Soil–Plant Carbon Pool Variations Subjected to Agricultural Drainage in Xingkai Lake Wetlands

Abstract

This study examines the responses of soil organic carbon (SOC) pools and their components to agricultural water drainage in paddy fields, with a focus on the wetland–paddy field ecotone of Xingkai Lake, a transboundary lake shared by China and Russia. Field investigations targeted three representative wetland vegetation types: Glyceria spiculosa (G), Phragmites australis (P), and Typha orientalis (T), across drainage durations ranging from 0 to over 50 years. SOC fractions, including light fraction organic carbon (LFOC), heavy fraction organic carbon (HFOC), dissolved organic carbon (DOC), and microbial biomass carbon (MBC), were systematically analyzed. The results revealed that SOC components in T and P wetlands steadily increased with drainage duration, whereas those in G wetlands exhibited a fluctuating pattern. SOC dynamics were primarily driven by LFOC, while MBC displayed species-specific variations. Correlation analyses and structural equation modeling (SEM) demonstrated that soil physicochemical properties, such as total nitrogen and moisture content, exerted a stronger influence on SOC fractions than microbial biomass. Overall, water drawdown significantly altered SOC dynamics, with distinct responses observed across vegetation types and wetland ages. This study provides critical data and theoretical insights for optimizing carbon sequestration and hydrological management in wetland–paddy field systems.

1. Introduction

Wetlands are critical ecosystems formed by the dynamic interaction between terrestrial and aquatic environments. Due to their high primary productivity and slower rates of organic matter decomposition, wetlands serve as significant carbon sinks in terrestrial ecosystems [1]. It is estimated that wetland ecosystems store approximately 20% to 30% of the global soil carbon in terrestrial environments [2,3]. Soil carbon cycling in wetlands is influenced by a range of environmental factors, with water level fluctuations playing a key role in determining carbon storage and release. This process is significantly influenced by numerous meteorological and environmental factors, and it can exert both direct and indirect effects on the global climate. Hydrological processes within wetlands influence biogeochemical cycles by regulating the soil environment and play a major role in determining the evolutionary trajectory of wetland ecosystems [4,5]. As a result, even subtle changes in the wetland soil carbon-cycling process—such as variations in total soil organic carbon content, shifts in the components of soil organic carbon, and changes in the rate of soil organic carbon mineralization—can lead to significant alterations in atmospheric CO₂ concentrations, thus having a profound impact on the global climate. Soil carbon cycling in wetland ecosystems is an inherently complex process, profoundly shaped by many meteorological and environmental factors, with hydrological rhythms being one of the key influences [6]. These rhythms impact numerous critical processes within the soil carbon cycle of lake wetlands, thereby affecting the broader wetland ecosystem [7]. Consequently, both domestic and international research efforts have extensively studied the effects of hydrological rhythms on the wetland soil carbon cycle [8].

Wetland vegetation is particularly sensitive to disturbances from anthropogenic activities and variations in environmental conditions [9], including changes in land use, fluctuations in water levels, and alterations in vegetation distribution and community composition [10]. As a result of the interactions among these multiple factors, the soil moisture dynamics exhibit seasonal variations [11]. Climate change is expected to affect the fluxes of organic carbon in wetland soils, influencing both the input and output of carbon pools [12]. Particularly under prolonged periods of precipitation or excessive soil moisture, there is a substantial decrease in the concentration of dissolved organic carbon (DOC) [13]. Both drought and increased precipitation, as extreme climate events, can significantly impact microbial community composition and activity, subsequently influencing the mechanisms of carbon cycling [14]. Soil microbial communities are actively involved in the decomposition of crop residues and play a crucial role as mediators in carbon sequestration, facilitating the mineralization and stabilization of soil organic carbon (SOC) through microbial processes [15]. Moreover, elevated soil temperatures may enhance the rate of evapotranspiration, resulting in a decrease in soil water content [16]. As a result, variations in the hydrological gradient of wetlands can significantly influence plant growth and the carbon sequestration capacity of wetland ecosystems [17]. Furthermore, investigating the characteristics of soil active carbon content is crucial for understanding the underlying mechanisms driving the dynamic fluctuations in soil carbon pool dynamics [18]. This is essential for evaluating the potential effects of future agricultural drainage on the functioning and integrity of wetland ecosystems.

The continuous expansion of agricultural production scale and scope, particularly with the implementation of recent hydraulic engineering and river–lake connectivity projects, has, on one hand, facilitated the spatial equilibrium of water resources at the regional scale, improved water use efficiency, and strengthened ecological security. On the other hand, it has led to shifts in the regional agricultural production structure, such as the rapid expansion of rice cultivation areas. These changes have also been accompanied by numerous environmental challenges, including alterations to wetland hydrological regimes and the influx of nutrients into wetlands, much of which have not been fully utilized by agricultural practices. Anthropogenic activities, such as drainage, could result in alterations to the carbon sequestration potential of wetland soils at a local scale [19]. Agricultural drainage has emerged as a critical environmental concern. From the perspective of ecohydrology, examining the hydrological processes of the Xingkai Lake wetland and the impacts of variations in wetland water source quality on the ecosystem is essential for analyzing the hydrological dynamics of wetland ecosystems and their influence on structural and functional components. This is also an emerging issue that must be considered in the ongoing efforts for black soil conservation in China. Additionally, it can offer scientific backing for the conservation of international transboundary lakes and for fulfilling international conventions. This study focused on three representative vegetation-type wetlands within the Xingkai Lake to examine the impact of the increasing duration of agricultural field drainage on the soil carbon fractions.

The objective of our research is (1) to examine the dynamic variations in the soil organic carbon fractions (i.e., HFOC, LFOC, DOC, and SOC) across three representative vegetation-type wetlands under the conditions of no drainage and prolonged drainage; (2) to quantify the relationship between the different soil carbon fractions and the overall soil organic carbon pool; and (3) to determine the key soil environmental factors driving the distribution of soil organic carbon fractions in these wetlands, thereby providing foundational data for future research on wetland organic carbon stocks and carbon sequestration processes.

2. Materials and Methods

2.1. Study Area

The Xingkai Lake wetland (132.3° E, 45.3° N), situated on the Sino-Russian border, is the largest freshwater lake in Northeast Asia, spanning an area of 4380 km² [20]. Xingkai Lake, a transboundary lake shared by China and Russia, is composed of two sections: the large and small Xingkai lakes, which are divided by a natural sandbar roughly one kilometer in width. The two lakes are interconnected by the first and second flood control sluice gates located on the sandbar. The lake basin is characterized by a temperate continental monsoon climate, with minimal precipitation in spring and autumn, higher rainfall during the summer, and a prolonged dry period throughout the winter. The experimental zone lies at the edge of a sand ridge on the northern shore of the small Xingkai Lake, where vegetation exhibits an irregular zonal distribution along the hydrological gradient. The Glyceria spiculosa (G) forms a meadow-like floating mat on the water surface, interspersed with wetland plants such as Phragmites australis (P) and Typha orientalis (T). Xingkai Lake typically freezes in December and thaws by late April. Paddy field irrigation generally takes place from May to October, and is divided into four stages. The first stage is the inundation phase, where water is applied to the paddy fields, resulting in a water depth of approximately 3 to 5 cm. The second stage is the transplanting phase, where the paddy fields are maintained with a water depth ranging from 1 to 3 cm. Following this, during the late tillering phase of rice, the paddy fields are drained to facilitate field drying. The third stage is the re-flooding developmental phase, where irrigation is applied to the paddy fields, resulting in a water depth of approximately 3 to 5 cm. The fourth stage is the drainage and harvest phase, where water is progressively drained from the paddy fields, culminating in the harvesting of the rice crop. This study was conducted in the marsh wetland of the Xingkai Lake. By monitoring the water quality and quantity in paddy fields, as well as examining the physiological and morphological traits of wetland plants (e.g, P, G, and T) and the transfer and allocation of aboveground and belowground biomass, we monitored and analyzed the seasonal dynamics of carbon pools in soil columns, plant growth, carbon fluxes, and soil microbial characteristics. This research aims to predict the changes in soil carbon pools and the dynamic response of carbon emissions, evaluating their feedback mechanisms and interactions with paddy field drainage, ultimately assessing the stability and trends of the carbon pools under such conditions.

2.2. Soil Sample Collections and Laboratory Analyses

On 7 July 2023, six sampling sites were selected around the Xiaoxingkai Lake region, representing wetlands with drainage durations of 0 years, 5–10 years, 10–20 years, 20–30 years, 30–40 years, and over 50 years (T and P were found in three wetlands, respectively). Using a 0.1 × 0.1 m² quadrat and ensuring the intactness of soil structure, plants were sampled with a soil block approximately 20 cm thick. To ensure the accuracy of the samples, three replicate samples were collected at each sampling point. Immediately after sampling, the specimens were transported to the laboratory, where surface litter and the obvious root layer were removed. The samples were dried at 70 °C to a constant weight (the difference between two consecutive weight measurements of the plant is less than 0.01 G), and the dry weight was recorded to calculate the aboveground biomass of the community (g·m⁻²). Subsequently, the soil samples were passed through a 100-mesh sieve (a 100-mesh sieve effectively removes roots, stones, and other large impurities while retaining the fine particles required for chemical analysis) to remove impurities. The soil was then divided into two portions: (i) One portion was air-dried and used for the determination of soil organic carbon (SOC), dissolved organic carbon (DOC), heavy fraction organic carbon (HFOC), light fraction organic carbon (LFOC), total nitrogen (TN), total phosphorus (TP), and available phosphorus (AP). (ii) The other portion was stored at 4 °C and passed through a 2 mm sieve [21] for the determination of microbial biomass carbon (MBC), microbial biomass nitrogen (MBN), and microbial biomass phosphorus (MBP). LFOC was quantified using density fractionation [22]. The sample weight recovery rate was calculated, with all recovery rates exceeding 95%. Recovery rate (%) = (initial sample weight/recovered sample weight) × 100%. MBC in soil was quantified using the chloroform fumigation−K₂SO₄ extraction technique. Fumigated and non-fumigated soil samples were extracted with 0.5 mol·L⁻¹ potassium sulfate (K₂SO₄) (Sinopharm Chemical Reagent Co., Ltd., Shanghai, China) for 30 min, and the organic carbon concentration in the extract was determined using a Shimadzu TOC-VCPH analyzer (Shimadzu Corporation, Kyoto, Japan) [21]. DOC in soil was extracted with 0.5 mol L⁻¹ K₂SO₄ and quantified using a Shimadzu TOC-VCPH analyzer [23]. SOC was measured using the potassium dichromate (Sinopharm Chemical Reagent Co., Ltd., Shanghai, China) oxidation method with external heating [24].

2.3. Statistical Analysis

Before conducting statistical analysis, the experimental data (including soil organic carbon fractions, soil physicochemical properties, microbial biomass, and water samples) were tested for normality (Shapiro–Wilk test) and homogeneity of variances (Levene’s test). For data that met both the normality and homogeneity of variance assumptions, an independent samples t-test was used to analyze differences between treatments. For data with unequal variances, Welch’s t-test was applied, and for data that did not follow a normal distribution, a non-parametric test (Wilcoxon rank-sum test) was employed. The significance level was set at p < 0.05. All data are presented as mean ± standard deviation (SD). Statistical analyses were performed using SPSS 25.0 (IBM CorP, Armonk, NY, USA). Graphs and figures were created using GraphPad Prism 9.0 (GraphPad CorP, La Jolla, CA, USA) and Origin Pro 2019 (Origin Lab CorP, Northampton, MA, USA). Experimental data were derived from three independent biological replicates to ensure the reliability and reproducibility of the results.

3. Results

3.1. The Variations in Surface Water Quality Characteristics Across Wetlands with Different Drainage Duration

The survey identified the dominant plant species in the wetland as G, P, and T. G was observed in wetlands with drainage durations of 0 years (A), 5–10 years (B), 10–20 years (C), 20–30 years (D), 30–40 years (E), and greater than 40 years (F). P was found in wetlands with drainage durations of 0 years (A), 20–30 years (D), and over 40 years (F), while T was present in wetlands with drainage durations of 10–20 years (C), 20–30 years (D), and over 40 years (F). Plants corresponding to the same drainage duration were sampled from the same wetland. The comparison between wetland water and agricultural drainage water is shown in Figure 1. Significant differences (p < 0.05) were observed in the concentrations of COD_Mn, BOD₅, TN, TP, TOC, Fe²⁺, and Fe³⁺ across wetlands with varying drainage durations. The concentrations of COD_Mn, BOD₅, TN, TP, TOC, TC, Fe²⁺, and Fe³⁺ varied within the following ranges: COD_Mn, 8.62–29.12 mg·L⁻¹; BOD₅, 4.45–34.83 mg·L⁻¹; TN, 0.82–2.51 mg·L⁻¹; TP, 0.090–0.603 mg·L⁻¹; TOC, 11.70–41.64 mg·L⁻¹; TC, 27.88–58.92 mg·L⁻¹; Fe²⁺, 0.046–0.21 mg·L⁻¹; and Fe³⁺, 0.034–0.0973 mg·L⁻¹ (Figure 2).

3.2. Variation in Plant Biomass Across Wetlands with Different Drainage Duration

The study results demonstrate significant differences (p < 0.05) in both aboveground and belowground biomass across wetlands with different drainage durations (Figure 3). As shown in Figure 3a, the aboveground biomass of G increases annually, with a decline observed in plot F. Similarly, the belowground biomass of G shows a progressive increase. Biomass dynamics were significantly influenced by drainage duration, with aboveground biomass showing a non-linear pattern, while belowground biomass exhibited a gradual increase. The smallest aboveground biomass was observed in plot F. No significant differences were found in aboveground biomass between plots A and F (p < 0.05), plots C and D (p < 0.05), and plots B and C (p < 0.05). Figure 3b illustrates that both the aboveground and belowground biomass of P increased progressively each year. However, no significant differences were observed between plots A, B, D, and E (p > 0.05). The aboveground biomass in plot F was 16.64% and 13.22% greater than that in plots A and B, respectively, with statistically significant differences (p < 0.05). Plot B exhibited a 125.86% increase in belowground biomass compared to plot A, and the difference was statistically significant (p < 0.05). Plot F exhibited a 162.45% and 16.2% increase in belowground biomass compared to plots A and B, respectively, with statistically significant differences (p < 0.05). The aboveground and belowground biomass of T (Typha angustifolia) (Figure 3c) increased annually. The aboveground biomass in plot D was 14.55% higher than in plot C, although the difference was not statistically significant (p > 0.05). The aboveground biomass in plot E was significantly higher by 114.3% and 87.09% compared to plots C and D, respectively, with statistically significant differences (p < 0.05). The belowground biomass in plot D exhibited a 52.83% increase compared to plot C, with the difference being statistically significant (p < 0.05). The belowground biomass in plot E exhibited increases of 156.18% and 69.59% compared to plots C and D, respectively, with statistically significant differences (p < 0.05). The biomass accumulation trends of the various plant species followed a similar pattern, with the duration of water recession significantly influencing plant biomass (p < 0.05). In all cases, biomass peaked during the longest water recession period. As shown in Figure 3, the root, stem, and leaf metrics of the three plant species growing in wetlands with extended water recession durations were markedly superior to those of G in wetlands with shorter recession periods. This suggests that plants in wetlands with longer water recession periods exhibit more vigorous growth. However, in plot F, the aboveground biomass of G showed a downward trend, indicating that prolonged water recession may suppress plant growth once a critical threshold is exceeded. Roots serve as both active absorption and synthesis organs in plants. Root growth directly affects the nutritional status and productivity of the aboveground biomass, making it a critical physiological parameter in plant development. The data presented in Figure 3 indicate that with prolonged water recession, plant root biomass increases. Soil carbon sequestration showed a positive correlation with both aboveground and root biomass, suggesting that the carbon fixation capacity of G, P, and T may improve with extended water recession periods, although this effect is likely to plateau once a certain threshold is reached.

3.3. Changes in Organic Carbon Content in the Topsoil (0–20 cm) of Dominant Plant Communities Across Varying Water Recession Durations

3.3.1. Variation in Soil Physicochemical Parameters

The physicochemical characteristics of G soil across varying water recession durations are presented in Figure 4a. The TN content in G is significantly higher than that in the other three species. In plot B, the concentrations of ammonium nitrogen (AN) and TN are the highest, whereas in plot D, these concentrations are the lowest, with statistically significant differences (p < 0.05). The total phosphorus (TP) content was highest in plot B, while the available phosphorus (AP) content showed a declining trend over the years. As shown in Figure 4b, the concentrations of TN and AN in the P wetland soil increased annually. The maximum TN concentration reached 8077 mg·kg⁻¹, which was 273.07% and 64.14% higher than in the other two plots, with statistically significant differences (p < 0.05). The maximum AN concentration reached 783.8 mg·kg⁻¹, 282.3% and 259.49% higher than in the other two plots, with statistically significant differences (p < 0.05). TP content followed the order: plot B < plot F < plot A, with significant differences (p < 0.05). The AP content in plot B peaked at 107.7 mg·kg⁻¹, which was 45.76% and 65.8% higher than in the other two plots, with statistically significant differences (p < 0.05). In the soil of the T wetland (Figure 4c), the concentrations of TN, AN, and AP showed an annual increase. The maximum TN concentration reached 2765.90 mg·kg⁻¹, which was 73.98% and 39.11% higher than in the other two plots, with statistically significant differences (p < 0.05). The maximum AN concentration was 203.86 mg·kg⁻¹, which was 25.18% and 12.99% higher than in the other two plots, with statistically significant differences (p < 0.05). The maximum AP concentration reached 53.67 mg·kg⁻¹, which was 46.86% and 33.71% higher than in the other two plots, with significant differences compared to plot C (p < 0.05), but no significant difference compared to plot D. The highest TP content was observed in plot D, which was 20.02% and 13.61% higher than in the other two plots, with statistically significant differences (p < 0.05). With the extension of water recession duration, the concentrations of TP (p < 0.05), AP (p < 0.001), and AN (p < 0.05) in T soil stabilized, while the TN concentration (p < 0.001) showed a consistent annual increase. Significant differences in the soil physicochemical properties were observed across different vegetation types. The correlation analysis between basic soil physicochemical properties and SOC is presented in Figure 3. The surface SOC of G showed a significant positive correlation with AN (p < 0.001) and TP (p < 0.01). The surface SOC concentration of P showed a significant positive correlation with moisture content (MC) (p < 0.001), TN (p < 0.001), and AN (p < 0.001). Additionally, the SOC concentration was significantly positively correlated with TN (p < 0.001), AP (p < 0.001), and AN (p < 0.001).

3.3.2. Dynamic Changes in the Organic Carbon Components of Surface Soil in Wetland Plants

The temporal variation in SOC content in the 0–20 cm surface soil of G wetlands under different water recession durations shows significant differences (Figure 5a). The SOC content is higher in plots B and F. In wetlands with a water recession duration exceeding 40 years (plot F), the DOC concentration is the highest, reaching 543.75 mg·kg⁻¹, significantly surpassing the levels observed in the other five wetland soils (p < 0.05). The DOC concentration in the 0-20 cm surface soil of G ranges from 246.81 mg·kg⁻¹ to 543.75 mg·kg⁻¹. The concentrations of LFOC and SOC in plot B are significantly greater than those in the other wetland soils (p < 0.05), with values of 351.45 g·kg⁻¹ and 275.8 g·kg⁻¹, respectively. For example, in the wetland soil of plot F, which has 53 years of water recession, the concentrations of HFOC, LFOC, DOC, and SOC are 68.38, 267.39, 656.5, and 139.22 g·kg⁻¹, respectively, which are significantly higher than those in the other two wetland soils (p < 0.05). The soil of T in plot E exhibits higher concentrations of organic carbon components. As shown in Figure 4c, the wetland soil in plot E exhibits HFOC, LFOC, and SOC concentrations of 224.06, 35.24, and 38.78 g·kg⁻¹, respectively, significantly surpassing those in the other two wetland soils, with statistically significant differences (p < 0.05). In plot D, the DOC concentration reaches 267.26 g·kg⁻¹, the highest among the three plots, which is 2.01% greater than in plot E, though no statistically significant difference was observed (p > 0.05), and 18.35% higher than in plot C, with no statistically significant difference (p > 0.05).

3.3.3. Temporal Dynamics of Microbial Biomass Variations

The contents of MBC, MBN, and MBP in G soil exhibit a progressively decreasing trend (p < 0.01, Figure 6a). Plot A shows the highest microbial biomass, with statistically significant differences (p < 0.05). In all plots, the MBC concentration is significantly higher than both MBP and MBN concentrations. In the P soil (Figure 6b), plot E exhibits the highest microbial biomass, with statistically significant differences (p < 0.05). The MBC concentration is 1489.27% and 3392.57% higher than in the other two plots, the MBN concentration is 1140.15% and 3160.91% higher, and the MBP concentration is 969.08% and 4764.41% higher. In the T soil (Figure 6c), the MBC concentration exhibits a progressively increasing trend (p < 0.001). Plot F displays the highest microbial biomass, which is 57% and 50% higher than in plots C and D, respectively, with statistically significant differences (p < 0.05). In all plots, the MBC concentration is significantly higher than both the MBP and MBN concentrations, with the latter two tending to stabilize.

Previous studies have shown a strong correlation between moisture content (MC) and MBC. The MBC in the surface soils of G, P, and T varies with increases in MC, indicating that a sustained increase in soil MC may lead to a decline in soil fertility. Microbial activity is strongly influenced by soil moisture, as enzymes involved in microbial processes are produced by microorganisms, and the transport of organic matter within the soil requires the liquid phase. In this study, we also observed a significant positive correlation between MBC and MC in the surface soils of the three plant species (p < 0.01).

3.3.4. Variation in the Concentrations of Total Iron (TFe) and Available Iron (Afe) in the Soil

The total iron (Tfe) content in G soil under different water recession durations exhibits a progressively increasing trend (p < 0.05, Figure 7a), with a decline observed in plot F. Plots D and E show the highest Tfe concentrations, while plot A has the lowest, which is 67.53% lower than in plot D, with statistically significant differences (p < 0.05). The Afe concentration in each plot is considerably lower than the Tfe concentration, with the highest Afe concentration in plot F and the lowest in plot A, which is 42.29% lower than in plot F, though the variation is relatively small. The Tfe content in the P soil exhibits a progressively decreasing trend (p < 0.001, Figure 7b), with a peak value of 33,537.25 mg·kg⁻¹, 107.51% and 83.27% higher than in the other two plots, showing statistically significant differences (p < 0.05). The Afe concentration decreases progressively, with the highest concentration reaching 854.17 mg·kg⁻¹, 90.07% and 312.4% higher than in the other two plots, with statistically significant differences (p < 0.05). In the T soil, the Tfe concentration shows a progressively decreasing trend (p < 0.05, Figure 7c), with a consistent annual decline in Tfe content, which is statistically significant (p < 0.001). The peak concentration of Tfe is 38,754.69 mg·kg⁻¹, which is 8.2% higher than in plot D, but no statistically significant difference is observed (p > 0.05), while it is 70.29% higher than in plot F, with statistically significant differences (p < 0.05). The Afe concentration decreases progressively, with a peak value of 677.69 mg·kg⁻¹, which is 0.3% higher than in plot C, but no statistically significant difference is observed (p > 0.05), and 19.39% higher than in plot F, showing statistically significant differences (p < 0.05).

3.4. Correlation Analysis of Soil Parameters for the Three Plant Species During the Water Recession Process

Figure 8a illustrates the Pearson correlation coefficients among various indicators of G. It presents the Pearson correlation coefficients between microbial biomass, plant biomass, soil carbon fractions, physicochemical properties, and 25 soil iron parameters.

3.5. Quantitative Analysis of the Soil Organic Carbon Fractions Associated with the Three Plant Species During the Water Recession Process

PLS-SEM was developed for G (Figure 9a). In PLS-SEM, direct effects are represented by the path coefficients (β) that connect two variables. This coefficient, analogous to the standardized regression coefficient, indicates both the strength and the direction of the relationship between the predictor and response variables. Figure 8 illustrates that soil physicochemical properties (β = −0.899, p < 0.001) exert a significant negative impact on soil iron content, and soil physicochemical properties (β = 0.882, p < 0.001) exert a significant positive effect on soil organic carbon fractions. The influence of soil physicochemical properties on plant biomass is not significant. Furthermore, water recession quality (β = −0.986, p < 0.001) has a significant positive effect on plant microbial biomass, while soil iron content negatively affects microbial biomass (β = −0.979, p < 0.001). Microbial biomass (β = 0.699, p < 0.05) significantly positively influences soil organic carbon fractions. Plant biomass and soil iron content do not exert a significant influence on soil organic carbon fractions.

PLS-SEM was developed for P (Figure 9b). Water recession quality (β = 0.717, p < 0.001) significantly positively influences soil iron content, water recession quality (β = 0.382, p < 0.05) has a significant positive effect on microbial biomass, and water recession quality (β = 0.650, p < 0.01) significantly positively affects soil physicochemical properties. Soil physicochemical properties (β = 0.69, p < 0.001) significantly positively affect microbial biomass, and soil physicochemical properties (β = 0.556, p < 0.01) significantly positively influence soil organic carbon fractions. Soil physicochemical properties (β = 0.960, p < 0.001) significantly positively impact plant biomass, microbial biomass (β = 0.297, p < 0.05) significantly positively affects soil organic carbon, while soil iron content (β = −0.268, p < 0.01) significantly negatively affects soil organic carbon. Soil physicochemical properties (β = −1.267, p < 0.001) have a significant negative effect on soil iron content; plant biomass does not significantly affect soil organic carbon fractions.

PLS-SEM was constructed for bulrush (Figure 9c). Water recession quality (β = 0.891, p < 0.001) exerts a significant positive impact on soil physicochemical properties, soil physicochemical properties (β = −1.315, p < 0.001) significantly positively influence soil iron content, soil physicochemical properties (β = 1.035, p < 0.01) significantly positively affect soil organic carbon fractions, and soil physicochemical properties (β = 0.749, p < 0.05) significantly positively influence plant biomass. Microbial biomass (β = 0.437, p < 0.05) significantly positively influences soil organic carbon. Soil iron content (β = −1.072, p < 0.001) significantly negatively affects microbial biomass. Neither plant biomass nor soil iron content significantly influences soil organic carbon fractions.

4. Discussion

4.1. Response of Plant Biomass Characteristics to Rice Paddy Water Withdrawal

Changes in water levels in wetlands during the water recession process are primarily influenced by the rate of water withdrawal and the total volume of water removed. When the volume of water withdrawn is large, the wetland water level decreases rapidly, potentially leading to a reduction in wetland area, a decline in aquatic vegetation, and, subsequently, affecting the survival and reproduction of aquatic fauna. On the other hand, if the water withdrawal volume is small, the wetland may have more time to adjust to the decreasing water level, which could result in a less-pronounced ecological impact. The total volume of water withdrawal is also a critical factor in determining the extent of water level changes in wetlands. A larger volume of water withdrawn can lead to a more significant decrease in the water level, with more substantial ecological consequences. In contrast, a smaller water withdrawal volume may result in a smaller reduction in water level, causing a relatively minor impact on the ecosystem.

Vegetation is highly sensitive to environmental variations [25]. Variations in plant biomass can influence the release of root exudates and the accumulation of organic carbon. Root exudates modulate soil microbial activity by altering the carbon input into the soil, which, in turn, affects the content of soil organic carbon [26]. In this study, the biomass growth trends of the various plant species were generally similar. The duration of water recession had a significant impact on plant biomass (p < 0.05), with biomass peaking at the longest water recession period. According to (Figure 2), the morphological indicators of roots, stems, and leaves of the three plant species growing in wetlands with longer water recession durations were significantly superior to those of G in wetlands with shorter recession durations. This indicates that the accumulation of nitrogen and phosphorus, with increased water recession time, significantly influences plant biomass, enhancing the growth of aboveground biomass. However, this growth is also constrained by the availability of bioavailable phosphorus in the soil [27]. Under nitrogen enrichment conditions, an increased amount of phosphorus is necessary to support accelerated plant growth [28]. Long-term nitrogen fertilization may further stimulate plant growth. Nitrogen enrichment, by reducing the soil carbon-to-nitrogen (C/N) ratio, accelerates litter decomposition [29]. Extended phosphorus fertilization has resulted in the significant accumulation of bioavailable phosphorus in the soil, which could, in turn, modify the aboveground plant community composition, leading to changes in the profile of root exudates [30].

In plot F, the aboveground biomass of G exhibited a decreasing trend, suggesting that with prolonged water recession and upon reaching a certain threshold, plant growth may be inhibited. However, research indicates that plant growth is more likely constrained by the availability of bioavailable phosphorus in the soil [31]. Research on nitrogen–phosphorus interactions commonly shows that plant growth is co-limited by both nitrogen and phosphorus. However, the extent and direction of this co-limitation can vary, leading to different outcomes. These studies primarily focus on the response of aboveground biomass, and the results of co-limitation are influenced by the nature of the nitrogen–phosphorus interaction [32]. Different ecosystems exhibit varying levels of nitrogen and phosphorus limitation, which results in differential responses to nitrogen–phosphorus interactions. The effects of nitrogen and phosphorus accumulation on plant biomass are influenced not only by nutrient supplementation, but also by factors such as ecosystem type, soil conditions, and plant species [33]. In addition, the growth of plant roots plays a critical role in determining the nutrient status and yield of the aboveground biomass, and is considered one of the primary physiological indicators of plant growth [34]. Plant roots serve as the primary source of organic carbon in soil, with root biomass acting as a key indicator of soil carbon input [35]. The measurements in (Figure 2) show that as the duration of water withdrawal increases, plant root biomass increases. Soil carbon sequestration exhibits a positive correlation with both aboveground plant biomass and root biomass [36]. This indicates that the carbon sequestration potential of G, P, and T may enhance as the duration of water withdrawal increases, albeit with a defined limit. Extensive research has shown that phosphorus addition alters species composition and enhances plant nitrogen uptake, thereby increasing the primary productivity of certain plant species and boosting belowground biomass [37]. The accumulation of nitrogen and phosphorus positively influences plant biomass, stimulating plant growth, especially in ecosystems where nitrogen and phosphorus are co-limiting factors [38]. Nevertheless, this enhancement of plant growth is limited by the bioavailability of phosphorus in the soil, with ecosystem-specific variations in the response to nitrogen–phosphorus accumulation, likely due to differences in nutrient limitation and ecological context [39]. Therefore, when developing agricultural management and ecological restoration strategies, it is essential to consider the combined effects of these factors.

4.2. Response of Soil Properties Characteristics to Rice Paddy Water Withdrawal

The findings indicate that the soil moisture content exhibited the greatest variation in G, with an increasing trend observed in the soils of P and T. This variation may be primarily due to the combined effects of precipitation and groundwater dynamics in the Xingkai Lake wetland ecosystem [40]. Conversely, the accumulation of organic matter in the surface layer enhances the soil’s water retention and absorption capacity, leading to an increase in soil moisture content [41]. The presence of water can notably modify the physical characteristics of the soil, thereby influencing its water-holding capacity and sustainability [42]. Plant litter is a principal contributor to soil organic matter in natural ecosystems, supplying essential nutrients that improve soil fertility and promote its overall nutrient cycling [43]. The study findings demonstrate that the addition of nitrogen and phosphorus primarily exerts a stimulatory effect on the total nitrogen and total phosphorus concentrations in the soil. The soil total nitrogen content in the three plant species is the highest. This can be attributed, in part, to the significant presence of humus in natural wetland soils, which undergoes ammonification to release substantial amounts of nitrogen [44]. Moreover, the drainage from rice paddies into the wetland, coupled with the year-on-year accumulation of organic fertilizers, results in a gradual increase in nitrogen and phosphorus concentrations in the soil. Nitrogen addition has been shown to decrease the decomposition rate of soil organic carbon [45], augment plant-derived carbon inputs [46], and consequently increase the total nitrogen content in the soil [47]. The input of nitrogen can stimulate the accumulation of microbial biomass, which in turn indirectly increases the levels of soil organic carbon and total nitrogen [44]. Additionally, nitrogen addition enhances the soil’s active phosphorus fraction by modifying soil pH and boosting microbial biomass, which in turn facilitates increased phosphorus uptake by plants [48]. As a result, this process leads to a reduction in the concentration of available phosphorus in the soil [49]. Furthermore, the addition of phosphorus can modulate the soil microbial community composition and enzymatic activity, potentially influencing nutrient cycling and soil fertility [50], which in turn can further affect the cycling process of total phosphorus in the soil, potentially altering nutrient availability and ecosystem functioning [51]. In forest ecosystems, the addition of phosphorus has been shown to boost vegetation productivity, thereby facilitating primary production and influencing nutrient dynamics [52], and enhance the carbon input from vegetation [53]; consequently, this process elevates the total nitrogen concentration in the soil [51].

4.3. Effects of Plant Biomass and Soil Properties on Soil Organic Carbon Component Changes

SOC exerts a significant influence on the functioning of terrestrial ecosystems and serves a pivotal role in the global carbon cycle [54]. Simultaneously, the composition of soil organic carbon holds significant ecological relevance and is potentially susceptible to environmental variations [55]. In this study, we observed that the dynamics of DOC content in G soil mirrored those of SOC. The DOC content in P soil exhibited an upward trend over the years, whereas no significant variation was observed in the DOC levels of T soil. DOC represents the fraction of soil organic carbon that is highly mobile and readily bioavailable, distinguished by its solubility, ease of leaching and decomposition, and elevated biological reactivity [56]. Despite comprising only a minor fraction of soil organic carbon, DOC serves as a readily accessible organic carbon source for soil biota and plants, playing a critical role in the soil carbon sequestration process. DOC, as the water-soluble fraction of soil organic carbon, is particularly sensitive to flooding. Prior research has demonstrated that variations in precipitation can significantly influence DOC concentrations, as rainfall facilitates the leaching of surface litter, root exudates, and other soluble organic compounds, driving their downward translocation under inundation. Furthermore, in water-saturated soils, DOC is more prone to being mobilized in aerobic conditions [57]. Our findings reveal that the DOC content in the surface soil of G initially increased, subsequently decreased, and then increased once more. In P surface soil, DOC content exhibited a continuous upward trend, while no significant changes were observed in T surface soil. The sampling occurred in August, coinciding with rising temperatures, which likely enhanced soil microbial activity, resulting in accelerated DOC degradation. Moreover, August represents the peak of the rainy season in the Xingkai Lake region, where increased precipitation releases DOC from the soil matrix. However, intense or prolonged rainfall may facilitate the leaching of soluble organic compounds into deeper soil horizons or their removal via runoff, thereby affecting DOC concentrations in the surface soil [58]. This season represents a phase of peak vegetative growth, characterized by substantial litterfall, which subsequently accelerates the decomposition rate under warm and humid conditions [59]. This process favors the enrichment of DOC within the soil. In comparison to G, both T and P exhibit significantly greater root biomass. Consequently, their abundant root exudates and litterfall contribute to the sustainability of soil DOC levels [60]. LFOC represents a low-density fraction predominantly consisting of undecomposed or partially decomposed plant and animal residues [61]. Variations in LFOC are likely driven by changes in plant biomass. In wetland ecosystems, fluctuations in water table levels significantly influence LFOC concentrations [62]; as water levels rise, wetland soils are characterized by elevated LFOC content and proportion, along with a marked increase in aboveground biomass [63].

MBC plays a direct role in the biogeochemical cycling of carbon and nitrogen within ecosystems [64]; it is intrinsically linked to ecosystem productivity. Climate conditions and land cover types are key determinants of the temporal dynamics of MBC reservoirs [65]. In our study, MBC concentrations across all wetland sites remained elevated during the early stages of plant growth in the typical growing season. This phenomenon is likely attributed to rising temperatures and moderate precipitation during the typical growing season, which accelerate the decomposition of dead roots and plant litter. These organic inputs supply an abundant source of substrates and energy, supporting soil microbial activity [66]. Moreover, the ample hydrothermal conditions observed in August intensified microbial metabolic processes, increased turnover rates, and led to substantial energy consumption, ultimately hindering the accumulation of MBC [67]. Hence, compared to the typical growing season, MBC content in G wetlands significantly declined during the rainy season (Figure 5), whereas the MBC levels in P and T soils exhibited an increasing trend. This divergence may be attributed to the water recession environment; given that inundation conditions impose constraints on microbial activity [68], the recession water level for Sporobolus pyramidalis (approximately 30–50 cm) exceeds that of P (approximately 0–10 cm) and T (approximately 20–40 cm). Under moderate flooding conditions, the soil microbial activity remains elevated, indicating a temporal adaptation mechanism of microbial communities to inundation. However, as the water level rises, resulting in a standing water environment, microbial activity becomes increasingly suppressed [69]. The relatively high biomass of P and T, coupled with its annual increase, results in substantial plant litter inputs, thereby elevating the organic matter content in surface soils. This abundant organic matter serves as a substrate for microbial decomposition, leading to the production of gases like carbon dioxide and methane, and subsequently enhancing the organic carbon content in the soil [70]. Soil moisture significantly influences SOC levels. Under optimal moisture conditions, microbial activity is elevated, leading to an accelerated decomposition of organic matter, which in turn enhances SOC accumulation [71]. Excessive soil temperatures or excessive moisture levels can suppress microbial activity, thereby decelerating the decomposition rate of organic matter, which ultimately results in a decline in SOC concentrations [72]. Relevant research indicates a strong positive correlation between soil moisture content (MC) and MBC [72,73,74]. Variations in MBC in the surface soils of Sporobolus pyramidalis, P, and T with increasing moisture content (MC) suggest that prolonged elevation in soil moisture may lead to a decline in soil fertility. Microbial activity is highly reliant on soil moisture, given that key enzymes are synthesized by microorganisms, and the transport of organic matter in the soil relies on the presence of a liquid phase. In our study, we also observed a significant positive correlation (p < 0.01) between MBC and MC in the surface soils of the three plant species.

Plant biomass serves as a mechanism for carbon sequestration within plant tissues, encompassing both aboveground and belowground biomass. The carbon stored in these plant components forms an integral part of the forest biomass carbon pool [75]. Belowground plant biomass acts as a primary regulator of plant-origin carbon fluxes [76]. Specifically, plants assimilate atmospheric carbon dioxide into organic matter via photosynthesis. The carbon embedded in these organic compounds constitutes a significant portion of plant biomass, thereby contributing to the establishment of a carbon reservoir [77]. Our findings indicate that the biomass of all three plant species exhibited a consistent upward trend, and this rise in plant biomass consequently contributes to the expansion of the carbon pool [78]. Furthermore, the influence of plant diversity on soil carbon stocks is intrinsically linked to plant biomass [79]. Studies indicate that plant diversity modulates plant biomass, thereby affecting the input of organic matter into the soil and, ultimately, determining soil carbon stocks [80]. This implies that by augmenting plant biomass, the soil carbon pool can be indirectly enhanced, contributing to overall carbon sequestration. Soil properties influence plant growth and vegetation structure, subsequently impacting the composition of soil organic carbon [81].

5. Conclusions

This study demonstrates that the dynamic variations in soil organic carbon components, such as HFOC, LFOC, and DOC, are strongly linked to the duration of water withdrawal in wetland soils. Wetlands with longer water withdrawal durations show a steady increase in SOC components in species like G, P, and T. In contrast, the SOC components in G wetlands exhibit a “first increase, then decrease” pattern, highlighting the significant influence of hydrological conditions and vegetation types on SOC dynamics. The study further reveals that the changes in SOC content, consistent with shifts in DOC, LFOC, and HFOC, play a key role in the wetland soil carbon pool’s dynamics. MBC also significantly impacts soil carbon stability and transformation. Additionally, N and P accumulation in the soil alters microbial activity and community structure, indirectly regulating the soil carbon pool’s processes. The study highlights that soil physicochemical properties—especially TN, TP, AP, and moisture content—affect the distribution of SOC components. The accumulation of nitrogen and phosphorus enhances plant biomass and SOC storage, while soil moisture content influences microbial activity, further affecting SOC cycling. MBC, particularly during water withdrawal, reflects the transformation and stability of soil carbon. Long-term monitoring reveals the critical role of water withdrawal duration, vegetation types, soil properties, and microbial activity in wetland carbon storage and cycling. The results show that water withdrawal significantly influences SOC storage by modifying soil’s hydrological and physicochemical properties. Given the combined effects of soil moisture and nutrient concentrations, wetland carbon sequestration capacity and microbial community structure undergo dynamic changes. This study underscores the importance of considering the integrated effects of hydrological conditions, soil nutrients, and microbial activity in wetland restoration and carbon sequestration management, offering a scientific foundation for future carbon sequestration strategies.