A laboratory experiment was conducted to determine the benefits of some soil organic and inorganic amendments on cadmium (Cd) sorption capacity in non-polluted calcareous soil. The non-polluted soil was separately incubated with organic [municipal solid waste compost (MSWC), rice husk biochars prepared at 300°C (B300) and 600°C (B600)] and inorganic amendments (coal fly ash, CFA), zero valent iron (Fe0), and zero valent manganese (Mn0) at 2% and 5% (W/W) for 90 days at 25°C. After incubation, sorption kinetics and sorption isotherms of Cd were determined using batch experiments. Results show that the adjustments enhanced the sorption capacity of soil for Cd, as indicated by generally higher KF and 1/n values (constants of Freundlich equation) and distribution coefficients (Kd) compared with the control. The increase of Cd sorption in the soils with different amendments followed the sequence Mn0>Fe0>B600>CFA>MSWC>B300. Among the sorption isotherms and kinetic equations, the linear forms of Freundlich and simple Elovich equations yielded good prediction of Cd sorption, respectively. Generally, all amendments improved and increased Cd sorption rate. Given that Mn0 treatment had the highest impact on increasing Cd sorption capacity, it can thus be recommended in the immobilization of Cd from polluted soil.
During the last decades, industrial activities have led to increasing heavy metals (HMs) concentration in the environment. Ground water contamination and ecotoxicological impact on plants and animals, which may adversely affect human health, are serious problems caused by the release of HMs to ecosystems. Cadmium (Cd), a well-known soil contaminant, has toxic elements that can harm plants and humans. Chemical fertilizers, sewage sludge, pesticides, and burning of fossil fuels are the major sources of Cd contamination in soils (Alloway, 2013). The adsorption (specific and nonspecific sorption), desorption, and precipitation–dissolution of metals are the main processes controlling the fate and mobility of HMs, including Cd (Strawn and Sparks, 2000; Loganathan et al., 2012). The effective assessment of sorption process may describe the dynamics of HMs in the environment (Boparai et al., 2011). Sorption and precipitation processes are the prevailing mechanisms at low and high concentrations of Cd, respectively (Loganathan et al., 2012). Several factors, such as solute composition and soil characteristics, affect the Cd distribution in soils. Retention of HMs by soils is usually expressed by mathematical equations, which is called sorption isotherms modeling (Maftoun et al., 2004).
In addition, to understand the fate and mobility of soil HMs with time, performing kinetic examinations is necessary (Sparks, 2003). HMs sorption onto soil components is affected by pH, ionic factors, types of sorbents, surface coverage, and redox reactions (Violante et al., 2008). Organic and inorganic amendments may increase, decrease, or have no effect on the sorption of HMs in soils. In recent years, solidification/stabilization (S/S), electrokinetic remediation, soil flushing, and phytoremediation as remediation methods are employed to reduce the negative impacts of HMs on human health in soils (Kumpiene et al., 2008; Vangronsveld et al., 2009; Thewys et al., 2010). The chemical stabilization technique is a remediation method of HMs in soils by sorption or precipitation (Kumpiene et al., 2008; Scanferla et al., 2009). To decrease Cd mobility, the stabilization of Cd using several amendments have been studied. Phosphate-based amendments (Hamon et al., 2002), biosolids (Basta et al., 2001), iron and/or manganese (hydrous oxides and Fe0) (Chen et al., 2000; Saffari et al., 2015a), and lime (Krebs et al., 1998) are common amendments that are used to stabilize Cd through several mechanisms, thus leading to the formation of stable compounds with lower Cd mobility (Kabata-Pendias, 2011).
Sorption of HMs on organic amendments is a result of ion exchange, surface sorption, chelation, coagulation, and peptization (Kabata-Pendias, 2011). In contrast, it seems that ion exchange, surface complexation, hydrophobic interaction, and electrostatic interaction are the major interaction processes between inorganic amendment and HMs (Wahba and Zaghloul, 2007; Saffari et al., 2015b).
The most extensively used amendments for Cd stabilization include various phosphorus-containing amendments, which reduce Cd mobility by ionic exchange mechanism. Cd sorption in calcareous soils has been studied by Maftoun et al. (2004) and Safarzadeh et al. (2011); however, the effects of organic and inorganic amendments on Cd sorption in calcareous soils have not been reported or documented. Therefore, the objectives of the present study are to examine the influences of organic (municipal solid waste compost, two types of biochar of rice husk) and inorganic amendments (coal fly ash, zero valent iron and zero valent manganese) on Cd sorption in calcareous soil as well as to evaluate different sorption isotherms and kinetic equations for describing the retention mechanism of Cd in calcareous soil.
Non-polluted soil sample was collected from surface depth (0–30 cm) of a calcareous soil (Fine, mixed, mesic, Fluventic Calcixerepts) from agricultural fields located at the College of Agriculture, Bajgah, Shiraz, Fars Province, Iran. The soil was air-dried and passed through a 2-mm sieve. Particle size analysis, pH, electrical conductivity (ECe), percentage of calcium carbonate equivalent (CCE), and organic matter (OM) were determined using standard methods (Sparks, 1996). Plant-available HMs were extracted by DTPA (Lindsay and Norvell, 1978) and determined by atomic absorption spectrophotometer. Total contents of HMs (Sposito et al., 1982) were determined using 4 m HNO3. The related properties of soil are presented in Table 1.
|pH||7.8||Soluble Fe in DTPA (mg kg−1)||4.1|
|CCE (%)||39.5||Soluble Cu in DTPA (mg kg−1)||0.92|
|Sand (%)||27||Soluble Mn in DTPA (mg kg−1)||5.6|
|Clay (%)||35||Soluble Pb in DTPA (mg kg−1)||Trace|
|OM (%)||1.4||Soluble Cd in DTPA (mg kg−1)||Trace|
|CEC [Cmol(+) kg−1]||15.8||Total Cd (mg kg−1)||Trace|
|EC (dS m−1)||0.65||Total Fe (mg kg−1)||30 554|
Six various amendments, including coal fly ash (CFA), municipal solid waste compost (MSWC), rice husk biochars prepared at 300°C (B300) and 600°C (B600), zero valent iron (Fe0), and zero valent manganese (Mn0) were used in the present study. CFA and MSWC were collected from the Zarand coal washing factory and the recycling and the municipal solid waste compost factory of Kerman, Iran, respectively. Biochars were prepared at 300°C and 600°C from rice husk. Husk samples were covered with aluminum foil (to simulate limited oxygen accessibility for the period of wildfires) and placed in a preheated muffle furnace for 4 h to produce biochars. Zero valent iron grit (Fe0:99.5%, Cl−≤20 mg kg−1, S2−≤100 mg kg−1, As ≤5 mg kg−1, Cu≤100 mg kg−1, Mn≤1000 mg kg−1, Ni≤500 mg kg−1, Pb≤20 mg kg−1, Zn≤50 mg kg−1), and zero valent manganese grit (Mn0:99.5%, Cu≤50 mg kg−1, Fe≤200 mg kg−1, Pb≤100 mg kg−1) were purchased from Sigma-Aldrich. The selected properties of amendments used are presented in Table 2.
|SIO2 (%)||Al2O3 (%)||TiO2 (%)||Fe2O3 (%)||CaO (%)||Bao (%)|
|SrO (%)||MgO (%)||K2 O (%)||Na2 O (%)||SO3 (%)||P2 O5 (%)|
|Mn3 O4 (%)||pH||C (%)||H (%)||N (%)||O (%)|
|MSWC||pH||EC (dS m−1)||OM (%)||Cu (mg kg−1)||Zn (mg kg−1)|
|Fe (mg kg−1)||Mn (mg kg−1)||Pb (mg kg−1)||Cd(mg kg−1)||Ni (mg kg−1)|
|pH||EC (dS m−1)||C (%)||H (%)||N (%)||H/C|
All amendments, including CFA, MSWC, biochars of rice husk (B300 and B600), Fe0, and Mn0, were added to each non-polluted soil sample separately, at two levels (2% and 5% W/W) (Table 3), and each soil sample was mixed thoroughly. The soil samples were incubated for 90 days at 25°C. The moisture was kept at about field capacity (FC) by adding distilled water to a constant weight. After the incubation period, samples were air-dried and used for Cd sorption studies.
|Treatment||Amendment (applied rate)||Treatment||Amendment (applied rate)|
|S1||MSWC (2%)||S7||B600 (2%)|
|S2||MSWC (5%)||S8||B600 (5%)|
|S3||CFA (2%)||S9||Fe0 (2%)|
|S4||CFA (5%)||S10||Fe0 (5%)|
|S5||B300 (2%)||S11||Mn0 (2%)|
|S6||B300 (5%)||S12||Mn0 (5%)|
To determine Cd sorption isotherms in a batch experiment, one g of soil sample was placed into centrifuge tubes and 25 mL of 0.01 m Ca(NO3)2 (to keep the ionic strength) containing different levels of Cd (50, 100, 200, 400, 600, 800, 1200, and 1800 mg L−1) as Cd(NO3)2 was added to each tube in three replicates. The suspensions were shaken for 1 h and left for 22 h at 25°C, shaken for another 1 h, and then centrifuged immediately at 3000 rpm. The samples were filtered using Whatman 42 filter paper, and Cd concentrations were determined by atomic absorption spectrophotometer (Shimatzu AA-670G). The amount of trace elements sorbed by soil is calculated using Equation (1)
where q is the amount of metal sorbed per unit mass of soil (mg kg−1), Ci is the initial concentration of the species in solution (mg L−1), Cf is the equilibrium concentration of the species in solution (mg L−1), V is the solution volume (mL), and W is the weight of air-dried soil samples (kg). The sorption data were fitted to linear forms of Freundlich and Langmuir, Temkin, Gunary, and Edie-Hofstee (Table 4).
|Linear form of Freundlich||Log X=Log KF+(1/n) Log C||KF: distribution coefficient (L kg−1), 1/n: Freundlich constant|
|Linear form of Langmuir||C/X=[1/(KL b)]+(1/b)C||KL: Langmuir coefficient (L kg−1), b: the maximum amount of Cd adsorbed by the soil (mg kg−1)|
|Temkin||X=K1T+K2T LnC||K1T and K2T: Temkin sorption constants|
|Gunary||C/X=K1G+K2G C+K3G C1/2||K1G, K2G and K3G: Gunary sorption constants|
|Edie-Hofstee||X=bE–X/KEC||bE, Edie-Hofstee adsorption maximum, KE: Edie-Hofstee sorption constant|
X and C are the amounts of Cd adsorbed per unit weight of soil (mg kg−1) and the concentration of Cd in equilibrium solution (mg L−1), respectively.
To assess Cd kinetic sorption, 1 g of each soil sample was placed in a polypropylene tube to which 30 mL of 80 mg L−1 Cd as Cd(NO3)2 solutions was added. Samples were shaken for time periods of 30, 60, 180, 360, 540, 720, 1440, 2160, 2880, 3600, and 6000 min at 25°C and then centrifuged at 3000 rpm. The supernatant was filtered, and the concentration of Cd in the clear extract solution was determined using atomic absorption spectrophotometer (Shimatzu AA-670G). The amount of Cd adsorbed was calculated as the difference between initial and final Cd concentrations [Equation (1)]. Different kinetic equations were used to describe the sorption kinetic capacity of Cd (Table 5). To determine the best-fitted model for Cd sorption, a standard error of estimate was calculated for each equation. Relatively high values of coefficients of determination (R2) and low values of standard errors of estimate (SE) were used as criteria for the selection of the best fitted models. The standard error is calculated with Equation (2)
where E and E′ are the measured and calculated amounts of sorbed Cd, respectively, and n is the number of measurements. The percentage of Cd removal was calculated from Equation (3)
|Zero order||qt=q0–K0 t||K0, zero order rate constant (mg Cd kg−1 h−1)|
|First order||ln qt=ln q0–K1t||K1, first-order rate constant (h−1)|
|Second order||1/qt=1/q0–K2t||K2, second-order rate constant [(mg Cd kg−1)−1]|
|Third order||1/qt2=1/q02–K3t||K3, third-order rate constant [(mg Cd kg−1)−2 h−2]|
|Parabolic diffusion||qt=q0–Kp t1/2||Kp, diffusion rate constant [(mg Cd g−1)−0.5]|
|Simple Elovich||qt=1/β ln (αsβs)+(1/βs) ln t||αs, initial desorption rate (mg Cd kg−1 h−1),
βs desorption constant [(mg Cd kg−1)−1]
|Two-constant rate||qt=atb||a, initial desorption rate constant (mg Cd kg−1 h−1)b
b, desorption rate coefficient [(mg Cd kg−1)−1]
q0 and qt are the amount of Cd desorption (mg Cd kg−1) at time zero and t, respectively.
The regressions of linear and other statistical analyses were performed using Microsoft Excel 2007 and SPSS V19.
Analysis of sorption kinetic experiment showed that 24 h was sufficient time to reach equilibrium. Hence, sorption isotherm experiment was conducted for 24 h. The changes in the amount of Cd adsorbed by the amended soils with different levels of amendments are shown in Figure 1. The trend of Cd sorption isotherm showed that the amount of sorbed Cd to soils increased as the equilibrium metal concentrations increased. As a result, the slope of the isotherm gradually decreased with increasing concentration of Cd (500–1000 mg L−1), which could be due to the decreasing empty sorption sites for Cd. On the basis of the Giles et al. (1974) classification, L-curve isotherms were observed in all soil samples. This type of curve indicates a relatively high affinity of solid phase at low concentration (Sparks, 2003). Similar results have been reported by others (Gao et al., 1997; Yang et al., 2008; Safarzadeh et al., 2011). The addition of amendments significantly increased curve slope compared with the control treatment. Organic matter (Moral et al., 2002), pH (Appel and Ma, 2002; Safarzadeh et al., 2011), and total soil Cd (Kabata-Pendias and Pendias, 2001) were the most important soil properties affecting Cd sorption. The addition of amendments in soils changed chemical properties during incubation time, which affected sorption capacity.
According to Figure 1, organic and inorganic amendments increased the sorption capacity of soil compared with the control treatment. In addition, Cd sorption increased with increasing rate of applied amendments (W/W). The highest Cd sorption capacity was observed in 5% (W/W) of added amendments. Several isotherm equations were evaluated to describe heavy metal sorption by soils. However, the coefficients of isotherm equations could not give details on the sorption mechanisms, although they may be useful in evaluating the sorption capacity of sorbents (Harter, 1991). The linear form of Freundlich and Langmuir, as indicated by high R2 values and low SE values, significantly described the fit of data for Cd sorption in all soil samples (Tables 6 and 7). Figure 2 shows the experimental sorption rate data, and the fitted curves obtained with the linear form of Freundlich and Langmuir equations for selected amended soil (soil treated with Mn05%) and control soil. Even though the R2 valued of the Gunary and Temkin equations (Tables 8–10 ) were high, the values of SE were also high, so these equations were not appropriate to describe Cd sorption process. The suitability of the Freundlich equation in describing Cd sorption has been reported by others (Usman, 2008; Shaheen, 2009).
Wong et al. (2007) studied Cd sorption on amended soils with dissolved organic matter, and reported that Langmuir and Freundlich equations were able to describe Cd sorption appropriately. Karak et al. (2014) studied the isotherms of Cd sorption from different soils of India. They observed that both Freundlich and modified Langmuir isotherms were the best-fitted equations that can be used to describe Cd sorption. Carrillo Zenteno et al. (2013) studied Cd sorption in selected amended soils by vermicompost, sugarcane filter cake, palm kernel pie, lime, phosphate rock, or zeolite. They observed that the most effective amendments for Cd sorption were lime and zeolite; furthermore, Langmuir sorption isotherms served as the best-fitted model to describe Cd retention in their research.
Freundlich equation parameters (KF and 1/n) are presented in Table 6. The calculated Freundlich KF and 1/n for amended soils ranged from 1536.6 to 2556.2 and from 0.354 to 0.394, respectively. In the case of control, these parameters were 1562.5 and 0.351, respectively. The higher slope (1/n) value of the Freundlich sorption isotherm indicated that small changes in concentration of Cd solution resulted in a major change in the amount of Cd adsorbed (Ali et al., 2013). The value of 1/n did not increase significantly with increasing rate of applied amendments (2% or 5% W/W). The obtained value of 1/n<1 in soil samples indicated least competition between Cd and water molecules for sorption sites on soils (Giles et al., 1974). Bhal and Toor (2002) explained that the value of slope of the Freundlich equation (1/n) indicates rate of adsorption of metal, and a decrease in 1/n value would indicate a reduction in metal sorption. The addition of amendments caused a considerable increase in 1/n and KF. The values of 1/n for amended soils were 0.84–10.91 times greater than those for control soil. The highest and lowest increase in 1/n were observed following the application of Mn05% (5% W/W) and Fe02% (2% W/W). Metal oxides adsorbed Cd on the surface OH groups via Cd-proton exchange or at the negatively charged surface sites (Naidu et al., 1997).
In general, amendments enhanced soil sorption capacity for Cd as indicated by higher KF and 1/n values compared with the control treatment. In addition, the Langmuir equation was used to estimate the sorption maximum (b) for predicting Cd sorption capacity and comparing different soils (Borling, 2003). Sorption maximum (b) for treated soils and control soil were 24 419 (average) and 21 288 mg kg−1, respectively (Table 7) indicating that the sorption capacity of treated soils was higher than untreated soil. The higher values of KL (Langmuir coefficients) were obtained for some amended soils (S3, S4, S11, and S12) compared with untreated soil. Such higher values indicate that these amended soils had stronger affinity for cadmium sorption than untreated soil.
According to the theory proposed by Bolt and Bruggenwert (1976), at a very low equilibrium concentration, 1+KLC=1, and so the nonlinear of Langmuir equation [X=(KLbC)/(1+KLC)] is rearranged to the form X/C=KLb. X/C is termed the distribution coefficient (Kd) and has been nominated as the maximum buffering capacity by Iyengar and Raja (1983); hence, Kd=KLb. Relatively high values of Kd indicate that the adsorbed Cd by solid phase occurred through sorption reactions (Shaheen, 2009). The calculated Kd values for amended soils and control were 191.9–347.5 L kg−1 and 211.32 L kg−1, respectively. The maximum Kd value was found in the Mn0 treatment, followed by those for the treatments CFA and MSWC. It seems that surface complexation plays an important role in the Cd immobilization in Mn0-treateded soils. In other words, Mn0 decreases the availability of Cd in soils because of the Cd sorption on the surface of this amendment, which possesses a large surface area. Chen et al. (2000) studied the effect of various amendments on Cd mobility, and showed that the addition of manganese oxide reduced the extractability of Cd in soils, and significantly reduced the uptake of Cd by wheat shoots. Alkaline materials, like CFA, showed high sorptive capacity for HMs (Iyer and Scott, 2001).
Increasing HM sorption by CFA is due to increasing pH and specific surface area, which in turn, lead to the precipitation of insoluble phases that promote metal sorption via surface complexation (Saffari et al., 2015a,b). Shaheen and Tsadilas (2010) reported that the applications of fly ash in acidic soil increased the Kd values for Cd and Pb. The process of Cd sorption in MSWC-treated soils could involve a number of possible mechanisms, such as 1 – Cd exchange with cations associated MSWC, which may cause the co-precipitation inner-sphere complexation with complexed humic matter and mineral oxides of MSWC as well as 2 – physical sorption and surface precipitation (Saffari et al., 2015a). Shuman et al. (2002) also showed that composted biosolid amendments could be useful in decreasing Cd uptake by plants from contaminated soil.
The results of Cd sorption on treated and untreated soils as a function of time are presented in Figure 3. The percentages of Cd sorption on amended and unamended soils increased with increasing contact time until equilibrium was obtained. The results indicated that all samples had the similar trend. As can be seen in Figure 3, Cd sorption includes two phases (biphasic process); the first phase is a rapid sorption (96%–99%) of Cd by the soils (in the first 180 min) followed by a slow sorption process. Hence, large amounts of sorption sites exist at the beginning of sorption, and subsequently the sorption sites are gradually filled up, thereby decreasing sorption (Huang Guan et al., 2012).
Trivedi and Axe (2000) reported that Cd sorption on hydrous Al, Fe, and Mn oxides is a biphasic process, a quick sorption to the external surfaces followed by slow intraparticle diffusion along the oxide micro pore walls on the internal surface (Loganathan et al., 2012). The removal percentage of Cd in soil samples after 1440 min (2160, 2880, 3600, and 6000 min) was 100%, and no Cd traces were detected in equilibrium solutions. Equilibrium was achieved after about 540 min, with sorption capacity of 2354–2391 mg kg−1. The percentage removal of Cd on amended soils in the beginning of sorption kinetic (At the end of the first 30 min) ranged from 96.8% to 98.2%. Whereas, the value of Cd removal for control soil was 96.5%. At the end of sorption kinetic (1440 min) the percentage of Cd removal increased to 98.5%–99.6% and 98.6% for amended and control soils, respectively. The maximum and minimum removals were observed in amended soil with Mn05% and control soil, respectively. In addition, there was no significant difference between the level of applied amendments (2% or 5%) in Cd sorption capacity.
High sorption affinities of manganese oxide for Cd have been reported by Fu et al. (1992) and Chen et al. (2000). The zero order, first order, second order, third order, power function, and simple Elovich equations were used to evaluate the sorption mechanism on amended and unamended soils. Based on of R2 and SE values (Table 11) the best-fitted equation for describing Cd sorption was the simple Elovich equation. Although power function equation yielded a similar R2, it also produced a higher SE value, which means that this equation is not suitable to describe Cd sorption (Table 11).
|Simple Elovich||Ln αs||198.35||189.52||198.35||199.01||128.30||168.48||237.00||241.88||194.98||269.61||416.49||323.22||158.90|
Figure 4 shows the experimental Cd sorption rate data and the fitted curves obtained with the Elovich and Power function equations for selected amended soil (soil treated with Mn0 5% W/W) and control soil. Given that numerical values of αs (Constant of simple Elovich) were very high, its value was shown as ln αs (Table 11). Cd Sorption kinetics were weakly described by zero order, first order, second order, and third order equations as indicated by low values of R2. Wahba and Zaghloul (2007) found that the Elovich equation is the best model to characterize Cd sorption on montmorillonite, kaolinite and calcite.
Ghasemi-Fasaei and Jarrah (2013) studied kinetics of Cd sorption from highly calcareous soils and reported that power function served as the best equation to describe Cd sorption. Chien and Clayton (1980) stated that the rate of release Cu increased when the value of αs increased or βs decreased. The result of the present study showed that application of amendments increased βs and was higher than the control soil, indicating that the addition of amendments increased Cd sorption capacity; furthermore, the values of ‘a’ constant (constant of power function equation) were lower in amended soils than in control soil. Kuo and Mikkelsen (1980) concluded that the decreasing value of ‘a’ probably indicate a decrease in the rate of zinc desorption in soils. This finding is in agreement with the assumptions made about the role of amendments in the increasing Cd sorption capacity. Heavy metals are adsorbed on the surface of colloidal particles in soils, including humus, hydrous oxides of Fe, Mn and Al, alumino-silicate clays, and some soluble salts, such as calcium carbonate (Alloway, 2013). Biochars can decrease the mobility of HMs (Liu and Zhang, 2009; Beesley and Marmiroli, 2011). The surface area of biochar can increase by increasing the pyrolysis temperature (Ladygina and Rineau 2013). Hence, it can be expected that Cd sorption capacity in B600 would be higher than B300.
Cd sorption kinetic in amended soils (by two types of biochars) showed that there was no significant difference between B300 and B600. Ahmad et al. (2012) studied the effects of a biochar on Pb sorption in soils and reported that application of biochar decreased Pb mobility. Iron compounds have a high sorption capacity for HMs (Lombi et al., 2002; Brown et al., 2005). The influence of MSWC on Cd sorption was similar with other amendments, and the value of βs was higher than in control soil. However, there was no significant difference between levels of applied amendment (2% or 5% W/W).
Meanwhile, the sorption of HMs by organic matter depends on soil pH and degree of organic matter humification (Kumpiene et al., 2007). Shuman (1999) explained that physical and chemical sorption are two main mechanisms of the sorption of HMs by solid organic matter. Cao and Ma (2004) applied compost to remediate chromated copper arsenate-contaminated soils and observed that arsenic (As) sorption by organic matter decreased the uptake of As by vegetables. Vaca-Paulin et al. (2006) studied the sorption of Cd and Cu in soils amended with sewage sludge or compost, and observed that application of sewage sludge or compost to soil increased sorption of Cu and Cd. According to the results, all organic and inorganic amendments generally improved the kinetic Cd sorption; however, application of Mn0 had the strongest effect on Cd sorption capacity. The comparison among amendments showed that Mn0, Fe0 and B600 showed the highest sorption capacity compared with other treatments. Generally, all amendments insignificantly increased Cd sorption capacity compared with control soil.
In the present study, the addition of soil organic and inorganic amendments have been evaluated to determine their abilities to decrease Cd release into the soil and the environment. The application of amendments to soil induced changes in Cd sorption according to the amendment properties. Zero valent manganese and zero valent iron increased Cd sorption in soil, due to Cd sorption on the surface of these amendments, which possess large surface areas. The addition of coal fly ash increased Cd sorption by increasing soil pH. Biochar prepared at 300°C was not so effective for Cd stabilization; it seems that a high amount of calcium carbonate in studied soil prevented effective Cd sorption in this case. However, the application of biochar prepared at 600°C in soils significantly increased Cd sorption rate, possibly due to the increasing specific surface area that promoted Cd sorption through surface complexation. As a result, soluble Cd compounds applied to soils are sorbed rapidly by soil components. Cd sorption onto amended soils appeared to follow L-shaped sorption model, indicating a relatively high affinity of solid phase at low concentration. The linear form of Freundlich and Langmuir well described the fit of data for Cd sorption in all soil samples.
Curve slopes in amended soils were significantly higher than the control soil, meaning that addition of amendments increased Cd sorption in soils as compared with control treatment. Overall, the amendments enhanced sorption capacity of soils for Cd as indicated by generally higher KF and 1/n values compared with those for the control treatment. Based on higher R2 and lowest SE values compared with other equations evaluated, the best equation for describing Cd sorption is the simple Elovich equation. High values of βs in amended soils obtained from simple Elovich equation confirmed the higher sorption capacity of Cd compared with the control treatment. Generally, all amendments (MSWC, CFA, B300, B600, Fe0, and Mn0) improved and increased Cd sorption, but Mn0 treatment had the highest impact on Cd sorption capacity.
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