Reduction and Immobilization of Movable Cu²⁺ Ions in Soils by Fe₇₈Si₉B₁₃ Amorphous Alloy

Abstract

The Fe-based amorphous alloy (Fe₇₈Si₉B₁₃^AP) is applied to the remediation of copper contaminated soil for the first time. The dynamic process of conversion of movable Cu to immobilized forms in the soil system is analyzed. In addition, the dynamic process of form transformation of Cu²⁺ ions in the soil system is analyzed. The morphology and phase composition of the reaction products are characterized by scanning electron microscopy (SEM) and X-ray diffraction analysis (XRD). Finally, the feasibility of recovering residual stabilizer particles and attached immobilized copper by the magnetic separation process is discussed. The results show that the apparent reaction rate constant of Fe₇₉Si₉B₁₃^AP with Cu²⁺ ions is higher than that of zero valent iron (ZVI) at all the experimental temperatures. According to the Arrhenius formula, the apparent activation energy of the reaction of Fe₇₈Si₉B₁₃^AP and ZVI with Cu²⁺ ions is 13.24 and 19.02 kJ/mol, respectively, which is controlled by the diffusion process. The lower apparent activation energy is one of the important reasons for the high reaction activity of Fe₇₈Si₉B₁₃^AP. After 7 days of reaction, a continuous extraction of the experimental soil shows that the main form of copper in the immobilized soil is Cu and copper combined with iron (hydroxide) oxide, and there is almost no soluble copper with a strong mobility, which effectively reduced the bioavailability of copper in the soil. The magnetic separation results of the treated soil show that the recovery rates of immobilized copper in Fe₇₈Si₉B₁₃^AP and soil are 47.23% and 21.56%, respectively, which reduced the content of iron and copper in the soil to a certain extent. The above experimental results show that Fe₇₈Si₉B₁₃^AP is a promising new material for the remediation of heavy metal contaminated soils, and provides more new references for the application of amorphous alloys in the field of remediation of water and soil contaminated by heavy metals and organic matter.

1. Introduction

Soil heavy metal pollution has a large base and strong mobility, which has become an important factor restricting national sustainable development and endangering human health. Copper is one of the essential trace elements for the human body, but the high content of copper has adverse effects on animals, plants, and the human body. Baryla et al. [1] found that excessive copper can cause the concentration difference between the two sides of the plasma membrane, resulting in the flow of potassium ions out of the cell membrane, thus damaging the plasma membrane of plant roots. Dijendra et al. [2] reported that when mice were fed with a diet containing copper (15 mg/kg), their liver morphology and function would show symptoms of poisoning, and the decrease of superoxide dismutase activity would lead to mitochondrial dysfunction, and eventually lead to the death of hepatocytes. Choudhary et al. [3] reported that under the stress of heavy metals such as Cu, the growth of cyanobacteria is inhibited by the increase in the amount of proline, malondialdehyde, and superoxide dismutase. The main source of copper pollution is exogenous copper, such as waste water discharged from mining and the storage of waste ore. These open-pit wastes continue to diffuse into the surrounding soil under the effects of weathering, acidification and rainfall, resulting in copper substances that enter the soil [4,5]. Water-soluble and exchangeable copper has good solubility and mobility, which can enter the plant through the root of the plant, affecting the normal metabolic activity of the plant [6,7,8]. From the point of view of ecological risk assessment, it is valuable to transform movable Cu²⁺ ions into an immobilized state with low bioavailability [9]. In recent years, researchers have developed a series of soil stabilizers for immobilization remediation of heavy metal contaminated soils. For example, Zhang et al. [10] studied the immobilization performance and mechanism of novel phosphorus-modified biochars on soil heavy metals. The experimental results show that by forming precipitates or complexes with Cu²⁺ ions, the phosphorus compounds in surface modified biochar can reduce the leaching amount of Cu²⁺ ions by 2–3 times and promote the conversion of Cu²⁺ ions from an acid solution to a stable state. Viglašová E. et al. [11] studied the adsorption process of Cu²⁺ by natural bentonite, Illite, and montmorillonite under different experimental conditions. The results show that the adsorption is realized by the ion exchange process, achieving a balance within 15 min. The adsorption capacity of the three materials for Cu²⁺ is bentonite 0.0658 mg/g, Illite 0.0663 mg/g, and montmorillonite 0.0667 mg/g, respectively. Cui et al. [12] remediated copper contaminated soil by adding 1–5% hematite to hydroxyapatite as a soil stabilizer. After 42 days of the flooding culture experiment, the soil redox potential decreases, pH increases, and extractable copper decreases by 53.7%. The immobilization process did not cause a secondary pollution to the environment. Therefore, immobilization is an efficient and economical method for soil remediation.

Among the many chemical immobilization methods, it is one of the most commonly used methods to immobilize heavy metal ions by the redox process, and then adsorb them to their corrosion products by gluing or precipitation [13,14]. Soil stabilizers should be reasonably selected according to different types of pollutants in soil immobilization remediation. Copper has a high electrode potential (Z(Cu²⁺/Cu⁰) = +0.34 V). It is an effective method to convert movable Cu²⁺ ions into immobilized copper by the redox reaction. Zero-valent iron (ZVI) is a low-cost environment-friendly reductant (Z(Fe²⁺/Fe⁰) = −0.44V), which is widely used in the degradation of environmental pollutants [15,16,17]. Frick et al. [18] remediation of chromated copper arsenates (CCA) contaminated soil with the ZVI-biochar combined stabilizer reduced the contents of water-soluble Cr, Cu, and As by 45%, 45%, and 43%, respectively. However, the treatment effect of using ZVI or biochar alone is not ideal. The main reason is that the adhesion of the reaction products to the surface of ZVI hinders the contact between Cu²⁺ ions and fresh ZVI surface, thus blocking the progress of the reaction. Kumpiene et al. [19] used ZVI to reduce the mobility and bioavailability of copper and arsenic in soils. The content of As⁵⁺ ions decrease by 99%, but the content of Cu²⁺ ions in the soil is still high, which has a certain residual toxicity to the soil. It can be seen that the immobilization effect of traditional stabilizers on Cu contaminated soil is not ideal. Therefore, there is an urgent need to develop new stabilizers with excellent performance and low cost in the field of soil remediation.

Amorphous alloy (MGs) is a thermodynamic metastable alloy formed by rapid solidification of molten metal at a cooling rate of more than 379 K/s [20]. The atomic arrangement of its components is far from the equilibrium position, showing the characteristics of long-range disorder and short-range order. This unique stacking structure of disordered atoms determines that it has excellent properties incomparable to crystalline alloys [21,22]. In 2010, Zhang et al. [23] used the Fe-Mo-Si-B amorphous alloy to degrade direct blue 2B for the first time. The results show that the reaction rate of direct blue 2B degradation of the Fe-Mo-Si-B amorphous alloy is four times higher than that of the same component crystalline alloy. This opens the research on the application of MGs in the field of environmental pollutant degradation. Zhang et al. [24] compared the reduction and removal ability of Cu²⁺ ions between Fe₇₉Si₁₃B₈^AR and 300 mesh iron powder. The results show that the normalization rate constant of the Fe₇₉Si₁₃B₈^AR surface area is 167 times higher than that of the 300 mesh iron powder. Liang et al. [25] used Fe₇₉Si₁₃B₈^AR/Na₂S to remove As⁵⁺ from the groundwater. As⁵⁺ is reduced to As³⁺ by Fe₇₉Si₁₃B₈^AR and then combined with S²⁻ to form the As₂S₃ precipitation. The removal rate of As⁵⁺ is more than 95% after 60 min. Some scholars attribute the high reactivity of MGs to the fact that the metastable structure of disordered arrangement of atoms makes the surface diffusion rate of MGs millions of times higher than that of the bulk phase [26,27].

Inspired by the previous research results, the Fe₇₉Si₁₃B₈^AP is applied to the immobilization and remediation of copper contaminated soil for the first time in this paper. The kinetic process of conversion of movable Cu to immobilized forms in soil and the effect of product morphology on the reaction rate are studied. Then, the feasibility of recovering residual Fe₇₉Si₁₃B₈^AP particles and copper immobilized on their surfaces by the magnetic separation process is explored. Providing more new references for the application of MGs in the field of environmental pollutant degradation is expected.

2. Materials and Methods

2.1. Materials

The reagents used in this experiment include CuSO₄·5H₂O, HCl, NaOH, MgCl₂, HNO₃, CH₃COONH₄, CH₃COOH, NH₄OH∙HCl, and H₂O₂. These reagents were purchased from Lanzhou Moer Chemical Products Co. Ltd. (Lanzhou, China). ZVI powders were purchased from China Zhongye Chemicals Co. Ltd. (Xingtai, Hebei, China), which have a mean grain size (d₅₀) of ~1 μm. The price of the powders is CNY 740/kg. Fe₇₈Si₉B₁₃^AP with an average particle size of 10 μm were purchased from Beijing Yijin Chemicals Co. Ltd. (Beijing China). The price is CNY 770/kg.

2.1.1. Soil Sample Collection

Surface loess soil without copper contamination was collected from Lanzhou suburb of Gansu Province (36°03′ N, 103°49′ E) and sifted through a 2 mm sieve. After screening, it was washed with distilled water three times to remove soluble compounds and suspended particles. The washed soil was air-dried and then passed through the 2 mm sieve again, and the obtained soil was put into a sample bag for backup.

2.1.2. Preparation of Contaminated Soil

The soil environmental quality standard of the People’s Republic of China stipulates that the upper limit of copper contaminated soil is 200 mg/kg [28]. In addition, the maximum content of copper in the soil around the Baiyin copper smelter, known as “Copper City of China”, is 658 mg/kg [29]. As a result, the content of copper in the contaminated soil was determined to be 600 mg/kg. The specific steps were as follows: Mixing a 1 L CuSO₄·5H₂O solution of 60 mg/L with 100 g washed soil and then mechanically stirring until air-dried. In order to determine the uniformity of Cu in soil when different volumes of water were added to soil samples of the same quality. The concentration of Cu²⁺ ions were inversely proportional to the volume of distilled water, indicating that Cu has been uniformly dispersed in the soil.

2.2. Immobilization of Cu²⁺ Ions in the Soil: Batch Test

The immobilization effect of Fe₇₉Si₁₃B₈^AP and ZVI on Cu²⁺ ions in the soil was evaluated under the experimental temperature of 293, 303, 313, and 323 K. In order to remove the surface oxide layer as much as possible, the stabilizer was leached with 2% HNO₃ before each batch of test. Each group of experiments was carried out in a series of 100 mL jars, and the preparation of slurry in each jar was as follows: 5 g of copper contaminated soil was mixed with 50 mL deionized water, then shaken in a water bath shaker for 5 min to completely wet the soil to form a soil suspension. According to the method of Shaaban et al. [30], the initial pH of the reaction system was adjusted to 6 by HCl and NaOH. In general, increasing the amount of stabilizer is beneficial to the immobilization of heavy metal ions [31]. However, when the content of iron in the soil is more than 5%, it will inhibit the growth of plants [32]. As a result, the dosage of Fe₇₈Si₉B₁₃^AP and ZVI was selected as 5% of the soil mass, that is, 0.25 g was added to the jar. When the soil reacted for 1, 2, 4, 8, 16, 24, 36, and 48 h, it was continuously extracted according to the method of Ure et al. [33]. The content of Cu²⁺ ions in each extract was determined by inductively coupled plasma atomic emission spectrometry (ICP-AES). After the remaining two slurries continued to react for 7 days, one of them was continuously extracted according to the above-mentioned method. The other slurry carried out magnetic separation of the treated soil according to the method described by Guan [34] to separate the immobilized products and stabilizer particles.

Table 1. Continuous extraction of metals in the soil.

Index	Speciation	Extraction Agents and Conditions
SE	soluble and exchangeable	0.01 mol/L MgCl2, 50 mL; 0.11 mol/L acetic acid (CH3COOH), 20 mL
OX	reducible state	0.5 mol/L NH4OH·HCI, 20 mL
OM	oxidizable state	Hydrogen peroxide (H2O2) at 85 °C and 1 mol/Lammonium acetate (CH3COONH4)
RE	residual	Aquaregia (HCl:HNO3, 1:3 v/v), 160 °C

shows all the reagents of the continuous extraction method.

In order to accurately analyze the copper in the immobilization reaction and the rate control process of the reaction of Fe₇₈Si₉B₁₃^AP and ZVI with Cu²⁺ ions in soil. The experimental results were fitted by the pseudo-first-order kinetic Equation (1) linear fitting:

C_t/C₀ = exp(−k_obst)

(1)

where t is the reaction time (h); C_t is the concentration of SE at the current time (mg/kg); C₀ is the concentration of SE at the initial time (mg/kg); and k_obs is the observed pseudo-first-order reaction rate constant.

The reaction activation energy (ΔE) is an important parameter to characterize the kinetic process of the chemical reaction. ΔE can be obtained according to the Arrhenius-type equation (Equation (2)):

lnk_T = −ΔE/RT + lnA

(2)

where k_T is the k_obs obtained by Equation (1) at different temperatures. R is the gas constant, R = 8.314 J/(mol·K); A is the constant.

2.3. Characterization Techniques

The surface morphology changes of Fe₇₈Si₉B₁₃^AP and ZVI before and after the reaction were observed by the scanning electron microscope (SEM, JSM-6700, JEOL, Akishima, Japan). The structural characterizations of the stabilizer and the composition of the reaction product were identified by X-ray diffraction (XRD, D8 ADVANCE, BRUKER, Karlsruhe, Germany) with Cu-Kα radiation (40 kV, 150 mA).

3. Results and Discussion

3.1. Characterization of Fe₇₈Si₉B₁₃^AP and ZVI

Figure 1 shows the structural and morphological differences between Fe₇₉Si₁₃B₈^AP and ZVI before the soil remediation experiments. In Figure 1a, the crystalline α-Fe phase is observed in the X-ray diffraction spectrum of ZVI, and there is no crystallization peak of iron oxide, indicating that the purity of ZVI is high and the surface is not oxidized. There is a diffuse scattering peak in the X-ray diffraction spectrum of Fe₇₉Si₁₃B₈^AP around 45°, and there is no crystallization peak in the rest of the X-ray diffraction spectrum, indicating that it is amorphous. Since there are no lattice fringes and lattice diffraction spots in the TEM and SAED images in Figure 1b, this conclusion is further confirmed. Figure 1c,d shows the SEM images of Fe₇₉Si₁₃B₈^AP and ZVI, respectively. The two kinds of powders have a spherical structure with a smooth surface and there is no adhesion between the particles. The particle size of Fe₇₈Si₉B₁₃^AP is 10–20 μm, and that of ZVI is 1–3 μm.

3.2. Cu²⁺ immobilization by Fe₇₈Si₉B₁₃^AP and ZVI

Figure 2a,b shows the exponential decay curve of the normalized concentration of Cu²⁺ reacting with Fe₇₈Si₉B₁₃^AP and ZVI at different temperatures. The reaction of each experimental group basically ended within 10 h, and there is no secondary dissolution after 10 h, indicating that the reduction products could stably exist in the reaction system. Figure 2c shows the k_obs obtained by fitting the quasi-first-order kinetic equation (Equation (1)). With the increase of the reaction temperature from 293 to 323 K, the k_obs of each experimental group shows an upward trend. The k_obs of Fe₇₈Si₉B₁₃^AP increases from 0.955 to 1.568 h⁻¹. The k_obs of ZVI is significantly lower than that of Fe₇₈Si₉B₁₃^AP. Generally speaking, the smaller the particle size and the higher the content of the reactant, the more conducive it is to speeding up the reaction [35,36]. However, in this experiment, the k_obs of ZVI with a relatively small particle size and higher iron content reacting with Cu²⁺ are lower than Fe₇₈Si₉B₁₃^AP with a relatively large particle size and lower iron content, which may be due to the metastable state of MGs caused by the structure.

As we all know, the activation energy (ΔE) reflects an important kinetic parameter of the difficulty of the chemical reaction [37]. On the basis of the above k_obs, the ΔE difference between crystalline and amorphous iron-based materials is further analyzed. The illustrations in Figure 2a,b are shown as straight lines fitted by the Arrhenius-type equation (Equation (2)) for the k_obs which was reacted by Fe₇₈Si₉B₁₃^AP and ZVI with Cu²⁺ at different temperatures. The correlation coefficients R² are all above 0.99. The ΔE of Fe₇₈Si₉B₁₃^AP is 13.24 kJ/mol, while the ΔE of ZVI which is significantly higher than that of Fe₇₈Si₉B₁₃^AP is 19.02 kJ/mol. In other words, Fe₇₈Si₉B₁₃^AP needs less energy to transform from a normal state to an active state where chemical reactions easily occur. This is due to the higher Gibbs free energy of MGs than crystalline alloys [38]. It also explains why the k_obs of Fe₇₈Si₉B₁₃^AP is higher than that of ZVI. The ΔE of the general chemical reaction is between 60–250 kJ/mol. The ΔE of the reaction between iron-based materials and Cu²⁺ is lower than that of 60 kJ/mol [39]. According to the diffusion theory, when Δ E is within the range of 8–22 kJ/mol, the reaction is controlled by an external diffusion [40]. The ΔE of the two materials is in this range, indicating that the adsorption of soil to Cu²⁺ hinders the effective contact between Cu²⁺ and the reaction surface.

3.3. The Fractions of Different Cu Forms in Soil after a 200 h Treatment

The existing form of heavy metals in soil is one of the important indicators for an environmental risk assessment. In order to determine the content of each component of the four forms of copper in soil after Fe₇₈Si₉B₁₃^AP and ZVI treatment, the method of Ure et al. [32] is used to extract the soil continuously. Figure 2d shows the influence of temperature on the fractions of different Cu forms after a 200 h treatment. The fractions of soluble and exchangeable (SE) in untreated soil is the highest, accounting for 68.94% of the total copper content. The second is the reducible state (OX) accounting for 16.50%. The fractions of oxidizable state (OM) and residual state (RE) is 9.14% and 5.41%, respectively. There is almost no SE in soil after Fe₇₈Si₉B₁₃^AP and ZVI treatment. Cu²⁺ is converted to OX and OM by the reduction and adsorption of iron-based materials. The source of OX is the Cu²⁺ combined with iron oxide with high surface energy and low coordination number [41]. The fractions of OX₁ in soil treated with Fe₇₈Si₉B₁₃^AP at different temperatures is 28.23%, 25.52%, 23.54%, and 22.36%, respectively. The fractions of OX₂ in the soil after ZVI treatment at different temperatures is 44.24%, 42.25%, 39.24%, and 36.20%, respectively. It can be seen that the fractions of OX₂ are significantly higher than that of OX₁ at all temperatures. This is due to the fact that the particle size of ZVI is much smaller than that of Fe₇₈Si₉B₁₃^AP, which provides more active sites for the adsorption reaction and is more conducive to the binding of Cu²⁺ ions. Since the extraction of OM is carried out in acidic H₂O₂, this process can not only oxidize the organic matter, but also reoxidize the reduced Cu⁰ to Cu²⁺ ions. Therefore, OM includes organically bound copper and elemental copper. The fractions of OM₁ in the soil after Fe₇₈Si₉B₁₃^AP treatment is 64.62%, 68.04%, 70.59%, and 71.37%, respectively. The fractions of OM₂ in soil treated with ZVI is 47.62%, 51.26%, 55.27%, and 57.34%, respectively. With the increase in temperature, the fractions of OM₁ and OM₂ show an upward trend, and the values of each group of OM₁ are higher than OM₂. Since the organic bound state does not participate in the immobilization process of Cu²⁺, the increase of OM fraction is caused by the conversion of Cu²⁺ ions to Cu⁰. As a result, the ability of Fe₇₈Si₉B₁₃^AP to reduce Cu²⁺ ions is higher than that of ZVI. Once the Cu²⁺ ions are reduced to Cu⁰, they cannot be dissolved even if they are leached by acid rain for a long time. Combined with the change of OX fraction, it can be determined that the process of immobilization and remediation of copper contaminated soil by two kinds of iron-based materials is mainly the reduction and adsorption reaction, and the increase in temperature is beneficial to the reduction and immobilization. The fraction of RE in the soil is about 6% before and after the treatment. This is due to the fact that the process of replacing silicon, aluminum, and other atoms in the soil lattice by copper is too slow to occur in this experiment.

3.4. Reaction Mechanism Analysis

According to the previous kinetic analysis, it is known that the reduction of Cu²⁺ ions by Fe₇₈Si₉B₁₃^AP and ZVI is controlled by the diffusion process. Therefore, the morphology and phase of the surface products directly affect the movement of Cu²⁺ ions to the reaction interface. The morphology and phase composition of the surface products before and after the reaction are analyzed by SEM and XRD. Figure 3a shows the morphology of the product on the surface of Fe₇₈Si₉B₁₃^AP. The originally smooth surface is seriously corroded and the product layer on the surface is thin. In order to observe the morphology of the product more clearly, the B region is magnified and displayed in Figure 3b. The product layer presents a three-dimensional flower-like structure composed of nano-lamellar intercalation perpendicular to the reaction interface, and the interstitial pores provide a channel for the diffusion of Cu²⁺ ions to the reaction interface. Figure 3c observed that ZVI is completely surrounded by a dense product layer. Some of the products fall off from the matrix, which is difficult to extract in the later magnetic separation process. Figure 3d shows a high-resolution image of the product layer overlaid on the ZVI. The morphology of the product is dendritic and grows randomly in all directions, which hinders the contact between Cu²⁺ ions and ZVI, which is the main reason for the low activity of ZVI. With regard to the explanation that the morphology of the products of amorphous and crystalline alloys affects the reaction rate, some scholars think that there is an active “liquid-like” layer on the surface of the amorphous alloy [42]. Figure 4 shows a Bragg peak of Cu⁰ in the XRD pattern of soil treated with Fe₇₈Si₉B₁₃^AP and ZVI. The products of Cu²⁺ ions reduction by two kinds of iron-based alloys are Cu⁰. Since the standard electrode potential of copper E(Cu/Cu²⁺) is higher than the standard electrode potential of hydrogen E(H₂/H⁺), even long-term acid rain leaching cannot dissolve Cu⁰ again, effectively reducing the potential risk of copper in the soil.

3.5. Magnetic Separation

The above experimental results show that Fe₇₈Si₉B₁₃^AP and ZVI can effectively reduce the potential risk of copper in the soil. However, when conditions such as soil pH change, the immobilized copper (such as OX) has the possibility of a secondary release. In addition, iron-based alloys added to the soil will have adverse effects on plant growth and microbial community [43]. In this study, the magnetic separation process is used to recover the stabilizer particles with immobilized products attached to the surface. The extraction result is shown in Figure 5. Since the iron content of ZVI is higher than that of Fe₇₈Si₉B₁₃^AP, the recovery rate of ZVI is relatively high, reaching 68.85%. The recovery rate of Fe₇₈Si₉B₁₃^AP is slightly lower, only 47.23%. After magnetic separation, the contents of Fe₇₈Si₉B₁₃^AP and ZVI in soil are 26.36 and 17.07 g/kg, respectively, which are less than the upper limit value. It can be seen that the remediation process does not cause a secondary pollution to the soil. Figure 5b shows that the recovery of copper is not satisfactory. The recovery rate of copper in the soil treated by Fe₇₈Si₉B₁₃^AP is 21.56%, and the remaining total copper content is 470.64 mg/kg. The recovery rate of copper in the soil treated by ZVI is 32.42%, and the remaining total copper content is 405.48 mg/kg. It is found that the recovery rate of copper in ZVI treated soil is higher than that of Fe₇₈Si₉B₁₃^AP, which could be explained by the SEM image in Figure 3. The products of ZVI and Cu²⁺ ions are attached to the surface of ZVI, while the adhesion amount of Fe₇₈Si₉B₁₃^AP surface products is less. Since Cu⁰ is not ferromagnetic, copper attached to the surface of the Fe-based alloy can only be extracted in the process of magnetic separation. Therefore, the recovery rate of copper in the soil treated with Fe₇₈Si₉B₁₃^AP is low. It can be seen from Figure 2d that when the temperature is 323 K, the content of OX and Cu⁰ in the soil after the Fe₇₈Si₉B₁₃^AP treatment is 22.36% and 54.87%, respectively. Under ideal conditions, the recovery rate of copper can reach 77.23%, while the experimental value only reaches 42.23%, the difference between the two is 35%. This is due to the shedding of the product layer and the adsorption of copper by the soil. In order to solve the problem of low magnetic separation rate of copper, the combination of strengthening loose products and stabilizers is being considered in the follow-up experiments.

4. Conclusions

In this paper, ZVI is used to compare and analyze the immobilization remediation effect of Fe₇₈Si₉B₁₃^AP on copper contaminated soil. The kinetic process of conversion of movable Cu to immobilized forms in the soil system is studied. The feasibility of the magnetic separation process for removing immobilized copper from the soil is verified. The main results and conclusions are summarized as follows:

Compared with the traditional stabilizer ZVI, Fe₇₈Si₉B₁₃^AP with a larger particle size and lower iron content can reduce movable copper in the soil more efficiently. According to the fitting analysis of the Arrhenius-type equation, the apparent activation energy of the reaction of Fe₇₈Si₉B₁₃^AP and ZVI with Cu²⁺ is 13.24 and 19.02 kJ/mol, respectively, which is controlled by the diffusion process.

The main forms of copper in the immobilized soil are OX and OM. There is almost no SE in the immobilized soil. The bioavailability of copper in the soil is effectively reduced.

The difference in the atomic arrangement between Fe₇₈Si₉B₁₃^AP and ZVI results in a great difference in the structure of the product layer. The surface products of Fe₇₈Si₉B₁₃^AP have a nano-pore structure, which provides a channel for the movement of Cu²⁺ to the reaction interface. However, ZVI is wrapped in a dense product layer, which hinders the sustainable progress of the reaction.

The recoveries of Fe₇₈Si₉B₁₃^AP and ZVI by the magnetic separation process are 47.23% and 68.85%, respectively, and the recoveries of copper in soil treated by Fe₇₈Si₉B₁₃^AP and ZVI are 21.56% and 32.42%, respectively. The content of iron and copper in soil is reduced to a certain extent, which has a certain engineering application value.

At present, the production technology of the amorphous alloy is quite mature, and the market price is close to that of micron ZVI, which has the basic conditions for an engineering application. The above research results show that the Fe-based amorphous alloy has a high activity in the treatment of copper-contaminated soil, which solves the problem of low treatment efficiency and material waste caused by ZVI passivation. The purpose of this paper is to provide some inspiration for the application of derived amorphous alloys to the field of soil remediation.