Large constructed treatment wetlands for decentralized wastewater treatment: an integrated decision-support framework for design, operation, and governance

1 Introduction

Decentralized water and wastewater treatment is no longer a peripheral topic reserved for remote villages or niche demonstration-scale projects. It is increasingly recognized as a strategic response to some of the most pressing water challenges of our time: aging infrastructure, rising energy and construction costs, climate variability, water scarcity, and the persistent mismatch between centralized planning models and local treatment realities. In this context, decentralization should not be reduced to a synonym for “small-scale solutions”, which tend to face governance and performance challenges in meeting effluent standards (Reymond et al., 2020). Rather, it refers more fundamentally to the treatment of water and wastewater at or near the point of generation, in ways that reduce dependence on extensive conveyance networks, improve adaptability to local conditions, and strengthen the resilience of communities and cities. Recent scholarship increasingly treats such distributed and hybridized arrangements not as provisional substitutes for conventional infrastructure, but as serious components of sustainable and resilient urban water management (Garrido-Baserba et al., 2024; Sha et al., 2024).

Within this evolving landscape, Treatment Wetlands (TWs) have regained a central role in sustainable water infrastructure. This technology actually has deep roots: The chinampas of pre-Columbian Mexico (engineered islands fertilized by city waste and lake water) already enabled a harmonious nature-based way of life hundreds of years ago (Becerril and Jiménez, 2007). Centuries later, the first modern engineer to document a treatment wetland was Cleophas Monjeau (United States), who applied for a patent of a vegetation-based water filter in 1900 (Monjeau, 1900). The modern application of TWs gained momentum in Germany about 50 years later. Botanist Käthe Seidel famously planted bulrushes in sewage and demonstrated dramatic water-cleaning effects in the 1950s (Seidel, 1955; Seidel, 1966; USEPA, 1993). Since then, TWs have gained importance not only because of their low external energy demand or comparatively simple operation, but because they represent a mature form of ecological engineering: a deliberately designed treatment environment in which hydraulic retention, filtration, sedimentation, sorption, microbial activity, plant–substrate interactions, and spatially differentiated redox conditions work synergistically to transform wastewater. Seven decades of modern research have demonstrated that TWs can reliably reduce organic matter, suspended solids, nutrients, pathogens, and a relevant share of emerging contaminants (Dotro et al., 2017; Sánchez et al., 2022; Overton et al., 2023). At the same time, they offer qualities that are increasingly valuable in a climate-constrained and resource-conscious era: modularity, landscape integration, biodiversity support, and the possibility of combining treatment with broader environmental and spatial functions. For precisely these reasons, TWs are now discussed not simply as low-cost treatment devices, but as multifunctional sanitation infrastructures with growing relevance for sustainable water management (Domínguez-Solís et al., 2025; Kumar et al., 2023; de Campos and Soto, 2024).

Yet the visibility of TWs in literature has also produced a persistent conceptual narrowing. Because many widely cited examples concern households, institutions, villages, or small communities, wetlands are still often framed as technologies whose natural domain is limited to the small or semi-rural scale. That perception is no longer adequate. Recent work on decentralized wastewater systems makes clear that decentralization is not defined solely by treatment capacity, but also by spatial logic, infrastructural arrangement, governance structure, and the degree to which treatment remains embedded in local service territories rather than being absorbed into remote metropolitan conveyance systems. In that sense, systems serving community, district, or even municipal populations may still be meaningfully decentralized when they are territorially proximate, operationally localized, and institutionally configured around local or regional management and reuse realities. This broader interpretation is especially relevant for treatment wetlands, whose technical and ecological strengths may be most compelling precisely where centralized expansion is economically, spatially, or institutionally constrained (Garrido-Baserba et al., 2024; Reymond et al., 2020; Sha et al., 2024).

For TW-based wastewater treatment, however, this shift in scale is not trivial. Large systems are not simply small systems drawn with a bigger ruler. Once wetlands are planned for higher and more variable flows, more heterogeneous influent conditions, and stronger regulatory expectations, the design problem changes qualitatively. Hydraulic and organic loading must be characterized with greater precision; pretreatment becomes more consequential; topography, hydrogeology, groundwater protection, and stormwater intrusion become more structurally decisive; and questions of land demand, substrate performance, flow distribution, oxygen transfer, solids accumulation, clogging risk, and maintenance logistics become integral to feasibility rather than secondary details. Recent literature also shows that common assumptions about TWs, particularly regarding land demand, can be misleading when assessed too narrowly, since direct footprint alone does not adequately capture broader infrastructure trade-offs relative to centralized systems. In parallel, the long-term operation of subsurface-flow (SSF) wetlands continues to be shaped by clogging processes and related hydraulic dysfunctions, making robust pretreatment, careful dimensioning, and explicit maintenance strategy indispensable from the outset (Fan et al., 2021; Liu et al., 2024; Mishra et al., 2024; Kataki et al., 2021).

At the same time, the technical horizon of TWs has expanded considerably through the development of hybrid and multistage configurations. Full-scale and review-based evidence shows that combinations of vertical-flow (VF), horizontal subsurface-flow (HF), and free water surface (FWS) stages can achieve more robust and differentiated treatment performance than single-stage systems required to perform every function alone (Parde et al., 2021). These hybrid arrangements allow designers to exploit complementary process strengths: intensified oxygen transfer and nitrification in vertical-flow units, denitrification and polishing functions in horizontal-flow stages, and additional buffering, finishing, or landscape integration in surface-flow components. Such configurations are especially important at larger scales, because they shift the design problem toward the logic of treatment trains tailored to site conditions, wastewater characteristics, operational priorities, and reuse goals. In this sense, hybridization is a planning principle that enlarges the feasible design space for wetland-based municipal and peri-urban treatment systems (Ávila et al., 2015; Parde et al., 2021; Masharqa et al., 2023).

Nevertheless, treatment performance alone does not secure infrastructural success. Wastewater systems do not fail only in the filter bed. They also fail in the institutional terrain around it (Reymond et al., 2020). The literature on decentralized and hybrid urban water systems repeatedly shows that long-term functionality depends on governance arrangements capable of supporting operation, monitoring, financing, maintenance responsibilities, training, compliance, and inter-agency coordination. This point is particularly relevant for distributed systems, where technical feasibility may outpace the development of clear institutional ownership and accountability. Inadequate governance can therefore undermine otherwise sound treatment concepts, whereas suitable governance frameworks can transform locally embedded systems into durable public assets. For large treatment wetlands, this means that engineering design and water governance should not be treated as separate conversations. They are co-constitutive dimensions of the same infrastructure question (Garrido-Baserba et al., 2024; Moretto et al., 2023).

Against this background, the present study positions large TWs not as isolated treatment units, but as integrated socio-technical infrastructures whose viability emerges from the interaction of design, site conditions, hydraulic reality, operational strategy, and institutional support. For the purposes of this article, large-scale refers to full-scale systems meeting at least one of the following criteria: a design capacity of at least 2,500 PE, an average or design inflow of at least 250 m³ d^-1, or a wetland treatment area of at least 5,000 m². This aligns the Introduction with Section 2.3 and Supplementary Table S1. The central premise is that the successful implementation of such systems cannot be secured through process design alone. Rather, it depends on the coordinated consideration of flow and wastewater characterization, hydrogeological and topographic suitability, pretreatment configuration, wetland treatment-train design, operation and maintenance requirements, and governance conditions capable of sustaining performance over time.

The contribution of this study therefore lies in clarifying how treatment wetlands should be planned when they are expected to function as reliable, scalable, and publicly relevant infrastructure. By framing large TW systems through the joint lenses of engineering design and water governance, the study seeks to move the discussion beyond generalized advocacy and toward a more rigorous implementation logic. In doing so, it addresses a question that is technically practical, scientifically important, and politically timely: Under which conditions can large wetland-based wastewater treatment systems become not just feasible projects, but resilient components of long-term water governance?

2 Materials and methods

2.1 Study design

This study was designed as a conceptual framework paper supported by a structured qualitative evidence synthesis and an embedded comparative case-study performance assessment. The objective was to integrate heterogeneous evidence on large treatment wetlands (TWs) into a practical decision-support framework for design, operation, performance assessment, governance, and long-term implementation. The methodological approach followed five steps: (i) definition of large-scale TW inclusion criteria; (ii) structured review of scientific, technical, regulatory, and project-level evidence; (iii) development of a six-dimensional conceptual planning framework; (iv) compilation and harmonization of project-level performance data; and (v) translation of the synthesized evidence into a decision-support matrix with an indicative readiness score.

The six framework dimensions were site and environmental factors, design principles, technological configurations, performance and efficiency, operational and governance aspects, and economic/environmental benefits. These dimensions were used as the analytical structure for evidence extraction and as the basis for the decision-support matrix presented in the Results section.

2.2 Evidence sources and literature search

Peer-reviewed literature was searched through Scopus, Web of Science, and ScienceDirect. Search terms included combinations of “large-scale constructed wetland,” “large-scale treatment wetland,” “hybrid treatment wetland,” “vertical flow constructed wetland,” “horizontal subsurface flow constructed wetland,” “free water surface wetland,” “decentralized municipal wastewater,” “wetland hydraulic design,” “constructed wetland performance,” “emerging contaminants,” “pharmaceuticals and personal care products,” “wetland governance,” “operation and maintenance,” “public-private partnerships,” and “life-cycle cost.”

The scientific search was complemented by technical standards, regulatory documents, design manuals, project reports, and project-monitoring records where these sources provided case-specific design, operational, or performance information. Key technical and regulatory sources included TW design guidance, wastewater legislation, and operation-and-maintenance manuals relevant to decentralized and large-scale wetland systems (USEPA, 1993; DWA, 2018; Dotro et al., 2017; Langergraber et al., 2020; European Union, 2024; SEMARNAT, 1997; SEMARNAT, 2021). Exact record counts for identification, screening, and exclusion are not reported; therefore, the study should be interpreted as a structured qualitative evidence synthesis with quantitative extraction for selected comparable performance indicators, not as a formal systematic review or meta-analysis.

Generative AI tools were used as assistive instruments for organizing literature-derived information, checking internal consistency, supporting table structuring, and improving manuscript language. All extracted values, calculations, classifications, citations, and interpretations were reviewed, corrected and verified by the authors.

2.3 Definition and inclusion criteria for large-scale treatment wetlands

For this study, large-scale TWs were operationally defined as full-scale systems meeting at least one of three criteria: (i) design capacity of at least 2,500 population equivalents (PE), (ii) average or design inflow of at least 250 m³ d⁻¹, or (iii) wetland treatment area of at least 5,000 m². The PE criterion was used as the primary threshold because PE is widely used in wastewater engineering to express biodegradable organic load per person (see Equation 1). However, PE definitions differ across design traditions and regulatory contexts. The OECD defines PE in relation to oxygen-demanding substances generated per person, von Sperling (2007) uses 54 g BOD d⁻¹, European practice commonly defines 1 PE as 60 g BOD₅ d⁻¹, and some U.S. design references use 80 g BOD d⁻¹ (OECD, 1997; European Parliament, 2022; Rowe and Abdel-Magid, 1995; von Sperling, 2007). Therefore, the 2,500 PE threshold was applied as a practical screening criterion.

P E = B O D l o a d k g d a y 0.054 0.06 0.08 k g i n h a b · d a y (1)

For comparability, two equivalent thresholds were added. Assuming 100 L PE⁻¹ d⁻¹, 2,500 PE correspond to 250 m³ d⁻¹. Assuming an indicative wetland area of 2 m² PE⁻¹, 2,500 PE correspond to 5,000 m² or 0.5 ha. These secondary thresholds allowed inclusion of large industrial, agricultural, stormwater, and polishing wetlands where PE was not reported but hydraulic or spatial scale clearly indicated full-scale implementation.

2.4 Project-level database and extracted variables

A project-level database of 64 large-scale TW cases was compiled from peer-reviewed studies, technical reports, project documentation, and project-monitoring records where available. For each project, the following variables were extracted where reported: location, wastewater source, treatment objective, treatment train, wetland type, inflow, PE, wetland area, dominant plants, substrate or media, pretreatment, post-treatment or disinfection, influent and effluent concentrations, reported removal efficiencies, evidence level, data limitations, special design or operational features, and source reference.

The conventional performance parameters considered were BOD₅, COD, TSS/SS, nitrogen indicators, and total phosphorus. Because nitrogen reporting differed among studies, TN, TKN, NH₄-N, or other explicitly defined nitrogen forms were grouped as “N indicator” only where interpretation within the individual case was possible. This avoided artificial comparability across studies with different nitrogen metrics.

2.5 Performance harmonization and statistical synthesis

Removal efficiencies were calculated from paired influent and effluent concentrations using Equation 2:where η is removal efficiency (%), C i n is influent concentration, and C o u t is effluent concentration.

η = C i n − C o u t C i n × 100 (2)

When original sources reported removal percentages without paired concentrations, these values were retained and marked accordingly. Values reported as below detection limits or thresholds were treated conservatively as the reported threshold. Missing values were not converted to zero, whereas true reported 0% removals were retained.

To avoid pooling unlike systems under the generic term “hybrid wetland,” projects were grouped according to wastewater source, treatment objective, and treatment-train sequence. This resulted in 18 detailed configuration groups, which formed the statistical basis of the conventional-pollutant performance synthesis. Broader synthesis families were used only for narrative interpretation, not as homogeneous statistical populations.

For each detailed configuration group and pollutant, the arithmetic mean, sample standard deviation, and interquartile range were calculated using Equations 3–5, respectively: where x ¯ is the arithmetic mean removal efficiency, x i is the project-level removal value, n is the number of available observations, s is the sample standard deviation, and I Q R is the interquartile range.

x ¯ = 1 n ∑ i = 1 n x i (3)

s = ∑ i = 1 n x i − x ¯ 2 n − 1 (4)

I Q R = Q 3 − Q 1 (5)

Minimum, maximum, median, interquartile range, mean, sample standard deviation, and number of observations were reported. Because the dataset includes different climates, wastewater sources, treatment objectives, monitoring periods, and reporting formats, the results were interpreted as a comparative performance synthesis. Boxplots were generated in RStudio from project-level removal values. Missing values were excluded, true zero-removal values were retained, and single-case groups were displayed as individual points.

2.6 Emerging-contaminant evidence synthesis

Emerging contaminant removal was synthesized separately from the conventional-pollutant case database because large-scale project-level evidence was insufficient for comparable statistical pooling. This synthesis focused on pharmaceuticals and personal care products, endocrine-disrupting compounds, antibiotics, steroid hormones, stimulants, analgesics and anti-inflammatory drugs, alkylphenols, bisphenols, phthalates, and hydrophobic organic contaminants. These compounds are known to show strongly compound-specific behavior in TWs, depending on sorption, biodegradation, photodegradation, plant uptake, redox conditions, substrate properties, and hydraulic retention time (Li et al., 2014; Verlicchi and Zambello, 2014; Ávila et al., 2015; Ilyas and van Hullebusch, 2020a; Ilyas and van Hullebusch, 2020b; Overton et al., 2023; Hou et al., 2025).

For each compound group, available information was extracted on TW configuration, removal efficiency, substrate or media, dominant plants, dominant removal mechanisms, hydraulic retention time, design-sensitive parameters, and planning implications. The resulting matrix therefore reports compound-group tendencies and design implications.

2.7 Governance, operation, financing, and economic synthesis

Operational, governance, and economic evidence was synthesized thematically because long-term TW performance depends not only on treatment processes but also on ownership, operation, maintenance, monitoring, financing, institutional accountability, and contract enforcement. Evidence was extracted from decentralized sanitation governance literature, TW operation and maintenance guidance, financing frameworks, public-private partnership literature, and life-cycle cost studies (Ross et al., 2014; Dotro et al., 2017; World Bank, 2019; von Sperling and Platzer, 2020; Lima et al., 2021; OECD, 2025).

The synthesis focused on recurring implementation risks, including unclear asset ownership, insufficient O&M budgets, weak monitoring arrangements, fragmented responsibilities, limited operator training, lack of sludge and biomass management, and weak accountability across political cycles. Economic and environmental evidence was evaluated using life-cycle logic, including capital expenditure, O&M costs, media renewal, land opportunity cost, energy demand, greenhouse-gas emissions, water reuse, biomass management, and ecosystem-service potential.

Where quantitative cost or greenhouse-gas data were not comparable across cases, the evidence was used qualitatively to identify planning implications and trade-offs.

2.8 Development of the decision-support matrix with readiness score

The synthesized evidence from the six-dimensional planning framework was further condensed into a decision-support matrix for large-scale TW planning.

For each framework dimension, key decisions, core KPIs/indicative thresholds, red flags/no-go risks and the planning output deliverables are illustrated in a hands-on one-pager matrix.

A four-level readiness score was added as a screening tool, and the total score can be calculated using Equation 6:where R j is the readiness score for dimension j, ranging from 0 to 3, and R t o t a l is the total readiness score with a maximum of 18.

R t o t a l = ∑ j = 1 6 R j (6)

3 Results

3.1 Integrated planning of large-scale decentralized treatment wetlands

3.1.1 Key dimensions of large-scale treatment wetlands projects

The design and implementation of large-scale treatment wetlands present both opportunities and challenges.

Figure 1 presents a synthesis of the six essential key dimensions that underpin the success of large-scale treatment wetlands projects: site and environmental factors, design principles, technological configurations, performance and efficiency, operational and governance aspects, and economic and environmental benefits. Together, these interrelated elements illustrate how technical robustness, ecological processes, and institutional arrangements converge to deliver resilient and cost-effective wastewater treatment solutions, setting the stage for the detailed discussion that follows.

3.1.2 Integrated decision-support matrix

Based on the key planning-dimensions-structured findings of the systematic review, the matrix for planning large, decentralized treatment wetlands in Table 1 could be extracted, to offer the reader a practical decision support.

TABLE 1. Decision-support matrix for planning large, decentralized treatment wetlands (compressed version).

Notes: 1) The thresholds are intended as decision-support benchmarks rather than universal design limits. Regulatory discharge limits, reuse standards, and site-specific risk assessments remain decisive. 2) Supplementary Table S1 provides the complete version of the decision-support matrix.

The matrix highlights that the success of large TWs is not determined by one isolated design parameter, but by the quality of interaction among the six planning dimensions. Site constraints shape the design basis; the design basis determines feasible technological configurations; technological configurations define achievable performance; and long-term performance depends on governance, financing, operation, and maintenance. Economic and environmental co-benefits can strengthen feasibility and public acceptance, but only when they are explicitly linked to measurable outputs, responsible actors, and realistic financing mechanisms.

For practical application, each dimension can be scored using a four-level readiness scale (Table 2): 0 = critical gap/no-go condition, 1 = partially addressed but requiring mitigation, 2 = sufficiently documented and institutionally secured, and 3 = robust/best-practice level. A project should proceed from conceptual planning to detailed design only if no dimension contains an unresolved no-go condition and if governance, O&M financing, and monitoring responsibilities are explicitly assigned.

TABLE 2. Readiness score.

Evaluation Recommendation: A project should not proceed from conceptual planning to detailed design if any of the six dimensions scores 0, or if the combined readiness score is below 12 out of 18. A score of ≥15 indicates high implementation readiness, provided that no regulatory or environmental no-go condition remains unresolved.

3.2 Site and environmental factors

3.2.1 Site selection and land availability

TWs designed for communities larger than 2,500 PE face challenges that differ from small-scale applications, particularly the need for extensive land areas, which can limit feasibility in densely populated or topographically complex regions (Fan et al., 2021). Although these systems are generally less demanding than conventional activated sludge plants, they still require skilled supervision to manage hydraulic loading, regulate vegetation growth, and prevent clogging of filtration beds (Rahman et al., 2020; Yadav et al., 2024). Their feasibility therefore depends heavily on the availability of suitable sites, ideally large, flat areas at the lowest geodetic altitude of a basin with appropriate soil, hydrology, and topography to ensure long-term stability and treatment efficiency. In urban or peri-urban contexts, securing such land often represents a decisive constraint for implementation (Fan et al., 2021; Rodriguez-Dominguez, 2020; Garcia-Chevesich et al., 2024). Furthermore, the land acquisition already interacts with local acceptance and long-term governance (e.g., who controls buffer zones and reuse corridors).

3.2.2 Climatic and geological resilience

Extreme weather events such as heavy rainfall, flooding, drought, or temperature fluctuations can significantly affect the performance of treatment wetlands. Designing for resilience therefore requires careful consideration of local climatic conditions, including bed depth, hydraulic retention time, and vegetation choice, to ensure year-round efficiency (Hube et al., 2023; Hou et al., 2025; Yadav et al., 2024). Large-scale systems must also be able to cope with seasonal variability in pollutant loads, while site selection should account for broader environmental risks. In addition to climatic extremes, tectonic factors such as seismic activity or soil instability are particularly relevant in earthquake-prone regions, as they may compromise the integrity of geomembranes, hydraulic structures, and substrate layers, reducing long-term reliability (Garcia-Chevesich et al., 2024). For this reason, suitable soil conditions, hydrology, topography, and geological stability are essential to support safe operation and ensure sustained performance over time (Rodriguez-Dominguez, 2020; Garcia-Chevesich et al., 2024).

3.3 Key design principles

3.3.1 Hydraulic load and hydraulic design

Proper hydraulic design is critical to ensure effective treatment. Systems must be dimensioned for appropriate hydraulic retention time (HRT) and hydraulic loading rate (HLR), with inlet and outlet structures engineered to distribute flow evenly and avoid short-circuiting or dead zones (Jiang and Chui, 2022; Wei et al., 2025). Equally essential is the rigorous determination of influent flows through detailed hydrological and wastewater generation studies, since accurate quantification of daily and seasonal variations directly conditions the sizing of each treatment stage. Flow characterization provides the basis for designing pretreatment units (screening, grit removal), anaerobic biological processes (sedimentation and clarification), and subsequent treatment wetland cells, ensuring hydraulic stability, protecting substrates from clogging, and maintaining long-term efficiency. Without such flow studies, design parameters risk being underestimated or oversized, leading to operational inefficiencies and regulatory non-compliance (USEPA, 1993; Ergaieg and Ben Miled, 2021; Didanovic and Vrhovsek, 2023).

3.3.2 Pollutant load and seasonal variability

Treatment wetlands must be designed to handle fluctuations in pollutant loads, which can vary seasonally or due to changes in wastewater composition. Flexible configurations and robust hydraulic design are necessary to maintain treatment efficiency under these variable conditions (Avila et al., 2015; Ishtiaq et al., 2025; Rahman et al., 2020). Equally critical is the systematic characterization of wastewater quality, since seasonal contributions from rainwater infiltration, stormwater runoff, and shallow groundwater intrusion into open conveyance systems can significantly alter influent composition. Accurate water quality studies provide the basis for determining pollutant loads, nutrient concentrations, and emerging contaminants, thereby guiding the sizing and staging of pretreatment, sedimentation, clarification, and wetland units. Experience further demonstrates that it is feasible to rely on national characterization datasets, such as those available in Mexico for urban communities of different sizes (CONAGUA, 2015), which offer valuable reference values for design and planning when site-specific monitoring is limited. Systems that integrate hybrid VF–HF stages have shown resilience in adapting to such variability, ensuring compliance with regulatory standards and long-term operational stability (Hou et al., 2025).

3.4 Treatment wetland technologies

3.4.1 Hybrid systems (VF + HF combinations)

Hybrid and modular configurations are particularly relevant at this scale, as they allow for flexible capacity expansion and ensure operational continuity during maintenance (Ishtiaq et al., 2025; Avila et al., 2015). Proper primary treatment is essential to reduce solids and protect the substrate from clogging (DWA, 2018; Pucher and Langergraber, 2019; Rahman et al., 2020). Screening and grit removal units, a pumping station, Imhoff sedimentation tanks, clarification tanks or lagoons, and drying and humification beds must be considered.

From the clarification stage, water is intermittently distributed into VF–HF treatment wetlands, composed of sand and gravel layers lined with geomembrane and planted with climate-appropriate species. Large-scale treatment wetlands commonly adopt hybrid configurations that combine vertical subsurface flow (VF), horizontal subsurface flow (HF), and free water surface (FWS) units (Ran, 2025; Chen et al., 2022). This multi-stage design enhances pollutant removal efficiency, balancing aerobic and anaerobic processes to achieve compliance with stringent discharge standards (Avila et al., 2015; Ishtiaq et al., 2025; Rahman et al., 2020). Subsurface flow systems are generally preferred for municipal wastewater treatment because they minimize odor emissions, reduce mosquito breeding, and offer higher treatment capacity per unit area compared to surface flow wetlands (USEPA, 1993; Halverson, 2004; Stefanakis, 2020). These advantages make them more suitable for large-scale applications in urban and peri-urban contexts (Rodriguez-Dominguez, 2020; Garcia-Chevesich et al., 2024). The modular design of TWs enables capacity to be scaled according to community needs while facilitating maintenance without disrupting overall operation. By dividing the system into multiple parallel cells that can be managed independently, these configurations provide flexibility, resilience, and operational continuity, ensuring that treatment performance is maintained even during maintenance or fluctuating loading conditions (Ishtiaq et al., 2025; Avila et al., 2015).

These modular units enhance oxygen transfer and microbial activity, thereby supporting organic matter degradation and nutrient removal, consistent with findings from hybrid VF–HF wetland applications (Avila et al., 2015; Yadav et al., 2024). VF beds are particularly effective for nitrification due to their high oxygen transfer capacity, while HF beds provide favorable conditions for denitrification and solids removal. The sequential use of VF followed by HF ensures robust treatment performance across diverse pollutant loads (Jiang and Chui, 2022; Yadav et al., 2024). In this stage, hydraulic design must ensure adequate retention times and uniform flow distribution to avoid short-circuiting and dead zones (Jiang and Chui, 2022; Wei et al., 2025).

Finally, the effluent is conveyed to a free water surface (FWS) wetland, which serves as a tertiary stage providing both polishing and natural disinfection. By incorporating meandering flow paths, these units extend hydraulic retention time and enhance pathogen removal through sedimentation, solar radiation, and microbial interactions, while simultaneously creating ecological habitat and biodiversity benefits (Vymazal, 2010; Eawag and Stauffer, 2020). This final stage consolidates international experience with hybrid treatment wetlands, demonstrating their ability to achieve high removal efficiencies for COD, nitrogen, and phosphorus while delivering ecological co-benefits and long-term resilience (Masharqa et al., 2023; Hou et al., 2025).

3.4.2 Media and substrate choices

The selection of media and substrate materials is a decisive factor in the performance and resilience of treatment wetlands. Conventional substrates such as gravel and sand provide structural stability and hydraulic conductivity, while lightweight aggregates and industrial by-products offer cost-effective alternatives with enhanced sorption capacity. Large basins are often lined with HDPE geomembranes as a cost-efficient substitute for concrete, ensuring impermeability and reducing construction costs (Rahman et al., 2020; Hube et al., 2023). Beyond structural considerations, substrate composition directly influences pollutant removal pathways. Reactive media such as zeolite, steel slag, and expanded clay aggregates have demonstrated high adsorption capacity for ammonium and phosphorus, thereby improving nutrient removal efficiency (Ran, 2025). Industrial by-products including blast furnace slag and fly ash have been successfully applied as low-cost sorbents, though their long-term stability and potential leaching require careful evaluation (Wu et al., 2022; Wang et al., 2024). Particle size distribution and porosity are equally critical, as they regulate oxygen transfer and microbial colonization, supporting nitrification–denitrification processes and organic matter degradation (Wu et al., 2023). Substrate clogging remains a major operational challenge in large-scale systems, underscoring the need for appropriate pretreatment and periodic regeneration or replacement strategies (Pucher and Langergraber, 2019). Recent studies further highlight the role of engineered substrates in removing emerging contaminants such as pharmaceuticals and personal care products, expanding the ecological function of TWs (Hou et al., 2025). Thus, media and substrate choices must be guided not only by cost and availability but also by their capacity to sustain hydraulic conductivity, enhance pollutant removal, and ensure long-term system resilience.

3.4.3 Plant selection and management

Selecting appropriate macrophyte species that thrive under wetland conditions is essential for pollutant removal, hydraulic stability, and long-term system resilience. Native or regionally adapted species are generally preferred, as they enhance nutrient uptake, stimulate microbial activity in the rhizosphere, and reduce risks of ecological imbalance (Yang et al., 2024; Yadav et al., 2024). Plant roots provide oxygen transfer to saturated substrates, supporting nitrification processes, while their biomass contributes to organic matter stabilization and pathogen reduction (Brix, 1997; Wu et al., 2022). Species such as Phragmites australis, Typha latifolia, and Scirpus spp. are widely documented for their tolerance to variable hydraulic and pollutant loads.

More recently, hydrophilic ornamental plants–including Canna indica, Heliconia spp., and Cyperus papyrus–have been incorporated into treatment wetlands, particularly in tropical and subtropical regions. These species not only contribute to pollutant removal through nutrient uptake and rhizosphere interactions but also create aesthetically pleasing environments that integrate into urban and peri-urban landscapes (Arias and Brown, 2009; Brix et al., 2011). Their use has been shown to improve public acceptance of TWs by combining ecological function with recreational and cultural value, thereby strengthening the socio-environmental legitimacy of decentralized wastewater treatment systems (Sowińska-Świerkosz and Stefanakis, 2026).

Ongoing management is required to control excessive growth, prevent substrate clogging, and avoid the spread of invasive species. Seasonal harvesting of above-ground biomass not only maintains hydraulic conductivity but also removes nutrients sequestered in plant tissues, reducing internal cycling of nitrogen and phosphorus (Vymazal, 2010; Ran, 2025). Mixed planting strategies are recommended to balance resilience against climatic variability, enhance biodiversity, and provide redundancy in case of species decline (Hou et al., 2025). Recent studies further highlight the role of macrophytes in attenuating emerging contaminants, including pharmaceuticals and personal care products, through uptake, rhizodegradation, and stimulation of microbial consortia (Avila et al., 2015; Yang et al., 2024).

Thus, plant selection and management must be guided not only by ecological suitability and pollutant removal capacity but also by operational considerations, regulatory compliance, and socio-cultural acceptance, positioning vegetation as a central design and management dimension in treatment wetlands.

3.4.4 Bioelectrochemical systems (MFCs and MECs)

Bioelectrochemical systems harness microbial redox reactions to enhance TWs. In a microbial fuel cell (MFC), exoelectrogenic bacteria oxidize organic pollutants at an anaerobic anode and transfer electrons to a cathode where oxygen (from air or plant roots) is reduced (Escapa et al., 2014). This creates a small electrical current while also improving pollutant removal. The power output is typically around 0.2 W/m² long term output to maximum 1.1 W/m² at a theoretical maximum 3.2 W/m² of projected anode area (European Commission, 2014). Pilot projects have demonstrated visible output, e.g., Plant-e’s “Starry Sky” system in the Netherlands lit hundreds of LEDs using wetland-grown plants (European Commission, 2014).

Integration of MFCs has been shown to boost wastewater treatment efficiency. Microcosm-scale reactor studies reported 6%–10% higher COD removal in an MFC-enhanced wetland compared to a benchmark wetland; Nitrate removal even increased by 40.2% (at almost neutral/non-acidic influent pH) (Wang et al., 2016; Zhu et al., 2022). This suggests that the electric circuit stimulated nitrifiers/denitrifiers on the cathode side and/or enhanced organic oxidation at the anode. Other lab studies even observed up to ∼20–30% lower effluent NH₄-N and COD under optimal MFC conditions (Corbella and Puigagut, 2018).

Practically, TW-MFCs in large systems might use submerged graphite or metal anodes and cathodes buried in zones with redox gradients. An MFC zone could be inserted at the bottom of a gravel bed, with electrodes connected to nearby planting filters. External circuit resistances and electrode spacing are design parameters. MFCs are not yet commercialized and have not been demonstrated at industrial scale (Ieropoulos et al., 2024). They are nevertheless discussed here because they may (i) reduce the land requirements of TWs, (ii) increase the capacity of existing TWs without requiring additional land, and (iii) reduce greenhouse-gas impacts. Zhu et al. (2022) reported a 24% reduction in global warming potential following MFC integration, which may support access to carbon-related finance.

Microbial electrolysis cells (MECs) also remain immature for large-scale wetland applications but may offer future potential under rising energy costs. MECs operate similarly to MFCs but require an external voltage to drive hydrogen production. In an MEC, the anode and cathode are both maintained in inert (often anaerobic) chambers; an external voltage (usually <1 V) is applied to drive hydrogen evolution at the cathode. Essentially, wastewater organics are oxidized at the anode by microbes, sending electrons through an external circuit to the cathode, where protons combine with those electrons to form H₂ gas (Escapa et al., 2014). This configuration treats organics while yielding a storable fuel. MECs have also shown to achieve high organic removal. Bench- and pilot-scale reactors often remove 80%–90% of COD (Escapa et al., 2014). A significant portion of the oxidized substrate ends up as hydrogen: theoretical energy recovery can approach ∼50% of the chemical energy in the COD. When accounting for the modest electrical input, net energy yields on the order of 1–2 kWh per kg COD removed have been demonstrated (Escapa et al., 2014). In practical terms, Escapa et al. (2014) estimated that an MEC pre-treatment (capturing H₂) could cut the overall aerobic treatment energy requirement by ∼30%, yielding ∼1.05 kWh/kg-COD (including H₂ credit). MECs require almost the same energy as activated sludge (0.9 kWh per kg COD removed) to split water, so the energy balance is only positive when the produced H₂ is captured (Gil-Carrera et al., 2013).

3.5 Performance and efficiency

3.5.1 Compliance with regulations

Properly engineered large-scale wetlands have demonstrated compliance with both international and national regulatory frameworks. In particular, hybrid VF–HF systems can meet the stringent effluent quality requirements established under the EU Directive 91/271/EEC for BOD, COD, nitrogen, and phosphorus removal, while comparable standards under USEPA guidelines and SEMARNAT (1997), SEMARNAT (2021) regulations in Mexico also emphasize nutrient and organic matter reduction, confirming the suitability of TWs as decentralized treatment solutions (DWA, 2018; Hendy et al., 2023; European Union, 1991). Meeting these regulatory standards is a decisive factor for success, as compliance not only validates technical performance but also requires navigating complex local, state, and federal permitting processes. Although time-consuming, such procedures are essential to ensure system legitimacy, sustainability, and long-term acceptance (Avila et al., 2015; Hendy et al., 2023).

This regulatory relevance has increased substantially with the recast of the EU Urban Wastewater Treatment Directive. Directive (EU) 2024/3019 introduces reinforced requirements for nutrient removal, micropollutant control, wastewater monitoring, water reuse, and energy neutrality. It requires quaternary treatment for urban wastewater treatment plants of 150,000 PE and above, and for agglomerations of 10,000 PE and above where discharges pose a risk in areas sensitive to micropollution (European Union, 2024). The Directive also establishes an 80% minimum removal target for selected organic micropollutant indicators and introduces extended producer responsibility, requiring producers of medicinal and cosmetic products to cover at least 80% of the full costs related to quaternary treatment and monitoring (European Union, 2024).

Within this new regulatory landscape, TWs are not a universal substitute for advanced treatment technologies such as ozonation, activated carbon, or membrane-based systems. However, they can function as important complementary or polishing systems, especially where land is available, energy costs are critical, reuse is intended, or small and medium-sized agglomerations require robust and lower-maintenance solutions. Recent policy-oriented analysis explicitly highlights that TWs align with several objectives of the revised EU framework, including nutrient control, micropollutant mitigation, stormwater management, water reuse, and energy-neutral wastewater treatment strategies (Gautier, 2025).

In regard to the capacity of treatment wetlands to attenuate emerging contaminants, it is important to note that removal is neither complete nor uniform across all compound classes. Reported efficiencies vary widely depending on physicochemical properties, hydraulic conditions, and plant–microbe interactions, with some pharmaceuticals and personal care products showing substantial reductions while others persist at detectable levels (Avila et al., 2015; Sgroi et al., 2018; Singh and Tiwari, 2026). Nevertheless, this partial but significant removal capacity aligns with the broader policy direction in which micropollutant control is becoming a regulatory driver and a financing issue—for example, through extended producer responsibility provisions for pharmaceuticals and cosmetics in the EU Urban Waste Water Treatment Directive recast (European Union, 2024).

3.5.2 Treatment efficiency synthesis of 64 large-scale TW projects

To provide an empirical basis for the performance expectations of large treatment wetlands, 64 projects meeting the large-scale definition were compiled from peer-reviewed studies, technical reports, project documentation, and project-monitoring records. Projects meeting the scale criteria were retained in the supplementary inventory for transparency; however, only pollutant-specific paired concentration data or explicitly reported removal values were included in the statistical performance synthesis. Based on wastewater source, treatment objective, and treatment-train configuration, the projects were assigned to 18 detailed configuration groups (Table 3; complete dataset in Supplementary Table S3). Mean removal efficiencies, sample standard deviations, minima, maxima, medians, interquartile ranges, and numbers of available observations were calculated for each pollutant and group where data availability allowed.

TABLE 3. Configuration-based synthesis of conventional pollutant removal in large-scale treatment wetlands.

As shown in Figure 2, the highest and most consistent removals of organic load were achieved by municipal/tourism VF→HF hybrids (G04), with a mean BOD₅ removal of 95% ± 5% (n = 2) and a COD removal of 93% in the single available COD case. Pretreated municipal VF/VSSF systems (G11), including the Cihuatlán-type configuration with primary clarification before intermittently loaded vertical-flow cells, also showed very high removal of BOD₅, COD, and TSS/SS, although the organic-matter and solids data are based on one paired case and should therefore be interpreted as lighthouse evidence rather than a generalizable group mean. The most data-rich municipal category, HSSF systems after primary treatment (G08), provided the strongest statistical basis, with a mean BOD₅ removal of 84% ± 14% (n = 10) and COD removal of 70% ± 16% (n = 7).

Nitrogen removal was most effective in VF → HF hybrids (G04) with 62% ± 33% TN removal (n = 2) and in pretreated VF systems (G11) with 66% TN removal. The highest reported phosphorus removals were found in municipal HSSF systems (G08) with 61% ± 13% TP removal (n = 3) and in a two-stage HSSF → VF hybrid (Group G06) where a single case achieved 82% TP removal. These two nutrient removal efficiencies are illustrated in Figure 3:

Several large-scale installations exemplify the achievable performance as illustrative lighthouse projects:

• Red Sea Global TW, Saudi Arabia (G04) – Designed for 16,000 m³/d, this two-stage VF → HF hybrid achieved 98.7% BOD₅, 92.9% COD, 98.9% TSS, 84.8% TN, and 76.8% TP removal under arid-climate conditions (Ishtiaq et al., 2025).
• Cihuatlán, Mexico (G11) – Serving 20,000 PE with six intermittently loaded vertical flow cells after a modified Imhoff clarifier, this system reported 93% BOD₅, 91% COD, 97% TSS, and 66% TN removal over multiple years of monitoring (Janisch and Schulz, 2021; CIATEJ, 2024 (internal report); Schmelzle, 2024).
• Orhei, Moldova (G05) – A French two-stage vertical flow reed bed treating raw municipal wastewater for 20,000 inhabitants achieved 86% BOD₅, 86% COD, 96% TSS, and 66% NH₄-N removal under cold-climate conditions (Masi et al., 2017).
• Merone CSO, Italy (G02) – This aerated vertical flow system for combined sewer overflow (peak hydraulic load 21,600 m³/d) removed 91% BOD₅, 89% COD, 95% TSS, and 54% NH₄-N, demonstrating that high-rate vertical flow wetlands can also handle stormwater-driven shock loads (Bresciani et al., 2026).
• Dakhla Oasis, Egypt (G08) – A horizontal subsurface flow wetland after Imhoff tank (design 750 m³/d, 4167 PE) shows expected BOD₅ removal of 86% (Gabr, 2022).

3.5.3 Removal of emerging contaminants

Treatment wetlands are increasingly recognized for their ability to remove pharmaceuticals and personal care products (PPCPs). Removal mechanisms include microbial degradation, adsorption onto substrates, and plant uptake. Studies highlight that hybrid systems can achieve significant reductions in these contaminants, contributing to improved aquatic health and water reuse safety (Avila et al., 2015; Hou et al., 2025; Yang et al., 2024). However, removal efficiencies are neither complete nor uniform across all compound classes; they vary widely depending on physicochemical properties, hydraulic conditions, substrate type, and plant-microbe interactions (Verlicchi and Zambello, 2014; Ilyas and van Hullebusch, 2020b).

A comprehensive compilation of recent peer-reviewed studies is presented in Supplementary Table S4, which compares removal efficiencies of a wide range of pharmaceutical compounds in activated sludge (AS) systems versus different configurations of treatment wetlands under secondary and tertiary treatment conditions. The table also includes information on dominant removal mechanisms, key advantages of TWs over AS, critical design parameters, and typical hydraulic retention times.

As shown in the complete Supplementary Table S4, TWs are particularly advantageous for compounds that are poorly eliminated by conventional activated sludge. For example, the antiepileptic drug carbamazepine is removed only by about 10% in AS, whereas planted HF wetlands with LECA (Lightweight Expanded Clay Aggregate) substrate can achieve up to 89% removal through adsorption and plant uptake (Özengin and Elmaci, 2016; Chen et al., 2018). Similarly, diclofenac is removed by 35% in AS, but vertical flow TWs with cocopeat or hybrid zeolite systems reach 97%–100% removal, demonstrating the critical role of specialized media and aerobic conditions (Sithamparanathan et al., 2024; Zapata-Morales et al., 2020). For ibuprofen and naproxen, which are already well removed by AS (90% and 75%, respectively), planted VF and HF TWs achieve equal or higher removal rates (up to 98%) without energy-intensive aeration, highlighting the low-energy polishing capacity of wetlands (Zhang et al., 2011; Zapata-Morales et al., 2020).

The ability of TWs to remove antibiotics is highly substrate-dependent. Standard gravel-based HF wetlands show moderate removal for sulfamethoxazole (43% ± 24%), but aerated vertical flow systems amended with bamboo biochar achieve >92% removal (Ameen et al., 2025; Zhang et al., 2020). For fluoroquinolones such as ciprofloxacin and ofloxacin, hybrid TWs filled with Jordanian zeolite reach >98% elimination, far exceeding the 80% removal of AS (Al-Mashaqbeh et al., 2024). Trimethoprim, a commonly co-administered antibiotic, is removed by 95% in AS; cocopeat-based vertical TWs and hybrid systems achieve 96%–100% removal, demonstrating that TWs can match or outperform conventional treatment while avoiding the continuous aeration demand of conventional activated-sludge treatment (Sithamparanathan et al., 2024; Ilyas and van Hullebusch, 2020b).

For steroid hormones, treatment wetlands are exceptionally effective. Progesterone is completely removed (100%) in VF TWs, and estriol is also eliminated entirely under aerobic conditions (Ilyas and van Hullebusch, 2020a). 17β-Estradiol and estrone show mean removals of 79% and 83% in HF TWs, respectively, which can be further enhanced by iron-carbon-modified substrates (Li et al., 2025). These findings underscore that TWs can serve as efficient polishing steps for endocrine-disrupting compounds, a capacity that is increasingly relevant as regulators begin to set discharge limits for such micropollutants (European Union, 2024).

The performance of TWs for emerging contaminants is governed by several critical parameters, as summarized in the last columns of Supplementary Table S4. Key factors include the choice of substrate (e.g., LECA, zeolite, cocopeat, biochar), the presence of plants (especially Typha spp., Phragmites australis, or Arundo donax), hydraulic retention time (typically between 1 and 10 days, depending on the compound), and redox conditions (aerobic conditions favor biodegradation of most compounds, while some hydrophobic substances are removed by sorption under anoxic conditions). Hybrid configurations that combine aerobic VF and anoxic HF stages are particularly advantageous because they provide both aerobic biodegradation and sorption-dominated removal pathways (Ilyas and van Hullebusch, 2020b; Avila et al., 2015).

In addition to conventional TW designs, intensified systems have been shown to further enhance micropollutant removal. For example, coupling a horizontal subsurface flow wetland with ozonation increased ibuprofen and naproxen removal to >97% (Lancheros et al., 2019). Integrating microbial fuel cells (MFCs) into TWs achieved >99% removal of sulfamethoxazole and tetracycline while simultaneously generating electricity (Wen et al., 2020). Electrolysis-integrated tidal flow TWs raised the removal of several pharmaceuticals by 28%–68% compared to non-electrolyzed controls (Liu et al., 2020). Activated carbon-amended vertical flow TWs completely removed (98%–100%) bisphenol A and other phenolic EDCs, demonstrating that targeted substrate amendments can transform TWs into highly efficient polishing units (Wirasnita et al., 2018).

Table 4 synthesizes the compound-specific evidence from Supplementary Table S4 at the compound-group level, highlighting dominant removal mechanisms and implications for planning and design.

TABLE 4. Group-level synthesis of emerging-contaminant removal in treatment wetlands.

This group-level synthesis is intended to summarize broad removal tendencies without replacing the compound-specific evidence provided in Supplementary Table S4. Because PPCP and EDC removal in treatment wetlands depends strongly on compound properties, media composition, hydraulic retention time, redox conditions, vegetation, and wastewater matrix, the table reports planning implications rather than universal removal guarantees.

3.6 Operational and governance aspects

Although TWs are widely regarded as low-input and operationally robust compared to conventional activated sludge plants, no treatment wetland is maintenance-free. The “passive” label becomes misleading when it encourages under-budgeting for routine inspection, inlet and outlet control, and preventive maintenance (Dotro et al., 2017). In large-scale TW systems, performance reliability depends on (i) execution of repeatable routines (inspection, cleaning, cutting/harvesting, sludge management), (ii) early diagnosis of hydraulic malfunctioning and incipient clogging, and (iii) a monitoring plan that is appropriately scaled to regulatory compliance rather than to research-grade measurement (Dotro et al., 2017; Pucher and Langergraber, 2019).

Routine tasks should be designed into the project during planning, not after construction. The owner has to be aware of the life-cycle-costs for their financial plannings. Practical guidance stresses that O&M planning should specify (a) how and how often upstream sludge is removed, transported, treated, and disposed/reused; (b) the timing and frequency of cutting/harvesting vegetation (and what is done with harvested biomass); (c) responsibilities and checklists/logbooks for routine operation and monitoring; (d) the expected time to major interventions (e.g., removal of accumulated solids, surface scraping, media cleaning/replacement, replanting); and (e) ensured physical access for periodic heavy maintenance (Langergraber et al., 2020). These “design-for-operations” elements are especially critical in decentralized settings where the local operator may be a small municipal unit with limited technical depth and staff turnover risk (von Sperling and Platzer, 2020).

Clogging is consistently identified as the major operational risk in SSF wetlands and is best treated as a governance issue as much as a technical one: it is frequently caused or accelerated by neglected pretreatment (e.g., delayed desludging of septic or settling tanks), uneven influent distribution, and loading rates that exceed recommended envelopes (Dotro et al., 2017; Pucher and Langergraber, 2019). For HF wetlands treating domestic wastewater, guidance highlights that clogging is often driven by excessive organic/solids loading onto the gravel bed, commonly linked to poor pretreatment maintenance and/or overloading; conversely, bed life can be extended by appropriate media selection, realistic loading rates (hydraulic and solids), and uniform distribution along the bed width (Dotro et al., 2017). For vertical-flow systems, additional pitfalls include unsuitable sand/gravel gradation (excess fines), inadequate runoff protection that washes sediments onto the filter surface, and uneven wetland surface slope that creates persistent ponding (Dotro et al., 2017).

Monitoring and compliance should follow a tiered logic: (i) operational monitoring to keep units functioning (visual inspection of inlet structures, water levels, distribution uniformity, ponding, vegetation status) plus simple field parameters where appropriate; and (ii) compliance monitoring to meet legal discharge standards, which requires defined sampling locations, agreed parameters, and specified frequency (Langergraber et al., 2020). Even when wetlands operate largely without mechanical complexity, pumps (if present) and flow-control components must be maintained to manufacturer specifications, and upstream coarse screening/grit removal must be inspected and cleaned because failures propagate into hydraulic short-circuiting and clogging (Dotro et al., 2017). In practice, modular layouts in large-scale systems can reduce operational risk because individual units can be temporarily taken offline for maintenance while parallel units continue service, supporting adaptive management without full plant shutdown (Dotro et al., 2017).

3.6.1 Governance models, accountability, and community participation

Operational success in decentralized wastewater treatment is frequently constrained less by treatment process limits than by “soft infrastructure” failures: unclear asset ownership, fragmented responsibilities, weak tariff collection, and insufficient capability for routine O&M budgeting and staffing (Ross et al., 2014; von Sperling and Platzer, 2020). Governance therefore needs to be defined explicitly as a system of roles and incentives that ensures: (i) continuous funding for O&M; (ii) transparency and accountability for performance; (iii) enforceable compliance with discharge or reuse standards; and (iv) mechanisms for conflict resolution and adaptive upgrading over the project lifetime (Ross et al., 2014).

A practical starting point is to formalize: (1) who owns the physical assets (land, liners, media, inlet/outlet structures, pumps), (2) who operates the plant and with what competency profile, (3) who regulates performance (permits, inspections), and (4) who pays for routine O&M and periodic refurbishment (Ross et al., 2014). Decentralized sanitation governance assessments emphasize that clarifying ownership of assets enables a natural division of responsibility to manage and maintain them, while coordination among multiple actors across scales is often an enabling condition rather than optional (Ross et al., 2014). This is particularly relevant for TWs because refurbishment (e.g., solids removal, media rehabilitation) may occur at multi-year to decadal intervals—meaning that governance must remain functional longer than a typical political cycle (Dotro et al., 2017).

Community engagement matters for TWs in two distinct ways. First, it supports site acceptance and reduces the risk that decentralized infrastructure becomes socially marginalized. Second, it can strengthen the local capacity to notice and report emerging operational issues early (e.g., odors, ponding, bypass events) before they translate into chronic failure. Empirical governance scans indicate that participation in mobilizing and decision-making has been associated with better performance for community-managed sanitation systems in some settings (as summarized in Ross et al., 2014). However, participation must be supported by ongoing technical support; otherwise, local enthusiasm can quickly disappear, especially when technical and financial burdens are unrealistic.

A useful analogy comes from community renewable energy systems: research on community energy emphasizes that the distinctive value of “community” often lies in local ownership, participation, and benefit-sharing–features that can improve legitimacy and acceptance compared with externally owned infrastructure (Walker and Devine-Wright, 2008; Bauwens et al., 2016). Translating this logic to wastewater, a community-embedded governance design can pair (i) professionalized O&M (municipal or contracted operators), with (ii) community oversight (transparent reporting, participatory goal setting, grievance mechanisms), and (iii) local benefit articulation (use of treated water, improved public space, reduced pollution). This does not require treating wastewater as a high-margin commodity; rather, it frames TWs as a local public-value infrastructure whose benefits are shared and therefore worth co-financing.

Decision-support methods can help align stakeholder incentives before contracts are signed and concrete is poured. Participatory modeling approaches synthesize technical and social knowledge, supporting scenario testing of governance and financing options under uncertainty (Voinov and Bousquet, 2010; Voinov et al., 2016). Agent-based modeling and simulation (ABMS) has demonstrated value in energy-system policy assessment (Klein et al., 2019) and has been operationalized in open, agent-based electricity market frameworks developed at the German Aerospace Center, with documented use in renewable integration studies (Schimeczek et al., 2023). While wastewater differs in revenue logic–treated effluent value may be indirect rather than market-priced–ABMS can still be adapted to explore how heterogeneous actors (households, municipal utilities, regulators, farmers, and potential private partners) behave under different tariff, subsidy, enforcement, and benefit-sharing schemes, thereby reducing the risk of governance “blind spots” that later become operational failures (Galán et al., 2009; Voinov et al., 2016).

3.6.2 Financing, procurement, and risk allocation

Even when TWs offer low O&M intensity, decentralized municipalities may still lack stable funding streams for routine costs (sampling, desludging, vegetation management) and for periodic refurbishment. Guidance for treatment wetlands in resource-constrained regions emphasizes that construction costs must be manageable and O&M costs must be low “in order to guarantee that the plant will be sustainable in the long run and not become neglected because of lack of funds” (von Sperling and Platzer, 2020). In other words, a TW can be technically sound and still fail financially and institutionally.

The future-oriented framing benefits from using the “3Ts” concept (tariffs, taxes, and transfers) to describe the funding sources for water and sanitation services (World Bank, 2019). In decentralized wastewater, tariffs alone are often insufficient (or politically constrained), implying that stable O&M requires a realistic mix of user tariffs, domestic public budgets, and external transfers (e.g., state/federal programs, development assistance, climate and sustainability finance), with explicit affordability provisions for low-income households (World Bank, 2019; Leflaive and Hjort, 2020).

Wastewater treatment produces public goods (health protection, ecosystem protection) and therefore warrants co-financing beyond household-level payments. Procurement and delivery models range from fully public provision to hybrid models involving private participation. Public–private partnerships (PPPs) represent a spectrum of organization types that differ primarily in which functions are transferred and how risks and payments are structured. The World Bank PPP resource distinguishes, among others, concessions (often user-pay), Build–Operate–Transfer (BOT) (typically asset-specific and often government-pay), and Design–Build–Operate (DBO) (publicly financed assets with private design, construction, and operation) (World Bank PPP Resource Center, 2026). The same guidance emphasizes that long-term PPP terms are intended to create incentives for a “whole-life” approach, where design choices reflect future maintenance and operating costs rather than only lowest-capex procurement (World Bank PPP Resource Center, 2026).

However, water-sector PPP outcomes are mixed and strongly contingent on regulatory capacity, contract design, risk allocation, and tariff politics. A systematic review of PPP research in the water sector identifies risk management, contractual arrangements, financing and tariffs, infrastructure, and governance as dominant themes–highlighting that no single “best” risk matrix exists and that governance capacity is central to performance (Lima et al., 2021). An OECD background note similarly stresses the need to balance risks and focus on results, indicating that PPPs are not a universal remedy for financing gaps and frequently fail when incentives, accountability, and affordability are misaligned (OECD, 2025). Detailed O&M plans must be developed and contractually stipulated, as well as the required discharge quality standards. Agreements on significant bonuses and penalties for compliance with contractual obligations are also essential to ensure the long-term success of the projects.

Real-world remunicipalization cases illustrate why governance and financing must be treated as co-determinants of operational sustainability. In Paris, the creation of Eau de Paris in 2010 followed a remunicipalization of water service delivery; analyses describe how the earlier delegated management model involved private operators including Veolia and Suez, and how the post-2010 public utility pursued a cost-recovery model rooted in tariffs and transfers within the French water-financing ecosystem (Butzbach and Spronk, 2022). In Berlin, empirical case analyses document the partial privatization of the Water and sewage company in Berlin (Berliner Wasserbetriebe) in 1999 through a PPP involving RWE and then Vivendi/Veolia, followed by buybacks in 2012–2014 restoring full public ownership (Schaefer and Warm, 2014; Beveridge and Naumann, 2013). These privatization cases resulted very expensive for the water consumers of the French and German capitals. However, this does not imply that private participation is inherently infeasible; rather, they demonstrate that transparency, obligations, long-term reinvestment incentives, and penalties for non-compliance with contractual agreements must be contractually agreed upon.

For TWs specifically, procurement should therefore prioritize (i) performance-based specifications (effluent targets, uptime, monitoring), (ii) explicit O&M and refurbishment provisions, and (iii) transparency requirements that reduce information asymmetry between operators and communities. A DBO (or design–build–operate–maintain variant) can be attractive for TWs where public authorities can finance capex but seek private or specialized operational support—provided the contract includes measurable performance requirements, monitoring, and enforceable penalties/bonuses tied to compliance outcomes (World Bank PPP Resource Center, 2026; Lima et al., 2021). In contrast, under poorly designed “lowest-cost wins” procurement, private incentive structures can rationally under-invest in maintenance and capacity building, with performance deterioration appearing only after warranty periods, by which time the municipality bears the failure risk.

3.7 Economic and environmental benefits

Economic appraisal of treatment wetlands should be framed as life-cycle performance rather than as a capex-only comparison. Open-access guidance explains that economic assessment comprises capital costs (land acquisition and construction) and O&M costs (management and operation on a yearly basis), and that annualized cost concepts rely on net present value logic across the asset lifetime (Dotro et al., 2017). Importantly, capital costs are site-dependent and can vary widely by local labor markets, liner/media availability, and land prices. Land acquisition can become the dominant capex component in high land-value settings because wetlands generally require higher surface area than intensive technologies for comparable objectives (Dotro et al., 2017). This economic reality is central for large-scale TW planning: if land opportunity cost is high, the economic case can shift from “saving money” to “buying multiple benefits,” i.e., coupling treatment with green and blue infrastructure functions.

By contrast, O&M costs are intrinsically low for wetland systems because of mechanical simplicity and reduced electricity consumption, especially when hydraulic design allows gravity feed (Dotro et al., 2017). The same guidance identifies O&M cost components as including energy for pumps when necessary, compliance monitoring, maintenance of access roads/berms, and repair/replacement of mechanical parts; vegetation management costs depend on harvesting and pest control requirements.

At large scales, cost advantages are often reported, because conventional activated sludge facilities incur significant electricity, mechanical maintenance, and skilled labor burdens. An analysis across nine full-scale TW facilities reported that operation costs can be approximately 60% lower than conventional treatment plants when assessed over a 25–40-year horizon, illustrating economy-of-scale patterns and the role of reduced energy/mechanical demands (Gkika et al., 2014). In a 600 population-equivalent hybrid TW case (Greece), total construction cost and expected annual operating costs were reported explicitly (Gikas and Tsihrintzis, 2014). These empirical anchors support the general recommendation that TW feasibility should be evaluated using life-cycle cost analysis (LCCA) rather than purely capex benchmarks (Dotro et al., 2017).

A compact way to report LCCA is given in Equation 7:

LCC = C 0 + ∑ t = 1 T C O & M , t + C rehab , t 1 + r t − V T 1 + r T (7)

Where:

LCC = life-cycle cost (present value); C ₀ = initial capital cost (land + construction); C _O&M,t = operating and maintenance cost in year t; C _*rehab,t* = rehabilitation/refurbishment cost in year t (e.g., solids removal, media restoration); V _*T* = residual value at end of analysis horizon; r = discount rate; T = analysis horizon (years).

Source concept: life-cycle economics framing and NPV logic for treatment wetlands are discussed in Dotro et al. (2017).

Framing benefits in terms of affordability requires linking LCCA to funding sources (“3Ts”) and to distributional outcomes. The World Bank emphasizes that WASH financing (Water Supply, Sanitation, and Hygiene) ultimately comes from tariffs, taxes, and transfers, and that tariffs, while crucial for long-term reliability, often require complementary taxes/transfers for sustainable cost recovery and affordability protection (World Bank, 2019). This underlines that TWs reduce the recurrent cost burden per unit of treatment but do not eliminate the need for stable O&M funding architecture.

3.7.1 Energy demand, greenhouse gases, and environmental trade-offs

TWs are often categorized as low-energy, nature-based wastewater treatment because they rely primarily on microbial degradation, plant-mediated processes, and filtration/sorption within the substrate, thereby minimizing continuous aeration and other electricity-intensive unit operations typical of activated sludge plants (Dotro et al., 2017; Garfí et al., 2026). This advantage matters because electricity can form a major fraction of operating costs in conventional plants; a Frontiers review of Waste Water Treatment Plant (WWTP) energy efficiency reports typical electricity consumption ranges on the order of 0.3–2.1 kWh per cubic meter of treated wastewater, depending on plant size and configuration, highlighting why energy reduction is economically meaningful (Gandiglio et al., 2017).

From a life-cycle perspective, LCAs frequently find that treatment wetlands have lower environmental impacts than conventional systems, largely due to low energy requirements and limited chemical use (Fuchs et al., 2011; Garfí et al., 2026). Nevertheless, it should also be acknowledged that TWs can emit methane (CH₄) and nitrous oxide (N₂O), meaning that climate benefits are not guaranteed without design and operational attention to redox management, substrate conditions, and loading (Yu et al., 2023; Li et al., 2024). In other words: TWs reduce indirect emissions by consuming less electricity, while potentially increasing direct biogenic emissions under certain conditions, an important trade-off to report transparently (Fuchs et al., 2011; Yu et al., 2023).

A concise accounting framework for reporting TW climate performance is given in Equation 8:

GHG total = E el · E F el + m CH 4 · GWP CH 4 + m N 2 O · GWP N 2 O (8)

Where:

GHG _*total* = total greenhouse gas impact (CO _2-*eq*); E _el = electricity consumed; EF _el = electricity emission factor; m _*CH*4, m _*N*2*O* = methane and nitrous oxide emissions (mass basis); GWP = global warming potential factors used for conversion to CO _2-*eq*.

Source concept: inclusion of direct wetland GHG emissions in CW LCA comparisons is emphasized in Fuchs et al. (2011), and methane-specific mitigation mechanisms are reviewed by Yu et al. (2023).

3.7.2 Water reuse, ecosystem services, and multifunctional value creation

Large-scale TWs can create “value” not only by lowering treatment costs, but by enabling reuse and by providing ecosystem services that strengthen the societal cost–benefit case. Practical guidance in wetland technology emphasizes multi-objective design. E.g., if the design objective includes irrigation reuse, nutrient removal targets should be aligned with the reuse purpose rather than automatically minimizing nutrients at all costs (Langergraber et al., 2020). Health-protective reuse must, however, follow risk-based frameworks; the World Health Organization guidelines for safe wastewater use in agriculture emphasize microbial risk assessment and context-specific health-based targets, providing a defensible structure for discussing reuse without claiming zero risk (WHO, 2006).

In regions facing water scarcity, the ability to reuse treated effluent for irrigation of green areas, agriculture, or managed aquifer recharge can function as a key co-benefit and can improve local willingness-to-pay and political support. Especially, when wastewater treatment is otherwise perceived solely as a cost center this becomes a decisive issue (WHO, 2006). Early evidence from ornamental wetland plant studies also suggests that the prospect of locally visible and economically valuable co-products can support adoption: a classic study from Mexico argued that rural communities may prioritize wastewater treatment only when they can derive economic benefit from improved water resources, motivating integration of ornamental plants such as Canna indica and related species into TW designs (Belmont et al., 2003). More recent literature synthesizes ornamental-plant TW applications and notes that ornamental flowering plants can improve aesthetic value while maintaining treatment performance, potentially increasing acceptance and providing ancillary economic opportunities (Karungamye, 2022; García-Ávila et al., 2023). The economic value of ornamental wetland plants must be verified case by case, because harvested biomass from wastewater systems may carry odor, contaminant, or marketability constraints. At large scales, biomass valorization may become relevant, but it should be treated as a conditional co-benefit requiring quality control, contaminant assessment, realistic off-taker demand, and a documented processing pathway.

Beyond reuse, TWs can deliver ecological services such as habitat creation and biodiversity support. They may also generate recreational and landscape benefits when integrated into urban and peri-urban green infrastructure. The large-scale wetland in Rayón, Mexico, designed by Janisch & Schulz Engineers, illustrates this multifunctional potential by combining wastewater treatment with habitat creation, a biotope observation platform, and a jogging trail, as illustrated in Figure 4. Such design elements can strengthen local acceptance and make nature-based wastewater treatment visible and accessible to the public.

4 Discussion

4.1 Large TWs as socio-technical wastewater infrastructure

The central contribution of this study is to reposition large treatment wetlands (TWs) from isolated wastewater-treatment units to socio-technical infrastructure systems. Existing TW literature often emphasizes pollutant removal, hydraulic loading, substrate selection, vegetation, or operation as separate topics. The synthesis developed here shows that large-scale TW success depends on the interaction among the assessed six dimensions. This supports the broader transition in decentralized wastewater research, where decentralized systems are increasingly understood as locally embedded infrastructure arrangements shaped by spatial, institutional, and governance conditions (Reymond et al., 2020; Garrido-Baserba et al., 2024; Sha et al., 2024).

This distinction matters because large TWs are not simply upscaled small wetlands. Once systems are designed for municipal, peri-urban, industrial, agricultural, or stormwater-related applications, the planning problem changes qualitatively. Higher and more variable flows increase the importance of influent characterization, hydraulic distribution, stormwater intrusion assessment, pretreatment, sludge management, clogging prevention, and modular operation. In this sense, the design basis is the first treatment process. If flow, pollutant load, land availability, hydrogeology, and institutional responsibility are poorly characterized, later optimization of media, plants, or wetland type cannot fully compensate for the initial error. This confirms previous concerns that land demand, clogging, and long-term hydraulic dysfunction remain decisive planning constraints for TW implementation (Fan et al., 2021; Pucher and Langergraber, 2019; Mishra et al., 2024).

4.2 Scale changes the meaning of decentralization

A second implication concerns the concept of decentralization itself. Decentralized wastewater treatment is often associated with households, small settlements, or remote rural communities. The cases reviewed here suggest a broader interpretation. A system may remain meaningfully decentralized even when serving several thousand or tens of thousands of people, provided that it treats wastewater near the point of generation, reduces dependence on long conveyance networks, remains embedded in a local service territory, and is governed through local or regional accountability structures. This interpretation aligns with recent decentralized-water literature emphasizing resilience, modularity, local adaptation, and distributed infrastructure rather than size alone (Reymond et al., 2020; Garrido-Baserba et al., 2024).

This is especially relevant for TWs because their comparative advantage is often strongest where conventional centralized expansion is constrained by cost, land, energy demand, topography, or institutional fragmentation. However, larger decentralized systems also face higher expectations. A large TW is public infrastructure. It must meet discharge or reuse standards, withstand seasonal loading variability, provide operational continuity, and remain governable beyond political cycles. Large-scale decentralization therefore requires a different planning standard than small community systems: it must combine ecological treatment logic with infrastructure-grade design, monitoring, financing, and institutional accountability.

4.3 Configuration-specific performance matters more than generic wetland labels

The decisive variable in treatment performance is the treatment-train sequence in relation to wastewater source and treatment objective. Vertical-flow stages support oxygen transfer, organic matter removal, and nitrification; horizontal subsurface-flow stages provide anoxic zones favorable for denitrification and polishing; and free water surface systems can contribute to buffering, suspended-solids removal, ecological integration, and tertiary polishing (Vymazal, 2010; Dotro et al., 2017; Parde et al., 2021).

This is why the 18 detailed configuration groups are more meaningful than a single aggregated “large TW” average. The strongest balanced municipal signals were observed in staged municipal/tourism VF→HF(/FWS) hybrids and selected pretreated VF/VSSF systems, while HSSF systems after primary treatment provided the broader and statistically more useful municipal evidence base. At the same time, reclaimed-water polishing systems, CSO filters, agricultural phosphorus-control wetlands, and river/lake restoration wetlands must be interpreted according to their specific objectives. A polishing wetland treating diluted secondary effluent cannot be judged by the same percentage-removal logic as a municipal secondary-treatment wetland treating raw or primarily treated wastewater. Low percentage removal under low influent concentrations may still represent effective polishing, while high percentage removal in event-based systems may not be transferable to continuously loaded municipal wastewater treatment.

The practical question is therefore, which treatment sequence works for which influent, site, regulatory target, climate, land constraint, and O&M capacity. In this sense, the framework advances the literature by linking treatment performance to configuration logic.

4.4 Governance is a performance condition

A major finding of this study is that operational governance should be treated as a performance condition. TWs are often described as low-input systems, but low-input does not mean input-/maintenance-free. Clogging, hydraulic short-circuiting, vegetation decline, inlet blockage, sludge accumulation, insufficient monitoring, and lack of operator training can progressively undermine performance even when the original design is technically sound (Dotro et al., 2017; Pucher and Langergraber, 2019; Langergraber et al., 2020).

For large TWs, governance becomes decisive because responsibility is distributed across actors: landowners, municipalities, operators, regulators, funders, communities, and sometimes private partners. If ownership, O&M budgets, monitoring responsibilities, sludge management, and refurbishment obligations are not defined and contractually secured before construction, the system is structurally vulnerable. This is particularly relevant in decentralized sanitation, where technically feasible systems often fail because institutional arrangements are weaker than engineering designs (Ross et al., 2014; von Sperling and Platzer, 2020).

The framework therefore supports a shift from “build-and-transfer” to “design-build-operate-maintain” thinking. Contractual models such as DBO or performance-based PPP arrangements can be useful when they specify effluent targets, monitoring frequency, maintenance obligations, response times, refurbishment responsibilities, and penalty/bonus mechanisms. However, PPPs are not a universal remedy. Water-sector evidence shows that private participation fails when risk allocation, tariff politics, transparency, and enforcement capacity are poorly designed (Lima et al., 2021; OECD, 2025). Thus, the relevant question is whether the governance model creates durable incentives (penalties and bonuses) for compliance, maintenance, reporting, reinvestment, and public accountability and thus creates win-win-situations for all stakeholders in the case of continuous outperforming of effluent requirements.

4.5 Regulatory transition, micropollutants, and multifunctionality

The regulatory context is also changing the role of TWs. Conventional indicators such as BOD₅, COD, TSS, TN, and TP remain central, but newer policy frameworks increasingly emphasize nutrient control, water reuse, micropollutant attenuation, energy neutrality, and extended producer responsibility. The recast EU Urban Wastewater Treatment Directive illustrates this shift by strengthening nutrient-removal requirements, introducing quaternary-treatment obligations for selected agglomerations, and setting removal expectations for organic micropollutant indicators (European Union, 2024).

This does not automatically mean that TWs should be designed without ozonation, activated carbon, membranes, or other advanced treatment technologies. Such a claim would overstate the evidence. A more defensible interpretation is that TWs can function as essential components of multi-barrier treatment strategies. For emerging contaminants, the evidence indicates partial and compound-specific attenuation driven by sorption, biodegradation, plant uptake, photodegradation, redox zoning, and media selection (Li et al., 2014; Verlicchi and Zambello, 2014; Ávila et al., 2015; Ilyas and van Hullebusch, 2020a; Ilyas and van Hullebusch, 2020b; Overton et al., 2023; Hou et al., 2025).

Similarly, multifunctionality should be treated as a design outcome. Large TWs may support biodiversity, recreation, landscape integration, heat mitigation, reuse, biomass production, and environmental education, but these benefits are realized only if they are planned, financed, monitored, and socially accepted. Economic evaluation should therefore move beyond capital expenditure and include life-cycle cost, land opportunity cost, O&M, monitoring, media renewal, sludge and biomass management, water reuse value, avoided energy use, and ecosystem-service potential (Dotro et al., 2017; Gkika et al., 2014). Environmental evaluation must also remain honest about trade-offs. TWs may reduce electricity demand relative to activated sludge systems, but they may also emit CH₄ and N₂O depending on redox conditions, organic loading, vegetation, and management (Fuchs et al., 2011; Yu et al., 2023; Li et al., 2024).

The framework guides practitioners to move from “treatment as a cost center” to “resource recovery hub” by systematically evaluating pathways for water reuse, biomass valorization, and biodiversity credits, and then linking these potential revenue streams back to the governance and financing model to ensure their realization. Coupled with circular economy approaches, harvested wetland biomass could be valorized into biochar, compost, or low-grade biofuels, while treated effluent supports urban agriculture or managed aquifer recharge. With an integral point of view, the byproducts of TWs (wastewater contaminants-absorbing biomass and gaseous emissions) must be integrated into a circular economy as well. Through further biotechnological processing of these residues, industrial raw materials are supposed to be recovered in a grave-to-cradle approach, to complement the cradle-to-grave industries in their water issues and close the matter cycles.

In short, TWs should be framed as potential “treatment-plus” infrastructures: they can create additional public value, but only when co-benefits are converted into measurable objectives, assigned responsibilities, and financing mechanisms in a Closed-Cycle Economy.

4.6 Limitations and future research

This study has several limitations. First, the review was structured and evidence-based, but not a fully PRISMA-compliant systematic review. Second, the performance synthesis is constrained by heterogeneous reporting formats, monitoring periods, wastewater sources, climates, treatment objectives, and nitrogen metrics. For this reason, the results are best interpreted as comparative configuration-specific evidence. Third, several configuration groups are supported by only one or two project-level observations. These cases are valuable as lighthouse examples, but they should not be overgeneralized as statistically stable benchmarks.

Fourth, emerging-contaminant as well as MFC and MEC evidence remains uneven and often comes from pilot, laboratory, or small-scale studies. The compound-group matrix should therefore be used as a design-oriented synthesis of removal tendencies and mechanisms. Fifth, governance, life-cycle cost, ecosystem services, and social acceptance remain less standardized than conventional effluent parameters. Future research should therefore develop comparable indicators for O&M reliability, governance continuity, asset ownership, tariff/tax/transfer financing structures, ecosystem-service valuation, and long-term climate performance.

Future work should also test the decision-support matrix through real project applications. This could include before–after analysis of TW planning processes, participatory modeling with municipalities and operators, agent-based simulation of governance and financing scenarios, and coupling hydraulic models with life-cycle cost and greenhouse-gas accounting. Large-scale TW operators should also have the legal responsibility to assess their micropollutant removal efficiencies in long-term studies, to evaluate to what extent which part of the TW will remove which pollutant with a high certainty. MFC and MEC TW have to demonstrate whether they will play an essential future role in performance-enhancing and micropollutant removal. Good governance guidelines and best-practice examples for post treatment of the byproducts of treatment wetlands aiming at a closed-loop circular economy need to be invented and investigated. Such research would move the field from descriptive case reporting toward predictive and decision-oriented planning support.

4.7 Planning implication

The main implication is that large TWs should be planned as long-term socio-technical infrastructure. Their success depends on three linked questions: What hydraulic and pollutant reality must be treated? Which staged treatment train is appropriate for that reality? And which governance and financing structure will keep the system functional for decades? The proposed framework does not replace detailed engineering design, but it helps identify whether a project is ready for such design. In this sense, it provides a bridge between wetland science, infrastructure planning, and water governance.

5 Conclusion

This study developed an integrated decision-support framework for large treatment wetlands (TWs) as decentralized wastewater infrastructure. Its main contribution is to shift the planning perspective from TWs as isolated treatment units toward TWs as lifecycle-based socio-technical systems. The synthesis shows that successful large-scale implementation depends on the coordinated interaction of six dimensions: site and environmental factors, design principles, technological configuration, performance and efficiency, operational governance, and economic/environmental benefits.

The comparative assessment of 64 large-scale TW projects further demonstrates that treatment performance must be interpreted according to wastewater source, treatment objective, and treatment-train sequence. The 18 configuration groups developed in this study provide a more precise basis for comparing large TWs and show that staged systems, especially well-designed VF/HF combinations and pretreated VF/VSSF systems, can achieve strong conventional pollutant removal when supported by appropriate pretreatment, hydraulic design, monitoring, and O&M structures.

Beyond treatment efficiency, the study highlights governance as a core performance condition. Risks and responsibilities for the whole lifetime of the project must be clearly allocated and linked to penalty/bonus regulations to motivate the actors to continuous best performance.

The developed framework is intended to translate this insight into practice. It supports early identification of no-go risks, structures trade-offs between land, treatment targets, reuse, micropollutant polishing, climate performance, and multifunctional value creation, and links technical design to governance and financing requirements. Large TWs should be planned in an integrative approach to become long-term success stories.

Generative AI statement

The author(s) declared that generative AI was used in the creation of this manuscript. During manuscript preparation, generative AI tools were used to support language editing, structural refinement, condensation of text, table organization, data-extraction assistance, consistency checking, and drafting of analytical summaries. The tools used by AR included OpenAI ChatGPT, including GPT-5.4 and GPT-5.5 Thinking versions; DeepSeek, current version available at the time of use; and J.d.A. used Microsoft Copilot tools during manuscript preparation. Earlier OpenAI ChatGPT versions available around March 2025 were also used during the initial large-scale treatment wetland project search and screening phase. Generative AI tools were used as assistive instruments for handling, structuring, checking, and summarizing large volumes of literature- and project-derived information. They were not used as authors, did not replace scientific judgment, and did not generate primary experimental data. All AI-assisted outputs, including text, tables, extracted values, calculations, interpretations, citations, and references, were critically reviewed, corrected, verified, and approved by the human authors. The authors take full responsibility for the accuracy, integrity, originality, and scientific content of the manuscript and Supplementary Material.

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Supplementary material

The Supplementary Material for this article can be found online at: https://www.frontiersin.org/articles/10.3389/fenve.2026.1857163/full#supplementary-material

Complete decision-support matrix for planning large decentralized treatment wetlands. This table expands the compressed matrix presented in the manuscript and includes framework dimensions, key planning questions, core indicators, red flags, planning outputs, and readiness-score interpretation. Supplementary material at the source

Configuration key and synthesis families for large-scale treatment wetland performance assessment. This table documents the classification logic used to assign projects to detailed configuration groups and broader synthesis families. Supplementary material at the source

Project-level database of 64 large-scale treatment wetland cases. This table includes location, wastewater source, inflow, population equivalent, wetland area, treatment train, wetland type, plants, pretreatment, post-treatment, conventional pollutant removal, evidence notes, special features, and source references. (Separate Excel-Table: “Table S3. Large-scale TWs.xlsx”). Supplementary material at the source

Emerging-contaminant removal evidence matrix. This table summarizes compound groups, individual substances, treatment wetland configurations, removal efficiencies, dominant removal mechanisms, substrate/media effects, plants, hydraulic retention time, design-sensitive parameters, and planning implications. Supplementary material at the source

Supplementary material Supplementary material at the source

Supplementary material Supplementary material at the source

Glossary

AS Activated sludge BOD _**5** Biochemical oxygen demand after 5 days CAPEX Capital expenditure COD Chemical oxygen demand CSO Combined sewer overflow DBO Design–Build–Operate DBO(M) Design–Build–Operate–Maintain EDC Endocrine-disrupting compound FWS Free water surface GHG Greenhouse gas GWP Global warming potential HF Horizontal subsurface flow HLR Hydraulic loading rate HRT Hydraulic retention time HSSF Horizontal subsurface-flow ImT Imhoff tank IQR Interquartile range IVCW Integrated vertical-flow constructed wetland LCA Life-cycle assessment LCCA Life-cycle cost analysis LECA Lightweight expanded clay aggregate MEC Microbial electrolysis cell MFC Microbial fuel cell N.D. No data O&M Operation and maintenance PE Population equivalent PPCPs Pharmaceuticals and personal care products PPP Public–private partnership RWR Reclaimed water reservoir SF Surface flow SSF Subsurface flow ST Septic tank TN Total nitrogen TP Total phosphorus TSS Total suspended solids TW Treatment wetland USEPA United States Environmental Protection Agency UWWTD Urban Waste Water Treatment Directive VF Vertical flow VSSF Vertical subsurface-flow WASH Water supply, sanitation, and hygiene WWTP Wastewater treatment plant

Footnotes

Edited by: Muhammad Bilal Asif, King Abdullah University of Science and Technology, Saudi Arabia

Reviewed by: Rosana Boglione, UTN, Argentina

Cangxiang Wen, Southeast University, China