Impact of Coastal Beach Reclamation on Seasonal Greenhouse Gas Emissions: A Study of Diversified Saline–Alkaline Land Use Patterns

Abstract

Reclaiming coastal wetlands for agricultural purposes has led to intensified farming activities, which are anticipated to affect greenhouse gas (GHG) flux processes within coastal wetland ecosystems. However, how greenhouse gas exchanges respond to variations in agricultural reclamation activities across different years remains uncertain. To address this knowledge gap, this study characterized dynamic exchanges within the soil–plant–atmosphere continuum by employing continuous monitoring across four representative coastal wetland soil–vegetation systems in Jiangsu, China. The results show the carbon dioxide (CO₂) and nitrous oxide (N₂O) flux exchanges between the system and the atmosphere and soil–vegetation carbon pools, which revealed the drivers of carbon dynamics in the coastal wetland system. The four study sites, converted from coastal wetlands to agricultural lands at different times (years), generally act as CO₂ sinks and N₂O sources. Higher levels of CO₂ sequestration occur as the age of reclamation rises. In terms of time scale, crops lands were found to be CO₂ sinks during the growing period but became CO₂ sources during the crop fallow period. Although the temporal trend of the N₂O flux was generally smooth, reclaimed farmlands acted as net sources of N₂O, particularly during the crop-growing period. The RDA and PLS-PM models illustrate that soil salinity, acidity, and hydrothermal conditions were the key drivers affecting the magnitude of the GHG flux exchanges under reclamation. This study demonstrates that GHG emissions from reclaimed wetlands can be effectively regulated through science-based land management, calling for prioritized attention to post-development practices rather than blanket restrictions on coastal exploitation.

1. Introduction

Greenhouse gas emissions are recognized internationally as the main cause of global warming [1]. Carbon dioxide (CO₂) and nitrous oxide (N₂O) are two important atmospheric greenhouse gases (GHGs) [2]. Among greenhouse gases, CO₂ maintains the highest atmospheric concentration and remains the primary focus of mitigation efforts. Concurrently, N₂O has garnered significant scientific attention due to its 100-year global warming potential (GWP100) exceeding 265 times that of CO₂ [3,4]. Amidst the impacts of global climate change, research on ecosystem GHG sources and sinks has attracted much attention [5,6]. Against the backdrop of the ecosystem carbon cycle being a focal point in global change research, disturbance-driven carbon exchange flux variability has drawn increasing attention, particularly in wetlands with heterogeneous disturbance legacies [7,8,9,10].

Human reclamation activities in wetlands have seriously disturbed ecosystems, which may lead to saline–alkali changes, species invasions, and the destruction of native vegetation [11,12]. The reclamation of coastal wetlands for urban expansion, industrial development, and agricultural production activities is common worldwide [13]. Ecosystem changes have occurred from natural wetlands to reclamation wetlands. It is therefore reasonable to assume that these unique coastal reclamation wetland settings are influenced by both natural siltation and human activities. The interplay between climatic variability and anthropogenic disturbances may exert nonlinear effects on the functioning of ecosystems in anthropogenically modified wetlands, potentially driving ecological systems beyond critical thresholds [14]. These influences contribute to local environmental changes and directly or indirectly result in greenhouse gas emissions [15,16]. Newly reclaimed coastal croplands represent a critical land resource, with converted wetlands used for continuous agricultural production. These systems exhibit advantages as strategic reserves with high developmental potential [17,18]. GHG emissions from aquaculture ponds converted from coastal wetlands have been well studied [4,19,20]. However, there is still a lack of quantitative studies on the GHG emissions from agricultural land use in coastal wetland reclamation areas under different land use patterns. Moreover, emerging challenges in newly reclaimed croplands—particularly low soil maturity, poor structural stability, and constrained productivity—impede agricultural development while also disrupting greenhouse gas emission regimes. There is a lack of monitoring and mechanism analyses of intra-annual GHG exchange for different reclamation years [21,22,23]. To reduce carbon emissions from coastal wetlands experiencing increasing human disturbances, it is necessary to quantify the possible responses of CO₂ fluxes in the reclamation of coastal wetlands for agriculture and explore the underlying impact mechanisms [24,25,26,27].

The Jiangsu coastal zone, representing a key region along China’s southeastern coast, is characterized by its rich silty beaches. Owing to its unique geographic environment, it has evolved into an important coastal wetland ecosystem on a global scale [28]. Under the combined pressures of climate change and anthropogenic disturbances, salt marshes along the Jiangsu coast have experienced persistent modifications [29]. Rapid urbanization in the economically powerful middle and lower Yangtze River region has intensified human activities in Jiangsu’s coastal wetlands, particularly reclamation [30]. Notably, China’s coastline is advancing at 25–30 cm/year, with most of the newly formed land reclaimed for agricultural development [31]. The coastal wetlands in Jiangsu exhibit a long history of reclamation, with an estimated total reclaimed area of approximately 107,420.0 hm². Over the past six decades, intensified reclamation has led to a >70% reduction in the natural wetland area along Jiangsu’s coast, coupled with severe degradation of native saltmarsh vegetation. This has resulted in spatially heterogeneous land use patterns shaped by varying reclamation chronosequences [32]. Following soil amelioration measures, agricultural cultivation—including crops such as wheat, rice, and maize—was initiated [33]. The conversion of coastal wetlands to arable land under saline–alkali conditions significantly alters the spatiotemporal dynamics of salt ion transport in surface soils across reclamation chronosequences. Notably, surface soil salinity decreases at an accelerated rate following prolonged reclamation [34]. Reclamation age gradients drive spatiotemporal differentiation in soil salt ion redistribution via ‘leaching-downward’ transport mechanisms. Newly reclaimed areas (<20 years) exhibit surface salt accumulation that constrains SOC sequestration, whereas long-term reclaimed sites (>40 years) demonstrate enhanced carbon–nitrogen turnover under amelioration practices [35]. Critically, soil salinization modulates multiple ecosystem parameters—including plant biomass accumulation, soil organic carbon (SOC) stocks, net photosynthetic rates, and CO₂ flux patterns [36]. Moreover, implementing systematic farmland management and sustained amelioration measures progressively enhances salinity reduction efficiency over extended reclamation periods [37]. Consequently, the carbon cycling mechanisms of the reclaimed agricultural systems evolve with both reclamation duration and associated physicochemical environmental shifts. Therefore, it is crucial to investigate specific environmental factors and pathways that influence GHG flux and balance under varying durations of reclamation.

Following reclamation, soil salinity exhibits a predictable temporal migration pattern. Initial hypersaline conditions strongly suppress vegetation establishment, whereas long-term anthropogenic amelioration reduces topsoil salt ion concentrations, forming a distinct vertical ‘top-down’ desalination gradient [38]. These salinity dynamics critically regulate soil–vegetation carbon interactions, yet systematic investigations into salinity–GHG coupling mechanisms across reclamation chronosequences remain limited. Notably, salinity exerts opposing controls on CO₂ and N₂O emissions—elevated salt content promotes CO₂ efflux via organic matter destabilization while concurrently inhibiting nitrification to reduce N₂O production. Agricultural interventions may disrupt this natural equilibrium, necessitating a quantitative assessment of how reclamation duration modulates agroecosystem carbon cycling. This study primarily utilized the reclaimed wetlands for dry agriculture, focusing on two greenhouse gases, CO₂ and N₂O. The research involved observations of GHG flux exchanges in four mudflat coastal wetland systems with 65, 49, 41, and 16 years of reclamation along the southeastern coast of China, to (I) quantitatively compare seasonal CO₂ and N₂O source–sink balances across reclamation chronosequences, establishing the GHG status of agroecosystems in reclaimed coastal wetlands; (II) identify the key environmental factors affecting CO₂ and N₂O balances under temporal changes and crop rotation systems; and (III) analyze the direct and indirect pathways through which environmental factors mediate greenhouse gas fluxes. This investigation addresses critical knowledge gaps in GHG flux dynamics during land use transition from coastal wetlands to cultivated croplands, providing novel insights for carbon sequestration strategies in reclaimed coastal agricultural ecosystems.

2. Materials and Methods

2.1. Study Area and Sampling Sites

This study is located in Rudong County (32°12′ N–32°36′ N, 120°42′ E–121°22′ E) in southeastern Jiangsu Province, situated on the northern flank of the Yangtze River Delta at the confluence of the Yangtze River and the South Yellow Sea (Figure 1). The region experiences a northern subtropical monsoon humid climate with abundant light, simultaneous rain, and heat. The average annual temperature is 15.9 °C, while the average yearly precipitation is 1155.6 mm. As a typical coastal reclamation area, the site features saline soils with expanding reclaimed farmland due to increasing human activities.

This study sampled the reclaimed farmland in the coastal wetland of Rudong County, Jiangsu Coastal area, which has a history of reclamation. Following representativeness (covering the range of reclamation histories), typicality (characteristic of regional land use patterns), and consistency (stable land use over the past decade) principles in sample point selection, this study used four reclamation areas from different reclamation eras as research subjects. Each reclamation area contained three replicate sampling points (Figure 1). Several sites show stable land use patterns over the last decade (

Table 1. Information and description of coastal wetland reclamation plots.

Site	Reclamation Years	Land Use Planting System
WCSP	65	Wheat–corn–soybean rotational cropping arable land system
FCP	49	Fava bean–corn rotational cropping arable land system
WCP	41	Wheat–corn rotational cropping arable land system
WP	16	Wheat continuous cropping arable land system

). Each sampling point was stably utilized, minimizing confounding effects from recent land use changes.

In this study, a soil chronosequence spanning 65 years was identified. To investigate temporal patterns of ecosystem development following reclamation, we employed a space-for-time substitution (SFTS) approach [39]. This method is widely used in ecological studies when long-term monitoring data are unavailable, allowing researchers to infer temporal dynamics by examining sites of different ages that share similar environmental conditions and management histories [40]. The validity of this approach rests on three key assumptions: (1) the selected sites share similar initial conditions, (2) the environmental gradients are minimal compared to the age gradient, and (3) the sites have experienced similar management regimes. In our study, all selected sites had comparable geomorphological settings, soil parent materials, and regional agricultural management practices, differing primarily in their reclamation history.

2.2. Gas Flux Measurements

Greenhouse gas (CO₂ and N₂O) fluxes were measured at 19 time points spanning the 2022–2023 agricultural cycle, using static chamber–gas chromatography techniques. Sampling intervals were stratified seasonally, including monthly sampling, bimonthly (transition periods), and weekly (growing season) measurements. All sampling events occurred between 09:00 and 11:00 local time to minimize diurnal variation effects. The seasonal settings discussed in this paper include spring (March–May), summer (June–August), autumn (September–November), and winter (December–February). Gas samples were collected and analyzed in a static box using gas chromatography. Transparent polycarbonate chambers (70 cm height × 70 cm diameter) equipped with internal fans were deployed for 40-min sampling periods, with gas samples collected at 10-min intervals (0, 10, 20, 30, and 40 min) using evacuated vials. Gas concentrations were analyzed within 24 h using a gas chromatograph (Agilent 8860, Santa Clara, CA, USA) equipped with electron capture and flame ionization detectors.

In this study, gas fluxes were measured using static box–gas chromatography. The GHG flux was determined as the change in GHG concentration per unit time per unit area of the static box [41,42]. The gas flux (mg·m⁻²·h⁻¹) was calculated using the following formula:

$$\mathrm{F}=\mathsf{\rho }\mathrm{H}{\displaystyle \frac{\mathrm{d}\mathrm{C}}{\mathrm{d}\mathrm{t}}}{\displaystyle \frac{273}{273+\mathrm{T}}},$$

(1)

where F represents the gas emission flux (mg·m⁻²·h⁻¹), indicating the amount of greenhouse gases released per unit of time per unit of area. ρ represents the density of the greenhouse gas (g·L⁻¹), calculated by the ideal gas equation of state, where CO₂ and N₂O are used as standard molecular weights, respectively. dc/dt is the rate of change of gas concentration in the chamber (mL·m⁻³·h⁻¹), obtained by linearly fitting the slope to the gas concentration in the box. H is the height of the sampling chamber (m). T is the average air temperature inside the box at the time of sampling (°C), recorded by the temperature sensor inside the box. The value 273 is the constant of the gas equation. Five gas samples were collected in each sampling set, and a test was performed by fitting a straight line to the gas concentration with corresponding time intervals (0, 10, 20, 30, and 40 min) with an R² value greater than 0.87 (p < 0.05). The gas emission fluxes can be calculated by combining parameters such as atmospheric pressure, air temperature, universal gas constant, effective height of the sampling box, and molecular weight of target gases. Cumulative gas emissions, total cumulative gas emissions, and global warming potential (GWP) are derived from the fluxes of gas emissions [43,44]. Global Warming Potential (GWP) was calculated using the 100-year timeframe as per IPCC AR6 guidelines [45]. The equations for calculating these values are provided below:

$$\overline{\mathbf{F}}\mathbf{=}\sum _{\mathbf{i}\mathbf{=}\mathbf{1}}^{\mathbf{n}}{\displaystyle \frac{{\mathbf{F}}_{\mathbf{i}}\mathbf{\times }{\mathbf{d}}_{\mathbf{i}}}{\mathbf{d}}} \mathbf{G}\mathbf{=}{\displaystyle \frac{\overline{\mathbf{F}}\mathbf{\times }\mathbf{24}\mathbf{\times }\mathbf{d}}{\mathbf{100}}}$$

(2)

$$\mathrm{F}=\mathsf{\rho }\mathrm{H}{\displaystyle \frac{\mathrm{d}\mathrm{C}}{\mathrm{d}\mathrm{t}}}{\displaystyle \frac{273}{273+\mathrm{T}}},$$

(3)

$$\mathbf{G}\mathbf{W}\mathbf{P}\mathbf{=}{\mathbf{G}}_{{\mathbf{C}\mathbf{O}}_{\mathbf{2}}}\mathbf{\times }{\displaystyle \frac{\mathbf{44}}{\mathbf{12}}}\mathbf{+}{\mathbf{G}}_{{\mathbf{N}}_{\mathbf{2}}\mathbf{O}}\mathbf{\times }{\displaystyle \frac{\mathbf{44}}{\mathbf{28}}}\mathbf{\times }\mathbf{265}$$

(4)

where ${\mathrm{F}}_{\mathrm{i}}$ represents the gas emission flux at the ith sampling, with CO₂ measured in C and N₂O in N (mg·m⁻²·h⁻¹), ${\mathrm{d}}_{\mathrm{i}}$ is the number of days between the ith sample and the next sample, while $\mathrm{d}$ represents the total number of days. $\mathrm{G}$ denotes the total emissions, with CO₂ measured in C and N₂O in N (kg·ha⁻¹). ${\mathrm{C}}_{\mathrm{i}+1}$ refers to the cumulative emissions of gases between the ith and i+1 sampling period (kg·ha⁻¹).

2.3. Vegetation Sampling and Analysis

This study involved the collection of vegetation from different crops during various cropping periods in four reclamation areas. The collection occurred during the crop harvesting season, and both above-ground vegetation components and below-ground root systems were collected. Biomass was estimated at each sample plot using 0.5 × 0.5 cm sample boxes for vegetation. All vegetation within the sample box was cut to separate its roots and leaves, and whole vegetation carbon was determined using a Vario EL cube organic element analyzer (Elementar, Langen, Germany). Furthermore, vegetation samples were then cleaned and dried, and aboveground biomass was determined. A few samples of each part of the plant were ground, and the carbon content of each part of the plant was determined using a C/N elemental analyzer (Vario PYRO, Elementar, Langen, Germany).

2.4. Soil Sampling and Analysis

Soil samples were collected from six different depths (0~10, 10~20, 20~30, 30~40, 40~50, and 50~60 cm) using stainless steel corers. Subsequently, soil bulk weight was measured using ring knife method [46]. Some of the soil samples were brought back to the laboratory, air-dried, ground, and passed through a 2 mm sieve for soil physico-chemical property tests. Some of the pre-treated soil samples were used to determine soil carbon and nitrogen content using the combustion oxidation–elemental analyzer method (vario-macroCN Elementar, Germany). The leaching–spectrophotometer method determined NH⁴⁺-N and NO³⁻-N (GB/T 42485) [47]. TN was determined by the Kjeldahl method (GB/T 6432) [48]. A digital pH meter was used to determine the pH value of the soil leachate with a 1:5 soil–water ratio to determine the soil acidity and alkalinity (pH). A Mettler Toledo Experimental Conductivity Meter (FE30, Mettler Toledo Instruments, Shanghai, China) was used to determine the electrical conductivity (EC), and soil water content (SWC) was measured using the drying and weighing method. This study employed the conductivity method to assess water-soluble total salts in saline soils. The conversion of conductivity to soil salt mass was conducted using the relationship between soil salt mass fraction and conductivity. The formula for salinity content ($\mathrm{S}\mathrm{C}$) (g·kg⁻¹) is provided below:

$$\mathrm{S}\mathrm{C}={\displaystyle \frac{\mathrm{E}\mathrm{c}-0.0033}{0.28989}}$$

(5)

where $\mathrm{E}\mathrm{c}$ refers to the soil’s electrical conductivity.

2.5. Statistical Analyses

Statistical analyses were performed using SPSS 26 and R 4.3.3. One-way analysis of variance (ANOVA) was conducted to assess differences in gas, soil, and vegetation component indices across reclamation years, followed by Fisher’s Least Significant Difference (LSD) post hoc test for pairwise comparisons (α = 0.05). The repeated-measures ANOVA model evaluated fixed effects of season (within-subject) and site (between-subject) on N₂O fluxes. Redundancy analysis (RDA) was implemented using the vegan package in R to examine relationships between greenhouse gas (GHG) fluxes and environmental factors [49,50]. Partial Least Squares Path Modeling (PLS-PM) was performed via the plspm package to evaluate direct and indirect effects of environmental variables on GHG emissions [51,52]. Direct pathways show immediate variable impacts, while indirect pathways operate through mediators. Model fit was assessed using the Goodness-of-Fit (GOF) index, where GOF > 0.50 was considered acceptable for exploratory research and GOF ≥ 0.70 indicated strong fit. Data visualization was generated using OriginPro 2023 (v9.9.0.225), R software (v4.3.3) and the ggplot2 package (v3.4.4).

3. Results

3.1. Description of Soil and Vegetation Traits

Soil and vegetation’s physical and chemical properties in the reclaimed area during the study period are presented below (Figure 2 and Figure 3). For some obvious factors, soil water content (SWC) varied from 6.64% to 22.30%, while soil salinity (SC) ranged from 0.0841 g·kg⁻¹ to 0.542 g·kg⁻¹. The pH of the soil ranged from 5.7 to 6.68. Additionally, the soil organic carbon content (SOC) ranged from 0.67% to 1.37%, and the soil total nitrogen content (STN) was between 0.008% and 0.102%. Soil structure and nutrient composition improved with increasing years of reclamation. For instance, BD, SOC, and STN were found to be higher at WCSP compared to other samples (p < 0.05), although their pH was significantly lower than that of other samples (p < 0.001). As years of reclaiming increased, there was an observed increase in soil maturity. For instance, soil total nitrogen in WP (the sample area was reclaimed in 2007) was significantly less than (p < 0.05) in the previously reclaimed area (Figure 2C). Moreover, there was also a decrease in salinity levels with a longer history of reclamation. Samples in WP showed significantly higher SC than other samples (p < 0.001). Significant differences were observed between surface and subsurface layers regarding various soil properties. However, most physicochemical parameters of samples from different depths in the subsurface were not statistically different (p > 0.05). In conclusion, for saline-amended cropland, the surface soil of arable land under the effect of reclamation activities has a more pronounced response.

Comparing the total biomass densities of the different sites, it was found that WCP had the highest total biomass density, of 2472.2 to 3124.1 g/m² (maize’s total biomass density ranged from 919.6 to 1328.3 g/m² and wheat’s total biomass density ranged from 1517.6 to 1795.8 g/m²). WP had the lowest total biomass density, of 1328.92–1958.16 g/m² (total biomass density of wheat crop only). Figure 3 illustrates the biomass density and total carbon index at harvest time for various vegetation types across different sites. However, no significant difference was observed when comparing the total carbon content of wheat crop leaves, fruits, and roots of vegetation (p > 0.05) among WCSP, WCP, and WP. Combined with the biomass data, the average carbon fixations of the crops in the four reclamation sample plots during a year were as follows: 1024.88 g C/m² for WCSP, 722.54 g C/m² for FCP, 1139.79 g C/m² for WCP, and 652.89 g C/m² for WP.

3.2. Analyses of the Dynamic Seasonal-Scale Gas Carbon Fluxes and Greenhouse Gas Emissions

The gas flux variation at different sampling times of the coastal reclamation during the study period is illustrated in Figure 4. CO₂ fluxes ranged from −7394.62 mg·m⁻²·h⁻¹ (WCP sample site, August 2023) to 1878.58 mg·m⁻²·h⁻¹ (FCP sample site, September 2023) for all sample sites during the sampling period (Figure 4). This contrasts with positive values denoting net carbon release to the atmosphere. The difference in monthly CO₂ flux status was found to be extremely significant (p < 0.001). The difference in mean CO₂ fluxes between corn and the other crop growth periods was found to be highly significant (p < 0.001). Furthermore, it was observed that the CO₂ fluxes during the crop fallow period were significantly greater than the CO₂ fluxes during the growing periods of both wheat and maize (p < 0.001), as well as being significantly greater than the CO₂ emissions during the growing period of fava bean (p < 0.05). In spring and winter, the CO₂ flux during the fallow period of WP was higher (Figure 4d) at 86.05 mg·m⁻²·h⁻¹, while the CO₂ flux during the fallow period of FCP was higher in summer and autumn (Figure 4b), at 1878.58 mg·m⁻²·h⁻¹. Overall, there were significant seasonal differences in CO₂ fluxes between spring/summer and autumn/winter (p < 0.01). However, no significant difference was found between autumn and winter (p > 0.05). The trend of the CO₂ fluxes in each reclamation area generally exhibited fluctuating changes throughout the months studied. In terms of total CO₂ emissions, the average CO₂ emissions of WCSP, FCP, WCP, and WP were −51,273.44 kg·ha⁻¹, −94,856.25 kg·ha⁻¹, −36,557.534 kg·ha⁻¹, and −9696.77 kg·ha⁻¹, respectively (

Table 2. The total CO₂ and N₂O greenhouse gas emissions (kg·ha⁻¹), as well as the corresponding GWP for different plots; different letters indicate significant differences between reclamation sites at p < 0.05.

Plot
	WCSP	FCP	WCP	WP
CO2	−51,273.55 ± 4746.99 bc	−81,181.77 ± 18,715.87 b	−35,678.02 ± 1086.91 ac	−7446.77 ± 6497.05 a
N2O	1.674 ± 0.598 b	5.016 ± 0.862 a	7.329 ± 0.260 a	0.701 ± 1.032 b
GWP	−187,305 ± 17,598.45 cb	−295,577 ± 68,741.63 cb	−127,767.24 ± 4064.92 ab	−27,012.72 ± 33,412.39 a

). Negative CO₂ flux values indicate net carbon uptake by the soil–plant system (i.e., the ecosystem functioning as a carbon sink), in which photosynthetic fixation exceeds respiratory emissions. During the growing periods of different sites, there were significant differences (p < 0.05) in CO₂ emissions between the FCP and WP sample plots (

Table 3. Comparison of average monthly CO₂ and N₂O greenhouse gas emissions from different plots in the reclamation area during the growing season and the fallow period (kg·ha⁻¹); different letters indicate significant differences between reclamation sites at p < 0.05.

Greenhouse Gas Emission	Period	Plot
		WCSP	FCP	WCP	WP
CO2	Fallow period	180.74 ± 16.43 a	106.86 ± 35.13 ab	70.17 ± 83.72 b	79.00 ± 29.85 b
	Crop growth period	−424.19 ± 57.21 ab	−542.93 ± 94.27 b	−438.46 ± 97.40 ab	−257.63 ± 64.85 a
N2O	Fallow period	0.0013 ± 0.0057 b	0.0120 ± 0.0028 b	0.0491 ± 0.0201 a	−0.0015 ± 0.0019 b
	Crop growth period	0.0147 ± 0.0058 b	0.0283 ± 0.0075 c	0.0325 ± 0.0052 abc	0.0142 ± 0.0042 abc

). Additionally, significant differences (p < 0.05) in CO₂ emissions were observed among the WSCP, WCP, and WP sample plots when considering different fallow periods of the same land (Table 3).

The trend of the N₂O fluxes remained relatively consistent throughout the study period. The N₂O fluxes at all the sample sites varied from −0.001491 mg·m⁻²·h⁻¹ (October 2023 at FCP sample plot) to 0.18217 mg·m⁻²·h⁻¹ (June 2023 at WCSP sample plot) (Figure 5). The N₂O emission status differed significantly (p < 0.05) from month to month for all the sample sites. Still, there was no significant seasonal difference (p > 0.05) in the overall N₂O flux exchange. In terms of total N₂O emissions, the average N₂O emissions for WCSP, FCP, WCP, and WP were 1.674 kg·ha⁻¹, 5.016 kg·ha⁻¹, 7.329 kg·ha⁻¹, and 0.701 kg·ha⁻¹, respectively (Table 2). Marginal differences existed among the reclamation sites (F = 3.84, p = 0.057, η² = 0.59), with FCP exhibiting the highest average flux (0.054 ± 0.021 mg·m⁻²·h⁻¹) and WP the lowest (0.007 ± 0.025 mg·m⁻²·h⁻¹). The WCSP emitted 34% less during growth periods than FCP (p = 0.03). For growing season and fallow period comparisons, the flux decreased by 52% during fallow periods compared to growth seasons (F = 5.18, p = 0.052, η² = 0.40), with consistent reduction patterns across all the sites (season × site: F = 1.08, p = 0.41). In terms of the N₂O emissions during the growing period in different sample plots, there was a significant difference (p < 0.05) between the N₂O emissions of WCSP and FCP. Additionally, significant differences (p < 0.01) were observed in the N₂O flux exchanges between WCP and other sample plots during the fallow period for different sites (Table 3). With a significant overall effect (F = 2.52, p = 0.042, η² = 0.043) but small effect size, the N₂O fluxes were found to be significantly lower in soybeans than in wheat and corn during the growth period (p < 0.05).

3.3. Environmental Drivers of GHG Flux Exchange in Coastal Reclamation Areas

The RDA method was utilized to elucidate the impact of environmental factors on GHG flux exchange. The selected environmental variables were chosen based on their established ecological relevance to GHG fluxes in restored ecosystems. Soil temperature (ST) and soil water content (SWC) are fundamental regulators of biogeochemical processes, while soil organic carbon (SOC) and nitrogen availability (nitrate nitrogen, NN; soil total nitrogen, STN) serve as key factors in CO₂ and N₂O production. Soil salinity (SC) and pH reflect critical saline–alkali soil characteristics. Plant total carbon (PTC) and vegetation biomass density (PVB) represent vegetation carbon status. Atmospheric temperature (AT) and precipitation (AWP) were included as essential climate drivers. The multicollinearity among the predictors was assessed using variance inflation factors (VIFs), with all the variables exhibiting VIF values < 10, indicating acceptable collinearity for multivariate analysis. The results of the redundancy analysis (RDA) (Figure 6) revealed that the spatial and temporal variations of the CO₂ and N₂O flux exchanges within the year were primarily influenced by temperature, soil hydrothermal environment, and soil salinity environment, which collectively accounted for 80.78% of the variance (Figure 6b). This suggests that temperature, soil moisture, soil temperature, soil pH, and soil salinity play a significant role in explaining GHG fluxes in coastal reclaimed wetlands alone, making them key drivers influencing greenhouse gas exchange. Specifically, temperature significantly impacts CO₂ flux exchange, while soil nitrate nitrogen has a greater influence on N₂O exchange. Additionally, both GHG flux exchanges are affected by coastal beach soil salinity, as well as overall salinity interference (Figure 6a).

PLS-PM was employed to elucidate further the direct and indirect pathways by which soil properties, vegetation carbon pools, and external climatic conditions impact CO₂ and N₂O gas flux exchanges, as well as the greenhouse gas source and sink status in coastal wetland reclamation areas (Figure 7). The results indicated that soil temperature (ST) had a significant direct effect on CO₂ flux exchange (direct effect = 7.94, p < 0.05), as did the saline and alkaline environment (SPS) (direct effect = 9.381, p < 0.01). Conversely, soil bulk density (SBD) showed a non-significant direct effect on CO₂ flux exchange in general (direct effect = −0.907, p = 0.09), as did soil carbon pools (Soil C) (direct effect = 0.258, p > 0.05). In addition, soil temperature, closely linked to climatic conditions, and saline and alkaline environments in the coastal reclamation have a more pronounced impact on CO₂ fluxes (significant path: Climate–ST-CO₂ flux; SPS-CO₂ flux). Based on PLS-PM analyses, for the soil nitrogen environment, soil total nitrogen (STN, direct effect = 3.301, p < 0.05) and nitrate nitrogen (SNN, direct effect = 0.748, p < 0.05) showed a significant positive effect on N₂O flux exchange. The saline environment (SPS, direct effect = −4.34, p < 0.01) and the soil hydrothermal environment (SWT, direct effect = −0.785, p < 0.05) showed a significant negative effect on N₂O flux exchange. In contrast, the vegetation carbon pool (Plant C, direct effect = −0.119, p > 0.05) did not show overall significance for N₂O flux exchange. In addition, nitrogen and saline environments influenced by climatic conditions have a more sensitive effect on N₂O flux exchange (significant pathways: Climate–STN-N₂O flux, Climate–SPS-N₂O flux, SNN-N₂O flux).

4. Discussion

4.1. Responses of CO₂ and N₂O Emissions to Wetland Reclamation and Saline–Alkali Dynamics

In this study, observations of greenhouse gases in coastal wetlands in southeastern China were carried out, contributing to targeted carbon sequestration and emission reduction in coastal reclaimed wetlands. The CO₂ fluxes of the reclaimed wetlands were all in obvious seasonal patterns (Figure 4). That is, the coastal reclaimed wetlands clearly showed a CO₂ gas sink during the crop growing period, and a clear CO₂ source during the crop intercropping fallow period (Figure 4; Table 3). Vegetated drylands significantly reduced CO₂ concentrations through crop-mediated carbon sequestration [53]. Photosynthesis converted atmospheric CO₂ into biomass, while deeper root systems enhanced soil carbon storage, consistent with He’s observations in Yangtze River wetlands [22].

Regarding the CO₂ sequestration capacity in different seasons, it was found to be strongest during the summer rainy season, which is consistent with Wang et al.’s study in the coastal reclaimed wetland of Chongming, Shanghai. The peak CO₂ sequestration occurred during summer rains, aligning with Wang et al.’s findings in Chongming wetlands [19]. Soil respiration increases with rising temperatures within a certain range. This reflects temperature-dependent respiration dynamics [54,55,56,57]. Higher temperatures increase systemic respiration, thereby increasing CO₂ flux exchange [58,59]. Accordingly, we identified a pathway of positive influence of soil temperature on CO₂ flux (direct effect of 7.94; Figure 7). In this study, the CO₂ flux balance showed a significant positive correlation with soil water content (Figure 6 and Figure 7). This could be attributed to the substantial impact of soil water content on the seasonal dynamics of soil respiration [60]. It is widely acknowledged that soil moisture has a threshold effect on soil respiration. When the soil moisture is low, it becomes the limiting factor for soil respiration; however, beyond a certain threshold, excessively high moisture content reduces soil respiration rates by impeding CO₂ diffusion through the soil and CO₂ release. This may be due to low gas diffusion rates and reduced aerated porosity resulting from excess moisture, which limits respiration and leads to increased anaerobic conditions [61]. Critically, soil temperature and moisture are recognized as the primary physical factors controlling carbon mineralization in soils. They influence gas diffusion through soil pores by affecting microbial and root activity, thereby impacting CO₂ exchange fluxes. Soil moisture exhibited threshold effects on respiration [45,46]: low moisture limited microbial activity, while saturation impeded CO₂ diffusion. As key controllers of mineralization, temperature and moisture jointly regulate gas diffusion via microbial and root activity [62]. An interesting discovery from this research was the presence of spatial regularity in total annual CO₂ emissions. The total annual CO₂ emissions from the different polder areas ranged from the ocean to the inland (Figure 1; Table 3). There was a decreasing trend in the total CO₂ emissions from sea to land, which may be attributed to the increasing years of reclamation from sea to land. The influence mechanism can be considered as the soil salinity and the soil acid–base environment playing an irreplaceable role (Figure 6 and Figure 7). The coastal soils in Jiangsu are typical saline soils, formed mainly by the deposition of river sediment into the sea through tidal power and its subsequent elevation, becoming land. The main characteristic of these soils is their high salinity due to leaching from the sea. Additionally, there is a significant positive correlation between soil salinity and CO₂ flux (direct effect of 9.381; Figure 7), indicating that higher soil salinity content and pH lead to increased CO₂ emissions [63,64]. This may be attributed to several factors. Firstly, salinization impacts the availability of soil unstable organic carbon, which subsequently affects microbially mediated soil organic mineralization processes [65]. Secondly, salinization has the potential to disrupt soil aggregates. Thirdly, salinization can alter the structure of microbial communities, leading to changes in nutrient patterns and availability, ultimately influencing CO₂ emissions [66]. After the natural wetland has been reclaimed, the soil water salt content will gradually decrease due to natural precipitation leaching and anthropogenic cultivation improvement. Since WP is a sample site from the Yudong Reclamation Area, reclaimed in 2007, it differs from the other sample sites reclaimed earlier in the year, with higher soil salt content and mildly salinized soils. As a result, it showed a lower CO₂ sequestration capacity. The small difference in annual mean CO₂ fluxes between WCSP and WCP may be attributed to the proximity of the two sites and the minimal difference between the reclaimed croplands.

While the CO₂ dynamics showed strong seasonality, the N₂O emissions displayed distinct patterns. It was observed that during crop growth, the reclaimed agricultural land acted as a clear source of N₂O (Figure 5). This finding is consistent with previous studies. Reclaimed farmland consistently acted as an N₂O source during crop growth (Figure 5), confirming agricultural soils as major anthropogenic N₂O emitters [67]. The mean annual N₂O emissions in different reclamation areas varied spatially, as follows: 7.329 kgN·ha⁻¹ (WCP); 5.016 kgN·ha⁻¹ (FCP); 1.674 kgN·ha⁻¹ (WCSP) and 0.701 kgN·ha⁻¹ (WP), respectively. This finding is consistent with previous studies indicating that agricultural soils, heavily impacted by human agricultural activities, are the primary anthropogenic sources of N₂O emissions [68]. Additionally, it was observed that saline environments affected by climatic conditions had a significant negative impact on N₂O fluxes (Figure 7; with a direct effect of −4.34), resulting in lower N₂O flux exchange for samples with higher salinity. This could be attributed to the following factors: (1) The low productivity of saline soils and the high intensity of agricultural production, both of which contribute to N₂O emissions. (2) The fact that as salinity levels increase, the stress effect also increases, leading to the inhibition of nitrification. This, in turn, results in high soil NH₄⁺-N retention, making it difficult for effective conversion to NO₃^—N to take place [69,70]. Low salinity has been found to stimulate the processes of nitrification and mineralization of soil nitrogen. Conversely, elevated levels of soil salinity have been shown to weaken the capacity for soil nitrification and may potentially lead to an increase in the intensity of soil ammonia volatilization. These findings highlight the significant impact that salinity levels can have on the nitrogen cycle within soils. Such insights are crucial for understanding and managing agricultural systems in saline environments. In coastal wetlands, environmental factors such as soil temperature, hydrological conditions, nutrients, pH, and salinity significantly impact the exchange of CO₂ and N₂O fluxes in soil. Salinity is particularly noteworthy among these factors due to its unique environmental influence.

4.2. Responses of GHG and Soil–Vegetation Carbon and Nitrogen Pools Under Different Years of Reclamation

The combination of the soil respiration and greenhouse gas sequestration capacity of coastal reclaimed cropland with its environment constitutes an agroecosystem. This special ecosystem is intricately linked to the functioning of its various components. In turn, this ecosystem process is driven by a range of soil physical and biogeochemical processes [35]. Soil organic carbon (SOC) serves as a crucial indicator of soil fertility and quality and plays a significant role in the global carbon cycle [71]. SOC can impact the exchange fluxes of atmospheric CO₂ by enhancing microbial activity, by providing material and energy to soil microorganisms. In this study, it was observed that SOC had a positive influence on CO₂ fluxes, although the effect was not statistically significant (p > 0.05) (Figure 6 and Figure 7). This could be attributed to the complex and variable environment of the reclamation area, which is influenced by both terrestrial and marine ecosystems to varying degrees. The reclaimed areas of the coastal mudflats in Rudong are of different ages, leading to variations in SOC affected by tides and climate before reclamation, as well as subsequent improvement measures, such as desalination of saline soils during conversion into agricultural land in different reclaimed areas [72]. Reclamation has resulted in significant environmental disturbances. Previous research has indicated that levels of soil organic carbon (SOC) mineralization in coastal soils exhibit significant disparities before and after reclamation [73]. The conversion of vegetated wetlands to alternative land uses may lead to a reduction in SOC, resulting in a relatively narrow range of spatial variation in SOC across different reclaimed areas [74]. However, it is important to consider the potential influence of the limited time scale for observation on establishing a more substantial correlation between SOC and CO₂ flux. In addition, when considering the synthesis of factors influencing the modeled pathways, it is important to note that the storage state of soil organic carbon (SOC) in coastal reclaimed areas is often influenced by a combination of soil properties [75]. It is also possible that changes in other properties, such as salinity and acidity/alkalinity, may buffer the sensitivity of SOC and its ability to alter CO₂ fluxes [76]. The study area has a complex history of land use and farming practices, which have led to environmental disturbances in different reclaimed areas. Additionally, different tillage management practices can impact CO₂ flux exchange [77], with CO₂ fluxes potentially increasing with higher tillage intensity [78].

The increase in soil nitrogen (N) content due to farming practices had a significant positive impact on N₂O emissions (direct effect of 3.301; Figure 7). Soil N₂O production primarily occurs through denitrification processes, in which NH₄⁺-N and NO₃⁻-N are transformed into each other [27]. The denitrification reaction is influenced by enzymatic reactions in the soil, with an increase in NO₃⁻-N accelerating the process. This study also found a significant correlation between N₂O flux and soil NO₃⁻-N content, which positively affected the N₂O flux balance (Figure 6 and Figure 7c,d), consistent with previous empirical studies showing a positive correlation between N₂O emissions and NO₃⁻-N concentrations [79,80]. Additionally, previous research has shown that the use of nitrogen fertilizers significantly increases N₂O emissions [81,82]. In agricultural ecosystems heavily impacted by human activities, the regularity of N₂O flux exchange is closely linked to the application rate of nitrogen fertilizer. According to our survey, in the study area, farmers typically make decisions regarding nitrogen fertilizer application based on crop growth. In conclusion, saline environments, diversity in the fertilizers used, and rates of nitrogen fertilizer application collectively influence the results for N₂O flux and balance within farmland ecosystems.

Different types of vegetation respond differently to temperature, particularly through their root systems. It has been observed that the root system of vegetation exhibits a greater sensitivity to temperature concerning soil respiration [83], which aligns with the findings depicted in this study (Figure 7). Additionally, it is worth noting that temperature can also impact soil respiration by affecting microbial activity [84]. In this study, the carbon sequestration capacity of each sample plot was found to be higher in summer than in winter and spring. This difference may be closely related to temperature, root system, and microbial activity status. The effect of a saline environment on soil carbon and nitrogen pools is also significant. In coastal saline wetlands, the soil matrix was affected by seawater immersion and had a high original pH value. As a result, the SOC content may be low [85]. Previous studies have indicated that soil salinity is significantly correlated with pH in terms of physical and chemical properties but negatively correlated with soil nutrients, the fractal dimension of soil particle fractions, and soil organic matter. It can be concluded that soil salinity is the most important factor influencing soil quality in coastal agricultural reclamation areas under widespread saline soils. Additionally, excessive salinity causes water loss to plants through osmotic stress, which affects crop survival, as well as carbon and nitrogen exchanges [66,86]. This finding aligns with the observation that salinity and pH had significant negative effects on vegetation carbon pools in this study (Figure 7). The coupling mechanism of CO₂ and N₂O fluxes with soil and vegetation nitrogen pools in this study exhibits a certain level of complexity, which is attributed to the unique environmental conditions resulting from reclamation. Before reclamation, the carbon and nitrogen contents of natural beach salt marsh wetlands were primarily influenced by tides, geomorphology, and saline vegetation. However, after reclamation, soil carbon and nitrogen pools are predominantly impacted by cropland management and agricultural practices [33]. With the transition from wetlands to agricultural land, the process of farmland management and crop growth during agricultural cultivation activities may alter the structure and diversity of biotechnology in sediments [87]. This alteration could potentially accelerate the decomposition of organic matter and greenhouse gas emissions in sediments.

Reclamation fundamentally alters coastal ecosystem equilibrium, with measurable impacts on biodiversity and biogeochemical cycling. Specifically, empirical evidence confirms tidal influence reduction in reclaimed zones [88], driving hydrological shifts, including water table depression. Furthermore, the reclamation of coastal wetlands also impacts the physiological processes and community structures of native salt-tolerant vegetation [89]. This, in turn, weakens soil drought biochemistry and anaerobic environments while affecting the composition and activity of soil microbial communities in response to changes in soil respiration processes at different levels. Additionally, soil salinity and pH are altered after coastal wetland reclamation, impacting not only the vegetation’s physiological processes and ecological status but also the CO₂ and N₂O balance within reclaimed wetland areas. When agricultural production is carried out on reclaimed land, unique types of coastal reclaimed wetlands can form. Due to their shallow water table, many shorter-reclaimed-year wetlands still maintain varying degrees of environmental characteristics typical for wetlands, further complicating soil respiration within these coastal wetland reclamation areas. Moreover, reclamation has resulted in stagnation of material and energy exchanges between coastal reclaimed areas and offshore regions [90], leading to anthropogenic interference in greenhouse gas fluxes within coastal wetlands.

4.3. Insights and Perspectives on Sustainable Planting in Coastal Wetland Development and Reclaimed Areas

Currently, there are significant changes in land use types along the southeastern coastline of China, particularly involving the transition between natural salt marshes and aquaculture ponds, as well as agricultural cropland [91]. The dynamics of wetland exploitation and coastal greenhouse gas fluxes along China’s coastline have been closely linked to government policies over the last decade or so [7,92]. This is evident in China’s large-scale reclamation of coastal saline soils to meet the growing demand for arable land for food security [33]. However, it is important to note that coastal wetland ecosystems are more fragile systems that are highly sensitive to climate change and human activities. It is widely recognized that large-scale reclamation and associated activities can disturb and damage coastal wetlands, leading to environmental risks such as soil salinization, coastal eutrophication, and reductions in species habitats. The natural salt marsh area in Jiangsu Province has significantly increased as a result of conservation and restoration policies implemented during development [30]. The presence of a substantial portion of wetlands with vegetative cover plays a crucial role in absorbing CO₂ and N₂O [23]. The advantageous regeneration and restoration rate of mud coasts in the coastal zone makes it feasible to carry out coastal zone development with suitable reclamation at an acceptable rate.

The term “carbon sink in agroecosystems” refers to the process through which atmospheric CO₂ is absorbed by agroecosystems and then fixed into crop and soil carbon pools. This involves the effective absorption of atmospheric CO₂ by crops, aided by planting and management techniques. While farmland soil organic carbon levels are relatively low compared to those found in natural landscapes, agriculture can still contribute to negative emissions through effective management [93]. Our research shows that in comparison to the greenhouse gases released by converting coastal salt marshes to aquaculture ponds [94,95,96], developing coastal salt marshes into agricultural land would be a more viable option. The farmland ecosystem in a reclamation area is an open system formed by the joint action of humans and nature, encompassing not only living and non-living elements on the farmland, but also their interactions. Human activities such as cultivation, fertilization, irrigation, weeding, and pest control have direct or indirect impacts on farmland ecosystems. However, farmland is also a significant source of N₂O emissions. When considering the global warming potential analysis of coastal reclaimed wetlands, it is important to note that while reclaimed farmland does contribute to N₂O emissions and has a global warming potential 265 times that of CO₂ gas, the sink function of CO₂ makes a substantial contribution to reclaimed farmland (Table 2). Therefore, coastal reclaimed wetlands still serve as sinks for greenhouse gases. Furthermore, differences in GHG flux exchange and yields on coastal reclaimed agricultural land can be attributed to natural conditions, farming systems in different reclaimed areas, and the types of crops grown [97,98]. Effective management of coastal reclamation agricultural land is therefore crucial.

Our analysis shows that corn has a superior ability to mitigate CO₂ emissions compared to other crops. During the growth of snap beans and soybeans, N₂O emissions were significantly reduced compared to the growth of non-legume crops. This suggests that the crops in the cropping system are of decision-making significance for the agroecosystem as a whole. The human activity of reclamation cultivation provides valuable information on various land use practices, farmland management techniques, and the cropping preferences of different farmers [31]. This includes practices such as fertilizer application, irrigation, plowing, and the return of straw to fields, among others. In addition, there are additional effects on soil respiration from different farming systems under varying climatic conditions. Numerous studies have demonstrated that the analysis of factors influencing CO₂ and N₂O fluxes, including the production, transportation, storage, and utilization of agricultural resources such as fertilizers and pesticides, as well as the impacts of agricultural activities on carbon balance [27], involves more intricate considerations. Research has indicated that diverse crop rotations and fertilization practices can directly alter the carbon sequestration and emission reduction capacity of soils [99]. In this study, different farming methods were employed to distinguish between wheat–corn–soybean rotation and the continuous cropping of wheat, which are managed by large professional households, and fava bean–corn rotation and wheat–corn rotation, which are operated by family farms. The two reclaimed plots managed by large professional households have a significant sown area, with activities such as fertilizer application, planting, pesticide spraying, and harvesting now being mechanized on a large scale. As a result, various disturbances exhibit complex mechanisms that require further support based on controlled and observational studies [100].

While this study provides novel insights into the GHG dynamics of reclaimed coastal agroecosystems, several limitations warrant consideration. Residual fertilizer impacts from pre-study years were not explicitly measured, although the soil N testing (0–60 cm) indicated minimal nitrate accumulation. Future studies should employ ¹⁵N isotopic tracing to quantify carry-over effects. Groundwater interactions and wind erosion effects were not quantitatively assessed due to methodological scoping. Their potential influence on carbon–nitrogen cycling warrants targeted investigation in future studies. Furthermore, while ANOVA and PLS-PM provided robust frameworks for our research questions, future studies could benefit from mixed-effects models to account for potential temporal autocorrelation in multi-seasonal measurements and plot-level random effects. Such approaches would be particularly valuable when analyzing high-frequency temporal dynamics within individual reclaimed plots. The limitations, however, do not invalidate our core findings regarding reclamation age effects, but instead highlight opportunities for mechanistic depth in coastal agroecology.

5. Conclusions

In this study, we comprehensively assessed greenhouse gas (GHG) fluxes from cultivated land in four typical reclaimed areas of wetland along China’s southeastern coast. By analyzing the drivers of the GHG fluxes in these ecosystems, we evaluated the importance of sustainable coastal reclamation development and GHG mitigation while ensuring food security. The key findings reveal distinct temporal patterns in CO₂ exchange. During crop growing periods, most agricultural lands function as CO₂ sinks, and during fallow periods, reclaimed lands predominantly become CO₂ sources. This pattern correlates strongly with the temperature-driven enhancement of soil respiration. Collectively, the coastal agroecosystems demonstrated significant carbon sequestration capacity, with annual mean CO₂ uptake values of −7446.77 kgC ha⁻¹, −35,678.02 kgC ha⁻¹, −51,273.55 kgC ha⁻¹, and −81,181.77 kgC ha⁻¹. However, the systems consistently acted as net N₂O emission sources, with annual fluxes of 7.329 kgN ha⁻¹, 5.016 kgN ha⁻¹, 1.674 kgN ha⁻¹, and 0.701 kgN ha⁻¹. Notably, when integrating all the GHGs using global warming potentials (GWP), these reclaimed wetlands functioned as net GHG sinks due to dominant CO₂ sequestration. Our study also demonstrated the carbon fixation effect of corn crops and the nitrogen fixation effect of legume crops. The RDA and PLS-PM analyses indicated that the exchange of CO₂ and N₂O gas fluxes was influenced by soil and vegetation physico-chemical properties, as well as various factors, including time, space, environment, and anthropogenic disturbances. The main driving factors were found to be soil temperature, moisture, salinity, and pH. Agricultural land reclaimed from coastal wetlands holds significant ecological importance. On one hand, the conversion of coastal wetlands into agro-ecosystems plays a crucial role in altering greenhouse gas emissions. On the other hand, farmland ecosystems also possess the capacity to sequester carbon. This is primarily demonstrated through ameliorating soil salinity and alkalinity over time following reclamation. Furthermore, adopting more rational agricultural management practices will further enhance crop carbon sequestration as the duration of reclamation increases. In conclusion, this study offers the following recommendations: (I) Effective measures should be implemented to enhance the carbon sequestration capacity of farmland ecosystems in coastal reclaimed areas by exploring reasonable agronomic practices and farmland management. (II) Deeper and longer-term research is needed to understand the carbon sequestration characteristics of farmland ecosystems in coastal wetland reclamation areas and their influencing factors, providing a theoretical basis for developing and protecting coastal zones.