Lead immobilization processes in soils subjected to freeze-thaw cycles

Ecotoxicology and Environmental Safety 192 (2020) 110288

Contents lists available at ScienceDirect

Ecotoxicology and Environmental Safety journal homepage: www.elsevier.com/locate/ecoenv

Lead immobilization processes in soils subjected to freeze-thaw cycles a,b

Lina Du a b

b

b

a

a,∗

b

, Miles Dyck , William Shotyk , Hailong He , Jialong Lv , Chad W. Cuss , Jingya Bie

T a

College of Natural Resources and Environment, Northwest A& F University, Yangling, Shaanxi, 712100, China Department of Renewable Resources, University of Alberta, Edmonton, AB, T6G 2G1, Canada

A R T I C LE I N FO

A B S T R A C T

Keywords: Lead Immobilization Sorption Adsorption Freeze-thaw cycles Soil

Soil freeze-thaw cycles (FTCs) change the physical and chemical properties of soils; however, information is limited about the consequences for heavy metal sorption and desorption. Lead (Pb) sorption isotherms and successive desorption tests were measured for three soils from North China (Chestnut, Lou and Black), following multiple freeze-thaw cycles (0, 1, 3, 6 and 9 FTCs) of −5 °C for 12 h and then +5 °C for 12 h. Lead adsorption dominated the sorption processes for all soils, and sorption capacity increased with additional FTCs. The Freundlich affinity parameter of soils for Pb sorption (i.e. A; Lβ mmol1-β kg−1), was linearly correlated with carbonate content for soils with multiple FTCs. The effects of FTCs on lead adsorption may be more dependent on carbonate and clay contents than organic matter (OM), cation exchange capacity (CEC) and amorphous iron content. Repeated FTCs increased the pH of soil solutions at applied Pb concentrations > 1.4 mmol L−1, which could facilitate formation of inner-sphere complexes of Pb in studied soils. Cation exchange, a weak association, could occupy specific adsorption sites with increasing Pb doses in soils and it can also be facilitated by FTCs. Our results demonstrate the great potential for increasing Pb immobilization with repeated FTCs, by facilitating the formation of both inner-sphere and outer-sphere complexes. Hence, these findings provide useful information on Pb immobilization in contaminated soils that undergo frequent FTCs and offer an additional insight into predicting Pb behavior in cold and freezing environments like the polar regions.

1. Introduction Lead occurs naturally in all soils, including those of temperate regions as well as cold and freezing environments; its occurrence is strongly related to the chemical and mineralogical composition. Lead can also enter the soil via atmospheric deposition or industrial contamination caused by anthropogenic activities (Shotyk et al., 1998; Feng et al., 2019). The mobility and bioavailability of lead in soils are predominantly governed by the reactions involved in metal retention on soil particle surfaces as reflected in sorption-desorption interactions on those materials (Abdelwaheb et al., 2019). The bioavailability of lead in soil has been widely studied on account of the hazards it posing to animal and human health (Kushwaha et al., 2018; Ming et al., 2012). Interactions between heavy metals and soil surfaces (particles or organic matter) are primary determinants of bioavailability, mobility and their fate. In general, the leaching and bioavailability of heavy metals in soils are decreased by inducing various sorption processes: ion exchange or specific adsorption to various solid phases, precipitation and the formation of stable complexes or chelates from the interaction with organic matter (Blaylock et al., 1997; Adriano, 2001; Kumpiene et al., 2008). Inner sphere complexes involve ∗

the formation of covalent bonds, so the adsorbed species are not readily displaced. Metallic ions, such as Pb2+, can form inner-sphere complexes with organic matter and variable-charged soil surfaces and therefore are strongly held (Evans, 1989). The formation of inner sphere complexes is based upon adsorption reactions with the surfaces of the phyllosilicate clay minerals, with OH-groups at the edges and surfaces of oxide and (oxy)hydroxide minerals and with certain functional groups (e.g., the carboxyl, carbonyl and phenolic groups) within humic substances (Evans, 1989; Bradl, 2004). Nonspecific adsorption (or cation exchange) is an electrostatic phenomenon in which cations do not form covalent bonds with the surface, thereby retaining their water of hydration to form only outer-sphere complexes which represent a weak association between the adsorbed ion and the soil particles (Evans, 1989; Bradl, 2004). Precipitation as secondary minerals may involve the formation of a mixed solid through inclusion or co-precipitation, or surface precipitation on a pre-existing solid phase (Sposito, 1986). Lead precipitation may occur in forms such as oxides, hydroxides, carbonates, sulfides, or phosphates. The complexation of heavy metal ions with hydroxyl ions necessitates considerations of the solubilities of metal-ion hydroxide formation. The intrinsic solubility of Pb(OH)2 is much lower

Corresponding author. College of Natural Resources and Environment, Northwest A& F University, Yangling, Shaanxi, 712100, China. E-mail address: [email protected] (J. Lv).

https://doi.org/10.1016/j.ecoenv.2020.110288 Received 9 August 2019; Received in revised form 21 December 2019; Accepted 31 January 2020 Available online 17 February 2020 0147-6513/ © 2020 Published by Elsevier Inc.

Ecotoxicology and Environmental Safety 192 (2020) 110288

L. Du, et al.

from the long-term record of atmospheric Pb deposition (15,000 years) and the widespread use of lead ores since ~ 5000 years ago (Shotyk et al., 1998, 2016). Numerous studies have also been conducted to better understand the mechanisms and factors that govern the adsorption of heavy metals to individual minerals as well as soils of temperate regions (Chen et al., 1997; Gomes et al., 2001; Martı́nez-Villegas et al., 2004; Lu and Xu, 2009; Fonseca et al., 2011). However, few studies have focused on lead sorption in soils of cold and some temperate regions where FTCs are commonplace (Christensen et al., 2013; He et al., 2015), and the impacts of increased FTCs caused by global warming on the geochemical cycling of lead remain unclear. The temperature changes that define soil FTCs differ considerably, but it is generally agreed that the daily maximum temperature must be greater than 0 °C, while the daily minimum temperature must be less than 0 °C (Baker and Ruschy, 1995; Ho and Gough, 2006). This common feature based on the observed diurnal cycle in air temperature was applied to define FTCs here. In this study, a series of controlled batch experiments were performed to investigate the effects of soil FTCs on Pb sorption in three typical different soils from northern China, by: 1) Measuring the sorption isotherms of Pb in different soils after 0, 1, 3, 6 and 9 FTCs; 2) Determining the amount of adsorbed Pb that was desorbed by NH4OAc after adsorption tests. The objectives of this work were to 1) study the variation of Pb sorption behavior caused by FTCs, and 2) estimate the soil properties that facilitate or impede Pb retention and how they are affected by a varying number of FTCs.

than that of Zn(OH)2 and Hg(OH)2, for example. Precipitation will take place only when the metal ion concentration exceeds the intrinsic solubility (Hahne and Kroontje, 1973). It has been suggested that the complexed Pb species are less bioavailable to living organisms including humans compared to free Pb2+ ions (Chlopecka, 1996; Adriano, 2001). It was observed that lead was the most strongly sorbed metal in competitive sorption experiments using Cd, Cu, Zn, and Pb with different soils in Eastern China (Lu and Xu, 2009). Soil pH significantly affects Pb speciation and the charge on soil particle surfaces. Lead undergoes hydrolysis at low pH values and exists primarily as Pb2+ from pH 4 to 7; PbOH+ predominates at pH ~8, while Pb(OH)20 is the major form at pH > 9 (Hahne and Kroontje, 1973; Schulthess and Huang, 1990). Martı́nez-Villegas et al. (2004) showed that Pb sorption capacity was sensitive to pH changes and increased with increasing pH. Over the past half-century, climate warming in the Northern hemisphere has significantly impacted snow cover, snow depth and snowfall events (Solomon et al., 2007). This warming trend is also apparent in northwestern China, where the snow season temperature fluctuated in a manner similar to the global temperature (Jones et al., 1999; Qin et al., 2006). The rise in air temperature will also lead to increased soil temperatures, which strongly influence physical, chemical and biological processes in soils (Sinha and Cherkauer, 2010; Urakawa et al., 2014; Connolly and Orrock, 2015). Associated changes in the dynamics of soil freezing are likely to vary by region, local winter temperature and snow cover; however, both warmer and drier winters can increase the frequency of soil freeze-thaw cycles (FTCs; Henry, 2008). It has been reported that soil warming can also increase the decomposition of soil organic matter (SOM), resulting in reduced cation exchange capacity (CEC) and thus a decreased capacity to retain soil nutrients and heavy metals (Rajkumar et al., 2013; Perelomov et al., 2018). Indeed, warmer temperatures in soils amended with sewage sludge increased the release of metals (Cd, Ni and Zn) that were complexed with organic matter, creating more labile forms (Antoniadis and Alloway, 2001). Freeze-thaw events alter soil structure and aggregate stability (Oztas and Fayetorbay, 2003; Kværnø and Øygarden, 2006; Dagesse, 2013). Frequent FTCs disrupt the structure and particle configuration in soils, reducing penetration resistance and bulk density in agricultural soils over winter (Unger, 1991; Yu et al., 2010). It has also been suggested that the expected increase in FTCs caused by climate change can influence ecosystem carbon and nutrient losses from soils, such as increased nitrogen (N) leaching (Henry, 2008; Urakawa et al., 2014). Multiple freeze-thaw treatments also caused a significant increase in the release of dissolved organic carbon (DOC) (Wang and Bettany, 1993; Grogan et al., 2004; Han et al., 2018). Soil FTCs can also have remarkable impacts on the stability of minerals in contaminated soils, which in turn increases leaching concentrations of heavy metals (Sanchez et al., 2009; Wei et al., 2015). Hafsteinsdóttir et al. (2011) found that pyromorphite (Pb5(PO4)3(Cl, Br, F, OH)) formation was stable over multiple FTCs. Hence, freeze–thaw events may substantially influence the sorption behavior and mobility of lead in soil in seasonally cold ecosystems. The global biogeochemical cycle of lead has been affected by human activities to a greater degree than any other potentially toxic heavy metal (Shotyk and Le Roux, 2005; Javed et al., 2017). This is apparent

2. Materials and methods 2.1. Site description and soil sampling Soil samples were collected from arable fields at three research stations in September 2010: 1) Lou soil from Wugong, Shaanxi Province (34°28′N; 108°22′E); 2) Black soil (silty loam) from Changchun, Jilin province (43°817′N; 125°324′E); and 3) Chestnut soil (silty sand) from Hohhot, Inner Mongolia Autonomous Region (40°83′N; 111°73′E). The mean annual temperature and precipitation from 2018 to 2010 were respectively 13.5 °C and 586 mm in Wugong, 5.4 °C and 645 mm in Changchun, and 7.7 °C and 393 mm in Hohhot (China meteorological administration http://data.cma.cn). December and January are the coldest months, with mean temperatures of 0.6 °C in Wugong, −17.7 °C in Changchun, and −9.8 °C in Hohhot. Winter temperatures in Wugong, Changchun and Hohhot areas are usually variable, with alternating periods of freezing and thawing and several snowmelt events. Freezing and thawing events occur frequently between November and April. Soil samples were collected from upper horizons (0–20 cm), airdried and ground to pass a 2-mm sieve. Basic physical and chemical properties were measured for all soil samples, including: particle-size distribution (dry-sieving and sedimentation method); soil pH (soil-solution-ratio of 1:1 using distilled water); carbonate content (gasometrical method); phosphate content (water-extractable phosphate and colorimetric analysis) (Murphy and Riley, 1962); organic matter (potassium dichromate-oxidation method) (Heanes, 1984); cation exchange capacity (NH4OAc extraction method and flame photometer) (Hajek et al., 1972), and; amorphous iron content (acid ammonium oxalate method and ICP-AES) (McKeague and Day, 1966) (Table 1).

Table 1 Physical and chemical properties of the studied soils. Soil type

pH

Carbonate (g kg−1)

Phosphate (μg L−1)

Organic matter (g kg−1)

CEC (cmol kg−1)

Amorphous iron (g kg−1)

% Clay (< 0.002 mm)

Chestnut soil Lou soil Black soil

8.5 8.1 7.1

28.0 118 1.99

47.0 35.5 29.5

16.3 8.53 32.4

11.6 23.3 31.1

0.72 1.20 1.84

27.1% 39.6% 19.9%

2

Ecotoxicology and Environmental Safety 192 (2020) 110288

L. Du, et al.

2.2. Sorption tests and 0, 1, 3, 6 and 9 FTCs treatments

2.5. Saturation state

Solutions of Pb(II) with concentrations of 0, 0.2, 0.5, 0.7, 1.0, 1.4 and 1.9 mmol Pb L−1 were prepared from reagent grade Pb(NO3)2, and adjusted to around pH 5 using HNO3 (0.01 M NaNO3 was used as background electrolyte). Batch sorption experiments were conducted in plastic centrifuge tubes (50 mL volume). For each soil type, 25 mL of each Pb(II) solution was added to 1 g of soil. Following the batch experiments, Pb(II) solution of 1.9, 2.4, 2.9, 3.9 and 4.8 mmol L−1 was also respectively applied to each 1 g of soil in order to test if the maximum sorption capacity of soils for Pb could be achieved. Each experiment was replicated 3 times. One FTC consisted of solutions being frozen at −5 °C for 12 h at constant temperature and humidity incubator, and then thawed at +5 °C for another 12 h. Freeze-thaw treatments consisted of 0, 1, 3, 6 or 9 FTCs. After each freeze-thaw treatment, the tubes were shaken continuously for 24 h at 120 oscillations per minute and at room temperature (≈25 °C) to reach sorption equilibrium. The samples were then centrifuged at frot = 4000 g for 10 min, and filtered through 0.45μm filter membrane by syringe. The dissolved concentration of Pb(Ⅱ) in each filtrate solution was measured using atomic emission spectrometry (AAS) (AA320CRT, Shanghai Analytical Instruments Co. Ltd., China). Sorbed Pb2+ was calculated as mmol Pb2+ per liter (initial solution) – mmol Pb2+ per liter (filtrate). The pH of the initial solutions and filtrates were also measured. The equilibrium sorption capacity (Qe) (mmol kg−1) for heavy metals was calculated according to the following equation.

The experimental sorption data and fitted isotherms provide no insight as to the actual mechanism of a sorption process in soil unless it can be determined whether a solution is super- or undersaturated (Sposito, 1984; Stumm and Morgan, 2012). As in the determination of the amount sorbed through Eq. (3), the measurements made solely on the aqueous solution phase can be utilized to characterize the saturation state of soil solution with respect to a solid formation (Sposito, 1986). The ion-activity product (IAP) for a proposed surface precipitate can be compared with the corresponding solubility product constant (Ksp) (Sposito, 1981, 1984, 1986). The saturation state (Ω) of a solution with respect to a solid is calculated and defined as follows (Stumm and Morgan, 2012):

Qe =

V (c0 − ce ) m

Ω=

IAP Ksp

(4)

Ω > 1, i.e., IAP > Ksp (supersaturated) Ω = 1, i.e., IAP = Ksp (equilibrium, saturated)

Ω < 1, i.e., IAP < Ksp (undersaturated) The IAP of the assumed precipitates can be estimated using the concentrations of Pb2+ in the equilibrium solution and potential ligands measured in the filtrates (Sposito, 1986). Oxides, oxyhydroxides, hydroxides, silicates, carbonates and phosphates are of very importance and the most probable for lead (Nriagu, 1972; Santillan-Medrano and Jurinak, 1975). Further, the Pb activity and solubility in many natural surface waters were regulated by PbCO3 and Pb phosphate formation (Hem and Durum, 1973). The concentration of carbonate in solution is controlled not only by the ionization of hydrated carbon dioxide, but also by pH value and the partial pressure of gaseous carbon dioxide at 25 °C (Butler, 1991). The concentration of carbonate and IAP with respect to PbCO3 was thereby estimated and calculated as:

(1)

where C0 and Ce are the concentrations of Pb in the initial solution and filtrate, respectively (mmol L−1); V is the solution volume (L), and; m is the mass of the soil sample (kg). 2.3. Desorption tests following sorption tests Following the sorption tests, the soil that remained after filtration was dried at 25 °C, weighed, and resuspended in 25 mL of 0.1 M NH4OAc. The tubes were then shaken for 24 h at room temperature (≈25 °C). The supernatant was then removed and Pb concentrations were analyzed using the same procedures as for the sorption tests. The NH4OAc desorption process was repeated three times (D1 – D3). The non-extractable fraction of the sorbed Pb2+ was calculated as mmol Pb2+ (adsorbed) – mmol Pb2+ (extracted using NH4OAc).

Log [CO32 −] = −18.1 + log PCO2 + 2pH + …

(5)

IAP = (Pb2 +)(CO32 −)

(6)

where [ ] and ( ) respectively represent concentration and thermodynamic activity. The thermodynamic activities can be estimated using the total molar concentrations measured in filtrates collected at sorption equilibrium, by extrapolating the concentration solubility product constants to zero ionic strength, at which point the concentrations and activities are equal (Sposito, 1981). Similarly, the IAP of Pb phosphate was calculated as follows using the measured concentrations of Pb and phosphate in the filtrates at 25 °C:

2.4. Data analysis

IAP = (Pb2 +)3 (PO43 −)2

The experimental isotherm data were analyzed in accordance with both the Langmuir and Freundlich models. The Langmuir equation has the form:

Log Ksp values of −44.6 for Pb3(PO4)2 (Nriagu, 1973; SantillanMedrano and Jurinak, 1975), and −12.9 for PbCO3 (Santillan-Medrano and Jurinak, 1975; Clark and Bonicamp, 1998) at 25 °C were used to compare with their corresponding IAPs.

Qe = Qmax

KCe 1 + KCe

(2)

3. Results

where Qmax is the maximum mass of lead sorbed per mass unit of adsorbent (mmol kg−1), and K is the equilibrium constant (L mmol−1). The Langmuir equation is applicable to homogeneous sorption and it obeys HenryÏs Law at low concentrations (McKay and Porter, 1997). The Freundlich model assumes that different sites with several adsorption energies are involved, according to the relationship:

Qe = ACe β

(7)

3.1. Sorption tests The pH of all three soils was close to neutral-alkaline/neutral (Table 1). The Lou soil was classified as silty-clay-loam soil and had highest carbonate content among three soils. The Black and Chestnut soils were silt-loam soils with higher content of organic matter (OM) than Lou soil. The contents of CEC and amorphous iron were higher for Black and Lou soils than Chestnut soil. The sorption isotherms (viz. Plots of sorbed Pb (Qe) vs. equilibrium Pb concentration (Ce)) were of the L-curve type for most treatments in all soils (Fig. 1), suggesting a high affinity of soil particles for Pb2+

(3)

where A (Lβ mmol1-β kg−1) and β are positive-valued parameters and β is constrained between 0 and 1 (Reddad et al., 2002; Martı́nez-Villegas et al., 2004). 3

Ecotoxicology and Environmental Safety 192 (2020) 110288

L. Du, et al.

Table 2 Freundlich and Langmuir parameters in different soils under different freezethaw treatments. Soil type

Chestnut soil

Lou soil

Black soil

FTCs

0 1 3 6 9 0 1 3 6 9 0 1 3 6 9

FTC FTC FTCs FTCs FTCs FTC FTC FTCs FTCs FTCs FTC FTC FTCs FTCs FTCs

Freundlich

Langmuir

β

A (Lβ mmol1−β kg−1)

R2

Qmax (mmol kg−1)

K (L mmol−1)

R2

0.38 0.41 0.43 0.42 0.40 0.42 0.49 0.44 0.67 0.55 0.39 0.47 0.40 0.43 0.39

4.29 5.18 5.90 6.17 6.95 8.64 8.98 11.2 19.0 23.5 4.89 5.03 6.34 6.18 6.99

0.87 0.83 0.83 0.78 0.82 0.92 0.91 0.69 0.71 0.76 0.96 0.89 0.97 0.98 0.93

0.38 0.45 0.53 0.52 0.81 1.89 1.21 – – – 0.70 0.41 1.04 0.74 1.36

19.1 21.7 22.9 24.8 23.1 20.3 23.2 – – – 18.0 22.2 19.7 21.0 19.7

0.79 0.79 0.82 0.65 0.83 0.55 0.75 – – – 0.76 0.63 0.73 0.75 0.74

Note: “-” indicates that the data could not be fitted using the Langmuir equation.

attributed to a combination of affinity and steric factors (i.e., the number of exchange sites available) (Sposito, 1981, 2008; Giles et al., 1960). The L-curve isotherms indicate a continuous increase in apparent sorption without saturation of the surface sorption sites. The sorption isotherms were best described by the Freundlich model (Table 2). The affinity parameter A (Lβ mmol1-β kg−1) represents Pb sorption capacity, and was highly correlated with different FTCs (0.89 < R2 < 0.98) and carbonate content for each FTC treatment (0.97 < R2 < 1.0), and hence was greatest for the Lou soil, followed by the Black and Chestnut soils. This correlation is consistent with other findings suggesting that CaCO3-rich soil had higher total sorption capacity than clay and sandy soils (Elkhatib et al., 1991). It is apparent from Fig. 2 that the affinity parameter was sensitive to the interaction of carbonates and FTCs.

3.2. Ion-activity product At sorption equilibrium, solubility calculations suggested that solutions were supersaturated with respect to Pb3(PO4)2 formation for all soils (i.e. Ω = IAP/Ksp > 1) (Fig. 3). Calculations also suggested that solutions were supersaturated with respect to PbCO3 at Pb concentrations below 2.4 mmol L−1 in Chestnut soil; however, Ω decreased to become < 1 and remained at similar levels as the Pb2+ concentration increased from 2.4 mmol L−1 to 4.8 mmol L−1, indicating that lead adsorption via formation of inner-sphere complexes or outer-sphere complexes was more probable than precipitation (Fig. S2a). In the Lou soil, Ω < 1 indicated that Pb adsorption occurred below PbCO3 saturation levels at concentrations of 0.2 mmol Pb L−1 with 0, 1, 3 and 9 FTCs, at 0.5 mmol Pb L−1 with 3 and 9 FTCs, and at 1.0 mmol Pb L−1 with 9 FTCs (Fig. 4b). However, Ω was > 1 at intermediate concentrations, and then decreased to be < 1 at the highest concentrations of 3.9 mmol Pb L−1 under 0, 1 and 3 FTCs, and at 4.8 mmol Pb L−1 under 0, 3 and 6 FTCs (Fig. 4b). In the Black soil, Ω < 1 at all Pb concentrations for all freeze-thaw treatments. Hence, it can be inferred that Pb may have been adsorbed via inner-sphere or outer-sphere complexes in Black soil, which had the lowest carbonate content, highest OM content, and highest CEC and amorphous iron among three studied soils (Table 1) (Elkhatib et al., 1991; Appel and Ma, 2002; Moreno et al., 2006).

Fig. 1. Isotherms of Pb sorption in the studied soils with 0, 1, 3, 6 and 9 FTCs in a. Chestnut soil; b. Lou soil; c. Black soil.

4

Ecotoxicology and Environmental Safety 192 (2020) 110288

L. Du, et al.

Fig. 2. Correlations between Freundlich affinity parameters (A) with FTCs (a.) and carbonate content under 0, 1, 3, 6 and 9 FTCs (b.), in Chestnut soil (28.0 g kg−1), Lou soil (118 g kg−1) and Black soil (1.99 g kg−1).

tests may have contributed to the low recovery of Pb. Such low recovery is expected given the large amount of Pb that was sorbed (~80–100%), which was illustrated to have been specifically bound to OM and mineral surfaces in soils, and may have precipitated as PbCO3 and Pb3(PO4)2 in Chestnut and Lou soils. In addition, Pb desorption also increased with higher Pb concentrations in Chestnut and Black soils, suggesting that the Pb may have been bound to the soil particle via weak associations such as outer-sphere complexes (Fig. 3a and c). The change in Pb desorption among 0, 1, 3, 6 and 9 FTCs was inconsistent in Chestnut and Black soils, whereas it clearly decreased with increasing FTCs in Lou soil (Fig. 3b). At Pb concentrations below 2.9 mmol L−1, the retained Pb2+ increased with the increasing number of FTCs in the Lou soil (Fig. 5). The amount of lead that was retained decreased with increasing Pb concentrations in Lou soil, since cation exchange may occur. With doses of Pb > 2.9 mmol L−1, repeated FTCs may facilitate the exchange of Pb2+ with other cations at such a concentration level in Lou soil, since the amount of retained Pb decreased with 1, 3 and 6 FTCs. But specific adsorption still dominated over other sorption processes, in such a concentration level, leading to the increased retention of Pb in Lou soil compared to other soils.

3.3. Effect of FTCs on pH during Pb sorption At applied Pb concentrations above 1.4 mmol L−1, the pH of equilibrium solutions clearly changed with the number of FTCs (Fig. 3). The adsorption of Pb increased with the pH of equilibrium solutions in Chestnut and Lou soils (pH ~ 5–7.5), and also in Black soil (pH ~ 5–7). Below Pb concentrations of 1.4 mmol L−1, the pH of equilibrium solutions were alkaline–neutral (i.e. ~ 7–8 for Chestnut soil and Lou soil and ~6.6–7.5 for Black soil). The pH also decreased slightly after repeated FTCs in soils with nearly 100% Pb sorption (Fig. 3). Accordingly, lead should be specifically bound to clay minerals, OM or iron oxides in these soils below this concentration level, at higher pH. The variability of equilibrium solution pH values increased with multiple FTCs, and pH tended to increase with repeated FTCs at Pb concentrations > 1.4 mmol L−1. Because many cations can form innersphere complexes with variable-charge soil surfaces and thus cations are strongly held, such as Pb2+ (Evans, 1989). With increasing doses, cation exchange may increasingly occupy the specific adsorption sites as the percent of Pb adsorption decreased, and it was also increased by FTCs. The number of FTCs had an inconsistent impact on pH. The pH increased with increasing number of FTCs at metal concentrations > 1.4 mmol L−1 in Lou soil (silty clay soil) (Fig. 3b). This more obvious increase may be is likely related to highest clay and carbonate contents compared to Chestnut and Black soils. As mentioned above, the affinity parameter was highly correlated with carbonate content, and the dispersion of clay microaggregates by multiple FTCs could contribute to increased specific adsorption sites for Pb2+. Furthermore, at Pb concentrations > 1.4 mmol L−1, the increased lead adsorption that was associated with repeated FTCs could be attributed to increasing pH. However, in the Black soil the pH decreased after 9 FTCs at Pb concentrations < 3.9 mmol L−1, while Pb adsorption increased. In addition, Pb adsorption generally increased with the number of FTCs, but its variability among different numbers of FTCs was relatively low in Black soil compared to the other soils.

4. Discussion 4.1. Importance of Pb immobilization via mineral precipitation versus adsorption Given their saturation state, the precipitation of Pb3(PO4)2 is likely to influence sorption reactions in three soils (Fig. S1). Lead phosphates are less soluble than oxides, hydroxides, carbonates, and sulfates under typical surface conditions on Earth (Ruby et al., 1994). Lead phosphate formation can serve as a sink for Pb in ecosystems (Santillan-Medrano and Jurinak, 1975). Therefore, the conversion of soil Pb to pyromorphite-type minerals by precipitation and ionic exchange to immobilize soil Pb and reduce its bioavailability is in widespread use. Various phosphorus-containing materials have also been employed to immobilize Pb in soils (Yang et al., 2001; Cao et al., 2008; Kumpiene et al., 2008). In respect to Pb carbonate, it was hypothesized to be a particularly likely end product for Pb in calcareous soils (Santillan-Medrano and Jurinak, 1975). In fact, the alkaline, calcareous soils of arid and semiarid regions are excellent sinks for Pb (Adriano, 2001). Hence, the formation of PbCO3 precipitates was probable for most treatments in Lou and Chestnut soils but it was less likely for Black soil (Fig. S2). The Lou and Chestnut soils had higher carbonate content than Black soil (Table 1). The highest Pb sorption was found in the Lou, probably

3.4. Desorption tests The sorption of Pb was almost irreversible, with less than 20%, 18% and 10% of sorbed Pb recovered over 3 successive NH4OAc extracts in Chestnut soil, Black soil and Lou soil, respectively (Fig. 4). Obviously, the percent of Pb desorption was observed in the opposite order with affinity parameter, A, from sorption tests of three soils. It is further implied that the carbonate-rich soil had the highest sorption capacity for Pb among these soils. Besides, the relatively low concentration of NH4OAc solution (0.01 M NH4OAc) that was used in the desorption 5

Ecotoxicology and Environmental Safety 192 (2020) 110288

L. Du, et al.

low and in some cases were negligible. Therefore, the adsorption of Pb2+ onto soil surfaces should predominate in these soil solutions. Due to the high affinity of clay minerals, organic matter and iron oxides for lead (Moreno et al., 2006; Covelo et al., 2007), they may be part of a cost-effective remediation strategy for reducing Pb solubility and bioaccessibility. 4.2. Dominant solid phases involved in removing Pb from solution via adsorption The Lou soil also contained the highest proportion of clay (Table 1). Clay content should be a good predictor of a soils’ ability to adsorb metals (Bradl, 2004). Clay minerals, including natural kaolinite, montmorillonite, and zeolite (i.e. synthetic mordenite), are commonly electrically charged, so that metals in the soil solution are attracted to their surfaces. The adsorption of the heavy-metal cations by kaolinite, montmorillonite and mordenite can occur via Si-related sites and Alrelated sites. The affinity of the metal ions towards these solids depends on the pH of the medium, the Lewis acid strength, and the type of surface sites (Schulthess and Huang, 1990). In this study, lead sorption increased with the increasing pH of soil solutions. Lead sorption as a function of pH is related to the decrease in concentration of H+ and the corresponding increase in negative charges on soil surfaces via the deprotonation of hydroxyl groups, decreasing the competition between H+ and metallic cations for sorption sites (Wang et al., 2011). Moreover, soil organic matter (SOM) may immobilize Pb via specific adsorption reactions, while the mobilization of Pb can also be facilitated by complexation with dissolved OM or fulvic acid extracts (Pinheiro et al., 1999; Adriano, 2001). The extent of metal retention by various OM has been the subject of numerous studies, and results typically demonstrate that Pb2+ can be more strongly retained by OM than many other elements (Cd, Ni, Zn, Co, Mn) (Evans, 1989). It also has been observed that Pb forms strong complexes with SOM, and that it can out compete most other metals for adsorption sites on SOM (Strawn and Sparks, 2000). Soil organic matter may also exhibit additive effects with the hydrous oxides of Fe, Mn, Al, which are also important soil constituents for metal sorption (Adriano, 2001). The CEC of soils is often correlated with the OM content and the higher the CEC of a soil, the greater the amount of metals a soil can retain. From the measured CEC (Table 1), and the maximum amount of sorbed Pb estimated from sorption tests in the concentration range of 0–4.8 mmol L−1 (115 mmol kg−1 in Chestnut soil, 116 mmol kg−1 in Black soil and 121 mmol kg−1 in Lou soil) (CEC > Qe), we hypothesize that lead was also occupying exchangeable sites by forming outer sphere complexes (Appel and Ma, 2002; Martı́nez-Villegas et al., 2004). Iron oxyhydroxides also have a high adsoprtion capacity for lead, with specific sorption sites at low pH values and negatively charged sites at high pH (Martı́nez-Villegas et al., 2004). The high affinity of iron oxides for lead allows them to act as long-term sinks (Forbes et al., 1976; Gadde and Laitinen, 1974; McKenzie, 1980). Besides, if iron hydrous oxides were precipitating with Pb2+ in the soil solution to form colloidal material with a relatively large surface area, additional Pb2+ could also adsorb on the surface of the precipitate (Sposito, 1984). Among the two basic adsorption mechanisms, specific binding is dominant for Pb and nonspecific binding is much less important (Pinheiro et al., 1999). Accordingly, the adsorption of Pb2+ on OM and iron oxyhydroxides should be the dominant mechanism of sorption in the Black soil. Based on the results of this study, OM, CEC and amorphous iron are important factors for predicting Pb adsorption capacity in soils, since Black soil adsorbed high amount of Pb and also contained both the greatest CEC and the most OM and amorphous iron of the three soils.

Fig. 3. Effects of 0, 1, 3, 6 and 9 FTCs on pH values in equilibrium solutions and percentage of Pb sorption in a. Chestnut soil; b. Lou soil; c. Black soil.

because it has the highest carbonate content among the three soils. If we assume that Pb phosphate and carbonate precipitates formed and accounted for the sorption process during the mixing of Pb2+ with Lou and Chestnut soils, the concentrations of Pb3(PO4)2, PbCO3 in the soil solution can be estimated (Table S1) by the amount of strongly adsorbed Pb as follows:

[Adsorbed Pb] = C0 − ([Pb3 (PO4 )2] + [PbCO3)]

(8)

4.3. Climate change, freeze-thaw cycles and metal immobilization

Based on this calculation, the amounts of Pb phosphate and carbonate precipitates were clearly higher when Pb concentrations were lower (Table S1). However, the amounts of both precipitates were very

In general, soil warming that is caused by global warming can 6

Ecotoxicology and Environmental Safety 192 (2020) 110288

L. Du, et al.

Fig. 4. Fractions of three successive extractions in the total sorbed Pb2+ with 0, 1, 3, 6 and 9 FTCs in a. Chestnut soil; b. Lou soil; c. Black soil. The terms D1, D2 and D3 represents the fraction of desorbed Pb2+ from each extraction, respectively.

2013). Indeed, warmer temperatures can increase the release of Cu, Ni and Zn from soil (Antoniadis and Alloway, 2001; Li et al., 2012), and increase the accumulation of Ca, Fe and Mo in plants (Sardans et al., 2008). Li et al. (2012) also reported that an increase of 3 °C temperature increased Cu, Zn and Fe concentrations in the leaves (Solanum tuberosum L.), and decreased Cd, Pb, Fe, Zn and Cu concentrations in tubers. Also, Pb mobility in soil was unaffected or very little affected by warmer temperature. However, lead sorption capacity significantly increased with the increasing number of FTCs for our investigated soils. Also, it was concluded that lead sorption is a process of exothermic sorption (Wang et al., 2011). During freezing, solutes tend to be largely excluded from ice and concentrated in the unfrozen water (Perfect et al., 1991; Marion, 1995). Consequently, the exclusion of ions by ice formation may result in high concentrations of Pb2+ and other ions in the unfrozen water near the surfaces of negatively charged clay surfaces, potentially increasing the number of interactions between Pb2+ and clay particles (Perfect et al., 1991). Freezing and thawing may result in either the dispersion or aggregation of clay particles in the flocculated and clay suspensions that were separated from soil solution via centrifugation in the sorption tests (Mostaghimi et al., 1988). On the one hand, the dispersion of clay microaggregates from repeated FTCs increases the specific surface area for Pb adsorption. On the other hand, fine particles may flocculate after FTCs, and even may form particles that are as large as sand. In addition to FTCs, soil solution chemistry affects dispersion/flocculation. In Na-rich alkali soils, clay particles may be present in soil solutions as dilute suspensions (Rowell and Dillon, 1972). However, the enhanced release of soil OM may be

Fig. 5. Percentages of residual Pb in the totally sorbed Pb2+ in the Lou soil after three successive desorption tests with 0, 1, 3, 6 and 9 FTCs.

increase the release of soluble metal ions into the soil solutions via the decomposition of SOM, the lysis of microbial cells, and the destruction of soil aggregates, thereby increasing bioavailability (Rajkumar et al., 7

Ecotoxicology and Environmental Safety 192 (2020) 110288

L. Du, et al.

tests demonstrated that FTCs can increase Pb adsorption capacity. FTCs can therefore have a major impact on the mechanisms of metal sorption, which may be caused by changes in the physical and chemical properties of soil constituents such as carbonate, clay and organic matter, and solution pH. Carbonate-rich soil had the highest sorption capacity for Pb among three soils. Our study provides additional insights into the effect of FTCs on the mechanisms of lead immobilization in soil, and also offers useful information for predicting the Pb sorption ability of soils in cold regions.

another explanation for the significant increase of Pb sorption capacity observed over multiple FTCs. Specific adsorption reactions and the formation of relatively stable complexes with soil OM are very important to the chemistry of Pb in soil. The possible sources of soil OM release by FTCs may include: 1) Physical disruption of soil aggregates, destroying the binding sites between soil and organic matter (Larsen et al., 2002; Han et al., 2018), and; 2) The destruction of hydrogen bonds in macro-molecular organic matter, caused by expansion and contraction processes (Han et al., 2018). A single freezing event stabilized the humus in the soil, whereas repeated freezing and thawing dispersed the humus and increased its mobility (Cheng et al., 1971). As the soil solution freezes and thaws, the shrinking and expansion of soil OM would enhance fragmentation and expose a larger surface area of the organic matter (Wang and Bettany, 1993), which may facilitate the specific adsorption of Pb2+ with organic matter or stable complexation of Pb with organic compounds. All of these processes could release either micro- or macro- organic carbon, and may also facilitate strong adsorption between OM and OM-associated Pb on the surface of soil particles. Total Pb concentrations cannot be used to distinguish adsorption reactions from the complexation or precipitation of Pb with soils and soil constituents. To better understand the fate of Pb in these soils, investigation of Pb speciation in soil solution is warranted. Furthermore, the increased sorption capacity produced by multiple FTCs can also be ascribed to the corresponding increase in pH values in equilibrium solutions. Freezing could increase adsorbed bases, while thawing could increase soil acidity (Marion, 1995). The observed increase in the percentage of lead that was adsorbed with increasing FTCs can thus be attributed to increasing pH after repeated FTCs (Wu et al., 2003; Moreno et al., 2006) (Fig. 3). Repeated FTCs had a more important influence on the Lou soil in this study, which had the highest carbonate and clay contents, suggesting that these soil components may play some role in modulating the pH of soil solutions. However, pH values decreased after 9 FTCs in Black soil, where OM content, CEC and amorphous iron were relatively high, and lead was strongly bound by specific sorption sites on OM and iron oxides, or by forming complexes of outer-sphere. Also, it is highly probable that FTCs facilitate the formation of more carbonates in soil solutions, as freezing is likely to cause the precipitation of CaCO3 in carbonate-rich solutions (Hallet, 1976; Marion, 1995). The increase in retained Pb2+ that was observed after desorption tests with repeated FTCs at Pb concentrations < 2.9 mmol L−1 in Lou soil (Fig. 5) is likely due to increased specific adsorption resulting from increased pH of soil solutions by repeated FTCs. Moreover, such increase may be related to the positive correlation between estimated sorption capacity of Pb and carbonate content in soils with 0, 1, 3, 6 and 9 FTCs. The formation of outer-sphere complexes may also result in the decrease of retained Pb2+ with increasing Pb concentrations of 0.2–2.9 mmol L−1 in Lou soil. The dissolution of metal carbonates in precipitates is another possible reason for the lower concentrations of retained Pb observed at total concentrations < 2.9 mmol L−1 in Lou soil. This could occur because dissolution would increase with decreasing pH, and the pH of soil solutions decreased with increasing Pb concentration (Evans, 1989). Furthermore, precipitation-dissolution and cation exchange are the most important chemical reactions that affected by freezing and thawing, and also crucial for the retention of Pb in soil. Solute exclusion that occurs during ice formation can lead to supersaturated solutions, which promotes the precipitation of secondary minerals in soil, alters solution-phase compositions (which may promote the dissolution of primary minerals), and shifts the equilibrium toward increased mineral weathering (Marion, 1995).

CRediT authorship contribution statement Lina Du: Conceptualization, Data curation, Formal analysis, Funding acquisition, Investigation, Methodology, Resources, Software, Visualization, Writing - original draft, Writing - review & editing. Miles Dyck: Funding acquisition, Supervision, Validation, Writing - review & editing. William Shotyk: Methodology, Supervision, Validation, Writing - review & editing. Hailong He: Funding acquisition, Project administration. Jialong Lv: Conceptualization, Methodology, Funding acquisition, Supervision. Chad W. Cuss: Validation. Jingya Bie: Data curation. Acknowledgments This work was supported by the National Key Research and Development Program of China (grant no. 2017YFD0200205); National Natural Science Foundation of China (NSFC) (grant no. 41501231 and 41877015); and the China Scholarship Council for Lina Du. Appendix A. Supplementary data Supplementary data to this article can be found online at https:// doi.org/10.1016/j.ecoenv.2020.110288. References Abdelwaheb, M., Jebali, K., Dhaouadi, H., Dridi-Dhaouadi, S., 2019. Adsorption of nitrate, phosphate, nickel and lead on soils: risk of groundwater contamination. Ecotoxicol. Environ. Saf. 179, 182–187. https://doi.org/10.1016/j.ecoenv.2019.04. 040. Adriano, D.C., 2001. In: Adriano, D.C. (Ed.), Lead - trace elements in terrestrial environments: biogeochemistry, bioavailability, and risks of metals. Springer New York, New York, NY, pp. 349–410. https://doi.org/10.1007/978-0-387-21510-5_10. Antoniadis, V., Alloway, B.J., 2001. Availability of Cd, Ni and Zn to ryegrass in sewage sludge-treated soils at different temperatures. Water Air Soil Pollut. 132 (3), 201–214. https://doi.org/10.1016/j.earscirev.2017.06.005. Appel, C., Ma, L., 2002. Concentration, pH, and surface charge effects on cadmium and lead sorption in three tropical soils. J. Environ. Qual. 31 (2), 581–589. https://doi. org/10.2134/jeq2002.0581. Baker, D.G., Ruschy, D.L., 1995. Calculated and measured air and soil freeze-thaw frequencies. J. Appl. Meteorol. 34 (10), 2197–2205. https://doi.org/10.1175/15200450. Blaylock, M.J., Salt, D.E., Dushenkov, S., Zakharova, O., Gussman, C., Kapulnik, Y., Raskin, I., 1997. Enhanced accumulation of Pb in Indian mustard by soil-applied chelating agents. Environ. Sci. Technol. https://doi.org/10.1021/es960552a. Bradl, H.B., 2004. Adsorption of heavy metal ions on soils and soils constituents. J. Colloid Interface Sci. 277 (1), 1–18. https://doi.org/10.1016/j.jcis.2004.04.005. Butler, J.N., 1991. Carbon Dioxide Equilibria and Their Applications. CRC Press. Cao, X., Ma, L.Q., Singh, S.P., Zhou, Q., 2008. Phosphate-induced lead immobilization from different lead minerals in soils under varying pH conditions. Environ. Pollut. 152 (1), 184–192. https://doi.org/10.1016/j.envpol.2007.05.008. Chen, Wright, J.V., Conca, J.L., Peurrung, L.M., 1997. Effects of pH on heavy metal sorption on mineral apatite. Environ. Sci. Technol. 31 (3), 624–631. https://doi.org/ 10.1021/es950882f. Cheng, B.T., Bourget, S.J., Ouellette, G.J., 1971. Influence of alternate freezing and thawing on the availability of some soil minerals. Can. J. Soil Sci. 51 (3), 323–328. https://doi.org/10.4141/cjss71-045. Chlopecka, A., 1996. Assessment of form of Cd, Zn and Pb in contaminated calcareous and gleyed soils in Southwest Poland. Sci. Total Environ. 188 (2–3), 253–262. https://doi. org/10.1016/0048-9697(96)05182-0. Christensen, A.F., He, H., Dyck, M.F., Lenore Turner, E., Chanasyk, D.S., Naeth, M.A., Nichol, C., 2013. In situ measurement of snowmelt infiltration under various topsoil cap thicknesses on a reclaimed site. Can. J. Soil Sci. 93 (4), 497–510. https://doi.org/ 10.4141/cjss2012-048. Clark, R.W., Bonicamp, J.M., 1998. The Ksp-solubility conundrum. J. Chem. Educ. 75 (9),

5. Conclusion The soils that were investigated in this work had an average pH of 8 and high affinity for Pb (Chestnut soil, Lou soil and Black soil). The Lcurve shape of adsorption isotherms and three successive desorption 8

Ecotoxicology and Environmental Safety 192 (2020) 110288

L. Du, et al.

Li, Y., Zhang, Q., Wang, R., Gou, X., Wang, H., Wang, S., 2012. Temperature changes the dynamics of trace element accumulation in Solanum tuberosum L. Climatic Change 112 (3–4), 655–672. https://doi.org/10.1007/s10584-011-0251-1. Lu, S.G., Xu, Q.F., 2009. Competitive adsorption of Cd, Cu, Pb and Zn by different soils of Eastern China. Environ. Geol. 57 (3), 685–693. https://doi.org/10.1007/s00254008-1347-4. Marion, G.M., 1995. Freeze-Thaw Processes and Soil Chemistry. Cold Regions Research and Engineering Lab Hanover NH. Martı́nez-Villegas, N., Flores-Vélez, L.M., Domı́nguez, O., 2004. Sorption of lead in soil as a function of pH: a study case in México. Chemosphere 57 (10), 1537–1542. https:// doi.org/10.1016/j.chemosphere.2004.08.099. McKay, G., Porter, J.F., 1997. Equilibrium parameters for the sorption of copper, cadmium and zinc ions onto peat. J. Chem. Technol. Biotechnol. 69 (3), 309–320. https://doi.org/10.1002/(SICI)1097-4660. McKeague, J.A., Day, Jh, 1966. Dithionite-and oxalate-extractable Fe and Al as aids in differentiating various classes of soils. Can. J. Soil Sci. 46 (1), 13–22. https://doi.org/ 10.4141/cjss66-003. McKenzie, R.M., 1980. The adsorption of lead and other heavy metals on oxides of manganese and iron. Soil Res. 18 (1), 61–73. https://doi.org/10.1071/SR9800061. Ming, H., He, W., Lamb, D.T., Megharaj, M., Naidu, R., 2012. Bioavailability of lead in contaminated soil depends on the nature of bioreceptor. Ecotoxicol. Environ. Saf. 78, 344–350. https://doi.org/10.1016/j.ecoenv.2011.11.045. Moreno, A.M., Quintana, J.R., Pérez, L., Parra, J.G., 2006. Factors influencing lead sorption–desorption at variable added metal concentrations in Rhodoxeralfs. Chemosphere 64 (5), 758–763. https://doi.org/10.1016/j.chemosphere.2005.10. 058. Mostaghimi, S., Young, R., Wilts, A.R., Kenimer, A.L., 1988. Effects of frost action on soil aggregate stability. Trans. ASAE 31 (2), 435–439. https://doi.org/10.13031/2013. 30727. Murphy, J., Riley, J.P., 1962. A modified single solution method for the determination of phosphate in natural waters. Anal. Chim. Acta 27, 31–36. https://doi.org/10.1016/ S0003-2670(00)88444-5. Nriagu, J.O., 1972. Lead orthophosphates. I. Solubility and hydrolysis of secondary lead orthophosphate. Inorg. Chem. 11 (10), 2499–2503. https://doi.org/10.1021/ ic50116a041. Nriagu, J.O., 1973. Lead orthophosphates—II. Stability of cholopyromophite at 25 °C. Geochim. Cosmochim. Acta 37 (3), 367–377. https://doi.org/10.1016/00167037(73)90206-8. Oztas, T., Fayetorbay, F., 2003. Effect of freezing and thawing processes on soil aggregate stability. Catena 52 (1), 1–8. https://doi.org/10.1016/S0341-8162(02)00177-7. Perelomov, L.V., Sarkar, B., Sizova, O.I., Chilachava, K.B., Shvikin, A.Y., Perelomova, I.V., Atroshchenko, Y.M., 2018. Zinc and lead detoxifying abilities of humic substances relevant to environmental bacterial species. Ecotoxicol. Environ. Saf. 151, 178–183. https://doi.org/10.1016/j.ecoenv.2018.01.018. Perfect, E., Groenevelt, P.H., Kay, B.D., 1991. In: Bear, J., Corapcioglu, M.Y. (Eds.), Transport phenomena in frozen porous media BT - transport processes in porous media. Springer Netherlands, Dordrecht, pp. 243–270. Pinheiro, J.P., Mota, A.M., Benedetti, M.F., 1999. Lead and calcium binding to fulvic acids: salt effect and competition. Environ. Sci. Technol. 33 (19), 3398–3404. https:// doi.org/10.1021/es990210f. Qin, D.H., Liu, S.Y., Li, P.J., 2006. Snow cover distribution, variability, and response to climate change in western China. J. Climate 19 (9), 1820–1833. Rajkumar, M., Prasad, M.N.V., Swaminathan, S., Freitas, H., 2013. Climate change driven plant–metal–microbe interactions. Environ. Int. 53, 74–86. https://doi.org/10.1016/ j.envint.2012.12.009. Reddad, Z., Gerente, C., Andres, Y., Le Cloirec, P., 2002. Adsorption of several metal ions onto a low-cost Biosorbent: kinetic and equilibrium studies. Environ. Sci. Technol. 36 (9), 2067–2073. https://doi.org/10.1021/es0102989. Rowell, D.L., Dillon, P.J., 1972. Migration and aggregation of Na and Ca clays by the freezing of dispersed and flocculated suspensions. J. Soil Sci. 23 (4), 442–447. https://doi.org/10.1111/j.1365-2389.1972.tb01675.x. Ruby, M.V., Davis, A., Nicholson, A., 1994. In situ formation of lead phosphates insSoils as a method to i mmobilize Lead. Environ. Sci. Technol. 28 (4), 646–654. https://doi. org/10.1021/es00053a018. Sanchez, F., Langley White, M.K., Hoang, A., 2009. Leaching from granular cement-based materials during infiltration/wetting coupled with freezing and thawing. J. Environ. Manag. 90 (2), 983–993. https://doi.org/10.1016/j.jenvman.2008.03.006. Santillan-Medrano, J., Jurinak, J.J., 1975. The chemistry of lead and cadmium in soil: solid phase formation. Soil Sci. Soc. Am. J. 39, 851–856. https://doi.org/10.2136/ sssaj1975.03615995003900050020x. Sardans, J., Peñuelas, J., Prieto, P., Estiarte, M., 2008. Changes in Ca, Fe, Mg, Mo, Na, and S content in a Mediterranean shrubland under warming and drought. J. Geophys. Res.: Biogeosciences 113 (G3). https://doi.org/10.1029/2008JG000795. Schulthess, C.P., Huang, C.P., 1990. Adsorption of heavy metals by silicon and aluminum oxide surfaces on clay minerals. Soil Sci. Soc. Am. J. 54 (3), 679–688. https://doi. org/10.2136/sssaj1991.03615995005500050054x. Shotyk, W., Appleby, P.G., Bicalho, B., Davies, L., Froese, D., Grant‐Weaver, I., Noernberg, T., 2016. Peat bogs in northern Alberta, Canada reveal decades of declining atmospheric Pb contamination. Geophys. Res. Lett. 43 (18), 9964–9974. https://doi.org/10.1002/2016GL070952. Shotyk, W., Le Roux, G., 2005. Biogeochemistry and cycling of lead. Met. Ions Biol. Syst. 43, 239–275. https://doi.org/10.1201/9780824751999.ch10. Shotyk, W., Weiss, D., Appleby, P.G., Cheburkin, A.K., Frei, R., Gloor, M., Van Der Knaap, W.O., 1998. History of atmospheric lead deposition since 12,370 14C yr BP from a peat bog, Jura Mountains, Switzerland. Science 281 (5383), 1635–1640. https://doi. org/10.1126/science.281.5383.1635.

1182. https://doi.org/10.1021/ed075p1182. Connolly, B.M., Orrock, J.L., 2015. Climatic variation and seed persistence: freeze–thaw cycles lower survival via the joint action of abiotic stress and fungal pathogens. Oecologia 179 (2), 609–616. https://doi.org/10.1007/s00442-015-3369-4. Covelo, E.F., Vega, F.A., Andrade, M.L., 2007. Heavy metal sorption and desorption capacity of soils containing endogenous contaminants. J. Hazard Mater. 143 (1–2), 419–430. https://doi.org/10.1016/j.jhazmat.2006.09.047. Dagesse, D.F., 2013. Freezing cycle effects on water stability of soil aggregates. Can. J. Soil Sci. 93 (4), 473–483. https://doi.org/10.4141/cjss2012-046. Elkhatib, E.A., Elshebiny, G.M., Balba, A.M., 1991. Lead sorption in calcareous soils. Environ. Pollut. 69 (4), 269–276. https://doi.org/10.1016/0269-7491(91)90117-F. Evans, L.J., 1989. Chemistry of metal retention by soils. Environ. Sci. Technol. 23 (9), 1046–1056. https://doi.org/10.1021/es00067a001. Feng, W., Guo, Z., Xiao, X., Peng, C., Shi, L., Ran, H., Xu, W., 2019. Atmospheric deposition as a source of cadmium and lead to soil-rice system and associated risk assessment. Ecotoxicol. Environ. Saf. 180, 160–167. https://doi.org/10.1016/j.ecoenv. 2019.04.090. Fonseca, B., Figueiredo, H., Rodrigues, J., Queiroz, A., Tavares, T., 2011. Mobility of Cr, Pb, Cd, Cu and Zn in a loamy sand soil: a comparative study. Geoderma 164 (3–4), 232–237. https://doi.org/10.1016/j.geoderma.2011.06.016. Forbes, E.A., Posner, A.M., Quirk, J.P., 1976. The specific adsorption of divalent Cd, Co, Cu, Pb, and Zn on goethite. J. Soil Sci. 27 (2), 154–166. https://doi.org/10.1111/j. 1365-2389.1976.tb01986.x. Gadde, R.R., Laitinen, H.A., 1974. Heavy metal adsorption by hydrous iron and manganese oxides. Anal. Chem. 46 (13), 2022–2026. https://doi.org/10.1021/ ac60349a004. Giles, C.H., MacEwan, T.H., Nakhwa, S.N., Smith, D., 1960. Studies in adsorption. Part XI. A system of classification of solution adsorption isotherms, and its use in diagnosis of adsorption mechanisms and in measurement of specific surface areas of solids. 786. J. Chem. Soc. 3973–3993. https://doi.org/10.1039/JR9600003973. 0. Gomes, P.C., Fontes, M.P.F., da Silva, A.G., de, S., Mendonça, E., Netto, A.R., 2001. Selectivity sequence and competitive adsorption of heavy metals by Brazilian soils. Soil Sci. Soc. Am. J. 65, 1115–1121. https://doi.org/10.2136/sssaj2001.6541115x. Grogan, P., Michelsen, A., Ambus, P., Jonasson, S., 2004. Freeze–thaw regime effects on carbon and nitrogen dynamics in sub-arctic heath tundra mesocosms. Soil Biol. Biochem. 36 (4), 641–654. https://doi.org/10.1016/j.soilbio.2003.12.007. Hafsteinsdóttir, E.G., White, D.A., Gore, D.B., Stark, S.C., 2011. Products and stability of phosphate reactions with lead under freeze–thaw cycling in simple systems. Environ. Pollut. 159 (12), 3496–3503. https://doi.org/10.1016/j.envpol.2011.08.026. Hahne, H.C.H., Kroontje, W., 1973. Significance of pH and chloride concentration on behavior of heavy metal pollutants: mercury(II), cadmium(II), zinc(II), and lead(II). J. Environ. Qual. 2, 444–450. https://doi.org/10.2134/jeq1973. 00472425000200040007x. Hajek, B.F., Adams, F., Cope, J.T., 1972. Rapid determination of exchangeable bases, acidity, and base saturation for soil characterization. Soil Sci. Soc. Am. J. 36, 436–438. https://doi.org/10.2136/sssaj1972.03615995003600030021x. Hallet, B., 1976. Deposits formed by subglacial precipitation of CaCO3. GSA Bullet. 87 (7), 1003–1015. https://doi.org/10.1130/0016-7606. Han, Z., Deng, M., Yuan, A., Wang, J., Li, H., Ma, J., 2018. Vertical variation of a black soil's properties in response to freeze-thaw cycles and its links to shift of microbial community structure. Sci. Total Environ. 625, 106–113. https://doi.org/10.1016/j. scitotenv.2017.12.209. He, H., Dyck, M.F., Si, B.C., Zhang, T., Lv, J., Wang, J., 2015. Soil freezing–thawing characteristics and snowmelt infiltration in Cryalfs of Alberta, Canada. Geoderma Reg. 5, 198–208. https://doi.org/10.1016/j.geodrs.2015.08.001. Heanes, D.L., 1984. Determination of total organic‐C in soils by an improved chromic acid digestion and spectrophotometric procedure. Commun. Soil Sci. Plant Anal. 15 (10), 1191–1213. https://doi.org/10.1080/00103628409367551. Hem, J.D., Durum, W.H., 1973. Solubility and occurrence of lead in surface water. J. (Am. Water Works Assoc.) 65 (8), 562–568. Retrieved from. http://www.jstor.org/stable/ 41267396. Henry, H.A.L., 2008. Climate change and soil freezing dynamics: historical trends and projected changes. Climatic Change 87 (3–4), 421–434. https://doi.org/10.1007/ s10584-007-9322-8. Ho, E., Gough, W.A., 2006. Freeze thaw cycles in Toronto, Canada in a changing climate. Theor. Appl. Climatol. 83 (1–4), 203–210. https://doi.org/10.1007/s00704-0050167-7. Javed, M.B., Cuss, C.W., Grant-Weaver, I., Shotyk, W., 2017. Size-resolved Pb distribution in the Athabasca River shows snowmelt in the bituminous sands region an insignificant source of dissolved Pb. Sci. Rep. 7, 43622. https://doi.org/10.1038/srep43622. Jones, P.D., New, M., Parker, D.E., Martin, S., Rigor, I.G., 1999. Surface air temperature and its changes over the past 150 years. Rev. Geophys. 37 (2), 173–199. https://doi. org/10.1029/1999RG900002. Kumpiene, J., Lagerkvist, A., Maurice, C., 2008. Stabilization of As, Cr, Cu, Pb and Zn in soil using amendments – a review. Waste Manag. 28 (1), 215–225. https://doi.org/ 10.1016/j.wasman.2006.12.012. Kushwaha, A., Hans, N., Kumar, S., Rani, R., 2018. A critical review on speciation, mobilization and toxicity of lead in soil-microbe-plant system and bioremediation strategies. Ecotoxicol. Environ. Saf. 147, 1035–1045. https://doi.org/10.1016/j.ecoenv. 2017.09.049. Kværnø, S.H., Øygarden, L., 2006. The influence of freeze–thaw cycles and soil moisture on aggregate stability of three soils in Norway. Catena 67 (3), 175–182. https://doi. org/10.1016/j.catena.2006.03.011. Larsen, K.S., Jonasson, S., Michelsen, A., 2002. Repeated freeze–thaw cycles and their effects on biological processes in two arctic ecosystem types. Appl. Soil Ecol. 21 (3), 187–195. https://doi.org/10.1016/S0929-1393(02)00093-8.

9

Ecotoxicology and Environmental Safety 192 (2020) 110288

L. Du, et al.

nitrogen cycling in ten temperate forest ecosystems throughout the Japanese archipelago. Soil Biol. Biochem. 74, 82–94. https://doi.org/10.1016/j.soilbio.2014.02. 022. Wang, F.L., Bettany, J.R., 1993. Influence of freeze-thaw and flooding on the loss of soluble organic carbon and carbon dioxide from soil. J. Environ. Qual. 22 (4), 709–714. https://doi.org/10.2134/jeq1993.00472425002200040011x. Wang, S., Nan, Z., Cao, X., Liao, Q., Liu, J., Wu, W., Jin, W., 2011. Sorption and desorption behavior of lead on a Chinese kaolin. Environ. Earth Sci. 63 (1), 145–149. https://doi.org/10.1007/s12665-010-0678-0. Wei, M.L., Du, Y.J., Reddy, K.R., Wu, H., 2015. Effects of freeze-thaw on characteristics of new KMP binder stabilized Zn-and Pb-contaminated soils. Environ. Sci. Pollut. Control Ser. 22 (24), 19473–19484. https://doi.org/10.1007/s11356-015-5133-z. Wu, Z., Gu, Z., Wang, X., Evans, L., Guo, H., 2003. Effects of organic acids on adsorption of lead onto montmorillonite, goethite and humic acid. Environ. Pollut. 121 (3), 469–475. https://doi.org/10.1016/S0269-7491(02)00272-5. Yang, J., Mosby, D.E., Casteel, S.W., Blanchar, R.W., 2001. Lead immobilization using phosphoric acid in a smelter-contaminated urban soil. Environ. Sci. Technol. 35 (17), 3553–3559. https://doi.org/10.1021/es001770d. Yu, X., Zhang, Y., Zhao, H., Lu, X., Wang, G., 2010. Freeze-thaw effects on sorption/ desorption of dissolved organic carbon in wetland soils. Chin. Geogr. Sci. 20 (3), 209–217. https://doi.org/10.1007/s11769-010-0209-7.

Sinha, T., Cherkauer, K.A., 2010. Impacts of future climate change on soil frost in the midwestern United States. J. Geophys. Res.: Atmosphere 115 (D8). https://doi.org/ 10.1029/2009JD012188. Solomon, S., Qin, D., Manning, M., Averyt, K., Marquis, M., 2007. Climate Change 2007the Physical Science Basis: Working Group I Contribution to the Fourth Assessment Report of the IPCC, vol. 4 Cambridge university press. Sposito, G., 1981. The Thermodynamics of Soil Solutions. Oxford University Press. Sposito, G., 1984. The Surface Chemistry of Soils. Oxford University Press. Sposito, G., 1986. Distinguishing adsorption from surface precipitation. Geochem. Process. Miner. Surface. 323, 217–228. https://doi.org/10.1021/bk-1987-0323. ch011. 1986. Sposito, G., 2008. The Chemistry of Soils, second ed. Oxford University Press, Oxford. Strawn, D.G., Sparks, D.L., 2000. Effects of soil organic matter on the kinetics and mechanisms of Pb(II) sorption and desorption in soil. Soil Sci. Soc. Am. J. 64, 144–156. https://doi.org/10.2136/sssaj2000.641144x. Stumm, W., Morgan, J.J., 2012. Aquatic Chemistry: Chemical Equilibria and Rates in Natural Waters, vol. 126 John Wiley & Sons. Unger, P.W., 1991. Overwinter changes in physical properties of no-tillage soil. Soil Sci. Soc. Am. J. 55 (3), 778–782. https://doi.org/10.2136/sssaj1991. 03615995005500030024x. Urakawa, R., Shibata, H., Kuroiwa, M., Inagaki, Y., Tateno, R., Hishi, T., Oyanagi, N., 2014. Effects of freeze–thaw cycles resulting from win ter climate change on soil

10