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Spieles, D. Wetland Construction, Restoration, and Integration. Encyclopedia. Available online: https://encyclopedia.pub/entry/21731 (accessed on 18 May 2024).
Spieles D. Wetland Construction, Restoration, and Integration. Encyclopedia. Available at: https://encyclopedia.pub/entry/21731. Accessed May 18, 2024.
Spieles, Douglas. "Wetland Construction, Restoration, and Integration" Encyclopedia, https://encyclopedia.pub/entry/21731 (accessed May 18, 2024).
Spieles, D. (2022, April 13). Wetland Construction, Restoration, and Integration. In Encyclopedia. https://encyclopedia.pub/entry/21731
Spieles, Douglas. "Wetland Construction, Restoration, and Integration." Encyclopedia. Web. 13 April, 2022.
Wetland Construction, Restoration, and Integration
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This entry is adapted from the peer-reviewed paper 10.3390/land11040554

In response to the global loss and degradation of wetland ecosystems, extensive efforts have been made to reestablish wetland habitat and function in landscapes where they once existed. The reintroduction of wetland ecosystem services has largely occurred in two categories: constructed wetlands (CW) for wastewater treatment, and restored wetlands (RW) for the renewal or creation of multiple ecosystem services. Where the spatial and biogeochemical contexts are favorable, Integrated Constructed Wetlands (ICW) present the opportunity to couple CW and RW functions, thereby enhancing the replacement of wetland services on the landscape. 

treatment wetlands ecological restoration socioecological systems coupled ecosystems integrated landscape approach

1. Introduction

Wetlands provide ecosystem services to a degree that is proportionately greater than their geographic extent [1]. Even so, wetlands are among people's most endangered ecosystems, having been drained, filled, diked, flooded, and converted to other land uses with impunity for much of modern history [2][3]. Socioecological systems around the world have suffered from wetland loss and degradation, as manifested in deteriorating fisheries, reduced water quality, loss of coastal storm abatement, biodiversity decline, increased flood intensity and frequency, aquifer depletion, and reduced carbon storage [4]. The remaining wetlands have consequently achieved a heightened degree of protection, and efforts around the world have attempted to replace some of the wetlands people have lost.
Wetland replacement generally occurs in two broad categories, which are the focus of this entry. The first is wetland construction. Constructed wetlands (CW) are typically designed to treat a particular wastewater stream (see [5][6][7] for representative designs). Often, but not always, CW are artificial ecosystems, in the sense that they are designed and managed primarily for wastewater treatment, and not intended to go through the adaptive cycle of succession [8][9]. CW are commonly engineered to treat particular wastes that are introduced at controlled concentrations, with carefully managed substrate, hydraulic path, retention time, oxidation, and a trophic structure. The second category, restored wetlands (RW), differs in several respects. RW are intended to reestablish multiple ecological functions on the landscape (see [10][11][12] for representative designs). In the broadest sense of the term, the restoration of wetlands may involve the rejuvenation of a wetland where it once existed, creation of new wetland habitat, or enhancement of a wetland that exists in a degraded state. The goals of wetland restoration vary by project, but they are typically not focused solely on water quality improvement. For example, habitat provision, flood water retention, aquifer recharge, carbon sequestration, and cultural services are all common desirable outcomes in RW [13]. In special cases, there are also political and economic goals, as wetlands are built for the compensatory mitigation of wetland losses [14][15]. Beyond their goals and objectives, RW differ from CW in that they are intended as ecosystems that self-organize, respond to disturbance, and change through succession [16].

2. Constructed Wetlands Performance and Management

Physical, chemical, and biological treatment mechanisms in CW vary by design, and thus CW conformations are specific to contaminant treatment objectives. Essentially, CW effectiveness relies on the interactions of the waste products with the vegetation, substrate, microbes, and water column of the CW [17]. Physical waste removal processes include flocculation, precipitation, sedimentation, and filtration, and are thus reliant on water column interactions with the substrate [6]. The substrate is also the primary locus of chemical removal processes, including interactions with ions and adsorbent surfaces, ion exchange, and redox processes. For this reason, substrate has been a focus of CW research, testing the relative efficacy of various adsorbent substrate materials, including alum sludge, limestone, coal slag, sand, rice husks, biochar, and many others [17][18]. The different physicochemical characteristics of substrate materials can result in a range of removal efficiencies for different contaminants; according to Patyal et al., substrate materials varied removal efficiency for oxygen-demanding wastes from 71.8–82%, for total phosphorus from 77–80%, and total nitrogen from 52–82% [17].
Biological processes, including microbial metabolism, phytoremediation, biosorption, and predation, also occur in the substrate, as well as in the water column [6]. Macrophyte stems, roots, and leaves reduce the flow rate and provide additional surface area for microbial biofilms. Macrophytes also exude carbon compounds and oxygen, particularly in the rhizosphere, stimulating chelation and aerobic microbial metabolism [19]. Macrophyte metabolism contributes to waste retention and removal through phytoaccumulation, phytodegradation, volatilization, and sequestration [20][21].
The physicochemical characteristics of CW are subject to careful management. For instance, hydraulic parameters are critical aspects of CW design [22]. The expected hydraulic loading rate (HLR) and contaminant concentration determine the necessary surface area and volume of the CW [23]. Equally critical is the frequency and duration of flooding and drawdown events, as these regulate the oxygen availability in the water column and substrate. The CW design, HLR, and hydrologic regime determine the hydraulic retention time (HRT), the average length of time that influent wastewater remains in the CW [24]. Generally, a longer HRT results in higher percentage of contaminant removal for a given system [25]. For example, Toet et al. [26] found a significant increase in retention of N and coliform bacteria as HRT was increased from 0.3 to 9.3 days in a FWS municipal wastewater treatment wetland. In a HF CW system, Ghosh and Gopal [25] similarly report a greater retention of nitrogen (83–100%) with an HRT of 4 days over an HRT of 1 day (21–77%). Many CW designs use long, sinuous paths, baffles, or multiple cells to lengthen the HRT and maximize the system’s efficiency [27].
Other factors can confound CW performance, however; season, temperature, pH, oxygen availability, changes in biomass productivity, bioturbation, and weather events can all influence the efficacy of contaminant removal [21]. These, too, can be managed to an extent. In cold-weather climates, temperatures may be increased with greenhouses [21] and bed heating [28]. Artificial aeration, pH buffering, insulation, and bio-augmentation are all used to maintain critical parameters to achieve the most efficient waste removal [21][29]. Vymazal et al. [18] and Ingrao et al. [30] review emerging challenges in new wastewater streams and novel pollutants amid the tighter regulation of effluent contaminant concentrations.
CW biota are also managed. The ideal macrophyte species for CW systems vary by geography, design, and waste stream, but generally the plants must be robust, with quick establishment, rapid growth, large biomass, and tolerance to the stressful conditions that wastewater presents [31]. While plant species selection is deemed an important part of CW design, the advantages or disadvantages of one macrophyte species over another are inconsistent in the literature [31][32]. Vymazal [32] surveyed over 640 studies on FWS systems that introduced a total of 150 macrophyte species. The most used species are Typha latifolia, Phragmites australis, Typha angustifolia, Juncus effusus, Scirpus lacustris, Scirpus californicus and Phalaris arundinacea. Others [33] have experimented with woody species in FWS CW. Subsurface CW often use emergent macrophytes for the enhanced treatment that the presence of a rhizosphere provides. Vyzmal [34] found that macrophytes of the genera Phragmites, Typha, and Scirpus are most used in subsurface systems. Of course, both FWS and subsurface CW provide habitat for volunteer species; Knight [35] notes that more than 600 plant species occur in CW systems in the US.
Nuisance animals, the subject of removal or control in many CW (particularly FWS) systems, include burrowing mammals (e.g., muskrat, beaver), mosquitos, bioturbators, and aggressive grazers [6][36]. However, wetland animals can also have a positive effect on CW performance. Li et al. [36] review an extensive list of invertebrates, fish, birds, reptiles, and mammals that can enhance contaminant retention through adsorption and bioaccumulation, by increasing the diversity of the microbial community, consuming pathogens, and stimulating plant growth.

3. Restored Wetlands Performance and Management

The choice of revegetation strategy is part of a larger philosophical debate on the degree to which anthropogenic management should be used to establish and maintain RW [11]. One approach, which has been called “self-design”, emphasizes the role of the biogeochemical and hydrological environment in determining how plant communities will develop in the RW [37][38]. In the extreme view, self-design can be implemented as passive management, relying solely on natural revegetation, successional development, and the capacity of ecological communities to self-assemble. However, self-design does not preclude plant introduction or vegetation management as a means of increasing the rate of development [39]. At the heart of the self-design approach is the idea that “the system itself will optimize its design by selecting for the assemblage of plants, microbes, and animals that is best adapted for existing conditions” [38]. An alternative perspective, sometimes called the “designer” approach to restoration [11], calls for attention to the life history strategies of target species in the restoration project. Through the careful introduction (and reintroduction) of plants and propagules, ecological engineering, and management of biotic and abiotic factors early in the restoration process, the restorationist can select individual species to be part of the assembled community [40]. The goal is not to achieve a particular species complement as an endpoint, but rather to encourage the establishment of desirable species that would otherwise be overwhelmed by undesirable invasive species [40]. As Middleton [11] notes, these two views are not mutually exclusive, but they do conceptualize ecosystem development in different ways. Ultimately, the degree of human intervention in a RW may depend upon the goals and performance standards of the project.
Many RW projects use one or more reference wetlands, or historic reference conditions, to evaluate success [41][42]. A reference wetland ostensibly represents the desirable structure and function of similar ecosystems in the region of interest. The idea has emerged from the Leitbild concept of river restoration in Germany and Austria; Leitbild refers to the ideal or undisturbed state of an ecosystem [43]. The identification and measurement of one or more such benchmarks can be used to set the parameters and expectations for the restoration work. Middleton [11] notes that, while the reference ecosystem approach can be useful in RW, the goal of matching RW conditions to reference conditions may be unachievable. This is because the RW and reference sites may have different land use legacies and different site impairments—leading to irreconcilable differences in ecological development—or changes in regional biota and climate that could make historical reference conditions untenable [11]. Moorhead [42] adds that the reference ecosystem concept can be problematic, given the differences in successional state (reference systems typically being mature, while restored systems are early successional), and the inherent variability that exists even among potential reference sites. According to Moorhead, RW goals should not be to duplicate the conditions of reference sites, but rather to establish “self-supporting and self-maintaining” ecosystems. The maturation of these systems, Moorhead suggests, is better measured with the general structural and functional attributes of early and late-stage ecosystems than with specific points of comparison to a reference ecosystem [42].
Even so, many RW—particularly mitigation wetlands—are undertaken with specific performance indicator goals. Targets for RW evaluation can be placed in several categories [44]. Structural metrics consider physical aspects of the wetland environment, such as the vegetated percentage, degree of habitat interspersion, or macrophyte biomass. Structural metrics can also be used to quantify landscape-level patterns of land use and connectivity. Measures of taxonomic diversity and evenness, particularly of macrophytes, are common, and often include indices of species quality. For example, the floristic quality index and associated coefficient of conservatism can be used to evaluate vegetation quality, based on species tolerance [45]. Indicator species are also commonly used to denote the presence or absence of particular communities or site conditions. Functional metrics are measures of ecosystem processes, such as nutrient processing, organic matter accumulation, and productivity, as well as process surrogates, such as trophic composition and functional diversity. Finally, taxonomic composition can be used as a direct comparison of species abundance with reference wetland composition [44].
Brudvig et al. [46] suggest that these attribute categories occur with different degrees of variation in RW systems. Structural metrics, according to Brudvig et al., may be established with the greatest degree of certainty. Diversity and function are subject to stochastic events and are more difficult to predict. The metric that is most subject to variation, according to Brudvig et al., is taxonomic composition. Developmental variation thus makes RW performance evaluation a challenge. RW are designed as open systems that are expected to progress through succession. The RW target goals are likely to be characteristics of late-successional reference systems that will emerge, if they emerge at all, many years after the initial RW design and management phases. Indeed, the ultimate goal of a self-maintaining ecosystem may well be incompatible with the rigid metrics of particular functional rates or species lists.

4. Integrated Constructed Wetlands

Thus far, this entry has demonstrated that CW and RW are contrasting and parallel approaches to the re-establishment of wetland services on the landscape. They differ in their objective, design, management, performance objectives, and in the ecosystem services they can provide. CW and RW services are not mutually exclusive; however, CW, especially FWS systems, can provide habitat and other ecosystem services in addition to nutrient retention, just as RW can sequester and transform influent contaminants. In general, though, CW are not designed for the long-term successional development of wetland habitat, nor are RW designed to process highly contaminated waste streams.
Many scholars and practitioners have observed that CW and RW are not only compatible, but synergistic, such that their use in combination may be able to achieve a broader range of ecosystem services on the landscape than either could alone [47][48][49]. CW and RW, in combination, are known as Integrated Constructed Wetlands (ICW); [47][50]. The ICW concept has roots in the whole-ecosystem studies of HT Odum and KC Ewel [51][52], and in the integrated small watershed research of Bormann and Likens [53][54]. Initially developed in Ireland [50], the ICW concept has since been applied elsewhere, but has yet to become the standard for wetland replacement [55]. Harrington et al. [56] present ICW as an ecosystem approach (after the Convention on Biodiversity 2010) to water quality and land-use management on a watershed scale. The ICW attributes suggested by Harrington et al. include: that influent contaminants are maintained below the threshold of macrophyte toxicity prior to entering the system; that the ICW design incorporates multiple sequential wetland cells; that the cells predominantly feature shallow habitat with dense emergent vegetation; and that the entire system be managed for ecosystem services beyond water quality improvement [56].

4.1. ICW Properties and Context

As have been seen, CW and RW both provide important functions to ecologically impoverished landscapes. Over the past five-plus decades, however, CW and RW research and development have rarely informed one another or been combined into holistic efforts to replace wetland services. McInnes et al. [55] speculate that this lack of integration may “stem from narrow disciplinary framing of legacy regulations or a lack of vision by, and appropriate support tools for, planners and managers”. Clearly, there are barriers to integration, including cost, land availability, regulatory requirements, technical difficulties, resistance to change, and disconnection between regulators, ecologists, engineers, land-use planners, and regulators [55][57]. Mitsch [58] suggests that the disconnect is primarily between ecologists and engineers, or more broadly between the disciplines of restoration ecology and ecological engineering. The benefits to CW and RW integration are equally clear. First, the functional capacity of integrated systems has the potential to exceed that of either system alone [59]. Second, integration allows for spatial and temporal heterogeneity, allowing for different functional loci in space and time [60]. Third, integrated systems allow for the coupling of ecological processes, such that the biological, chemical, and physical processes of one ecosystem are linked to processes of other ecosystems within the landscape [61]. Fourth, and encompassing the previous three, landscapes that are designed to integrate wetland functions optimize and maximize ecological services by supporting both human and natural systems [62][63]. These benefits of ICW systems are not merely theoretical; they have been demonstrated in practice.

4.2. ICW Objectives, Design, and Performance

The purpose of ICW systems is to ameliorate the pollutants from a wastewater stream while also providing a long-term, successional habitat for wetland-dependent species and providing provisional, supporting, regulating, and cultural services as appropriate to the setting. The goal is not to provide all ecological services at all times, but rather to couple treatment processes with wetland ecosystem functions in ways that are integrated with the socioecological character of the landscape. Scholz et al. [47] describe ICW objectives as follows: “the explicit integration of (a) the containment and treatment of influents within emergent vegetated areas using (wherever possible) local soil material; (b) the aesthetic placement of the wetland structure into the local landscape with the intention of enhancing the site’s ancillary values; and (c) enhanced habitat diversity and nature management”.
ICW can take different forms. Commonly, a CW is built to receive and treat wastewater by some combination of subsurface and/or free surface flow cells. The CW effluent then flows into an RW before entering a receiving body of water. Scholz et al. [47] describe an ICW in Ireland consisting of sequential FWS cells that intercept agricultural runoff before discharging into a stream. Similar designs are described in the US by Ludwig and Wright [64] and in Sudan by Ladu et al. [65]. Yan et al. [66] use an alternative ICW hybrid design in which domestic wastewater from a septic system enters a subsurface CW before discharging into an RW. Zhang et al. [67] similarly use an intricate series of HSF, VF, and FSW cells to treat domestic wastewater while providing a wetland habitat in China. Boets et al. [68] describe a hybrid system for treating pig manure in Belgium that incorporates eight cells of both FWS and subsurface design. Alternative ICW designs feature RW with floating treatment wetlands, either in the RW itself or in the river, lake, or lagoon water that supplies the RW [69].
The multi-cell designs that are common to many ICW share some characteristics with hybrid CW systems. Indeed, hybrid CW have been shown to support wetland biodiversity and associated ecosystem services [35][70][71][72]. ICW differ from hybrid CW in that they include cells which are: (1) hydrologically connected to ground or surface water other than the wastewater stream; (2) designed primarily to support biological diversity; (3) connected with other ecosystems in the landscape; and (4) open to the adaptive cycle of succession. While the multi-cell design facilitates the multi-faceted nature of ICW, this is not always necessary. ICW that receive pollutants in low concentrations may combine processes in a single basin. For example, Mitsch et al. [38] designed a FSW system for ambient river water that showed both effective improvement in water quality and habitat provision [73][74].
At the broadest conception of performance, ICW systems may be evaluated according to the IUCN Integrated Wetland Assessment Toolkit, developed in 2009 [75]. In this approach, wetlands are evaluated not only for their physical (e.g., hydrological and water quality) and biological (e.g., biodiversity) attributes, but also for their contributions to local livelihoods, to the regional political economy, and to regional socio-ecological systems. Such analyses require a significant investment of time and resources and are seldom incorporated into the evaluation of a single ICW system. Most ICW systems are measured in terms of water-quality improvement, biodiversity, and/or socioeconomic benefits. For example, van Biervliet et al. [76] report a significant nutrient retention, along with a significant increase in avian species richness for an ICW system in the UK. Becerra-Jurado et al. [77] show that ICW support a macroinvertebrate species richness similar to nearby natural aquatic environments; Boets et al. [68] also report macroinvertebrate richness increases in an ICW in Belgium. Harrington and McInnes [78] note significant nutrient retention and numerous cultural, provisioning, supporting, and regulating services provided by ICW systems in Ireland.

5. CW, RW, and ICW: Summary Assessment

Replacing what have been lost, and continue to lose, in terms of wetland services is no simple task, but the last 50 years of research have provided some valuable lessons. Clearly, there is a place for traditional CW, especially in urban environments, where land area is limited and/or the contaminant in question is particularly toxic. CW have proven to be quite effective in ideal conditions; however, their efficacy may decline over time, and the stressful CW environment may limit ecosystem services. Continued research is needed on the best designs for particular contaminants and best practices to maximize the CW lifespan. There is also a place for stand-alone RW, particularly when they are situated where wetlands once existed. Careful design and early successional management have proven to be quite successful in yielding a high-quality and self-sustaining wetland habitat. Stressful landscape matrices can make an invasion trajectory a concern; further research on successional variability and best management practices is needed.
Despite their individual success, CW and RW can offer only partial wetland functions and services to degraded landscapes. This is due to differences in objective, design, and management, but also a result of the ecological attributes of these two types of replacement wetlands. In the parlance of EP Odum, CW share some characteristics with early successional and stressed ecosystems [79][80]. They are dominated by r-type species, with simple trophic and habitat structure. They feature rapid nutrient turnover of predominantly extrabiotic nutrients. Additionally, CW are high-management systems, owing to their relatively low stability and narrow operating parameters. RW share some of these attributes in early succession, but, as they mature, RW ideally become more diverse and trophically complex, with a high degree of spatial heterogeneity. Nutrients in a mature RW are predominantly intrabiotic, and organic detritus play an important role in both nutrient and trophic dynamics [79][80]. As a combination of early and late successional systems—or high- and low-stress systems—ICW have the potential to offer a synergistic relationship between CW and RW (Table 1). Modular, multicell designs offer options for coupling CW cells with RW habitat, thereby providing both wastewater treatment, long-term habitat development, functional diversity, and ancillary ecosystem services.
Table 1. Comparison of idealized ecosystem characteristics for constructed wetlands for wastewater treatment (CW), young restored wetlands (RWy), mature restored wetlands (RWm), and integrated constructed wetlands (ICW). Attributes adapted from Odum [79][80].
  CW RWy RWm ICW
Hydrologic regime closed open open open
Ecological Stress high low low moderate
Inorganic nutrients extrabiotic extrabiotic intrabiotic extra/intra
Dominant life strategy r r r, K r, K
Trophic structure simple simple complex complex
Habitat heterogeneity simple simple complex complex
Nutrient exchange rate rapid rapid slow moderate
Role of detritus low low high high
Stability low low high moderate
Temporal variability low high high moderate
Management effort high high low moderate
Ecosystem services narrow broad broad broad
While the ICW concept has been successfully demonstrated around the world, it is still in its infancy. There remains a disconnect between CW and RW in terms of theory, experimental design, management, and evaluation; this gulf will need to be bridged for ICW to become mainstream. Further, there is a need for more long-term, landscape-level studies on the integration of wetland services into the socioecological matrix, and for a cost–benefit analysis of wetland ecosystem services in these different systems. Finally, methods for the holistic evaluation of ecosystem services as they apply to ICW need to be refined and standardized to facilitate assessment within and among sites. These goals are attainable, and indeed are already being achieved by scholars and practitioners who seek to connect the built and natural environments for the benefit of both.

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Douglas Spieles
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