Aging mechanism of copper added to bentonite

Geoderma 147 (2008) 86–92

Contents lists available at ScienceDirect

Geoderma j o u r n a l h o m e p a g e : w w w. e l s ev i e r. c o m / l o c a t e / g e o d e r m a

Aging mechanism of copper added to bentonite☆ Shi-Wei Zhou a,b, Ming-Gang Xu b, Yi-Bing Ma b,⁎, Shi-Bao Chen b, Dong-Pu Wei b a

State Key Laboratory of Soil and Sustainable Agriculture, Institute of Soil Science, Chinese Academy of Sciences, Nanjing 210008, China Key Laboratory of Plant Nutrition and Nutrient Cycling, Ministry of Agriculture, Institute of Agricultural Resources and Regional Planning, Chinese Academy of Agricultural Sciences, Beijing 100081, China

b

a r t i c l e

i n f o

Article history: Received 27 December 2007 Received in revised form 22 July 2008 Accepted 11 August 2008 Available online 7 September 2008 Keywords: Aging Copper Micropore diffusion Bentonite

a b s t r a c t The extractability, bioavailability and toxicity of metals in soils decline with time owing to aging processes. The processes have important impact on the ecological risk assessment of metals in soils. In this study, a sequential extraction procedure was used to investigate the aging processes of copper added to bentonite and the effects of temperature and pH on the aging processes; X-ray diffraction technique was used to study the changes of basal spacing of bentonite saturated with Cu at different pH. The results from the sequential extraction demonstrated that the proportions of the most labile Cu fractions (viz. water-soluble, NH4NO3 extractable and EDTA extractable Cu) to total added Cu decreased with increasing incubation time, whereas the potentially labile fraction extracted by 0.5 mol L− 1 HCl and inert fractions (6 mol L− 1 HCl extractable Cu and residual Cu) exhibited the opposite trend. Also the aging of added Cu was a slow process (more than one year) where the easily extractable Cu species on soil solid surfaces gradually transformed into less extractable Cu forms which possibly diffused into interlayer of bentonite. The aging experimental data were well fitted using a parabolic diffusion equation (R2 = 0.690–0.943, p b 0.0001) with apparent diffusion rate coefficients (D/r2) of 6.43–20.0 × 10− 6 d− 1 and activation energy of diffusion (Ea) of 38.3–56.1 kJ mol− 1, indicating that micropore diffusion could be the main mechanism of aging processes of Cu added to bentonite. The results of X-ray diffraction showed that the basal spacing of bentonite decreased with increasing pH, which was from 1.51 nm at pH 4.42 and 5.51, 1.37 nm at pH 6.40, to 1.26 nm at pH 7.68. It was proposed that Cu2+ was dominant diffusive ions at low pH (b5.5), while CuOH+ became dominant at relatively high pH. During longterm aging, copper ions in interlayer of bentonite might diffuse into hexagonal cavities of clay mineral by dehydration. © 2008 Elsevier B.V. All rights reserved.

1. Introduction The term “aging” describes the processes by which the extractability, bioavailability and toxicity of copper added to soils decrease with time (Scott-Fordsmand et al., 2000; Bruus Pedersen and Van Gestel, 2001; McLaughlin, 2001; Lock and Janssen, 2003; TomPetersen et al., 2004; Lu et al., 2005; Ma et al., 2006a). These processes, which are commonly observed for several elements, are sometimes also termed “natural attenuation” or “irreversible sorption”, etc. (McLaughlin, 2001; Brady et al., 2003; Ma et al., 2006b). This means that there is great difference in bioavailability/toxicity of copper between long-term field soils and freshly spiked soils, thus, it is much important and necessary to understand and predict the aging processes in ecological risk assessment of metals in soils. However, it

☆ A statement:The manuscript is original work, not published elsewhere or under consideration for publication elsewhere. All authors have seen the manuscript and agree to its submission to Geoderma. ⁎ Corresponding author. Postal address: South Street 12, Zhongguancun, Haidian District, Beijing 100081, China. Tel.: +86 10 6211 2103; fax: +86 10 6897 5161. E-mail address: [email protected] (Y.-B. Ma). 0016-7061/$ – see front matter © 2008 Elsevier B.V. All rights reserved. doi:10.1016/j.geoderma.2008.08.003

is always be considered that the reaction of added Cu in soils is rapid in equilibrium, and the current soil quality criteria are often derived from the toxicological effects of freshly spiked soils, which may be overestimated owing to ignoring the long-term behavior of metals in soils (Alexander, 2000; McLaughlin, 2001; Renella et al., 2002; Lock and Janssen, 2003; Tom-Petersen et al., 2004). So far several explanations have been proposed for aging reactions of metals added to soils, including (i) diffusion into micropores and interstices on soil minerals or organic materials, or solid-state diffusion into the crystalline structure of soil minerals (Barrow, 1986; Bruemmer et al.,1988; He et al., 2001; Axe and Trivedi, 2002; Bourg et al., 2003; AlQunaibt et al., 2005; Karmous et al., 2006; Ma et al., 2006b); (ii) changes in the forms of surface complexes, e.g., outer-sphere vs. innersphere complexes, mononuclear vs. multinuclear complexes, or formation of surface clusters or precipitates (McBride, 1991; Du et al., 1997; Karthikeyan et al., 1999; Sparks, 2003; Alvarez-Puebla et al., 2005; Hyun et al., 2005); and (iii) occlusion by soil minerals or organic materials via recrystallization or coprecipitation (Martínez and McBride, 1998, 2000; Schosseler et al., 1999; Elzinga and Reeder, 2002). In brief, the aging mechanisms of metals in soils are basically micropore diffusion, surface nucleation/precipitation and occlusion, as

S.-W. Zhou et al. / Geoderma 147 (2008) 86–92

speculated by McLaughlin (2001). Furthermore, based on these suppositions, Ma et al. (2006a,b) developed a semi-mechanistic model, where the short-term (30 days) and long-term (2 years) aging of copper added to 19 European soils were modeled well. In the model, the calculated activation energy (33–36 kJ mol− 1) and apparent diffusion rate coefficient (0.66–20 × 10− 10 s− 1) suggested the micropore diffusion was main aging process of added Cu (Ma et al., 2006a,b). However, it is needed that the more direct evidence to support the mechanism of Cu aging in soils. Under heating copper ions adsorbed on montmorillonite may diffuse into hexagonal cavities of clay mineral by dehydration, which was proved by X-ray diffraction (XRD) (He et al., 2001; Karmous et al., 2006). But, under normal soil temperature, it is unclear whether this diffusion occurs. Obviously, the species of copper ions on soil mineral surfaces are essential to explain the aging process. Since it was used by Tessier et al. (1979) to investigate the fractionation of heavy metals in sediments, sequential extraction procedure (SEP) has become a conventional method for assessing the bioavailability of heavy metals in soils and sediments (Shuman, 1985; Davidson et al., 1994, 2006; Ahnstrom and Parker, 1999; Kaasalainen and Yli-Halla, 2003; Tüzen, 2003; Pueyo et al., 2008). Moreover, in recent years, it has been widely used to study the aging of added Cu in soils (Brady et al., 2003; Inaba and Takenaka, 2005; Lu et al., 2005; Arias-Estevez, 2007; Jalali and Khanlari, 2008). Especially, Ma and Uren (1998) developed a SEP to study the long-term (430 days) aging of zinc in bentonite, by which combined with XRD they approved the diffusion of zinc ions into hexagonal cavities of bentonite. As a low cost and ubiquitous expandable montmorillonite in soils, bentonite could be utilized to reduce the release of Cu2+ from soils into solution, and it was more effective to retain heavy metals for aged soils (Ling et al., 2007). Thus, in this study, SEP and XRD were used to investigate the Cu aging in bentonite as a function of incubation time, pH and temperature in order to elucidate the aging mechanism of added Cu, based on dynamic and thermodynamic parameters and the changes of basal spacing of clay minerals. 2. Materials and methods 2.1. Bentonite Ca-bentonite (National Standard Substance GBW 070049) was supplied by Testing Center for Rock and Mineral, Jiangsu Province, China, which was sampled from Chunhua town, Jiangning District, Nanjing, Jiangsu Province, China. Main mineral is montmorillonite (91%), and minor minerals are pyrophyllite, kaolinite, feldspar, and christobalite. The proportion of grain less than 10 μm accounts for 71.8%, and that of grain between 10 and 50 μm is 27.1%. According to the methods of Lu (2000), the pH of bentonite measured at solid/water ratio of 1:2.5 was 8.3, cation exchange capacity (CEC) measured using ammonium chloride–ammonium acetate method was 1034 mmol kg− 1, specific surface area measured using glycol ether adsorption method was 651 m2 g− 1, and its main chemical composition measured using lithium metaborate fusion method was listed as follows: SiO2 (60.44%), Al2O3 (18.62%), Fe2O3 (2.48%), CaO (3.52%), MgO (4.43%). In addition, Cu content measured using HF–HNO3–HClO4 digestion method was 28.4 mg kg− 1. 2.2. Aging experiments The aging of Cu was conducted using batch experiments. For the kinetics experiment, bentonite (2.0 g) was weighed and arranged in two groups, and laid in 50 mL centrifugal tubes of known weight. After 10 mL of de-ionized water was added, one group was added with 10 mL of 0.001 mol L− 1 CuCl2 solution, the other was added with 10 mL of de-ionized water (as control). All the centrifugal tubes were capped and incubated at room temperature for 2 h, 1, 3, 10, 30, 90, 180, and

87

370 days, respectively. In this experiment, there were two duplicates in each treatment. For the experiment of pH and temperature effects, bentonite (1.0 g) was weighed and laid in 50 mL centrifugal tubes of known weight. After 10 mL of de-ionized water was added, the pH of suspensions was measured using pH-meter (PB-10, Sartorius Instrument Corporation, Germany) and adjusted with 1 mol L− 1 KOH/HCl to the appropriate values (initial pH 4.5, 5.5, 6.5, and 7.5), and the volume of added KOH/ HCl was recorded. Then all the centrifugal tubes were added with 10 mL of 0.001 mol L− 1 CuCl2 solutions, capped and incubated for 90 days at 4 °C, 25 °C, and 50 °C, respectively. There were two duplicates in each treatment. 2.3. Sequential extraction procedure After the incubation, the suspensions incubated in centrifugal tubes were centrifuged at 4000 rpm for 30 min, and then filtered. The concentrations of copper in the filtrates as water-soluble Cu were measured. At the same time, the centrifugal tubes with residues were weighed (for deduction in calculating Cu fractions). Then Cu fractions in solid-phase of bentonite were measured by a modified sequential extraction procedure of Ma and Uren (1998). The chemical reagents, extraction conditions and corresponding fractions are listed in Table 1. Between each successive extraction, the supernatant was centrifuged at 4000 rpm for 30 min, and then filtered, and the centrifugal tubes with residues were weighed. Finally, Cu concentrations in all fractions including digestion solution were measured by a flame atomic absorption spectrophotometer (F-AAS, WEX-120, Beijing Ruili Analytic Instrument Company, China). 2.4. Preparation of bentonite saturated with Cu at different pH According to the method of Ma and Uren (1998), bentonite (10 g) was saturated with copper ions by shaking for 4 h with 50 mL of 0.1 mol L− 1 CuCl2 solutions and centrifuged, and then the supernatant was decanted. The residue was washed twice for 1 h each time with 25 mL of de-ionized water. The pH of bentonite suspension was measured using a pH-meter and adjusted with 1 mol L− 1 KOH/HCl to the appropriate values (initial pH 4.5, 5.5, 6.5, and 7.5), and the suspension was allowed to stand several days to achieve pH equilibrium when it was adjusted to the values of initial pH again and again. Then, the final pH was measured, and the suspension was centrifuged and the supernatant was decanted. Finally, the residue was air-dried at room temperature, ground and passed through a 0.2 mm sieve, and stored in desiccators.

Table 1 Sequential extraction procedure and the corresponding fractionsa Fraction

Extraction procedure

20 mL of 1 mol L− 1 NH4NO3 at pH 7.0 by shaking for 1 h at room temperature 2 successive 20 mL of 1% NaHCaEDTAb in 1 mol L− 1 ammonium acetate at pH 8.3 by shaking for 2 h at room temperature 0.5 mol L− 1 HCl extractable 20 mL of 0.5 mol L− 1 HCl by shaking for 2 h at room Cu (0.5 M HCl-Cu) temperature 6 mol L− 1 HCl extractable 20 mL of 6 mol L− 1 HCl by shaking for 2 h at room Cu (6 M HCl-Cu) temperature Residual Cu (RES-Cu) Digesting with mixed solutions of 7.5 mL HCl, 2.5 mL HNO3, and 5 mL HClO4

NH4NO3 extractable Cu (NH4NO3-Cu) EDTA extractable Cu (EDTA-Cu)

a

Based on the procedure of Ma and Uren (1998). Preparation of 1% NaHCaEDTA in 1 mol L− 1 ammonium acetate at pH 8.3: Following the method of Beckwith (1955), a 2% solution of disodium EDTA was saturated with CaCO3. It was then filtered, and the filtrate mixed with an equal volume of 2 mol L− 1 ammonium acetate. The pH was adjusted to 8.3 using ammonia solution and if necessary, using acetic acid. b

88

S.-W. Zhou et al. / Geoderma 147 (2008) 86–92

2.5. Measurements of water content and basal spacing of Cu-bentonite Cu-bentonite prepared above was placed in desiccators at water activities (aw) of 0, 0.12, 0.35, 0.55, 0.65, and 1.0, respectively, at room temperature for more than 2 months. Following the method of Ma and Uren (1998), P2O5, LiCl·H2O, CaCl2·6H2O, Ca(NO3)2·4H2O, NaNO2, and H2O, respectively, were used to prepare saturated solutions for constant aw mentioned above. Finally, the water content of samples at different aw was measured by weighing. At the same time, some of the samples were dried at 105 °C for 24 h to determine humidity in order to make the corresponding correction. In addition, some of the samples were stored in desiccators for at least 2 months above a saturated solution of Mg(NO3)2·6H2O to obtain constant relative humidity (55.5%) prior to XRD analysis, as proposed by Ma and Uren (1998). Then the basal spacing was measured by X-ray diffraction meter (D/MAX-2400, Rigaku International Corporation, Japan), where Cu Kα (λ = 0.154056 nm) was selected, and scan mode (2θ) ranged from 3° to 10°, with the step of 0.01°. 2.6. Data analysis and quality control Standard errors (SE) were calculated using Microsoft Office Excel 2007 (Microsoft Corporation, USA); graphics and associated curvefitting were conducted with SigmaPlot for Windows Version 10.0 (Systat Software Inc., USA); and statistics was conducted using SAS 9.0 System for Windows (SAS Institute Inc., USA). During digesting with mixed solutions of 7.5 mL HCl, 2.5 mL HNO3, and 5 mL HClO4 to determine residual Cu, a National Standard Substance (Heilongjiang dark-brown earth, GBW07401) was used as quality control. The value determined was 21.22 mg kg− 1 (SE = 0.23, n = 20), not significantly different from that certified (21 mg kg− 1) (p b 0.001). Based on measured concentration and weight, each Cu fraction was corrected. Then, the accumulative total of exogenous Cu in bentonite was 317.3 mg kg− 1 (SE = 1.60, n = 16), not significantly different from that added (317.6 mg kg− 1) (p b 0.001). 3. Results and discussion 3.1. Changes of Cu fractions with time Endogenous copper in bentonite mainly existed as residual Cu (RES-Cu) (91.0%), and unobvious changes of Cu fractions with time

Table 2 The estimated parameter values of parabolic diffusion equation and correlation coefficient Less extractable Cu fractiona

−3

−1.59 × 10 (5.23 × 10− 3)c 2.26 × 10− 2 (9.02 × 10− 3) 1.58 × 10− 1 (2.48 × 10− 2)

RES-Cu 6 M HCl-Cu 0.5 M HCl-Cu

−6

6.43 × 10 (8.48 × 10− 7) 7.06 × 10− 6 (1.53 × 10− 6) 2.00 × 10− 5 (7.00 × 10− 6)

R2b 0.9427⁎⁎⁎ 0.8586⁎⁎⁎ 0.6899⁎⁎⁎

a 0.5 M HCl-Cu, 6 M HCl-Cu, and RES-Cu mean 0.5 mol L− 1 HCl extractable Cu, 6 mol L− 1 HCl extractable Cu and residual Cu, respectively. b ⁎⁎⁎ indicates significance at p b 0.0001. c The numbers in brackets are standard error.

occurred (data not shown). However, exogenous Cu fractions changed greatly with incubation time (Fig. 1). Because it was not almost detected, water-soluble Cu was joined with NH4NO3 extractable Cu (NH4NO3-Cu) into exchangeable copper (EXC-Cu). It was shown in Fig. 1 that EXC-Cu decreased from 18.0% to 1.1% after 370 days of incubation time, whereas EDTA extractable Cu (EDTA-Cu) increased from 63.1% to 70.3% in initial 3 days of aging, which may be ascribed to the transformation from EXC-Cu, and then it continually decreased with aging reaction, with 22.5% after 370 days of incubation. Contrarily, the potentially labile (0.5 mol L− 1 HCl extractable Cu) and inert (6 mol L− 1 HCl extractable Cu and residual Cu) fractions increased continually with incubation time. This indicated that the aging of added Cu was a slow process (more than one year) where the easily extractable Cu species on soil solid surfaces gradually transformed into less extractable forms which possibly diffused into interlayer of bentonite. Arias-Estevez et al. (2007) approved that 500 days of incubation in an acid soil were still inadequate for the aging of added Cu more than 500 mg kg− 1. It is obvious that the aging process of Cu usually lasted for several months or years, gradually reducing its bioavailability or toxicity. 3.2. Modeling the aging process of added Cu When water-soluble metals were added to soils, the initially rapid reaction was followed by a much slower reaction which could be described well with a solution to Fick's second law (Bruemmer et al., 1988; Ma and Uren, 1997; Lu et al., 2005; Ma et al., 2006b). As a result, in this work, the aging process of added Cu was modeled by the following parabolic diffusion equation: Yn ¼Mþ6 Ym

Fig. 1. Changes in Cu fractions from sequential extraction procedure as percentages of total amount of added copper (317.6 mg kg− 1) in bentonite with incubation time. EXCCu is the sum of water-soluble Cu and NH4NO3 extractable Cu; EDTA-Cu, 0.5 M HCl-Cu, 6 M HCl-Cu, and RES-Cu mean EDTA extractable Cu, 0.5 mol L− 1 HCl extractable Cu, 6 mol L− 1 HCl extractable Cu, and residual Cu, respectively.

D/r2 (d− 1)

M

rffiffiffiffiffiffiffiffiffiffi D t πr 2

ð1Þ

where Yn is less extractable Cu fractions (mg kg− 1), Ym is the concentration of added Cu (mg kg− 1), and M is a constant which is believed to represent the effect of rapid reactions on the loss of extractability of added Cu. D is the diffusion coefficient (cm2 d− 1) and r is the radius of the spherical particle (cm), and the term D/r2 expresses the apparent diffusion rate coefficient (d− 1). And t is incubation time (d) (Ma and Uren, 1997; Ma et al., 2006b). The less extractable Cu fractions were fitted well with Eq. (1) (R2 = 0.690–0.943, p b 0.0001), and the calculated apparent diffusion rate coefficient (D/r2) ranged between 6.43 × 10− 6 d− 1 and 2.0 × 10− 5 d− 1 (viz. 7.44 × 10− 11–2.31 × 10− 10 s− 1) (Table 2), approximate to those of copper and zinc in soils (Ma and Uren, 1997; Ma et al., 2006b), which indicated micropore diffusion was dominant aging mechanism. It may be deduced from Table 2 that because M was approximate to zero, residual Cu (RES-Cu) was mainly ascribed to diffusion, whereas larger proportions to 0.5 mol L− 1 HCl extractable Cu (0.5 M HCl-Cu) was derived from surface nucleation/precipitation in view of higher value of M. Further, the contribution of diffusion to each less extractable Cu

S.-W. Zhou et al. / Geoderma 147 (2008) 86–92

89

Fig. 2. The effects of pH (a) and temperature (b) on apparent diffusion rate coefficient (D/r2) after 90 days of incubation through multiple comparisons. According to Eq. (1), D/r2 was obtained from the inert fractions (6 mol L− 1 HCl extractable Cu and residual Cu). Here different capital letters indicate significant difference at p b 0.05 level.

fraction could be calculated and assessed according to Eq. (1). For example, after 90 days of incubation 47.7% of 0.5 M HCl-Cu and 79.1% of 6 mol L− 1 HCl extractable Cu (6 M HCl-Cu) were owed to diffusion reaction; whereas, after 370 days of incubation the proportions increased to 64.9% and 88.5%, respectively. Therefore, during longterm (more than 90 days) aging of Cu added to bentonite, the inert fractions (viz. RES-Cu and 6 M HCl-Cu) might be considered to almost completely derive from micropore diffusion.

specific adsorption of copper ions on mineral surface increased with increasing pH due to the increase in EDTA-Cu and 0.5 M HCl-Cu at lower extent. Under high amount of Cu addition this adsorption would be probably strengthened, as mentioned above. Based on these results, it is obvious that during 90 days of incubation the surface nucleation/precipitation was promoted whereas the diffusion was weakened accordingly with increasing pH. Sari et al. (2007a,b) also found that at higher pH value than 5, metal precipitation appeared. It has been evidenced that D/r2 obeys the two equations as follows:

3.3. Effects of temperature and pH

Ea

D=r 2 ¼ A  e−RT As mentioned above, after 90 days of incubation the inert fractions were almost completely regarded as the result of diffusion reaction. Then, D/r2 may be calculated according to Eq. (1), by which the effects of temperature and pH could be assessed quantitatively. Factor analysis of variance showed that there were significant effects of temperature, pH, and their interaction on D/r2 (p b 0.01), and the effect of temperature was more remarkable than that of pH (data not shown). But, multiple comparisons suggested that these differences were insignificant until lower pH (4.37) or higher temperature (50 °C) (Fig. 2). The rise in temperature would increase D/r2 (Fig. 2), which agreed with results in literatures (Barrow, 1986, 1992; Bruemmer et al., 1988; Ma and Uren, 1997; Ma et al., 2006b). However, during 90 days of incubation the significant change of D/r2 did not occur at low temperature (b25 °C), indicating that higher energy barrier exists in the diffusion of copper ions into the interlayer of clay mineral. Probably, under higher amount of Cu addition (635.5 mg kg− 1) copper ions were more associated to mineral surface as specific and nonspecific adsorption species. Contrarily, D/r2 decreased with increasing pH. However, the effect of pH on D/r2 was significant only between pH 4.37 and pH 5.44 (Fig. 2). Ma et al. (2006b) used an isotopic dilution technique to study the change of isotopically exchangeable Cu (E value) in soils with time when water-soluble Cu was added to different soils and found that D/r2 increased with the rise of soil pH. In this experiment here, D/r2 was obtained from the inert fractions (viz. 6 M HCl-Cu and RES-Cu). If the potentially labile fraction (0.5 M HClCu) was involved in, that is, EXC-Cu and EDTA-Cu was regarded as labile fractions, there would be no significant effect of pH on calculated D/r2; further, if only EXC-Cu was regarded as labile fraction, the calculated D/r2 exhibited an opposite trend, where it increased with increasing pH, with significant change at pH N 5.44 (data not shown). The D/r2 reported by Ma et al. (2006b) was derived from isotopically exchangeable Cu (E value). Thus, it was considered that the effects of pH on D/r2 depended on which fractions or species of Cu in soils or minerals were used. Moreover, it also could be deduced that

ð2Þ

ΔG

ΔH

ΔS

D=r 2 ¼ ðkb T=hÞ  e− RT ¼ ðkb T=hÞ  e− RT  e R

ð3Þ −1

2

where D/r is the apparent diffusion rate coefficient (s ), A is the frequency or pre-exponential factor, representing the unchanged parts of diffusion coefficients with temperature (s− 1); Ea is the activation energy of diffusion (kJ mol− 1); kb is the Boltzmann constant (1.38 × 10− 23 J K− 1); h is Planck's constant (6.626×10− 34 J s); ΔG is the standard Gibbs free energy of activation (kJ mol− 1); ΔH is the standard enthalpy of activation (kJ mol− 1); ΔS is the standard entropy of activation (J mol− 1 K− 1); R is the gas constant (8.31 J mol− 1 K− 1); and T is the absolute temperature in Kelvin (K) (Bruemmer et al., 1988; Scheckel  2  and Sparks, 2001). By plotting ln(D/r2) vs. 1/T and ln D=r vs. 1/T, linear relationships T are obtained and one can determine thermodynamic parameters such as Ea, ΔG, ΔH and ΔS (Table 3). Furthermore, Scheckel and Sparks (2001) considered that these parameters obtained may be assessed according to the following relationship:   Ea ¼ ΔH þ RT T ¼ 25 C :

ð4Þ

o

−1

The y-intercept (actual RT = 2.48 kJ mol ) was excellent in agreement to our experimental data (2.51 kJ mol− 1).

Table 3 Summary of reaction parameters derived from Eqs. (2) and (3) for Cu added to bentonite during 90 days of incubation at different pH pH

4.37 5.44 6.54 7.60 a

Ea

A

ΔH

ΔS

ΔG at 25 °C

kJ mol− 1

s− 1

kJ mol− 1

J mol− 1 K− 1

kJ mol− 1

56.1 (0.9)a 46.1 (10.1) 39.0 (3.4) 38.3 (12.2)

2.09 × 10− 1 (7.06 × 10− 2) 1.94 × 10− 3 (7.43 × 10− 3) 1.04 × 10− 4 (1.37 × 10− 4) 5.90 × 10− 5 (2.73 × 10− 4)

53.5 (0.8) 43.5 (10.1) 36.5 (3.4) 35. 8 (12.1)

−266.4 (2.7) −305.4 (31.7) −329.6 (10.7) − 334.4 (38.6)

133.0 (0.2) 135.3 (0.3) 135.0 (0.2) 136.4 (0.1)

The numbers in brackets are standard error.

90

S.-W. Zhou et al. / Geoderma 147 (2008) 86–92

The results in Table 3 showed that when pH increased from 4.37 to 7.60, Ea apparently reduced from 56.1 to 38.3 kJ mol− 1, but significantly higher than free energies of adsorption for metal ions on minerals (about 10 kJ mol− 1), where Sari et al. (2007a,b) found that the adsorption was feasible, spontaneous and exothermic at 20–50 °C, and suggested it may be carried out via chemical ion-exchange mechanism. Apparently, the diffusion of copper ions into interlayer of bentonite was much more difficult than the adsorption on bentonite surfaces. Based on results in Fig. 2 and Table 3, it might be deduced that at lower pH (4.37), free Cu2+ ions enough to diffuse because of weaker adsorption on surface of bentonite, so D/r2 was higher. But Cu2+ possesses bigger hydrated radius, which caused higher energy barrier of diffusion (viz. Ea) and consequent restraint of diffusion reaction. Contrarily, at higher pH (N5.44), less free copper ions enough to diffuse due to more strongly specific adsorption on surface, which would result in lower D/r2. However, CuOH+ at high pH with smaller hydrated radius became dominant diffusive ions. Because CuOH+ ions diffused more easily than Cu2+ into the interlayer of clay mineral, the energy barrier was lower. In conclusion, at lower pH there were higher values of D/r2 and Ea; whereas at higher pH there were lower values of D/r2 and Ea. Ultimately, the reciprocal effects of D/r2 and Ea would result in the standard Gibbs free energy of activation (ΔG) constant at different pH (Table 3). Sari et al. (2007a,b) also found that at higher pH values than 5, metal precipitation appeared and adsorbent was deteriorated with accumulation of metal ions, which greatly approximates to the value obtained here, that is, pH 5.44. The effect of increasing temperature on slow reaction of metals added soil minerals was suggested to be equivalent to that of prolonging incubation time, and equivalent time (teq) for one temperature such as 25 °C could be calculated according to the following equation:

Fig. 4. XRD patterns and the basal spacing of the bentonite saturated with copper at different pH at a relative humidity of 55.5%.

3.4. Changes of basal spacing of bentonite saturated with Cu at different pH

where teq is equivalent time (d) at 25 °C, t and T are actual reaction time (d) and temperature (K), respectively; R is the gas constant (8.31 J mol− 1 K− 1); and Ea is activation energy of diffusion (kJ mol− 1) (Barrow, 1986; Ma and Uren, 1997; Scheckel and Sparks, 2001; Ma et al., 2006b). pffiffiffiffiffiffi By plotting Yn/Ym vs. teq , linear relationships should be obtained, as shown in Fig. 3, where it was obvious that the effects of temperature became weak with increasing pH. Because the diffusion of copper ions was an endothermal change, existing higher energy barrier (Ea) needs to be overcome at lower pH such as pH 4.37 (Table 3), which caused greater effects of temperature.

The basal spacing (001) of bentonite saturated with copper chloride at different pH was measured by XRD at a relative humidity of 55.5% (Fig. 4). The results showed that the basal spacing of Cubentonite decreased obviously with increasing pH, from 1.51 nm at pH 4.42 and 5.51 to 1.37 nm at pH 6.40, and to 1.26 nm at pH 7.68. So, it may be explained that copper ions adsorbed on surfaces could diffuse into the interlayer of clay minerals, and the dominant diffusive ions were different species with different hydrated radius. The water contents of bentonite saturated with Cu at different water activity (aw) also decreased with increasing pH. For example, at aw = 0.55 it decreased from 12.2% at pH 4.47 to 7.2% at pH 7.62 (Fig. 5). This supported further that diffusive ions possessed different hydrated shells and hydrated radius. Based on this study, it may be assured firmly that at lower pH (b5.5) diffusive ions were mainly Cu2+, with more hydrated shells and hydrated radius, whereas at higher pH (N6.5) CuOH+ became dominant diffusive ions, with less hydrated shells and hydrated radius. He et al. (2001) showed that the hydrated Cu2+ ion in the interlayer of montmorillonite loses the coordinating water on

Fig. 3. The proportion of inert species (Yn/Ym in fraction) to added Cu in bentonite at different pH as a function of square root of equivalent time (d) adjusted to 25 °C. Inert species is the sum of 6 mol L− 1 HCl extractable Cu and residual Cu.

Fig. 5. The water content of Cu-bentonite at different pH as a function of water activity (aw).

teq ¼ t  expðEa =298R−Ea =RT Þ

ð5Þ

S.-W. Zhou et al. / Geoderma 147 (2008) 86–92

heating and then enters into the hexagonal cavities. When the heating temperature further increase, dehydroxylation occurs, which facilitates Cu2+ ion in the hexagonal cavities to penetrate into the octahedral vacancies. In another experiment, Karmous et al. (2006) found that Cu cation is located in the middle of the interlayer space of montmorillonite, with one water sheet, where the basal spacing is 1.24 nm. After thermal treatment (until 250 °C) the basal spacing decreases up to 1.0 nm, which is attributed to the loss of interlayer water. When increasing temperature up to 350 °C, a proportion of Cu cations diffuses into the octahedral vacancies and this proportion increases when increasing the heating duration. Though it is still unclear how copper ions diffuse and penetrate into clay minerals under normal soil temperature, we suppose it is probable to enter the hexagonal cavities by dehydration after longterm incubation. In our experiment, the basal spacing measured at pH 7.68 was 1.26 nm (Fig. 4), very close to the reported (1.24 nm) (Karmous et al., 2006), in which copper ions existed in the interlayer of clay mineral possessed one water sheet. Thus, it might be deduced that CuOH+ with one water sheet mainly existed in the interlayer of bentonite at pH 7.68, whereas Cu2+ with two water sheets became dominant ions in the interlayer of bentonite at pH less than 5.5. And that, with increasing incubation time these copper ions could diffuse into hexagonal cavities of bentonite by dehydration. Ma and Uren (1998) had similar results in bentonite saturated with zinc, where the basal spacing of ZnOH+–bentonite at high pH (6.9 and 8.8) ranged from 1.21 to 1.26 nm, and that of Zn2+–bentonite at low pH (5.6 and 6.3) was 1.51 nm. Furthermore, according to these results, they developed a structural model which explained ZnOH+ ions diffused at higher pH into hexagonal cavities and were consequently entrapped and fixed by clay minerals. Because of the completely approximate ionic radius between Cu2+ (r = 0.073 nm) and Zn2+ (r = 0.074 nm) (Dean, 1999), the chemical characteristics of the two metal ions were probably alike. As a result, it may be considered that there was the same behavior of diffusion for Cu added to bentonite as that for Zn found by Ma and Uren (1998). 4. Conclusions The aging of copper added to bentonite as a function of incubation time, pH and temperature was examined to determine dynamic and thermodynamic parameters. It was concluded that the aging of added Cu was a slow process (more than one year) where the easily extractable Cu species on soil solid surfaces gradually transformed into less extractable Cu forms which possibly diffused into interlayer of bentonite. The aging process can be well fitted using a parabolic diffusion equation (R2 = 0.690–0.943, p b 0.0001) with apparent diffusion rate coefficients (D/r2) of 6.43–20.0 × 10− 6 d− 1 and activation energy of diffusion (Ea) of 38.3–56.1 kJ mol− 1, indicating that micropore diffusion could be the main mechanism of aging processes of Cu added to bentonite. At lower pH free Cu2+ ions were enough to diffuse because of weaker adsorption on surface of bentonite, so D/r2 was higher. But Cu2+ possesses bigger hydrated radius, which caused higher energy barrier of diffusion (Ea), and accordingly greater effect of temperature. Contrarily, at higher pH there were less free copper ions to diffuse due to more strongly specific adsorption on surface, which would result in lower D/r2. However, CuOH+ with smaller hydrated radius became dominant diffusive ions, which caused lower Ea, and accordingly smaller effect of temperature. In conclusion, the reciprocal effects of D/r2 and Ea would result in the constant standard Gibbs free energy of activation (ΔG) at different pH. In a word, sequential extraction and XRD approved that Cu2+ was main diffusive ions at lower pH (b5.5), whereas CuOH+ became dominant diffusive ions at higher pH. Moreover, after long-term incubation these copper ions might diffuse into hexagonal cavities of bentonite by dehydration.

91

Acknowledgements The authors wish to thank Prof. Zhen-Yu Zhang for XRD analysis, Dr. Xiao-Li Bi for data analysis and Dr. Hua Zhang for the constructive comments. This study was funded by the National Natural Science Foundation of China (40571071), Major State Basic Research and Development Program of China (2002CB410809), and State Key Laboratory of Soil and Sustainable Agriculture, China (0551000014).

References Ahnstrom, Z.S., Parker, D.R., 1999. Development and assessment of a sequential extraction procedure for the fractionation of soil cadmium. Soil Sci. Soc. Am. J. 63, 1650–1658. Alexander, M., 2000. Aging, bioavailability, and overestimation of risk form environmental pollutants. Environ. Sci. Technol. 34, 4259–4265. Al-Qunaibt, M.H., Mekhemer, W.K., Zaghloul, A.A., 2005. The adsorption of Cu (II) ions on bentonite—a kinetic study. J. Colloid Interface Sci. 283, 316–321. Alvarez-Puebla, R.A., dos Santos Jr., D.S., Blanco, C., Echevreria, J.C., Garrido, J.J., 2005. Particle and surface characterization of a natural illite and study of its copper retention. J. Colloid Interface Sci. 285, 41–49. Arias-Estevez, M., Novoa-Munoz, J.C., Pateiro, M., Lopez-Periago, E., 2007. Influence of aging on copper fractionation in an acid soil. Soil Sci. 172, 225–232. Axe, L., Trivedi, P., 2002. Intraparticle surface diffusion of metal contaminants and their attenuation in microporous amorphous Al, Fe, and Mn oxides. J. Colloid Interface Sci. 247, 259–265. Barrow, N.J., 1986. Testing a mechanistic model. II. The effects of time and temperature on the reaction of zinc with a soil. J. Soil Sci. 37, 277–286. Barrow, N.J., 1992. A brief discussion on the effect of temperature on the reaction of inorganic ions with soil. J. Soil Sci. 43, 37–45. Beckwith, R.S., 1955. Studies on soil manganese. I. The use of disodium calcium versenate for the extraction of divalent manganese from soils. Aust. J. Agr. Res. 6, 299–307. Bourg, I.C., Bourg, A.C.M., Sposito, G., 2003. Modeling diffusion and adsorption in compacted bentonite: a critical review. J. Contam. Hydrol. 61, 293–302. Brady, W.D., Eick, M.J., Grossl, P.R., Brady, P.V., 2003. A site-specific approach for the evaluation of natural attenuation at metals-impacted sites. Soil Sediment Contam. 12, 541–564. Bruemmer, G.W., Gerth, J., Tiller, K.G., 1988. Reaction kinetics of the adsorption and desorption of nickel, zinc and cadmium by goethite. Ι. Adsorption and diffusion of metals. J. Soil Sci. 39, 37–51. Bruus Pedersen, M., Van Gestel, C.A.M., 2001. Toxicity of copper to the collembolan Folsomia fimetaria in relation to the age of soil contamination. Ecotoxicol. Environ. Safety 49, 54–59. Davidson, C.M., Thomas, R.P., McVey, S.E., Perala, R., Littlejohn, D., Ure, A.M., 1994. Evaluation of a sequential extraction procedure for the speciation of heavy metals in sediments. Anal. Chim. Acta 291, 277–286. Davidson, C.M., Urquhart, G.J., Ajmone-Marsan, F., Biasioli, M., da Costa Duarte, A., DíazBarrientos, E., Grčman, H., Hossack, I., Hursthouse, A.S., Madrid, L., Rodrigues, S., Zupan, M., 2006. Fractionation of potentially toxic elements in urban soils from five European cities by means of a harmonised sequential extraction procedure. Anal. Chim. Acta 565, 63–72. Dean, J.A., 1999. Lange's Handbook of Chemistry (Fifteenth Edition). McGraw-Hill, Inc., New York. Du, Q., Sun, Z., Forsling, W., Tang, H., 1997. Adsorption of copper at aqueous illite surfaces. J. Colloid Interface Sci. 187, 232–242. Elzinga, E.J., Reeder, R.J., 2002. X-ray absorption spectroscopy study of Cu2+ and Zn2+ adsorption complexes at the calcite surface: implications for site-specific metal incorporation preferences during calcite crystal growth. Geochim. Cosmochim. Acta 66, 3943–3954. He, H.P., Guo, J.G., Xie, X.D., Peng, J.L., 2001. Location and migration of cations in Cu2+adsorbed montmorillonite. Environ. Int. 26, 347–352. Hyun, S.P., Cho, Y.H., Hahn, P.S., 2005. An electron paramagnetic resonance study of Cu (II) sorbed on kaolinite. Appl. Clay Sci. 30, 69–78. Inaba, S., Takenaka, C., 2005. Changes in chemical species of copper added to brown forest soil in Japan. Water Air Soil Pollut. 162, 285–293. Jalali, M., Khanlari, Z.V., 2008. Effect of aging process on the fractionation of heavy metals in some calcareous soils of Iran. Geoderma 143, 26–40. Kaasalainen, M., Yli-Halla, M., 2003. Use of sequential extraction to assess metal partitioning in soils. Environ. Pollut. 126, 225–233. Karmous, M.S., Rhaiem, H.B., Naamen, S., Oueslati, W., Amara, A.B.H., 2006. The interlayer structure and thermal behavior of Cu and Ni montmorillonites. Z. Kristallogr. Suppl. 23, 431–436. Karthikeyan, K.G., Elliott, H.A., Chorover, J., 1999. Role of surface precipitation in copper sorption by the hydrous oxides of iron and aluminum. J. Colloid Interface Sci. 209, 72–78. Ling, W.T., Shen, Q., Gao, Y.Z., Gu, X.H., Yang, Z.P., 2007. Use of bentonite to control the release of copper from contaminated soils. Aust. J. Soil Res. 45, 618–623. Lock, K., Janssen, C.R., 2003. Influence of ageing on copper bioavailability in soils. Environ. Toxicol. Chem. 22, 1162–1166. Lu, A., Zhang, S., Shan, X.Q., 2005. Time effect on the fractionation of heavy metals in soils. Geoderma 125, 225–234.

92

S.-W. Zhou et al. / Geoderma 147 (2008) 86–92

Lu, R.K., 2000. Analysis Methods in Soil Agrochemistry (In Chinese). Chinese Agricultural Science and Technology Press, Beijing. Ma, Y.B., Uren, N.C., 1997. The effects of temperature, time and cycles of drying and rewetting on the extractability of zinc added to a calcareous soil. Geoderma 75, 89–97. Ma, Y.B., Uren, N.C., 1998. Dehydration, diffusion and entrapment of zinc in bentonite. Clays Clay Miner. 46, 132–138. Ma, Y.B., Lombi, E., Oliver, I.W., Nolan, A.L., McLaughlin, M.J., 2006a. Long-term aging of copper added to soils. Environ. Sci. Technol. 40, 6310–6317. Ma, Y.B., Lombi, E., Nolan, A.L., McLaughlin, M.J., 2006b. Short-term natural attenuation of copper in soils: effects of time, temperature and soil characteristics. Environ. Toxicol. Chem. 25, 652–658. Martínez, C.E., McBride, M.B., 1998. Solubility of Cd2+, Cu2+, Pb2+, and Zn2+ in aged coprecipitates with amorphous iron hydroxides. Environ. Sci. Technol. 32, 743–748. Martínez, C.E., McBride, M.B., 2000. Aging of coprecipitated Cu in alumina: changes in structural location, chemical form, and solubility. Geochim. Cosmochim. Acta 64, 1729–1736. McBride, M.B., 1991. Processes of heavy and transition metal sorption by soil minerals. In: Bolt, G.H., De Boodt, M.F., Hayes, M.H.B., McBride, M.B. (Eds.), Interactions at the Soil Colloid–Soil Solution Interface. Kluwer Academic Publishers, Dordrecht, pp. 149–175. McLaughlin, M.J., 2001. Ageing of metals in soils changes bioavailability. Fact Sheet Environ. Risk Assess. 4, 1–6. Pueyo, M., Mateu, J., Rigol, A., Vidal, M., López-Sánchez, J.F., Rauret, G., 2008. Use of the modified BCR three-step sequential extraction procedure for the study of trace element dynamics in contaminated soils. Environ. Pollut. 152, 330–341. Renella, G., Chaudri, A.M., Brookes, P.C., 2002. Fresh additions of heavy metals do not model long-term effects on microbial biomass and activity. Soil Biol. Biochem. 34, 121–124.

Sari, A., Tuzen, M., Citak, D., Soylak, M., 2007a. Adsorption characteristics of Cu (II) and Pb (II) onto expanded perlite from aqueous solution. J. Hazard. Mater. 148, 387–394. Sari, A., Tuzen, M., Citak, D., Soylak, M., 2007b. Equilibrium, kinetic and thermodynamic studies of adsorption of Pb (II) from aqueous solution onto Turkish kaolinite clay. J. Hazard. Mater. 149, 283–291. Scheckel, K.G., Sparks, D.L., 2001. Temperature effects on nickel sorption kinetics at the mineral–water interface. Soil Sci. Soc. Am. J. 65, 719–728. Schosseler, P.M., Wehrli, B., Schweiger, A., 1999. Uptake of Cu2+ by the calcium carbonates vaterite and calcite as studied by continuous wave (CW) and pulse electron paramagnetic resonance. Geochim. Cosmochim. Acta 63, 1955–1967. Scott-Fordsmand, J.J., Weeks, J.M., Hopkin, S.P., 2000. Importance of contamination history for understanding toxicity of copper to earthworm Eisenia fetica (Oligochaeta: Annelida), using neutral-red retention assay. Environ. Toxicol. Chem. 19, 1774–1780. Shuman, L.M., 1985. Fractionation method for soil microelements. Soil Sci. 140, 11–22. Sparks, D.L., 2003. Environmental Soil Chemistry (Second Edition). Academic Press, New York. Tessier, A., Campbell, P.G.C., Bisson, M., 1979. Sequential extraction procedure for the speciation of particulate trace metals. Anal. Chem. 51, 844–851. Tom-Petersen, A., Hansen, H.C.B., Nybroe, O., 2004. Time and moisture effects on total and bioavailable copper in soil water extracts. J. Environ. Qual. 33, 505–512. Tüzen, M., 2003. Determination of trace metals in the River Yesilirmak sediments in Tokat, Turkey using sequential extraction procedure. Microchem. J. 74, 105–110.