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Article

Irrigation Promotes Arsenic Mobilization via Goethite: Insight from the Perspective of the Solid–Liquid Interface Interaction Process

by
Hong Xu
1,†,
Yaru Wang
1,†,
Jiankang Wang
1,2,*,
Xin Liu
1,
Cuizhong Chen
1,
Chang Zhou
1 and
Qingyuan Tian
1
1
College of Water Conservancy & Architectural Engineering, Shihezi University, Shihezi 832000, China
2
Key Laboratory of Water Resources Efficient Utilization in Arid Areas, Shihezi University, Shihezi 832000, China
*
Author to whom correspondence should be addressed.
†
These authors contributed equally to this work.
Water 2025, 17(7), 1058; https://doi.org/10.3390/w17071058
Submission received: 8 March 2025 / Revised: 28 March 2025 / Accepted: 29 March 2025 / Published: 3 April 2025
(This article belongs to the Section Water, Agriculture and Aquaculture)
Browse Figures
Versions Notes

Abstract

:
Dramatic changes in the farmland soil–groundwater environment caused by irrigation affects the geochemical behavior of arsenic (As). However, the mechanism of As mobilization during soil–groundwater interactions remains unclear. This study explored the effects of phosphate (PO43−), fulvic acid (FA), and oxic/anoxic conditions on As mobilization through batch and column experiments. The results indicated that a saline–alkali environment and the involvement of PO43−/FA suppressed the adsorption capacity of goethite for As(III) in the water environment and that PO43− had a primary effect. An increase in the PO43−/FA concentration further increased its inhibition. Notably, oxic/anoxic conditions did not affect the adsorption capacity of goethite for TAs in the presence of high concentrations of PO43−/FA. In the soil environment, periodic irrigation led to regular fluctuations in the As content in the soil pore water. The addition of PO43− to irrigation water resulted in a higher content of As in the pore water in the short term. In contrast, the addition of FA greatly increased the long-term mobility of As. This study highlighted that irrigation amplifies As mobility in soil–groundwater systems, particularly in saline–alkali environments, and PO43−/FA addition exacerbates this effect, posing risks to agricultural safety. The systematic management of irrigation practices is recommended to mitigate these risks.

Graphical Abstract

1. Introduction

Arsenic (As) pollution in soil and water has emerged as a significant global challenge and environmental concern. Approximately 240 million people around the world are estimated to face health problems, such as skin cancer, cardiovascular disease, and respiratory disease, due to exposure to high-As environments [1,2]. Arsenic enter human production and life through natural geological processes and human activities. In particular, irrigation activities promote the enrichment of As in groundwater by increasing the interactions between soil, surface water, and groundwater, thereby increasing the risk to agricultural production and human health [3,4]. Therefore, a comprehensive understanding of the geochemical behavior of arsenic in agricultural irrigation systems is essential for the development of effective mitigation methods and regulatory frameworks.
The behavior of groundwater profoundly affects the distribution of various ions. For example, due to the wide application of irrigation technology, the groundwater dynamic field has changed significantly, which promotes the migration of soluble salts in the aquifer [5,6]. In this process, a variety of ions migrate in different media, and the geochemical behavior of As changes accordingly. The co-occurrence of groundwater salinization and As contamination has been reported in many parts of the world, including the Red River Delta, the delta in southwestern Bangladesh, the Datong Basin, the Junggar Basin, and the Pannonia Basin [7,8,9,10]. The mobilization of arsenic in subsurface aqueous systems is primarily governed by two fundamental processes, the reductive breakdown of iron oxides hosting arsenic species and the attenuation of Fe-As complexation stability, which collectively facilitate arsenic release into groundwater [11]. Saline–alkali ions such as sulfate (SO42−) and bicarbonate (HCO3−) have competitive adsorption effects on As, which affects the fixation of As on the surface of iron minerals [12,13]. Excessive contents of carbonate (CO32−) and HCO3− lead to increases in the pH and alkalinity of groundwater and weaken the binding ability of As to iron minerals [14]. Arsenic can also form secondary minerals with some dissolved components, changing their spatial distribution [15]. In arid and semiarid regions, the high saline–alkali content and As enrichment in groundwater are affected by the chemical evolution process of groundwater. However, there has been little research on the connection between these two issues until now.
Irrigation plays a vital role in sustaining agricultural production in arid and semiarid regions. However, high irrigation intensity promotes the release of As from the solid phase to the water phase and its potential environmental impact on agricultural systems cannot be ignored [16]. Irrigation activities introduce oxygen or dissolved oxygen (DO) in surface water into soil and groundwater, affecting the redox conditions of the soil environment [17,18]. When the soil environment is in a reducing state, some iron minerals undergo morphological transformation or reduction dissolution and, because As is mainly present in the form of As(III), As is mainly enriched in groundwater. After oxygen enters the soil environment, the redox potential increases. The dominant species of Fe and As become Fe(III) and As(V). As is fixed in the soil due to the stronger binding force between As(V) and Fe. To improve the yield and quality of crops, some exogenous substances, such as phosphate (PO43−) and organic matter (OM), are added. These substances enter the soil through irrigation and affect the distribution and redistribution of As in agricultural systems through competitive adsorption, ligand exchange, and redox mechanisms [19,20,21]. PO43− competes for active adsorption sites on the surface of iron minerals, reducing the binding capacity of As and iron minerals [22]. Some functional groups of OM have high activity and strong binding ability with As [23]. These functional groups affect the mobilization of As via complexation or by changing the existing forms and chemical forms of As and Fe [24]. How does the oxic/anoxic cycle caused by irrigation affect As mobilization? Is the effect of exogenous substances on As mobilization consistent in the solid phase and water phase under irrigation conditions?
Therefore, in this study, batch experiments and column experiments were designed to investigate the process of As transformation under irrigation. The objectives of this study were to (1) assess the effects of a saline–alkali environment and PO43− and FA input conditions on As adsorption via goethite; (2) elucidate the difference and mechanism of the interaction between As and goethite under oxic/anoxic conditions; and (3) reveal the mobilization and transformation of As in soil under the influence of cyclic irrigation.

2. Materials and Methods

2.1. Materials

All the reagents used in this study were of analytical grade or higher. Sodium arsenite (>90.00%) was used to prepare a 250 mg/L As(III) stock solution, which was diluted in the subsequent experimental process. All the solutions were mixed with deionized water (18.25 MΩ·cm). Before the experiment, all the containers were soaked in 10% HNO3 for 24 h and rinsed with deionized water.

2.2. Batch Experiments

2.2.1. Saline–Alkali Environment and PO43− and FA Input

To simulate a groundwater environment with high salinity, this study selected HCO3−, SO42−, Cl−, and Ca2+ as background ions to carry out a series of experimental studies with the water-quality parameters of groundwater in the Kuitun area of Xinjiang as a reference [25]. Moreover, PO43− and FA were selected as typical substances introduced by irrigation.
Batch experiments were carried out in a saline–alkali environment with different concentrations of PO43− (5, 10, or 20 mg/L) and FA (10, 100, or 500 mg/L). Goethite (0.1 g) was added to 100 mL of an As-containing solution (initial pH = 8.5 ± 0.1) and shaken. All the experiments were conducted three times. The specific experimental steps are described in Text S1.

2.2.2. Oxic and Anoxic Experiments

The preparation method for the deoxygenated water was as follows: after boiling the deionized water, it was cooled in a glove box filled with high-purity N2 (99.999%). All the solutions were prepared under anoxic conditions with deoxygenated water. The anoxic atmosphere was maintained throughout the experiment and subsequent sampling operations. All other experimental steps were the same as those described in Text S1.

2.3. Column Experiments

2.3.1. Synthesis of Goethite-Coated Quartz Sand

Goethite-coated quartz sand was synthesized via the heterogeneous suspension method [26,27]. Briefly, 1 g goethite was added to a 0.01 M NaNO3 (pH = 2.5 ± 0.1) solution and the pretreated quartz sand was mixed for 24 h; after being washed with deionized water, it was dried at 110°C and placed into a self-sealing bag for use. The pretreatment method for quartz sand and further synthesis details are provided in Text S2 and Text S3. The coating density of goethite was approximately 3.5 mg/g quartz sand.

2.3.2. Method of Column Experiments

The column experiments were set up using an inverted T-shaped plexiglass column. As shown in Figure S1, the inner diameters of the transverse column and the vertical column were both 2.5 cm, the length of the transverse column was 20 cm, and the height of the vertical column was 10 cm. A porous plate was set at the junction of the vertical column and the horizontal column to support the simulated soil and 200 mesh (75 μm) nylon was wrapped on the porous plate to prevent sand leakage.
The column experiments were divided into three groups (Table S2). The simulated groundwater was prepared in an anoxic environment and a peristaltic pump was used to pump it into the horizontal column at a speed of 1 mL/min. After the water outlet of the horizontal column stabilized and the liquid level of the vertical column stabilized (2 cm above the simulated soil), sampling was initiated. The upper, middle, and lower pore water samples were collected every 8 h and the water samples were acidified by adding acid after passing through the membrane. The simulated irrigation process was conducted every 60 h for 30 min and the irrigation water flow rate was 3 mL/min. Each group of experiments was conducted for 20 days, totaling seven irrigations.
After the column experiments, the simulated soils in the upper, middle, and lower parts of the column were sampled and the exchangeable As and Fe on the surface were analyzed after freeze-drying for 24 h. The analysis steps were as follows: 0.2 g of the dried solid sample was mixed with 30 mL of an As desorption solution (1 M KH2PO4) and continuously oscillated for 24 h to extract the exchangeable As from the simulated soil surface. Next, 1.5 g of the dry solid sample was mixed with 30 mL of an Fe desorption solution (0.4 M HCl) and continuously oscillated for 10 min to extract the exchangeable Fe from the simulated soil surface [28,29].

2.4. Analytical Methods

2.4.1. Aqueous Phase Analysis

The concentrations of As(III) and total As (TAs) were analyzed via hydride generation atomic fluorescence spectrometry (AFS-9700; Beijing Haiguang Instrument Co., Ltd., Beijing, China). The concentrations of Fe(II) and total Fe (TFe) were measured via the phenanthroline spectrophotometry method (UV-1800, Shanghai Mapada Instrument Co., Ltd., Shanghai, China). DO was measured via a portable water-quality multiparameter tester (SL1000, HACH, Loveland, CO, USA). The pH and redox potential (Eh) were measured via a pH meter (PHS-25, Shanghai Instrument Electrical Science Instruments Co., Ltd., Shanghai, China). The surface charge of the iron minerals was determined via a zeta potential analyzer (DLS15Z, Shanghai Yimai Instrument Technology Co., Ltd., Shanghai, China).

2.4.2. XPS and FTIR Analyses

After reacting in a saline–alkali environment in the presence of 20 mg/L PO43− and 500 mg/L FA, goethite was selected for the characterization analysis. The changes in the surface functional groups of goethite under different conditions were analyzed via FTIR (Escalab250Xi, Thermo Fisher Scientific, Waltham, MA, USA). The chemical forms of the elements on the surface of the goethite were analyzed via XPS (Nicolet 6700, Thermo Fisher Scientific, USA). XPS data were analyzed via Advantage 5.9 software. The detailed information used to characterize the goethite method is shown in Text S4.

2.5. Statistical Analysis

The batch experiments were repeated three times and the results were expressed as the mean ± standard deviation. A one-way analysis of variance (ANOVA) was performed using SPSS 27.0.1 and then the least significant difference test was performed to compare the significant differences in the effects of different factors on the adsorption capacity of goethite. In addition, the batch experiments were fitted to an exponential function model [6,29], as shown in Equation (1):
C t = A + B e − K t
where C t is the As concentration at a certain time in the solution (mg/L) and K is the rate constant (min−1).

3. Results and Discussion

3.1. Adsorption of As(III) on Goethite Surface

3.1.1. Effects of the Saline–Alkali Environment

Figure 1 shows that the adsorption process of As(III) by goethite conformed with the exponential decay process [30]. In the initial stage of adsorption (0–30 min), As(III) was rapidly adsorbed. At this stage, the adsorption capacity of goethite for As(III) reached 80% of the final adsorption capacity. Subsequently, the adsorption rate of goethite gradually decreased and finally reached the adsorption equilibrium state. The concentration of As(III) no longer significantly changed.
After the adsorption equilibrium of goethite was reached (Figure 1a,b), the concentration of As(III) in the control group was similar to that in the saline and saline–alkali environments (p > 0.05), while the concentration of TAs differed (p < 0.01). In the control group (blank), the adsorption capacity of TAs by goethite was 0.45 mg/g. It was 0.37 mg/g after adding saline and alkali ions, which decreased by 17.83%. In addition, the concentration of As in the solution exponentially decreased, so this study fitted the exponential function model and compared the rate constants under different environments. As shown in Table S3, compared with the control group (blank), the rate constant decreased after the addition of salt ions. This showed that these salt ions had an inhibitory effect on the adsorption of goethite. The addition of SO42− and HCO3− could form outer-sphere complexes on the surface of iron minerals and affect As(III) adsorption [13,31]. Moreover, Saalfield et al. [32] reported that, in the presence of bicarbonate, a combination of bicarbonate and Ca2+ or Mg2+ induced the desorption of As on an iron ore surface. In addition, as shown in Figure 2, the zeta potential of the goethite surface decreased in a saline–alkali environment and the electrostatic repulsion between As and iron minerals increased, indicating the inhibitory adsorption of As. Notably, a small amount of As(III) was oxidized to As(V) during the adsorption of As(III) by goethite. In addition to As(III) being slowly oxidized by O2 in the air [33,34], Fe(III) in goethite also has a certain oxidative effect [35,36]. Under the conditions of these experiments, As(V) in the solution had a stronger negative charge than As(III), so the inhibition of As(V) was stronger, which further led to an increase in the TAs concentration in the saline-alkali environment.

3.1.2. Effects of PO43− and FA

In a saline–alkali environment, the addition of PO43− and FA significantly inhibited the adsorption of As(III) on the goethite surface (Figure 1c–f). After PO43− and FA were added, the adsorption capacities of goethite for TAs were reduced by 49.62% and 24.04% (p < 0.01), respectively. This result indicated that the inhibitory effect of PO43− on goethite adsorption was stronger than that of FA and that its inhibitory ability increased as the initial concentration of PO43− increased. In addition, with the increase in its concentration, the adsorption capacity of goethite for As(III) also decreased. When the concentration of PO43− increased from 5 mg/L to 20 mg/L, the adsorption capacity decreased by 40.06~57.95% (p < 0.001) and the rate constant decreased from 0.17 to 0.13 min−1. Similarly, when the concentration of FA increased from 10 mg/L to 500 mg/L, the adsorption capacity decreased by 15.15~33.49% (p < 0.001). When the FA concentration was 500 mg/L, its rate constant exceeded the low-concentration FA conditions (10 and 100 mg/L). This was because FA occupied too many adsorption sites, greatly reducing the adsorption capacity of goethite for As and making the adsorption process reach an equilibrium faster. PO43− and As(III) have similar physicochemical properties and can form inner-sphere complexes with iron oxides. Thus, both of these compounds strongly compete for adsorption on a goethite surface [37]. Distinct from PO43−, FA molecules are characterized by abundant carboxyl and phenolic functional groups that promote stable surface complexation, consequently suppressing the As(III) adsorption capacity on iron mineral substrates through surface site competition [38]. A low surface charge is not conducive to the adsorption of As(III) [39]. Moreover, FA can form an As(III)–FA complex with As(III) and inhibit As(III) from binding to iron mineral surfaces [40]. Through an FA-induced redox reaction, goethite underwent reductive dissolution and the density of the active adsorption sites on its surface also decreased. At the same time, As(III) was oxidized through the Fenton reaction induced by Fe(II) (Section 3.1.3). Dissolved Fe can form colloidal or dissolved Fe–organic-matter complexes with FA [41] and can further form ternary As-Fe–organic-matter complexes with As [3,38].

3.1.3. XPS and FTIR Analyses

The functional groups on the surface of goethite changed in a saline–alkali environment in the presence of PO43− and FA (Figure 3). The peak near 3412 cm−1 was attributed to surface-bound H2O molecules [42] and the weak peak at 2875 cm−1 was caused by C–H stretching vibrations [43]. The characteristic band observed at approximately 556 cm−1 was identified as corresponding with the vibrational mode of the iron–oxygen covalent bond [44]. The characteristic infrared absorption bands observed at wavenumbers 876 and 710 cm−1 were assigned to hydroxyl group deformation vibrations in the α-FeO(OH) crystalline structure, confirming the presence of goethite in the analyzed sample [45]. The peak at 475 cm−1 was attributed to the complexation of As and iron oxides, which proved that As was adsorbed on the surface of the goethite [46]. The peak near 1422 cm−1 belonged to the carbonate species [47], which was caused by the formation of inner-sphere complexes on the surface of the goethite [48]. The presence of carbonate characteristic peaks was also observed in the control group without HCO3−, which was mainly due to the adsorption of CO2 in the air on the surface of the goethite [49].
The frequency of PO43− molecules was in the range of 900~1200 cm−1, which was mainly adsorbed on the surface of the goethite through the bidentate coordination of bridge bonds and monoprotonated monodentate coordination [50]. The weak peaks near 1050 and 1180 cm−1 represented the C-O-C bond and O-H bond, respectively, indicating C-O stretching and O-H deformation in the carboxyl group [40]. This phenomenon indicated that FA had complexed with goethite, affecting the adsorption of As(III) in the goethite via competitive adsorption.
XPS was used to analyze goethite under four experimental conditions to explore the changes in the chemical forms of elements after the adsorption of goethite in a saline–alkali environment and in the presence of PO43− and FA. The analysis of the O 1s energy spectrum (Figure S3) indicated that the peak at 529.60 eV corresponded with metal oxides (Fe-O), and two new peaks were observed in the spectrum at 531.51 and 532.98 eV following goethite adsorption corresponding with HCO3− and As (As-O), respectively. The results indicated that HCO3− had a competitive adsorption effect. The As 3d energy spectrum analysis, as shown in Figure 4, revealed two peaks under FA conditions. These were an As(III) peak at 43.71 eV and an As(V) peak at 44.98 eV (Figure 4d). In contrast, the saline–alkali and PO43− environments exhibited only an As(III) peak at 43.71 eV (Figure 4b–d). Figure 5 shows the Fe 2p energy spectrum, where the peaks at 710.24 eV, 723.74 eV, 719.22 eV, and 729.26 eV correspond with the characteristic peaks of goethite (Fe2O3). In contrast, two new peaks appeared in the Fe 2p spectra at 710.37 eV and 723.78 eV under FA conditions (Figure 5d), which were attributed to FeO, indicating that part of the Fe(III) had converted to Fe(II). These findings revealed that the presence of FA induced the redox reaction of goethite with As(III). In this process, Fe(III) was first reduced to Fe(II) by FA. Dissolved Fe(II) reacted with O2 to form a variety of reactive oxygen species, such as H2O2, · OH, etc., which oxidized As(III) in the solution to As(V) and was then adsorbed by the goethite [51].

3.2. Adsorption of As(III) Under Oxic and Anoxic Conditions

In general, the adsorption capacity of goethite for As(III) under oxic conditions was greater than that under anoxic conditions (Figure 6). In the control group (Figure 6a,b), the adsorption capacity of goethite under anoxic conditions was 0.42 mg/g (p < 0.05), which was 7.51% lower than that under oxic conditions. The inhibitory effect under anoxic conditions in saline–alkaline conditions was slightly stronger than that under oxic conditions (Figure 6c,d). The adsorption capacity of As(III) by goethite was 0.36 mg/g (p < 0.05), a decrease of 9.20%. Overall, the combined effect of saline–alkali and anoxic conditions significantly inhibited the adsorption of As(III) on the surface of iron minerals (the combined effect of the two decreased the adsorption capacity by 25.38%). However, different results were shown when PO43− and FA were separately present. Unexpectedly, the adsorption capacity of goethite in the presence of PO43− was 0.16 mg/g under oxic and anoxic conditions (Figure 6e,f; p > 0.05). Similarly, the adsorption capacity of goethite in the presence of FA was 0.25 mg/g and 0.24 mg/g under oxic and anoxic conditions, respectively (Figure 6g,h; p > 0.05). This indicated that the effect of the oxic/anoxic conditions became weak. When comparing the rate constants under oxic and anoxic conditions (Table S3), we observed that the rate constants under anoxic conditions in different environments were lower than those under oxic conditions, which proved that anoxic conditions could significantly reduce the adsorption rate of goethite and prolong the time to reach an equilibrium. However, in the presence of high concentrations of PO43− and FA, the adsorption capacity of TAs by goethite was almost unchanged.
The zeta potential on the surface of the iron minerals decreased under anoxic conditions regardless of the addition of exogenous substances (HCO3−, PO43−, FA, or HA) [47]. The possible reason was that the increase in pH under anoxic conditions led to the deprotonation of hydroxyl groups on the mineral surface [52], thereby reducing the surface charge of the goethite. This phenomenon was verified in the subsequent column experiments, which are described in Section 3.3 (Table S4). The deeper the depth, the lower the DO content and the higher the pH of the soil pore water. A higher pH weakened the binding ability of As to goethite, and this inhibition was weakened due to the presence of PO43− or FA. Therefore, in the presence of PO43− and FA, the concentration of As(III) in the solution under anoxic conditions was slightly higher than that under oxic conditions. At the same time, more As(III) was oxidized under oxic conditions, while PO43− or FA had a stronger inhibitory effect on As(V). This resulted in similar TAs concentrations in the solution under oxic and anoxic conditions in the presence of PO43− and FA.

3.3. As Mobilization During Irrigation

3.3.1. Variations in As Content in Pore Water

Irrigation activities not only introduce exogenous substances but also change the redox environment of the soil layer at different depths, affecting the release of As in the soil [7,53]. The column experiments were designed using deionized water (Column A), PO43− (Column B), and FA (Column C) for irrigation to elucidate the potential effects of irrigation activities and different irrigation water components (PO43− and FA) on As mobilization.
As shown in Figure S4 and Figure 7, before the first irrigation (0–60 h), the simulated groundwater entered the soil and the As in the water was adsorbed by the soil. The As content in the middle and upper pore water tended to stabilize, whereas that in the lower pore water gradually increased through the exchange of substances with the groundwater. After each irrigation event, the As content in the pore water sharply decreased but As was subsequently desorbed from the soil surface. A large amount of As was released, which increased the risk of groundwater pollution. As can form a variety of complexes of varying strengths with iron(hydr)oxides. With some weaker-bound As being more sensitive to environmental perturbations and is more easily removed from the soil suiface [16,54,55]. Moreover, the magnitude of the fluctuation in As at different depths within the soil column was lower > middle > upper, which was related to the As content in the soil and the redox conditions at different depths. There were fewer interactions between the upper soil and groundwater, the As adsorbed on the soil surface was less, and the upper soil was in a relatively oxidized environment (Table S4). The mobility of As was weakened in this environment [56,57]. The results of the simulated irrigation experiments revealed that irrigation caused fluctuations in the As content in the soil pore water and that the fluctuations in As increased with an increase in depth.

3.3.2. Influence of Irrigation Water Composition

Similar to that in Column A, after PO43− was added to the irrigation water (Figure S4b and Figure 7b), the fluctuation in the As content in the soil column gradually stabilized with an increase in irrigation time. Compared with that in Column A, the As content in the upper and middle soil pore water was always greater when PO43− was added to the irrigation water. The competitive adsorption mechanism induced by PO43− at the solid–water interface weakens the binding capacity of As with minerals and promotes the activation of As [58]. This competitive phenomenon is similar to the competitive mobilization behavior of Cu2+ and Zn2+ in soil [30]. The lower soil layer was in an anoxic environment (Section 3.2) and the competitive adsorption of PO43− was inhibited in this environment. Therefore, the fluctuation range of the As content in the lower soil of Column A and Column B was similar.
In contrast, after adding FA to the irrigation water, the variation in the TAs content in the pore water was different from that in Columns A and B. The fluctuation range of the TAs content in the middle and lower soil pore water slowly decreased and then rapidly increased with an increase in irrigation time. However, the fluctuation range of As(III) in the middle and lower pore water of the soil column did not increase with an increase in irrigation time (Figure S4c). The proportion of As(V) in the pore water increased with an increase in irrigation time. This phenomenon indicated that, in addition to the competitive adsorption of FA, there were other mechanisms that controlled the mobilization of As.
Further studies revealed that, compared with Columns A and B, a large amount of exchangeable Fe accumulated on the bottom soil surface of Column C (Figure 8). The contents of exchangeable Fe and As in each layer of soil in Column C were greater than were those in Columns A and B, indicating that FA could promote the reduction and dissolution of Fe. Dissolved Fe and As in the pore water migrated to the bottom of the soil column with irrigation water. The previous study (Section 3.1.3) revealed that goethite could be reduced to Fe(II) by FA and that partially dissolved Fe(II) was transferred to a lower layer with irrigation water. Owing to the DO carried by the irrigation water, As(III) was oxidized to As(V) by Fe(II) and the formed Fe(III) reattached to the underlying soil. Moreover, the anoxic environment of the lower soil promoted the release of As(V) [59] and reduced the adsorption capacity of As(V) through the high negative charge on the surface of the iron minerals [60]. Therefore, when FA was added to the irrigation water, the fluctuation range of the TAs content in the soil pore water gradually increased with an increase in irrigation time.

3.4. Mechanism of As Mobilization During Irrigation

The mechanism of the mobilization of As in soil by irrigation is shown in Figure 9. Periodic irrigation triggered the redistribution of ions in the soil pore water at different depths as well as the introduction of exogenous substances and DO, which altered the hydrochemical properties of the pore water, leading to periodic fluctuations in the As content in the soil pore water.
In the column experiments, the Fe in the upper layer was in an oxidizing environment and its form was relatively stable, which limited the migration ability of As. Thus, As mobilization in this layer of soil was mainly regulated by competitive adsorption mechanisms. Under the influence of periodic irrigation, the redox conditions of the middle and lower soils exhibited periodic changes and the fluctuation in As content increased with an increase in the number of irrigation events. However, when FA was added to the irrigation water, the fluctuation range of the As(III) content in the pore water remained unchanged, while the fluctuation range of the TAs content increased with an increase in irrigation times. Therefore, the reason for the increasing fluctuation in the As content was because the anoxic conditions inhibited the adsorption of As(V) on the surface of the iron minerals under the condition of FA being added to irrigation water.
The input of FA and the periodic changes in redox conditions also led to morphological changes in Fe and As. A large amount of oxygen entered the soil with the irrigation water. The irrigation water contained FA, which could induce the reduction of Fe(III) to Fe(II) and promote the oxidation of As(III) through a Fenton-like reaction. Partially dissolved Fe and As migrated to the lower soil layer with the irrigation water and then Fe reattached to the surface of the lower soil layer and adsorbed As in the pore water. With the consumption of dissolved oxygen, the redox potential of the middle and bottom soils gradually changed to a lower value. With the reductive dissolution of iron minerals in the reducing environment, As on the surface was released into the pore water. In addition, the decrease in the surface charge of iron minerals could also release As on the surface into the pore water, especially As(V). Compared with FA, PO43− mainly exhibited a competitive adsorption mechanism, resulting in a greater fluctuation range of the As content in the short term. Owing to their different mechanisms of mobilization, PO43− had a greater risk of As mobilization in the short term but FA had a greater ability to mobilize As in the long term.

4. Conclusions

Irrigation activities significantly affected As mobilization and transformation in saline–alkali environments. In an aqueous environment, the adsorption capacity of As(III) by goethite was inhibited in a saline–alkali environment. The inhibitory effect was more significant after the addition of PO43− or FA and the inhibitory effect of PO43− was stronger than that of FA. Notably, anoxic conditions inhibited the adsorption of As(III) by goethite. The surface charge of goethite decreased and the oxidation of As(III) decreased under anoxic conditions. Therefore, the adsorption of TAs by goethite in the presence of PO43− or FA was not affected by a change in the redox environment.
In a soil environment, periodic irrigation caused regular changes in the As content in the soil pore water. Irrigation activities promoted the release of As from the solid phase to the liquid phase due to changes in the hydrochemical properties of the soil pore water. With the addition of different components to the irrigation water, the fluctuation in the As content in the pore water exhibited two trends. When deionized water was used as irrigation water or when PO43− was added to the irrigation water, the fluctuation range of the As content gradually stabilized with an increase in irrigation time. The addition of PO43− led to greater fluctuations in the As content in the short term through competitive adsorption. When FA entered the soil with irrigation water, the reductive dissolution of goethite mediated by FA led to more Fe migrating to the bottom layer and As also migrated in this process. At the same time, the dissolved Fe(II) promoted the transformation of As(III) to As(V). These changes led to the fluctuation range of the As content gradually increasing with an increase in irrigation times, finally exceeding the fluctuation range of the As content under the irrigation of PO43−-containing irrigation water. Therefore, during the long-term irrigation process, FA showed a stronger sustained effect on the mobilization of As and had a more significant effect on the long-term mobilization of As. This study clarified the mobilization behavior and mechanism of As in irrigation activities and provided theoretical support for the development of targeted prevention and control measures. It is suggested that relevant management departments should implement precise agronomic regulation schemes, including the reasonable application of fertilizers and improvements to irrigation systems, to effectively inhibit the activation process of As and reduce its ecological environment risk and impact on agricultural production.

Supplementary Materials

The following supporting information can be downloaded at: https://www.mdpi.com/article/10.3390/w17071058/s1, Text S1: Batch experiments method; Text S2: Quartz sand pretreatment method; Text S3: Synthesis of goethite coated sand; Text S4: FTIR and XPS analysis; Figure S1: Simulated soil column experimental device; Figure S2: XPS full spectrum scanning energy spectrum of goethite, goethite in saline-alkali environment, goethite in saline-alkali + PO43−, goethite in saline-alkali + FA; Figure S3: Goethite O 1s XPS plots (a) goethite, (b) saline-alkali environment, (c) saline-alkali + PO43−, (d) saline-alkali + FA systems; Figure S4: Variation of As(III) in pore water of soil column during alternate irrigation. (a) Deionized water, (b) 20 mg/L PO43−, (c) 500 mg/L FA; Table S1: Batch experimental setup; Table S2: Column experimental setup; Table S3: Table of rate constant fitting; Table S4: Simulated groundwater, simulated irrigation water and soil redox potential and DO at different depths after the column experiment. References [26,27] are cited in the Supplementary Materials.

Author Contributions

H.X.: Conceptualization, Writing—Original Draft, Data Curation, and Formal Analysis; Y.W.: Writing—Review and Editing and Data Curation; J.W.: Methodology, Investigation, Writing—Review and Editing, and Data Curation; X.L.: Data Curation, Formal Analysis, and Methodology; C.C.: Resources, Investigation, and Data Curation; C.Z.: Investigation; Q.T.: Data Curation. All authors have read and agreed to the published version of the manuscript.

Funding

This research was funded by the National Natural Science Foundation of China (No.42107414) and the Basic Research Program of Shihezi University (MSPY202404).

Data Availability Statement

The original contributions presented in this study are included in the Supplementary Materials. Further inquiries can be directed to the corresponding author.

Conflicts of Interest

The authors declare no conflicts of interest.

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Water 17 01058 g001
Figure 1. The As adsorption kinetics of goethite in (a,b) saline and saline–alkali environments, (c,d) saline–alkali + PO43−, and (e,f) saline–alkali + FA systems. As(III) = 1 mg/L, goethite dosage = 1 g/L, and initial pH = 8.5 ± 0.1.
Figure 1. The As adsorption kinetics of goethite in (a,b) saline and saline–alkali environments, (c,d) saline–alkali + PO43−, and (e,f) saline–alkali + FA systems. As(III) = 1 mg/L, goethite dosage = 1 g/L, and initial pH = 8.5 ± 0.1.
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Figure 2. The zeta potential of goethite surfaces after reactions. As(III) = 1 mg/L, goethite dosage = 1 g/L, initial pH = 8.5 ± 0.1, PO43− = 20 mg/L, and FA = 500 mg/L. ** p < 0.001.
Figure 2. The zeta potential of goethite surfaces after reactions. As(III) = 1 mg/L, goethite dosage = 1 g/L, initial pH = 8.5 ± 0.1, PO43− = 20 mg/L, and FA = 500 mg/L. ** p < 0.001.
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Figure 3. FTIR spectra of (a) blank, (b) saline–alkali environment, (c) saline–alkali + PO43−, and (d) saline–alkali + FA systems.
Figure 3. FTIR spectra of (a) blank, (b) saline–alkali environment, (c) saline–alkali + PO43−, and (d) saline–alkali + FA systems.
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Figure 4. Goethite As 3D XPS plots: (a) goethite, (b) saline–alkali environment, (c) saline–alkali + PO43−, and (d) saline–alkali + FA systems.
Figure 4. Goethite As 3D XPS plots: (a) goethite, (b) saline–alkali environment, (c) saline–alkali + PO43−, and (d) saline–alkali + FA systems.
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Figure 5. Goethite Fe 2p XPS plots: (a) goethite, (b) saline–alkali environment, (c) saline–alkali + PO43−, and (d) saline–alkali + FA systems.
Figure 5. Goethite Fe 2p XPS plots: (a) goethite, (b) saline–alkali environment, (c) saline–alkali + PO43−, and (d) saline–alkali + FA systems.
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Figure 6. The As(III) and TAs adsorption kinetics of goethite in (a,b) blank, (c,d) saline–alkali, (e,f) saline–alkali + 20 mg/L PO43−, and (g,h) saline–alkali + FA addition systems under oxic and anoxic conditions. Initial As(III) = 1 mg/L, goethite dosage = 1 g/L, and initial pH = 8.5 ± 0.1.
Figure 6. The As(III) and TAs adsorption kinetics of goethite in (a,b) blank, (c,d) saline–alkali, (e,f) saline–alkali + 20 mg/L PO43−, and (g,h) saline–alkali + FA addition systems under oxic and anoxic conditions. Initial As(III) = 1 mg/L, goethite dosage = 1 g/L, and initial pH = 8.5 ± 0.1.
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Figure 7. Variation in TAs in the pore water of soil columns during alternate irrigation: (a) deionized water irrigation, (b) 20 mg/L PO43− irrigation, and (c) 500 mg/L FA irrigation. The experiment lasted for 20 days and irrigation was performed once every 60 h for 30 min (represented by the gray vertical bar). The pore water samples were collected every 8 h after the simulated groundwater flow stabilized.
Figure 7. Variation in TAs in the pore water of soil columns during alternate irrigation: (a) deionized water irrigation, (b) 20 mg/L PO43− irrigation, and (c) 500 mg/L FA irrigation. The experiment lasted for 20 days and irrigation was performed once every 60 h for 30 min (represented by the gray vertical bar). The pore water samples were collected every 8 h after the simulated groundwater flow stabilized.
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Figure 8. Contents of exchangeable (a) As(III), (b) TAs, (c) Fe(II), and (d) TFe in simulated soil at different depths at the end of the column experiments.
Figure 8. Contents of exchangeable (a) As(III), (b) TAs, (c) Fe(II), and (d) TFe in simulated soil at different depths at the end of the column experiments.
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Figure 9. Mechanisms of As mobilization under irrigation.
Figure 9. Mechanisms of As mobilization under irrigation.
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Xu, H.; Wang, Y.; Wang, J.; Liu, X.; Chen, C.; Zhou, C.; Tian, Q. Irrigation Promotes Arsenic Mobilization via Goethite: Insight from the Perspective of the Solid–Liquid Interface Interaction Process. Water 2025, 17, 1058. https://doi.org/10.3390/w17071058

AMA Style

Xu H, Wang Y, Wang J, Liu X, Chen C, Zhou C, Tian Q. Irrigation Promotes Arsenic Mobilization via Goethite: Insight from the Perspective of the Solid–Liquid Interface Interaction Process. Water. 2025; 17(7):1058. https://doi.org/10.3390/w17071058

Chicago/Turabian Style

Xu, Hong, Yaru Wang, Jiankang Wang, Xin Liu, Cuizhong Chen, Chang Zhou, and Qingyuan Tian. 2025. "Irrigation Promotes Arsenic Mobilization via Goethite: Insight from the Perspective of the Solid–Liquid Interface Interaction Process" Water 17, no. 7: 1058. https://doi.org/10.3390/w17071058

APA Style

Xu, H., Wang, Y., Wang, J., Liu, X., Chen, C., Zhou, C., & Tian, Q. (2025). Irrigation Promotes Arsenic Mobilization via Goethite: Insight from the Perspective of the Solid–Liquid Interface Interaction Process. Water, 17(7), 1058. https://doi.org/10.3390/w17071058

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