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Article

Potential Release of Phosphorus by Runoff Loss and Stabilization of Arsenic and Cadmium in Mining-Contaminated Soils with Exogenous Phosphate Fertilizers

by
Meng Zhang
1,
Chaoyang Wei
2,
Fen Yang
2,*,
Yujian Lai
1,
Xuemei Wang
1,
Menglu Wang
1,
Wei Han
1,
Xinlian Zhong
1,
Jian Wang
1,
Hongbing Ji
1,* and
Zhiling Guo
3
1
Beijing Key Laboratory of Resource-Oriented Treatment of Industrial Pollutants, School of Energy and Environmental Engineering, University of Science and Technology Beijing, No. 30 Xueyuan Road, Haidian District, Beijing 100083, China
2
Key Laboratory of Land Surface Pattern and Simulation, Institute of Geographic Sciences and Natural Resources Research, Chinese Academy of Sciences, A11 Datun Road, Chaoyang District, Beijing 100101, China
3
School of Geography, Earth & Environmental Science, University of Birmingham, Edgbaston, Birmingham B15 2TT, UK
*
Authors to whom correspondence should be addressed.
Sustainability 2024, 16(22), 9783; https://doi.org/10.3390/su16229783
Submission received: 5 September 2024 / Revised: 3 November 2024 / Accepted: 7 November 2024 / Published: 9 November 2024
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Abstract

:
Phosphate has been proven to be effective in remediating soils contaminated with potentially toxic elements (PTEs); however, the potential release of phosphorus (P) through runoff and the impact on PTEs’ transport in this process have never been assessed. A rainfall simulation study was conducted to investigate P runoff loss and its impact on the stability of arsenic (As) and cadmium (Cd) after applying potassium dihydrogen phosphate (PDP), superphosphate (SSP), and ground phosphate rock (GPR) in soil trays packed with As–Cd-contaminated soil. The phosphorus loss through runoff and sedimentary phases followed the order of SSP > PDP > GPR > control. Phosphate fertilizers’ application reduced the mobility of As and Cd. In the first rainfall, the enrichment ratios (ERs) of As and Cd in the sedimentary phase after PDP, SSP, and GPR treatment were 0.12, 0.04, and 0.08 and 0.24, 0.16, and 0.07 units lower than the control, respectively. The <53 μm fraction in the sedimentary phase accounted for 53.06–75.95%, and phosphate fertilizers significantly enhanced the As and Cd stability in this fraction. The XPS analysis showed that the conversion of As(III) to As(V) and the generation of Cd–phosphate compounds were important reasons for enhancing As and Cd stability. This study demonstrated that PDP might be capable of the remediation of As–Cd contamination with the least release of P to watersheds.

Graphical Abstract

1. Introduction

Soil is an important medium for material cycling and energy exchange between plants and the natural environment. However, the contamination of soil with PTEs has become a global problem [1,2]. Mines are an important source of PTEs in soils [3,4]. Over-exploitation, smelting, and the improper management of mines have resulted in the accumulation of PTEs in soil, which migrate to the surrounding soil, rivers, and even groundwater through a series of processes such as runoff, leaching, and sedimentation, causing environmental pollution and threatening regional environmental ecology [3]. In addition, PTEs also accumulate in human organs and organisms through the food chain, posing a serious threat to human health [3,5,6,7]. Among them, the co-contamination of As and Cd in soil is a particularly serious problem. The anionic metalloid As and the cationic metal Cd show opposite biogeochemical behaviors, which makes the remediation of As and Cd co-contaminated soil challenging [8,9].
Phosphorus is an essential element for plant growth. Applying phosphate fertilizers to the soil can significantly improve the soil nutritional status and have a positive effect on crop growth. At the same time, phosphate fertilizer can also be used as a fixative for PTEs in soil [8] and have received increasing attention due to their high efficiency, environmental friendliness, low hazard risk, and low corrosiveness [10,11]. Studies have shown that the application of phosphate fertilizer is one of the most effective and relatively low-cost methods to stabilize Cd in situ [12,13,14]. Deng et al. [15] applied calcium magnesium phosphate in acidic paddy soil, resulting in lower extractable levels of Cd in the soil and reduced Cd accumulation in all tissues of rice (Oryza sativa). Azzi et al. [16] added superphosphate to column experiments, reducing Cd mobility in soil and accumulation in lettuce (Lactuca sativa). The mechanism of remediation effects of phosphate fertilizers usually involves precipitation and adsorption between Cd and phosphate fertilizers [17]. Therefore, the application of phosphate fertilizers in Cd-contaminated soil–crop systems has become a practical way to reduce Cd extraction from soils and accumulation in crops [13,18,19]. However, unlike the nature of Cd, As and P have similar chemical properties and exhibit a highly competitive relationship between adsorption in soil and surface complexation with iron minerals [20,21]. In soil systems, phosphorus addition increases As desorption, thereby improving its bioavailability [22]. In this case, the potential for As migration in soil is greatly increased, as is the risk of As contamination of the surrounding environment [23].
In addition, long-term phosphorus application can also significantly increase phosphorus levels in arable soils, increasing the risk of phosphorus loss through soil erosion, leaching, or runoff [24,25]. The loss of phosphorus from soil to receiving waters is mediated by hydrological conditions (surface and subsurface flow) in solutions or sediments [26]. Due to the strong adsorption of soil, phosphorus is usually lost in surface runoff to a greater extent than in groundwater flow paths [27,28]. The form in which P is present during this process and the adsorption–desorption behavior have a major impact on the stability of PTEs in soil. Therefore, it is particularly important to explore the runoff loss of P in soil after the application of phosphate fertilizers and its impact on the transport of As and Cd in the process. Many previous studies have focused on exploring the main factors influencing soil P loss, such as the soil type [27,29], tillage practices [24], rainfall intensity [30], P sources, and P forms [31]. Additionally, some research has utilized P as a remediation agent to immobilize and stabilize PTEs in soil. For instance, Mignardi et al. [32] used hydroxyapatite and phosphate rock to significantly reduce the water-soluble Cd, Cu, Pb, and Zn content of soil. Similarly, Jiang et al. [33] used slow-release phosphate materials to simultaneously stabilize As, Cd, and Pb in soil, achieving stabilization efficiencies of water-soluble As, Cd, and Pb up to 99% within seven days. However, there is still a lack of systematic research on the effects of P loss on the migration and transformation of As and Cd in soil and the related mechanisms.
In this study, different phosphate fertilizers such as PDP, SSP, and GPR were applied to the soil of mine-affected agricultural lands, and an artificial rainfall experiment was used to explore the following: (i) the runoff loss of P after the application of different exogenous phosphate fertilizers, and its impact on As and Cd concentration, transport forms, and accumulation in surface runoff from contaminated soil; (ii) the effect of different phosphate fertilizers on the stabilization of As and Cd; (iii) the relevant mechanisms of different phosphate fertilizers on As and Cd runoff losses. The research findings are of significant theoretical and practical importance for the rational use of phosphate fertilizers, improving resource utilization efficiency, reducing the environmental risk of PTEs in soils, and achieving the goal of eco-friendly sustainable development.

2. Materials and Methods

2.1. Soil and Determination of Basic Soil Properties

The experimental soil was collected from the farmland around an abandoned lead–zinc ore smelting plant in Hezhang County, Guizhou Province, China (104° 30′ 53″ E, 26° 58′ 0″ N), where As–Cd composite contamination is the main environmental problem. The checkerboard sampling method was used to collect 12 soil samples in the area, with a sampling depth of 0–20 cm. Soil samples were air-dried, rocks and roots were removed, and samples were gently crushed to pass through a 2 mm nylon fiber sieve for subsequent experiments.
Three phosphate fertilizers were selected for this study including potassium dihydrogen phosphate (PDP), superphosphate (SSP), and ground phosphate rock (GPR). The phosphate fertilizers of PDP and SSP were purchased from Hebei Xianzheng Agricultural Technology Co., Ltd., Baoding, China and GPR was purchased from Taizhou Changpu Chemical Reagent Co., Ltd., Taizhou, China. The three phosphate fertilizers were homogenized and passed through a 2 mm sieve, sealed, and stored for subsequent experiments.

2.2. Experimental Design

The artificial rainfall experiments were carried out in a PVC runoff box of 1 m in length, 20 cm in width, and 7.5 cm in depth, with the back and side walls of the runoff box 2.5 cm above the soil surface and nine 5 mm drainage holes at the bottom [30,34]. The bottom of the box was covered with a piece of cheesecloth. The runoff experiment was divided into four treatments: three phosphate fertilizers, PDP, SSP, and GPR, were added into 13 kg of air-dried soil samples, respectively, so that the exogenous phosphate fertilizer content in the soil was 2 g·kg−1 (calculated as P2O5), mixed well and spread in a runoff box with slight compaction, and air-dried soil without added phosphate fertilizer was used as the control treatment (Control). Urea and potassium sulfate were added at the same time to make the soil N and K2O levels up to 200 and 150 mg·kg−1, respectively (no additional potassium sulfate was added to the PDP group). The soil was watered regularly to ensure that the soil moisture content was 70% of the field capacity but did not cause surface runoff. After 30, 60, and 90 d of cultivation, artificial rainfall experiments were conducted on the four groups of treatments, which were recorded as the 1st, 2nd, and 3rd experiments, respectively.
Calibrations were conducted prior to the artificial rainfall experiment to maintain uniform and intense rainfall. The developed artificial rainfall simulator consisted of a water supply, water reservoir, water pump, pressure gage, sprayers, detachable bracket, wind protection equipment, and sample collectors. The rainfall height was 1.0 m from the ground, and the effective area of rainfall was 2 m × 2 m. The simulated rainfall was similar to the natural rainfall, and the experiment avoided wind interference to ensure accuracy. During the experiment, the rainfall intensity was strong and the duration was short. Therefore, the influence of evaporation during rainfall was not considered. Before each rainfall experiment, the soil in the runoff box was premoistened using a sprayer so that runoff started from already moist conditions. The artificial rainfall experiment lasted for 60 min, and the rainfall intensity was set as 83 mm·h−1 based on the rainfall intensity in Guizhou Province. The slope of the runoff box was 5°.
In the experiment, surface runoff was collected every 3 min in the first 30 min and every 5 min in the last 30 min. After the rainfall, the runoff samples were immediately transported to the laboratory with the accurate volume of each sample recorded, and then filtered through 0.45 μm filter membrane to separate the runoff samples into the runoff phase and the sedimentary phase. The runoff phase was stored at 4 °C under refrigeration after the addition of appropriate amount of nitric acid and was measured as soon as possible. The sedimentary phase samples were dried and weighed, sealed, and stored for subsequent experiments.

2.3. Chemical Analysis

All chemicals used were of analytical reagent grade. The soil pH was measured at a soil-to-water ratio of 1:2.5 (m:V) by using an automated pH meter. The soil was divided into clay (<0.002 mm), silt (0.002–0.02 mm), and sand (>0.02 mm) by the wet sieving method, and the percentage of each particle size was determined. The available P in soils was extracted with 0.5 mol·L−1 NaHCO3 and determined using the molybdenum antimony colorimetric method (700 nm) [8]. The concentrations of As, Cd, and dissolved phosphorus (DP) in the runoff phase were determined by inductively coupled plasma mass spectrometry (ICP-MS, PE DRC-e, PerkinElmer, Waltham, MA, USA). The sedimentary phase was combined in pairs to obtain new sedimentary samples at 6, 12, 18, 24, 30, 40, 50, and 60 min since the content in the experiment was low. To determine the contents of As, Cd, and P in the raw soil and the particulate As, Cd, and P in the sedimentary phase, 0.1 g of dried and ground samples was digested with HNO3–HF–HClO4, and then the content determined by inductively coupled plasma optical emission spectrometry (ICP-OES, PE 5300DV, PerkinElmer, Waltham, MA, USA).
The speciation of As was determined using the Wenzel method [8,35,36], including the non-specifically bound (F1-As), specifically bound (F2-As), non-crystalline hydrous oxide-bound (F3-As), crystalline hydrous oxide-bound (F4-As), and residual form (F5-As). The speciation of Cd was determined using the Tessier five-step extraction method [8,37], including the exchangeable (F1-Cd), carbonate-bound (F2-Cd), Fe-Mn oxide-bound (F3-Cd), organic-bound (F4-Cd), and residual form (F5-Cd). The modified Hedley method [38,39,40] was used to determine the speciation distribution of P in the sedimentary phase, including the adsorbed (A–P), iron/aluminum-bound (Fe/Al–P), calcium-bound (Ca–P), and inert inorganic form (Pi). The sedimentary phase samples in the third rainfall were combined for 0–24 min, 25–40 min, or 41–60 min and wet-sieved through 53 and 250 μm sieves to obtain three particle size fractions, namely <53, 53–250, and >250 μm. The sieved sedimentary phase samples were weighed after drying and grinding, and the contents of As, Cd, and P were determined by ICP-OES in order to analyze the elemental distributions of different particle size fractions treated with different phosphate fertilizers.
Three particle size samples at the initial stage (0–24 min) of runoff under different treatments (Control, PDP, SSP, and GPR) were selected for characterization. The phase structure of the soils was determined by X-ray diffractometer (XRD, Rigaku Ultima IV, Rigaku, Tokyo, Japan) in the range of 10–80° (2 theta). The possible existing forms of surface elements (P, As, and Cd) after different phosphate fertilizers’ application were characterized using X-ray photoelectron spectrometer (XPS, Thermo Scientific K-Alpha, Waltham, MA, USA). All spectra were energy-calibrated using the C1s peak position measured at 284.8 eV.

2.4. Statistical Analysis

The enrichment ratio (ER), the ratio of the concentration of a soil constituent in eroded sediment to that of the original soil, is widely used to reflect the transportation and enrichment of PTEs in sediment particles [41,42,43]. And the ER was calculated as described by Rao et al. [23].
E R T i = ∑ j = 1 i c i · m i ∑ j = 1 i m i / C 0
where C0 is the concentration of PTE in the original soil, Ti represents the time of the ith runoff sample collection, and mi and ci are the mass of the sedimentary phase at time i and the concentration of PTE in this sedimentary phase, respectively.
All data were recorded and organized in Microsoft Excel. The Shapiro–Wilk test (W-test) was used to check the normality of the primary data. The significance level α = 0.05 was set as the threshold for determining whether a difference was significant. If the p-value was less than α, the difference was considered statistically significant, referred to as a “significant difference”. SPSS 19.0 was used to perform one-way ANOVA (with Duncan’s test to determine statistical differences between treatments) and statistical analysis, and OriginPro 2019 software was used to produce figures.

3. Results and Discussion

3.1. Soil Characteristics

Selected physical and chemical properties of the experimental soil are listed in Table 1. The experimental soil was slightly alkaline with a pH of 7.63, and the total phosphorus (TP) content was 1.02 g·kg−1. It contained 49.97% sand, 43.82% silt and 6.2% clay and was classified as a sandy loam according to the American System of Soil Texture Classification (ASSC) standards, which is susceptible to rainfall erosion.
The average contents of As and Cd in the soil were 1429.57 and 43.93 mg·kg−1, which were 11.91 and 14.64 times higher than the soil pollution risk control values (120 mg·kg−1 for As and 3 mg·kg−1 for Cd) stipulated in China’s Soil environmental quality—Risk control standard for soil contamination of agricultural land (GB 15618-2018) [44], respectively. Meanwhile, the As and Cd contents in the soil were 71.48 and 66.56 times higher than the background values of soil As (20 mg·kg−1) and Cd (0.66 mg·kg−1) in Guizhou Province, respectively [45]. Therefore, the soil and rivers surrounding this site are facing the problem of runoff pollution caused by PTEs.

3.2. Effect of Phosphate Fertilizers on P, As, and Cd in the Runoff Phase

The total runoff of each treatment of the control, PDP, SSP, and GPR caused by artificial rainfall was 12.62–13.26, 11.20–12.42, 10.84–12.98, and 10.92–12.72 L·m−2, respectively (Table 2). In the control, the DP losses in the three rainfall experiments were 1.45, 0.02, and 0.56 mg·m−2, respectively. The application of phosphate fertilizers could significantly increase the soil P loss, among which SSP was the most significant (p < 0.05). The three rainfall experiments increased 11.23, 710.5, and 13.59 times, respectively, compared with the control. In addition, the application of SSP also caused the largest losses of As and Cd, which were 1.05–1.69 and 0.06–0.07 mg·m−2, respectively, and 6.47–11.07 and 0.12–1.57 times that of the control during the same period.
Figure 1 shows the variation in the P, As, and Cd contents in the runoff phase with time after the application of different phosphate fertilizers. All three elements were maximum at the initial moment of runoff and decreased with increasing rainfall time, which was similar to the previous study [46,47]. The decline is due to the dilution of P on the surface and the sorption of P translocated into the soil by infiltration [30]. The DP fluctuated at 0–30 min and was basically stable after 30 min. The DP concentration in the runoff phase among the four treatments was SSP > PDP > GPR > Control in descending order, which might be related to the available phosphorus concentration of different phosphate fertilizers. In addition, the P concentration in the runoff phase decreased with the increase in incubation time.
All three types of phosphate fertilizers can increase the As content in runoff, among which SSP has the most obvious effect. In the first experiment, the DP concentration decreased rapidly from 0 to 12 min and was 28.99 μg·L−1 at 12 min. Subsequently, the As concentration fluctuated in the range of 28.61–32.34 μg·L−1 from 12 to 35 min. After 35 min, the As concentration in the runoff phase decreased, and the As concentration was 19.14 μg·L−1 at the end of the runoff. After the application of PDP, the As concentration in the runoff fluctuated slightly, ranging from 9.63 μg·L−1 to 17.21 μg·L−1, which was 5.05 to 9.07 times that of the control. The application of GPR had the smallest impact on As concentration in the runoff phase, which was 1.26–3.02 times that of the control. In addition, with the increase in cultivation time, the concentration of As in the runoff phase decreases (Table S1). The results show that the application of phosphate fertilizers could increase the As concentration in the runoff phase, which could be attributed to the fact that P has similar chemical properties as As, and PO43− competes with AsO43− for adsorption sites in the soil, leading to the release of As [8,48]. Although the concentration of As in the runoff phase increased to varying degrees after the application of phosphate fertilizers, they were all lower than the limit of the Class I water quality standard of Environmental Quality Standards for Surface Water (GB 3838-2002) [49].
The Cd concentration in the runoff phase decreased in the initial stage of rainfall (0–15 min) and then basically plateaued. Compared with the control, the application of SSP and GPR can increase the Cd concentration in the runoff phase, while the application of PDP can reduce the Cd concentration to a certain extent. In the first experiment, the concentration of Cd in the 60 min rainfall samples was 0.93 μg·L−1 with the application of PDP, which was 7.74% lower than that of the control. The application of PDP enhanced the adsorption of Cd by phosphate and the formation of Cd3(PO4)2, which may also contribute to reducing the extractability of Cd in the soil [13]. On the contrary, SSP and GPR are acidic fertilizers that can lead to a decrease in pH, and soil pH is the most important factor affecting Cd extractability [13,50]. With the extension of culture, the Cd concentration in the runoff phase of each treatment decreased significantly (p < 0.05) compared with the first experiment (Table S1). This indicates that the migration of As and Cd in soil with runoff was gradually decreased and stabilized with the increase in incubation time. In addition, the concentration of Cd in the runoff phase in all four treatments was lower than the limit value (5 μg·L−1) of the Class II water quality standard of Environmental Quality Standards for Surface Water (GB 3838-2002). The application of PDP reduced the average Cd concentration in the three experiments to 0.753, 0.727, and 0.634 μg·L−1, respectively, which are all lower than the limit value (1 μg·L−1) of the Class I water quality standard.

3.3. Loss of P and Migration of As and Cd in Sedimentary Phases

3.3.1. Contents of P, As, and Cd in Sedimentary Phases

The 60 min rainfall resulted in sediment amounts of 234.31, 417.43, and 388.00 g·m−2 in the control treatment in three experiments, respectively (Table 2). The sediment yields were similar to those of typical loess soils [51], suggesting that the tested soils were susceptible to erosion. Compared with the runoff phase, the losses of P, As, and Cd in the sedimentary phase accounted for more than 96%, 99%, and 98%, respectively, which also shows that the losses of soil P, As, and Cd are mainly particle losses [27,52,53].
The curves of changes in the As, Cd, and P contents in the sedimentary phase with runoff time are shown in Figure 2. In the three experiments, the application of phosphate fertilizers can significantly increase the P content in the sedimentary phase. Among them, SSP has the most obvious promotion effect on P in the sedimentary phase, followed by PDP and GPR. Overall, the particulate P content fluctuates little with the extension of rainfall duration, which is consistent with Ma et al. [27].
In the first experiment, the As and Cd contents increased significantly in the initial stage (6–12 min) of rainfall, and gradually decreased in the subsequent 12–30 min. This reaction was similar to the study by Shi and Schulin [46]. The difference is that the As and Cd contents in the sedimentary phase increased at 6–12 min, which may be related to the fact that the soil is not yet saturated and the production volume is low in the early stage of the reaction. In the later stage of rainfall (30–60 min), unlike the study by Shi and Schulin [46], the contents of As and Cd in the sedimentary phase increased to varying degrees. Studies have shown that fine particles are more likely to be carried away than coarse particles because they have lower settling velocities. As the rainfall time increases, clay and silty particles in the sediment become more abundant, and the size of the output sediment also increases [23,41]. In this study, the >250 μm fraction soil was enriched with more PTEs elements (See Section 3.4), which led to an increase in As and Cd contents in the sedimentary phase during the late rainfall period with rainfall time.
The effects of the application of the three phosphate fertilizers on the treatment of As and Cd in the sedimentary phase varied greatly. Compared with the control, all three phosphate fertilizers could increase the As content in the sedimentary phase. The application of SSP had the most significant cumulative effect on As content in sediments, followed by GPR, and PDP had the smallest effect on As in the sedimentary phase. In the first experiment, when it rained for 60 min, the As content in the sediment phase after applying SSP, GRP, and PDP was 1.19, 1.11, and 1.05 times that of the control, respectively. Unlike As, the application of three kinds of phosphate fertilizers can reduce the Cd content in the sedimentary phase to a certain extent. Among them, the application of PDP had the most obvious effect on the reduction in Cd content in the sedimentary phase. After 60 min of rainfall in the three experiments, the application of PDP reduced the Cd content in the sedimentary phase by 10.58%, 17.79%, and 15.89%, respectively, compared with the control, and this is consistent with the study of Yan et al. [54]. The application of PDP can lead to the maximum release of available P in the soil in a short period of time, which can be easily retained by soil particles through physical attraction, chemical adsorption, and cation bridging. This process binds PO43- in the soil solution to mineral surface species by forming partial covalent bonds at hydroxyl sites associated with Fe and Al cations at hydroxyl sites [27]. The major initial reaction product of monocalcium phosphate (MCP), which is the main P component of SSP, is dicalcium phosphate dihydrate (DCPD) in the experiment soil because of the pH. Because DCPD is more soluble than other Ca–P compounds, SSP is also more effective in stabilizing Cd in calcareous soils [55,56]. In contrast, GPR has the lowest apparent solubility [57], and its effectiveness largely depends on the soil pH. Its effect is significantly reduced when applied to slightly alkaline soils. It is worth mentioning that the Cd content in the sedimentary phase decreased with the increase in incubation time, regardless of whether phosphate was applied or not. At 60 min of rainfall, the Cd content in the sedimentary phase of the four treatments in the third experiment decreased by 20.87%, 26.74%, 22.39%, and 23.59%, respectively, compared with the first experiment.

3.3.2. The ERs of As and Cd in Sedimentary Phases

The ERs of PTEs in the sedimentary phase with rainfall time were calculated and the results are shown in Figure 3. The application of phosphate fertilizer reduced the risk of As migration, and at the end of the first rainfall, the ER values under the treatments of PDP, SSP, and GPR were reduced by 0.12, 0.04, and 0.08 units, respectively, compared with the control. At the reaction time of 6–18 min, the ER of Control-As was greater than the enrichment critical value of 1.0, which indicates a potential risk of As enrichment at this time [23,43]. As the culture time increased, the Control-ER decreased significantly. At 60 min of rainfall in the third experiment, the Control-ER was 0.71, which was 0.21 units lower than the same time of the first rainfall. However, the ERs for the application of the three phosphate fertilizers, PDP, SSP, and GPR, increased slightly in the second rainfall (0.77–0.82, 0.88–0.89, and 0.83–0.86, respectively) and were lower than that of the first rainfall in the third rainfall (0.73–0.78, 0.75–0.86, and 0.7–0.81, respectively). This suggests that the application of phosphate fertilizers has the potential risk of increasing the migration of As in soil, but since the overall ERs were all less than the critical value of 1.0, the migration and enrichment of As by phosphate fertilizer was under control.
Different from As in soil, the ERs of Cd in the control, SSP, and GPR treatments were all greater than the enrichment threshold value of 1.0 in the first experiment, indicating that there was a potential risk of migration and enrichment of Cd in soil at this time. The application of phosphate fertilizers can reduce the risk of Cd migration and accumulation. At the end of the first rainfall, the application of PDP, SSP, and GPR reduced the risk of Cd by 0.24, 0.16, and 0.07 units, respectively, compared with the control. The potential risk of Cd migration and enrichment decreased with increasing incubation time. During the second rainfall, only the ER at 6 min of GPR is greater than 1.0, and the rest have no risk of migration and enrichment. At the end of the third rainfall, the ERs of the four treatments of the control, PDP, SSP, and GPR were 0.90, 0.73, 0.82, and 0.86 respectively, which were 0.33, 0.21, 0.25, and 0.30 units lower than at the end of the first rainfall. Overall, the ER of As and Cd is not only related to the type of fertilization, but also closely related to the duration and frequency of runoff. Proffitt et al. [58] found that at the onset of surface runoff, eroded sediments composed mainly of clay and silt particles became coarser and coarser over time until they approached equilibrium. It may be that rainfall creates a loose layer of coarse particles on the soil surface, which shields the fine particles below, making them less likely to be entrained by surface runoff [41,59]. Therefore, the relationship between ER and runoff time and frequency can be summed up as the relationship between particle size. Shi et al. [41] showed that in the final stage of rainfall, the ER of large-sized (>250 μm) sediments is usually higher than that of intermediate-sized portions, resulting in a bimodal (i.e., 0–20 μm and 250–1000 μm) enrichment pattern. This is consistent with our research.

3.4. Particle Size Distribution of Sedimentary Phase and Relationship Between P, As, and Cd Stability

In order to explore the factors affecting the application of different phosphate fertilizers on soil As and Cd runoff, the samples of the sedimentary phase of the third runoff were sieved for particle size, and the distribution diagram is shown in Figure S1. The sedimentary phase soils were mainly dominated by fine particles (<53 μm), followed by coarse particles (>250 μm), and the medium particles (53–250 μm) fraction was the least. The three particle sizes accounting for 53.06–75.95%, 8.83–27.44%, and 10.84–19.55% of the total sedimentary phase, respectively.
Figure 4 shows the contents of P and As and Cd in each particle size in the initial, middle, and final rainfall. In the Control treatment (Figure 4a), the contents of P in the three fractions <53, 53–250, and >250 μm were 690.14–807.71, 1218.13–1280.84, and 1177.64–1232.00 mg·kg−1, respectively. After applying three kinds of phosphate fertilizers, the P content in the three fractions of <53, 53–250 and >250 μm all increased. The most significant increase was in SSP (Figure 4c), which increased by 2.44–3.48, 1.91–2.01, and 1.79–2.01 folds, respectively, compared with that of the control. Followed by PDP, and the least increase in GPR. This is consistent with the runoff pattern of P in the runoff phase (Figure 1) and sedimentary phase (Figure 2). This indicates that SSP is easily transported with rainfall while GPR is more stable after application.
The distribution of As in each particle size in the sedimentary phase was similar to that of P. The content was higher in the >250 and 53–250 μm fractions, while the distribution in the <53 μm fraction was low, with the content being 1604.40–1945.56, 1705.00–1829.81, and 37.76–869.18 mg·kg−1, respectively. After the application of phosphate fertilizers, the As content in each particle size increased, and the SSP group was also the highest. The As contents in the <53, 53–250, and >250 μm fraction were 1282.87–1735.23, 1817.13–2040.16, and 1707.41–2291.19 mg·kg−1, respectively, which were 25.65–86.39%, 0.41–12.94%, and 4.11–33.55% higher than that of the control, respectively. This was followed by GPR, and PDP had little effect on the runoff of As of various particle sizes, which is consistent with the runoff results of As in the sedimentary phase (Figure 2). In general, the effect of phosphate fertilizer on As in the sedimentary phase is mainly achieved by affecting the As content in the <53 μm fraction.
The effects of the three phosphate fertilizers on Cd varied greatly. After the application of PDP, the Cd contents of the <53, 53–250, and >250 μm fractions were 26.65–27.26, 50.72–53.97, and 59.37–61.61 mg·kg−1, respectively, which were 6.20–24.10%, 3.00–4.69%, and 7.10–12.00% lower than that of the control, respectively. Different from PDP, after the application of SSP, the Cd content in the >250 μm fraction decreased by 6.97–24.41% compared with the control, while it increased in the <53 μm and 53–250 μm fractions. The changes in Cd content in different particle sizes after GPR application were similar to those of SSP, with increases of 2.46–15.76% and 0.34–9.12% in the <53 μm and 53–250 μm fractions, respectively, and decreased of 14.04–23.59% in the >250 μm fraction. The P, As, and Cd in the sedimentary phase are mainly concentrated in the >250 μm fraction, followed by the 53–250 μm fraction, with the least enrichment in the <53 μm fraction, which is different from previous studies [46]. In general, many chemicals, including PTEs such as As and Cd, are mainly attached to clay particles with a large specific surface area [23]. However, when coarse sand contains coarse minerals or heavy minerals with a strong ability to hold PTEs, the content of PTEs in the coarse particle size will also increase [60]. The experimental soil was taken from the vicinity of the mining area. Anthropogenic activities such as mining and smelting, as well as natural activities such as ore weathering, are the main sources of soil pollution in this area, which has led to the concentration of PTEs in the coarse fraction.

3.5. Speciation Distribution of Elements in Sedimentary Phase and the Effect of P on the Stability of As and Cd

In order to further explore the impact of phosphate fertilizers application on the migration of As and Cd, speciation extraction of P, As, and Cd in different particle sizes of the third experiment was carried out, and the results are shown in Figure 5. In the control treatment, P was mainly in the form of Pi, Fe/Al–P was the second most abundant form after Pi, and as the most active form of phosphorus, A–P accounted for the highest proportion in <53 μm fraction, accounting for 5.37–9.12% of the total soil phosphorus, which was 1.25–2.33 and 1.11–1.82 times higher than that in the 53–250 and >250 μm fractions, respectively. This shows that compared with the <53 μm fraction, although the content of P in the coarse fraction is higher (Figure 4a), it is more stable and not easily absorbed and utilized by the plant because it exists in the form of Pi and Ca–P.
The distribution of P in the soil changed significantly after the application of the three phosphate fertilizers, in which A–P increased substantially, with SSP being the most significant. The percentage of A–P in the <53, 53–250, and >250 μm fractions was 3.75–5.17, 5.19–6.99, and 5.02–6.05 times higher than that of the control, respectively. Followed by PDP, the proportions of A–P in the three particle sizes are 2.82–4.00, 4.17–5.61, and 4.11–4.47 times that of the control, respectively. The proportion of GPR increase was the smallest but still 1.75–2.98, 3.22–5.17, and 2.39–2.88 times greater than that of the control. This is consistent with the relationship with the amount of DP in the runoff phase (Figure 1). The application of phosphate fertilizer significantly reduced the distribution of Pi, and the reduction in SSP was the most significant, with the three fractions of <53, 53–250, and >250 μm reduced by 56.30–67.28%, 58.55–67.41%, and 52.14–60.16%, respectively, compared with that of the control. This was followed by PDP, and the GPR dropped the least, being 23.83–29.08%, 39.85–45.56%, and 25.61–30.48% lower than the control, respectively. Phosphate fertilizer application significantly increased the A–P ratio in the soil and decreased the stability of P, which in turn led to the loss of P in the sedimentary phase. The analysis of XPS with three particle sizes of the sedimentary phase in the initial stage of rainfall after phosphate fertilizers application is shown in Figure 6. In the control treatment, the characteristic peaks were not shown due to the low soil P content. In addition, all the phosphorus fertilizer treatments had two sets of peaks, P2p1/2 and P2p3/2, with binding energies of 133.91–134.38 eV and 133.04–133.51 eV, respectively, which were attributed to HPO42− and PO43− [61,62]. The binding energy of P2p1/2 and P2p3/2 in the particle size of <53 μm is increased, which may be due to the stronger activity of P in the fine particle size and its combination with PTEs [61].
There are large differences in the speciation distribution of soil As. In the control, As mainly existed as F5-As in the 53–250 and >250 μm fractions, accounting for 64.55–65.70% and 61.87–67.37% of the total As, respectively. The distributions of F4-As, F3-As, and F2-As decreased in sequence. F1-As, the most migratory component of the As form, accounted for only 0.02% of the total amount of As in each particle size. Different from this, the F4-As form was dominant in the <53 μm fraction, accounting for 56.72–62.20% of the total amount of As in the fine particle size. This was followed by F3-As, and the F5-As content only accounts for 7.19–15.44% of the total As. After the application of three types of phosphate fertilizers, the distribution of As in the <53 μm fraction changed the most significantly. The proportion of F5-As increased after applying PDP and SSP, increasing 1.29–3.88 and 1.08–3.52 times, respectively, compared with the control, while the contents of F4-As and F3-As decreased to varying degrees. The application of GPR also showed an increase in F4-As occupancy and a decrease in F3-As, although there was no significant change in the F5-As form. This shows that the application of phosphate fertilizer promotes the transformation of the As form from an unstable to a stable state, and the reaction at the fine particle level is the most significant. Research shows that the phosphate precipitation formed in the iron–aluminum (hydr)oxide system will lead to the immobilization of As, and the generated minerals such as FeOOH and FePO4 can form an As–P–Fe structure and control the concentration of As in the solution [20,63]. In addition, the concentration of As(V) in the soil solution decreases with the increase in Ca2+ concentration. This is because the Ca2+ adsorbed by soil particles promotes the fixation of As(V) on soil oxides. In this study, the application of phosphate fertilizers changed the ratio of Fe/Al–P and Ca–P in the soil, thereby increasing the adsorption of As. In addition, correlation analysis of different forms of P, As, and Cd in three particle sizes (Figure S2) showed that F5-As had a significant correlation (p < 0.05) with A–P, Fe/Al–P, and Ca–P in the <53 μm fraction, whereas there was no such correlation in the other two fractions. The XPS analysis of As (Figure 7) showed two characteristic peaks at binding energies of 45.20–46.63 eV and 50.07–50.17 eV in the control treatment, which represent As(III) and As(V), respectively [48]. The percentages of As(III) and As(V) in the three particle sizes <53 μm, 53–250 μm, and >250 μm were 40.05 and 59.95%, 36.61 and 63.39%, and 37.65 and 62.35%, respectively. After phosphate fertilizers were applied, the proportion of As(III) in the particle size of <53 μm decreased, while As(V) increased. In the PDP, SSP, and GPR treatments, the percentage of As(III) was 32.49, 35.86, and 37.88%, respectively, which were 18.88, 10.46, and 5.42% lower than the control. As(III) is difficult to adsorb, and its biological toxicity is higher than that of As(V). The oxidation of As(III) into As(V) can significantly reduce the biotoxicity and mobility of As in the environment [48], which is another important reason for the enhancement of soil As stability after phosphate fertilizer application.
In the control, the distribution of soil Cd in each particle size was similar to that of As, which was also dominated by the F5-Cd in the 53–250 and >250 μm fractions, accounting for 45.46–49.44% and 43.59–47.84% of the total, respectively. This was followed by F3-Cd, and the contents of F2-Cd, F1-Cd, and F4-Cd decreased in descending order. Among them, the F1-Cd content accounted for 6.84–7.67% and 7.78–7.98%, respectively. In the <53 μm fraction, F3-Cd was the predominant form, accounting for 32.63–43.62% of the total Cd, while F5-Cd accounted for only 11.18–20.77%, which was slightly higher than the F1-Cd form (11.16–12.37%). The application of the three phosphate fertilizers, PDP, SSP, and GPR, resulted in a significant increase in F5-Cd content in the <53 μm fraction, which increased 1.01–2.87, 1.74–4.48, and 2.46–5.31 times, respectively, compared to the control, while the remaining four forms were lower than the control. Cadmium can be passivated by phosphates, and the mechanisms by which phosphates immobilize Cd include interactions such as ion exchange, surface complexation, dissolution, and precipitation [13,64]. Overall, the application of phosphate can transform Cd in the fine-grained fraction into a stable state, while the effect on the medium-grained and coarse-grained fractions is not significant. This may be because, on the one hand, clay particles are dominant in the fine fraction, which has a larger surface area and more easily adsorbs chemical substances, and then a series of adsorption and desorption reactions occur. On the other hand, the existing forms of elements such as P and PTEs in the fine fraction are unstable, and their biological effectiveness is stronger, which makes them more prone to transformation. The XPS spectra of Cd (Figure 7) under different phosphate treatments showed that there was only one characteristic peak in the particle size of <53 μm after SSP application, and there were two groups of characteristic peaks at 406.99 ± 0.5 eV and 413.79 ± 0.5 eV in the other samples, which could be attributed to the existence of cadmium carbonate and cadmium phosphate compounds [61,65]. In the control treatment, the percentages of two groups of peaks with particle sizes of <53 μm, 53–250 μm, and >250 μm were 7.71% and 92.29%, 11.49% and 88.51%, and 9.88% and 90.12%, respectively. The content of cadmium phosphate compounds in the particle size of <53 μm increased after applying phosphate fertilizers, in which the content of cadmium phosphate compounds in the SSP treatment reached 100%. The existence of cadmium phosphate compounds was also found in XRD analysis (Figure S3), which showed that phosphate fertilizer application can form insoluble cadmium salts with Cd through precipitation or co-precipitation reactions, which enhanced stability.
In summary, the application of phosphorus fertilizer in As–Cd co-contaminated soil can not only increase the stability of As and Cd, but also enhance the soil fertility. In this study, only the effect of artificial rainfall on As and Cd runoff after indoor application of phosphorus fertilizer was investigated, and it is necessary to carry out field experiments with a long-term span in the future.

4. Conclusions

In this paper, the runoff loss of soil P and its influence on the stability of As and Cd after applying different phosphate fertilizers were studied through artificial rainfall experiments. The experimental results show that the vast majority of the runoff loss of P, As, and Cd in the soil exists in the particulate state, and although the application of phosphate fertilizers will increase the risk of the loss of As and Cd in the runoff phase, they are lower than the limits of the Class I and Class II water quality standards of the Environmental Quality Standards for Surface Water, respectively. In this study, the runoff sedimentary phase is dominated by the <53 μm fraction, which had a higher bioavailability of As and Cd, although this fraction had lower levels of As and Cd. The application of PDP and SSP significantly reduced the bioavailability of As and Cd and increased their stability in this particle size, but SSP also increased the risk of P loss. The application of GPR increased Cd stabilization but increased the risk of As migration. Therefore, PDP application was found to be effective in reducing P loss from experimental soils and enhancing soil As and Cd stability. The research results are of great practical significance and theoretical value for the remediation of PTEs pollution in mines, the rational application of soil phosphate fertilizers, and the sustainable utilization of phosphate rock resources. Future research could further explore the relevant mechanism of different soil texture, rainfall intensity, and slope on the influence of phosphate fertilizers on the runoff erosion of As and Cd.

Supplementary Materials

The following supporting information can be downloaded at: https://www.mdpi.com/article/10.3390/su16229783/s1, Table S1. Concentrations of As and Cd in the runoff phase in three rainfall experiments. Figure S1. Particle size distribution in the sedimentary phase in the third runoff experiment. Figure S2. Correlation analysis of P, As, and Cd forms in the (a) <53 μm, (b) 53–250 μm, and (c) >250 μm fractions. Figure S3. XRD patterns of sedimentary phases with particle sizes of <53, 53–250, and >250 μm after control and SSP treatment.

Author Contributions

Conceptualization, H.J., F.Y.; methodology, M.Z., C.W., F.Y., Y.L., X.W., H.J.; software, M.W., W.H.; investigation, M.Z., Y.L., X.Z., J.W.; data curation, M.Z.; writing—original draft preparation, M.Z.; writing—review and editing, C.W., H.J., Z.G.; supervision, H.J.; funding acquisition, C.W., X.W., H.J. All authors have read and agreed to the published version of the manuscript.

Funding

This work was supported by the National Natural Science Foundation of China (NSFC) grants (Nos. 41473122, 41173113), the Ministry of Science and Technology of the People’s Republic of China (MOST) and the Fundamental Research Funds for the Central Universities (FRF-TP-19-020A1), Strategic Priority Research Program of Chinese Academy of Sciences (XDA23010401), and the Hundred Talents Program of the Chinese Academy of Sciences.

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Data Availability Statement

The data that support the findings of this study are available from the corresponding author upon reasonable request.

Acknowledgments

Thanks to Lingqing Wang from Institute of Geographic Sciences and Natural Resources Research, Chinese Academy of Sciences, for his guidance on runoff devices.

Conflicts of Interest

The authors declare no conflicts of interest.

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Figure 1. Variation curves of P (a–c), As (d–f), and Cd (g–i) concentrations in the runoff phase with rainfall duration. (PDP—potassium dihydrogen phosphate, SSP—superphosphate, GPR—ground phosphate rock). The points in the figure are average values, and the error bar represents the standard deviation (SD) (n = 3).
Figure 1. Variation curves of P (a–c), As (d–f), and Cd (g–i) concentrations in the runoff phase with rainfall duration. (PDP—potassium dihydrogen phosphate, SSP—superphosphate, GPR—ground phosphate rock). The points in the figure are average values, and the error bar represents the standard deviation (SD) (n = 3).
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Figure 2. Variation curves of P (a–c), As (d–f), and Cd (g–i) concentrations in the sedimentary phase with rainfall time (PDP—potassium dihydrogen phosphate, SSP—superphosphate, GPR—ground phosphate rock). The points in the figure are average values, and the error bar represents the standard deviation (SD) (n = 3).
Figure 2. Variation curves of P (a–c), As (d–f), and Cd (g–i) concentrations in the sedimentary phase with rainfall time (PDP—potassium dihydrogen phosphate, SSP—superphosphate, GPR—ground phosphate rock). The points in the figure are average values, and the error bar represents the standard deviation (SD) (n = 3).
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Figure 3. Variation curves of As (a–c) and Cd (d–f) enrichment ratio in sedimentary phase with rainfall time. (PDP—potassium dihydrogen phosphate, SSP—superphosphate, GPR—ground phosphate rock) (The points in the figure are average values, and the error bar represents the standard deviation (SD) (n = 3)).
Figure 3. Variation curves of As (a–c) and Cd (d–f) enrichment ratio in sedimentary phase with rainfall time. (PDP—potassium dihydrogen phosphate, SSP—superphosphate, GPR—ground phosphate rock) (The points in the figure are average values, and the error bar represents the standard deviation (SD) (n = 3)).
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Figure 4. P (a–d), As (e–h), and Cd (i–l) content of each particle size at different periods in the third runoff experiment (I = initial: 0–24 min; M = middle: 25–40 min; F = final: 41–60 min, PDP—potassium dihydrogen phosphate, SSP—superphosphate, GPR—ground phosphate rock). The points in the figure are average values, and the error bar represents the standard deviation (SD) (n = 3). Different letters indicate significant differences among the treatments at p < 0.05 according to Duncan’s test.
Figure 4. P (a–d), As (e–h), and Cd (i–l) content of each particle size at different periods in the third runoff experiment (I = initial: 0–24 min; M = middle: 25–40 min; F = final: 41–60 min, PDP—potassium dihydrogen phosphate, SSP—superphosphate, GPR—ground phosphate rock). The points in the figure are average values, and the error bar represents the standard deviation (SD) (n = 3). Different letters indicate significant differences among the treatments at p < 0.05 according to Duncan’s test.
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Figure 5. Speciation of P (a–d), As (e–h), and Cd (i–l) in different particle sizes during different rainfall periods in the third experiment (I = initial: 0–24 min; M = middle: 25–40 min; F = final: 41–60 min).
Figure 5. Speciation of P (a–d), As (e–h), and Cd (i–l) in different particle sizes during different rainfall periods in the third experiment (I = initial: 0–24 min; M = middle: 25–40 min; F = final: 41–60 min).
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Figure 6. XPS spectra of P with particle sizes of <53 μm, 53–250 μm, and >250 μm in the sedimentary phase at the initial stage (0–24 min) of rainfall under different phosphate fertilizers treatments: (a,e) Control, (b,f) PDP, (c,g) SSP, (d,h) GPR (PDP—potassium dihydrogen phosphate, SSP—superphosphate, GPR—ground phosphate rock).
Figure 6. XPS spectra of P with particle sizes of <53 μm, 53–250 μm, and >250 μm in the sedimentary phase at the initial stage (0–24 min) of rainfall under different phosphate fertilizers treatments: (a,e) Control, (b,f) PDP, (c,g) SSP, (d,h) GPR (PDP—potassium dihydrogen phosphate, SSP—superphosphate, GPR—ground phosphate rock).
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Figure 7. XPS spectra of As and Cd with particle sizes of <53 μm, 53–250 μm, and >250 μm in the sedimentary phase at the initial stage (0–24 min) of rainfall under different phosphate fertilizer treatments ((a,e) Control, (b,f) PDP, (c,g) SSP, (d,h) GPR) (PDP—potassium dihydrogen phosphate, SSP—superphosphate, GPR—ground phosphate rock).
Figure 7. XPS spectra of As and Cd with particle sizes of <53 μm, 53–250 μm, and >250 μm in the sedimentary phase at the initial stage (0–24 min) of rainfall under different phosphate fertilizer treatments ((a,e) Control, (b,f) PDP, (c,g) SSP, (d,h) GPR) (PDP—potassium dihydrogen phosphate, SSP—superphosphate, GPR—ground phosphate rock).
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Table 1. Selected physicochemical properties of the soil.
Table 1. Selected physicochemical properties of the soil.
PropertiespH (H2O)Moisture content (%)Clay (<0.002 mm) (%)Silt (0.002–0.02 mm) (%)Sand (>0.02 mm) (%)Total P (g·kg−1)Available P (mg·kg−1)Total As (mg·kg−1)Total Cd (mg·kg−1)
Value7.63 (±0.03)25.19 (±3.28)6.243.8249.971.02
(±0.03)		24.98
(±0.17)		1429.57 (±57.23)43.93 (±2.36)
Table 2. Cumulative runoff, sediment yield, and loss of As, Cd, and P under different phosphate fertilizer treatments under simulated rainfall conditions.
Table 2. Cumulative runoff, sediment yield, and loss of As, Cd, and P under different phosphate fertilizer treatments under simulated rainfall conditions.
TreatmentsControlPDPSSPGPR
Rainfall Times1st2nd3rd1st2nd3rd1st2nd3rd1st2nd3rd
Cumulative runoff (L)13.26 a12.61 abc13.20 a11.20 bc11.41 abc12.42 abc12.98 ab11.17 bc10.84 c10.92 c12.22 abc12.72 abc
Dissolved P (mg·m−2)1.45 g0.02 h0.56 h6.46 d5.72 e6.40 de17.73 a14.23 b8.17 c2.57 f1.33 g1.88 fg
Water–As (mg·m−2)0.14 f0.17 ef0.12 f0.77 d0.69 d0.70 d1.69 a1.27 b1.05 c0.25 e0.27 e0.19 ef
Water–Cd (mg·m−2)0.07 cd0.07 cd0.06 d0.04 e0.04 e0.04 e0.18 a0.11 b0.06 d0.08 c0.08 c0.07 cd
Sediment yield (g·m−2)234.31 d417.43 b388.00 b232.62 d320.49 c226.82 de238.25 d250.74 d166.57 f196.20 ef498.96 a88.99 g
Particulate P
(mg·m−2)		181.65 g373.07 d351.19 de323.91 e507.40 b376.54 d451.20 c496.09 b332.80 e223.36 f565.74 a114.36 h
Particulate As
(mg·m−2)		271.56 f477.11 b396.20 c258.74 fg365.58 d247.22 g298.51 e315.32 e195.26 h242.51 g609.27 a99.84 i
Particulate Cd
(mg·m−2)		11.81 d17.22 b15.47 c10.10 ef11.25 d7.21 g11.10 de9.58 f6.02 h9.91 f20.56 a3.44 i
The data in the table are average values (n = 3). Different letters indicate significant differences among the treatments at p < 0.05 according to Duncan’s test.
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Zhang, M.; Wei, C.; Yang, F.; Lai, Y.; Wang, X.; Wang, M.; Han, W.; Zhong, X.; Wang, J.; Ji, H.; et al. Potential Release of Phosphorus by Runoff Loss and Stabilization of Arsenic and Cadmium in Mining-Contaminated Soils with Exogenous Phosphate Fertilizers. Sustainability 2024, 16, 9783. https://doi.org/10.3390/su16229783

AMA Style

Zhang M, Wei C, Yang F, Lai Y, Wang X, Wang M, Han W, Zhong X, Wang J, Ji H, et al. Potential Release of Phosphorus by Runoff Loss and Stabilization of Arsenic and Cadmium in Mining-Contaminated Soils with Exogenous Phosphate Fertilizers. Sustainability. 2024; 16(22):9783. https://doi.org/10.3390/su16229783

Chicago/Turabian Style

Zhang, Meng, Chaoyang Wei, Fen Yang, Yujian Lai, Xuemei Wang, Menglu Wang, Wei Han, Xinlian Zhong, Jian Wang, Hongbing Ji, and et al. 2024. "Potential Release of Phosphorus by Runoff Loss and Stabilization of Arsenic and Cadmium in Mining-Contaminated Soils with Exogenous Phosphate Fertilizers" Sustainability 16, no. 22: 9783. https://doi.org/10.3390/su16229783

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