were sampled using a 10-m grid system established for the wetlands after they were constructed. Permanent field markers were established at each grid point that allowed for direct comparisons of soil samples collected between years. At each grid point, a sample was collected at the surface (0–8 cm depth) and subsurface (8–16 cm depth). A total of 102 wetland soil samples were collected and analyzed in 1993 and 1995 (Nairn and Mitsch, 1996), and a total of 166 in 2003. Soil color was determined at multiple sites in each wetland basin by comparing soil color to a standard Munsell color
Fig. 4. Pattern of pumped inflow to the planted (Wetland 1) and naturally colonizing (Wetland 2) experimental wetlands, 1994–2003.
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chart and the percentage of samples ≤2 was determined. Percent organic matter was estimated for each sample through loss on ignition at 550 ◦ C for 1 h. 2.5. Water quality Sampling of water temperature, dissolved oxygen, pH, conductivity, and redox have been measured twiceper-day (dawn and dusk) for 10 years with water quality sondes at the inflow of the wetlands and outflows of both wetland basins. One-hundred-milliliter samples are also taken dawn and dusk each day at the inflow and two outflows for turbidity analyses in the laboratory with a Hach ratio turbidimeter. In addition to the twice-per-day manual sampling, weekly water samples are taken at inflow, middle, and outflows of the wetlands for nutrients (SRP, TP, NO3 ) that are determined by standard methods. Samples are split into filtered (0.45 m) and unfiltered samples, frozen until analysis, and analyzed for total phosphorus, soluble reactive phosphorus, and nitrate–nitrogen (USEPA, 1983; APHA, 1989). Both total phosphorus and soluble reactive phosphorus methods employ the ascorbic acid and a molybdate color reagent method with a Lachat QuikChem IV automated system. Total phosphorus samples are first digested by adding 0.5 ml of 5.6 N H2 SO4 and 0.2 g (NH4 )2 S2 O8 to 25 ml of sample and exposing the samples to a heated and pressurized environment for 20 min in an autoclave. Nitrate + nitrite were analyzed on a Lachat QuikChem IV automated system with the cadmium reduction method. Samples from April 1994 through July 1995 were run by similar methods by Heidelberg College Water Quality Laboratory using a Traacs 800 autoanalyzer. The accuracy of the nutrient analysis was checked every 10–20 samples with a known standard, and the samples are redone if the accuracy is off by 5%. 2.6. Hydrology Water stages and inflow rates are manually read from staff gauges and flow meters twice per day. Outflow is estimated from water stages. Water budgets for the wetland basins have been calculated for the past 10 years and published these results each year in our annual reports.
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2.7. Gas fluxes CH4 fluxes were quantified twice per month during a year of pulsing flow from March to September, 2004, in six 0.26-m2 plots in each wetland basin; half were located in the edge zones and contained emergent vegetation, and half were located in permanentlyinundated areas. Portable, non-steady-state chambers (Livingston and Hutchinson, 1995) were employed for gas flux measurements. Permanent PVC frames and HDPE bases were installed at edge-zone sampling locations; 4-mil, fitted polyethylene bags affixed with sampling ports were attached to the frames at the time of sampling. Permanently inundated areas were sampled with floating chambers made from the same materials. Sampling was conducted in the morning, afternoon, and after dark on each sampling date. Sampling took place during flood pulses, and at the end of the lowflow period before the subsequent pulse. Each flux rate was determined from three to five samples collected from each chamber at regular intervals over 30–60 min. Samples were injected into evacuated 10-ml autosampler vials and stored at 4 ◦ C until analysis, which was completed within 1 week. Gas samples were analyzed with a Shimadzu GC 14-A flame ionization detector (150 ◦ C) using a 40 position HTA Autosampler, a 1.8 m Porapak-Q column, and helium as the carrier gas (25 ml min−1 ). Flux rates were determined by the linear change in concentration by mass of CH4 –C over time (Holland et al., 1999). N2 O fluxes were measured using a closed chamber (Plexiglas chamber 24.5 cm × 24.5 cm × 70 cm) technique (Smith et al., 1983) from June 2003 through December 2004. Nine plots were sampled in each wetland near the inflow, middle, and outflow, with three plots in a transitional upland zone, three plots in the edges with alternate wet and dry conditions, and three plots in the marsh, which were regularly flooded. During gas collection the chambers were sealed for 2 h using a water seal, gas samples were withdrawn every 30 min. Samples were taken weekly, three times a month, when pulsing events occurred and once a month when no pulses were performed. All samples were taken between 11:00 a.m. and 3:30 p.m. for uniformity within the diurnal period. Nitrous oxide gas samples were analyzed using a gas chromatograph (Shimadzu GC-14-A) fitted with a 2 ml sampling loop, two Porapak-Q 1.8 m columns and an electron capture
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Ni-63 detector. The instrument used ultra pure nitrogen carrier gas (10 ml min−1 ) and operated at temperatures of 70, 80, and 300 ◦ C for column, injector, and detector, respectively. Nitrous oxide fluxes were calculated using a closed chamber flux equation (Holland et al., 1999).
3. Results and discussion 3.1. Macrophyte cover Fig. 5 presents dominant vegetation community patterns from 1994 through 2003. Patterns of dominance and subsidence of macrophyte communities is summarized in Fig. 6. By the end of the 10th growing season in 2003, Wetlands 1 (W1) and Wetland 2 (W2) had approximately 62 and 38% macrophyte cover respectfully. These values were considerably lower than the 73 and 74% cover in the two wetlands in 2002—the highest percent coverage in the basins since they were created in 1994. The decrease is probably due to the spring pulsing of water through both wetlands and some herbivory. From 1994, when there was no significant macrophyte vegetation cover in either basin, macrophyte coverage increased yearly in both basins through 1999. Wetland 1 had a greater percent coverage than Wetland 2 until 1997, when that trend reversed. Coverage decreased in both Wetlands 1 and 2, from 2000 to 2001, probably as a result of a combination of muskrat herbivory and possibly increased water levels due to outflow sedimentation. The overall pattern of vegetation can be summarized in several distinct periods. Wetland 1 was planted in 1994 and a distinct pattern of vegetation development around the edge of the wetland was observed in 1995 as a result while the “unplanted” wetland remained relatively free of macrophytes except for an edge of Populus seedlings beginning on the interior mudflat. By the third year, soft-stem bulrush Schoenoplectus tabernaemontani (a.k.a. Scirpus validus) had colonized in the unplanted wetland, and by the end of the third growing season, it appeared that the planted and unplanted wetlands had converged with a domination of cover by Schoenoplectus. Typha spp. dominance increased dramatically in the naturally colonizing Wetland 2 from 1996 to 1999. It generally has remained less than 17% or less cover in the planted Wetland 1. By 1999, the naturally coloniz-
ing wetland (W2) was totally dominated by a very productive cover of Typha while the planted wetland (W1) had a diversity of communities including ones dominated by four communities: Sparganium eurycarpum, S. tabernaemontani, Typha spp., and Scirpus fluviatilis. Wetland macrophyte coverage began to significantly erode in 2000 and, by 2001, the wetlands had only 27.6 and 17.4% macrophyte cover in W1 and W2, respectively. This macrophyte vegetation loss was caused primarily by a peak in muskrat activity that followed the year of maximum plant coverage in 1999. We believe that after the muskrat activity peak, as measured by winter lodge counts, the absence of vegetation cover for food, protection, and lodge construction caused the muskrat population to drop significantly. Predation undoubtedly was one reason but movement of muskrats to the river and other wetlands is also likely. A significant drawdown of both basins was done in spring and early summer 2002 to allow the seed bank to reset. The approach was successful. At the end of the 2002 growing season, vegetation coverage was the highest it had ever been (73–74% cover in the basins) and Typha coverage was only 9% of the total area of Wetland 2 and 5% of the total area of Wetland 1. This was considerably reduced from peak year 1999 when it was 56% of the cover in Wetland 2. One of the most significant changes in 2002 was the increased coverage by S. tabernaemontani in both wetlands, apparently from the marsh seedbank. S. tabernaemontani dominance increased in both basins in 2002 from 0.3 to 52% coverage in W1 and from 0 to 63% in W2. In January 2003, we started a pulsing experiment in W1 and W2 where several 7-day duration floods were pulsed through the wetlands, mostly in late winter and spring. This pulsing was one of the apparent reasons for the reduction in macrophyte cover in the wetland basins in that year. 3.2. Macrophyte community diversity We developed a macrophyte community diversity index (CDI) to quantify spatial diversity in these experimental wetlands (Mitsch et al., 2005a). The index is expressed as: CDI =
N i=1
(Ci ln(Ci ))
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Fig. 5. Vegetation community cover in the planted (Wetland 1) and naturally colonizing (Wetland 2) experimental wetlands at Olentangy River Wetland Research Park, 1994–2003.
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CDI increased in W1 as a good balance among five communities developed. Four of those communities were dominated by plants introduced in the planting. Diversity dropped in 2000 and 2001 during the years of extensive herbivory but dramatically increased again in 2002 and 2003 (years 9 and 10) when the wetland vegetation reset. 3.3. Macrophyte productivity Fig. 6. Typha and other macrophyte community cover in the planted (Wetland 1) and naturally colonizing (Wetland 2) experimental wetlands at Olentangy River Wetland Research Park, 1994–2003. Episodes of high muskrat herbivory in 2000 and 2001 are also indicated.
where Ci is the percent cover of community “i” (0–1) and N is the number of plant/aquatic communities. Overall, there were 13 different communities identified from ground surveys supported by aerial photography during the study, based on the dominant species in communities (see Fig. 5). When the wetlands are viewed in terms of our community diversity index, which includes evenness of plant cover as well as number of dominant communities, an interesting pattern of macrophyte community diversity in the two wetlands is seen (Fig. 7). The data show a general pattern of higher diversity in the planted wetland with the exception of years 3–5 when community diversity was similar in the two basins. By the sixth year (1999), after the 3 years of similar plant cover in the two basins, different spatial community diversity developed. The CDI in W2 dropped as Typha formed close to a monoculture; the
Fig. 7. Macrophyte diversity in the planted (Wetland 1) and naturally colonizing (Wetland 2) experimental wetlands at Olentangy River Wetland Research Park, 1994–2003. CDI stands for community diversity index, which is defined in the text.
Macrophyte net aboveground primary productivity (NAPP), as estimated by peak aboveground biomass in August each year, decreased in both wetlands in 2003 compared to similar plot readings in 2002 (Fig. 8a). When paired sites were compared between the two wetlands (16 sites were paired in 2003) macrophyte aboveground biomass was statistically higher in the planted Wetland 1 than in the naturally colonized Wet-
Fig. 8. Indicators of macrophyte productivity in the planted (Wetland 1) and naturally colonizing (Wetland 2) experimental wetlands, 1997–2003: (a) peak aboveground biomass in vegetation zones and (b) cumulative organic matter production, kilograms per year in each experimental wetland basin.
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land 2 (t = 0.0002; α = 0.05) for the first time in 7 years. This is because of a significant decrease in macrophyte cover in the southern half of Wetland 2 in 2003 due, we believe, to a combination of higher herbivory and effects of temporary high water levels caused by pulsing. Before that, peak aboveground biomass was significantly higher in plots in Wetland 2 than in Wetland 1 for 4 straight years from 1998 to 2001. Based on the aboveground biomass and estimates of macrophyte cover presented above, aboveground productivity by macrophytes in each wetland basin was estimated for 1997–2003 (Fig. 8b). Overall macrophyte organic productivity decreased 46% in Wetland 1 and 81% in Wetland 2 from 2002 to 2003. These numbers are significant for two reasons. First, although the macrophyte community substantially recovered in 2002 from the herbivory and subsequent macrophyte losses of 2000 and 2001, productivity once more decreased in 2003, due to herbivory in Wetland 2 and possibly due to high water levels that were a result of a pulsing experiment in both basins in spring 2003. Second, 2003 is the second year in a row where the planted Wetland 1 had a higher estimated macrophyte carbon sequestration than the naturally colonized Wetland 2. The total organic matter production by macrophytes over the last 7-year period is now almost the same in the two wetland basins at 22–23 tonnes per basin. 3.4. Soil development Soil changes at the wetland surface have been substantial since the addition of water to the basins in 1994 (Fig. 9). Most changes have occurred at the surface (0–8 cm depth) because of the prevailing influence of sedimentation and organic accumulation. Soil organic matter (SOM) increased by 63% in the upper 8 cm of wetland soils over the decade to an average of 8.6%. On average, surface organic matter content increased by approximately percentage point each 3 years. In contrast, Craft et al. (2003) found an increase in the amount of carbon per unit area in created salt marshes after 10 years, but to levels 18% of those found in natural salt marshes. Color in the soils changed dramatically after 10 years of inundation (Figs. 9b and 10) with 94% of the samples from both 0–8 cm depth and 8–16 cm depth having soil chromas of 2 or less in 2003. Before the wetlands were created in 1993, no soil samples indicated soil chroma of 2 or less. In 1995, 17 months after
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Fig. 9. Experimental wetlands soil data pooled for both experimental wetlands: (a) percent organic matter in 1993, 1995, and 2003, and (b) percent of soil samples that exhibited chroma values of 2 or less in 1993, 1995, and 2003. Soil data indicated as 2003 were collected between December 2002 and March 2003.
water was added to the wetlands, 78% of the samples taken in the surface layer had soil chromas of 2 or less while only 24% in the 8–16 cm depth had chromas of 2 or less. Five years after the creation of a freshwater marsh, Vepraskas et al. (1999) found soil matrices depleted of iron and low chromas. 3.5. Water quality Patterns of water quality changes through both wetlands have been pronounced and variable over the 10 years of wetland development, reflecting changes from characteristics of eutrophic shallow ponds in the early years (Wu and Mitsch, 1998) to functioning wetlands in latter years. The changes in water quality are presented as “percent change” in all subsequent graphics. Even though the results are presented as only one number per
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Fig. 10. General pattern of surface soil color change, as indicated on a Munsell Color Chart, averaged for both experimental wetlands at Olentangy River Wetland Research Park from 1993 (prior to wetland basin flooding) to 2002. Note that the soil chroma decreased to values of 2 or less after 1996, or 2 years after the basins were flooded.
wetland for each year, statistical differences have been determined between the wetlands by paired tests of the outflow concentrations. Water temperature increases through the wetlands (Fig. 11a) showed a consistent pattern of decreasing (there was less net warming from inflow to outflow each year) over the first 6 years in each wetland as vegetation developed to maximum cover by 1999. Then muskrat herbivory removed macrophyte cover in 2000 and especially in 2001 (see Fig. 5), exposing the water surface to full sunlight that resulted in a net effect of warmer temperatures. The wetland vegetation recovered dramatically in 2002 and temperature patterns once again were less dramatic. That pattern reversed in 2003 when our hydrologic pulsing experiment decreased vegetation cover but also changed the basin retention rates, causing higher temperature increases again. Dissolved oxygen change through the wetlands (Fig. 11b) showed a fairly consistent pattern over the 10 years of decreasing in importance, probably due to
the continued decrease in the importance of aquatic primary productivity. The site was an experimental farm for decades prior to the construction of the wetlands and initial algal blooms were long in duration and intense (Wu and Mitsch, 1998). Gradually, high concentrations of nutrients have been removed from the wetland soils, resulting in lower aquatic primary productivity. The low change in dissolved oxygen persisted in the years 1999–2003 despite the fact that macrophyte cover was lower in most of those years. In 2003, the flood pulsing may have decreased water column productivity in the basins. Patterns of pH change through the wetland basins (Fig. 11c) mirrored those seen for dissolved oxygen. Early in the study it was common to see dramatic increases in pH in the late afternoon due to high productivity in the water column. In latter years, that was no longer the case. Conductivity (Fig. 11d) decreased dramatically through the wetlands in the second and third years when water column productivity was the
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Fig. 11. Percent changes from inflow to outflow of water quality parameters in the planted (Wetland 1) and naturally colonizing (Wetland 2) experimental wetlands at the Olentangy River Wetland Research Park, 1994–2003 for: (a) water temperature, (b) dissolved oxygen, (c) pH, (d) conductivity, (e) turbidity, and (f) redox potential.
highest. We documented significant precipitation of calcium carbonate in those years (Liptak, 2000). Conductivity decreased again in 2003 (the pulsing year) by almost 12% in both basins. Suspended sediments, as reflected in turbidity measurements taken over the 10 years of wetland function, showed an interesting pattern of ecosystem succession (Fig. 11e). With each succeeding year, the wetlands appear to be removing fewer suspended sediments. By the ninth year, both wetlands were exporting more sediments than were coming in the wetland. Redox potential (Fig. 11f) shows an oscillating 10-year pattern that made little ecological sense. The year with
the hydrologic pulses (2003) led to one of the highest drops in redox potential through the wetlands since measurements began. There could be related to the fact that when the wetland “expands in size” with the pulses, the surface water comes in contact with more anaerobic soil waters on the edges of the wetlands that are otherwise isolated from the wetlands. 3.6. Nutrient retention These created wetlands have been fairly significant in removing nutrients over the past decade. Patterns of NO3 –N reduction through the wetlands have averaged
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Fig. 12. Percent changes from inflow to outflow in the planted (Wetland 1) and naturally colonizing (Wetland 2) experimental wetlands at the Olentangy River Wetland Research Park, 1994–2003 for: (a) nitrate + nitrite-nitrogen; (b) soluble reactive phosphorus (SRP); (c) total phosphorus (TP).
close to 35% by concentration and mass (Mitsch et al., 2005b; Fig. 12a). Nitrate removal patterns in the naturally colonizing wetland have been erratic, ranging from removal rates in the 20% range 1 year, and high 40% range the next. The planted and more diverse wetland (WI) has removal rates more consistent from year to year. The spring pulsing in 2003 did not appear to reduce nitrate–nitrogen removal. In fact, the average nitrate removal by the wetlands was the highest in the two wetlands (42% reduction) as it has been since the beginning of the study a decade earlier. The retention of soluble reactive phosphorus (SRP) in the wetland basins (Fig. 12b) continued to be significant at around 70% through the 10th year 2003 (no nutrient data are reported for 2002). In the 10th year after wetland construction and during the pulsing year, the planted wetland retained 63% of SRP while the naturally colonizing wetland retained 73% (not significantly different, α = 0.05). Total phosphorus (TP) showed a much different picture with little retention and even export by the
10th year (Fig. 12c). TP decreased by only 2% in W2 and increased by 14% in W1 after a consistent pattern of total phosphorus retention during the first 8 years since wetland construction. This pattern is consistent with the pattern seen for turbidity in the 2003 pulsing year where there was either a very small reduction or slight increase in turbidity in that year. The sediments that are discharging from the wetlands have significant phosphorus bound to the soil particles. 3.7. Pulsing effects 3.7.1. Macrophyte productivity In 2003, flood pulses of 7-day durations and inflow of approximately 500 gpm were pumped into each of the two experimental wetlands. Preliminary results of the 2003 flood pulses are described here. Macrophyte vegetation NAPP appeared to be lower during the pulsing in 2003 compared to the previous year (Fig. 8a). High water levels in the spring caused by the pulses may
W.J. Mitsch et al. / Ecological Engineering 25 (2005) 510–527 1.0 (102)* 0.52 (92)* 23 (109)* 0.05 (94)* 9 (113)* 3 (57) ± ± ± ± ± ± 13.6 11.42 535 8.23 252 23 1.0 0.55 (89) 21 (103)* 0.07 (87)* 10 (102) 3 (61)*
± ± ± ± ± ± 12.2 11.94 723 8.17 246 15 1.8 0.73 (25)* 54 (32) 0.07 (28)* 15 (32) 4 (22) Significant different at α = 0.05 between inflow vs. outflow. *
± ± ± ± ± ± 11.9 11.96 729 8.15 264 19 (◦ C)
Temperature DO (mg/l) Conductivity (/cm) pH Redox (mV) Turbidity (NTU)
10.3 10.79 726 7.97 261 24
± ± ± ± ± ±
1.5 (31) 0.57 (24) 51 (32) 0.06 (27) 10 (31) 5 (22)
Outflow 1
(31)*
Outflow 2
(31)*
1.8 1.02 (25) 50 (32) 0.09 (28)* 13 (31) 2 (22)
11.5 10.69 650 7.93 281 29
± ± ± ± ± ± Inflow Inflow
0.8 (94) 0.38 (80) 23 (101) 0.06 (86) 9 (105) 4 (63)
13.2 10.75 530 8.25 257 23
± ± ± ± ± ±
(95)* Outflow 1 Non-pulsing Pulsing Parameters
3.7.3. Gas exchange Studies are underway to evaluate the effects of hydrologic pulsing on gas exchange from the wetlands, particularly for methane and nitrous oxide. Preliminary results of the effects of flood pulsing on methane flux in the experimental wetlands (Fig. 13) show that methane emission in vegetated areas increased significantly when flooding occurred in zones that previously did not have standing water. However, preliminary results also suggest that overall “average” methane emission from pulsed sites was less than that from those sites where flooding was permanent. A seasonal pattern in N2 O emissions was observed during the study period with highest rates in summer. In summer 2003 and winter 2004, N2 O emissions were higher when no pulses occurred than when flood pulsing occurred. In spring 2004, N2 O fluxes were higher during pulsing than during non-pulsing conditions.
Table 1 Comparison of two-per-day water quality samples in the two experimental wetlands during flood pulsing and non-pulsing periods in 2003
3.7.2. Water quality Some preliminary effects of the flood pulses or artificial “floods” in 2003 on water quality in the two experimental wetlands are shown in Tables 1 and 2. Water quality changes during the pulsing events of 2003 are immediate. When the two wetlands are investigated with frequent two-per-day sampling, water temperature and pH significantly increased in both basin outflows during flood pulses (Table 1, α = 0.05). Between flood pulses (low-flow conditions), temperature and pH were again higher and conductivity was lower in both wetland outflows compared to the inflows (Table 1). When data are investigated with both wetland basins combined at the outflow swale (Table 2), there is a significant difference between pulsed and non-pulsed conditions for conductivity and soluble reactive phosphorus. Both conductivity and soluble reactive phosphorus decreased more rapidly during non-pulsing conditions than during pulses.
Outflow 2
have had a detrimental effect on emergent macrophyte productivity but the lower productivity could have also been due to other causes. The pattern was particularly acute in W2 where macrophytes disappeared from the southern half of the basin in 2003 (Fig. 5). Higher community diversity in W1 may have insulated it from this stress caused by the flooding whereas the lower diversity in W2 made it more susceptible to the combined stress of flooding and grazing.
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Table 2 Effects of flood pulsing on water quality in the experimental wetlands in 2003 Inflow
Swale outflow
Percent change
Temperature (◦ C) Pulse 12.4 ± 2.8 (11) Non-pulse 15.6 ± 1.8 (16)
13.6 ± 2.6 (13) 16.6 ± 1.9 (16)
+9.7 +6.4
Dissolved oxygen (mg/l) Pulse 7.47 ± 1.33 (10) Non-pulse 9.29 ± 0.70 (14)
6.98 ± 1.37 (11) 8.66 ± 0.84 (16)
+6.6 +6.8
Conductivity (/cm) Pulse 705 ± 76 (11) Non-pulse 627 ± 33 (16)
640 ± 82 (13) 556 ± 37 (18)
−9.2* −11.3*
pH Pulse Non-pulse
8.28 ± 0.23 (7) 7.80 ± 0.14 (15)
8.28 ± 0.15 (7) 7.88 ± 0.08 (16)
0 +1.0
Redox (mV) Pulse Non-pulse
237 ± 28 (11) 263 ± 26 (16)
292 ± 35 (13) 230 ± 20 (18)
+23.2 −12.5
2.88 ± 0.38 (15) 2.01 ± 0.44 (26)
−31 −55
Nitrate–nitrogen (mg N/l) Pulse 4.17 ± 0.50 (15) Non-pulse 4.48 ± 0.51 (18) SRP (g P/l) Pulse Non-pulse
27 ± 3 (3) 46 ± 26 (3)
15 ± 6 (2) 5 ± 1 (6)
−44* −89*
Total P (g P/l) Pulse 54 ± 12 (3) Non-pulse 103 ± 38 (3)
47 ± 2 (3) 44 ± 9 (5)
−13 −57
Data for SRP and total P are for July–September and cover only one pulse in early August. Swale outflow is combined outflow from two experimental wetlands. * Significant difference (α = 0.05) in swale concentrations during pulsing and non-pulsing.
Fig. 13. Average (±S.E.) of methane flux for plots pooled for both experimental wetlands containing emergent vegetation, when inundated vs. when they were exposed, in 2004 growing season (April 25–September 29, 2004). Methane flux from permanently-inundated areas is shown for comparison.
had low productivity. Seven years of productivity data for each of the two wetlands (14 wetland-years) provides results that are not as obvious (Fig. 15). Data from the planted wetland W1, which has never been dominated by one species, appears to have a pattern of highest productivity at a mid-level of community diversity; inclusion of productivity and diversity data from the Typha-dominated wetland W2 does not fit the same pattern.
Overall, when marsh and edge plots were inundated, nitrous oxide fluxes were lower than when the same plots were exposed to air (Fig. 14). In the upland plots, nitrous oxide fluxes were not different between flooded and dry periods. 3.8. Wetland productivity and diversity Our 10 years of whole-ecosystem measurements give us the opportunity to investigate the connections between diversity and productivity. Previously, we showed that there appeared to be a negative relationship between the two (Mitsch et al., 2005a). Monospecific wetlands were highly productive and diverse wetlands
Fig. 14. Average (±S.E.) of N2 O flux pooled for both experimental wetlands during flooding and non-flooding periods from June 2003 through December 2004. Marsh refers to areas normally flooded; edge refers to areas that are flooded seasonally with flood pulses; T-upl refers to uplands that are rarely flooded.
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Fig. 15. Net aboveground primary productivity, estimated from aboveground biomass, as a function of community diversity index (CDI) for experimental wetlands, 1997–2003.
4. Conclusions After 10 years of observing these experimental wetlands, the following conclusions can be drawn regarding the influence of planting of ecosystem succession and on the general development of a wetland when it is created on formerly non-wetland soil: • Contrary to conventional wisdom, wetlands can be created if the correct hydrologic conditions are available. Created wetlands with the proper hydrology can develop appropriate biota (e.g., wetland plants) and physiochemistry (e.g., hydric soils) relatively rapidly without the need for planting if the proper hydrologic conditions are present and plant propagules are continually introduced as they were in our riverine wetlands. • Planting has a profound effect on ecosystem function of created wetlands, even several years after planting. The naturally colonizing wetland and the planted wetland diverged in macrophyte cover that affected ecosystem function on several occasions during the 10 years of this study. Differences in plant compostion and productivity, caused in all probability by the introduction of plants to one of the wetlands, led to differences in water quality and carbon accumulation in the wetlands. • Hydric soils as valid indicators of natural wetlands can develop within 2–3 years of wetland creation. We found that soil organic matter can increase by 1% every 3 years in temperate zone created marshes.
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• Water quality changes significantly as a wetland develops in primary succession and as perturbations occur. Some changes in water quality are direct and the immediate result of macrophyte cover and aquatic metabolism; other changes occur over longer periods due to sediment accumulation, and soil and redox changes. • Wetlands, if they are not overloaded with nutrients, can be effective nutrient sinks for many years. Our systems were subjected to nutrient concentrations and loadings that were significantly less than those usually added to treatment wetlands. Thus, our wetlands continued to be nutrient sinks for 10 years and support a reasonable plant diversity as well. • The addition of species to enhance biodiversity in wetland creation actually can lead to lower productivity of macrophytes and subsequent changes in the food web and water quality. • There are desirable values from both “diverse” marshes and “high productivity” marshes and to make generalizations that one is better than the other is wrong. • The continual introduction of plant, animal, and microbial species through water flows, atmospheric transport, and biological vectors give self-design a significant opportunity to manifest itself. • Flood pulses have significant influences on water and gaseous fluxes from wetlands. Continued study of the importance of periodic floods is needed on the combined effects on water and gas fluxes in these and other wetlands.
Acknowledgements We appreciate the many post-docs, graduate students, undergraduate students, and staff who assisted in this research over the 11 year period described here. Particular appreciation is given to Xinyuan Wu and Naiming Wang for their extraordinary assistance in the 1990s. Olentangy River Wetland Research Park paper 05-010.
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Creating riverine wetlands: Ecological succession, nutrient retention, and pulsing effects William J. Mitsch ∗ , Li Zhang, Christopher J. Anderson, Anne E. Altor, Maria E. Hern´andez Olentangy River Wetland Research Park, School of Natural Resources, The Ohio State University, 352 W. Dodridge Street, Columbus, OH 43202, USA Received 17 February 2005; accepted 17 April 2005
Abstract Successional patterns, water quality changes, and effects of hydrologic pulsing are documented for a whole-ecosystem experiment involving two created wetlands that have been subjected to continuous inflow of pumped river water for more than 10 years. At the beginning of the growing season in the first year of the experiment (1994), 2400 individuals representing 13 macrophyte species were introduced to one of the wetland basins. The other basin was an unplanted control. Patterns of succession are illustrated by macrophyte community diversity and net aboveground primary productivity, soil development, water quality changes, and nutrient retention for the two basins. The planted wetland continued to be more diverse in plant cover 10 years after planting and the unplanted wetland appeared to be more productive but more susceptible to stress. Soil color and organic content continued to change after wetland creation and wetlands had robust features of hydric soils within a few years of flooding. Organic matter content in surface soils in the wetlands increased by approximately 1% per 3-year period. Plant diversity and species differences led to some differences in the basins in macrophyte productivity, carbon sequestration, water quality changes and nutrient retention. The wetlands continued to retain nitrate–nitrogen and soluble reactive phosphorus 10 years after their creation. There are some signs that sediment and total phosphorus retention are diminishing after 10 years of river flow. Preliminary results from the beginnings of a flood pulsing experiment in the two basins in 2003–2004 are described for water quality, nutrient retention, aboveground productivity, and methane and nitrous oxide gaseous fluxes. © 2005 Elsevier B.V. All rights reserved. Keywords: Created wetlands; Wetland restoration; Floodplain wetlands; Wetland succession; Olentangy River Wetland Research Park
1. Introduction
∗ Corresponding author. Tel.: +1 614 292 9774; fax: +1 614 292 9773. E-mail address: [email protected] (W.J. Mitsch).
0925-8574/$ – see front matter © 2005 Elsevier B.V. All rights reserved. doi:10.1016/j.ecoleng.2005.04.014
Wetlands are being created and restored at great frequency around the world both as “mitigation” wetlands that are meant to replace or compensate for wetland habitat loss and as wetland treatment systems for
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improving water quality. As important as the functions of wetlands are for providing values such as habitat structure and water quality improvement, there are scarcely any long-term data from thousands of created and restored non-tidal freshwater wetlands. In contrast, there are some results of long-term (>10-year) in the literature for created salt marshes and tidal freshwater marshes (Craft et al., 2002, 2003; Leck, 2003). Further, few studies have investigated how macrophyte diversity and cover affect ecosystem function in created and restored wetlands, despite the frequent use of macrophyte cover and species requirements as determinants of legal and ecological success of these wetlands in mitigating wetland loss (Mitsch et al., 1998; NRC, 2001). This paper presents the results of 10 years of study of two created riverine wetlands developed “from scratch” on non-wetland soils and maintained for those 10 years by pumped river water. These experimental wetlands allowed the simultaneous long-term study of three different aspects of newly created wetlands: (1) measuring the importance of wetland plant introduction on ecosystem function; (2) investigating the time it takes for the development of hydric soils at a site where no hydric soils previously existed; and (3) determining the long-term patterns of water quality changes of a flow-through wetland as it develops from open ponds of water to vegetated, hydric-soil, marshes. For the plant introduction study, our hypothesis in this study was that planted and unplanted wetlands would initially diverge in structure but eventually converge in structure. Three major questions in self-design were also part of this study: 1. Does human introduction of propagules have any measurable effect on ecosystem function? 2. At what rate and to what degree will a diverse biological community develop in newly created hydrologically open wetland ecosystems in which little biological life initially existed? 3. Does biodiversity affect ecosystem function? We believe that introducing macrophytes in created and restored wetlands could be regarded as a humaninduced switch that changes the resiliency of the wetlands and that this resiliency or lack thereof may be an important factor in allowing systems to cross thresholds that alter ecosystem function, recovery trajectory, and successional patterns. Because created and restored wetlands have rarely been monitored beyond the 5-year
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Fig. 1. Four possible outcomes of propagule introduction (e.g., planting of macrophytes) in newly created wetlands. Pathways 1 and 2 lead to healthy wetlands; Pathways 3 and 4 lead to a failed ecosystem state.
periods, the importance of species introduction in general and macrophyte planting in particular on ecosystem function remains poorly understood. Changes that are long-term in nature, e.g. woody plant invasion or soil carbon increases, can have significant effects on shorter-term dynamics of wetlands, e.g. algal and macrophyte vegetation dynamics. For example, Fig. 1 illustrates four possible ecosystem states possible for wetlands with two different initial conditions. In one case, a wetland is planted (high propagule introduction) and in the second case, a wetland is left to natural colonization (low propagule introduction). Achieving a healthy ecosystem state for shorter-term variables is possible in both situations (Pathways 1 and 2) but unexpected threshold shifts can occur (Pathways 3 and 4); these short-term dynamics couple with long-term ecosystem processes (Carpenter and Turner, 2001). We have 10 years detailed data on ecosystem structure and function in two 1-ha wetland basins maintained under strict and well documented hydrologic conditions that, when analyzed in a retrospective analysis, will reveal: (1) the importance of macrophyte introduction and subsequent macrophyte community diversity on ecosystem function and (2) the effects of this diversity on ecosystem resiliency to stressors and coupling of fast and slow turnover dynamics. Our hypothesis is that in wetland ecosystems with essentially identical forcing functions, the introduction of propagules can lead to alternative ecosystem states. These alternate states may or may not provide the functions necessary to meet restoration and regulatory needs for wetlands or provide the right conditions for water quality improvement.
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1.1. A whole-ecosystem approach Studies that attempt to link wetland function with structure are often done at inappropriate spatial and temporal scales to assist those responsible for the management of wetland landscapes. While there is no single optimum scale for ecosystem experimentation, it is well know that it is easier to apply statistical methods successfully when many small replicated systems or plots are used (Carpenter, 1998; Carpenter et al., 1998b). As a result, these scales are often chosen to make inferences about wetland management. Unfortunately, when results of small-scale, short-term studies are applied to full-scale conditions, conclusions are questionable at best. For example, Engelhardt and Ritchie (2001) manipulated seventy 1.5-m diameter wading pools with one, two, and three species of the submersed pondweed (Potamogeton spp.) over one growing season and found that higher algal biomass and higher phosphorus uptake occurred in the pools with highest macrophyte species richness. They concluded that higher species richness created up to 25% higher algal biomass and caused 30% more phosphorus uptake and thus would support more wildlife and fish. They further concluded that a wetland with high richness or diversity due to disturbance might better “sustain ecosystem functioning and promote the services of those wetlands to humans.” We consider this extrapolation of short-term results from 1.5-m pools to national wetland policy in a prestigious scientific journal to be infelicitous. Extrapolation of short-term microcosm experiments across temporal and spatial scales for adaptive management of ecosystems is questionable at best. Alternatives to the replicated small-scale mesocosms for wetland study are large-scale, long-term whole ecosystem studies that include more components of the ecosystem. Examples of such studies are described by Mitsch and Day (2004) and include systems such studies of tropical rain forests (Odum and Pigeon, 1970); forested watersheds (Likens et al., 1977); lakes (Schindler, 1977; Schindler et al., 1997; Carpenter et al., 1996,1998a), and wetlands (Odum et al., 1977; Mitsch et al., 1995). Whole-ecosystem studies are often criticized because the size, cost, and logistics do not allow for much if any replication. Yet the lack of replication can be compensated for by the decrease in variances that has been shown to result
when large-scale experiments are used. Results from large-scale experiments are less stochastic and thus more homeostatic, and often allow for the demonstration of ecosystem properties that otherwise would not appear in smaller scale experiments (Pomeroy et al., 1988; Odum, 1990; Carpenter et al., 1995; Carpenter, 1998). Kemp et al. (2001) illustrate in plankton experiments that as the scale of the experiment increases, the relative variance decreases. This may be a general principle and if it is, there is less need for great numbers of replications with increasing size of the experiment. There needs to be more understanding and acceptance of large-scale experiments and observations in the literature, if for no other reason than that they serve as a check on theories and management recommendations being published from smaller-scale studies chosen primarily because of elegant replications and statistics. 2. Methods 2.1. Experimental design Two 1-ha experimental wetlands and a river water delivery system were constructed in 1993–1994 at The Olentangy River Wetland Research Park, a 12ha wetland research site on the campus of The Ohio State University in Columbus (Fig. 2). Over 2400 plant propagules (mostly root stock and rhizomes) representing 13 species typical of Midwestern USA marshes were planted in one wetland (Wetland 1, W1) in May 1994 (Fig. 3). Wetland 2 (W2) remained unplanted. Both wetlands have received the same amount and quality of pumped river water and remained essentially identical hydroperiods over the 10-year period 1994–2003 (Fig. 4). River water is pumped into the wetlands continuously, day and night, except for occasional short-term unscheduled electrical failures. After start-up trials in 1994, a pumping protocol was developed that involves changing the pumping rate two or three times per week based on a formula that relates pumping rate to river discharge. In 2003, extramural funding from Ohio and Federal agencies was obtained for a pulsing study whereby artificial “floods” were introduced to the wetland basins; each wetland was administered with the same hydrologic conditions, so the “planting experiment” was not violated. Pumped inflow to each wetland has averaged 20–30 m/year. Water depths in the major portions of the wetland are
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Fig. 2. Olentangy River Wetland Research Park at Ohio State University, showing two experimental wetlands used for long-term whole-ecosystem study described in this proposal.
generally 20–40 cm in the shallow areas where most of the emergent macrophytes grow and 50–80 cm in the deepwater areas that were constructed in the wetland to allow overwintering of fish (for mosquito control) and long-term sediment storage. Early results illustrated system divergence and convergence (Mitsch et al., 1998) and water quality changes caused by the wetlands (Nairn and Mitsch, 2000; Spieles and Mitsch, 2000a) in the first 3 years of this study. There was a clear pattern of ecosystem divergence 6 years after planting (Mitsch et al., 2005a). Other studies published on the ecosystemscale experimental wetlands include those on aquatic system modeling (Metzker and Mitsch, 1997), algal
dynamics (Wu and Mitsch, 1998), hydrology (Koreny et al., 1999; Zhang and Mitsch, 2005), water quality (Kang et al., 1998), benthic invertebrates (Spieles and Mitsch, 2000b), sedimentation patterns (Harter and Mitsch, 2003), and Typha hybridization (Selbo and Snow, 2004). Methods for some of the prominent data collected annually at the experimental wetlands are described here. 2.2. Macrophyte coverage and community diversity Macrophyte coverage by dominant community has been estimated each year from aerial color photogra-
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to the same size basin map utilizing geographic information system software. 2.3. Macrophyte productivity Aboveground peak biomass has been used as an estimate of net aboveground primary productivity in the experimental wetlands since 1997 by direct aboveground harvesting of random 1-m2 plots in 16 general locations in each wetland along sampling boardwalks.
Fig. 3. Macrophyte species planted in experimental Wetland 1 (W1) in May 1994. Mudflat was designation used for shallow slightly sloping convex portion of experimental wetlands where the most diversity of plants were introduced. Graphic shows survival of individual plants, as marked by flagging, in June and August 1994 and August 1995.
phy taken at the period of peak biomass (late August), coupled with ground truth surveys. Ground surveys involved mapping plant communities along 500 m of seven transects in each wetland. Transects are on permanent walkways that are about 1.5 m above the wetland, thus giving a good perspective even with 3-m tall plants. A 10 m × 10 m grid system marked with permanent numbered PVC poles is used to identify the location of plant communities in each wetland. Vegetation community maps for each year are normalized
2.4. Soil sampling Soils have been comprehensively analyzed for color and organic content on three occasions: 1993 (prior to flooding); 1995 (17 months after flooding began), and in 2003 (approximately 10 years after construction). Surface soils were sampled using a 10-m grid system established for the wetlands after they were constructed. Permanent field markers were established at each grid point that allowed for direct comparisons of soil samples collected between years. At each grid point, a sample was collected at the surface (0–8 cm depth) and subsurface (8–16 cm depth). A total of 102 wetland soil samples were collected and analyzed in 1993 and 1995 (Nairn and Mitsch, 1996), and a total of 166 in 2003. Soil color was determined at multiple sites in each wetland basin by comparing soil color to a standard Munsell color
Fig. 4. Pattern of pumped inflow to the planted (Wetland 1) and naturally colonizing (Wetland 2) experimental wetlands, 1994–2003.
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chart and the percentage of samples ≤2 was determined. Percent organic matter was estimated for each sample through loss on ignition at 550 ◦ C for 1 h. 2.5. Water quality Sampling of water temperature, dissolved oxygen, pH, conductivity, and redox have been measured twiceper-day (dawn and dusk) for 10 years with water quality sondes at the inflow of the wetlands and outflows of both wetland basins. One-hundred-milliliter samples are also taken dawn and dusk each day at the inflow and two outflows for turbidity analyses in the laboratory with a Hach ratio turbidimeter. In addition to the twice-per-day manual sampling, weekly water samples are taken at inflow, middle, and outflows of the wetlands for nutrients (SRP, TP, NO3 ) that are determined by standard methods. Samples are split into filtered (0.45 m) and unfiltered samples, frozen until analysis, and analyzed for total phosphorus, soluble reactive phosphorus, and nitrate–nitrogen (USEPA, 1983; APHA, 1989). Both total phosphorus and soluble reactive phosphorus methods employ the ascorbic acid and a molybdate color reagent method with a Lachat QuikChem IV automated system. Total phosphorus samples are first digested by adding 0.5 ml of 5.6 N H2 SO4 and 0.2 g (NH4 )2 S2 O8 to 25 ml of sample and exposing the samples to a heated and pressurized environment for 20 min in an autoclave. Nitrate + nitrite were analyzed on a Lachat QuikChem IV automated system with the cadmium reduction method. Samples from April 1994 through July 1995 were run by similar methods by Heidelberg College Water Quality Laboratory using a Traacs 800 autoanalyzer. The accuracy of the nutrient analysis was checked every 10–20 samples with a known standard, and the samples are redone if the accuracy is off by 5%. 2.6. Hydrology Water stages and inflow rates are manually read from staff gauges and flow meters twice per day. Outflow is estimated from water stages. Water budgets for the wetland basins have been calculated for the past 10 years and published these results each year in our annual reports.
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2.7. Gas fluxes CH4 fluxes were quantified twice per month during a year of pulsing flow from March to September, 2004, in six 0.26-m2 plots in each wetland basin; half were located in the edge zones and contained emergent vegetation, and half were located in permanentlyinundated areas. Portable, non-steady-state chambers (Livingston and Hutchinson, 1995) were employed for gas flux measurements. Permanent PVC frames and HDPE bases were installed at edge-zone sampling locations; 4-mil, fitted polyethylene bags affixed with sampling ports were attached to the frames at the time of sampling. Permanently inundated areas were sampled with floating chambers made from the same materials. Sampling was conducted in the morning, afternoon, and after dark on each sampling date. Sampling took place during flood pulses, and at the end of the lowflow period before the subsequent pulse. Each flux rate was determined from three to five samples collected from each chamber at regular intervals over 30–60 min. Samples were injected into evacuated 10-ml autosampler vials and stored at 4 ◦ C until analysis, which was completed within 1 week. Gas samples were analyzed with a Shimadzu GC 14-A flame ionization detector (150 ◦ C) using a 40 position HTA Autosampler, a 1.8 m Porapak-Q column, and helium as the carrier gas (25 ml min−1 ). Flux rates were determined by the linear change in concentration by mass of CH4 –C over time (Holland et al., 1999). N2 O fluxes were measured using a closed chamber (Plexiglas chamber 24.5 cm × 24.5 cm × 70 cm) technique (Smith et al., 1983) from June 2003 through December 2004. Nine plots were sampled in each wetland near the inflow, middle, and outflow, with three plots in a transitional upland zone, three plots in the edges with alternate wet and dry conditions, and three plots in the marsh, which were regularly flooded. During gas collection the chambers were sealed for 2 h using a water seal, gas samples were withdrawn every 30 min. Samples were taken weekly, three times a month, when pulsing events occurred and once a month when no pulses were performed. All samples were taken between 11:00 a.m. and 3:30 p.m. for uniformity within the diurnal period. Nitrous oxide gas samples were analyzed using a gas chromatograph (Shimadzu GC-14-A) fitted with a 2 ml sampling loop, two Porapak-Q 1.8 m columns and an electron capture
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Ni-63 detector. The instrument used ultra pure nitrogen carrier gas (10 ml min−1 ) and operated at temperatures of 70, 80, and 300 ◦ C for column, injector, and detector, respectively. Nitrous oxide fluxes were calculated using a closed chamber flux equation (Holland et al., 1999).
3. Results and discussion 3.1. Macrophyte cover Fig. 5 presents dominant vegetation community patterns from 1994 through 2003. Patterns of dominance and subsidence of macrophyte communities is summarized in Fig. 6. By the end of the 10th growing season in 2003, Wetlands 1 (W1) and Wetland 2 (W2) had approximately 62 and 38% macrophyte cover respectfully. These values were considerably lower than the 73 and 74% cover in the two wetlands in 2002—the highest percent coverage in the basins since they were created in 1994. The decrease is probably due to the spring pulsing of water through both wetlands and some herbivory. From 1994, when there was no significant macrophyte vegetation cover in either basin, macrophyte coverage increased yearly in both basins through 1999. Wetland 1 had a greater percent coverage than Wetland 2 until 1997, when that trend reversed. Coverage decreased in both Wetlands 1 and 2, from 2000 to 2001, probably as a result of a combination of muskrat herbivory and possibly increased water levels due to outflow sedimentation. The overall pattern of vegetation can be summarized in several distinct periods. Wetland 1 was planted in 1994 and a distinct pattern of vegetation development around the edge of the wetland was observed in 1995 as a result while the “unplanted” wetland remained relatively free of macrophytes except for an edge of Populus seedlings beginning on the interior mudflat. By the third year, soft-stem bulrush Schoenoplectus tabernaemontani (a.k.a. Scirpus validus) had colonized in the unplanted wetland, and by the end of the third growing season, it appeared that the planted and unplanted wetlands had converged with a domination of cover by Schoenoplectus. Typha spp. dominance increased dramatically in the naturally colonizing Wetland 2 from 1996 to 1999. It generally has remained less than 17% or less cover in the planted Wetland 1. By 1999, the naturally coloniz-
ing wetland (W2) was totally dominated by a very productive cover of Typha while the planted wetland (W1) had a diversity of communities including ones dominated by four communities: Sparganium eurycarpum, S. tabernaemontani, Typha spp., and Scirpus fluviatilis. Wetland macrophyte coverage began to significantly erode in 2000 and, by 2001, the wetlands had only 27.6 and 17.4% macrophyte cover in W1 and W2, respectively. This macrophyte vegetation loss was caused primarily by a peak in muskrat activity that followed the year of maximum plant coverage in 1999. We believe that after the muskrat activity peak, as measured by winter lodge counts, the absence of vegetation cover for food, protection, and lodge construction caused the muskrat population to drop significantly. Predation undoubtedly was one reason but movement of muskrats to the river and other wetlands is also likely. A significant drawdown of both basins was done in spring and early summer 2002 to allow the seed bank to reset. The approach was successful. At the end of the 2002 growing season, vegetation coverage was the highest it had ever been (73–74% cover in the basins) and Typha coverage was only 9% of the total area of Wetland 2 and 5% of the total area of Wetland 1. This was considerably reduced from peak year 1999 when it was 56% of the cover in Wetland 2. One of the most significant changes in 2002 was the increased coverage by S. tabernaemontani in both wetlands, apparently from the marsh seedbank. S. tabernaemontani dominance increased in both basins in 2002 from 0.3 to 52% coverage in W1 and from 0 to 63% in W2. In January 2003, we started a pulsing experiment in W1 and W2 where several 7-day duration floods were pulsed through the wetlands, mostly in late winter and spring. This pulsing was one of the apparent reasons for the reduction in macrophyte cover in the wetland basins in that year. 3.2. Macrophyte community diversity We developed a macrophyte community diversity index (CDI) to quantify spatial diversity in these experimental wetlands (Mitsch et al., 2005a). The index is expressed as: CDI =
N i=1
(Ci ln(Ci ))
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Fig. 5. Vegetation community cover in the planted (Wetland 1) and naturally colonizing (Wetland 2) experimental wetlands at Olentangy River Wetland Research Park, 1994–2003.
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CDI increased in W1 as a good balance among five communities developed. Four of those communities were dominated by plants introduced in the planting. Diversity dropped in 2000 and 2001 during the years of extensive herbivory but dramatically increased again in 2002 and 2003 (years 9 and 10) when the wetland vegetation reset. 3.3. Macrophyte productivity Fig. 6. Typha and other macrophyte community cover in the planted (Wetland 1) and naturally colonizing (Wetland 2) experimental wetlands at Olentangy River Wetland Research Park, 1994–2003. Episodes of high muskrat herbivory in 2000 and 2001 are also indicated.
where Ci is the percent cover of community “i” (0–1) and N is the number of plant/aquatic communities. Overall, there were 13 different communities identified from ground surveys supported by aerial photography during the study, based on the dominant species in communities (see Fig. 5). When the wetlands are viewed in terms of our community diversity index, which includes evenness of plant cover as well as number of dominant communities, an interesting pattern of macrophyte community diversity in the two wetlands is seen (Fig. 7). The data show a general pattern of higher diversity in the planted wetland with the exception of years 3–5 when community diversity was similar in the two basins. By the sixth year (1999), after the 3 years of similar plant cover in the two basins, different spatial community diversity developed. The CDI in W2 dropped as Typha formed close to a monoculture; the
Fig. 7. Macrophyte diversity in the planted (Wetland 1) and naturally colonizing (Wetland 2) experimental wetlands at Olentangy River Wetland Research Park, 1994–2003. CDI stands for community diversity index, which is defined in the text.
Macrophyte net aboveground primary productivity (NAPP), as estimated by peak aboveground biomass in August each year, decreased in both wetlands in 2003 compared to similar plot readings in 2002 (Fig. 8a). When paired sites were compared between the two wetlands (16 sites were paired in 2003) macrophyte aboveground biomass was statistically higher in the planted Wetland 1 than in the naturally colonized Wet-
Fig. 8. Indicators of macrophyte productivity in the planted (Wetland 1) and naturally colonizing (Wetland 2) experimental wetlands, 1997–2003: (a) peak aboveground biomass in vegetation zones and (b) cumulative organic matter production, kilograms per year in each experimental wetland basin.
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land 2 (t = 0.0002; α = 0.05) for the first time in 7 years. This is because of a significant decrease in macrophyte cover in the southern half of Wetland 2 in 2003 due, we believe, to a combination of higher herbivory and effects of temporary high water levels caused by pulsing. Before that, peak aboveground biomass was significantly higher in plots in Wetland 2 than in Wetland 1 for 4 straight years from 1998 to 2001. Based on the aboveground biomass and estimates of macrophyte cover presented above, aboveground productivity by macrophytes in each wetland basin was estimated for 1997–2003 (Fig. 8b). Overall macrophyte organic productivity decreased 46% in Wetland 1 and 81% in Wetland 2 from 2002 to 2003. These numbers are significant for two reasons. First, although the macrophyte community substantially recovered in 2002 from the herbivory and subsequent macrophyte losses of 2000 and 2001, productivity once more decreased in 2003, due to herbivory in Wetland 2 and possibly due to high water levels that were a result of a pulsing experiment in both basins in spring 2003. Second, 2003 is the second year in a row where the planted Wetland 1 had a higher estimated macrophyte carbon sequestration than the naturally colonized Wetland 2. The total organic matter production by macrophytes over the last 7-year period is now almost the same in the two wetland basins at 22–23 tonnes per basin. 3.4. Soil development Soil changes at the wetland surface have been substantial since the addition of water to the basins in 1994 (Fig. 9). Most changes have occurred at the surface (0–8 cm depth) because of the prevailing influence of sedimentation and organic accumulation. Soil organic matter (SOM) increased by 63% in the upper 8 cm of wetland soils over the decade to an average of 8.6%. On average, surface organic matter content increased by approximately percentage point each 3 years. In contrast, Craft et al. (2003) found an increase in the amount of carbon per unit area in created salt marshes after 10 years, but to levels 18% of those found in natural salt marshes. Color in the soils changed dramatically after 10 years of inundation (Figs. 9b and 10) with 94% of the samples from both 0–8 cm depth and 8–16 cm depth having soil chromas of 2 or less in 2003. Before the wetlands were created in 1993, no soil samples indicated soil chroma of 2 or less. In 1995, 17 months after
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Fig. 9. Experimental wetlands soil data pooled for both experimental wetlands: (a) percent organic matter in 1993, 1995, and 2003, and (b) percent of soil samples that exhibited chroma values of 2 or less in 1993, 1995, and 2003. Soil data indicated as 2003 were collected between December 2002 and March 2003.
water was added to the wetlands, 78% of the samples taken in the surface layer had soil chromas of 2 or less while only 24% in the 8–16 cm depth had chromas of 2 or less. Five years after the creation of a freshwater marsh, Vepraskas et al. (1999) found soil matrices depleted of iron and low chromas. 3.5. Water quality Patterns of water quality changes through both wetlands have been pronounced and variable over the 10 years of wetland development, reflecting changes from characteristics of eutrophic shallow ponds in the early years (Wu and Mitsch, 1998) to functioning wetlands in latter years. The changes in water quality are presented as “percent change” in all subsequent graphics. Even though the results are presented as only one number per
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Fig. 10. General pattern of surface soil color change, as indicated on a Munsell Color Chart, averaged for both experimental wetlands at Olentangy River Wetland Research Park from 1993 (prior to wetland basin flooding) to 2002. Note that the soil chroma decreased to values of 2 or less after 1996, or 2 years after the basins were flooded.
wetland for each year, statistical differences have been determined between the wetlands by paired tests of the outflow concentrations. Water temperature increases through the wetlands (Fig. 11a) showed a consistent pattern of decreasing (there was less net warming from inflow to outflow each year) over the first 6 years in each wetland as vegetation developed to maximum cover by 1999. Then muskrat herbivory removed macrophyte cover in 2000 and especially in 2001 (see Fig. 5), exposing the water surface to full sunlight that resulted in a net effect of warmer temperatures. The wetland vegetation recovered dramatically in 2002 and temperature patterns once again were less dramatic. That pattern reversed in 2003 when our hydrologic pulsing experiment decreased vegetation cover but also changed the basin retention rates, causing higher temperature increases again. Dissolved oxygen change through the wetlands (Fig. 11b) showed a fairly consistent pattern over the 10 years of decreasing in importance, probably due to
the continued decrease in the importance of aquatic primary productivity. The site was an experimental farm for decades prior to the construction of the wetlands and initial algal blooms were long in duration and intense (Wu and Mitsch, 1998). Gradually, high concentrations of nutrients have been removed from the wetland soils, resulting in lower aquatic primary productivity. The low change in dissolved oxygen persisted in the years 1999–2003 despite the fact that macrophyte cover was lower in most of those years. In 2003, the flood pulsing may have decreased water column productivity in the basins. Patterns of pH change through the wetland basins (Fig. 11c) mirrored those seen for dissolved oxygen. Early in the study it was common to see dramatic increases in pH in the late afternoon due to high productivity in the water column. In latter years, that was no longer the case. Conductivity (Fig. 11d) decreased dramatically through the wetlands in the second and third years when water column productivity was the
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Fig. 11. Percent changes from inflow to outflow of water quality parameters in the planted (Wetland 1) and naturally colonizing (Wetland 2) experimental wetlands at the Olentangy River Wetland Research Park, 1994–2003 for: (a) water temperature, (b) dissolved oxygen, (c) pH, (d) conductivity, (e) turbidity, and (f) redox potential.
highest. We documented significant precipitation of calcium carbonate in those years (Liptak, 2000). Conductivity decreased again in 2003 (the pulsing year) by almost 12% in both basins. Suspended sediments, as reflected in turbidity measurements taken over the 10 years of wetland function, showed an interesting pattern of ecosystem succession (Fig. 11e). With each succeeding year, the wetlands appear to be removing fewer suspended sediments. By the ninth year, both wetlands were exporting more sediments than were coming in the wetland. Redox potential (Fig. 11f) shows an oscillating 10-year pattern that made little ecological sense. The year with
the hydrologic pulses (2003) led to one of the highest drops in redox potential through the wetlands since measurements began. There could be related to the fact that when the wetland “expands in size” with the pulses, the surface water comes in contact with more anaerobic soil waters on the edges of the wetlands that are otherwise isolated from the wetlands. 3.6. Nutrient retention These created wetlands have been fairly significant in removing nutrients over the past decade. Patterns of NO3 –N reduction through the wetlands have averaged
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Fig. 12. Percent changes from inflow to outflow in the planted (Wetland 1) and naturally colonizing (Wetland 2) experimental wetlands at the Olentangy River Wetland Research Park, 1994–2003 for: (a) nitrate + nitrite-nitrogen; (b) soluble reactive phosphorus (SRP); (c) total phosphorus (TP).
close to 35% by concentration and mass (Mitsch et al., 2005b; Fig. 12a). Nitrate removal patterns in the naturally colonizing wetland have been erratic, ranging from removal rates in the 20% range 1 year, and high 40% range the next. The planted and more diverse wetland (WI) has removal rates more consistent from year to year. The spring pulsing in 2003 did not appear to reduce nitrate–nitrogen removal. In fact, the average nitrate removal by the wetlands was the highest in the two wetlands (42% reduction) as it has been since the beginning of the study a decade earlier. The retention of soluble reactive phosphorus (SRP) in the wetland basins (Fig. 12b) continued to be significant at around 70% through the 10th year 2003 (no nutrient data are reported for 2002). In the 10th year after wetland construction and during the pulsing year, the planted wetland retained 63% of SRP while the naturally colonizing wetland retained 73% (not significantly different, α = 0.05). Total phosphorus (TP) showed a much different picture with little retention and even export by the
10th year (Fig. 12c). TP decreased by only 2% in W2 and increased by 14% in W1 after a consistent pattern of total phosphorus retention during the first 8 years since wetland construction. This pattern is consistent with the pattern seen for turbidity in the 2003 pulsing year where there was either a very small reduction or slight increase in turbidity in that year. The sediments that are discharging from the wetlands have significant phosphorus bound to the soil particles. 3.7. Pulsing effects 3.7.1. Macrophyte productivity In 2003, flood pulses of 7-day durations and inflow of approximately 500 gpm were pumped into each of the two experimental wetlands. Preliminary results of the 2003 flood pulses are described here. Macrophyte vegetation NAPP appeared to be lower during the pulsing in 2003 compared to the previous year (Fig. 8a). High water levels in the spring caused by the pulses may
W.J. Mitsch et al. / Ecological Engineering 25 (2005) 510–527 1.0 (102)* 0.52 (92)* 23 (109)* 0.05 (94)* 9 (113)* 3 (57) ± ± ± ± ± ± 13.6 11.42 535 8.23 252 23 1.0 0.55 (89) 21 (103)* 0.07 (87)* 10 (102) 3 (61)*
± ± ± ± ± ± 12.2 11.94 723 8.17 246 15 1.8 0.73 (25)* 54 (32) 0.07 (28)* 15 (32) 4 (22) Significant different at α = 0.05 between inflow vs. outflow. *
± ± ± ± ± ± 11.9 11.96 729 8.15 264 19 (◦ C)
Temperature DO (mg/l) Conductivity (/cm) pH Redox (mV) Turbidity (NTU)
10.3 10.79 726 7.97 261 24
± ± ± ± ± ±
1.5 (31) 0.57 (24) 51 (32) 0.06 (27) 10 (31) 5 (22)
Outflow 1
(31)*
Outflow 2
(31)*
1.8 1.02 (25) 50 (32) 0.09 (28)* 13 (31) 2 (22)
11.5 10.69 650 7.93 281 29
± ± ± ± ± ± Inflow Inflow
0.8 (94) 0.38 (80) 23 (101) 0.06 (86) 9 (105) 4 (63)
13.2 10.75 530 8.25 257 23
± ± ± ± ± ±
(95)* Outflow 1 Non-pulsing Pulsing Parameters
3.7.3. Gas exchange Studies are underway to evaluate the effects of hydrologic pulsing on gas exchange from the wetlands, particularly for methane and nitrous oxide. Preliminary results of the effects of flood pulsing on methane flux in the experimental wetlands (Fig. 13) show that methane emission in vegetated areas increased significantly when flooding occurred in zones that previously did not have standing water. However, preliminary results also suggest that overall “average” methane emission from pulsed sites was less than that from those sites where flooding was permanent. A seasonal pattern in N2 O emissions was observed during the study period with highest rates in summer. In summer 2003 and winter 2004, N2 O emissions were higher when no pulses occurred than when flood pulsing occurred. In spring 2004, N2 O fluxes were higher during pulsing than during non-pulsing conditions.
Table 1 Comparison of two-per-day water quality samples in the two experimental wetlands during flood pulsing and non-pulsing periods in 2003
3.7.2. Water quality Some preliminary effects of the flood pulses or artificial “floods” in 2003 on water quality in the two experimental wetlands are shown in Tables 1 and 2. Water quality changes during the pulsing events of 2003 are immediate. When the two wetlands are investigated with frequent two-per-day sampling, water temperature and pH significantly increased in both basin outflows during flood pulses (Table 1, α = 0.05). Between flood pulses (low-flow conditions), temperature and pH were again higher and conductivity was lower in both wetland outflows compared to the inflows (Table 1). When data are investigated with both wetland basins combined at the outflow swale (Table 2), there is a significant difference between pulsed and non-pulsed conditions for conductivity and soluble reactive phosphorus. Both conductivity and soluble reactive phosphorus decreased more rapidly during non-pulsing conditions than during pulses.
Outflow 2
have had a detrimental effect on emergent macrophyte productivity but the lower productivity could have also been due to other causes. The pattern was particularly acute in W2 where macrophytes disappeared from the southern half of the basin in 2003 (Fig. 5). Higher community diversity in W1 may have insulated it from this stress caused by the flooding whereas the lower diversity in W2 made it more susceptible to the combined stress of flooding and grazing.
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Table 2 Effects of flood pulsing on water quality in the experimental wetlands in 2003 Inflow
Swale outflow
Percent change
Temperature (◦ C) Pulse 12.4 ± 2.8 (11) Non-pulse 15.6 ± 1.8 (16)
13.6 ± 2.6 (13) 16.6 ± 1.9 (16)
+9.7 +6.4
Dissolved oxygen (mg/l) Pulse 7.47 ± 1.33 (10) Non-pulse 9.29 ± 0.70 (14)
6.98 ± 1.37 (11) 8.66 ± 0.84 (16)
+6.6 +6.8
Conductivity (/cm) Pulse 705 ± 76 (11) Non-pulse 627 ± 33 (16)
640 ± 82 (13) 556 ± 37 (18)
−9.2* −11.3*
pH Pulse Non-pulse
8.28 ± 0.23 (7) 7.80 ± 0.14 (15)
8.28 ± 0.15 (7) 7.88 ± 0.08 (16)
0 +1.0
Redox (mV) Pulse Non-pulse
237 ± 28 (11) 263 ± 26 (16)
292 ± 35 (13) 230 ± 20 (18)
+23.2 −12.5
2.88 ± 0.38 (15) 2.01 ± 0.44 (26)
−31 −55
Nitrate–nitrogen (mg N/l) Pulse 4.17 ± 0.50 (15) Non-pulse 4.48 ± 0.51 (18) SRP (g P/l) Pulse Non-pulse
27 ± 3 (3) 46 ± 26 (3)
15 ± 6 (2) 5 ± 1 (6)
−44* −89*
Total P (g P/l) Pulse 54 ± 12 (3) Non-pulse 103 ± 38 (3)
47 ± 2 (3) 44 ± 9 (5)
−13 −57
Data for SRP and total P are for July–September and cover only one pulse in early August. Swale outflow is combined outflow from two experimental wetlands. * Significant difference (α = 0.05) in swale concentrations during pulsing and non-pulsing.
Fig. 13. Average (±S.E.) of methane flux for plots pooled for both experimental wetlands containing emergent vegetation, when inundated vs. when they were exposed, in 2004 growing season (April 25–September 29, 2004). Methane flux from permanently-inundated areas is shown for comparison.
had low productivity. Seven years of productivity data for each of the two wetlands (14 wetland-years) provides results that are not as obvious (Fig. 15). Data from the planted wetland W1, which has never been dominated by one species, appears to have a pattern of highest productivity at a mid-level of community diversity; inclusion of productivity and diversity data from the Typha-dominated wetland W2 does not fit the same pattern.
Overall, when marsh and edge plots were inundated, nitrous oxide fluxes were lower than when the same plots were exposed to air (Fig. 14). In the upland plots, nitrous oxide fluxes were not different between flooded and dry periods. 3.8. Wetland productivity and diversity Our 10 years of whole-ecosystem measurements give us the opportunity to investigate the connections between diversity and productivity. Previously, we showed that there appeared to be a negative relationship between the two (Mitsch et al., 2005a). Monospecific wetlands were highly productive and diverse wetlands
Fig. 14. Average (±S.E.) of N2 O flux pooled for both experimental wetlands during flooding and non-flooding periods from June 2003 through December 2004. Marsh refers to areas normally flooded; edge refers to areas that are flooded seasonally with flood pulses; T-upl refers to uplands that are rarely flooded.
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Fig. 15. Net aboveground primary productivity, estimated from aboveground biomass, as a function of community diversity index (CDI) for experimental wetlands, 1997–2003.
4. Conclusions After 10 years of observing these experimental wetlands, the following conclusions can be drawn regarding the influence of planting of ecosystem succession and on the general development of a wetland when it is created on formerly non-wetland soil: • Contrary to conventional wisdom, wetlands can be created if the correct hydrologic conditions are available. Created wetlands with the proper hydrology can develop appropriate biota (e.g., wetland plants) and physiochemistry (e.g., hydric soils) relatively rapidly without the need for planting if the proper hydrologic conditions are present and plant propagules are continually introduced as they were in our riverine wetlands. • Planting has a profound effect on ecosystem function of created wetlands, even several years after planting. The naturally colonizing wetland and the planted wetland diverged in macrophyte cover that affected ecosystem function on several occasions during the 10 years of this study. Differences in plant compostion and productivity, caused in all probability by the introduction of plants to one of the wetlands, led to differences in water quality and carbon accumulation in the wetlands. • Hydric soils as valid indicators of natural wetlands can develop within 2–3 years of wetland creation. We found that soil organic matter can increase by 1% every 3 years in temperate zone created marshes.
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• Water quality changes significantly as a wetland develops in primary succession and as perturbations occur. Some changes in water quality are direct and the immediate result of macrophyte cover and aquatic metabolism; other changes occur over longer periods due to sediment accumulation, and soil and redox changes. • Wetlands, if they are not overloaded with nutrients, can be effective nutrient sinks for many years. Our systems were subjected to nutrient concentrations and loadings that were significantly less than those usually added to treatment wetlands. Thus, our wetlands continued to be nutrient sinks for 10 years and support a reasonable plant diversity as well. • The addition of species to enhance biodiversity in wetland creation actually can lead to lower productivity of macrophytes and subsequent changes in the food web and water quality. • There are desirable values from both “diverse” marshes and “high productivity” marshes and to make generalizations that one is better than the other is wrong. • The continual introduction of plant, animal, and microbial species through water flows, atmospheric transport, and biological vectors give self-design a significant opportunity to manifest itself. • Flood pulses have significant influences on water and gaseous fluxes from wetlands. Continued study of the importance of periodic floods is needed on the combined effects on water and gas fluxes in these and other wetlands.
Acknowledgements We appreciate the many post-docs, graduate students, undergraduate students, and staff who assisted in this research over the 11 year period described here. Particular appreciation is given to Xinyuan Wu and Naiming Wang for their extraordinary assistance in the 1990s. Olentangy River Wetland Research Park paper 05-010.
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