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Article

Passivation Remediation of Cd-Contaminated Farmland in Yongkang, China by CaAl-LDH: A Mechanism and Application Study

1
College of Environmental Science and Engineering, Yangzhou University, Yangzhou 225009, China
2
Zhejiang Institute of Geosciences, Hangzhou 310023, China
3
State Key Laboratory of Soil Pollution Control and Safety, Chinese Academy of Environmental Planning, Beijing 100041, China
4
Innovation Center for Soil Remediation and Restoration Technologies, College of Environment and Safety Engineering, Fuzhou University, Fuzhou 350108, China
5
Jiangsu Longchang Chemical Co., Ltd., Rugao 226532, China
6
Center for Environmental Science in Saitama, Kazo 347-0115, Japan
7
School of Ecological Environment and Urban Construction, Fujian University of Technology, Fuzhou 350118, China
*
Authors to whom correspondence should be addressed.
These authors contributed equally to this work.
Agronomy 2025, 15(10), 2354; https://doi.org/10.3390/agronomy15102354
Submission received: 15 July 2025 / Revised: 2 October 2025 / Accepted: 7 October 2025 / Published: 7 October 2025
(This article belongs to the Special Issue Heavy Metal Pollution and Prevention in Agricultural Soils)

Abstract

The enrichment of cadmium (Cd) in farmland soil poses serious risks to agricultural safety and remains challenging to remediate. This study evaluated CaAl-layered double hydroxide (CaAl-LDH) as a highly efficient and stable passivator for Cd-contaminated soil. Laboratory adsorption tests demonstrated that Cd2+ adsorption on CaAl-LDH followed pseudo-second-order kinetics and the Langmuir model, indicating monolayer chemisorption, with a maximum capacity of 469.48 mg·g−1 at pH 6. The adsorption mechanisms include surface complexation, interlayer anion exchange, dissolution–precipitation, and isomorphic substitution. A three-year field trial in Yongkang City, China showed that CaAl-LDH promoted the transformation of Cd in rhizosphere soil from the ion exchange state (F2) to the residual state (F7) and Fe–Mn oxidized state (F5), reducing the exchangeable Cd content by 26.71%. Consequently, Cd content in rice grains decreased by 68.42% in the first year and remained over 37% lower in the second year, consistently below the national food safety limit. Future research should focus on the optimization of material’s stability and application protocol. The results demonstrate that CaAl-LDH provides a cost-effective and sustainable strategy for the in situ passivation remediation of Cd-contaminated farmland, contributing to food safety and sustainable agriculture.

1. Introduction

The heavy metal cadmium (Cd) poses a serious threat to the safety of soil [1,2]. This is a pervasive global issue, evidenced by a recent study documenting over 1.16 million contaminated sites across 131 countries, underscoring the widespread nature of Cd pollution in soil [3,4]. The situation is particularly pressing in China, the largest developing country, where more than 280,000 hectares of farmland are susceptible to Cd pollution. This contamination leads to the annual production of over 1.5 million tons of Cd-contaminated agricultural products [5], which not only signifies substantial economic losses but also poses a direct and severe risk to food security and public health. Presently, strategies for remediating and controlling Cd-contaminated soil primarily include electrokinetic remediation [6], phytoremediation [7], and leaching technology [8]. However, each of these remediation technologies exhibits distinct advantages and limitations in specific applications. The in situ passivation method offers significant advantages for remediating Cd-contaminated soil, which typically include low biological toxicity, simple operation, low cost, and strong practicality. This method aids in the safe utilization of soil by reducing the activity of Cd in the environment. Although traditional passivation amendments such as lime, phosphate rock, and biochar are widely employed and operate through well-documented mechanisms including surface complexation, co-precipitation, and solid solution formation [9,10,11], their ability to passivate Cd remains vulnerable to environment fluctuations, particularly soil acidification. In the remediation practices of diverse soil types, these materials often demonstrate lower efficiency and stability than anticipated. Consequently, repeated applications are frequently necessary, leading to risks of secondary pollution and deterioration of soil quality [12].
Layered double hydroxides (LDHs) represent a broad class of both natural and synthetic layered materials [13,14]. LDHs are characterized by the general formula [M2+1−xM3+x (OH)2]x+ [(An−)x/n·mH2O]x−, where M2+ and M3+ denote divalent and trivalent cations, respectively, while A represents an n-valent interlayer anion [15,16]. Recently, advancements in preparation technology have diversified the types of metal elements in LDHs’ main layers, evolving their elemental valence states from solely divalent and trivalent to a polyvalent combination [17]. LDHs’ unique structure and properties offer potential for the adsorption and precipitation of heavy metal ions [18,19]. Numerous studies indicate that LDHs employ various mechanisms to remove metal ions from the environment: (1) as hydroxides, LDHs create an alkaline environment that facilitates the chemical precipitation of metal ions [20]; (2) coordination between heavy metal ions and LDHs’ surface hydroxyl groups for adsorption [21,22]; (3) isomorphic substitution of metal ions in LDHs by heavy metal ions [23,24]. For instance, in LDHs’ layered structure, M2+ can be replaced by environmental ions such as Cd2+, Cu2+, Ni2+, Zn2+, etc., whereas M3+ may be substituted by Cr3+; (4) Introducing special functional group ligands into LDHs’ interlayer structure, like citric acid, malic acid, EDTA, and other organic acids, enables these ligands to chelate with heavy metal ions, facilitating adsorption [25].
Layered double hydroxides (LDHs) have garnered extensive study and application in the field of environmental remediation [26,27]. Especially for cationic heavy metals. Scholars have synthesized an expanded graphite/MgAl-LDH composite with a Mg/Al molar ratio of 3:1, which achieved a high Cu2+ removal efficiency of 99.89% when the pH value reached 6 [28]. This process’s adsorption kinetics and isotherm adhered to the pseudo-second-order kinetic model and Langmuir equation, respectively [29]. Lehmann et al. successfully removed Zn2+ from solutions using MgAl-LDH, particularly in solutions with a pH above 4 [30]. Tarasov et al. incorporated the organic chelating ligand EDTA into the interlayer structure of LiAl-LDHs, thereby endowing them with the ability to chelate Ni2+ [25]. Rojas achieved enhanced removal rates of Cu2+, Cd2+, and Pb2+ using NO3 intercalated in CaAl-LDH [31]. Recent focus has mainly been put on the removal of Cd by LDH. For example, the CaAl-LDH prepared by Kong et al. showed both a high adsorption rate (~5 min) and a high capacity (592 mg/g) for Cd2+ [23]. Shan et al. demonstrated that the adsorption of Cd2+ by MgAl-LDH was feasible, spontaneous and endothermic in nature [32]. Chen et al. reported that CaAl-Cl LDH could significantly reduce the migration coefficient of soil Cd by 29.2%, ensuring the reduction in crop uptake and the relevant ecological risk of Cd [12].
Identified sources of soil Cd pollution encompass geological background, industrial discharge, and sewage irrigation [33,34]. Research indicates that, compared to areas with high geological backgrounds, the activity of heavy metal elements like Cd in soils contaminated by anthropogenic factors is elevated, resulting in greater harm to crops [35]. Furthermore, mitigating pollution from human sources is comparatively straightforward. Consequently, the application of in situ fixation and other remediation technologies in soils contaminated by anthropogenic heavy metals is feasible. The toxicity of heavy metals in soils to plants is primarily determined by the occurrence state of these metals [36,37,38]. Highly active, water-soluble Cd and ion exchange Cd are readily absorbed by plants, exhibiting bioavailability [39]. Transforming heavy metals from highly bioavailable active states to stable states with low bioavailability can mitigate the harm of these metals in soils, thereby achieving the goal of soil remediation [40]. On this basis, Mao et al. proposed the concept of stable mineralization structure, which involves anchoring heavy metal ions within the lattice of LDHs under an in situ mineralization strategy. Through this process, M-LDH is formed (M denotes a heavy metal element), the structure of which exhibits extremely low solubility and high stability [41]. As a promising passivator for Cd, LDHs are expected to reduce its mobility in soil, thereby enhancing crop safety. Despite this promise, the long-term stability of LDHs in practical soil remediation remains a significant challenge. Their effectiveness is highly contingent on soil-specific conditions, such as the presence of competitive cations (e.g., Ca2+, Mg2+), fluctuations in pH, ionic strength, organic matter content, and redox conditions [42,43,44]. These dynamic properties can undermine the stability of passivated Cd, increasing the risk of its remobilization over time. Consequently, evaluating the persistence of LDHs’ passivation effect under realistic field conditions is essential. However, research on their long-term performance is still notably limited.
Based on the foregoing background, it was hypothesized that CaAl-LDH could effectively passivate Cd in contaminated farmland soil via various mechanisms, thereby reducing the proportion of bioavailable Cd and its accumulation in grains, and maintaining remediation stability over the years. To test this hypothesis, the specific objectives of the study included: (1) investigating the adsorption behavior and underlying mechanisms of Cd2+ adsorption by CaAl-LDH via laboratory adsorption test and material characterization; and (2) assessing the practical remediation efficiency and long-term performance of CaAl-LDH under field conditions by analyzing changes in Cd fractions in soil and Cd content in grains over a three-year period. This study is expected to provide both theoretical and practical insights into the sustainable remediation of heavy metal-contaminated farmland, contributing to the maintenance of agricultural productivity and safety.

2. Materials And Methods

2.1. Preparation

Among various methods for preparing CaAl-LDH, the co-precipitation method is operationally straightforward, making it an ideal choice for laboratory synthesis [45]. The preparation of CaAl-LDH proceeded as follows: 28.34 g of Ca(NO3)2·4H2O and 24.00 g of Al(NO3)3·9H2O were dissolved in 200 mL of deionized water to form solution A, followed by dissolving 17.6 g of NaOH in another 200 mL of deionized water to form solution B. Solutions A and B were then introduced into the colloid mill reactor at a speed of 3000 rpm. After reacting for 3 h, the product was collected, washed several times with deionized water and anhydrous ethanol, and subsequently dried in a 40 °C oven for 12 h prior to sample collection.
As for the subsequent field remediation experiment, the CaAl-LDH used was industrially synthesized by Jiangsu Longchang Chemical Co., Ltd (Rugao, China). Its quality complied with the following industrial standards issued by the Ministry of Agriculture and Rural Affairs of China: Soil Amendments-Determination of Calcium, Magnesium and Silicon Content (NY/T2272—2012) [46], Soil Amendments-Determination of Phosphorus and Potassium Content (NY/T2273—2012) [47], Soil Amendments-Determination of Aluminum and Nickel (NY/T3035—2016) [48], and Soil Amendments-General Regulations (NY/T3034—2016) [49].

2.2. Adsorption Test of Cd by CaAl-LDH

2.2.1. Adsorption Kinetics

To investigate the adsorption efficacy of CaAl-LDH on Cd2+, four Cd(NO3)2 solutions with a concentration of 150 mg·L–1 were prepared, maintaining pH levels at 4, 5, 6, and 7 for closely simulating actual soil conditions in the subsequent field studies. For each solution, 50 mL was transferred into a conical flask, to which 10 mg of CaAl-LDH was added, followed by agitation at room temperature at a stirring rate of 150 rpm. Samples were collected at 5, 10, 30, 60, 90, 120, 240, 480, 720, 960, 1500, and 2220 min, respectively. At regular intervals, a 2.5 mL needle tubing was employed to extract the supernatant, which was then filtered through a 0.22 μm membrane. Subsequently, the concentration of Cd2+ in the supernatant was quantified using inductively coupled plasma–optical emission spectrometry. The adsorption capacity of CaAl-LDH for Cd2+ was calculated using Equation (1), and the corresponding adsorption curve was plotted.
q t = ( c 0 c t ) × V / m
In this context, qt represents the adsorption capacity of the adsorbent for Cd2+ at time t (mg·g–1), c0 denotes the initial concentration of Cd2+ in the solution (mg·L–1), ct refers to the concentration of Cd2+ at time t (mg·L–1), V signifies the volume of the Cd2+ solution (L), and m is the mass of the adsorbent (g).
Adsorption kinetic models are utilized to delineate the relationship between adsorption rate and various parameters, commonly characterized through adsorption kinetics equations, reflecting both the performance and mechanism of adsorbents [50]. The pseudo-first-order kinetic model demonstrates that the adsorption rate is directly proportional to the concentration of the adsorbate [51]. Contrarily, the pseudo-second-order kinetic model does not exhibit a simple linear relationship between adsorption rate and adsorbate concentration, suggesting a chemical reaction process between the adsorbent and adsorbate, and the formation of chemical bonds [52].
ln q e q t = k 1 t / 2.303 + l n q e
t / q t = 1 / k 2 q e 2 + t / q e
In this context, qe denotes the adsorption capacity of the adsorbent at equilibrium (mg·g–1); qt denotes the adsorption capacity at time t (mg·g–1); k1 denotes the reaction rate constant (min–1) of the pseudo-first-order kinetic model; k2 denotes the reaction rate constant (g·min–1) of the pseudo-second-order kinetic model; t denotes the adsorption time (min).

2.2.2. Adsorption Isotherms

To investigate the relationship between CaAl-LDH’s adsorption capacity and Cd2+ concentration, Cd(NO3)2 solutions of varying mass concentrations were prepared, including 0, 50, 150, 200, 250, 300, 400, 500, and 600 mg·L−1 maintained at pH 6. For each solution, 50 mL was transferred into a conical flask, to which 10 mg of CaAl-LDH was added, followed by agitation at room temperature at a stirring rate of 150 rpm. Upon reaching adsorption equilibrium at 1500 min, a 2.5 mL needle tubing was used to extract the supernatant, which was then filtered through a 0.22 μm membrane. Subsequently, ICP-OES was employed to quantify the Cd2+ content in the supernatant. The adsorption capacity of CaAl-LDH for Cd2+ was calculated utilizing Equation (4).
q e = ( c 0 c e ) × V / m
In this equation, qe represents the adsorption capacity of the adsorbent (mg·g–1), c0 denotes the initial concentration of Cd2+ in the solution (mg·L–1), ce signifies the concentration of Cd2+ at adsorption equilibrium (mg·L–1), and V is the volume of the Cd2+ solution (L), and m is the mass of the adsorbent (g).
The isothermal adsorption model delineates the relationship between an adsorbent’s adsorption capacity (qe) and the adsorbate’s concentration (ce) at adsorption equilibrium. Based on the monolayer adsorption theory, the Langmuir isotherm adsorption model presupposes a uniform distribution of adsorption sites [53]. This model suggests the presence of multiple adsorption sites with uneven distribution [54]. In this study, the Langmuir isotherm adsorption model (5) and the Freundlich isotherm adsorption model (6) were utilized to characterize the experimental data. The equations for these models are as follows:
c e / q e = c e / q m + 1 / k L q m
l n q e = l n k F + l n c e / n
In these models, qe denotes the adsorption capacity at equilibrium (mg·g–1), qm denotes the maximum adsorption capacity of the adsorbent (mg·g–1), ce denotes the equilibrium concentration of the adsorbate in the solution (mg·L–1), kL denotes the Langmuir adsorption constant (L·mg–1), kF denotes the Freundlich adsorption constant (mg·g–1).

2.2.3. Characterization of CaAl-LDH

The crystal structure of the CaAl-LDH before and after Cd adsorption was analyzed via X-ray powder diffraction (XRD, Mini Flex 600, Rigaku Corporation, Akishima, Japan) with a scanning range of 5–80° and a scan rate of 5°·min−1. In addition, the morphology and structure of CaAl-LDH were analyzed by scanning electron microscopy (SEM, Supra 55, Carl Zeiss AG, Oberkochen, Germany).

2.3. Experimental Site and Design

Zhiying Town, located in Yongkang City, Zhejiang Province, China, experiences a subtropical monsoon climate, characterized by an average monthly rainfall of approximately 120 mm. It is significantly affected by Cd contamination, primarily due to electroplating and metal processing activities in the small hardware industry. The area affected by Cd contamination encompasses up to 1.27 km2, constituting 26.9% of Zhiying Town’s total area. The study area is situated within a large expanse of farmland in Zhiying Town, covering approximately 13 hm2. Land use data indicates that this area consists of basic farmland predominantly used for rice cultivation over an extended period. Soil analysis revealed an average Cd concentration of 0.32 mg·kg−1, which is 1.16 times higher than the local background value. Corresponding Cd levels in rice grains averaged 0.17 mg·kg−1, with 25% of the samples exceeding the national food safety limit (GB 2762-2022) [55]. Detailed characteristics of the experimental site are provided in Tables S1 and S2. A total of 76 surface soil samples were collected from this farmland, and the results were employed for spatial interpolation analysis using the inverse distance weighting algorithm in ArcGIS 10.8 software (Redlands, CA, USA), as illustrated in Figure 1.
In 2020, CaAl-LDH was utilized in a remediation experiment on Cd-contaminated farmland within the experimental field. A 1.5 m protective boundary was established around the experimental field, which was subdivided into six plots, comprising three for the experimental group and three for the control group. Each plot encompassed an area of 20 m2, with the ridges between plots sealed by a 1 m wide film extending to the plow pan, ensuring isolated rows and irrigation for each plot. Guided by the safety utilization and control plan for polluted farmland in Zhejiang Province, and taking into account restoration costs and effectiveness, the initial treatment commenced on 4 June 2020. As per the experimental design (Table 1), soil conditioner and base fertilizer were simultaneously applied to the fields, followed by manual plowing and uniform mixing. The initial application rate of CaAl-LDH was 1500 kg·hm–2, with a cost of 3 CNY/kg (about 0.4 USD/kg). One week subsequent to this, test variety rice seedlings were transplanted into the experimental field, and pollution-free irrigation water was utilized to saturate the field. Consistent field management practices, including irrigation, fertilization and weeding, were conducted in line with local agricultural traditions. In 2021 and 2022, soil amendments were no longer applied, with other treatment measures remaining consistent with those in 2020. The duration of the experiment extended from January 2020 to November 2022.

2.4. Sample Analysis and Quality Assurance

Rhizosphere soil samples of rice from the experimental plots were collected annually during the rice maturity period using the five-point sampling method. The fraction of Cd in the soil was determined using a sequential extraction procedure based on the seven-step method outlined in the analytical standard DD2005-03 [56]. The procedure, detailed in Table S3, was performed to quantify the following fractions: water-soluble Cd (F1), ion exchange Cd (F2), carbonate-bound Cd (F3), humic acid Cd (F4), Fe-Mn oxidized Cd (F5), strongly organic-bound Cd (F6), and residual Cd (F7). Typically, the sum of the water-soluble (F1) and ion-exchangeable (F2) fractions is defined as the bioavailable Cd, which is readily absorbed by plants. Before measurement of total Cd content, soils were digested by using HNO3, H2O2, HF and HClO4, and grains were digested by using HNO3 and H2O2 according to the standard methods described in Tables S4 and S5 [13,57,58]. The concentration of Cd in liquid samples was quantified using either inductively coupled plasma mass spectrometry (ICP-MS, XSERIES 2, Thermo Fisher Scientific, Waltham, MA, USA) or inductively coupled plasma–optical emission spectrometry (ICP-OES, iCAP7400, Thermo Fisher Scientific, Waltham, MA, USA). The method detection limits and quantification limits for Cd were 0.002 mg·kg−1 and 0.005 mg·kg−1 for ICP-MS, and 1.000 mg·L−1 and 3.000 mg·L−1 for ICP-OES, respectively. Concentrations of other heavy metals were also determined using either ICP-MS or ICP-OES.
Quality assurance was carried out using the national primary reference materials: GBW07405-GSS-5 for soil and GBW10010-GSB-1 for rice grains. The recovery rates for Cd ranged from 90% to 110% for solid samples. The accuracy of sample processing and analysis strictly adhered to the Regional Geochemical Sample Analysis (DZ/T0279–2016) [59]. Based on the quality assessment of the test results, both the relative error and relative deviation were within the acceptable limits, confirming that the data quality met the required standards.

2.5. Data Processing

Data reduction and analysis in this study were conducted using Microsoft Office Excel 2019 (Redmond, Washington, DC, USA). Statistical analysis was performed with the software Statistical Product and Service Solutions (SPSS) 25 (IBM, Armonk, NY, USA), and comparison between treatments regarding heavy metal contents in soils and rice grains was performed by using one-way ANOVA and multiple comparison. Statistical charts were generated using Origin Pro 2022 (OriginLab, Northampton, MA, USA).

3. Results and Discussion

3.1. Characteristics of CaAl-LDH

3.1.1. Contents of Heavy Metals

The primary component of CaAl-LDH is a silica–calcium matrix, with the constituents in descending order of proportion as follows: CaO (56%), SiO2 (26%), MgO (6%), total Fe expressed as Fe2O3 (5%), and Al2O3 (1%). Additional components include SOC (<5%), Na2O (<1%), and K2O (<1%). The pH of a 20 g/L aqueous suspension of CaAl-LDH is 12. The concentrations of heavy metal (loid) elements in CaAl-LDH are below the limits specified in Soil Environmental Quality Risk Control Standard for Soil Contamination of Agricultural Land (GB 15618-2018) [60], issued by the Ministry of Ecology and Environment of China. It is suitable for direct application.

3.1.2. SEM Analysis

The SEM image of CaAl-LDH (Figure 2) reveals a distinct lamellar structural morphology, consistent with previous studies [61,62]. This two-dimensional layered structure, along with its abundant edge sites and interlayer pores, provides critical morphological evidence that helps explain the high Cd adsorption capacity of CaAl-LDH discussed in the following sections.

3.2. Adsorption of Cd by CaAl-LDH

3.2.1. Adsorption Kinetics

Figure 3 displays the adsorption curves of CaAl-LDH for Cd2+ at various pH values. These curves illustrate that the adsorption rate of CaAl-LDH for Cd2+ varied over time. Within the initial 20 min of the experiment, the adsorption rate was at its highest, displaying a pronounced upward trend and rapid adsorption. During this phase, the surface of CaAl-LDH had ample active adsorption sites capable of rapidly adsorbing nearby Cd2+ ions. As the adsorption process progressed, these active sites became occupied and coated by an abundance of Cd2+ ions or new compounds. Simultaneously, some CaAl-LDH particles in the solution agglomerated, leading to a decrease in total active adsorption sites and a gradual leveling of the adsorption curve. Beyond 120 min, there was a notable decrease in adsorption rate, transitioning into a slower adsorption phase. Following 1500 min, the increase in adsorption capacity became negligible, as the adsorption sites neared saturation, effectively reaching the adsorptive capacity of CaAl-LDH.
CaAl-LDH’s adsorption capacity exhibited significant variation across different pH values. At pH values ranging from 4 to 7, the adsorption capacities were 341.73 mg·g–1, 336.18 mg·g–1, 450.05 mg·g–1, and 468.52 mg·g–1, respectively. With an increase in pH value, there was a significant improvement in the total adsorption amount. This suggests that the pH value significantly influences the adsorption capacity of CaAl-LDH. In contrast to acidic conditions, a pH range of 6–7 could help diminish the activity of Cd2+ and enhance its adsorption and removal in the medium.
A linear model representing ln(qeqt) versus time for pseudo-first-order reaction kinetics was constructed using experimental data, with the results depicted in Figure 4. The goodness-of-fit values for the Cd sorption at pH of 4, 5, 6 and 7 were 0.8410, 0.9081 0.8404, and 0.8396, respectively. The plot of t/qt versus time, based on the pseudo-second-order reaction kinetic model, was constructed and its results are illustrated in Figure 5. Goodness-of-fit values for different pH levels were 0.9984, 0.9933, 0.9740, and 0.9535, respectively.
The model fitting outcomes indicated a higher goodness of fit for the pseudo-second-order reaction kinetic model within a pH range of 4–7, implying that this model more accurately describes the adsorption reaction of CaAl-LDH against Cd2+. This implies the presence of a chemical reaction in the adsorption process, encompassing the coordination of Cd2+ ions with CaAl-LDH’s surface hydroxyl groups and the replacement of Ca2+ by Cd2+ within the CaAl-LDH lattice [24,25].

3.2.2. Adsorption Isotherms

The Langmuir and Freundlich models were applied to fit the experimental data, as illustrated in Figure 6. The resulting regression equations were c e / q e = 0.00213 × c e + 0.14156   ( R 2 = 0.98427 ) and ln q e = 0.49373 × ln c e + 3.18797   ( R 2 = 0.82113 ) .
The fitting results of the models indicate that the adsorption process of CaAl-LDH for Cd2+ closely aligns with both the Langmuir and Freundlich models, with the Langmuir model being more representative. This suggests that CaAl-LDH’s adsorption of Cd2+ is predominantly monolayer with a uniform distribution of active adsorption sites. As per the Langmuir model, the maximum adsorption capacity of CaAl-LDH for Cd2+ is calculated to be 469.48 mg·g–1, under conditions of room temperature and a pH of 6, which is consistent (of the same order of magnitude) with values reported in the literature [63].
According to previous studies, the adsorption mechanisms for Cd2+ on LDH could include the following: (1) surface functional groups formed complexes with Cd; (2) interlayer anions of LDH exchanged with waterborne anions and subsequently bound Cd via electrostatic adsorption [64]; (3) partial dissolution of the LDH released OH and interlayer anions, and formed precipitates with Cd [21]; and (4) isomorphic substitution of Ca2+ with Cd2+ led to stable adsorption [23].

3.2.3. XRD Analysis

XRD patterns of CaAl-LDH before and after Cd adsorption are shown in Figure 7. The diffraction peaks of the synthesized CaAl-LDH closely matched those in the standard card PDF#42-0558 (Ca2Al(OH)6Cl·2H2O), with peaks at 2θ = 11.21°, 23.71°, and 31.67° corresponding to the (002), (004), and (041) crystal planes, respectively. In addition, the peak at 2θ = 29.51° corresponded to the (104) crystal plane of PDF#05-0586 (CaCO3). These results confirmed the successful bench-scale synthesis of CaAl-LDH with high purity and crystallinity [65], and the presence of CaCO3 might just be attributed to the dissolution of atmospheric CO2 into the reaction solution during synthesis [66]. Following Cd adsorption, the XRD pattern showed good agreement with the standard reference PDF#42-1471 (Cd4Al2O6(CO3)·10H2O), and no impurities peaks were detected. The characteristic diffraction peaks observed at 2θ = 11.77°, 23.67°, and 31.50° were indexed to the (002), (004), and (110) crystal planes, respectively. Based on the Bragg’s law [67], the interlayer spacing of CaAl-LDH was calculated to be 0.788 nm prior to adsorption, which decreased to 0.752 nm after the adsorption process. The structural transformation likely resulted from the isomorphic substitution of Ca2+ by Cd2+ in the CaAl-LDH lattice, maintaining the original crystal framework and consequently facilitating stable fixation of Cd.

3.3. Effect of CaAl-LDH on Soil pH

Table 2 presents the test results of soil pH values from the experimental field. In 2020, the soil pH value of the control group significantly increased compared to pre-experiment levels, consistent with previous report [12]. A continued decrease in pH was observed over the long-term from 2020 to 2021 (p < 0.05). For the experimental group, soil pH values remained relatively stable both before and after the experiment (p < 0.05), generally exhibiting weak acidity. In acidic soils, the increase in soil pH further had a positive effect on the passivation of Cd through precipitation.

3.4. Effect of Passivation Remediation on Occurrence Forms of Soil Cd

Figure 8 displays the occurrence forms of Cd in the soil before and after the experimental field test. Analysis of the control group over three consecutive years revealed significant changes in the proportions of water-soluble Cd, humic acid-bound Cd, and Fe-Mn oxidized Cd prior to the comparative experiment in 2020 (p < 0.05), but no significant difference in the proportions of other Cd forms (p > 0.05). Throughout the observation period, there were no significant differences in the proportions of all Cd forms compared to those before the comparative experiment (p > 0.05). The soil in the experimental field experienced disturbances from natural or artificial plowing, resulting in dynamic changes in the proportions of certain forms of Cd. Overall, the forms of Cd were relatively stable, offering some reference value. The proportions were as follows: ion exchange Cd ranged from 63.35 to 70.92%, water-soluble Cd from 1.18 to 2.38%, and carbonate-bound Cd from 6.23 to 9.74%. Generally, the proportion of bioavailable Cd was relatively high.
Over three consecutive years, analysis of the experimental group showed that the proportion of water-soluble Cd at the end of 2020 was 0.79%, marking a decrease of 1.59 percent from the control group in the same period (p < 0.05), though a slight increase was noted during the extended observation period. The proportion of ion exchange Cd was 44.21%, signifying a substantial reduction of 26.71 percent from the control group (p < 0.05), in agreement with the findings of Kong et al. [23]. Furthermore, long-term observations align closely to pre-experiment levels. Fe-Mn oxidized Cd increased by 5.68 percentage points relative to the control group (p < 0.05), yet long-term observations indicated a decline to pre-experimental levels. Residual Cd experienced an 18.49 percentage point increase by the end of the 2020 experiment (p < 0.05), with subsequent long-term observations revealing a decrease. The proportions of carbonate-bound Cd, humic acid-bound Cd, and strongly organic-bound Cd showed no significant changes compared to the control group (p > 0.05).
The 2020 experimental and long-term observation results primarily indicated a transformation of ion exchange Cd into residual Cd and Fe-Mn oxidized Cd following the application of CaAl-LDH in the soil. Similar results were reported by other researchers [68,69] which may be due to the passivation of soil Cd by CaAl-LDH through a variety of mechanisms, such as surface complexation, interlayer anion exchange-adsorption, dissolution–precipitation, and isomorphic substitution. Particularly, this mirrors the findings of batch adsorption experiments, wherein ion exchange Cd was incorporated into the CaAl-LDH layered structure, replacing Ca2+ and becoming fixed within the CaAl-LDH lattice, effectively facilitating the passivation of available soil Cd. Long-term observations revealed that the proportions of ion exchange Cd and residual Cd exhibited a trend toward reverting to pre-experimental levels, discounting the input of Cd from irrigation water. A plausible explanation is that plowing brought deeply contaminated soil to the surface, while the adsorption capacity of CaAl-LDH gradually reached saturation, thereby diminishing its passivation effect. According to the experimental findings, CaAl-LDH has the potential to control the availability of Cd in soils of industrially contaminated areas for 1 to 2 years.

3.5. Effect of Passivation Remediation on Cd Content in Rice Grains

Table 3 presents the Cd content in rice grains from both the control and experimental groups in the experimental field. From 2020 to 2022, the Cd content in rice grains of the control group ranged from 0.16 to 0.23 mg·kg–1, with no significant difference (p > 0.05). These results suggest that the Cd content in rice grains remained stable throughout the experimental period and was influenced by the application of CaAl-LDH. Given that other environmental factors had minimal impact, the experimental results are deemed reliable. In 2020, Cd content in rice grains of the experimental group was 0.06 mg·kg–1, a reduction of 68.42% compared to the control group (0.05 < p < 0.1). Research on the correlation between soil Cd morphology and rice absorption indicates a significant positive relationship between exchangeable soil Cd content and Cd content in rice grains [70,71]. Apparently, the reduction in Cd content in grains could mainly be attributed to the transformation of ion exchange Cd to the most inert and least bioavailable residual state, as shown in Figure 8. Consequently, the remediation impact of CaAl-LDH proved significant.
In 2021, the long-term observation group’s rice grains exhibited a Cd content of 0.10 mg·kg–1. With an increase in the proportion of ion exchange Cd in the root soil, the Cd content in rice grains increased compared to that in 2020, yet remained significantly lower than the control group (p < 0.05). This indicated a weakening Cd remediation effect of CaAl-LDH’s. For the long-term observation group in 2022, rice grains had a Cd content of 0.14 mg·kg–1, still significantly different from that of the control group in the corresponding period (p < 0.05). However, this represented a continued increase from the previous year, indicating a further diminution of CaAl-LDH’s remediation effectiveness. As shown in Figure 8, some of the residual Cd gradually reactivated and became ion exchange Cd, which could well explain the observed phenomenon in grains.
CaAl-LDH exhibited its most effective remediation in 2020, significantly reducing the Cd content in rice grains from the experimental area. Between 2021 and 2022, the tilling of deeper contaminated soil led to the saturation of CaAl-LDH’s adsorption sites, resulting in a resurgence of soil Cd activity and a gradual decline in its passivation effect. Consequently, the Cd content in rice grains of the experimental group increased annually. As per the National Food Safety Standard for Maximum Levels of Contaminants in Foods (GB 2762-2022) [55], the Cd content in rice grains of the control group exceeded the permissible limit by 24% over the three-year trial, whereas the experimental group’s rice met the standard.
In general, CaAl-LDH demonstrated long-term effectiveness in passivating Cd in farmland, as evidenced by the consistently lower Cd content in rice grains compared to the control group. However, a gradual decline in its performance over time was observed. This reduction in efficacy could be attributed primarily to the following mechanisms: (1) Competitive adsorption from coexisting cations in the soil limited the accessibility and occupancy of active sites on LDH, thereby constraining its sustained passivation capacity [44]; (2) Changes in soil redox conditions, particularly those resulting from alternating wetting and drying cycles, could reactivate previously passivated Cd [72,73]; (3) Soil components such as organic acids, along with changes in the microenvironment, may reduce LDH stability, thereby affecting the long-term stability of passivated Cd forms [62,74]; and (4) upward migration of Cd from deeper layers through capillary action and root penetration could further increase the risk of late-stage uptake. These alterations in soil physical, chemical, and biological properties, combined with competitive ion adsorption, were likely responsible for the progressive decline in Cd passivation performance. Further investigation is expected to guide the design and synthesis of modified materials, enabling adjustments and optimization for sustainable passivation performance.

3.6. Limitations and Prospect in Practical Application

While this study confirmed that CaAl-LDH application effectively reduced Cd uptake in rice, a more comprehensive understanding of its environmental impacts and underlying mechanisms requires further investigation. The present work focused primarily on crop uptake responses; subsequent research should place greater emphasis on soil ecological processes, particularly the interactions among LDH, heavy metals, and soil microbial communities, to better evaluate the passivator’s impact on soil health. In terms of mechanistic insight, this study identified key adsorption behavior of Cd on LDH through aqueous-phase adsorption experiments. Due to technical challenges, in situ mechanistic studies within complex soil environments remain limited. Therefore, future work should prioritize elucidating reaction pathways under realistic soil conditions.
Furthermore, the gradual decline in passivation efficacy over time was likely attributable to multiple factors, especially for the decrease in active sites and the fluctuations in soil conditions. These insights highlight the need to enhance the environmental stability of LDH through material innovation and to optimize application strategies based on local soil properties and contamination levels, moving beyond empirically determined way of application toward more efficient and sustainable remediation practices.

4. Conclusions

This study has shown that CaAl-LDH effectively passivates Cd in contaminated farmland soil through multiple mechanisms, resulting in sustained reduction in Cd bioavailability and accumulation in rice grains. Laboratory adsorption experiments and material characterization revealed that Cd2+ adsorption by CaAl-LDH follows pseudo-second-order kinetics and the Langmuir model, with a maximum adsorption capacity of 469.48 mg·g−1 at pH 6. The adsorption mechanisms include surface complexation, interlayer anion exchange, dissolution–precipitation, and isomorphic substitution. A three-year field trial in Yongkang City, China demonstrated the practical remediation performance and long-term stability of CaAl-LDH. The passivator promoted the transformation of soil Cd from the ion exchange state (F2) to the residual state (F7) and Fe-Mn oxidized state (F5). The Cd content in rice grains decreased by 68.42% in the first year and remained significantly lower in the subsequent years, consistently below the national food safety limit.
Although a gradual attenuation of the passivation effect was observed beyond the second year, this study demonstrates that CaAl-LDH provides a cost-effective technology for the safe utilization of lightly to moderately Cd-contaminated farmland. Future research should focus on the material’s stability under complex field conditions, particularly its behavior in the presence of competitive cations and fluctuating redox environments, as well as the development of optimized application protocols. These findings offer both theoretical and practical support for sustainable agricultural production and food safety.

Supplementary Materials

The following supporting information can be downloaded at: https://www.mdpi.com/article/10.3390/agronomy15102354/s1, Table S1: Basic physicochemical properties of the experimental soil; Table S2. Average concentrations of heavy metals in soil and rice grains of the experimental field; Table S3. Detailed Procedures for the Seven-Step Sequential Exaction Method for Cd Fractionation in Soil; Table S4. Digestion Procedure for Soil Samples; Table S5. Digestion Procedure for Grain Samples.

Author Contributions

Conceptualization, N.W., J.Z., X.C., J.N. and X.W.; Methodology, X.L., H.F., X.C. and X.W.; Validation, X.L., N.W., H.F., X.C., J.N. and X.W.; Formal Analysis, X.L. and H.F.; Investigation, X.L., H.F., F.H., J.C., R.S., Y.C. (Yining Chen), Y.C. (Yanfang Chen) and X.Z.; Resources, J.Z., K.O. and T.Y.; Writing—Original Draft Preparation, X.L., F.H., J.C., R.S., J.Z., Y.C. (Yanfang Chen), X.Z., K.O., T.Y. and X.W.; Writing—Review and Editing, N.W., H.F., Y.C. (Yining Chen), X.C., J.N. and X.W.; Supervision, N.W., X.C., J.N. and X.W.; Project Administration, J.Z. and X.W.; Funding Acquisition, X.L., N.W., Y.C. (Yining Chen) and X.C. All authors have read and agreed to the published version of the manuscript.

Funding

This work was supported by the National Key Research and Development Program of China [2024YFD1701800; 2022YFC3702500]; the National Natural Science Foundation of China [41807116]; the Open Project of State Key Laboratory of Soil Pollution Control and Safety [SPCS-2025-02012]; the Ecological Environment Research Project of Zhejiang Province [2021HT0031]; the Geological Special Fund Project of Zhejiang Province [2023021]; the Scientific and Technological Innovation Project of China Metallurgical Geology Bureau [CMGBKY202301]; the National College Student Innovation Training Program [202510386029]; and the Japan Society for the Promotion of Science Research Foundation KAKENHI [16H05633].

Data Availability Statement

The data presented in this study are available upon request from the corresponding authors.

Conflicts of Interest

Jianyu Zhang and Xuchuan Zhang were employed by Jiangsu Longchang Chemical Co., Ltd. The remaining authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.

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Figure 1. Location of the experimental field.
Figure 1. Location of the experimental field.
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Figure 2. SEM image of CaAl-LDH.
Figure 2. SEM image of CaAl-LDH.
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Figure 3. Adsorption curve of CaAl-LDH for Cd2+.
Figure 3. Adsorption curve of CaAl-LDH for Cd2+.
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Figure 4. Fitting of pseudo-first-order kinetic model for Cd2+ adsorption by CaAl-LDH.
Figure 4. Fitting of pseudo-first-order kinetic model for Cd2+ adsorption by CaAl-LDH.
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Figure 5. Fitting of pseudo-second-order kinetic model for Cd2+ adsorption by CaAl-LDH.
Figure 5. Fitting of pseudo-second-order kinetic model for Cd2+ adsorption by CaAl-LDH.
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Figure 6. Fitting of Langmuir (a) and Freundlich (b) isothermal adsorption models.
Figure 6. Fitting of Langmuir (a) and Freundlich (b) isothermal adsorption models.
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Figure 7. XRD images of LDH before and after Cd adsorption.
Figure 7. XRD images of LDH before and after Cd adsorption.
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Figure 8. Proportion of Cd fraction to total Cd. Note: F1–F7 represent different fractions of Cd, which are mentioned in Section 2.4. Pre-test: The control group in 2019; Test and CK: The remediation group and the control group in 2020; Post-test and CK: The remediation group and the control group in 2022.
Figure 8. Proportion of Cd fraction to total Cd. Note: F1–F7 represent different fractions of Cd, which are mentioned in Section 2.4. Pre-test: The control group in 2019; Test and CK: The remediation group and the control group in 2020; Post-test and CK: The remediation group and the control group in 2022.
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Table 1. Experimental design for remediation of Cd-contaminated farmland.
Table 1. Experimental design for remediation of Cd-contaminated farmland.
YearTestingControl GroupTest Group
2020TestBase fertilizerCaAl-LDH+Base fertilizer
2021Post-testBase fertilizerBase fertilizer
2022Post-testBase fertilizerBase fertilizer
Base fertilizer: N, P, K compound fertilizer 630 kg·hm–2, urea 270 kg·hm–2
CaAl-LDH: 1500 kg hm–2
Table 2. Soil pH values of the experimental fields.
Table 2. Soil pH values of the experimental fields.
YearTestingControl GroupExperimental Group
2019Pre-test5.04 ± 0.24 B5.04 ± 0.24 A
2020Test5.75 ± 0.48 A a5.04 ± 0.09 A b
2021Post-test5.38 ± 0.61 AB a5.19 ± 0.16 A a
2022Post-test5.07 ± 0.06 B a5.14 ± 0.18 A a
Note: In the context of statistical analysis, ‘a’ and ‘b’ denote the differences between the control and experimental groups, while ‘A’ and ‘B’ indicate differences across years (p < 0.05).
Table 3. Cd content in rice grains from the experimental fields.
Table 3. Cd content in rice grains from the experimental fields.
YearTestingCd Content in Rice (mg/kg)
Control GroupExperimental Group
2020Test0.19 ± 0.10 A a0.06 ± 0.02 B a
2021Post-test0.16 ± 0.03 A a0.10 ± 0.02 AB b
2022Post-test0.23 ± 0.04 A a0.14 ± 0.03 A b
Note: In the data analysis, ‘a’ and ‘b’ denote differences between the control and experimental groups, while ‘A’ and ‘B’ indicate differences across years (p < 0.05).
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Lu, X.; Wei, N.; Fang, H.; Hu, F.; Cheng, J.; Sun, R.; Chen, Y.; Zhang, J.; Chen, Y.; Zhang, X.; et al. Passivation Remediation of Cd-Contaminated Farmland in Yongkang, China by CaAl-LDH: A Mechanism and Application Study. Agronomy 2025, 15, 2354. https://doi.org/10.3390/agronomy15102354

AMA Style

Lu X, Wei N, Fang H, Hu F, Cheng J, Sun R, Chen Y, Zhang J, Chen Y, Zhang X, et al. Passivation Remediation of Cd-Contaminated Farmland in Yongkang, China by CaAl-LDH: A Mechanism and Application Study. Agronomy. 2025; 15(10):2354. https://doi.org/10.3390/agronomy15102354

Chicago/Turabian Style

Lu, Xinzhe, Nan Wei, Haochen Fang, Feng Hu, Jianjun Cheng, Rui Sun, Yining Chen, Jianyu Zhang, Yanfang Chen, Xuchuan Zhang, and et al. 2025. "Passivation Remediation of Cd-Contaminated Farmland in Yongkang, China by CaAl-LDH: A Mechanism and Application Study" Agronomy 15, no. 10: 2354. https://doi.org/10.3390/agronomy15102354

APA Style

Lu, X., Wei, N., Fang, H., Hu, F., Cheng, J., Sun, R., Chen, Y., Zhang, J., Chen, Y., Zhang, X., Oh, K., Yonekura, T., Chen, X., Niu, J., & Wang, X. (2025). Passivation Remediation of Cd-Contaminated Farmland in Yongkang, China by CaAl-LDH: A Mechanism and Application Study. Agronomy, 15(10), 2354. https://doi.org/10.3390/agronomy15102354

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