Geochemical Regulation of Heavy Metal Speciation in Subtropical Peatlands: A Case Study in Dajiuhu Peatland

Abstract

Heavy metals in peatland pose significant ecological risks due to their persistence, bioaccumulation, and dynamic mobilization under fluctuating environmental conditions. Understanding heavy metal dynamics in subtropical peatlands is critical for addressing global gaps in wetland metal cycling, as these ecosystems face intensified organic decomposition and climatic fluctuations that amplify mobilization risks—contrasting starkly with stable northern counterparts. This study investigates the geochemistry of heavy metals (Cr, Cu, Cd, and Pb) of Dajiuhu peatland in central China, using sequential extraction, gradient diffusion (DGT), and random forest modeling. The mean concentrations of Cr, Cu, Cd, and Pb in peat samples were 24.6 ± 13.7 mg/kg, 14.9 ± 2.51 mg/kg, 1.15 ± 0.62 mg/kg, and 54.9 ± 16.16 mg/kg. Principal component analysis identified three sources: plant-derived litter, bedrock weathering, and atmospheric deposition. Metal speciation revealed the predominance of residual fractions (Cr: 64%, Cu: 61%, Pb: 65%, Cd: 35%), with Cd exhibiting higher mobility (exchangeable: 20%, reducible: 25%). DGT measurements further confirmed distinct migration behaviors, as Cd stored in peat actively diffuses into the surrounding environment, while Pb present in the environment becomes immobilized within the peat matrix. Environmental factors regulate heavy metal speciation through distinct mechanisms. The exchangeable fractions of Cu and Cr are primarily controlled by the C/N ratio, whereas their oxidizable forms are significantly associated with Al content and pH levels. The exchangeable fractions of Pb and Cd are largely influenced by oxidation-reduction potential (ORP) and Ca concentrations, and their reduced forms are closely linked to total sulfur (TS) content. Furthermore, the reducible fractions of Cr and Cd are not only regulated by ORP but also modulated by TS. Our study highlights that the mobility of heavy metals in subtropical peatlands is likely to increase substantially as a result of environmental changes.

1. Introduction

Heavy metals represent a significant global environmental pollutant, being widely distributed in water, soil, and air [1]. These metals can transfer through the food chain, enter the human body via ingestion, and cause adverse health effects when critical thresholds are exceeded [2]. In addition to their high toxicity, most heavy metals exhibit persistence, non-biodegradability, and substantial bioaccumulation potential [3,4,5]. Peatland sediments, which are rich in organic matter, contain functional groups such as carboxyl, phenolic, and thiol, which bind with metals, enabling peatlands to effectively accumulate heavy metals and act as an important sink [6,7]. Since the Industrial Revolution, intensified anthropogenic activities have resulted in considerable deposition of heavy metals in peatlands [3], with approximately 85% of these metals accumulating in surface sediments [8]. Under periodic flooding conditions, dynamic biological and physicochemical processes may lead to the mobilization and release of heavy metals from these sediments [9,10]. Previous research has shown that heavy metal concentrations in peatland sediments can vary by 5 to 8 times under periodic dry–wet alternation conditions [11]. Hence, understanding the mechanisms of heavy metal accumulation, storage, and release in peatlands is crucial for mitigating their ecological risks.

Recent investigations into heavy metals in peatlands have predominantly focused on northern peatlands at high altitudes or latitudes, while research on tropical and subtropical peatlands remains relatively limited [11,12,13]. Tropical and subtropical regions exhibit more complex ecosystems and experience more pronounced climatic fluctuations compared to high-latitude and high-altitude areas [11,12]. Elevated temperature and precipitation conditions not only enhance plant productivity but also accelerate the decomposition of organic matter in peat [14,15]. The reduced heavy metal adsorption capacity of peat, attributed to the release of decomposed organic matter and the production of small-molecule organic acids, results in a significant increase in mobile heavy metal concentrations within peatlands [16,17]. Previous research has shown that the transformation of organic matter significantly influences the Hg content and stock in peat, with Hg concentrations gradually decreasing from the surface to deeper layers [18]. Investigations in Indonesian peatland demonstrated that increased levels of dissolved organic matter markedly enhance the proportions of dissolved As, Cd, and Zn [19]. The high-temperature and rainy climate conditions prevalent in tropical and subtropical regions significantly modify the biogeochemical cycles of heavy metals, thereby questioning the applicability of heavy metal fixation theories developed for northern peatlands. Although tropical and subtropical peatlands occupy a relatively small area, they account for approximately 20% of global carbon storage [20]. Therefore, elucidating the migration and transformation mechanisms of heavy metals in these regions is essential for mitigating their ecological risks.

The migration and transformation of heavy metals in peatlands are influenced by their speciation [3]. Using conventional chemical extraction methods, heavy metals can be divided into four distinct fractions: the exchangeable fraction, which is highly sensitive to pH fluctuations and carbonate binding; the reducible fraction and oxidizable fraction, which are primarily associated with Fe/Mn oxides as well as organic matter and are influenced by redox potential; and the residual fraction, which is closely related to soil-forming minerals and exhibits minimal mobility [21,22,23]. The speciation of heavy metals in sediments is largely determined by the environmental characteristics of wetlands. For instance, in Sundarben wetland soils, Cd and Pb predominantly exist in exchangeable and reducible forms [24]. In Severn River wetlands, Cu, Pb, Zn, and Cr are mainly bound to carbonates and Fe/Mn oxides [25]. Recent studies have demonstrated that under rapidly changing geochemical conditions in wetlands, traditional chemical measurement methods may be significantly affected [26,27]. Rapid changes in dissolved organic matter, redox potential, acidity, and root exudates alter the distribution of heavy metals in wetland sediments and influence measurement results [26,27]. The gradient diffusion film technique (DGT) provides an effective method for analyzing in situ heavy metal migration because it is unaffected by pH, ionic strength, substrate composition, or flow velocity [28,29]. This technology has been widely applied in dynamic systems such as mangrove wetlands and estuarine wetlands [22,29]. Therefore, the integrated use of traditional chemical analysis methods and gradient diffusion film technology can more effectively elucidate the behavior of heavy metals in rapidly changing systems, such as subtropical peatlands.

To investigate the geochemical behavior of heavy metals in subtropical peatlands, a representative subtropical peatland (Dajiuhu, DJH), located in central China, was selected as the study site. Traditional chemical extraction methods and DGT were combined to elucidate the distribution patterns, sources, and migration/transformation mechanisms of different heavy metals within the peatland system, incorporating various geochemical factors and data analysis techniques.

2. Materials and Method

2.1. Study Area and Sample Preparation

The DJH peatland is situated in a subalpine mountain basin (31°28′50″ N, 110°00′09″ E) within the Shennongjia region of Hubei Province in central China. This region lies at the ecotone between the subtropical and warm temperate zones and is characterized by a monsoon-influenced climate [11]. The peatland ranges in elevation from 1700 m to 1760 m above sea level, covers a total area of 16 km², and has an average annual temperature of 7.4 °C. Vegetation is predominantly composed of Carex spp., Sanguisorba officinalis, Juncus spp., Gentiana pseudo-aquatica, and Sphagnum spp. [11].

The sampling points are depicted in Figure 1. Pore water, peat, and vegetation samples were systematically collected during the wet season in July 2023, when water levels and temperatures reached their maximum values.

Pore water samples (n = 16) were systematically extracted using a pore water sampler (Rhizon MOM, Rhizosphere, Wageningen, The Netherlands). Sampling was conducted at the sediment surface layer (0–10 cm depth), with each collection lasting 12 h and yielding approximately 100–150 mL per sample. The samples were then transferred to 100 mL constant-volume bottles, where measurements of pH, oxidation-reduction potential (ORP), temperature, electrical conductivity (EC), and salinity (SAL) were performed using a water quality parameter meter (Bante 900P, Shanghai, China). To ensure appropriate preservation, 5% HNO₃ was added immediately after collection.

Peat samples (0–10 cm depth, n = 16) were collected from the same locations as the pore water samples, with each sample weighing approximately 200 g. Immediately after collection, the samples were stored at −4 °C and then transported to the laboratory for freeze-drying. After sieving through a 200-mesh sieve, the samples were prepared for subsequent analysis.

The green parts of Sphagnum were meticulously harvested using Teflon scissors and subsequently rinsed with ultra-pure water to remove surface contaminants (n = 10). The samples were then stored at −4 °C in incubators for short-term preservation prior to transportation. Upon arrival at the laboratory, the samples were subjected to freeze-drying, followed by grinding to a 200-mesh size, and subsequently stored under appropriate conditions for further analysis.

The DGT device (DGT, Lancaster, UK) was employed to enrich elements such as Cd, Pb, Cr, and Cu [29]. This sampler utilized Chelex gel as the adsorption film material, with a surface area of 4.9 cm² and a volume of 0.15 mL. The diffusion membrane had a thickness of 0.078 cm, while the filter membrane was 0.014 cm thick. Prior to deployment, all DGT units were individually packaged in clean, sealed plastic bags and stored refrigerated at 4 °C. Six hours before deployment, the devices were removed from refrigeration and allowed to equilibrate to ambient temperature. Each DGT device was marked 1–2 cm from the top of the sampling window (near the handle) and vertically inserted into the sediment until the mark aligned with the sediment/water interface (n = 5). After 24 h of on-site exposure, the DGT device was recovered with care taken to avoid contact with the filter membrane on the device surface.

The DGT device was then rinsed with distilled or deionized water and prepared for further analysis.

2.2. Experimental Methods

For the determination of elemental concentrations (Ca, Mg, Al, Cd, Pb, Cr, and Cu), peat samples (50 mg, n = 16) were digested using a mixture of concentrated HF (1 mL, 40% v/v) and HNO₃ (2 mL, 65% v/v) in PTFE-lined stainless-steel vessels heated in an oven [30]. Moss samples (20 mg) were digested with a combination of HF (1 mL, 40% v/v), HClO₄ (0.5 mL, 70% v/v), and HNO₃ (2 mL, 65% v/v) [30]. Filtered pore water samples (10 mL) were analyzed directly without additional pretreatment [30]. The concentrations of the elements (Ca, Mg, Al, Cd, Pb, Cr, and Cu) in the digestion solutions were measured using an inductively coupled plasma mass spectrometer (ICP-MS, Agilent 7900, Agilent, Santa Clara, CA, USA). Analytical quality and accuracy were ensured by analyzing reagent blanks and certified reference materials (GBW-07423 for peat and GBW-10020 for moss; Institute of Geophysical and Geochemical Exploration, Langfang, China). Recovery rates for the elements in the reference materials ranged from 95% to 110%, and the relative standard deviations of triplicate measurements were below 5%. Water quality was evaluated using internal standards, blanks, and triplicates, with recovery rates ranging from 90% to 115% and relative standard deviations of ±5%.

The DGT technique effectively simulates the dynamic migration process of bioavailable heavy metals and quantifies the exchange flux of heavy metals at the water-sediment interface [29]. Specifically, the binding membrane selectively adsorbs free or weakly complexed heavy metals, excluding colloidal-bound or strongly bound species [31]. The adsorption film at the bottom of the DGT device (n = 5) was carefully transferred into a 50 mL clean centrifuge tube. Subsequently, 2 mL of HNO₃ (1 mol/L) was added as the elution solution to ensure that the entire adsorption film was fully immersed and completely unfolded within the solution. The centrifuge tube was then placed on a shaker and oscillated for 8 h to achieve efficient elution. After oscillation, the eluent was carefully transferred to a 15 mL centrifuge tube and diluted with ultra-pure water to a final volume of 5 mL for heavy metal analysis. Heavy metal concentrations were quantified using ICP-MS [29].

Heavy metal extraction from sediments was performed according to the BCR sequential extraction procedure [32]. Due to the high content of plant roots and water in the two surface peat sediments, which failed to meet the requirements for the BCR sequential extraction test, we were limited to analyzing only 14 surface sediment samples. For the exchangeable fraction, peat sample (0.5 g, n = 14) was mixed with HAc (20 mL, 0.11 mol/L), shaken for 16 h at 250 r·min⁻¹ and 22 °C, and subsequently centrifuged at 4000 r·min⁻¹ for 20 min. The supernatant was collected and stored at 4 °C, while the residue was washed with 10 mL of ultra-pure water. For the oxidizable fraction, NH₂OH·HCl (20 mL, 0.5 mol/L) was added to the residue, shaken under identical conditions for 16 h, and then centrifuged to separate the supernatant. The residue was washed again. For the reducible fraction, H₂O₂ (5 mL, 30% v/v) was incrementally added and allowed to dissolve at room temperature for 1 h. The mixture was subsequently heated in a water bath at 85 °C until near dryness, followed by repeated dissolution and cooling. Then, NH₄OAc (25 mL, 1 mol/L) was added, and the mixture was centrifuged. For the residual fraction, the material was sequentially dissolved in HF (1 mL, 40% v/v), HClO₄ (0.5 mL, 70% v/v), and HNO₃ (2 mL, 65% v/v) at temperatures ranging from 120 to 200 °C. After silicon removal, the solution volume was adjusted to 100 mL. If the residue was not fully dissolved, acid treatment was repeated. Finally, the residue was dissolved and reformulated using 0.5 mL HNO₃ (50%, v/v). The concentrations of heavy metal elements in the extraction solutions from each step were determined by ICP-MS. GBW-07423 was employed as the certified reference material. The recovery rate was determined by comparing the total heavy metal content obtained from the analysis with the certified value of the reference material. The recovery rates ranged from 85% to 115%, and the relative difference between duplicate analyses was less than 6%.

Peat samples were dried, ground, and sieved through a 2 mm mesh before pH analysis in the laboratory. The pH was determined using a digital pH meter (Orion Star A111, Thermo Scientific, Waltham, MA, USA) with an integrated combined electrode (Shanghai LeiCi 65–1 C, Shanghai, China), maintaining a soil-to-water ratio of 1:2.5 [33]. The carbon (C), nitrogen (N), and sulfur (S) contents in peat samples (n = 16) were quantified using the combustion method [34]. Approximately 5 mg of peat sample was accurately weighed, wrapped in tin foil (4 × 6 mm), and combusted at 1020 °C. The resulting gases (CO₂, N₂, and SO₂) were subsequently dried at 70 °C using anhydrous Mg(ClO₄)₂, and their concentrations were determined using an elemental analyzer (DeltaV Advantage, Thermo Fisher, Waltham, MA, USA). Sulfadiazine (C = 41.81%, N = 16.25%, S = 18.62%; Thermo Fisher, USA) was used as the quality control sample for the determination of the elemental compositions (% C, % N, and % S) of carbon, nitrogen, and sulfur. The measurement accuracies for C, N, and S contents were 0.1‰, ±0.1‰, and ±0.2‰, respectively.

2.3. Statistical Analyses

The value of heavy metal content of DGT in peat was calculated according to (1,2):

M = C_e × (V_acid + V_gel)/fe

(1)

C_DGT = M × Δg/D_d × A × t

(2)

where Ce represents the concentration of the target substance in the eluent, which can be calculated by multiplying the measured concentration by the dilution ratio; V_acid denotes the volume of the eluent used, with a value of 2 mL; V_gel refers to the volume of a single adsorption film (Chelex adsorption film, 0.15 mL); fe indicates the elution efficiency of the target substance, with values of 0.8 for Cd, Pb, Cr, and Cu in the Chelex film; Δg represents the thickness of the diffusion layer, which is the sum of the thickness of the diffusion film (0.078 cm) and the filter film (0.014 cm); and Dd represents the diffusion coefficients of the target substances in the diffusion layer at the DGT experimental temperature (22 °C). The diffusion coefficients of Cd, Pb, Cr, and Cu are 5.61 × 10⁻⁶ cm²/s, 7.40 × 10⁻⁶ cm²/s, 4.65 × 10⁻⁶ cm²/s, and 5.74 × 10⁻⁶ cm²/s, respectively [29]. t denotes the placement time (s), recorded during the field DGT experiment, and A represents the sampling window area of the DGT device, with a value of 4.9 cm².

We utilized the compensation coefficient (K) to quantify heavy metal concentrations in peat, thereby mitigating the impact of regional background values on the evaluation of heavy metal accumulation in peat [35]. The calculation of the K was undertaken according to (3):

$$K=1+0.5\times \left({\displaystyle \frac{\mathrm{Regional}\mathrm{measured}\mathrm{values}{CV}_B}{20\mathrm{\%}}}\right)$$

(3)

Elemental content analysis was conducted using Origin 2019 software. Principal component analysis (PCA) was subsequently employed to determine the relationships among various elements based on the standardized correlation matrix. Cluster analysis (CA) was performed using the average linkage method, with Euclidean distance and the sum of squared deviations serving as measures of data similarity. Furthermore, random forest modeling was applied to assess the relative contributions of Al, Mg, Ca, SAL, ORP, pH, C, N, and S to heavy metal concentrations.

3. Results and Discussion

3.1. Descriptive Statistics of Heavy Metals and Chemical Parameters in Soil, Moss, and Pore Water

The descriptive statistics for metal parameters in peat, pore water, and moss are presented in Figure 2. The mean concentrations of Cr, Cu, Cd, and Pb in peat samples were 24.6 ± 13.7 mg/kg, 14.9 ± 2.51 mg/kg, 1.15 ± 0.62 mg/kg, and 54.9 ± 16.16 mg/kg, which were higher than those in the soil of Shennongjia (Cu: 13.53 mg/kg, Pb: 9.89 mg/kg, Cd: 0.49 mg/kg), except Cr (32.39 mg/kg) [36]. The concentrations of Cr, Cu, and Pb were below the risk screening thresholds specified in the Chinese Soil Environmental Quality Risk Control Standard (GB15618-2018) [37]. In contrast, the Cd concentration in peat exceeded the threshold outlined in the same standard [37]. The compensation coefficients of heavy metals in the peat of Dajiuhu are as follows: Cr (2.39), Cu (1.42), Cd (2.35), and Pb (1.73). These results indicated that the Dajiuhu peatland has undergone significant enrichment and accumulation processes relative to the surrounding soil. The high coefficients of variation (CVs) for Cr (56%), Cu (17%), Cd (54%), and Pb (29%) suggest significant spatial heterogeneity in their distributions. Similarly, the average values of water level (2.92 ± 7.95 cm), ORP (3.99 ± 1.90 mV), SAL (0.04 ± 0.02 ppt), and EC (3.99 ± 1.90 μS/cm) exhibited substantial variability across sampling points, with CVs of 272% (water level), 48% (ORP), 50% (SAL), and 42% (EC). Notably, the pH values (4.52 ± 0.34) at various points in the wetland were slightly acidic, with minimal variability (CV: 7.7%). Despite no statistically significant correlation being observed between heavy metal content and changes in water level, ORP, SAL, pH, or EC (based on the results of the correlation analysis, p > 0.05), these factors may collectively influence the geochemical behavior of heavy metals in wetlands. Previous studies in the DJH region have shown that variations in water levels driven by peatland topography significantly affect the distribution of Hg in peatlands [38]. Furthermore, the percentage differences in C (56 ± 11%) and N (3.7 ± 0.59%) contents at each point were not significant, indicating that C and N inputs from plant sources are distributed relatively uniformly throughout the peatland. This uniformity suggests that humus, which is closely associated with plant input, may play a buffering role in stabilizing pH levels in peatland.

The concentrations of Cr and Cu in the pore water were relatively low, falling below the detection limit. The average concentrations of Cd and Pb were 0.09 ± 0.06 μg/L and 0.67 ± 0.17 μg/L, respectively (Figure 2). The average metal concentrations in DJH water samples were markedly lower than the Class I standard limits specified in GB3838-2002 [39], indicating that the heavy metal content in this region is at a pristine level. Moreover, the heavy metal concentrations in pore water were lower than those in the surface water of the Shennongjia area (DJH peatland located within the Shennongjia basin) [30]. In the relatively unpolluted Shennongjia region, the sources of heavy metals in surrounding soils may be analogous to those in the peat of DJH; however, the preservation forms of heavy metals may differ substantially. Organic-bound states represent a critical occurrence form of heavy metals in soil [40]. Under conditions of high organic matter content, free heavy metals in the soil may bind with organic matter, forming fewer mobile organic-bound states [41]. Conversely, the organic matter content in the soils of the Shennongjia area is considerably lower than that in the peat of DJH, leading to a greater likelihood of heavy metal release from the soil. By contrast, in the peatlands of DJH, heavy metals tend to form stable complexes with peat organic matter and are therefore less likely to be released.

The content of heavy metals in Sphagnum is shown in Figure 2. The average contents of Cr, Cu, Cd, and Pb were 1.76 ± 0.43, 5.20 ± 1.07, 0.37 ± 0.10, and 5.34 ± 2.92 mg/kg. The heavy metal content of Sphagnum in DJH was also significantly lower than that in nearby mountains and nearby cities (Chengdu) [42,43]. Heavy metals in moss are mainly derived from atmospheric deposition rather than root absorption [44]. DJH peatland is a rain-fed peatland, and heavy metals from dry and wet atmospheric sedimentation are the most important sources of the peatland [11]. This also shows that heavy metals transported from the atmosphere over long distances have not caused pollution to the sphagnum moss in DJH peatland. In addition, the content of heavy metals in Sphagnum is significantly lower than that in peat. The difference in heavy metal content between sphagnum moss and peat indicates that Sphagnum accumulates gradually during the process of cumulative degradation. The coefficients of variation of heavy metals (Cr, Cu, Cd, and Pb) were 24%, 21%, 26%, and 55%, respectively. This indicates that the heavy metals from atmospheric deposition sources received by the whole peatland are relatively average, except Pb.

3.2. Sources of Heavy Metals in Peat

The entire dataset (Ca, Mg, Al, C, N, Cr, Cu, Cd, and Pb) was analyzed using a combination of PCA and CA (Figure 3). The results of PCA and CA exhibit consistency. PCA identified three main factors influencing the concentration of heavy metals. Factor 1 accounts for 39.76% of the total variance and is characterized by C, N, and the C/N ratios, as well as positive loadings of certain metals (Cr and Cu). The contents of C and N in peat may be influenced by the amount of litter supplied by plant death, the decomposition process of plants, and accumulation conditions [45]. The strong correlation between heavy metals and carbon and nitrogen suggests that heavy metals may originate from plant absorption and subsequently enter peatland via litter deposition. Factor 2 explains an additional 22.05% of the variance and is distinguished by the positive loading of Ca. Combined with the results of cluster analysis, Ca, Mg, and Al form a group. Carbonate rocks serve as the bedrocks in DJH, with Ca and Mg being key components of these rocks [46]. Previous research has demonstrated that the weathering of carbonate rocks can lead to increased transport of Ca and Mg in wetland [47]. The input of Ca and Mg may reflect contributions from the weathering of bedrocks. Notably, Factor 3 explains 16.26% of the total variance and is characterized by positive factor loadings of Pb, Cd, and S. In conjunction with the results of CA, a group is also formed by Pb, Cd, and S, indicating their common origin. Pb and S emitted from anthropogenic and natural sources enter the atmosphere in gaseous and particulate forms (e.g., SO₂, SO₄²⁻, PbO, and PbSO₄) [48]. Once in the atmosphere, S and Pb can remain suspended for several days to weeks and, through atmospheric circulation, undergo dry and wet deposition before entering wetlands [49]. Isotopic characteristics and contribution rates of Pb in the soil of the Shennongjia area suggest that Pb may originate from long-distance atmospheric transport and subsequent deposition [50]. In the atmosphere, compared with Pb, Cd shows a greater tendency to adsorb onto finer particles and is more prone to forming volatile compounds, such as CdCl₂ [51], which facilitates its retention in the atmosphere (10 to 15 days) and long-distance transport. The Shennongjia area is a natural scenic region. Despite its relatively low levels of industrial and agricultural emissions, previous studies have indicated that garbage produced by local residents, vehicle emissions, and increased tourist activities collectively constitute significant sources of Cd pollution. The emitted Cd subsequently enters the soil and atmosphere, leading to a substantial increase in the Cd content of peat [36]. Compared with heavy metal absorption by plants through leaves and complexation with organic matter in litter, heavy metals entering peatland via direct atmospheric deposition may exhibit higher reactivity. Therefore, based on the above analysis, the content of heavy metals in the peat of DJH may be jointly influenced by the input of plant residues (indirect atmospheric input), the weathering of bedrocks, and direct atmospheric deposition.

3.3. Heavy Metal Speciation in Sediments

As depicted in the Figure 4, the distribution of various forms of different heavy metals demonstrates substantial variation. Specifically, Cr and Cu were predominantly found in the residual fractions (64% and 61%, respectively), followed by the oxidizable fractions (35% and 37%), reducible fractions (0.72% and 0.87%), and exchangeable fractions (0.26% and 1.64%). A small proportion of Cu (0.72%) and Cr (0.87%) may be adsorbed onto the surfaces of goethite and Al hydroxide or co-precipitated with Fe/Mn oxides, thereby forming reducible states [52]. Under sulfur-rich conditions, most Cu and Cr form stable complexes with humic acid and fulvic acid through interactions with carboxyl groups and phenolic hydroxyl groups, or precipitate as sulfides, thus forming oxidizable fractions [53]. Consequently, the mobility of Cu and Cr may increase under reoxidation and pH changes. During the transformation of sulfides into sulfates and the degradation of organic matter, these metals are released from peat, potentially posing a threat to the ecosystem. Pb was primarily present in the residual fractions (65%), followed by the oxidizable fractions (17%), reducible fractions (16%), and exchangeable fractions (1.6%). Notably, the proportion of reducible Pb was significantly higher than that of Cr and Cu. Pb can combine with the surface of Fe/Mn oxides through adsorption or co-precipitation [54]. The ionic radius of Pb (1.19 Å) is closer to that of Fe (1.20 Å), leading to higher adsorption efficiency on oxide surfaces (e.g., inner layer complexation) compared to Cu²⁺ (0.73 Å) and Cr³⁺ (0.615 Å), and Pb more readily combines with Fe/Mn oxides [55,56]. The lower proportion of oxidizable Pb relative to Cr and Cu can be attributed to the lower solubility of PbS compared to CuS and the lower complexation stability constant of Pb with organic matter compared to Cr [57]. Cu tends to form sulfides more readily than Pb, while Cr is more dependent on combining with organic matter, resulting in a higher oxidizable state ratio for Cu relative to Cr than for Pb. The forms of Cd were predominantly in the residual fractions (35%), followed by the reducible fractions (25%), oxidizable fractions (21%), and exchangeable fractions (20%). The exchangeable fraction of Cd comprised 21% of the total components, which is much higher than that of Cu (1.64%), Cr (0.26%), and Pb (1.6%); these findings indicate that Cd exhibits relatively high mobility. The charge density of Cd²⁺ (2.13 C·A⁻²) is lower than that of Pb²⁺ (1.39 C·A⁻²) and Cr³⁺ (7.87 C·A⁻²) [54]. Higher charge density enhances the electrostatic attraction of ions to surrounding ligands, promoting the formation of more stable complexes. The low charge density of Cd results in its stability constant being lower than that of Cu and Cr when combined with organic substances and Fe/Mn oxides, contributing to its relatively higher mobility [58]. Collectively, the geochemical properties of Cd contribute to its high mobility in DJH.

The DGT-measured concentrations of Cr, Cu, Cd, and Pb in peat were 0.23 ± 0.07 μg/kg, 0.32 ± 0.08 μg/kg, 0.13 ± 0.09 μg/kg, and 0.45 ± 0.04 μg/kg, respectively. Compared with pore water, the DGT-measured concentrations of Cr, Cu, and Cd were significantly higher, while the concentration of Pb was lower. This suggests that Cr, Cu, and Cd tend to diffuse from sediment into the water body, whereas Pb migrates from pore water to sediment. In conjunction with speciation analysis, Cr and Cu are predominantly present in oxidizable forms (excluding residual states that exhibit limited mobility). Cr³⁺ preferentially forms stable complexes with the carboxyl group (-COOH) in humic acid, while Cu²⁺ is readily chelated by organic ligands containing thiol groups (-SH), such as cysteine. Upon environmental oxidation (e.g., increasing Eh and decreasing pH), the carboxyl group undergoes degradation, and the thiol group is oxidized to disulfide bonds or sulfonic acid groups [58]. These transformations lead to the exposure and subsequent oxidation of Cr³⁺ to Cr⁵⁺, as well as the release of Cu²⁺ from its organic complex [26]. The DGT data for Cd further demonstrate its high mobility relative to other heavy metals, indicating continuous migration [29]. For Pb, atmospheric dry and wet deposition represents the primary pathway for its direct entry into peatland [46,47]. Although Pb’s fixed preservation form exhibits relatively higher mobility compared to Cr and Cu, it tends to form more stable complexes or sulfides, leading to continuous accumulation in peat. The variations in DGT-measured concentrations of Cr, Cu, Pb, and Cd reflect their distinct behaviors in peatlands. Among these, Cd shows the highest propensity for migration, while Cr and Cu exhibit a tendency for slow release from peat into the surrounding environment. In contrast, Pb is predominantly retained and continuously accumulates in peat.

3.4. Geochemical Factors Affecting the Distribution of Heavy Metals in Peat

Random forest analysis was utilized to evaluate the effects of various geochemical factors on the speciation of heavy metals in peatlands, as illustrated in Figure 5.

The exchangeable fraction of Cu is positively influenced by C/N, SAL, and Ca. A higher C/N ratio indicates lower peat decomposition, which enhances the exchangeable Cu fraction due to rapid degradation of labile organic matter [59]. Levels of Ca and SAL are associated with parent material weathering; in the DJH area, carbonate rocks such as limestone and dolomite release Ca, Mg, and HCO₃⁻ through weathering processes, indirectly elevating the exchangeable Cu fraction [46,47]. TS negatively affects Cu exchangeability because CuS precipitates form under anaerobic conditions, immobilizing Cu. The reducible fraction of Cu correlates with organic carbon. High organic matter promotes microbial activity, reducing Fe/Mn oxides and releasing Cu [60]. Secondary minerals re-adsorb Cu, forming a new reducible fraction [61]. Al content positively influences the oxidizable Cu fraction, while increasing pH reduces it. Al³⁺ complexes with organic matter (e.g., humic acid), exposing coordination sites that adsorb Cu²⁺, thereby enhancing organic binding [62,63]. Higher pH decreases adsorption sites or causes deprotonation, weakening Cu binding and reducing the oxidizable fraction [64,65].

Cr exhibits geochemical behaviors similar to those of Cu. The exchangeable Cr fraction correlates positively with C/N and C content but is less responsive to Ca and SAL. Reducible Cr is significantly modulated by redox potential (ORP), showing greater redox variation (trivalent to pentavalent) than Cu [26]. Pentavalent Cr forms CrO₄²⁻ under high ORP, where Fe and Mn oxides enhance Cr coating and increase its reducible fraction [66]. The oxidizable Cr fraction correlates positively with Al content but negatively with pH, consistent with Cu migration and transformation patterns.

The exchangeable fraction of Pb is positively influenced by electrical conductivity (EC), ORP, and Ca levels. Under reducing conditions, Pb forms refractory PbS with sulfides. An increase in ORP promotes lead sulfate formation, disrupting PbS precipitation and transitioning Pb into a dissolved state. Higher Ca and EC levels reflect enhanced bedrock weathering, increasing the exchangeable Pb fraction. The reducible fraction of Pb is negatively correlated with Al, C, N, and the C/N ratio, but positively correlated with Ca and Mg. Al oxides compete with Fe/Mn oxides for Pb²⁺ adsorption, reducing Pb binding on Fe/Mn surfaces. Organic matter (rich in C and N) forms stable organic–Pb complexes (e.g., humic acid–Pb), competing with Fe/Mn oxides and decreasing the exchangeable Pb fraction [67]. A high C/N ratio indicates a reducing environment, slowing organic matter decomposition and indirectly suppressing Fe/Mn oxide stability. The oxidizable fraction of Pb is primarily influenced by Ca. Increased Ca input enhances deprotonation of organic acids, strengthening Pb complexation and increasing Pb binding to organic matter [68].

The exchangeable fraction of Cd is positively influenced by ORP. In reducing environments, Cd primarily exists as CdS; increased ORP may partially oxidize CdS, releasing Cd²⁺ ions. Elevated ORP also promotes organic matter decomposition, further increasing the exchangeable Cd fraction. The exchangeable Cd fraction correlates positively with C/N. Under high C/N conditions, degraded organic substances release aromatic hydrocarbons and carboxylic acids, providing adsorption sites [69,70]. Polyphenolic compounds bind to Fe/Mn oxide surfaces, enhancing Cd²⁺ adsorption [71]. Additionally, iron-reducing bacteria under high C/N conditions produce Fe²⁺, which co-precipitates with Cd, increasing the reducible Cd fraction [72]. The oxidizable Cd fraction correlates positively with C, N, C/N, and Ca, but negatively with TS. Carboxyl groups form organometallic complexes with phenolic hydroxyl groups, while amino groups form stable five-membered ring chelates with Cd. Increased carbon and nitrogen contribute to Cd fixation [73]. Elevated Ca content forms ternary organic–Ca–Cd complexes, strengthening Cd binding and increasing its oxidizable fraction [74]. Under high S conditions, thiol groups exhibit stronger complexing ability than hydroxyl or carboxyl groups, resisting oxidative decomposition.

The occurrence forms of heavy metals are regulated by variations in environmental factors and exhibit distinct differentiation patterns.

4. Conclusions

This study elucidates the intricate interplay between heavy metal dynamics and subtropical peatland ecosystems, revealing distinct mechanistic pathways that govern metal origin, speciation, and environmental mobility. Heavy metal accumulation in the DJH peatland is primarily driven by natural biogeochemical processes, including organic matter decomposition of plant litter, weathering of carbonate substrates, and atmospheric deposition, with anthropogenic inputs playing a negligible role. Metal speciation analysis highlights contrasting environmental behaviors: while Cr, Cu, and Pb are predominantly sequestered in stable residual phases, Cd exhibits pronounced mobility through labile exchangeable fractions. DGT further corroborates phase-specific migration patterns, demonstrating active cadmium mobilization from peat matrices and enhanced lead retention via geochemical binding mechanisms. Cd mobility is predominantly controlled by redox-sensitive sulfide dissolution (enhanced under oxidizing conditions) and C/N-mediated organic degradation, whereas Pb mobility is influenced by calcium-promoted organic binding and sulfide oxidation. For Cu and Cr, speciation balances organic binding (driven by C/N ratios and aluminum bridging) against pH-induced dissociation of organic–metal complexes. Elevated C/N ratios amplify exchangeable states for Cu, Cr, and Cd, and rising pH systematically destabilizes organic–metal associations for Cu and Cr. Importantly, the accelerated organic matter turnover characteristic of subtropical peatlands, exacerbated by warm, humid climates, counteracts traditional metal immobilization pathways, promoting episodic metal release during hydrological perturbations. These findings collectively challenge the applicability of temperate-region metal retention models to subtropical systems and emphasize the need for management strategies that integrate rapid organic cycling, climatic stressors, and geochemical feedback to mitigate long-term ecological risks.