Environmental Geochemistry and Health

, Volume 39, Issue 2, pp 345–352 | Cite as

Effect of chemical amendments on remediation of potentially toxic trace elements (PTEs) and soil quality improvement in paddy fields

  • Sung Chul Kim
  • Young Kyu Hong
  • Se Jin Oh
  • Seung Min Oh
  • Sang Phil Lee
  • Do Hyung Kim
  • Jae E. Yang
Original Paper

Abstract

Remediation of potentially toxic trace elements (PTEs) in paddy fields is fundamental for crop safety. In situ application of chemical amendments has been widely adapted because of its cost-effectiveness and environmental safety. The main purpose of this research was to (1) evaluate the reduction in dissolved concentrations of cadmium (Cd) and arsenic (As) with the application of chemical amendments and (2) monitor microbial activity in the soil to determine the remediation efficiency. Three different chemical amendments, lime stone, steel slag, and acid mine drainage sludge, were applied to paddy fields, and rice (Oryza sativa L. Milyang 23) was cultivated. The application of chemical amendments immobilized both Cd and As in soil. Between the two PTEs, As reduction was significant (p < 0.05) with the addition of chemical amendments, whereas no significant reduction was observed for Cd than that for the control. Among six soil-related variables, PTE concentration showed a negative correlation with soil pH (r = −0.70 for As and r = −0.54 for Cd) and soil respiration (SR) (r = −0.88 for As and r = −0.45 for Cd). This result indicated that immobilization of PTEs in soil is dependent on soil pH and reduces PTE toxicity. Overall, the application of chemical amendments could be utilized for decreasing PTE (As and Cd) bioavailability and increasing microbial activity in the soil.

Keywords

Potentially toxic trace elements Chemical amendments Paddy soil Microbial activity Soil quality 

Introduction

Pollution of potentially toxic trace elements (PTEs) in agricultural fields near abandoned metal mines is a critical problem (Baker et al. 2011; Garau et al. 2007; Sun et al. 2015). PTE contamination in soil can be detrimental to soil ecosystems, crops, and even human health (Bade et al. 2012). Among various PTEs, cadmium (Cd) and arsenic (As) are very toxic elements and can cause cancer in humans (de Mora et al. 2005; Ko et al. 2013; Kosolsaksakul et al. 2014; Moon et al. 2004). Cd is a relatively rare element in earth and is associated with igneous and metamorphic rocks at lower concentrations (0.02–0.2 mg kg−1) and zinc ores at higher concentrations (up to 14,500 mg kg−1) (Kosolsaksakul et al. 2014). Similarly, As is naturally released in the environment through weathering, but anthropogenic activities such as pesticide and insecticide usage, smelting industries, and mining activities can result in environmental exposure (Leist et al. 2000; Moon et al. 2004).

Various remediation techniques have been studied previously for reducing the concentration of bioavailable PTEs in soil. Remediation techniques in contaminated soil can be classified into ex- and in situ techniques. Since ex situ techniques have disadvantages, such as high cost and soil structure deterioration, in situ techniques, such as stabilization with chemical amendments, have been widely adapted (Asensio et al. 2013; Garau et al. 2007; Ko et al. 2013). Chemical amendment applied to PTEs contaminated soil can be divided into three categories: pH adjustment, organic adsorbents, and minerals (US EPA 2007). Among them, lime-containing materials, steel slag, and acid mine drainage represent amendments for pH adjustment, minerals, and inorganic by-product, respectively. Lime is a pH-neutralizing soil amendment including limestone (CaCO3), hydrated lime (Ca(OH)2), and cement kiln dust. The application of lime to PTE-contaminated soil can increase soil pH, inducing the precipitation of metal carbonates, oxides, and calcium (Ca) silicates, and eventually decrease PTE solubility (Adriano et al. 2004; Chlopecka and Adriano 1996). Steel slag (SS) is an industrial by-product containing oxide and a silicate formation of Ca, magnesium (Mg), and iron (Fe) (Cobo et al. 2009; Kim et al. 2008). Various mechanisms such as hydrolysis, precipitation, and adsorption can play an important role to reduce PTSs when SS is applied in soil (Qiu et al. 2012). Acid mine drainage sludge (AMDS) is a by-product of coal mining and is mainly composed of amorphous Fe/aluminum (Al) hydroxide (Tsang et al. 2013). Inorganic by-products that contain abundant Fe, Al, and Ca can sequester oxyanions through adsorption onto surface hydroxyl groups, co-precipitation/incorporation into an oxide matrix, and Fe/Ca arsenate precipitation (Komárek et al. 2013; Miretzky and Cirelli 2010).

Paddy fields are a unique environment where floods are maintained for a certain period of time for rice growth. During the flood period, paddy soil is generally in a reduced condition and the physicochemical properties of the soil are affected, including soil pH, soil redox potential (Eh), and contents of soil organic matter (Van den Berg and Lock 2000). Soil pH becomes neutral, and Eh decreases because of microbial respiration during the flood period, resulting in the reduction of Fe and Mn in the soil (Mcbride 1994). Consequently, in the case of As, arsenate (Asv) strongly adsorbed onto Fe oxide in soil is reduced to arsenite (AsIII) under flooding and accumulated in the rice grain by a silicate transporter. In contrast, Cd accumulation in rice grains during flooding generally decreases owing to increased soil pH.

Soil microbial activity is an important factor for organic matter decomposition, nutrient recycling, and maintaining environmental quality (Garau et al. 2007; Nwachukwu and Pulford 2011). Soil microbial activity can be an indicator of PTE toxicity since high concentrations of PTEs can inhibit microbial activity and negatively impact microbial community diversity (Hu et al. 2014). Therefore, researchers have included soil microbial activity assessments in the remedial analyses of PTEs to evaluate remedial efficiency (Baker et al. 2011; de Mora et al. 2005; Garau et al. 2007; Hu et al. 2014). However, the response of soil microbial activity to polluted PTEs in soil is not always the same. Soil microbial activity might not be affected when soil is exposed to PTEs on a long term possibly leading to a strong immobilization of PTEs in soil and/or the adaptation of the microbial community to the pollution (Hu et al. 2014).

Among various soil microbial activities, soil respiration (SR), and enzyme activities are major parameters for assessing soil microbial activity. SR is a measure of C mineralization in soil, and generally, PTEs can potentially significantly reduce the ability of bacteria to decompose organic substrates (Nwachukwu and Pulford 2011). Consequently, measuring emitted CO2 in soil is a reliable method to evaluate the effect of PTEs on microbial activity. Soil enzymes are also involved in C and N mineralization, and ecological reactions in soil are affected by enzyme catalysis (Hu et al. 2014). Sucrase, urease, dehydrogenase, and phosphatases mainly break down polymers (sucrose) into monomers (glucose or fructose) or hydrolyze organics to inorganics for providing energy to microorganisms (Dick et al. 1996; Edwards 2002; Kaschuk et al. 2010; Mikanova 2006; Nannipieri et al. 1978; Pant and Warman 2000).

The main purpose of the present research was to evaluate three different chemical amendments (limestone, SS, and acid mine drainage sludge) in decreasing the bioavailable fraction of Cd and As in paddy fields. Furthermore, the effect of chemical amendments on soil microbial activity, soil respiration (SR), and microbial biomass C (MBC) was assessed to evaluate remediation efficiency and soil quality improvement.

Materials and methods

Site description

Field experiments were conducted in a paddy field located near an abandoned metal mine (N 37.01.52.32, E 128.14.36.32) in Korea. The mine waste section was located 0.6 km north of the study area, and historically, mine waste, such as mine tailings and overburden, had an effect on heavy metal pollution near the study area. The metal mine was operated from 1939 to 1995, and the major ores were gold (Au), silver (Ag), Fe, and zinc (Zn). The geology of the study area was mainly pegmatite and felsite. The study area had a humid continental/subtropical climate with a dry winter, and annual average precipitation is about 1130–1500 mm. Soil texture in the paddy field was silt loam (sand: 36.9%, silt: 45.5%, clay: 17.6%), and chemical properties such as soil pH (6.06 ± 0.17), electrical conductivity (291 ± 13.7 μS cm−1), organic matter (27.25 ± 0.78 g kg−1), and cation exchange capacity (18.58 ± 0.32 cmol kg−1) were measured before plot construction.

Field plot construction

Field plots were constructed with a dimension of 2 × 2 × 0.6 m (L × W × D) and a ridge surrounded each plot to separate them. Three chemical amendments: limestone (LS), SS, and AMDS, were applied using a complete randomized block design with three replicates for each treatment. A control plot, with no chemical amendments applied, was constructed for comparison. To mix the surface soil and chemical amendments, 40 cm of the surface soil was excavated and 3% chemical amendments were thoroughly mixed with the surface soil in a container. A previous study showed that a mixing ratio of 3% is the optimum ratio for soil pH, electric conductivity (EC), and immobilization of Cd and As (Oh et al. 2012). Therefore, a 3% mixing ratio was adapted in the present study. Surface soil mixed with chemical amendment was then transferred to each plot and compacted with a shovel. After constructing each plot, soil was equilibrated for two weeks and then moistened with non-contaminated groundwater for two weeks. Prior to planting, fertilizer was applied according to the guidelines of the Rural Development of Agriculture (RDA) in Korea to each plot, and then, plots were irrigated with non-contaminated groundwater. The rice cultivar (Oryza sativa L. Milyang 23) was supplied from the National Academy of Agricultural Science (NAAS), and five stands (3–4 plants for a stand) of 50-day-old seedlings were planted in each plot with 30 cm between each seedling.

Sample collection and pre-treatment

Soil samples in each plot were collected in March 2013 directly after plot construction and in October 2013 after the rice was harvested. Surface (0–15 cm) and subsurface (15–30 cm) soil was collected at the same locations with hand augers after removing organic debris on the surface. Five different soil samples were collected and combined to make one representative sample for each plot, contained in plastic bags, and transferred to the laboratory for further analysis. After collection, each soil sample was divided into two parts. One part of the soil was stored in a refrigerator at 4 °C to keep it moistened for soil microbial analysis. The other part of the soil was air-dried at room temperature (20 °C) and passed through a 2 mm sieve for chemical and Cd and As analysis.

Analysis

Soil analysis

Soil chemical properties were analyzed based on standard methods proposed by the National Academy of Agricultural Science and the Korean Ministry of Environment. Briefly, soil pH and EC (1:5) were analyzed with a pH (Mettler Toledo, MP 200, Switzerland) and EC (Mettler Toledo, S230, Switzerland) meter after 5 g soil and 25 mL deionized water were thoroughly mixed in a 50-mL flask for 1 h. Organic matter (OM) was determined following the Walkley–Black method (Nelson and Sommers 1996), and cation exchange capacity (CEC) was analyzed using a 1 M CH3COOH extraction method (Sumner and Miller 1996).

The bioavailable fraction of Cd and As in soil was determined using 0.1 N HCl for Cd and 1.0 N HCl for As. Briefly, 10 g soil and 50 mL extractant were placed into 100-mL polypropylene centrifuge tubes and shaken for 1 h at 30 °C in an incubation shaker (JSR, JSOS-500, Korea). After shaking, samples were filtered through 0.45-µm filter paper and filtrates with the metal concentrations were measured using atomic absorption spectrometry (AA6800, Shimazu, Japan) with a hydride generator for As analysis. For QA/QC, certified reference materials (CRM) of soil samples (BAM-U112, Federal Institute for Materials Research and Testing, Berlin, Germany) were measured in every 30 samples. In addition, blank and spiked aqueous samples with known concentrations were periodically measured. The average recovery ratio of CRM was 96.5 ± 12.7% for examined PTSs through the measurement.

Soil biological properties, such as SR and MBC, were analyzed following the carbon dioxide (CO2) trap and fumigation-extraction methods, respectively (Vance et al. 1987).

Statistical analysis

The significance of results was determined with an ANOVA test at the p value of 0.05. Results between concentration of Cd and As and soil properties were evaluated using a correlation analysis to examine the effect of chemical amendments in terms of Cd and As reduction and changing soil properties.

Results and discussion

Chemical properties of soil

The chemical properties of the study plots are summarized in Table 1. Soil pH increased in all treated plots, including the control in October, except in the subsurface soil of the AMDS-treated plot. However, no significant differences were observed in treated plots except for the control in October. A previous study concluded that soil pH of paddy fields tended to be neutral because of the effect of microbial respiration on the redox potential and H+ consumption under reducing conditions that affect pH. The result of our experiment also showed that soil pH of all treated plots, including the control, was close to neutral (Zheng and Zhang 2011).
Table 1

Summary of chemical properties of each plot applied with chemical amendments

 

Amendments

PH

EC DS m−1

OM %

P2O5 Mg kg−1

March

Oct.

March

Oct.

March

Oct.

March

Oct.

Surface

Control

6.51 ± 0.26b

7.32 ± 0.37a

2.19 ± 0.92a

0.93 ± 0.24a

3.84 ± 0.14a

4.22 ± 0.33a

273.19 ± 78.42a

193.00 ± 63.79a

Lime stone (LS)

7.28 ± 0.16a

7.76 ± 0.14a

1.67 ± 0.24a

0.78 ± 0.09a

3.51 ± 0.53a

3.79 ± 0.13a

208.91 ± 39.21a

183.28 ± 28.54a

Steel slag (SS)

6.96 ± 0.06a

7.64 ± 0.33a

1.74 ± 0.35a

1.36 ± 0.51a

3.31 ± 0.46a

4.35 ± 0.57a

228.00 ± 70.69a

274.09 ± 112.43a

Acid mine drainage sludge (AMDS)

6.53 ± 0.14b

7.10 ± 0.26a

2.20 ± 0.02a

1.57 ± 0.41a

3.06 ± 0.57a

4.00 ± 0.71a

217.95 ± 61.04a

184.07 ± 100.01a

Subsurface

Control

6.52 ± 0.26a

6.95 ± 0.52a

1.68 ± 0.37b

1.55 ± 0.34a

2.34 ± 0.31b

3.98 ± 0.76a

125.55 ± 18.41a

159.90 ± 9.84a

Lime stone (LS)

7.31 ± 0.04b

7.56 ± 0.33a

1.34 ± 0.25ab

1.75 ± 0.54a

1.78 ± 0.09a

3.80 ± 0.84a

213.94 ± 114.97a

282.88 ± 95.67a

Steel slag (SS)

6.90 ± 0.28ab

7.66 ± 0.46a

0.90 ± 0.14a

1.38 ± 1.00a

1.67 ± 0.36a

3.93 ± 0.42a

107.72 ± 32.34a

262.91 ± 129.95a

Acid mine drainage sludge (AMDS)

7.01 ± 0.19b

6.90 ± 0.16a

1.23 ± 0.17ab

1.40 ± 0.39a

1.74 ± 0.11a

3.79 ± 0.63a

104.46 ± 13.92a

212.82 ± 39.09a

All values are average of triplicate measurement

Statistical analysis was conducted with all triplicate measurements

Same letters are not significantly different (p < 0.05)

Electrical conductivity (EC), organic matter (OM), and concentration of phosphorus (P2O5) were 1.67–2.20 dS m−1, 3.06–3.84%, 208.91–273.19 mg kg−1 in surface soil and 0.90–1.68 dS m−1, 1.67–2.34%, 159.0–282.88% in subsurface soil, respectively. The data from soils sampled in October were not significantly different from those derived from soils sampled in March.

Bioavailable fraction of Cd and As in soil

The Cd and As concentration as a bioavailable fraction is summarized in Table 2. The initial total average Cd concentration in the study plot was 6.26 ± 2.69 mg kg−1, which exceeded the “threshold of danger level” (4.0 mg kg−1) in Korea. However, the bioavailable fraction of Cd in the chemically amended plots showed no significant difference from that of the control both in March and in October. This result might indicate that the Cd immobilization occurred immediately after mixing the chemical amendments and no further redistribution of bioavailable Cd occurred during the experimental period. A recent study revealed soil pH and Ca2+ ion concentration play an important role to decrease the bioavailable fraction of Cd in paddy soil (Wu et al. 2016). Increased soil pH produced OH ions, which precipitated with heavy metals resulting in a reduction in the bioavailable fraction of Cd. Soil pH of lime and SS-treated plots in March (2 weeks after mixing) was 7.28, 6.96 and 7.31, 6.90 in surface and subsurface soil, respectively. These soil pH values were higher than other plots, and increased soil pH is the main mechanism for reducing bioavailable Cd in lime and SS-treated plots. In addition, high Ca2+ concentrations in lime and SS have an antagonistic effect on heavy metals in soil and decrease bioavailable Cd (Wu et al. 2016).
Table 2

Bioavailable fraction of Cd and As in each plot applied with chemical amendments

 

Amendments

Cd Mg kg−1

As Mg kg−1

March

Oct.

March

Oct.

Surface

Control

0.70 ± 0.03a

0.69 ± 0.06a

26.81 ± 6.51a

22.15 ± 1.37b

Lime stone (LS)

0.48 ± 0.09a

0.59 ± 0.09a

11.51 ± 5.74b

9.15 ± 3.24a

Steel slag (SS)

0.65 ± 0.21a

0.61 ± 0.23a

18.99 ± 1.33ab

11.38 ± 2.15a

Acid mine drainage sludge (AMDS)

0.76 ± 0.17a

0.72 ± 0.30a

17.37 ± 4.23ab

6.80 ± 1.06a

Subsurface

Control

0.55 ± 0.21a

0.54 ± 0.13a

24.76 ± 1.96a

17.18 ± 2.40b

Lime stone (LS)

0.34 ± 0.12a

0.40 ± 0.21a

22.65 ± 1.22a

2.43 ± 2.10a

Steel slag (SS)

0.35 ± 0.09a

0.41 ± 0.10a

25.14 ± 2.08a

3.97 ± 1.78a

Acid mine drainage sludge (AMDS)

0.44 ± 0.10a

0.61 ± 0.11a

23.21 ± 1.58a

1.73 ± 0.57a

All values are average of triplicate measurement

Statistical analysis was conducted with all triplicate measurements

Same letters are not significantly different (p < 0.05)

For As, significantly lower As concentrations were observed in the LS-treated plot in March, and in all chemically amended plots in October, than that in the control. The total average As concentration extracted with aqua regia in the study plot was 803.84 ± 18.36 mg kg−1, which exceeded the “corrective action level” for As (75 mg kg−1). Since the study plot was heavily polluted with As, the effect of chemical amendments on decreasing the bioavailable As fraction was significant. Among the three chemical amendments, AMDS and LS showed a high efficiency for decreasing As concentrations during experimental periods. Previous research showed that the general mechanism for As immobilization with lime and AMDS was the formation of Ca–As precipitates (Moon et al. 2004). When lime is applied to As-contaminated soil, the formation of Ca3(AsO4)2 and CaHAsO3 precipitates controls As immobilization. For AMDS, iron oxide can play an important role in adsorbing As in the soil, resulting in As immobilization.

Biological properties of soil

Results of the soil biological properties such as SR and MBC are summarized in Table 3. Initially, higher SR rates were observed in all chemically amended plots than in the control plot. Among them, SS-treated plots showed the highest SR rate followed by LS- and AMDS-treated plots. After rice cultivation, the SR rate increased in all treated plots, including the control, by more than 10 times. The highest increase in SR rate was observed in the AMDS-treated plot. For MBC, higher concentrations were observed in all chemically amended plots than in the control plots. However, the MBC concentration decreased between 58 and 76% after cultivation in chemically amended plots, although there were no significant differences to the control plot. This result is in contrast to previous studies showing increased MBC concentrations in plots mixed with four different organic amendments: municipal waste compost, biosolid compost, leonardite, and litter (de Mora et al. 2005). Generally, the MBC concentration is highly correlated with dissolved organic carbon (DOC), and previous studies that applied organic amendments showed an increase in DOC. However, the organic matter concentration decreased after rice cultivation in the present study (Table 1). This could explain the decreasing MBC in chemically amended plots.
Table 3

Soil biological properties in chemical amended plots

 

Amendments

Soil respiration Mg CO2 100 g−1 day−1

Microbial biomass carbon Mg kg−1

March

Oct.

March

Oct.

Surface

Control

153.64 ± 27.24a

170.33 ± 30.49a

145.72 ± 33.00a

124.07 ± 10.99a

Lime stone (LS)

147.95 ± 21.96a

192.67 ± 24.62a

144.79 ± 28.01a

146.90 ± 19.10a

Steel slag (SS)

157.90 ± 41.69a

221.06 ± 45.69ab

127.08 ± 7.22a

137.52 ± 16.25a

Acid mine drainage sludge (AMDS)

158.06 ± 29.42a

297.60 ± 67.04b

141.15 ± 10.16a

152.70 ± 14.73a

Subsurface

Control

126.69 ± 24.43a

155.63 ± 27.81a

146.35 ± 31.30a

155.30 ± 12.29a

Lime stone (LS)

143.12 ± 37.89a

191.85 ± 84.79a

142.19 ± 34.55a

161.60 ± 2.38a

Steel slag (SS)

144.09 ± 38.12a

156.75 ± 58.07a

173.96 ± 14.77a

136.43 ± 36.22a

Acid mine drainage sludge (AMDS)

179.47 ± 26.26a

199.64 ± 50.11a

138.54 ± 26.90a

128.17 ± 17.40a

All values are average of triplicate measurement

Statistical analysis was conducted with all triplicate measurements

Same letters are not significantly different (p < 0.05)

Correlation between soil properties and bioavailable PTEs

The statistical correlation between soil properties and the bioavailable fraction of PTEs is shown in Table 4. The PTE concentration in soil is commonly associated with physicochemical properties of soil, and consequently, toxicity of PTEs is correlated with microbial activity in soil (Kizikaya et al. 2004).
Table 4

Correlation coefficient between soil properties and bioavailable fraction of potentially toxic elements (Cd and As) in surface soil

 

PH

EC

OM

P2O5

SR

MBC

As

Cd

pH

1

−0.87**

−.033

−0.17

0.49

−0.87**

−0.70

−0.54

EC

1

0.37

0.53

−0.65

0.77*

0.89**

0.45

OM

1

0.37

−0.67

0.66

0.50

0.16

P2O5

1

−0.35

0.22

0.50

−0.23

SR

1

−0.74*

−0.88**

−0.45

MBC

1

0.78*

0.62

As

1

0.56

Cd

1

* indicates that correlation is significantly different at p < 0.05

** indicates that correlation is significantly different at p < 0.01

In general, the bioavailable fraction of Cd and As is negatively correlated with soil pH and SR rate, indicating that increased pH can reduce the bioavailable fraction of PTEs resulting in an enhanced SR rate in soil among various soil physicochemical properties. Therefore, soil pH is the most effective parameter to control bioavailable fractions of potentially toxic trace elements in soil. The mobility of most cationic PTEs, including Cd, generally decreased in soil as soil pH increased. The main mechanism of enhancing immobilization in soil is deprotonation on the soil surface, resulting in the generation of more negative binding sites as soil pH is increased (Hale et al. 2012). In addition, adsorption and precipitation immobilize As in alkaline soil pH conditions. Previous research has demonstrated that precipitation of either CaHAsO3 or Ca3(AsO4)2 occurred at high pH levels, depending on the presence of As(III) or As(V), respectively, in soil (Dutré et al. 1999; Vandecasteele et al. 2002). Therefore, Ca–As precipitation is the dominant immobilization mechanism for As in high soil pH conditions.

The negative correlation between the bioavailable concentration of PTEs and SR rate clearly showed that chemical amendment application effectively reduced the bioavailable concentration of PTEs in the soil. Since 70% of added carbon in soil is lost as CO2, SR is an effective measure for monitoring microbial activity in soil (Nwachukwu and Pulford 2011).

Furthermore, the soil organic matter (OM) showed a positive correlation with MBC and negative correlation with the SR rate. This is because the MBC concentration increased owing to a high biodegradable organic matter concentration. In addition, increased OM concentrations can reduce the bioavailable fraction of Cd and As in soil (Kizikaya et al. 2004).

Conclusion

The application of LS, AA, and AMDS in Cd- and As-contaminated paddy fields could reduce the bioavailable fraction at a range of 5–60% for up to 7 months. Among various soil properties, soil pH had a negative correlation with the bioavailable fraction of Cd and As, indicating that controlling the soil pH was the most efficient strategy for reducing PTE solubility. The application of chemical amendments also significantly enhanced the SR rate. This result might indicate that chemical amendments increased Cd and As immobilization, resulting in the diminished toxicity of PTEs in the soil. Overall, in situ stabilization techniques with chemical amendments could be adapted in contaminated soil with PTEs to decrease the bioavailable fraction and furthermore enhance the soil quality for crop productivity in paddy fields.

Notes

Acknowledgements

This study was supported by research grants from the Kangwon National University (C 1009703-01-01) and the Korean Ministry of Environment (MOE) as the “Development of Korean Evaluation and Management System of Surface Soil Resources” in the GAIA Project (201400054003).

References

  1. Adriano, D. C., Wenzel, W. W., Vangronsveld, J., & Bolan, N. S. (2004). Role of assisted natural remediation in environmental cleanup. Geoderma, 122(2–4), 121–142.CrossRefGoogle Scholar
  2. Asensio, V., Covelo, E. F., & Kandeler, E. (2013). Soil management of copper mine tailing soils—sludge amendment and tree vegetation could improve biological soil quality. Science of the Total Environment, 456–457, 82–90.CrossRefGoogle Scholar
  3. Bade, R., Oh, S., & Shin, W. S. (2012). Assessment of metal bioavailability in smelter-contaminated soil before and after lime amendment. Ecotoxicology and Environmental Safety, 80, 299–307.CrossRefGoogle Scholar
  4. Baker, L. R., White, P. M., & Pierzynski, G. M. (2011). Changes in microbial properties after manure, lime, and bentonite application to a potentially toxic trace elements-contaminated mine waste. Applied Soil Ecology, 48(1), 1–10.CrossRefGoogle Scholar
  5. Chlopecka, A., & Adriano, D. C. (1996). Mimicked in situ stabilization of metals in a cropped soil: Bioavailability and chemical form of zinc. Environmental Science and Technology, 30(11), 3294–3303.CrossRefGoogle Scholar
  6. Cobo, M., Gálvez, A., Conesa, J. A., & de Correa, C. M. (2009). Characterization of fly ash from a hazardous waste incinerator in Medellin Colombia. Journal of Hazardous Materials, 168(2–3), 1223–1232.CrossRefGoogle Scholar
  7. de Mora, A. P., Ortega-Calvo, J., Cabrera, F., & Madejón, E. (2005). Changes in enzyme activities and microbial biomass after “in situ” remediation of a potentially toxic trace elements-contaminated soil. Applied Soil Ecology, 28(2), 125–137.CrossRefGoogle Scholar
  8. Dick, R. P., Breakwell, D. P., & Turco, R. F. (1996). Soil enzyme activities and biodiversity measurements as integrative microbiological indicators. In J. W. Doran & A. J. Jones (Eds.), Methods for assessing soil quality (Vol. 49, pp. 247–271). Fitchburg: Soil Science Society of America.Google Scholar
  9. Dutré, V., Vandecasteele, C., & Opdenakker, S. (1999). Oxidation of arsenic bearing fly ash as pretreatment before solidification. Journal of Hazardous Materials, 68(3), 205–215.CrossRefGoogle Scholar
  10. Edwards, C. A. (2002). Assessing the effects of environmental pollutants on soil organisms, communities, processes and ecosystems. European Journal of Soil Biology, 38(3–4), 225–231.CrossRefGoogle Scholar
  11. Garau, G., Castaldi, P., Santona, L., Deiana, P., & Melis, P. (2007). Influence of red mud, zeolite and lime on potentially toxic trace elements immobilization, culturable heterotrophic microbial populations and enzyme activities in a contaminated soil. Geoderma, 142(1–2), 47–57.CrossRefGoogle Scholar
  12. Hale, B., Evans, L., & Lambert, R. (2012). Effects of cement or lime on Cd Co, Cu, Ni, Pb, Sb and Zn mobility in field-contaminated and aged soils. Journal of Hazardous Materials, 199–200, 119–127.CrossRefGoogle Scholar
  13. Hu, X. F., Jiang, Y., Shu, Y., Hu, X., Liu, L., & Luo, F. (2014). Effects of mining wastewater discharges on potentially toxic trace elements pollution and soil enzyme activity of the paddy fields. Journal of Geochemical Exploration, 147(Part B), 139–150.CrossRefGoogle Scholar
  14. Kaschuk, G., Alberton, O., & Hungria, M. (2010). Three decades of soil microbial biomass studies in Brazilian ecosystems: Lessons learned about soil quality and indications for improving sustainability. Soil Biology & Biochemistry, 42(1), 1–13.CrossRefGoogle Scholar
  15. Kim, D. H., Shin, M. C., Choi, H. D., Seo, C. I., & Baek, K. T. (2008). Removal mechanisms of copper using steel-making slag: Adsorption and precipitation. Desalination, 223(1–3), 283–289.CrossRefGoogle Scholar
  16. Kizikaya, R., Aşkin, T., Bayrakli, B., & Sağlam, M. (2004). Microbiological characteristics of soils contaminated with potentially toxic trace elements. European Journal of Soil Biology, 40(2), 95–102.CrossRefGoogle Scholar
  17. Ko, M. S., Kim, J. Y., Lee, J. S., Ko, J. I., & Kim, K. W. (2013). Arsenic immobilization in water and soil using acid mine drainage sludge. Applied Geochemistry, 35, 1–6.CrossRefGoogle Scholar
  18. Komárek, M., Vaněk, A., & Ettler, V. (2013). Chemical stabilization of metals and arsenic in contaminated soils using oxides—a review. Environmental Pollution, 172, 9–22.CrossRefGoogle Scholar
  19. Kosolsaksakul, P., Farmer, J. G., Oliver, I. W., & Graham, M. C. (2014). Geochemical associations and availability of cadmium (Cd) in a paddy field system, northwestern Thailand. Environmental Pollution, 187, 153–161.CrossRefGoogle Scholar
  20. Leist, M., Casey, R. J., & Caridi, D. (2000). The management of arsenic wastes: Problems and prospects. Journal of Hazardous Materials, 76(1), 125–138.CrossRefGoogle Scholar
  21. McBride, M. B. (1994). Environmental chemistry of soils (pp. 127–333). London: Oxford University Press.Google Scholar
  22. Mikanova, O. (2006). Effects of potentially toxic trace elements on some soil biological parameters. Journal of Geochemical Exploration, 88(1–3), 220–223.CrossRefGoogle Scholar
  23. Miretzky, P., & Cirelli, A. F. (2010). Remediation of arsenic-contaminated soils by iron amendments—a review. Critical Reviews in Environmental Science and Technology, 40(2), 93–115.CrossRefGoogle Scholar
  24. Moon, D. H., Dermatas, D., & Menounou, N. (2004). Arsenic immobilization by calcium-arsenic precipitates in lime treated soils. Science of the Total Environment, 330(1–3), 171–185.CrossRefGoogle Scholar
  25. Nannipieri, P., Ceccanti, B., Cervelli, S., & Sequi, P. (1978). Stability and kinetic properties of humus-urease complexes. Soil Biology & Biochemistry, 10(2), 143–147.CrossRefGoogle Scholar
  26. Nelson, D. W., & Sommers, L. E. (1996). Total carbon, organic carbon, and organic matter. In A. L. Page et al. (Eds.), Methods of soil analysis, part 2. Chemical analysis (2nd ed., pp. 961–1110). Madison: Soil Science Society of America.Google Scholar
  27. Nwachukwu, O. I., & Pulford, I. D. (2011). Microbial respiration as an indication of metal toxicity in contaminated organic materials and soil. Journal of Hazardous Materials, 185(2–3), 1140–1147.CrossRefGoogle Scholar
  28. Oh, S. J., Kim, S. C., Kim, R. Y., Ok, Y. S., Yun, H. S., Oh, S. M., et al. (2012). Change of bioavailability in potentially toxic trace elements contaminated soil by chemical amendment. Korean Journal of Soil Science and Fertilizer, 45(6), 973–982.CrossRefGoogle Scholar
  29. Pant, H. K., & Warman, P. R. (2000). Enzymatic hydrolysis of soil organic phosphorus by immobilized phosphatases. Biology and Fertility of Soils, 30(4), 306–311.CrossRefGoogle Scholar
  30. Qiu, H., Gu, H. H., He, E. K., Wang, S. Z., & Qiu, R. L. (2012). Attenuation of metal bioavailability in acidic multi-metal contaminated soil treated with fly ash and steel slag. Pedosphere, 22(4), 544–553.CrossRefGoogle Scholar
  31. Sumner, M. E., & Miller, W. P. (1996). Cation exchange capacity and exchange coefficients. In D. L. Sparks et al. (Eds.), Methods of soil analysis part 3. Chemical methods. Madison: Soil Science Society of America.Google Scholar
  32. Sun, Y., Li, Y., Xu, Y., Liang, X., & Wang, L. (2015). In situ stabilization remediation of cadmium (Cd) and lead (Pb) co-contaminated paddy soil using bentonite. Applied Clay Science, 105–106, 200–206.CrossRefGoogle Scholar
  33. Tsang, D. C. W., Olds, W. E., Weber, P. A., & Yip, A. C. K. (2013). Soil stabilisation using AMD sludge, compost and lignite: TCLP leachability and continuous acid leaching. Chemosphere, 93(11), 2839–2847.CrossRefGoogle Scholar
  34. US EPA. (2007). The use of soil amendments for remediation, revitalization, and reuse. http://nepis.epa.gov/Exe/ZyPDF.cgi/60000LQ7.PDF?Dockey=60000LQ7.
  35. Van den Berg, G. A., & Loch, J. (2000). Decalcification of soils subject to periodic waterlogging. European Journal of Soil Science, 51(1), 27–33.CrossRefGoogle Scholar
  36. Vance, E. D., Brookes, P. C., & Jenkinson, D. S. (1987). An extraction method for measuring soil microbial biomass. Soil Biology & Biochemistry, 19, 703–707.CrossRefGoogle Scholar
  37. Vandecasteele, C., Dutré, V., Geysen, D., & Wauters, G. (2002). Solidification/stabilisation of arsenic bearing fly ash from the metallurgical industry. Immobilisation mechanism of arsenic. Waste Management, 22(2), 143–146.CrossRefGoogle Scholar
  38. Wu, Y. J., Zhou, H., Zou, Z. J., Zhu, W., Yang, W. T., Peng, P. Q., et al. (2016). A three-year in situ study on the persistence of a combined amendment (limestone + sepiolite) for remedying paddy soil polluted with heavy metals. Ecotoxicology and Environmental Safety, 130, 163–170.CrossRefGoogle Scholar
  39. Zheng, S., & Zhang, M. (2011). Effect of moisture regime on the redistribution of heavy metals in paddy soil. Journal of Environmental Science, 23(3), 434–443.CrossRefGoogle Scholar

Copyright information

© Springer Science+Business Media Dordrecht 2017

Authors and Affiliations

  • Sung Chul Kim
    • 1
  • Young Kyu Hong
    • 1
  • Se Jin Oh
    • 2
  • Seung Min Oh
    • 2
  • Sang Phil Lee
    • 2
  • Do Hyung Kim
    • 3
  • Jae E. Yang
    • 2
  1. 1.Department of Bio-Environmental ChemistryChungnam National UniversityDaejeonRepublic of Korea
  2. 2.Department of Biological EnvironmentKangwon National UniversityChuncheonRepublic of Korea
  3. 3.Soil Environment CenterKorea Environmental Industry and Technology InstituteSeoulRepublic of Korea

Personalised recommendations