Impact of Anthropogenic Activities on Microbially Mediated Carbon Dioxide and Methane Emissions in Wetlands: A Review and Prospects

Abstract

Global wetlands play a significant role as “blue carbon sinks”. Despite their relatively small coverage, they have enormous potential for carbon capture and sequestration, and also serve as an important natural source of atmospheric carbon dioxide (CO₂) and methane (CH₄). Wetland ecosystems are characterized by complex microbial interactions that mediate carbon (C) cycling processes, and also directly influence the dynamic changes of CO₂ and CH₄, underscoring the crucial role of microorganisms in these systems. Understanding the ecological significance of these gases and their response mechanisms to environmental changes is vital for mitigating the greenhouse effect and conserving ecosystems. This paper reviewed the major environmental challenges facing wetlands globally, such as salinization, over-fertilization, heavy metal input, and microplastic pollution, all influenced by human activities. Additionally, it examined their impact on microbial interactions that mediate the carbon cycle and related greenhouse gas emissions. This review highlighted the crucial role of microorganisms in these cycles and provided a microbial ecological perspective and theoretical foundation for promoting sustainable development and reducing greenhouse gas emissions in wetland areas.

1. Introduction

Wetlands are areas inundated with water either permanently or intermittently, encompassing swamps, peatlands, marshes, and coastal intertidal zones. They are extensively distributed across the globe, covering approximately 5–10% of the Earth’s land surface [1]. Acting as ecotones between terrestrial and aquatic environments, wetlands provide critical ecosystem services such as hydrological regulation, water purification, flood mitigation, and biodiversity maintenance [2], and possess irreplaceable ecological and economic values [3]. In particular, wetlands play a critical dual role in the global carbon cycle, functioning as both natural sources and sinks for the major greenhouse gases carbon dioxide (CO₂) and methane (CH₄) [4]. Additionally, wetlands are core sites for the long-term accumulation and sequestration of soil organic carbon (SOC) [5]. However, this crucial ecological function is facing severe threats due to anthropogenic activities.

In recent years, with increasingly intensive anthropogenic activities, wetlands have started to degrade at an accelerated pace. It is estimated that approximately 54–57% of global wetlands have disappeared since 1900 [6]. Although wetland degradation has persisted throughout human history, with anthropogenic activities continually triggering crises such as salinization and soil degradation, historical measures—including the cultivation of salt-tolerant crops and improved management practices—have partially alleviated these challenges in some regions [7]. Nevertheless, the impacts of contemporary anthropogenic activities on wetlands remain widespread and severe [8]. Under the current context of over-exploitation, anthropogenic hydrological alterations leading to wetland salinization, nutrient enrichment from over-fertilization and heavy metal pollution due to the conversion of wetlands into agricultural and industrial lands, and the massive input of microplastics (MPs) from the use of novel materials have profoundly impacted the ecological health and biogeochemical cycles of wetland areas. These environmental issues significantly influence carbon cycling processes related to CO₂ and CH₄ emissions.

The conservation and restoration of wetland ecosystems have become a critical issue in global ecological governance. Clarifying the response patterns of wetland carbon cycling to anthropogenic disturbances is central to achieving wetland sustainability and greenhouse gas mitigation. However, existing research has largely focused on the impact of single environmental stressors on wetland greenhouse gas emissions, lacking a systematic synthesis of multiple concurrent pressures. This paper explores urgent environmental threats to global wetlands from a microbial ecological perspective. By conducting a systematic review of the literature on anthropogenic impacts on CO₂ and CH₄ emissions, it evaluates the severity of these effects, especially regarding wetland salinization, nutrient enrichment from over-fertilization, heavy metal pollution and MP input. The clustering function of CiteSpace (version 6.1.R2) was also used to reflect the distribution of research hotspots in each field to better inform conservation policies and integrated management strategies. The findings ultimately aim to establish a framework for sustainable wetland development and greenhouse gas mitigation moving forward.

2. Methods

2.1. Literature Search Strategy

This study conducted a systematic literature search to synthesize current knowledge on anthropogenic impacts on carbon cycling and greenhouse gas dynamics in wetland ecosystems. We used two authoritative databases (Web of Science Core Collection and Google Scholar) to search for literature, covering peer-reviewed studies published from 2014 to 2024. Specifically, the search strategy combined “CO₂, wetlands” OR “CH₄, wetlands” with the four anthropogenic stressor terms. Searches were applied to titles, abstracts, and author keywords to ensure broad coverage of relevant studies. In addition to database searches, Google Scholar was used to retrieve complementary literature not fully captured by predefined keyword combinations. Detailed search strings for each database are provided in the Supplementary Materials.

2.2. Study Selection Criteria and Screening

The screening and selection of studies followed a PRISMA-style workflow [9], including duplicate removal, title and abstract screening, full-text eligibility assessment, and final inclusion. Full-text articles were assessed for eligibility based on predefined inclusion and exclusion criteria. Studies were included if they met the following criteria:

(1)

Focused on natural, artificial, or restored wetland ecosystems;

(2)

Addressed CO₂ and/or CH₄ emissions, carbon cycling, or closely related biogeochemical or microbial processes;

(3)

Examined the influence of anthropogenic stressors, including salinization, nutrient enrichment, heavy metal pollution, or microplastic contamination.

Following screening and eligibility assessment, 352 studies related to salinization, 284 related to nutrient enrichment, 132 related to heavy metal pollution, and 87 related to microplastics were retained and included in the final qualitative synthesis and bibliometric analysis. Detailed screening criteria and step-by-step statistics are available in the Supplementary Materials.

2.3. Bibliometric and Hotspot Analysis Using CiteSpace

To identify major research hotspots and development trends within each anthropogenic stressor category, a bibliometric analysis was conducted using CiteSpace (version 6.1.R2), which was developed by Professor Chaomei Chen from Drexel University, Philadelphia, USA [10]. In CiteSpace, the time slicing was set to one year. Keywords were used as the node type to construct co-occurrence networks.

Clustering analysis was based on the log-likelihood ratio (LLR) algorithm. Values greater than 0.7 indicate robust and well-defined clusters [10].

In this study, the average silhouette values for all four stressor categories ranged from 0.7422 to 0.8014, confirming reasonable cluster divisions. The resulting keyword clusters (as shown in Tables S1–S4) provide a scientific basis for the narrative structure of each corresponding section in this review. Detailed CiteSpace parameter settings and hotspot interpretation criteria are presented in the Supplementary Materials.

3. Carbon Dioxide and Methane Emissions in Wetlands

After the last glacial maximum, wetlands began to accumulate organic carbon continuously and formed a dense carbon pool [11]. However, this dense carbon pool has been significantly disturbed by anthropogenic activities since the pre-industrial period, leading to significant changes in the carbon cycle in wetlands, including in the emissions of CO₂ and CH₄ [12]. It was estimated that the degradation of wetlands could cause greenhouse gas emissions equivalent to about 408 gigatons of CO₂ between 2021 and 2100 [12]. Specially, wetlands are the largest natural source of CH₄ emissions, accounting for about 62% of the natural CH₄ emissions, and representing the most important source of uncertainty in the CH₄ budget [13].

In wetland soils, CO₂ emissions are primarily mediated through autotrophic root respiration and heterotrophic microbial decomposition processes. Conversely, wetlands serve as natural CO₂ sinks by sequestering atmospheric carbon in soils and vegetation biomass over long timescales [14]. CH₄ production and release in wetlands are driven by methanogenic archaea under anoxic conditions (

Table 1. Classification of different trophic methanogenic archaea in wetland soils at the level of order, family, and genus.

Type	Order	Family	Genus	Reference
Acetotrophic methanogens	[not assigned]	Methanosaetaceae	Methanosaeta	[17,18,19,20]
	Methanosarcinales	Methanosarcinaceae	Methanosarcina
Hydrogenotrophic methanogens	Methanobacteriales	Methanobacteriaceae	Methanobacterium Methanobrevibacter Methanosphaera Methanothermobacter	[19,20]
		Methanothermaceae	Methanothermus
	Methanococcales	Methanococcaceae	Methanococcus Methanothermococcus
		Methanocaldococcaceae	Methanocaldococcus Methanotorris
	Methanomicrobiales	Methanomicrobiaceae	Methanomicrobium Methanoculleus Methanofollis Methanogenium Methanolacinia Methanoplanus
		Methanoregulaceae	Methanosphaerula Methanoregula
		Methanospirillaceae	Methanospirillum
		Methanocorpusculaceae	Methanocorpusculum
	Methanosarcinales	Methanosarcinaceae	Methanimicrococcus
	Methanopyrales	Methanopyraceae	Methanopyrus
	Methanocellales	Methanocellaceae	Methanocella
Methanotrophic methanogenic	Methanosarcinales	Methanosarcinaceae	Methanococcoides Methanohalobium Methanohalophilus Methanolobus Methanomethylovorans Methanimicrococcus Methanosalsum Halomethanococcu Methanosarcina	[20,21,22]
Oxymethanotrophic methanogenic	Methanosarcinales	Methermicoccaceae	Methermicoccus	[23]
Long-chain alkyl hydrocarbon trophic methanogenic	“Candidatus Methanoliparales”	“Candidatus Methanoliparaceae”	“Candidatus Methanoliparum”	[23]

), via acetate fermentation, CO₂/hydrogen (H₂) reduction, and methylotrophic pathways [15]. The quantity and pathway of CH₄ production depends on substrate type and availability, pH, temperature, and methanogenic community structure [16].

Before being emitted to the atmosphere, a large proportion of produced CH₄ undergoes anaerobic oxidation in anoxic wetland soils by consortia of archaea and bacteria utilizing alternate electron acceptors like sulfate (SO₄²⁻), nitrate (NO₃⁻)/nitrite (NO₂⁻), manganese (Mn), or iron (Fe) [24]. This anaerobic methane oxidation (AOM) modulates a major CH₄ sink, estimated to consume over 30% and 90% of freshwater and marine wetland CH₄ emissions, respectively [25]. Biogeochemical cycling of CO₂ and CH₄ in wetlands is intricately coupled and strongly mediated by microorganisms (Figure 1) [26]. Therefore, sustainable wetland management incorporating microbiology is key to regulating greenhouse gas emissions and sinks.

4. The Impact of Wetland Salinization on CO₂ and CH₄ Emissions

Human activities have disrupted hydrological patterns and altered soil salt distribution [27]. Over-irrigation emerges as a significant driver of wetland salinization. For instance, extensive groundwater pumping for irrigation in the Murray River Basin, Australia, has significantly lowered groundwater levels since 2000. A total of 76 reaches (21%) had groundwater levels 10 m below the surface level from 2000 to 2019, disrupting groundwater–wetland connectivity and reducing water flow to downstream reaches, further increasing the risk of salinization of wetland soils [28]. In addition, water infrastructure has also dramatically altered wetland hydrology. For example, the construction of the Three Gorges Dam in China reduced the connection between the Yangtze River and its floodplain wetlands, decreasing the frequency of freshwater inputs, while increasing the chances of seawater intrusion, leading to an increase in soil salinity in downstream wetlands [29,30]. Notably, numerous naturally occurring saline–alkaline habitats exist in nature (e.g., coastal salt marshes, salt lakes, and mangroves), where the vegetation and microbial communities have evolved strong salt tolerance, enabling them to adapt to high-salinity environments and maintain the stability of their own ecosystems [31]. However, it is clear that human activities have extensively altered global wetland hydrology through a variety of mechanisms, inducing widespread secondary salinization. This anthropogenically driven salinization process often exceeds the adaptive capacity of the original wetland biota [32], thereby jeopardizing agricultural productivity, ecological stability, and water resource sustainability. These problems are likely to intensify as climate change continues, necessitating integrated measures to mitigate and adapt to changes in wetland salinity and ensure the resilience and vitality of these important ecosystems.

Based on the CiteSpace clustering analysis of 352 publications, research on wetland salinization has shifted from isolated salinity monitoring toward a more integrated understanding of system-level feedbacks in carbon cycling. One dominant hotspot concerns blue carbon dynamics and vegetation regulation. Studies are highly concentrated on the roles of Spartina alterniflora and Phragmites australis in regulating greenhouse gas fluxes and blue carbon storage. Changes in plant composition and distribution are increasingly recognized as key mediators of wetland carbon sequestration and emissions under salinization stress [33,34,35]. An emerging and rapidly growing hotspot is peat collapse induced by sea-level rise. This process has been identified as a critical risk factor that accelerates the release of previously stable carbon pools, highlighting the vulnerability of coastal and peatland wetlands to salinity-driven geomorphic and biogeochemical disturbances [36,37]. At the mechanistic level, soil respiration and litter decomposition responses to climate change, together with extracellular enzyme activity as an indicator of saltwater intrusion effects on organic matter mineralization, represent core research frontiers [38,39]. These micro-scale processes provide essential insights into how salinization alters carbon turnover and greenhouse gas emissions in wetlands.

4.1. Impact of Wetland Salinization on CO₂ Emissions

Salinization severely disrupts the carbon cycle in coastal wetland ecosystems through interconnected pathways. Among them, microbe-mediated mineralization of organic matter is the main source of CO₂ emissions from wetland soils [40]. In wetlands, mild salinization with a sodium chloride (NaCl) concentration of 3‰ inhibits microbial activity and thus reduces CO₂ production. A salinity level of 3‰ NaCl was shown to reduce cumulative CO₂ emissions by 15%, from 856 mg C/kg in an unsalted control to 723 mg C/kg [41]. This inhibitory effect increases with increasing salinity. Severe salinization with a NaCl concentration of 10‰ significantly inhibits microbial respiration, reducing cumulative CO₂ emissions by 35% from 856 mg C kg⁻¹ to 558 mg C kg⁻¹ [41].

Concurrently, salinization diminishes primary production and carbon fixation by imposing osmotic stress on plant tissues. High salinity inhibits chlorophyll synthesis, with subsequent inhibition of photosynthesis, and the degree of inhibition positively correlated with salinity levels [42]. Additionally, high salt stress prompts competition between sodium (Na⁺) and chloride (Cl⁻) ions and essential nutrients such as potassium (K⁺), calcium (Ca²⁺), and magnesium (Mg²⁺), resulting in reduced nutrient uptake and consequently affecting the plant’s nutrient balance [42]. Despite these impacts, certain specialized microbial communities, such as gram-negative bacteria and anaerobic methane-oxidizing bacteria, may adapt to high salt environments and sustain a degree of CO₂ release [42].

In summary, the salinization process disrupts the carbon cycle of wetland ecosystems through multiple pathways. It inhibits microbial activity, organic matter mineralization, and plant growth and photosynthesis, thereby reducing CO₂ emissions and carbon fixation. Although some specialized microbial communities show some adaptability, they have a limited role in regulating the whole carbon cycle. In addition, changes in salinity can affect the extent of the damaging effects. Therefore, controlling salinity and maintaining appropriate salinity levels are essential to ensure the stable functioning of coastal wetland ecosystems and their carbon sink capacity.

4.2. Impact of Wetland Salinization on CH₄ Emissions

Salinity could regulate the process of CH₄ production mainly by affecting substrate supply (e.g., quantity and quality of soil organic matter, plant secretion, and residue substrates) [43]. There were differences in the responses of different types of wetlands to CH₄ fluxes during salinization. This was mainly attributed to differences in vegetation composition, primary productivity, root oxygen release, and availability of electron acceptors in different wetland types [44]. Additionally, the sensitivity of distinct methanogen taxa to salinity changes could elicit contrasting feedbacks. For example, low concentrations of salinity (0–7.5‰) were not thought to inhibit methanogens. In contrast, CH₄ production rates could be significantly inhibited when salinity is 10‰ and above, with a 24% reduction in both potential CH₄ production, and thus a number of studies have pointed to 10‰ as an important threshold [45]. However, some halophilic archaea (belonging to the family Methanosarcinaceae, Methanosarcinales) can thrive and produce CH₄ in salinities up to 120 g L⁻¹ [46]. High salinity also increased soil pH, further inhibiting methanogens [47]. Methanogens are most active at pH 6.5–7.5 [48] and are suppressed under highly alkaline conditions above pH 9 [49]. In summary, salinity could affect CH₄ production through interconnected pathways. Although hyper-salinity suppresses methanogenesis, some microorganisms have adapted to sustain the process.

AOM is a process in which CH₄ is reduced under anaerobic conditions, mediated by microorganisms. The capacity and rates of AOM are affected by salinity through multiple aspects. High salinity exceeding 292 g L⁻¹ severely suppresses AOM, likely by inhibiting the metabolisms of the microorganisms involved [50]. This salinity effect may be attributed to ionic stress on cellular structures, osmotic stress on molecular pathways, and reduced bioenergetic yields at high salt levels [51]. Comparing the upper limit of salinity tolerated by methanogenic and methane-oxidizing bacteria, salinity inhibited methanogenic archaea more significantly than methane-oxidizing bacteria [52]. Thus, higher salinity could reduce CH₄ emissions, and the optimal salinity for sequestering CH₄ was hypothesized to be 24 g L⁻¹ based on Chen et al. [4].

5. The Impact of Wetland Over-Fertilization on CO₂ and CH₄ Emissions

Since 1900, more than 50% of natural wetlands have been lost globally [6]. The main type of conversion was agricultural land [53]. Over-fertilization in agricultural lands severely disrupted the biogeochemistry of wetland soils. For example, Mississippi wetlands have received 60% and 56% of nitrogen (N) and phosphorus (P) inputs, respectively, from agricultural sources [54]. Studies have shown that higher rates of N/P additions lead to soil acidification, decreased pH, and accelerated SOC mineralization, resulting in the loss of SOC [55]. It is clear that over-fertilization destroyed the native habitat and microbial community structure of wetland soils. It also stimulated climate forcing through increased soil respiration and nitrous oxide emissions [16]. These changes disrupted the nutrient mitigation capacity of wetlands and have implications for carbon emissions.

The CiteSpace clustering analysis of 284 publications reveals several key dimensions that currently dominate research on nutrient enrichment and over-fertilization in wetland ecosystems. Stoichiometric regulation emerges as a major research focus, particularly regarding how changes in carbon-to-nitrogen (C:N) ratios under nitrogen deposition influence ecosystem stability [56]. This topic has been intensively studied in representative regions such as the Sanjiang Plain, where long-term fertilization experiments provide insights into nutrient–carbon coupling under external nutrient loading [57,58]. Another prominent hotspot centers on rice-based wetland ecosystems. Large research clusters address the responses of rice growth, microbial biomass, and 16S rRNA gene diversity to long-term fertilization [59,60]. These studies collectively aim to mitigate CH₄ emissions by regulating the microbial community structure while maintaining agricultural productivity in managed wetlands.

5.1. Impact of Wetland Over-Fertilization on CO₂ Emissions

Currently, nitrogen is the most widely used fertilizer worldwide. Several studies have shown that both short-term and long-term N fertilizer applications enhance microbial respiration in wetlands and promote CO₂ emissions. This is because N fertilizer provides more available substrate and stimulates an increase in the soil microbial population [61]. However, excessive N fertilization can have a suppressive effect and reduce CO₂ release [62]. For example, the application of NH₄NO₃ fertilizer gradually increases soil acidity over time, which reduces enzyme activity and CO₂ production [61]. At the same time, a Finnish study found that N fertilizers may increase CO₂ uptake by vegetation and offset the increase in soil emissions [63]. However, the effect depends on site conditions such as soil properties and water table. Notably, in the context of global warming, temperature and nitrogen addition may interact to weaken each other’s effect on CO₂ production [64]. This suggests that it is difficult to predict the combined effects of climate change and eutrophication on the wetland carbon cycle.

Similar to nitrogen fertilizer additions, both short-term and long-term phosphorus fertilizer inputs promote soil CO₂ emissions. Results from phosphorus application trials in Florida wetlands, USA, showed that the respiration rate of microorganisms increased significantly by 857.14% at a phosphorus application rate of 4.6 mg P g⁻¹ under the condition of low water levels and the provision of an available carbon source [65]. Meanwhile, temperate wetlands yielded similar results in relatively low-phosphorus-addition trials (0.2 mg P g⁻¹) [66]. Compared to short-term phosphorus application, long-term high-phosphorus inputs increased soil CO₂ emissions. The study shows that high-phosphorus inputs emit 14–40% more CO₂ than low-phosphorus controls. Long-term high-phosphorus inputs may promote soil CO₂ emissions. On the one hand, a higher microbial biomass P in high-phosphorus soils indicates activated microbial activity. On the other hand, high-phosphorus treatments can increase the aboveground biomass, enhancing autotrophic respiration.

However, excessive or prolonged phosphorus application inhibited microbe-driven CO₂ release. Studies have shown that nutrient overloading due to excess phosphorus is an important inhibitory mechanism [67]. This is due to the large accumulation of phosphorus in the soil; more than 75–90% of the added phosphorus will form insoluble phosphorus salts. These phosphorus salts are immobilized, and it is difficult for them to be utilized by the plants and microorganisms, which inhibits the respiration of the organisms [68]. At the same time, long-term over-application of phosphorus will inhibit the activity of soil enzymes that can dissolve fixed phosphorus, such as acid phosphatase [69]. In fact, with 481.9 kg⋅ha⁻¹ of P application, the acid phosphatase activity was reduced by about 24.81% [69], so that the conversion of fixed phosphorus to effective phosphorus was weakened, resulting in a reduction in effective phosphorus.

In conclusion, moderate nitrogen and phosphorus fertilization can stimulate the amount and activity of wetland microorganisms and promote CO₂ emission; however, excessive or long-term fertilization can lead to soil acidification and inhibit enzyme activity or cause nutrient overload and inhibit microbial metabolism.

5.2. Impact of Wetland Over-Fertilization on CH₄ Emissions

The impact of long-term heavy nitrogen fertilizer application on CH₄ emissions in soil has become a focal point for ecologists, but research findings are inconclusive. A meta-analysis by Aronson and Helliker [70] reported that small amounts (<100 kg N ha ⁻¹) of N tended to increase CH₄ oxidation, whereas large amounts tended to inhibit CH₄ oxidation. Specifically, a small amount of N application reduced CH₄ emissions from the marsh by 87%, especially during the hot summer months [71]. This may be due to the fact that the nitrogen application increased the activity of methanotrophic bacteria using NO₂⁻/NO₃⁻ as electron acceptors, thereby reducing CH₄ emissions [71]. However, excess nitrogen fertilization can lead to soil acidification, and the optimal pH for Methylomirabilis oxyfera, the main bacterium involved in the AOM denitrification-dependent anaerobic methane oxidation (DAMO) process, is 7.5 [72]. This suggests that excess nitrogen fertilization could hinder the activity of functional bacteria, impede the AOM process, and ultimately exacerbate CH₄ emissions.

The long-term application of phosphorus fertilizer (around 5.6 kg P ha⁻¹) could suppress CH₄ emissions [73]. In the Dipper Harbour wetland, CH₄ emissions from the high-phosphorus and low-nitrogen treatment groups were significantly lower than those from the control group, with a reduction of about 38%. This difference may be attributed to the fact that phosphorus fertilization enhanced vegetation growth, increased root density, and inhibited CH₄ production by increasing soil aeration. Moreover, the luxuriant vegetation may have altered the transport pathway of CH₄ from the deep soil to the atmosphere, reducing CH₄ emissions. In contrast, the effect of phosphorus fertilizer on CH₄ appeared to be weaker compared to the stimulatory effect of nitrogen fertilizer [73].

It is worth noting that in actual farmland, the effects of the combined application of chemical and organic fertilizers may be different from the effects of single-fertilizer applications. Yuan et al. [60] found that the application of organic fertilizers increased CH₄ emissions. The application of organic fertilizer favored acetoclastic methanogens as the primary group but suppressed hydrogenotrophic methanogens. This shift in the methanogenic community structure, coupled with increased carbon substrate availability, likely contributed to higher CH₄ production in the organic-treated compared to other treatments. However, the hybrid treatment approach had the potential to balance the forces of methanogenic and methane-oxidizing bacteria to partially reduce CH₄ emissions while maintaining high rice yields. Therefore, it is important to consider the interaction effect of fertilizer type and concentration when evaluating the effect of fertilizer application on CH₄ release.

6. The Impact of Wetland Heavy Metal Pollution on CO₂ and CH₄ Emissions

In recent years, urbanization and industrialization have led to the accumulation of heavy metals such as iron (Fe), manganese (Mn), copper (Cu), cadmium (Cd), and arsenic (As) in wetlands through industrial waste [74], resulting in serious environmental pollution. In addition, livestock manure from wetland farming used as agricultural fertilizer can also lead to heavy metal accumulation in soil [75]. This not only jeopardizes the soil properties, but also threatens the atmosphere, water bodies, and biosphere, and may endanger human health through water, food, and dermal contact [76]. In addition, heavy metal pollution also directly affects the key ecological processes in mudflat wetlands [77]. Studies have shown that different bacteria have different sensitivities to heavy metals, and thus heavy metals affect wetland ecosystems and carbon emission processes by altering microbial functions [78]. However, the diverse effects of heavy metals on the ecosystem and the high complexity of the environment make the mechanism of this process controversial, and it needs to be further explored.

Clustering analysis of 132 publications indicates that research on heavy metal pollution in wetlands is increasingly moving toward an integrated framework combining ecological assessment and remediation. One major hotspot involves synergistic remediation strategies, particularly the interactions between constructed wetlands and arbuscular mycorrhizal fungi [79,80]. These approaches are widely investigated as effective means to reduce heavy metal contamination while maintaining carbon stability and ecosystem functioning. At the mechanistic frontier, the application of metagenomic and metatranscriptomic techniques has become a high-frequency research focus. These molecular tools are used to elucidate microbial tolerance mechanisms during heavy metal transformation and to reveal how microbial functional shifts drive carbon cycling under metal stress conditions [81].

6.1. Impact of Wetland Heavy Metal Pollution on CO₂ Emissions

The expansion of industrial activities often leads to the pollution of various heavy metals, which can significantly inhibit the CO₂ release from wetland soils. Cd, Hg, Cr, Cu, Zn, and Pb are the key heavy metal pollutants that have attracted much attention in wetland ecosystems [82] (Table 2). There is a clear nexus between metal contamination and the inhibition of microbial respiration, with the strength of inhibition being influenced by soil characteristics [83]. It is noteworthy that the inhibitory effect is most pronounced in soils characterized by low pH values and limited organic matter content. Conversely, it is less obvious in soils with higher pH values and greater organic matter content [84]. This can be ascribed to the disparate reactivity of metals in different soil matrices. Soils with higher organic matter content exhibit a greater resistance to metal contamination. The presence of ample organic matter facilitates the adsorption of metal ions, effectively mitigating their deleterious impact on microorganisms [83]. Simultaneously, a low pH value (around 5) promotes the desorption of heavy metals from soil constituents and leads to a significant increase in the dissolution of heavy metals in the soil solution [85], whereas a higher pH has the opposite effect. Consequently, these soils provide a more favorable environment for microbial communities, enabling them to sustain metabolic activity even in the presence of heavy metal pollutants.

It should be noted that the effect of mixed contamination with different heavy metals may be different from that of single-metal contamination. For example, Zn pollution did not inhibit soil respiration when a single heavy metal such as Zn at a moderate level (5018 $\mathrm{m}\mathrm{g}·{\mathrm{k}\mathrm{g}}^{-1}$) was polluted according to Beata et al. [86]. However, heavy metal pollution with Zn at a lower level (234.82 ± 15.61 $\mathrm{m}\mathrm{g}·{\mathrm{k}\mathrm{g}}^{-1}$) and Cr (102.99 ± 1.71 $\mathrm{m}\mathrm{g}·{\mathrm{k}\mathrm{g}}^{-1}$) enriched in Futian National Mangrove Nature Reserve suppressed microbe-mediated CO₂ emissions, which were estimated to have decreased by 34.18% [87]. In these two studies, Zn, functioning as a trace element, may induce short-term stimulation of soil respiration in the case of singular moderate Zn pollution. However, when experiencing mixed pollution comprising Zn (at a low level) and Cr, it manifested as an inhibition of microbial activity, resulting in a reduction in the soil respiration rate. It is important to acknowledge that Beata et al. used a glucose induction method to measure soil respiration, which may have overestimated the respiration rate of contaminated soil.

Table 2. Comparative analysis of heavy metal pollutants: soil impact, human health effects, microbial responses, and greenhouse gas emission influences.

Type	Object			C Emissions	Reference
	Soil	Human	Microbe
Hg: CH3Hg+, Hg2+	Present in soil through volatilization, dissolution or adsorption to the surface of soil particles.	Susceptible to mercury toxic encephalopathy, mental retardation, vomiting and diarrhea, and increased risk of cancer.	Low Hg concentrations are tolerated by microorganisms, reducing toxicity, while high concentrations inhibit their growth and reduce numbers.	Hg and Cd promote methanogenic archaea, increasing CH4 emissions but reducing CO2 emissions.	[87,88,89]
Cd: Cd2+	Excessive cadmium adsorption alters soil properties, causing particle cementation and impeding water infiltration.	Susceptible to chronic obstructive pulmonary disease, lung cancer, cadmium bone disease, kidney cancer, and damage to the cardiovascular system.	Cd disrupts microbial proteins/enzymes via sulfhydryl binding, inhibiting activity, damaging cells, and affecting metabolism. Some microbes tolerate and absorb Cd.
Cu: Cu2+	High concentrations reduce soil pH, fertility, disrupt soil structure and reduce soil water retention.	Moderate Cu intake is necessary, but excess can lead to Cu toxicity, Wilson’s disease, and interference with micronutrient absorption.	Impacts urease activity, reduces oxidative potential, inhibits microbial growth; Methylomonas, Methylobacter, and Methylomicrobium exhibit tolerance.	Increased emissions of CH4 and CO2 from wetland sediments.	[88,89,90,91,92]
Zn: Zn2+	High concentration of Zn alters soil bicarbonate and organic matter content and affects soil pH.	Zn deficiency affects the sense of taste, leads to anemia and dermatitis, and weakens the immune system. Excessive intake can lead to nerve and liver damage.	Zn deficiency impairs energy metabolism, enzyme activity, and protein and DNA synthesis. Excess Zn inhibits enzyme activity and alters nutrient metabolism.	Insignificant effect on CH4, promotes CO2 emissions.	[88,89,93,94]
Pb: Pb2+	Reduces soil pH, affects soil adsorption capacity, and reduces soil fertility.	Causes chronic kidney disease, bone pain, and anemia. Learning disabilities, mental decline irritability, and memory loss.	Lead disrupts cell membranes and binds to protein and nucleic acid groups, causing conformational changes, inhibiting cell division, and inducing protein denaturation.	Pb inhibits soil microorganisms, reducing both microbial respiration and CH4 production.	[83,89,95,96]
Cr: Cr6+	Reduces soil pH, changes soil color, and reduces the rate of organic matter decomposition and nutrient cycling.	Chronic bronchitis, pulmonary fibrosis, dermatitis, increased risk of cancer.	Decreased urease, nitrate reductase, and catalase activities, inhibited microbial metabolism, and reduced the abundance of microbial communities.	Stress prompts bacteria to use more energy, increasing CO2 production.	[89,97]

Table 2. Comparative analysis of heavy metal pollutants: soil impact, human health effects, microbial responses, and greenhouse gas emission influences.

Type	Object			C Emissions	Reference
	Soil	Human	Microbe
Hg: CH3Hg+, Hg2+	Present in soil through volatilization, dissolution or adsorption to the surface of soil particles.	Susceptible to mercury toxic encephalopathy, mental retardation, vomiting and diarrhea, and increased risk of cancer.	Low Hg concentrations are tolerated by microorganisms, reducing toxicity, while high concentrations inhibit their growth and reduce numbers.	Hg and Cd promote methanogenic archaea, increasing CH4 emissions but reducing CO2 emissions.	[87,88,89]
Cd: Cd2+	Excessive cadmium adsorption alters soil properties, causing particle cementation and impeding water infiltration.	Susceptible to chronic obstructive pulmonary disease, lung cancer, cadmium bone disease, kidney cancer, and damage to the cardiovascular system.	Cd disrupts microbial proteins/enzymes via sulfhydryl binding, inhibiting activity, damaging cells, and affecting metabolism. Some microbes tolerate and absorb Cd.
Cu: Cu2+	High concentrations reduce soil pH, fertility, disrupt soil structure and reduce soil water retention.	Moderate Cu intake is necessary, but excess can lead to Cu toxicity, Wilson’s disease, and interference with micronutrient absorption.	Impacts urease activity, reduces oxidative potential, inhibits microbial growth; Methylomonas, Methylobacter, and Methylomicrobium exhibit tolerance.	Increased emissions of CH4 and CO2 from wetland sediments.	[88,89,90,91,92]
Zn: Zn2+	High concentration of Zn alters soil bicarbonate and organic matter content and affects soil pH.	Zn deficiency affects the sense of taste, leads to anemia and dermatitis, and weakens the immune system. Excessive intake can lead to nerve and liver damage.	Zn deficiency impairs energy metabolism, enzyme activity, and protein and DNA synthesis. Excess Zn inhibits enzyme activity and alters nutrient metabolism.	Insignificant effect on CH4, promotes CO2 emissions.	[88,89,93,94]
Pb: Pb2+	Reduces soil pH, affects soil adsorption capacity, and reduces soil fertility.	Causes chronic kidney disease, bone pain, and anemia. Learning disabilities, mental decline irritability, and memory loss.	Lead disrupts cell membranes and binds to protein and nucleic acid groups, causing conformational changes, inhibiting cell division, and inducing protein denaturation.	Pb inhibits soil microorganisms, reducing both microbial respiration and CH4 production.	[83,89,95,96]
Cr: Cr6+	Reduces soil pH, changes soil color, and reduces the rate of organic matter decomposition and nutrient cycling.	Chronic bronchitis, pulmonary fibrosis, dermatitis, increased risk of cancer.	Decreased urease, nitrate reductase, and catalase activities, inhibited microbial metabolism, and reduced the abundance of microbial communities.	Stress prompts bacteria to use more energy, increasing CO2 production.	[89,97]

In general, heavy metal pollution significantly reduced CO₂ emissions from wetlands, with the inhibitory effect being positively correlated with the type and concentration of pollution, and influenced by soil conditions. Since microorganisms are sensitive to heavy metals, the respiration rate can be greatly reduced by low-dose pollution.

6.2. Impact of Wetland Heavy Metal Pollution on CH₄ Emissions

Cadmium (Cd) and copper (Cu) could inhibit CH₄ production, and their inhibition effects were correlated with suppressed activity of methanogenic archaea [98]. It was published that 4 mg kg⁻¹ Cd(II) treatment reduced CH₄ production by 16–99% [99]. This inhibition was likely due to Cd(II) suppressing the activity of enzymes involved in organic matter degradation, reducing the substrates available for methanogenic archaea [99]. Cu(II) also markedly inhibited CH₄ production. Cu(II) toxicity is more pronounced in hydrogenotrophic methanogens, with 8.9 mg L⁻¹ Cu(II) causing a 50% decline in CH₄ production [100]. Mechanistically, Cu(II) disrupts the integrity and selective permeability of archaeal cell membranes by binding to amino, carboxyl, and phosphate groups, compromising archaeal cells [101]. Additionally, Cu(II) can competitively inhibit the activity of key archaeal enzymes like methane monooxygenase, further impeding CH₄ production [100]. However, methanogenic archaea exhibit higher tolerance to heavy metals [87]. This relative resilience allows methanogenic archaea to persist even under heavy metal stress.

Heavy metals also have significant effects on the AOM process. For example, iron (Fe) and manganese (Mn), which are important metastable metal elements in wetland environments, can act as electron acceptors in the AOM process. Beal et al. [102] presented preliminary evidence for AOM-dependent Fe and Mn in California watersheds. Egger et al. [103] found that moderate amounts of Fe(III) (48 $\mu \mathrm{m}\mathrm{o}\mathrm{l} {\mathrm{y}}^{-1} {\mathrm{c}\mathrm{m}}^{-3}$) and Mn(IV) (56 $\mu \mathrm{m}\mathrm{o}\mathrm{l} {\mathrm{y}}^{-1} {\mathrm{c}\mathrm{m}}^{-3}$) stimulated the rate of AOM by 5 and 3.5 times, respectively.

In addition to the involvement of Fe(III) and Mn(IV), numerous other metals including As(V), selenium (Se(VI)), Cu(II), Cr(VI), alum (V(V)), antimony (Sb(V)), and tellurium (Te(IV)) have also been found to mediate AOM, as documented in

Table 3. Research progress on Metal–AOM (anaerobic oxidation of methane) including metal types, conclusions, year, and references.

Metal Type	Conclusions	Year	Reference
Fe(II), Fe(III), Mn(II)	The first discovery of Fe- and Mn-dependent AOM has important global implications.	2009	[102]
Fe(II), Cu(II)	Fe(II) and Cu(II) (20 μmol·L−1 and 10 μmol·L−1, respectively) significantly promoted the activity and growth of N-DAMO bacteria.	2015	[107]
Cr(VI)	The bioreduction of chromate (Cr(VI)) with CH4 as the only electron donor was demonstrated for the first time.	2016	[108]
Sb(V)	CH4 was used as the sole electron donor for the reduction of antimonate (Sb(V)) in an anaerobic membrane biofilm interstitial reactor (MBBR). The results suggest that CH4 may play an important role in the Sb(V) reduction process.	2018	[109]
Cr(VI)	In a membrane biofilm reactor (MBfR), methanotrophic bacteria can reduce chromate. High Cr(VI) loading reduces the Cr(VI) flux by inhibiting CH4 oxidation.	2019	[110]
Te(IV)	This study confirms the ability of a CH4 matrix biofilm reactor (MBfR) in tellurium reduction and generation.	2019	[111]
V(V)	Vanadate reduction can be achieved via the combination of Methylomonas with CH4 oxidation or heterotrophic vanadate reducers with volatile fatty acids (VFAs) as intermediates, while the introduction of nitrate can inhibit vanadate reduction.	2020	[112]
Mn(IV)	Evidence of Mn(IV)-coupled methane anaerobic oxidation was found in freshwater ecosystems, and the microorganism that may be associated with Mn(IV)-coupled methane anaerobic oxidation was proposed as ANME-2d.	2020	[25]
Fe(III)	Cultured sediments from Kinneret Lake (Israel) showed the presence of Fe–AOM in the methanogenic zone, which removed about 10% to 15% of the CH4 produced in the lake sediments.	2022	[113]

. Cantera et al. [90] found that increasing the Cu concentration from 0.05 uM to 25 uM led to a threefold increase in the maximum methane consumption rate (qmax), primarily regulating the methane-oxidizing bacterial community structure and methane monooxygenase expression. This significantly improved CH₄’s biodegradation kinetics, enhanced the CH₄ oxidation capacity, and reduced microbial CH₄ emissions. The copper (Cu) concentration has complex effects on methane-oxidizing bacteria; higher concentrations could inhibit soluble methane monooxygenase (sMMO) but enhance particulate methane monooxygenase (pMMO) expression, increasing CH₄ oxidation rates, while lower Cu concentrations promote sMMO expression and boost CH₄’s degradation rates [104]. Shi et al. [105] provided evidence that As can act as an alternative electron acceptor for AOM, where AOM bacteria reduce As through reverse methanogenesis and respiration. They further found that the genes responsible for these functions were prevalent in the natural environment, thereby indicating that this process may be a potentially global process that has been previously overlooked. Furthermore, Lai et al. [106] demonstrated the complete reduction of selenate to SeO in a membrane bioreactor using CH₄ as the sole electron donor, and suggested that Methylomonas could directly reduce SeO₄²⁻ to SeO while oxidizing CH₄, thereby eliminating the toxicity of SeO₄²⁻.

7. The Impact of Wetland Microplastic Input on CO₂ and CH₄ Emissions

With the widespread use of new plastic materials, microplastic (MP) pollution has become a global environmental problem [114]. Studies have shown that MP contamination is also present to varying degrees in wetland soils [115]. For example, the MP content in the surface soil of wetlands in the Yangtze River Delta, China, ranged from 1.62 ± 0.61 × 10⁵ to 4.25 ± 3.87 × 10⁶ items km⁻² [116]. The MP content in the wetlands of Park’s Estuarine Sanctuary, USA, was as high as 1 × 10⁵ to 7 × 10⁵ items km⁻², mainly in the form of fragment and foam [117]. This is mainly due to the input of MPs from wastewater effluent discharge, land application of biosolids, agricultural runoff containing plastic mulch fragments, and atmospheric deposition [118]. The impact of MPs on the environment is multifaceted (Figure 2). First, MPs can cause mechanical damage to organisms upon ingestion and release chemicals such as plasticizers during migration and transformation, resulting in toxic effects including reduced feeding efficiency, effects on development and reproduction, and even death [119]. Second, MPs have the ability to adsorb heavy metals, antibiotics, and other contaminants, leading to their long-term coexistence and subsequent enrichment in organisms. Consequently, the toxic effects of these pollutants may be altered, with compounding effects on organisms and changes in the physicochemical properties of the environment, such as soil pH and bulk weight [120]. In addition, MPs can act as carriers for microorganisms, algae, and insects, positively and negatively affecting these organisms as they undergo bioenergetic transport, including biological invasions [121]. Thus, MPs in soil have the potential to pose varying degrees of risk to the environment, plant and animal health, and further threaten human health through food chain bioconcentration. Although MPs have been detected in many wetlands around the world, the effects of MPs on C-cycling microorganisms are not fully understood.

Taken together, the widespread use of plastics has introduced MPs into wetland soils, affecting microorganisms, soil properties, and potentially human health through the food chain. We thus detail the impact of MP input on wetlands’ carbon cycles, especially on CO₂ and CH₄ emissions. However, it must be noted that research on the influence of microplastics on wetland carbon cycling processes, particularly greenhouse gas emissions, is still in the preliminary stages of investigation. Most current studies on microplastics are mostly conducted under controlled laboratory conditions. The ecological effects of microplastics and their specific mechanisms of impact on carbon-cycling microorganisms require further field-based empirical research and systematic integration.

As an emerging environmental threat, the CiteSpace analysis of 87 publications indicates that research on microplastic inputs to wetlands is rapidly expanding in both scope and depth. Interactions between microplastics and wetland carbon pools represent a leading research frontier. In particular, growing attention has been paid to how microplastics regulate the transport and transformation of dissolved organic carbon (DOC), with implications for carbon mobility and greenhouse gas emissions [122,123]. Another key hotspot relates to the biodegradation potential of microplastics. Recent studies increasingly explore microbe- and photocatalytic-based degradation technologies and assess how these degradation processes feed back into greenhouse gas emissions [33]. However, some studies indicate that higher biodegradability may be associated with a greater greenhouse gas (GHG) emission burden [124,125]. This line of research is widely regarded as a critical direction for future mitigation strategies in wetland ecosystems.

7.1. Impact of Wetland Microplastic Input on CO₂ Emissions

With the extensive use of new plastic materials, various types of MP pollution have become a major environmental pressure on wetland ecosystems. There are differences in the effects of different types of MPs on soil microbial activity and CO₂ emissions [126]. MPs are synthetic substances mainly composed of carbon, hydrogen, and oxygen, which are man-made substances with high carbon content [127]. Their characteristic of not being easily decomposed allows them to accumulate in the soil. Studies have shown that their abundance is positively correlated with the organic carbon content in all study areas where MP pollution exists [128]. In addition, the readily decomposable fraction of MPs can be utilized by microorganisms as a source of soil microbial carbon or organic substrate to promote the growth of associated functional bacteria [129]. For example, Auta et al. [130] demonstrated that Bacillus sp. strain 27 and Rhodococcus sp. strain 36 can utilize PP-MPs for growth in the sediment of mangroves.

Different types of MPs affect the activity of soil microorganisms through a variety of mechanisms, thus altering CO₂ emissions [131]. These variations could be influenced by factors such as the type of MP, the dosage, exposure duration, and the specific soil properties [132]. MPs generally increase soil respiration by 18.2%, with polypropylene (PP) causing the most pronounced effect (58.8%). Alkaline soils show a 77.9% increase in soil respiration without plants [132]. PP microplastics significantly enhance enzyme activity, particularly β-glucosidase, acid phosphatase, and FDA hydrolase. Conversely, polyethylene terephthalate (PET) and polyethylene (PE) microplastics suppress soil enzyme activity, including urease, β-glucosidase, and hydrogen peroxidase [132].

Industrial plastics contain various additives, including flame retardants, plasticizers, and ultraviolet (UV) stabilizers, some of which have been shown to affect the composition and function of microbial communities [133]. Plasticizers, such as phthalates, can serve as carbon and energy sources for microorganisms, undergoing degradation and consumption by microbial communities [134]. Photosynthetic microorganisms, such as cyanobacteria, are early colonizers of MP surfaces, contributing to biofilm formation and subsequent CO₂ fluxes. However, specific data on the effects of different types and concentrations of plasticizers on microbes and CO₂ emissions are still insufficient, necessitating further research. Additionally, the degradation of additives may complicate the assessment of plastic biodegradation.

The available data demonstrate that UV stabilizer-treated MPs exhibit 136–148% greater surface adhesion, facilitating initial microbial attachment. UV stabilizers induce the formation of reactive oxygen species (ROS) within microbial cells, promoting biofilm formation [135]. While the toxicity of UV stabilizers can initially kill attached microorganisms, these deceased microorganisms can form a protective layer, inhibiting further UV stabilizer release and enabling subsequent microorganisms to grow on the biofilm of the deceased. Since biofilm formation stimulates microbial growth and metabolism, MP additives may indirectly increase CO₂ emissions by promoting biofilm formation. However, specific quantitative data on this phenomenon require further investigation. In contrast to plasticizers and UV stabilizers, flame retardants have been found to inhibit microbial growth and metabolic activities, potentially reducing CO₂ emissions. For instance, a study demonstrated that the presence of polybrominated diphenyl ethers (PBDEs), a common flame retardant, significantly decreased the abundance and diversity of soil microbial communities, leading to a corresponding reduction in CO₂ emissions [136].

Overall, the impact of MP additives on CO₂ emissions is multifaceted. Additives such as plasticizers and UV stabilizers may indirectly contribute to CO₂ emissions by increasing microbial growth and biofilm formation, whereas other additives (e.g., flame retardants) may have an inhibitory effect. Further studies are needed to quantify the specific effects of different additives and their concentrations on microbial communities and associated CO₂ emissions. In addition, potential interactions between the various additives and their combined effects deserve to be investigated.

7.2. Impact of Wetland Microplastic Input on CH₄ Emissions

MPs exhibit diverse impacts on microbially mediated CH₄ emissions, primarily influenced by their type, dosage, and exposure duration [137,138]. Polystyrene (PS) and polyvinyl chloride (PVC) microplastics are strongly associated with the highest CH₄ yield [139,140], while other MP types, such as PE and PP, typically result in lower CH₄ yields [139]. This discrepancy is primarily attributed to the varying interaction mechanisms between different MPs and microorganisms, which consequently impact the methanogenic pathway. For instance, PVC microplastics significantly promoted the accumulation of protein/phenol-like substances in soil dissolved organic matter, enriching methanogenic rice cluster bacteria, thereby elevating CH₄ emissions. PS microplastics stimulated hydrolytic enzyme activities and upregulated genes associated with both acetotrophic and hydrogenotrophic methanogenesis pathways, ultimately enhancing CH₄ yield [138]. In contrast, PE and PP microplastics tend to enrich denitrifying and anaerobic ammonium-oxidizing bacteria, which may compete with methanogens for substrates, suppressing CH₄ production [137].

Increasing the MP concentration inhibits CH₄ production, with CH₄ yield decreasing by 16–27% for every 0.1 g L⁻¹ increase in the MP concentration. This inhibition is partially attributed to the increased production of oxidative stress factors at higher MP concentrations, which negatively impact methanogenic archaea [140].

Furthermore, MPs can impair methane-associated microorganisms through various mechanisms [141]. For instance, under acidic soil conditions (pH 5.0), MPs significantly reduced methanogenic archaea activity by inhibiting the abundance of the CH₄ synthase gene mcrA [142], leading to a 16.9% reduction in CH₄ production. Concurrently, MPs can decrease the copy number of genes related to electron transfer, reduce the content of cytochrome, and decrease the activity of the electron transfer system, indirectly affecting CH₄ production and oxidation processes [131,143].

The impact of MPs extends beyond their physical presence due to the leaching of chemical additives. For instance, PVC microplastics significantly affect methane-associated microorganisms by releasing plasticizers, mainly bisphenol A (BPA). At low concentrations (10 particles g⁻¹ TS), PVC microplastics promote CH₄ production by inducing solubilization and increasing CH₄ yield by approximately 5.9% [144]. However, at high concentrations (≥20 particles g⁻¹ TS), BPA inhibits CH₄ production, leading to a 9.4–24.2% reduction in yield [144].

In summary, MPs exhibit diverse impacts on microbially mediated CH₄ emissions, primarily influenced by their type, dosage, and exposure duration. These impacts are mediated through various mechanisms, including alterations in microbial community composition, the inhibition of methanogenic pathways, reduction of key gene copy numbers, impairment of electron transfer systems, and the release of chemical additives. Understanding these mechanisms is crucial for developing strategies to mitigate the adverse effects of MP pollution on greenhouse gas emissions from agricultural fields and anaerobic digesters.

8. Synergistic Bioengineering Approaches for Wetland Restoration and Carbon Emission Reduction

8.1. Pathways and Mechanisms of Bioremediation for Salinization and Carbon Emission Reduction

Mitigating wetland salinization and achieving carbon reduction is a systematic ecological engineering task that requires a variety of synergistic measures (

Table 4. Pathways and mechanisms for salinization remediation, including methods, categories, microbial mechanisms, and references.

Method	Category	Microbial Mechanism	Reference
Long-term rice cultivation	Phytoremediation	Improve soil physicochemical properties, promote microbial functions related to N, C, and S cycles.	[139]
Microbial fertilizer	Bioremediation	Increasing the mass ratio of large aggregates and decreasing the mass ratio of silt and clay particles resulted in the accumulation of SOC mainly in large aggregates and silt and clay particles.	[145]
Bioorganic fertilizer	Bioremediation	Enrich beneficial microbes, alter microbial network structure. Lignite-based fertilizer enhances alkali grass’ multi-metal saline soil remediation.	[146]
Humic acid and organic fertilizer	Bioremediation	Reduce microbial restrictions on N and P.	[147]
Cotton straw return and subsoiling	Combination of agricultural practices and bioremediation	Increase soil microbial abundance and diversity, enhance Bacillus, Pseudomonas, Fusarium, and Alternaria in carbon and nitrogen fixation.	[148]
Plant microbe desalination cell	Combination of electrochemical remediation and bioremediation	Dominant electrogenic bacteria and Spirochaetes in the anode, reduces soil electrical conductivity and pH, removes ions like sodium, potassium, and chloride.	[149]

). In recent years, biotechnological approaches to ecological restoration have garnered widespread attention due to their sustainability and efficiency advantages.

Bio-organic fertilizers, particularly those based on lignite, are an effective biotechnological measure that can enhance the multi-metal saline soil remediation ability of alkali grass [147]. Both manure-based organic fertilizer (MOF) and lignite-based organic fertilizer (LOF) can enrich beneficial microorganisms and modulate the microbial network, enhancing plant adaptation to heavy metals and salt stress, with bacteria playing a crucial role, especially in LOF treatment. In essence, bio-organic fertilizers, particularly those based on lignite, can regulate the soil microbial community, improving the remediation effectiveness of alkali grass in multi-metal saline soils.

Microbial fertilizers, primarily composed of fermented organic matter, can mitigate soil salinization in low-salinity soils (EC < 1.45 dS m⁻¹) by increasing the soil organic carbon content and promoting soil aggregate formation [145]. However, their effects are limited under high-salinity conditions (EC ≥ 1.45 dS m⁻¹), possibly due to the inability of some microorganisms to tolerate high-salt environments.

In addition to the application of bio-organic fertilizers, certain agricultural practices can also be used for the amelioration of saline soils. Long-term straw mulching can increase microbial diversity and the carbon and nitrogen fixation capacity in coastal saline cotton fields, while deep plowing, although reducing soil pH and salinity and increasing the abundance of dominant microorganisms, does not further affect microbial carbon and nitrogen fixation [148]. In this process, specific plant-associated bacteria and fungi, such as Bacillus, Pseudomonas, Fusarium, and Alternaria, play a key role in enhancing carbon and nitrogen fixation. In summary, these practices can optimize the microbial communities in coastal saline soils, improving carbon and nitrogen utilization and ensuring sustainable land use. Furthermore, long-term rice cultivation has also been proven to significantly improve the physicochemical properties of saline soils and promote microbial energy metabolism functions, especially those related to nitrogen, carbon, and sulfur cycles, with these effects peaking at 11–16 years of cultivation. Long-term rice cultivation can effectively restore the physicochemical balance and microbial functions of saline soils without other chemical or physical remediation measures, exhibiting sustainability and synergy, and providing a new perspective on saline soil management. However, it generally requires a longer time and may not be applicable to lands that require rapid restoration. Considering the growth characteristics of rice, this approach may be limited by geographic and climatic conditions in some areas where rice cultivation is impractical or inapplicable.

For the amelioration of saline–alkali soils, researchers have proposed plant microbe desalination cell (PMDC) technology [149]. Compared to soil microbe desalination cells (SMDCs), PMDC technology more effectively reduces soil electrical conductivity and pH, removes ions such as sodium, potassium, and chloride, and improves soil physicochemical properties. PMDC technology combines plant remediation, microbial cells, and electrodialysis effects, representing a more effective new method for ameliorating saline–alkali soils.

8.2. Pathways and Mechanisms of Bioremediation for Over-Fertilization and Carbon Emission Reduction

Currently, the DAMO reaction, which utilizes CH₄ as an electron donor for NO₃⁻/NO₂⁻ reduction, is widely regarded as a crucial link between global C and N nutrient cycles [150]. The ability of M. oxyfera and Anammox bacteria to jointly utilize CH₄ and NH₄⁺ as electron donors suggests that the shared functional role of both could be utilized for soil denitrification in soils experiencing severe nitrogen pollution [151]. This approach could significantly address nitrate and ammonium nitrogen pollution resulting from excessive use of nitrogen fertilizers over extended periods in wetland soils and could help mitigate the greenhouse effect.

The study [152] proposes two methods to mitigate excessive nitrate accumulation in soils and plants: the use of nitrogen oxide inhibitors dicyandiamide (DCD) and 3,4-dimethylpyrazole phosphate (DMPP), and their combination with graphene oxide (GO). The research reveals that DCD/DMPP reduces the soil nitrate content by inhibiting the activity and abundance of functional nitrogen-cycling microorganisms, whereas GO contributes to nitrate reduction in soil by enhancing certain soil enzyme activities and altering the microbial composition. Furthermore, DCD/DMPP alters the nitrogen form ratio, hindering the conversion of ammonium nitrogen to nitrate, thus lowering nitrate levels in plants.

The anther mitigation of excessive nitrogen fertilization through reduced subsequent chemical N input (RCN) and organic fertilization (OF) improves soil health by leveraging microbial mechanisms [153]. RCN reverses soil N enrichment, alleviates microbial N inhibition, and enhances the beneficial bacterial community structure and interactions, thereby increasing community stability. OF, on the other hand, substitutes chemical fertilizers, promotes microbial diversity, and enhances soil multifunctionality. Together, these strategies synergistically counteract the adverse effects of excessive N fertilization, fostering a resilient and sustainable soil microbiome and reducing chemical fertilizer dependence.

8.3. Pathways and Mechanisms of Bioremediation for Heavy Metal Pollution and Carbon Emission Reduction

Heavy metal pollution and greenhouse gas emissions are two major environmental challenges we face today. Traditional physical and chemical remediation methods are not only costly but may also lead to secondary pollution, prompting the need for efficient, cost-effective, and environmentally friendly bioremediation techniques [154]. In recent years, microbe-based remediation strategies have shown great potential due to their unique mechanisms, not only for remediating heavy metal-contaminated soils but also contributing to climate change mitigation (

Table 5. Pathways and mechanisms for heavy metal pollution remediation, including methods, categories, microbial mechanisms, and references.

Method	Category	Microbial Mechanism	Reference
Nano-Manganese Dioxide-Coated Microbial Communities (NMCMP)	Phytoremediation	Coating with nano-manganese dioxide preserves enzyme activity. Supports microbial vitality and functionality under adverse conditions. Catalytic enzymes enable reduction and oxidation reactions simultaneously. Promotes key remediation microbes (e.g., Flavisolibacter, Arthrobacter). Enhances remediation efficiency of hexavalent chromium. Inhibits release of trivalent arsenic.	[155]
Microbial Fuel Cells (MFCs)	Bioremediation	Efficient removal of heavy metals like Cr3+, Cu2+, and Pb2+. Influences heavy metal speciation and bioavailability. Supports heavy metal transformation and immobilization.	[156]
Combination of Inorganic Amendments, Organic Fertilizers, and Biochar	Bioremediation	Increases abundance of bacterial phyla (e.g., Chloroflexi, Rokubacteria, and Firmicutes). Reduces bioavailability of heavy metals like cadmium, lead, and chromium. Achieves heavy metal stabilization and insolubilization. Increases functional microbes involved in organic matter decomposition and denitrification.	[157,158]

).

One innovative technique for remediating heavy metal-polluted soils is the use of Nano-Manganese Dioxide-Coated Microbial Communities (NMCMP) [155]. The NMCMP technology significantly alters the microbial community composition, preserving soil microbial diversity and promoting the proliferation of key remediation microorganisms such as Flavisolibacter and Arthrobacter. Studies have shown that NMCMP technology not only enhances the remediation efficiency of hexavalent chromium but also inhibits the release of trivalent arsenic, achieving combined detoxification.

Another promising remediation technique is the use of Microbial Fuel Cells (MFCs) [156]. In an MFC system, electrogenic microorganisms at the anode utilize readily degradable organic compounds to produce electrons, which are transferred through an external circuit to the cathode, generating an electric field between the electrodes. This electric field drives cationic heavy metal ions to migrate towards the cathode, where some ions (e.g., Cu²⁺) are reduced and deposited, achieving removal. Apart from removing heavy metals, MFC systems can also alter soil properties like Eh and pH, influencing heavy metal speciation and bioavailability; compete with other microorganisms for organic compounds, limiting their mineral reduction activity; and modulate microbial community structure, facilitating heavy metal transformation and immobilization.

Moreover, optimizing the combination of inorganic amendments, organic fertilizers, and biochar has proven to be an effective remediation strategy. For instance, the S1A0.5B1.5 combination (seaweed fertilizer 1 part: phosphate rock 0.5 parts: biochar 1.5 parts) can optimize the soil bacterial community structure, significantly increasing the abundance of bacterial phyla such as Chloroflexi, Rokubacteria, and Firmicutes [157]. The enhanced bacterial activity further reduces the bioavailability of heavy metals like cadmium, lead, and chromium in the soil, achieving heavy metal stabilization and insolubilization. Notably, the S1A0.5B1.5 combination also increases the relative abundance of functional microbes involved in organic matter decomposition and denitrification, improving overall soil functionality and contributing to carbon sequestration and greenhouse gas mitigation.

These bioremediation techniques achieve the efficient remediation of heavy metal-contaminated soils by modulating microbial activity, community composition, functional diversity, and interactions with heavy metals. Simultaneously, they contribute to climate change mitigation through novel bioremediation pathways. During the heavy metal remediation process, microorganisms enrich functionally relevant strains closely associated with carbon sequestration (such as those involved in organic matter decomposition and denitrification) through mechanisms like redox reactions, electron transfer, and enzyme production. Consequently, these remediation techniques achieve a dual effect of heavy metal pollution control and greenhouse gas mitigation to a certain extent.

8.4. Pathways and Mechanisms of Bioremediation for Microplastic Input and Carbon Emission Reduction

Microorganisms exert intricate mechanisms to combat MP pollution in soil, constituting a multifaceted approach crucial for environmental sustainability (

Table 6. Pathways and mechanisms for microplastic input remediation, including methods, categories, mechanisms, and references.

Method	Category	Microbial Mechanism	Reference
Microbial degradation	Bioremediation	Enzyme-mediated degradation by microorganisms (e.g., Pseudomonas spp., Bacillus spp., Bacillus thuringiensis etc.). Microbial biofilm formation on microplastic surfaces. Direct microbial utilization of microplastics as carbon source modulation of soil microbial community structure and function.	[159,161]
Microbial fertilizer	Bioremediation	Increasing the mass ratio of large aggregates and decreasing the mass ratio of silt and clay particles resulted in the accumulation of SOC, mainly in large aggregates and silt and clay particles.	[145]
Soil fertilizer system	Bioremediation	Organic fertilizers increase oxidized functional groups on microplastic surfaces. Enrichment of microbial taxa like Ascomycetes and Actinobacteria.	[141]
Biochar remediation	Integrated remediation	Improvement of soil physicochemical properties. Mitigation of detrimental effects of PVC microplastics on soil microorganisms.	[162]
Microbial worm degradation	Bioremediation	Microbial biofilm formation and enzymatic degradation. Worms facilitate degradation and provide suitable microenvironment.	[163]
Policy and management	Administrative remediation	Development of biodegradable plastic. Improvement of agricultural film recovery rate. Proper treatment of agricultural wastes.	[164]

). One of the primary mechanisms is enzyme-mediated degradation. Microorganisms secrete hydrolytic enzymes and oxidoreductases that catalyze the breakdown of plastic polymers into smaller, metabolizable molecules [159]. For instance, bacteria such as Pseudomonas, Bacillus, and Arthrobacter are proficient in degrading phthalate-based plasticizers [160]. This illustrates that the microbial enzymatic activity has efficiency and potential in reducing plastic pollution.

Another pivotal strategy employed by microorganisms is biofilm formation. Microorganisms aggregate and adhere to MP surfaces, creating microenvironments that facilitate enhanced enzymatic degradation [163]. This process significantly increases the surface area available for microbial activity, thereby boosting degradation efficiency. The role of organic fertilizers in promoting biofilm formation has been well documented. For instance, Zhang et al. [141] found that the addition of organic fertilizers in black soil significantly increased the diversity of biofilms on MP surfaces. These biofilms, composed mainly of Ascomycetes and Actinomycetes, enhance the degradation process by creating favorable conditions for microbial activity [141].

In addition to enzymatic activity and biofilm formation, microorganisms can directly utilize MPs as a carbon and energy source [161]. Microorganisms metabolize microplastic-derived carbon through various pathways, contributing to their growth and the further degradation of MPs. Some degrading microorganisms, such as Bacillus spp., Bacillus thuringiensis, and degrading enzymes involved in hydrolysis and oxidation reactions were identified in studies [163]. This direct utilization not only aids in reducing MP pollution but also supports microbial proliferation and ecosystem health.

Furthermore, the incorporation of biochar into soil emerges as a promising strategy for mitigating MP pollution. Research shows that biochar-amended soils exhibit increased microbial diversity and enzymatic activity, which in turn enhances MP degradation rates. Moreover, biochar improves soil fertility and water retention, creating a more conducive environment for microbial processes [162].

Microbial degradation methods also involve the adhesion of microorganisms to plastic surfaces, secretion of hydrolytic enzymes, and subsequent metabolization of plastic-derived carbon sources into CO₂ and H₂O [161]. This comprehensive process is facilitated by the physicochemical properties of MPs, which influence microbial interactions and degradation efficiency. Additionally, environmental conditions like pH and salinity can be adjusted to optimize these microbial processes.

In conclusion, the detailed understanding of microbial mechanisms for MP degradation, coupled with empirical data on enzymatic activities, biofilm formation, and direct carbon utilization, provides valuable insights for the development of effective mitigation strategies. By leveraging microbial processes and soil amendment techniques, it is possible to address soil MP contamination comprehensively and sustainably. These strategies not only enhance the degradation of MPs but also contribute to overall soil health and ecosystem functionality, presenting a promising avenue for environmental remediation.

9. Conclusions and Prospects

Global wetlands, as the core carriers of the blue carbon cycle, have been continuously influenced by various anthropogenic activities. These anthropogenic activities have altered the source–sink roles of wetlands in terms of carbon dioxide and methane emissions through different mechanisms (

Table 7. The impact of various anthropogenic activities on greenhouse gas emissions.

Anthropogenic Activity	Gas Type	Influence	The Main Driving Factors	Condition of the Occurrence	Reference
Salinization	CO2	Inhibition	Inhibition of microbial activity and organic matter mineralization; impairment of plant photosynthesis; ionic competition for essential nutrients	Salinity 3‰ (mild): 15% reduction in CO2 emissions; salinity 10‰ (severe): CO2 emissions decreased to 558 mg C kg−1; more significant inhibition in soils with low pH and low organic matter content	[40,41,42]
	CH4	Inhibition (above 10‰ salinity threshold)	Suppression of methanogen activity; alteration of substrate supply; pH increase induced by high salinity; inhibition of AOM microbial metabolism	Salinity ≥10‰ (threshold): 24% reduction in CH4 production; salinity >292 g L−1 severely inhibits AOM; halophilic archaea can produce CH4 at salinity up to 120 g L−1	[43,44,45,46,47,48,49,50,51,52]
Over-fertilization	CO2	Promotion at appropriate levels, inhibition at excessive levels	Activation of microbial activity by appropriate N/P inputs; soil acidification caused by excessive N; formation of insoluble phosphate salts by excessive P	Nitrogen fertilizer: promotion in the short term, inhibition due to acidification in the long term; phosphorus fertilizer at 4.6 mg P g−1: 857% increase in respiration rate; high phosphorus at 481.9 kg ha−1: 24.81% reduction in acid phosphatase activity	[61,62,63,64,65,66,67,68,69]
	CH4	Nitrogen: emission reduction at low doses, emission increase at excessive doses; phosphorus: long-term inhibition	Enhancement of methanotroph activity by low- nitrogen inputs; inhibition of DAMO process by excessive nitrogen; promotion of vegetation growth and soil aeration by phosphorus; fertilizer type	Nitrogen <100 kg ha−1: 87% reduction in CH4 emissions from swamps; high phosphorus and low nitrogen: 38% reduction in CH4 emissions from Dipper Harbour wetland; organic fertilizer increases methane emissions; combined organic-inorganic fertilizer reduces methane emissions	[60,70,71,72,73]
Heavy Metal Pollution	CO2	Inhibition in most cases	Inhibition of microbial activity; damage to cell membranes and enzyme activity; enhanced heavy metal bioavailability at low pH	Combined pollution of Zn and Cr (Zn 234.82 mg kg−1 + Cr 102.99 mg kg−1): 34.18% reduction in CO2 emissions	[82,83,84,85,86,87]
	CH4	Inhibition in most cases, AOM-facilitated emission reduction by some metals	Inhibition of methanogens by Cd and Cu; Fe(III), Mn(IV), etc., serving as electron acceptors for AOM; regulation of effects by metal concentration and speciation	Cd 4 mg/kg: 16–99% reduction in CH4 production; Fe(III) 48 μmol y−1 cm−3: 5-fold increase in AOM rate; Cu 25 μmol L−1 increased the rate of CH4 consumption	[87,90,98,99,100,101,102,103,104,105,106]
Microplastic Input	CO2	Promotion or inhibition depending on type/additives	PP activates enzyme activity, while PET/PE inhibits enzyme activity; plasticizers and UV stabilizers promote microbial growth; flame retardants inhibit community metabolism	PP microplastics: 58.8% increase in soil respiration; alkaline soil without plants: 77.9% increase in respiration, PBDE flame retardants reduce microbial abundance	[126,127,128,129,130,131,132,133,134,135,136]
	CH4	Promotion or inhibition depending on type/concentration	PS/PVC upregulates methanogenesis-related genes; PE/PP competes with methanogens for substrates; high concentration induces oxidative stress; effects of additive leaching	PS/PVC significantly increase CH4 production; 0.1 g L−1 increase in microplastic concentration: 16–27% reduction in CH4 production; low-concentration PVC (10 particles g−1 TS): 5.9% increase, high-concentration (≥20 particles g−1 TS): 9.4–24.2% reduction	[137,138,139,140,141,142,143,144]

).

Salinization can inhibit wetland CO₂ and CH₄ emissions to varying degrees by inhibiting soil microbial activity and vegetation photosynthesis. However, individual saline microorganisms can partially compensate for the emissions. Microbial responses to salinity varied according to wetland types.

Over-fertilization destroys the microbial structure of wetland soil and affects carbon emissions. Nitrogen fertilizer can stimulate CO₂ emissions in the short term, but inhibit it in the long term. Phosphorus fertilizer can stimulate emissions at a low level, but inhibit it at a high level. The interaction effect increases the uncertainty of prediction.

Different heavy metals can significantly inhibit microbial activity and reduce CO₂ and CH₄ emissions. However, they can also participate in anaerobic CH₄ oxidation within a moderate range. Assessing their overall impact on greenhouse gas emissions is complex.

MP pollution has also become an emerging environmental pressure affecting the carbon cycle in wetlands. The impact on carbon emissions in wetlands varies depending on the type, concentration, exposure time, and additives. Different concentrations and types of MPs can have varying effects on soil microbial activity and CO₂ emissions. They can also influence CH₄ production and oxidation through chemical toxicity and physical effects.

Although existing studies have separately revealed the microbial regulatory mechanisms of single anthropogenic activities on carbon emissions, the interactive effects of multiple anthropogenic activities and the characteristics of microbial functional responses at the global scale remain unclear. This leads to the problem that it is difficult to integrate the conclusions of regional studies. Therefore, future research should establish a systematic observation and experimental framework, focusing on analyzing how complex anthropogenic activities such as salinity, nutrients, heavy metals, and microplastics synergistically regulate the activities of key microbial functional communities, such as methanogens, methane-oxidizing bacteria, and denitrifying bacteria, as well as the complex interactions among them. Additionally, the expression of functional genes such as mcrA and pmoA should be utilized to uncover the carbon cycle process and its microbe-driven mechanisms.

Moreover, the standardization of global wetland carbon emission research should be promoted. This includes unifying the specifications for measuring and calculating greenhouse gas fluxes, establishing a multi-monitoring system that covers key environmental variables such as redox potential, sulfate/nitrate, and dissolved organic carbon, as well as the abundance and diversity of microbial functional genes. Meanwhile, the standards for the sampling, characterization, and quantification of microplastics should be standardized to enhance the standardization and usability of data.

Future research should focus on formulating wetland restoration and management strategies based on microbial functions. At the same time, it is necessary to evaluate the applicability and thresholds of different management measures, such as nutrient interception, salinity regulation, and combined plant–microbe remediation, in various types of wetlands. In summary, future research should provide more standardized and sustainable management approaches for wetlands to cope with human disturbances and climate change.