Ecological Functions of Microbes in Constructed Wetlands for Natural Water Purification

Abstract

Constructed wetlands (CWs) are sustainable and cost-effective systems that utilise plant–microbe interactions and natural processes for wastewater treatment. Microbial communities play a pivotal role in pollutant removal by crucial processes like nitrogen transformations, phosphorus cycling, organic matter degradation and the breakdown of emerging contaminants. Dominant phyla, such as Proteobacteria, Bacteroidetes, Actinobacteria and Firmicutes, collectively orchestrate these biogeochemical functions. Advances in molecular tools, including high-throughput sequencing and metagenomics, have revealed the diversity and functional potential of wetland microbiomes, while environmental factors, i.e., temperature, pH and hydraulic retention time, strongly influence their performance. Phosphorus removal efficiency is often lower than nitrogen, and large land requirements and long start-up times restrict broader application. Microplastic accumulation, the spread of antibiotic resistance genes and greenhouse gas emissions (methane, nitrous oxide) present additional challenges. The possible persistence of pathogenic microbes further complicates system safety. Future research should integrate engineered substrates, biochar amendments, optimised plant–microbe interactions and hybrid CW designs to enhance treatment performance and resilience in the era of climate change. By acknowledging the potential and constraints, CWs can be further developed as next-generation, nature-based solutions for sustainable water management in the years to come.

1. Introduction

The increasing global water crisis intensifies the urgent need for sustainable and cost-effective wastewater treatment technologies that can address current demands as well as future environmental challenges [1,2]. Conventional water treatment systems face challenges from rising energy costs, complex operational requirements and emerging contaminants, highlighting the need for natural and sustainable approaches [3]. Constructed wetlands are a probable solution [4]. Constructed wetlands have gained prominence as viable alternatives that harness natural processes for comprehensive water purification [5]. The microbial ecology of constructed wetlands encompasses an immense diversity of bacteria, archaea, fungi and other microorganisms that collectively orchestrate pollutant degradation, nutrient cycling and ecosystem functioning across various spatial and temporal scales [6]. These engineered ecosystems represent a paradigm shift toward nature-based solutions that merge ecological principles with engineering design to create controlled environments optimised for pollutant removal [7]. Constructed wetlands function through synergistic interactions between three primary components: macrophytic vegetation, substrate media and diverse microbial communities [8]. Among these components, microbial communities serve as the biological engines driving the complex biogeochemical transformations essential to effective water purification [9]. An ideal constructed wetland for cleaning wastewater is shown in Figure 1.

Recent advances in molecular biology and omics technologies have revolutionised our understanding of microbial communities in constructed wetlands, enabling researchers to move beyond traditional culture-based approaches to comprehensive ecosystem-level analysis of microbial structure, function and dynamics [10]. The significance of microbial ecology in constructed wetlands extends beyond basic pollutant removal to cope with emerging challenges, such as antibiotic resistance gene management, climate change adaptation and the treatment of micropollutants [11,12]. The spatial and temporal dynamics of these microbial communities directly influence the efficiency and stability of pollutant removal processes, making their study crucial for developing resilient and sustainable water treatment solutions [13,14]. Understanding the relationship between microbial communities, environmental conditions and treatment performance is essential for optimising the constructed wetland design and operation. In this review, we discuss the comprehensive analysis of current knowledge regarding microbial ecology in constructed wetlands, exploring fundamental mechanisms through which microbial communities drive natural water purification processes while identifying critical knowledge gaps and future research directions.

2. Microbial Community Structure and Diversity

2.1. Taxonomic Composition and Dominant Phyla

The taxonomic dominance of bacterial populations reflects the metabolic versatility required to process diverse pollutants and the adaptation to the dynamic environmental conditions characteristic of constructed wetland systems. The microbial communities in constructed wetlands exhibit remarkable diversity and complexity. The bacterial populations in constructed wetlands are typically dominated by different major phyla (Figure 2), which collectively account for 60–80% of total bacterial abundance across different wetland configurations and operational conditions [10].

Proteobacteria emerge as the most abundant bacterial phylum in constructed wetlands. Their population ranges from 25 to 48% of the total bacterial communities, depending on system configuration and operational parameters [10]. The dominance of proteobacteria reflects their metabolic versatility and adaptation to the fluctuating redox conditions that characterise constructed wetland environments. The role of this phylum in constructed wetlands is very diverse. This phylum consists of diverse functional groups, including ammonia-oxidising bacteria (such as nitrosomonas and nitrobacter), denitrifying bacteria (such as pseudomonas and dechloromonas), and sulfur-cycling bacteria that help trace metal immobilisation.

Gamma proteobacteria within the proteobacteria play critical roles in nitrogen cycling and organic matter degradation. The most common members of this class are Pseudomonas and Acinetobacter; they possess diverse metabolic capabilities that enable them to utilise various carbon sources. The other class, Betaproteobacteria, includes key nitrifying bacteria, such as Nitrosomonas and Nitrosospira. They are essential for ammonia oxidation in constructed wetlands. The other class, Alphaproteobacteria, contributes to the degradation of complex organic compounds and plant–microbe interactions in the rhizosphere [10].

Bacteroidetes is the second-most abundant phylum. Their population ranges from 10 to 28% of bacterial communities. Bacteroidetes are particularly important in the initial stages of organic matter processing, breaking down complex polymers into simpler compounds that other microbial groups can utilise, thereby playing a crucial role in the degradation of complex organic matter and cellulose decomposition [16]. This phylum includes specialised bacteria that are associated with the breakdown of plant litter and other recalcitrant organic compounds, contributing significantly to overall carbon cycling in wetland ecosystems.

The Actinobacteria phylum accounts for 5–15% of bacterial communities. Their primary role is to degrade complex organic compounds and metabolise antibiotics and other emerging contaminants [9]. Recent studies have highlighted the importance of Actinobacteria in the degradation of emerging contaminants, with specific genera showing enhanced capabilities for antibiotic metabolism [10]. This phylum produces diverse secondary metabolites and enzymes capable of breaking down recalcitrant compounds, making them particularly valuable for treating wastewater containing pharmaceuticals, personal care products and other persistent organic pollutants.

The Firmicutes phylum varies significantly with environmental conditions, particularly temperature and organic loading, ranging from 4 to 12% of total bacterial communities [13]. Firmicutes contribute to the initial breakdown of complex organic matter through fermentation, producing intermediate compounds that serve as substrates for other microbial groups in the treatment process [10]. This phylum includes important anaerobic bacteria that are involved in fermentation and the production of organic acids.

Table 1. Dominant microbial phyla in constructed wetlands.

Phylum	Abundance (%)	Major Role	Representative Genera	Source
Proteobacteria	25–48.7	N cycling, P removal, S cycling, COD oxidation	Pseudomonas, Nitrosomonas, Acinetobacter	[10]
		Consists of diverse functional groups like ammonia-oxidising, denitrifying and sulfur-cycling bacteria.	Gamma-proteobacteria (Pseudomonas, Acinetobacter), Betaproteobacteria (Nitrosomonas, Nitrosospira) and Alpha-proteobacteria	[10]
Bacteroidetes	10–28.2	Cellulose and protein hydrolysis	Flavobacterium, Cytophaga	[10]
		Crucial for the initial breakdown of complex organic matter and cellulose decomposition.	Specialised bacteria are involved in the breakdown of plant litter and other recalcitrant organic compounds.	[16]
Actinobacteria	5–15	Degradation of complex organic compounds	Streptomyces, Mycobacterium	[9]
		Metabolise antibiotics and other emerging contaminants.	Specific genera enhanced capabilities for antibiotic metabolism.	[10]
Firmicutes	4.1–20	Fermentation, volatile fatty acid production	Clostridium, Bacillus	[13]
		Contribute to the initial breakdown of complex organic matter through fermentation.	Includes important anaerobic bacteria involved in fermentation processes and organic acid production.	[10]
Others	1–8	Nitrification, P storage, biofilm stability	Nitrospira, Gemmatimonas, Chloroflexi	[10]

summarises the dominant bacterial phyla commonly found in constructed wetlands and highlights their relative abundance, primary ecological functions, and representative genera. The category labelled ‘others’ includes phyla that play supportive but significant roles in nitrification, phosphorus storage and biofilm stability within wetland systems.

2.2. Spatial Distribution and Environmental Gradients

The microbial communities in constructed wetlands exhibit distinct spatial patterns that reflect environmental gradients and ecological processes operating at different scales within the system [17]. These biogeographic patterns are crucial for understanding treatment mechanisms and for optimising system design to achieve maximum efficiency and stability. Horizontal gradients across constructed wetlands, from the point where water enters to the point where it exits, typically show systematic changes in microbial diversity and community composition as pollutant concentrations decline and environmental conditions shift [16]. At the beginning of treatment, when nutrient concentrations are high, fast-growing bacteria (copiotrophs) tend to grow more. At the same time, later stages support more oligotrophic populations capable of scavenging residual nutrients. These gradients reflect the successional dynamics of microbial communities as they process pollutants and create increasingly oligotrophic conditions downstream.

Vertical stratification within the wetlands is important. The substrate profiles create distinct microbial zones corresponding to redox gradients and oxygen availability [18]. The upper layer, enriched by plant roots and air exposure, provides more oxygen and supports aerobic microbes. In deeper layers, oxygen fades away, and microbes that live without it, such as those involved in fermentation or methane production, become more common. The overall functional capacity of the microbial community is influenced by the thickness and stability of these zones and also determines the types of biogeochemical processes that can occur. Rhizosphere effects create localised hotspots (high-density regions) of microbial activity and diversity around plant roots, with microbial biomass 2–10-fold higher than in bulk sediments [18].

3. Biogeochemical Processes and Microbial Functions

3.1. Nitrogen Transformation Pathways

Nitrogen cycling in constructed wetlands consists of complex microbial-mediated transformations that require specific environmental conditions and microbial populations [19]. These transformations help in achieving substantial nitrogen removal from wastewater through coordinated processes operating under varying redox conditions [20]. The efficiency of nitrogen removal depends on the successful coordination of multiple microbial processes. Ammonification is the initial step in nitrogen processing [21]. Heterotrophic bacteria convert organic nitrogen compounds to ammonia using extracellular enzymes that break down proteins, nucleic acids, and other nitrogenous organic compounds [22]. The rate of ammonification depends on the ambient air temperature, pH, carbon availability and microbial community composition [23]. The optimal conditions for this typically occur in moderate anaerobic environments where organic matter decomposition is active [24].

Nitrification is a two-step process where ammonium ions are oxidised to nitrite by ammonia-oxidising bacteria, followed by oxidation of nitrite to nitrate by nitrite-oxidising bacteria [25]. Bacteria, mostly nitrosomonas and nitrosospira, catalyse the oxidation of ammonia into nitrite, followed by other bacteria, including Nitrobacter and Nitrospira, that oxidise nitrite to nitrate [26]. Additionally, recent studies have confirmed the presence of ammonia-oxidising archaea and comammox bacteria in constructed wetlands, which are essential for understanding the nitrification community. Nitrifying bacteria are taxonomically and physiologically diverse, but their activity is closely linked to spatial distribution, which is strongly influenced by oxygen availability [27]. Changes in temperature and plant growth activities can affect nitrification rates, leading to the highest levels during warmer seasons when plants are moving oxygen reserves most effectively.

Denitrification is the dominant nitrogen removal process in constructed wetlands, in which nitrate is stepwise reduced to nitrogen gas under anaerobic conditions [5]. Denitrification is a distinct metabolic process that requires organic carbon as an electron donor and proceeds through successive reduction steps mediated by different bacterial communities, and it also depends on the availability of suitable electron donors, which may limit denitrification efficiency in constructed wetlands [28]. A feature of denitrifying bacteria is their pronounced metabolic flexibility, with plant exudates, dead plant material, and wastewater organic compounds among the best carbon sources [29]. Denitrifying bacteria are diverse, and some, such as those in the genera pseudomonas, paracoccus or thauera, are facultative anaerobes that allow complete denitrification of nitrates to nitrogen gas [30]. The complete nitrogen cycle is shown in Figure 3. All the Nitrogen removal mechanisms and key microbial players are shown in

Table 2. Nitrogen removal mechanisms and key microbial players.

Process	Condition	Key Microbial Genera	Pathway Description	Source
Ammonification	Anaerobic	Bacillus, Clostridium	Organic N → NH4+	[22]
		Heterotrophic bacteria convert organic nitrogen to ammonia through extracellular enzymes.	Optimal conditions occur in moderate anaerobic environments.	[23]
Nitrification	Aerobic	Nitrosomonas, Nitrobacter	NH4+ → NO2− → NO3−	[25]
		Includes ammonia-oxidising archaea and comammox bacteria.	Two-step process: oxidation of ammonium to nitrite and then nitrite to nitrate.	[26]
Denitrification	Anaerobic	Pseudomonas, Paracoccus	NO3− → NO2− → NO → N2O → N2	[5]
		Also includes Thauera.	Stepwise reduction of nitrate to nitrogen gas under anaerobic conditions using organic carbon as an electron donor.	[29]
Anammox	Anaerobic	Brocadia, Kuenenia	NH4+ + NO2− → N2 + H2O	[20]

.

3.2. Phosphorus Cycling and Removal Mechanisms

The remediation of phosphorus in constructed wetlands is mediated by abiotic and biotic processes in which microbial communities directly contribute to phosphorus transformation, immobilisation and cycling [1]. Phosphorus does not have a gaseous phase under environmental conditions, unlike nitrogen, making its permanent removal much more complicated and management more complex [31]. Phosphorus is removed biologically via phosphorus-accumulating organisms that take up phosphorus and store it as polyphosphate granules under alternating anaerobic and aerobic conditions [32]. This cyclical phosphorus absorption and discharge at the cellular level is mainly carried out by specific Proteobacteria phyla, with important genera including Candidatus Accumulibacter, Tetrasphaera, and some Gamma-proteobacteria, which play crucial roles in enhanced biological phosphorus removal processes.

In microbial reactions, precipitation could occur, immobilising phosphorus in sediments via mineral phases formed by bacterial metabolism [33]. Under anaerobic conditions, iron-reducing bacteria promote the formation of iron–phosphorus complexes, and re-oxidation may lead to stable iron–phosphate precipitates [34]. It is possible due to the contribution of sulfate-reducing bacteria, which form a long-term storage mechanism through metal sulfides (which can bind phosphorus). Bacteria specialised in phosphorus solubilisation and remobilisation can release previously immobilised phosphorus, especially when redox conditions change [35]. These bacteria release organic acids and extracellular enzymes that solubilise mineral phosphates, making it essential to keep proper environmental conditions for stable phosphorus elimination over a long operational time [36] (Figure 4). Phosphorus removal mechanisms in constructed wetlands are presented in

Table 3. Phosphorus removal mechanisms in constructed wetlands.

Mechanism	Description	Microbial Involvement	Source
Biological P removal	Polyphosphate accumulation via PAOs	Accumulibacter, Tetrasphaera	[32]
Microbial-induced precipitation	Formation of iron-, aluminium- or calcium-bound P	Iron-reducing, sulfate-reducing bacteria	[37]
Solubilisation/Remobilisation	Release of stored P under changing redox/pH	Bacillus, Aspergillus	[35]

.

3.3. Carbon Cycling and Organic Matter Processing

Carbon cycling in constructed wetlands involves complex interactions between autotrophic and heterotrophic microbial processes that determine the fate of organic matter and affect overall system performance [32]. Due to the diverse carbon sources and metabolic flexibility of microbial communities, constructed wetlands efficiently treat various organic pollutants and play an essential role in carbon sequestration. Specifically, aerobic organic matter degradation primarily occurs in well-aerated regions, where heterotrophic bacteria utilise dissolved oxygen as the terminal electron acceptor for rapid and complete mineralisation of biodegradable organic compounds [38]. Initial treatment stages are critical for the removal of indicators like biochemical oxygen demand (BOD) and chemical oxygen demand (COD), in which several aerobic bacteria, such as pseudomonas and acinetobacter, play an essential role in organic matter degradation [39]. In oxygen-limited zones, anaerobic degradation pathways predominate, with sequential fermentation, acetogenesis and methanogenesis steps [40]. Fermentative bacteria break down more complex organic compounds into less complex organic acids and alcohols, and acetogenic bacteria convert these intermediates to acetate to be used for methanogenic archaea [41]. Efficiency is based on interspecific product exchange between groups in a syntropic association and on environmental regulation. Specialised degradation pathways enable constructed wetlands to eliminate recalcitrant organic compounds and emerging contaminants by sustaining a unique population of bacteria that can degrade pharmaceuticals, pesticides and industrial chemicals [42]. Specialised degraders are often minor players in the community but have a disproportionate effect on removing certain pollutants that conventional treatment pathways cannot remove [43]. Carbon cycle illustrating the flow of carbon through the environment is shown in Figure 5.

4. Pollutant Removal Mechanisms and Efficiency

4.1. Comprehensive Pollutant Removal Performance

The integrated physical, chemical and biological operations of wetlands for the treatment of different pollutants are driven mostly by microbial activities [1]. In summary, the different elimination mechanisms that constructed wetlands rely on give these systems an advantage in treating complex wastewaters contaminated by multiple classes of chemical pollutants. The multi-faceted approach to pollutant removal makes constructed wetlands particularly effective for treating complex wastewaters containing multiple contaminant classes [44]. Pollutant removal efficiency ranges achieved by constructed wetland systems for major contaminant categories (Figure 6).

Reductions in BOD and COD for organic matter removal range from 85 to 98% depending on the configuration of the constructed wetlands [45]. Heterotrophic bacteria primarily act as organic matter decomposers, utilising aerobic and anaerobic pathways according to local redox conditions. This high removal efficiency indicates the concerted activity of multiple bacterial populations, acting synergistically to degrade organic matter with varying degrees of complexity and biodegradability [46]. Due to the integration of physical filtration and biological processes mediated by biofilm communities, suspended solids are removed at 80–95% [22]. Physical filtration occurs through the substrate matrix, and biofilm communities stabilise and aggregate particles. The formation of biofilm on the substrate surface acts as a biological filter, increasing particle capture and providing spaces for the attachment of specialised microbial populations [47].

Ammonia nitrogen has a removal efficiency of 70–99%, while total nitrogen removal ranges from 60 to 85% depending on system configuration and operational parameters [20]. Such variation illustrates the complexity of nitrogen cycling processes and the specific environmental requirements for nitrification and denitrification [48]. Multi-zone or hybrid designs usually provide better nitrogen removal [49].

Phosphorus does not have a gaseous removal mechanism; removal is even more difficult, with total phosphorus removal ranging from 40–80% depending on the substrate used and operational strategies [1]. Biological phosphorus removal is contributed to by phosphate-accumulating bacteria in addition to chemical precipitation and adsorption mechanisms [50]. Phosphorus removal over the long term depends on careful selection of substrates and on routine maintenance of installed systems to prevent saturation [51].

4.2. Heavy Metals, Sulfate and Iron Reduction

4.2.1. Heavy Metal Detoxification

Microbial communities in constructed wetlands play a significant role in detoxifying heavy metals through biosorption, bioprecipitation and enzymatic redox transformations. Biosorption occurs when metal ions bind to negatively charged sites on microbial cell walls, extracellular polymeric substances (EPS) and biofilms, providing an effective passive removal mechanism [52]. Microbial activity also promotes the precipitation of metals as insoluble sulfides, phosphates, or carbonates, thereby reducing their solubility and ecological risk [53]. In addition, many bacteria catalyse redox transformations that convert toxic metals to less mobile forms, such as the reduction of Cr(VI) to Cr(III) or the interconversion of As(V) and As(III) [54]. Genera such as Pseudomonas, Bacillus, Clostridium and cyanobacteria are frequently reported in these processes. Collectively, these mechanisms immobilise metals within wetland sediments, protecting aquatic ecosystems from downstream contamination [55].

4.2.2. Sulfate and Iron Reduction

In addition to direct detoxification, sulfate reduction and iron reduction are key anaerobic microbial processes in wetland sediments that indirectly influence heavy metal dynamics. Sulfate-reducing bacteria (SRB), particularly Desulfovibrio and Desulfobacter, utilise sulfate as a terminal electron acceptor, producing sulfide (H₂S) under anoxic conditions [56]. The generated sulfide readily reacts with dissolved metal ions, such as Fe²⁺, Pb²⁺ and Cd²⁺, to form highly insoluble metal sulfides, which are buried in sediments and effectively immobilised. Similarly, iron-reducing bacteria, notably Geobacter and Shewanella, couple the oxidation of organic matter to the reduction of Fe(III) to Fe(II), thereby contributing to redox buffering and influencing nutrient and metal availability [57]. Iron reduction can release bound phosphorus under anoxic conditions, but it can also facilitate secondary mineral formation that immobilises contaminants. Together, sulfate reduction and iron reduction stabilise heavy metals, regulate nutrient cycling and highlight the critical role of anaerobic microbial guilds in maintaining redox balance and treatment efficiency in constructed wetlands.

4.3. Emerging Contaminants and Micropollutants

The treatment of emerging contaminants, including pharmaceuticals, personal care products and industrial chemicals, represents a growing application area for constructed wetlands, with specialised microbial communities providing effective degradation pathways for many persistent compounds [13]. The diversity of metabolic capabilities present in constructed wetland microbial communities enables the degradation of compounds that resist conventional treatment processes [58]. Antibiotic removal occurs through biodegradation, plant uptake, adsorption and photodegradation, with microbial biodegradation serving as the primary mechanism for most antibiotics [59]. Specialised bacteria capable of utilising antibiotics as carbon or nitrogen sources have been identified in constructed wetland systems, with removal efficiencies typically ranging from 50 to 90% depending on antibiotic class and environmental conditions. Hormone and endocrine disruptor removal involves microbial communities capable of degrading steroid compounds through ring cleavage and metabolic transformation [60]. Proteobacteria and Actinobacteria possess enzymes for steroid metabolism, with hormone removal efficiency typically ranging from 60 to 95% depending on system design and operational conditions.

Industrial chemicals and persistent organic pollutants require specialised anaerobic bacteria for dehalogenation reactions, with Dehalococcoides and related bacteria performing reductive dechlorination under strictly anaerobic conditions [42]. Establishing appropriate redox conditions is crucial for supporting these specialised bacteria and achieving effective treatment of chlorinated compounds [61].

4.4. Microplastics and Microbial Interactions in Constructed Wetlands

Microplastics (MPs; particles <5 mm) are increasingly recognised components of wastewater and runoff that accumulate in wetland systems and interact strongly with microbial communities [62]. Constructed wetlands can trap MPs via vegetation, sedimentation and filter media; however, retained particles become colonisation substrates for dense biofilms (plastisphere), which alter particle buoyancy, fate and associated pollutant transport. Recent reviews and field studies showed that CWs can reduce MP loads by physical retention and sedimentation, but removal efficiency depends strongly on particle size, density, vegetation type and hydraulic conditions [63,64].

Microbial biofilm formation is the first and critical step in the fate of MPs. Bacterial and fungal cells attach to plastic surfaces quickly after exposure, forming structured biofilms whose composition differs from surrounding water and sediments [65]. Biofilm colonisation can alter the surface properties of MPs (hydrophobicity and density) and promote aggregation or sinking, thereby enhancing retention in CW sediments [66]. The plastisphere also concentrates other pollutants and can act as hotspots for horizontal gene transfer, including antibiotic resistance genes (ARGs) [67,68].

Beyond colonisation, specific microbes contribute to partial biodegradation of plastics through extracellular enzymes. Genera such as Pseudomonas, Bacillus, Rhodococcus, and Actinobacteria, along with fungi like Aspergillus and white-rot species, have been reported to secrete enzymes, including PETase, MHETase, laccases, and cutinases, that can oxidise or depolymerise polymers, especially polyesters and PET. In constructed wetlands, this translates into a dual role, i.e., CWs primarily act as sinks that retain MPs. At the same time, trapped particles also become hotspots for limited microbial transformation or fragmentation into smaller particles, such as nanoplastics. Although laboratory and mesocosm studies confirm enzymatic activity and surface oxidation, substantial mass loss of common plastics (e.g., PE, PP) in field-scale CWs remains scarce. Practical implications include designing CWs with engineered substrates or biochar to improve retention and promote diverse biofilms while ensuring monitoring of MP-associated antibiotic resistance genes and pathogens, since MPs can serve as vectors of microbial risks if not appropriately managed [67,69,70].

4.5. Regulation of Methane and Greenhouse Gases in Constructed Wetlands

Constructed wetlands (CWs) have attracted attention for their potential as significant sinks for pollutants and as biogeochemical hotspots of greenhouse gas (GHG) production, such as methane (CH4), nitrous oxide (N2O), and carbon dioxide (CO₂). Methanogenesis occurs in the anaerobic zones of CW sediments, where acetate, CO2, and H2 are produced by fermentation and acetogenesis of available organic matter, which can then serve as substrates for methanogenic archaea inhabiting these zones [71]. These processes are accelerated at high temperatures, high organic loading and long retention periods in anaerobic conditions, all of which can increase CH₄ fluxes [13].

The interdependence between methanogens and methanotrophic bacteria in CWs primarily determines methane dynamics. For instance, methanotrophs inhabiting oxygenated microsites, especially at the rhizoplane of emergent macrophytes, oxidise much of the methane produced prior to atmospheric emission [71]. Accordingly, from a mechanistic perspective, root-mediated radial oxygen loss to the soil solution thus represents a key mitigation pathway. Likewise, wetland designs that favour aerobic–anaerobic alternation, such as tidal-flow wetlands or hybrid vertical–horizontal systems, enhance net methane emission suppression by stimulating methane oxidation and displacing electron-accepting pathways, such as denitrification and sulfate reduction [72].

In addition to CH₄, CWs produce N₂O and CO₂. Nitrous oxide is produced during ineffective denitrification in oxygen-restricted situations or as a result of nitrifier-denitrification when ammonium is oxidised in a heterotrophic process using a reduced amount of oxygen [20]. Carbon dioxide (CO₂) is produced naturally from heterotrophic bacteria oxidising organic matter during aerobic respiration and can be the main GHG when well-aerated conditions are maintained [73]. Some of this balance will vary based on hydrology, plants, substrate and even seasonal activity of microbes [74].

The reduction in GHG emissions in CWs needs an ecological and engineering integration. These include (i) the selection of macrophytes with enhanced radial oxygen release to promote methane oxidation, (ii) amendment of substrates with biochar or iron-rich media to suppress methanogenesis or stimulate alternative electron acceptors, (iii) operational approaches, such as intermittent aeration or tidal-flow cycles to prevent strictly anaerobic conditions, and (iv) monitoring frameworks to pair wastewater treatment performance with GHG flux assessments [75,76].

In summary, CWs serve as key components in wastewater purification and carbon sequestration, but they can also be a source of GHGs. Policymakers should seek to balance increased removal of specific pollutants against climate regulation benefits through low-emission designs and microbial management strategies in future research. Removal efficiency of emerging contaminants are presented in

Table 4. Removal efficiency of emerging contminants.

Contaminant Type	Removal Mechanism	Efficiency Range (%)	Responsible Microbial Groups	Source
Antibiotics	Biodegradation, Adsorption	50–90	Actinobacteria, Proteobacteria	[59]
Hormones/EDCs	Biotransformation	60–95	Proteobacteria, Mycobacterium	[60]
Industrial Solvents	Reductive dechlorination	70–95	Dehalococcoides, Clostridium	[42]

.

5. Environmental Factors Influencing Microbial Communities

5.1. Temperature Effects and Seasonal Dynamics

Ambient air temperature represents one of the most significant environmental factors influencing microbial community structure, activity and functional performance in constructed wetlands, creating dynamic conditions that shape microbial succession patterns and affect treatment efficiency throughout the year [74]. Seasonal temperature variations have profound effects on community composition and metabolic rates of key microbial processes [77]. Seasonal community shifts have been documented across various constructed wetland systems, with distinct winter and summer microbial assemblages reflecting adaptation to different thermal regimes [13]. Winter conditions typically feature reduced microbial diversity and altered community composition, with cold-adapted bacteria becoming more prominent, while mesophilic populations decline. Proteobacteria abundance often decreases significantly during winter months, while specific cold-tolerant groups may increase in relative abundance [78].

Temperature-dependent metabolic processes exhibit varying sensitivities to seasonal changes, with nitrification and denitrification rates typically decreasing substantially at low temperatures [71]. This leads to reduced nitrogen removal efficiency during the cold season when biological activity is suppressed. Methanogenic activity shows similar temperature dependence, with peak methane production occurring during warmer periods when anaerobic degradation processes are most active [71]. Cold-climate adaptations have been observed in constructed wetland microbial communities from northern latitudes, where specialised bacteria maintain activity at low temperatures through cold-shock proteins (CSPs) and modified enzyme systems. Some studies have identified psychrotolerant bacteria that become dominant during winter months, enabling continued treatment performance under challenging thermal conditions [79].

5.2. pH and Chemical Environment

The pH environment significantly influences microbial community composition, enzymatic activity and overall treatment performance in constructed wetlands, with most microbial processes operating optimally within circumneutral pH ranges [32]. Deviation from optimal pH conditions can severely impact specific microbial populations and alter the balance of biogeochemical processes [80]. pH exerts a strong selective filter on microbial community composition in wetland and soil systems; acidification or alkalinisation shifts relative abundances of major bacterial groups (e.g., Acidobacteria, Proteobacteria) and often reduces overall bacterial diversity away from neutral pH [81,82]. These shifts can significantly alter functional capacity and treatment efficiency, particularly for pH-sensitive processes like nitrification.

Nitrification is strongly pH-dependent; ammonia-oxidising microorganisms show reduced activity under acidic conditions, and typical optimum activity is often observed near neutral to slightly alkaline pH (commonly reported ~7.0–8.5), such that acidic wastewaters can suppress nitrifiers and cause transient ammonia accumulation [83,84]. Conversely, highly alkaline conditions can promote ammonia volatilisation, reducing substrate availability for biological nitrification and altering nitrogen cycling dynamics. Buffer capacity and pH stability influence the establishment and maintenance of stable microbial communities, with systems possessing adequate buffering capacity resisting pH fluctuations and supporting more diverse and stable microbial populations [76]. Substrate selection can significantly influence pH buffering, with limestone and other alkaline materials helping maintain circumneutral conditions conducive to optimal microbial activity [85].

5.3. Hydraulic Conditions and Retention Time

Hydraulic conditions, notably flow velocity, hydraulic retention time (HRT) and flow pattern, strongly shape microbial community assembly, biofilm stability and pollutant removal by controlling contact time and mass transfer; multiple experimental and review studies document HRT as a key design parameter for optimising nitrification, denitrification and organic matter removal [86]. These hydraulic factors represent critical design parameters that can be optimised to enhance microbial treatment performance. Hydraulic retention time is a critical parameter influencing microbial community establishment and treatment performance, with longer retention times generally promoting the development of more diverse and stable microbial communities by providing sufficient time for slow-growing bacteria to establish [86]. However, an excessively long retention time can lead to stagnation and reduced treatment efficiency due to oxygen depletion and the accumulation of inhibitory compounds.

Different microbial processes require varying retention times for optimal performance, with nitrification typically requiring a longer retention time than organic matter removal due to slower growth rates of nitrifying bacteria [38]. Denitrification efficiency often improves with increased retention time, remarkably when carbon is limited, as additional time allows the utilisation of slowly biodegradable organic compounds. Flow patterns and distributions affect the uniformity of microbial community development and the occurrence of preferential flow paths, which can reduce treatment efficiency [42]. A uniform flow distribution promotes consistent microbial establishment across the treatment area, while short-circuiting creates zones of reduced microbial activity and compromises overall treatment performance. Intermittent operation and flow cycling can enhance microbial diversity and treatment performance by creating alternating environmental conditions that support aerobic and anaerobic processes [1]. Tidal flow wetlands that cycle between saturated and unsaturated conditions enable optimisation of different treatment processes within the same system, improving overall efficiency while reducing greenhouse gas emissions [72]. Key environmental factors affecting microbial ecology are presented in

Table 5. Key environmental factors affecting microbial ecology.

Factor	Optimal Range	Impact on Microbial Activity	Source
Temperature	20–35 °C	Affects metabolic rates, species dominance	[74]
pH	6.5–8.5	Influences enzyme activity, nitrifier stability	[32]
Hydraulic retention time	3–10 days	Determines treatment contact and biofilm growth	[86]
DO concentration	1–5 mg/L	Aerobic vs. anaerobic zone distribution	[87]

.

6. Molecular Techniques and Community Analysis

6.1. Advanced Molecular Approaches

The revolution in molecular biology techniques has fundamentally transformed our understanding of microbial communities in constructed wetlands, enabling the comprehensive analysis of community structure, function and dynamics without the limitations of traditional culture-based approaches [13]. These advanced techniques have revealed the extraordinary complexity and functional diversity of wetland microbiomes while providing tools for system optimisation and performance monitoring. 16S rRNA gene sequencing has emerged as the gold standard for bacterial and archaeal community analysis in constructed wetlands, exploiting the universal distribution and sequence conservation of this gene for accurate taxonomic identification [88]. High-throughput sequencing technologies enable the analysis of thousands of sequences per sample, providing unprecedented resolution of community structure and revealing previously unrecognised microbial diversity in wetland systems.

Pyrosequencing and Illumina platforms have generated extensive datasets demonstrating the consistent dominance of Proteobacteria, Bacteroidetes, Actinobacteria and Firmicutes across different constructed wetland configurations while identifying system-specific variations in community composition [53]. These studies have also revealed the presence of rare but functionally important microbial populations that contribute disproportionately to specific treatment processes. Functional gene analysis targets genes encoding key enzymes in biogeochemical cycles to assess the functional potential of microbial communities, with nitrogen cycling genes, including amoA, nirS/nirK and nosZ, being extensively studied to understand nitrogen transformation processes [10]. Quantitative PCR (qPCR) techniques enable the quantification of specific genes and provide estimates of functional population sizes that can be correlated with process rates and treatment performance.

6.2. Omics Technologies and Systems Biology

Advanced omics technologies have opened new frontiers in constructed wetland microbial ecology by providing comprehensive insights into community function, metabolism and interactions that move beyond taxonomic identification to reveal actual functional activities [71]. These approaches enable systems-level understanding of microbial community function and provide mechanistic insights into treatment processes. Metagenomics involves sequencing total community DNA to characterise both taxonomic composition and functional gene content, revealing extensive functional diversity in constructed wetland microbial communities and identifying novel genes and metabolic pathways [89,90]. Metagenomic studies have been particularly valuable for understanding antibiotic resistance gene dynamics and identifying specialised degradation pathways for emerging contaminants.

Metatranscriptomics analyses community RNA to identify actively expressed genes and metabolic pathways, providing insights into functional activities under specific environmental conditions [91]. This approach has revealed dynamic expression of nitrogen cycling genes in response to environmental variations and identified key metabolic pathways active during different operational phases. Multi-omics integration enables systems-level understanding of microbial community function through combined analysis of genomic, transcriptomic, proteomic and metabolomic data [92]. These integrated approaches provide mechanistic insights into treatment processes and reveal complex metabolic networks that drive pollutant removal in constructed wetlands [93].

7. Climate Change Impacts and Future Challenges

7.1. Temperature and Precipitation Changes

Climate change is fundamentally altering environmental conditions globally, with significant implications for microbial communities and treatment performance in constructed wetlands that require adaptive management strategies to maintain efficiency under changing climatic conditions [12,73,90,94]. Rising temperatures, extreme weather events and altered precipitation patterns present both challenges and opportunities for constructed wetland operation. Warming impacts on microbial communities include shifts in community composition, altered metabolic rates and changes in biogeochemical cycling patterns [38]. Higher temperatures generally accelerate microbial metabolic processes, potentially improving treatment rates for some pollutants while creating challenges for temperature-sensitive processes. However, extreme temperatures can disrupt established microbial communities and reduce treatment performance through thermal stress and altered competitive relationships.

Climate change-driven alterations in precipitation patterns influence the stability and resilience of constructed wetlands, requiring ecosystem-based adaptation approaches to maintain treatment performance under extreme conditions [95]. Precipitation changes and hydrological impacts can alter hydraulic loading, dilution rates and microbial community stability, with extreme rainfall events potentially overwhelming treatment capacity and washing out slow-growing microbial populations [94,96]. Design strategies, including overflow structures, storage capacity and modular configuration, can improve flood resilience while maintaining treatment performance during extreme weather events. Drought adaptation challenges require systems capable of maintaining treatment performance under reduced water availability with concentrated pollutants and extreme environmental conditions that stress microbial communities [38]. Water recycling, alternative source and drought-tolerant plant species can improve system resilience, while specialised microbial management strategies help maintain community stability.

7.2. Emerging Contaminants and Resistance Genes

The proliferation of antibiotic resistance genes and emerging contaminants in aquatic environments represents a significant challenge for constructed wetland systems, requiring enhanced understanding of fate and transport mechanisms to optimise treatment efficiency while minimising environmental risks [97]. These challenges demand innovative approaches that integrate traditional engineering with advanced molecular monitoring and management strategies. Antibiotic resistance gene dynamics in constructed wetlands involve complex interactions between selection pressure, horizontal gene transfer and environmental conditions that influence ARG abundance and dissemination [98]. Studies have demonstrated that constructed wetlands can serve as sinks and sources for antibiotic resistance, with removal efficiency typically ranging from 0.5 to 3.0 log units depending on system design and operational conditions. Emerging contaminant treatment capabilities of constructed wetlands continue to expand as specialised microbial populations are identified and cultivation strategies are developed [42]. Pharmaceutical compounds, personal care products and industrial chemicals require specific degradation pathways that can be present in diverse microbial communities but require optimisation for consistent treatment performance.

8. Technological Innovations and Future Directions

8.1. Advanced System Configurations

Innovation in constructed wetland design continues to evolve through integration of advanced technologies and novel system configurations that enhance microbial treatment capabilities while addressing emerging challenges [99]. These innovations leverage improved understanding of microbial processes to create more efficient and resilient treatment systems. Hybrid system integration combines different wetland configurations to optimise treatment for specific pollutants through complementary microbial processes [1]. Vertical flow wetlands for nitrification, followed by horizontal flow wetlands for denitrification achieve enhanced nitrogen removal, while additional specialised stages can target specific contaminants as needed.

Biochar and engineered substrates offer enhanced surface area and specific chemical properties that improve microbial colonisation and treatment performance [75]. Research has demonstrated that biochar amendments can increase microbial diversity and improve treatment efficiency through enhanced biofilm formation and electron transfer processes. Artificial intelligence and process optimisation applications enable real-time monitoring, predictive modelling and adaptive control of treatment processes through integration of sensors, data analytics and machine learning algorithms [32]. These technologies can identify patterns in microbial community dynamics and optimise operational parameters for enhanced treatment performance.

8.2. Circular Economy Integration

Future constructed wetland development increasingly focuses on integration with circular economy principles, including resource recovery, energy generation and waste minimisation, that transform treatment systems into integrated resource recovery facilities [38]. These approaches maximise value extraction while minimising environmental impacts through comprehensive resource utilisation strategies. Nutrient recovery technologies can capture valuable nutrients from wastewater for agricultural applications through processes that concentrate and recover nitrogen and phosphorus in forms suitable for fertiliser use [32]. Integration with constructed wetlands creates opportunities for sustainable nutrient management while reducing the environmental impacts of conventional fertiliser production. Energy generation and storage through bioelectrochemical systems and biogas production can offset operational energy requirements while contributing to renewable energy supplies [76]. Microbial fuel cell integration with constructed wetlands has demonstrated power generation capabilities while maintaining treatment performance.

8.3. Limitations of Constructed Wetlands

Although they offer many environmental benefits and low operating costs, CWs have some limitations that should be considered for a fair assessment. The need for a large space compared to traditional treatment technologies is one of the main limitations of the A2O process, making its implementation impractical in urban and densely populated areas [1]. CWs also have a long time for start-up because the microbial communities and the plant systems also need months to establish steady performance for treatment [11]. Phosphorus is not removed from wastewater by a gaseous pathway and is mostly retained through hydrolysis, adsorption or precipitation, leading to relatively lower and variable phosphorus removal efficiencies [31]. Likewise, emerging contaminant and micropollutant removal, even if promising, is inconsistent and relies heavily on the presence of specific microbial assemblages, type of substrate and hydraulic conditions [42].

A related major drawback is the risk of GHG emissions, primarily methane and nitrous oxide, and, if not carefully managed, these emissions can mitigate the climate benefits of CWs [71,72]. In turn, seasonal changes in temperature, hydrology, and plant growth affect microbial activity, often decreasing efficiency during winter or extreme weather [71,76]. In addition, CWs can act as sinks for antibiotic resistance genes (ARGs) and pathogenic microorganisms, which can have serious ecological and public health consequences due to the reuse of effluents in agriculture or their disposal into sensitive ecosystems [75,76,89,98]. However, technological innovations regarding biochar amendments, hybrid systems, and artificial intelligence have been identified as important solutions, but their applications have been restricted in large-scale activities due to costs, scalability issues, and regulatory impediments. Thus, although CWs are a sustainable, nature-based solution to wastewater treatment, the widespread implementation must take these limitations into account through future research exploring scalable innovations and adaptive management strategies.

9. Conclusions

Constructed wetlands are a sustainable and inexpensive approach to treat wastewater based on natural processes driven by naturally complex microbial–vegetation ecosystems. Microbial communities in these systems contribute to pollutant removal and drive critical biogeochemical processes for nitrogen cycling, phosphorus removal, organic matter degradation and the treatment of emerging contaminants. CWs contain remarkable microbial diversity dominated by a number of bacterial phyla, including Proteobacteria, Bacteroidetes, Actinobacteria and Firmicutes, playing diverse functional roles that contribute to the treatment process. The microbial communities are also affected by environmental factors, such as temperature, pH, hydraulic retention time and plant species composition, and play an essential role in the efficiency of the wastewater treatment (WWT) process.

The rapid development of molecular biology tools (especially high-throughput sequencing and metagenomic analysis) has largely advanced our comprehension of microbial functions and ecology in CWs. The future will be bright with technology, as optimisation of CW system design in terms of engineering configurations (hybrid CW), media (biochar and engineered substrates) and microbiology (inoculum enrichment) can significantly improve removal efficiencies and restore ecosystem functions while providing more resilience against climate change. Improving operational parameters through real-time monitoring using molecular tools, such as metagenomics and metatranscriptomics, is also essential, and this will help to improve predictions for treatment performance, especially for new contaminants or antibiotic resistance. With climate change and all its problems, future CWs should aim for more resilience against extreme weather events such as floods and dry spells.