Abstract
Eroded mine soils are a serious environmental problem in many parts of the world. Coal gasification slag, a byproduct of the coal gasification process, was modified with sludge-derived amino acids for the purpose of remediating eroded mine soils. Amino acid–modified magnetic coal gasification slag (AMS) was evaluated for its ability to decrease heavy metal erosion, improve soil aggregate structure and plant growth, and enhance microbial enzymatic activity. Alhagi sparsifolia (A. sparsifolia) was grown in greenhouse experiments under different AMS concentrations (0%, 1%, 5%, and 10%). The growth rate of A. sparsifolia increased with increasing AMS concentrations, and the optimal portion of AMS in soil was 7.5%. A. sparsifolia grown in mine soil amended with 7.5% AMS significantly (p < 0.05) decreased bioavailability of all heavy metals studied in this work (except for iron [Fe]) and increased (p < 0.05) activity of three soil enzymes: phosphatase (PA), urease (UA), and β-glucosidase (β-GA). In addition, the reusability of AMS was very good. AMS-amended mine soil may interact with growing plants to cycle needed nutrients while immobilizing toxic metals. AMS is thus a promising amendment for remediating eroded mine soils.
Introduction
Faced with increasingly severe soil erosion problems in mining areas, more and more soil management techniques had been developed, such as soil washing, solidification, and electroremediation techniques (Muhammad et al. 2017). Many of these techniques needed to use chemical additives, such as lime, phosphate (PO43−), sulfur (S), and iron (Fe) salts (Jiang et al. 2018a). These chemical additives played important roles in improving soil properties (Liu et al. 2018a). However, long-term excessive use of the chemical additives had caused the degradation of soil quality and the pollution of soil environment (Ning et al. 2017). Therefore, it is of great importance to find new reagents that are viable replacements for these chemical agents—replacements that are nontoxic, abundant, recyclable, and easy to produce.
Coal gasification slag (CGS) is a solid waste from coal gasification process. It has high proportions of calcium (Ca), iron (Fe), and silica (SiO2), and small amounts of aluminum (Al), potassium (K), zinc (Zn), and molybdenum (Mo) (Li et al. 2017). Recently, CGS has received more and more interests in various fields due to its unique properties, such as high surface area, low cytotoxicity, and cost effectiveness (Xiang et al. 2018b). CGS has been used to govern the stabilization and accumulation of heavy metals in soil (Francis et al. 1985; Seul et al. 2013). In addition, the capability of CGS to hold various functional molecules has opened the door for numerous applications, and these multifunctional materials have been widely used in various fields, such as environmental protection and adsorbent materials (Liu et al. 2018b; Liu et al. 2019). However, CGS may have traces of toxic elements (e.g., chromium [Cr], copper [Cu]) and soluble salt, and this property restricts its land application (Liu et al. 2018b). If the CGS can be endowed with magnetism, it will be easy to recycle (Xie and Ma 2018), and then the risks of CGS will be reduced substantially. In addition, the magnetic materials can play a positive role in the rehabilitation of soil eroded by heavy metals and improvement of soil aggregate structure (Deng et al. 2017).
On the other hand, surfactants can promote the desorption of contaminants from soil and enhance the removal of heavy metals by complexation, dissolution, and ionic exchange (Mao et al. 2015). Amino acids with carboxyl and amino groups are widely available, and they can improve soil enzyme activity by increasing the level of organic matter in soil and improving the physical and chemical properties of soil (Lv et al. 2018). Soil enzyme activity is an important indicator of soil quality, and it is also a sensitive indicator of variation in soil fertility, especially for nutrient availability (Li et al. 2017). Therefore, amino acids are beneficial for the improvement of soil quality. Tan et al. (2010) indicated that excess sewage sludge contained abundant proteins (about 2,602 ± 134 mg L−1). These proteins can be easily degraded to small molecule amino acids (Su et al. 2014). Therefore, excess sewage sludge can be used as a raw material for amino acid production.
Alhagi sparsifolia (A. sparsifolia) is a plant that can tolerate drought, low fertility, and salinity. A. sparsifolia is hearty and highly productive. In addition, A. sparsifolia can not only resist sand but also fix sand (Li et al. 2014). Therefore, A. sparsifolia is a suitable plant for the control of soil erosion and desertification in a mining area.
To the best of our knowledge, until now, studies on the effect of sludge amino acid–modified magnetic coal gasification slag (AMS) amendment on A. sparsifolia growth, metal availability, and soil enzyme activity have not been reported from the literature review. If the risk of heavy metals can be controlled and plant growth can be improved, the application of the sludge AMS will achieve the dual benefit of the ecological restoration of mining areas and the treatment and disposal of wastes. Therefore, the objectives of this investigation were to (1) determine the effect of AMS on A. sparsifolia growth and metal availability, and (2) reveal the ability of AMS to improve soil enzyme activity.
Materials and Methods
Collection and Preparation of Materials. The original mine soil (MS) was collected at a depth of 0 to 20 cm from an abandoned coal mine in the Jurassic Coalfields. A soil profile pit was dug at the sampling site. The pit was divided into three layers—the first layer was the surface, the second layer was within 10 cm of the surface, and the third layer was within 20 cm of the surface. Soil samples were taken from each different soil horizon down to the third horizon, and four samples were taken at each layer with three replicates. The samples were naturally dried and ground to pass through a 2 mm sieve for determining soil pH and texture. A quarter of the sieved sample was further crushed to pass through a 1 mm sieve for physico-chemical analysis. CGS (pH = 9.6 ± 0.3) was collected from Daliuta, Shaanxi Province, China, and the main compositions are shown in table 1.
For extraction of sludge amino acids (SAA), excess sewage sludge was collected from the municipal waste water treatment plant of Yulin, China. Moisture content of the sludge was adjusted to 98% (w w−1) by moving supernatant or diluting with ultrapure water, and then SAA was produced based on a previously published report by Xiang et al. (2018a). Ion exchange method was used for the separation and purification of SAA (Zhang et al. 2012). The extracted SAA was sampled and characterized with a Fourier-transform infrared spectrometer (FTIR; IRPrestige-21) under the range of 400 to 4,000 cm−1 using KBr pellet. After that, the extraction solution of SAA was dried at 30°C for one week to produce SAA powder.
For synthesis of AMS, first, the iron (II, III) oxide (Fe3O4) nanoparticles were synthesized according to the report published by Shan et al. (2014). Iron (III) chloride hexahydrate (FeCl3·6H2O; 2 g) was dissolved in 20 mL ethylene glycol solvent. Under magnetic stirring, 6 mL of hydrazine hydrate (N2H4·H2O) (80% by weight) and 3 g of sodium hydroxide (NaOH) were added. After stirring, the mixture was sealed in a 50 mL teflon-lined stainless steel autoclave and heated at 90°C for 10 hours. After being cooled to room temperature, the black powders were washed five times with deionized water and vacuum dried at 55°C ± 5°C. Second, the magnetic Fe3O4/CGS (MCGS) was prepared. CGS was dipped in 40% by weight NaOH solution (liquid-solid ratio 4:1), and the mixture was heated in 600°C for 3 hours. The solid (A) was cleaned, filtered, and dried at 100°C. A solution that contained 18.69 g of solid (A) and 31.70 g of NaNO3 was prepared by dissolution of these materials into 300 mL of distilled water, and then the prepared Fe3O4 (2.14 g) was added to the solution. The mixed solution was rapidly stirred with a magnetic mixer in a 65°C water bath for 15 seconds. Following that, the mixed solution was slowly added into 400 mL of alkaline solution containing sodium carbonate (Na2CO3; 6.69 g) and NaOH (7.86 g) and centrifuged at 1,500 rpm for 60 minutes. The suspension was aged at 100°C for 10 hours under stirring. The precipitate was collected by filtration and then washed with decarbonated water until the effluent solution was neutral. The sample was dried at 65°C (oil bath) for two days and then milled and sieved (200 mesh). MCGS with layer structure was obtained. Third, to synthesize AMS, the SAA intercalated with MCGS (Chuang et al. 2010). The 100 mL of 0.1 M SAA and 2 g of MCGS were mixed with carbon dioxide (CO2) free of double-distilled water and shaken for three days at 55°C. The pH of the mixture was about 8.9, and later the pH was adjusted to 8.5 with 0.05 M HCl solution. The mixture was washed with 100 mL of 50% alcohol and centrifuged several times at 13,400 × gravity for 10 minutes in order to remove residual SAA. The remaining AMS solid was freeze dried and then milled and sieved (100 mesh). The obtained AMS was stored under dark and cool conditions.
Greenhouse Experiment. Treatments consisted of 0% (control), 2.5%, 5%, 7.5%, and 10% (w w−1) soil amendments for SAA, MCGS, and AMS each, and 1% (w w−1) monopotassium phosphate (KH2PO4) as fertilizer into prepared 2 mm sieved MS, and artificial soil (AS) with soil amendments were obtained. The prepared samples (10 kg of every pot) were put into clay pots (50 cm height and 30 cm diameter). A. sparsifolia plants were collected from the same suburb of Yulin in the same condition (including height, biomass, age, and good health), and then the plants were planted in the prepared pots. These pots were placed in a greenhouse (55% ± 8% relative humidity, 60% ± 2% of the water-holding capacity, 27°C ± 5°C, and 10 to 12 hours of natural daylight for 90 days). The pots for each planting experiment were randomly divided into five groups, including one control group (without A. sparsifolia) and four replicates (with A. sparsifolia), and numbered in sequence. Ninety days later, the plants were separated from the pots for chemical analysis. Cation exchange capacity (CEC), calcium carbonate (CaCO3), moisture content (MC), and organic matter (OM) were determined with standard laboratory procedures (Lu 1999). Heavy metal availability was measured by using a four-stage Community Bureau of Reference sequential procedure (BCR) (Xu et al. 2012).
Analysis. The fraction of water-stable aggregates (WSA) was measured with the wet-sieving method (Six et al. 2002). The particle-size distribution of WSA was 1 to 3 mm, 0.25 to 1 mm, 0.05 to 0.25 mm, and below 0.05 mm. One hundred grams of air-dried raw soil sample was placed on 2 mm sieves and immersed in water for 10 minutes. After that, the soil sample was graded by Aggregate Analyzer (TPF-100, Hangzhou, China) with the grading frequency of 30 strokes per minute (SPM) and the grading time of 5 minutes. After wet screening, the soil aggregates in each particle-size distribution were collected and weighed after oven drying (50°C).
For the enzyme activity analysis, three enzymes were tested: phosphatase (PA; EC 3.1.3.2), urease (UA; EC 3.5.1.5), and β-glucosidase (β-GA; EC 3.2.1.21). Soil UA was assayed by estimating the ammonium (NH4+) produced by use of colorimetric methods after adding urea to soil subsamples (Allison 2001), while PA and β-GA were measured based on the method proposed by Li (2017). The standard solutions and substrate were freeze dried on the plates and stored at −20°C. The enzyme activity analysis was implemented with 200 μL of the diluted soil sample (1:100) injected directly into freeze-dried substrates to produce substrate concentrations of 500 μM. The 200 μL of the diluted and homogenized sample was used in the standard measurements.
In order to investigate the component of the AMS, FTIR spectra were recorded on a FTIR IRPrestige-21 spectrometer in the range of 500 to 4,000 cm−1 using KBr pellet. The structure variations of CGS were examined by a scanning electron microscope (SEM, Quanta-600, FEI).
Results were statistically analyzed by one-way analysis of variance (ANOVA) with a significance level of 5% for all tests using the IBM SPSS Statistics 19.
Results and Discussion
Characterization of Amino Acid–Modified Magnetic Coal Gasification Slag and Soil in Abandoned Mine Area. Figure 1 illustrates X-ray diffraction patterns of CGS, Fe3O4, and magnetic MCGS. The peaks for MCGS appeared at (003), (006), (015), (018), (110), and (113) planes, which were attributed to the CGS characteristic peaks (Shan et al. 2015). The peaks for Fe3O4 appeared at (220), (311), (400), (442), (511), and (440) planes, which were the Fe3O4 characteristic peaks (Saievar et al. 2014). The magnetic MCGS had apparent superparamagnetism because the Fe3O4 nanoparticles had been efficiently integrated with the CGS.
The FTIR spectrums of MCGS, SAA, and AMS are shown in figure 2. The broad and strong absorption peak at 3,445 cm−1 for MCGS was the stretching vibration band of O–H from water molecules of the layer surface and/or interlayer. The peaks at 1,564 cm−1 and 2,352 cm−1 were due to the in-plane stretching of nitrogen (N)–hydrogen (H), −NH2, or ammonia (−NH3) from the Fe3O4 synthesis process (Guo et al. 2013). The peaks at 1,375, 785, and 678 cm−1 for the composite were vibrational modes of the carbonate (CO32−) anion from the prepared CGS (Shan et al. 2015). For SAA (figure 2b), strong peaks at 3,417 cm−1 and 3,220 cm−1 were attributed to N–H vibrations. The absorption peak at 2,959 cm−1 was the stretching vibration of carbon (C)–H. The peak at 1,720 cm−1 was associated with the absorption peak of carboxyl (−COOH), and the peaks at 1,116 cm−1 and 1,646 cm−1 were the C–oxygen (O) stretching vibration and the C=O adsorption of acid amide, respectively. These results indicated that the product was amino acids (Su et al. 2014). For AMS (figure 2c), the N–H stretching modes of amino acids shifted to 3,430 cm−1, the C–H stretching modes shifted to 2,957 cm−1, the absorption peak of −COOH shifted to 1,721 cm−1, and the peaks corresponding to the C–O stretching vibration and the C=O adsorption of acid amide shifted to 1,117 and 1,647 cm−1, respectively. The results indicated that the interactions such as van der Waals forces, hydrogen bonding, and electrostatic attraction had taken place between the MCGS and SAA molecules (Zhang et al. 2017a).
The sand, silt, and clay contents in the MS were 48.9%, 46.5%, and 3.89%, respectively. Table 2 shows the physico-chemical properties of MS and AS with SAA, MCGS, and AMS, respectively (p < 0.05). With the increasing of SAA and AMS, pH of AS decreased, while its pH increased with the increasing of MCGS. CEC of AS increased with the increasing of the soil amendments, and the effect of AMS was the most prominent for CEC. The decrease of pH after SAA application might lead to the increase of exchangeable cations in soil, which resulted in the increase of CEC of AS, and superparamagnetism of CGS might contribute to CEC increase (Gao et al. 2007). The OM, MC, total N, and available N of AS increased with the increments of SAA, MCGS, and AMS. The CaCO3, total K, and available K of AS increased with the increase of MCGS and AMS. The benefits might be attributed to the high content of nutrients in the soil amendments. In addition, the superparamagnetism of MCGS and AMS could reduce permanent charge value of soil, which would play a role in the formation of soil aggregates (Yi et al. 1991; Yi 2000). Yi (2000) indicated that the superparamagnetism could also improve soil enzyme activity and soil water retention characteristics and decrease the soil cohesion by reducing the specific surface area of soil and improving the degree of soil aggregation. Organic matter and superparamagnetism could effectively alleviate soil erosion, improve soil quality (such as structural properties, porosity, soil temperature, and gas permeability), and enhance soil infiltration rate and soil water-holding power (Li et al. 2009). Therefore, the high content of nutrients or superparamagnetism from SAA, MCGS, and AMS is favorable to MS. In addition, total P and available P of AS remained little changed after adding the soil amendments.
The effects of the soil amendments (SAA, MCGS, and AMS) on heavy metals of MS and AS are shown in table 3. As the SAA increased, lead (Pb), Zn, Cr, Cu, and Fe slightly increased. As the MCGS increased, the heavy metals (except for Fe, Cu, and Cr) slightly decreased (p < 0.05). As the AMS increased, Pb, Zn, Cr, and Cu slightly increased (p < 0.05) and the Fe obviously increased (p < 0.05). Table 3 also shows that these changes were irregular. Even so, the contents of heavy metals in all AS samples were lower than the soil environmental limited values standard in China (natural background), except for Cd (GB15618-1995). Therefore, the AMS would not lead to soil heavy metal contamination, and the heavy metal risks of AS could be controlled by decreasing Cd content.
Figure 3 shows the effect of AMS on the WSA distribution. By comparison, with control check (CK), the proportions of macroaggregates (>0.25 mm) increased and that of microaggregates (<0.25 mm) decreased with the increment of AMS. The increases of stable 0.25 to 1 mm soil aggregates were about 1.53 times (from 20.21% to 51.16%) with 5% AMS and 1.73 times (from 20.21% to 55.12%) with 7.5% AMS (figure 3). Occurrence of bigger proportion of WSA as macroaggregates compared to microaggregates in AMS-amended plots could be attributed to the addition of amino acid through AMS, resulting in improvement of soil microbial activity, which was conducive to binding of aggregates (Sodhi et al. 2009). In addition, AMS with layer structure had good absorbability and cohesion ability, which also helped in cementing of aggregates (Bi et al. 2007; Ji et al. 2019).
Alhagi Sparsifolia Growth and Heavy Metals Uptake. The effects of AS on the biomass and growth of A. sparsifolia are shown in figure 4. Obviously, there was poor A. sparsifolia growth in the raw soil (figure 4). The group of MS with 5% AMS was best for improving the biomass and growth of plant, followed by 5% MCGS + 5% SSA, 5% SSA, 5% MCGS, and MS. It was interesting to note that in the group of MS with 5% AMS, the increases of the biomass and growth of A. sparsifolia were higher than the values obtained by the simple combination of 5% MCGS and 5% SSA. In other words, AMS as a soil amendment produced a significantly greater effect than the simple combination of SAA and MCGS. It could also be observed that the growing trend of A. sparsifolia in AS with SAA was not pronounced after 54 days (figures 4a and 4b). The cause might be that the amino acids were assimilated and decomposed by bacteria in soil. Zhang et al. (2017b) indicated that amino acids could provide necessary nutriments for plant growth, but amino acids were easily assimilated and decomposed by bacteria in the soil. In this work, SAA was inserted into the layer of MCGS, and MCGS with the similar layer structure of hydrotalcite could shield amino acids from the assimilation and decomposition of bacteria in soil (Liu et al. 2011; Ma 2016). The biomass of the plant from the MS group was scarce. In contrast, after soil amendments (especially AMS) were added to soil, there were obvious increases in the biomass and growth of plant (figure 4c). The plants in the AS were better than in the MS. Fresh and dry weight yields of A. sparsifolia in the soil with AMS were highest, and the lowest values appeared in the MS (see control group in figure 4c). AMS could not only effectively improve soil structural characteristics, but also provide rich nutrients such as N to soil. Abundant N could promote plant growth and increase plant biomass. The maximum growth rate of the plant might have been reached when the proportion of AMS was 7.5%, so the further increase of the AMS (>7.5% w w−1) would have no obvious effect on the plant growth. Therefore, if the AMS was 7.5%, a more satisfying effect could be obtained (figure 4d).
The concentration of total metals in A. sparsifolia aboveground parts and roots in AS (with AMS) is shown in table 4. With the increase of AMS, the heavy metals (except for Fe) in shoots decreased, especially Pb and Cd (table 4). The heavy metals in roots were higher than in aboveground parts, except for Zn. The heavy metal uptake by the A. sparsifolia roots displayed mixed results possibly because of the high standard deviation about the means. It appeared that the AMS could obviously decrease heavy metals at 7.5% and 10% (table 4).
Species Distribution of Heavy Metals in the Rhizosphere Soil. Species distribution of heavy metals in soil often determines their bioavailability and ecotoxicity. BCR method divides species distribution into residual form (stable and safe form), oxidizable form (potential effect), reducible form (assimilable), and exchangeable form (assimilable). Figure 5 shows species distribution of heavy metals in the MS and AS (with 7.5% AMS) before and after the planting tests. As figure 5 shows, mercury (Hg) and Fe mainly appeared in the residual fraction (68.9% to 73.1% and 60.3% to 80.1% in the MS and 80.2% to 94.7% and 50.2% to 69.7% in the MS) in all tests. It indicated that the MS and AS with 7.5% AMS were safe and had low bioavailability to the environment in the aspects of Hg and Fe. Residual fraction of Cr, nickel (Ni), cadmium (Cd), arsenic (As), and Hg increased after planting, which revealed that A. sparsifolia could decrease the bioavailability of these metals. Residual fraction of Pb, Zn, Cu, and Fe decreased in the MS after planting. Residual fraction of all heavy metals in this work (except for Fe) increased in the AS after planting, and the percentages of reducible fraction and exchangeable fraction of Pb (4.1%, 0%), Cr (2.9%, 1.1%), Cd (0%, 2.3%), and Hg (0%, 0%) were very low in the AS after planting. Therefore, AMS and A. sparsifolia could be used to decrease the bioavailability of all heavy metals studied in this work (except for Fe).
We had found in previous study that the bioavailability of Ni, vanadium (V), and Mo was decreased by A. sparsifolia, and A. sparsifolia could also decrease the bioavailability of Cd, Cu, Cr, and manganese (Mn) in artificial soil (modified excess sludge: aeolian sandy soil = 1:2) (Xiang et al. 2016). In this study, the above effects of A. sparsifolia were further enhanced due to the addition of AMS. Some studies have indicated the metal ions could be adsorbed to the pores and surface of MCGS materials (Wen et al. 2013; Suresh et al. 2018). The active sites were evenly distributed throughout MCGS, which had appropriate energy to promote spontaneous surface diffusion of heavy metals (Cao et al. 2016). On the other hand, the heavy metals could be removed from the soil through surfactant-associated complexation (Ochoa-Loza et al. 2001) and ionic exchange (Swarnkar et al. 2012). Organic matter such as SAA could convert the exchangeable metals into organic bond fraction, thus decreasing uptake (Muhammad et al. 2017). In addition, Jiang et al. (2018b) indicated that the magnetism of material could improve the adsorption of heavy metals. Consequently, surfactant, Al compounds, and magnetic substance made the AMS suitable for the removal of heavy metals in soil environment.
Figure 6 shows the changes of AMS before and after soil application. Scanning electron microscope (SEM) analyses showed that there were some variations in the appearances of AMS before and after application (figures 6a and 6c). There seemed to be many more small particles in figure 6c than in 6b. This might be because some heavy metal compounds were adsorbed on AMS surface (figure 6c). Energy dispersive spectrometer (EDS) analyses before and after application were obviously different (figures 6b and 6d). Compared to the original AMS, the peaks of heavy metals (such as Cr, Pb, Cu, Cd, Ni, Mn, and As) appeared in the AMS sample after land use. This further proved that the adsorption of heavy metals might occur in both the interlayer and on the surface of AMS.
Effect of Amino Acid–Modified Magnetic Coal Gasification Slag on Enzyme Activities. Compared to the corresponding bulk soils, the activities of UA, PA, and β-GA in the rhizosphere of A. sparsifolia planted in the AS were significantly improved (figure 7). With the increase of the AMS, the activities of UA, PA, and β-GA all increased significantly (p < 0.05) compared to the control group. It was reported that soil enzyme activities were closely related to vegetation and physiochemical properties of soil (Song et al. 2016; Yang et al. 2017). In general, soil enzyme activities in the rhizosphere could be improved because of the high microbial biomass from root exudates (Ning et al. 2017). In this work, AMS could promote the growth of A. sparsifolia (figure 4), and the growth of A. sparsifolia could generate more root exudates. Furthermore, MCGS with special sandwich structure and controllability of layer interval could provide microenvironment for enzymes, which might be useful in retaining and prolonging the activity of the enzymes (Ji et al. 2004). The activities of the three enzymes of the control group (without AMS) were all very low. The inhibition of heavy metals in soils might be one possibility for the low activity. Novak et al. (2018) indicated that microbial activity was inhibited by heavy metals in soil. The addition of AMS to soil greatly decreased the activity of heavy metals and relieved their toxicity on soil enzymes; therefore, the enzyme activities were significantly (p < 0.05) enhanced.
Mechanism Analysis and Reusability of Amino Acid–Modified Magnetic Coal Gasification Slag. Obviously, the growth of A. sparsifolia and rehabilitation of MS were both improved under the effect of AMS. First, the AMS not only improved the immobilization of heavy metals through adsorption and reduced the bioavailability of heavy metals, but also offered nutrients to the growth of plant. Second, plant growth also helped improve the soil organic matter, microorganism, and soil enzyme activity (Wang et al. 2019). Low molecular weight organic matters supplied dissolved ligand for some heavy metals (except for Fe in this work) to form soluble complexes (figure 8), while the macromolecule part decreased some heavy metals availability (such as Cr, Ni, Cd, As, and Hg) (Du et al. 2009). Zhang et al. (2014) also indicated that organic matter (especially fine particles) had a significant effect on trace metals solubility and bioavailability in soil. Consequently, the MS environment could be improved by the AMS and A. sparsifolia.
Based on the studies stated above, reusability of AMS (7.5% w w−1) was tested for 8 cycles (figure 9). After each cycle, soil remained unchanged, plant was replaced with a new one, and the AMS was separated from the AS by magnetic separation technology and washed with deionized water. The heavy metals in AMS were desorbed with SAA (Huang and Li 2009). The procedure was the same as the procedure of SAA intercalated with MCGS, and the obtained AMS was dried at 40°C in the shade for further use. Results showed that the growth of plant was better than in the control group (without AMS, other conditions being equal) for the first seven experiments, while the growth trend after eight experiments decreased to a similar extent as in the control group (figure 9a). The effects of repeated use of AMS on Cd content in plant and soil were shown in figure 9b. Cadmium content in soil decreased with the increase of AMS recycle times. Cadmium content in plant fluctuated in the range of 0.20 to 0.25 mg kg−1. AMS structure might be destroyed when reuse number was excessive (>7), so the plant growth decreased. Through the “memory effect” of AMS, the invalid AMS could be recycled by calcination at low temperature (below 500°C) and soaking in the mixed solution that was obtained in AMS preparation stage (Chen et al. 2016; Shui et al. 2018).
Summary and Conclusions
This study tested the ability of AMS to improve plant growth and soil enzyme activity and decrease heavy metals. Results showed the positive effects of AMS on the remediation of the MS and growth of A. sparsifolia. Paramagnetic AMS with high nutrient content could change the physico-chemical properties of the soil and help in the formation of soil macroaggregates. When AMS was 7.5%, A. sparsifolia presented the highest growth potential. AMS and A. sparsifolia could effectively decrease the bioavailability of heavy metals and improve soil enzyme activities. AMS improved the immobilization of heavy metals through adsorption, and the unique structure of AMS also improved and retained the activity of soil enzymes. The plant growth could help to improve the soil enzymes and plant biomass. Therefore, both growing plants and AMS interact to improve nutrient cycling and MS remediation. In addition, A. sparsifolia growth was still better than that of control group when the AMS was used for 7 times repeatedly. Consequently, the recyclable AMS is promising for the remediation of contaminated soil.
Acknowledgements
The authors are grateful for the funding and support provided by the National Natural Science Foundation of China (41967022) and the General Project (Youth), Natural Science Basic Research Program of Shaanxi Science and Technology Department, Shaanxi, China (2019JQ-413).
- Received May 23, 2019.
- Revision received October 7, 2019.
- Accepted October 21, 2019.
- © 2020 by the Soil and Water Conservation Society