Various forms of heavy metals, when added to soils, slowly redistribute and repartition among their solid-phase components (Sposito et al., 1983; Chang et al., 1984; Shuman, 1985; McGrath and Cegarra, 1992; McLaren and Ritchie, 1993; Soon, 1994; Banin, et al., 1995; Han et al., 1995; Han and Banin, 1997 and 1999, Han et al., 2000). The soil moisture regime controls both soil redox potential and biological activity in soils. Generally under saturated conditions, because of the lower redox potential in a soil, Fe and Mn oxides are reduced under anaerobic conditions and, as a consequence, heavy metals are released and redistributed among the solid-phase components (Silviera and Sommers, 1977; Schwab and Lindsay, 1983; Mandal and Mandal, 1986; Hazra et al., 1987; Patrick and Jugsujinda, 1992; Han and Banin, 1996, 1997, and 2000). Observations under field conditions show that heavy metals added to soils are transferred with time from the more labile fractions (such as soluble and exchangeable fraction) into less labile fractions (McGrath and Cegarra, 1992; McLaren and Ritchie, 1993; Soon, 1994). In salt- and sludge-amended soils, heavy metals in the carbonate, hydrous oxide-associated, and the residual fraction tend to increase with time, whereas metals in the exchangeable and organic matter fraction decrease (Sposito et al., 1983; McGrath and Cegarra, 1992; McLaren and Ritchie, 1993; Soon, 1994).
The pathways and kinetics of transformation of Cd, Cu, Cr, Pb, Ni, and Zn added to arid-zone soils incubated in both saturated paste and field capacity moisture regimes have been studied in detail (Han and Banin, 1997 and 1999). In both moisture regimes, transformations of heavy metals added as soluble salts were characterized by an initial fast retention process (such as adsorption), followed by a slower, long-term redistribution process. In both processes, the heavy metals were transformed with time from more labile fractions into more stable fractions. However, during the subsequent slow, long-term process, the moisture regime strongly affected the transformation pathways (Han and Banin, 1997 and 1999).
Irrigation with reclaimed sewage water is the most readily available and economically feasible way to supplement fresh water in arid-zone areas (Feigin et al., 1991; Avnimelech, 1993). However, prolonged irrigation with treated sewage effluents resulted in the accumulation of heavy metals (Cd, Cu, Ni, and Pb) in the top 0 to 15-cm layer of arid-zone soils, leading to increased uptake of certain metals by crops and their possible introduction to animals and humans through the food chain (Banin et al., 1981). The wetting-drying moisture regime is one of the most important factors in controlling the physical, chemical, and biological properties of irrigated soils. The redistribution of heavy metals added at low, realistic levels to arid-zone soil under a moisture regime involving wetting and drying cycles, which simulates irrigated soils, is not well understood. This study investigated solid-phase partitioning of added Cr, Cu, Ni, and Zn in arid-zone soils during a 1-year period of incubation, under a wetting (field capacity) and drying (hygroscopic moisture) cycle regime.
MATERIALS AND METHODS
Bulk samples (0-20 cm) of two soils were collected. The soils were air-dried, ground, and sieved through a 2-mm sieve. Properties of the two soils have been reported earlier by Han and Banin (1996, and 1997) and are briefly presented in Table 1. Soil organic matter was determined by the "loss-on-ignition" method (Ben-Dor and Banin, 1989) and specific surface area (SSA, in m2 g−1) by EGME (ethylene glycol monoethyl ether) adsorption (Ratner-Zohar et al., 1983). Carbonate was measured by the gasometric method (Nelson, 1982), and included calcite, siderite, and rhodochrosite. X-ray powder diffraction (Philips PW 1720 X-ray diffractometer) showed that calcite was a dominant carbonate in the soils (Han and Banin, 1995) and montmorillonite was a major clay mineral (data not shown).
Samples of 1 kg of soil (dry weight) were thoroughly mixed in plastic pots with a combination of metal nitrate salts in powder form (Table 2). A mixture of metal salts was ground and mixed well with a small portion of soil, and this metal/soil mixture was then mixed well with a large amount of soil. In both soils, metals were added in amounts reflecting the total (T) native contents of metals in the soil. Nitrate salts of Cr, Cu, Ni, and Zn were added at rates designed as 0T, 0.5T, 3T, and 5T (Table 2). After mixing with the salts, the soils with the 0.5T and 3T treatments were stored in a controlled temperature room at 25 ± 1 °C under air-dry conditions for 6 months, and treatments with 0.5T and 5T were maintained under the same conditions for 12 months. One of original purposes was to investigate the time effects on metal redistribution in soils at hygroscopic moisture. The results indicated that there was very limited redistribution of added metals in soils at this low moisture regime and fewer time effects (data not shown). The soil moisture at the air-dry condition ranged from 1.49% to 3.41% for these two soils. After storage, the soils were incubated in a moisture regime entailing periodic wetting-drying cycles. Each cycle lasted for 2 weeks and included 2 days of wetting (field capacity) and a subsequent 12 days of air-drying. Each treatment was replicated in two pots.
Subsamples were taken from each pot for sequential fractionation analyses after the 1st, 2nd, 4th, 8th, 16th, and 24th cycles. In each cycle, soil was sampled both after 2 days of wetting and after 12 days of drying. A control soil was also analyzed during the extraction sequence throughout the experiment to check the repeatability of sequential techniques.
In each cycle, pots were covered to prevent evaporation during the first 2 days of incubation at the field capacity regime,. During the subsequent 12 days, the cover was removed and the soil was allowed to air-dry at 25 °C. Before sampling for fractionation analysis, the soil was mixed thoroughly to ensure homogeneity and representability of samples and subsampled for soil moisture content. Analyses showed that moisture content of both soils remained fairly constant throughout the experiment. Soil moisture contents for each soil were similar for subsequent sampling dates during the course of the experiment. During 1 year of incubation, the average moisture for the sandy soil was 14.34 ± 2.9% and 1.49 ± 0.4% after 2 days of wetting and 12 days of drying, respectively. For the loessial soil, the average moisture was 16.44 ± 1.2% and 3.41 ± 0.9% after 2 days of wetting and 12 days of drying, respectively. It should be pointed out that the moisture contents in both soils after 2 days of wetting were comparatively lower than the normal field capacity regime.
Selective Sequential Dissolution Procedures
Heavy metals in soils were assumed to be present in six operationally defined solid-phase fractions, which were obtained by selective sequential dissolution. This method is based both on the solubility of individual solid-phase components and the selectivity and specificity of chemical reagents. The procedure provides a gradient for the physicochemical association between trace elements and solid particles rather than actual speciation (Martin et al., 1987), thus providing semiquantitative indication of their relative availability to plants or to further migration to underground water. The terms of all fractions are more likely to be operationally than chemically defined (Kheboian and Bauer, 1987). However, each extractant in the sequential selective procedures effectively targets one major solid-phase component. It is recognized that in no case can an extractant remove all of a targeted solid-phase component with no attack on other components. No selective dissolution scheme can be considered completely accurate in distinguishing between different forms of an element, i.e., various organic-inorganic solid-phase components. Readsorption of metals (Cd, Cu, Pb, and Zn) during sequential extraction were reported to be minimal (Belzile et al., 1989; Kim and Fergusson, 1991). However, some reports on re-adsorption during sequential extraction may be possibly due to the use of either large spikes or simple model materials (Belzile et al., 1989). Despite these shortcomings, common to any chemical extraction procedure, sequential dissolution techniques still furnish useful information on metal binding, mobility, and availability.
The procedures employed in this study were based on the procedures developed by Tessier et al. (1979), Shuman (1985), and Han and Banin (1997). Compared with the Tessier and Shuman procedures, the exchangeable, carbonate and residual fractions, as well as the total measurements, were modified. The modified selective sequential dissolution procedure for arid-zone soils removed the carbonate-bound fraction more efficiently and introduced microwave digestion techniques in the residual fraction and total measurement (Han and Banin, 1995, 1997, and 1999). The modified sequential procedure is summarized as follows:
1. Exchangeable Metals: This fraction includes soluble plus exchangeable metals. Twenty-five milliliters of 1 M NH4NO3 solution (pH adjusted to 7.0 with NH4OH) was added to 1.2 g of wet soil (equivalent to 1 g of dry soil) in a 50-mL Teflon centrifuge tube, and the mixture was shaken for 30 min at 25 °C. The mixture was then centrifuged at 11,950 g for 10 min. The supernatant was then decanted and filtered through a 0.45-μm filter and the soil residue kept for the next step. This reagent avoids a strong complexation of chloride and strong exchangeability of Mg in the original procedures.
2. Carbonate-Bound Metals: This may extract metal bound to various carbonate, including calcite, siderite and rhodochrosite. Twenty-five milliliters of 1 M NaOAc-HOAc solution at pH 5.0 was added to the soil residue from the previous step, and the mixture was shaken for 6 h. Excess CO2 was released by opening the tube cap during the first 2 h (Tessier et al., 1979; Han and Banin, 1995). The detailed techniques for various arid-zone soils were reported earlier by Han and Banin (1995).
3. Easily Reducible Oxide-Bound Metals: Metals in this fraction are commonly referred to as Mn oxide bound (Chao, 1972; Shuman, 1982 and 1985; Tipping et al., 1985; Han and Banin, 1995). Twenty-five milliliters of 0.04 M NH2OH·HCl in 25% HOAc solution was added to the soil residue and shaken for 30 min (Han and Banin, 1996 and 1997).
4. Organic Matter-Bound Metals: Three milliliters of a 0.01 M HNO3 and 5 mL of 30% H2O2 were added to the soil residue. The mixture was digested on a water-bath at 80 °C for 2 h. An additional 2 mL of H2O2 was added, and the mixture was heated for 1 h. Fifteen milliliters of a 1 M NH4NO3 solution was then added, and the sample was shaken for 10 min (Tessier et al., 1979; Han and Banin, 1996). It was reported that attack of major silicate phases by hydrogen peroxide was minimal (Tessier al., 1979). These arid-zone soils contained no sulfide, and the Mn oxide-bound fraction in our sequential procedure was extracted before this step; therefore, metals released in this fraction were associated primarily with organic matter. The subsequent extraction of residual soil by a neutral salt (NH4NO3) may largely recover the metal re-adsorbed during previous process.
5. Reducible Oxide-Bound Metals: Reagents in this step primarily extract metals bound to reducible oxides, such as amorphous and crystalline iron oxides. Twenty-five milliliters of 0.04 M NH2OH·HCl in 25% HOAc solution was added to the soil residue and the sample digested in a water bath at 90 °C for 3 h (Gupta and Chen, 1975; Banin et al., 1990). For iron-rich soils or mine tailings, a high concentration of the reducing agent and longer reaction period may be required (Ribet et al., 1995).
6. Residual Metals and Total Metals: We introduced a microwave digestion technique to the sequential dissolution procedures (Han and Banin, 1995). Twenty-five milliliters of 4 M HNO3 was added to the soil residue (Sposito et al., 1982), and the sample (about 1.0 g soil) for the total metals was transferred to a Teflon microwave digestion vessel. Digestion was conducted at 90% of power (850 ± 50 watts) and 100 psi of pressure for 30 min. Control of energy input is by the internal pressure. The temperature of the mixture reaches 160 to 180 °C (Han and Banin, 1995, 1996).
Multielemental analyses in soil extracts were conducted by ICP-AES (Spectro, Germany) (Han and Banin, 1997). The detection limits for Cu, Cr, Ni, and Zn on our ICP-AES were 6, 7, 15, and 2 μg/L−1, respectively. All readings were background-corrected. Calibration was done before measurement of samples using a series of standard solutions containing mixtures of the various elements. Quality control was checked every 15 samples.
Comparisons of metal concentrations in individual fractions after the first and after the last cycle of incubation were made with t tests. A significance level of 0.05, chosen a priori, was used in all analyses.
Mass Balance of the Sequential Analyses
In general, our sequential procedure provided a reasonable degree of recovery and reproducibility (Table 2). Han and Banin (1997 and 1999) also reported the similar metal recovery of the sequential procedure from two previous studies. The total recovery (sum of fractions) of the metals Cr, Cu, Zn, and Ni extracted by the sequential procedure during 1 year of incubation was relatively constant (CV of the sum: <10-15% and mostly <10%, Table 2). This indicated that the sums of the metals were relatively constant over 1 year of incubation. Thus, both percentages of metal in the fractions and absolute concentrations were used to compare the redistribution and transformation among treatments, between the two soils and among different moisture regimes. The sum of the metals by sequential analysis closely approximates the total content of the metals (Table 2). The relative errors of Cu, Ni, and Zn between the sum and the total content are mostly < 10%, and Cr had the larger relative errors (10-20%, mostly) because the total concentration of Cr was relatively low.
RESULTS AND DISCUSSIONS
Pathways of Long-Term Redistribution of Metals
The distributions of metals in both sandy and loessial soils before incubation are presented in Fig. 1. In the unincubated soils, Cr is mostly in the residual fraction and Cu, Ni and Zn in the residual and reducible oxide-bound fractions. However, in the sandy soil, Cr and Ni in the organic matter-bound and Zn in the carbonate-bound fractions are also important solid-phase components.
After the initial redistribution (during storage and, mainly, the first 2 days of wetting stage), heavy metals added were primarily present in the carbonate fraction for Cu, Ni, and Zn and the organic matter-bound fraction for Cr. The long-term redistribution and transformations with time of added heavy metals during the subsequent wetting-drying regime are exemplified (using 3T treatment) in Fig. 2. Only small changes were found for most of the metals in the control soils during incubation in the wetting-drying cycle regime (data not shown). Redistribution of added heavy metals continued with time, slowly, under the wetting-drying cycle regime from the soluble plus exchangeable fraction and the carbonate fraction into the more stable fractions. Statistical analyses of metal concentrations in individual fractions after the first and the last cycle of incubation are included in Fig. 2. The detailed pathways of transformation depended primarily on the nature of the metal and soil properties. In general, Cu, Ni, and Zn were transferred from the carbonate and, to some extent, exchangeable (soluble plus exchangeable) fractions into the organic matter fraction in the sandy soil and into the reducible oxide fraction in the loessial soil. Chromium was transferred from the carbonate and exchangeable fractions into the organic matter fraction. In general, more metal was in the organic matter-bound fraction in the sandy soil, whereas more metal was in the residual and reducible oxide-bound fractions and less in the organic matter fraction in the loessial soil (Fig. 2). The loading level had a secondary effect on the pathway and rate of transfer and redistribution for each metal.
Soluble chromium added to the two soils moved mainly into the organic matter fraction after the initial redistribution (Fig. 2). During subsequent incubation involving wetting and drying cycles, additional Cr was transferred slowly from the carbonate fraction into the organic matter and, to a lesser extent, reducible oxide fractions (Fig. 2, Table 3). This was similar to Cr redistribution in soils under the field capacity regime (Han and Banin, 1999). However, compared with the wetting-drying cycle and field capacity regimes, more Cr was transferred into the carbonate fraction under the saturated paste regime, resulting in less Cr in the organic matter fraction (Table 3). At the end of the year, the quasi-equilibrium state present in control soils was not attained in the two soils, mainly because of the limited transfer of Cr into the residual fraction. Drying and remoistening air-dry soils lowered greatly their ability to oxidize Cr (Bartlett and James, 1980). Since Cr3+ has an ionic radius (0.64 × 10−10 m) similar to Mg (0.65 × 10−10 m) and trivalent Fe (0.65 × 10−10 m), it is possible that Cr3+ could substitute readily for Mg in silicates and for Fe3+ in iron oxides. This helps explain the high proportion of Cr in the residual fraction in the native soil (Fig. 1). On the other hand, humic acids have a high affinity for Cr (III) (Adriano, 1986). Thus, our results showed when soluble Cr was added to soils, added Cr3+ was initially and immediately bound to the organic matter fraction. As a result of its slow conversion into the reducible oxide and residual fractions, Cr in the amended soils largely departed from and remained removed from the quasi-equilibrium, but it did approach it with time.
The addition of soluble Cu, Ni, and Zn to the two soils resulted in initial binding in the carbonate fraction (Fig. 2). During prolonged incubation under the wetting-drying cycle regime, the three metals continued to be transferred from the more labile fractions (the exchangeable and carbonate fractions) into the organic matter fraction in the sandy soil and reducible oxide fractions in the loessial soil (Fig. 2, Table 3). In addition, Ni was also moved into the organic matter fraction in the loessial soil (Fig. 2, Table 3). Compared with the wetting-drying cycle regime, more added Cu, Ni, and Zn moved into the easily reducible oxides at the field capacity regime and more Cu, Ni, and Zn moved into the carbonate and reducible oxide fractions at the saturated paste regime (Table 3).
Statistics showed that in both soils, concentrations of Cu, Cr, Ni, and Zn in the carbonate (CARB) fraction decreased significantly and those in the organic matter fraction significantly increased after 1 year of incubation (at P < 0.05 probability, Fig. 2). In addition, Ni and Zn in the exchangeable fraction in both soils and Cu in the exchangeable fraction in the loessial soil decreased significantly after 1 year (Fig. 2). Chromium in the reducible oxide fraction in the sandy soil and in the easily reducible oxide fraction in the loessial soil, Ni and Zn in the easily reducible oxide and reducible oxide fractions in both soils, and Cu in the reducible oxide fraction in the loessial soil increased significantly after 1 year of incubation.
Zinc adsorption can occur via exchange of Zn2+ and Zn(OH)+ with surface-bound Ca2+ on calcite (Zachara et al., 1988). Zinc and Ni form surface complexes on calcite as hydrate until they are incorporated into the structure via recrystallization (Zachara et al., 1991). The selectivities of metal sorption on calcite were reported as follows: Cd > Zn > Ni (Zachara et al., 1991). The easily reducible oxide-bound metals were mainly from Mn oxides (Chao, 1972; Shuman, 1982, 1985). At pH > 6, Zn sorption on Mn oxide increased abruptly because of hydroxylation of the ions (Loganathan et al., 1977), and a high soil pH favored Zn sorption on Mn oxides as a result of the greater proportion of Zn hydroxy species (Zasoski and Burau, 1988). On the other hand, Cr and Cu have higher stability constants with organic matter. These may explain, in part, the observation that there was a higher concentration of Zn (20-25% of the total) in the easily reducible oxide fraction than Cu, Ni, and Cr in the loessial soil at the 3T level after 1 year of incubation (Fig. 2).
Other soil properties are also important in determining sinks for redistributions of metals in soils. Since sandy soil is relatively poor in oxides, carbonate, and clay minerals, organic matter played an important role in binding added heavy metals. The higher affinity of Cu for organic matter compared with Ni and Zn, caused Cu to be preferentially bound by the organic matter fraction. This helps explain the higher amounts of Ni and Zn and the lower Cu amount in the exchangeable fraction in the sandy soil. Taking the 5T treatment in the sandy soil as an example, copper in the organic matter fraction at the 5T treatment in the sandy soil was 28% (14 mg kg−1) of the added copper, whereas Ni was 21% (7 mg kg−1) and Zn 10% (14 mg kg−1) (Note: the total of added Zn was much higher than added Cu). On the other hand, the loessial soil is rich in carbonate, Mn and Fe oxides, and clay minerals. Zinc and Ni competed successfully with Cu, and, thus, both were preferentially bound by carbonate and Mn oxides, resulting in lowering Ni and Zn concentrations in the exchangeable fraction (compared with the higher amount of Cu in the soluble and exchangeable fraction) in the loessial soil.
Changes of Overall Lability of Metals in Soils under Varying Moisture Regimes
We further grouped the metal fractions among solid-phase components into the readily labile, potentially labile, and less labile groups (Han and Banin, 2000). Metals in the exchangeable fraction are readily labile, whereas those in the carbonate, organic matter-, and readily reducible oxide-bound fractions are potentially labile and highly dependent on soil properties and environmental factors (such as pH and Eh change). Metals in the reducible oxide and residual fractions are regarded as less labile fractions. The overall lability of heavy metals in two soils under a wetting-drying cycle moisture regime for 1 year decreased, in general, with time (Fig. 3). However, the over-all lability of metals is strongly dependent on the nature of the metal and soil properties (Fig. 3). In sandy soil, lability of the metal decreased as metals were redistributed from the readily labile fraction into the potentially labile fractions. In loessial soil, however, metal lability decreased as metal redistribution took place from both readily labile and potentially labile fractions into the less labile fractions due to higher amounts of clay minerals and oxides than in sandy soil. Overall, lability changes of metals in loessial soil were larger than those in the sandy soil (Fig. 3). This implies that reactivities of heavy metals in loessial soil are stronger than those in sandy soil. Furthermore, changes of overall lability of Ni, Zn, and Cu were larger than for Cr (Fig. 3).
Soil moisture regime affects strongly the over-all lability of metals in soils as a result of its influence on metal redistribution among solid-phase components (Fig. 4). In general, soils at the saturated paste regime had stronger reactivity than those at the wetting-drying cycle, which, in turn, had higher reactivity than those at the field capacity regime, especially in the loessial soil and for Zn, Ni, and Cu. (Fig. 4). The moisture effects on metal reactivity are influenced strongly by soil properties and the nature of the metal (Fig. 4). After 1 year of incubation, loessial soil at the saturated paste regime had higher Cu, Ni, and Zn concentrations in the less labile fractions than those at the wetting-drying cycle regime. The loessial soil of the field capacity regime had the lowest Cu, Ni, and Zn concentrations in the less labile fractions. Moisture effects in the sandy soils were not significant because of their lesser amounts of clay minerals and oxides.
In conclusion, heavy metals added to aridzone soils in soluble form under wetting-drying cycle moisture regimes are redistributed slowly from the soluble and exchangeable and carbonate bound fractions into more stable fractions, resulting in a lowering of the overall lability of metals in soils. The pathways of metal redistribution are strongly affected by the nature of the metal and soil properties. Soil moisture regime also affect redistribution and overall lability of added soluble metals in arid-zone soils. Soils of both saturated paste and wetting-drying cycle moisture regimes had, in general, higher metal reactivity than soils at field capacity, resulting in metal redistribution into more stable fractions.
This study is a part of research performed at the Department of Soil and Water Sciences, The Faculty of Agricultural, Food, and Environmental Quality Sciences, The Hebrew University, Rehovot, Israel, and of a thesis submitted to the Hebrew University by Han F.X. as partial fulfillment of the requirements for the Ph.D. degree. The study was supported, in part, by funds from the Ministry of Science and the Arts, Israel, and the European Community (ISC-8911 ISR (ENV) #3006), and from the Hebrew University of Jerusalem.
Adriano, D. C. 1986. Trace Elements in the Terrestrial Environment, Springer-Verlag, New York.
Avnimelech Y. 1993. Irrigation with sewage effluents: The Israeli experience. Environ. Sci. Technol. 27:1281-1281.
Banin, A., J. Navrot, Y. Noi, and D. Yoles. 1981. Accumulation of heavy metals in arid-zone soils irrigated with treated sewage effluents and their uptake by Rhodes grass. J. Environ. Qual. 10:536-540.
Banin, A., Z. Gerstl, P. Fine, Z. Metzger, and D. Newrzella. 1990. Minimizing soil contamination through control of sludge transformations in soil. Joint German-Israel research projects report. Wt 8678/458. Hebrew Univ. of Jerusalem, Israel.
Banin, A., S. Nir, G. W. Brummer, F.X. Han, C. Serban, and J. Krumnohler. 1995. Cd pollution in soils: Long-term processes in the solid phase, their characterization and models for their prediction. Joint Israel-Commission of the European Communities research projects. ISC-8911-ISR (ENV). Hebrew Univ. of Jerusalem, Israel.
Bartlett, R., and B. James. 1980. Studying dried, stored soil samples-some pitfalls. Soil Sci. Soc. Am. J. 44:720-724.
Ben-Dor, E., and A. Banin. 1989. Determination of organic matter content in arid-zone soils using a simple "loss-on-ignition" method. Commun. Soil Sci. Plant Anal. 201:1675-1695.
Belzile, N., P. Lecomte, and A. Tessier. 1989. Testing readsorption of trace elements during partial chemical extractions of bottom sediments. Environ. Sci. Technol. 23:1015-1020.
Chang, A. C., A. L. Page, J. E. Warneke, and E. Grgurevic. 1984. Sequential extraction of soil heavy metals following a sludge applications. J. Environ. Qual. 13:33-38.
Chao, T. T. 1972. Selective dissolution of manganese oxides from soils and sediments with acidified hydroxylamine hydrochloride. Soil Sci. Soc. Am. Proc. 36:704-768.
Feigin, A., I. Ravina, and J. Shalhevet. 1991. Irrigation with treated sewage effluent. Springer-Verlag, New York.
Han, F. X., A. T. Hu, and H. Y. Qin. 1995. Transformation and distribution of forms of zinc
in acid, neutral and calcareous soils of China. Geoderma 66: 121-135.
Han, F. X., and A. Banin. 1995. Selective sequential dissolution techniques for trace metals in arid-zone soils: The carbonate dissolution step. Commun. Soil Sci. Plant Anal. 26:553-576.
Han, F. X., and A. Banin. 1996. Changes of manganese fractionation among solid-phase components in saturated arid-zone soils: Pathways and short- and intermediate-term kinetics. Soil Sci. Soc. Am. J. 60:1072-1080.
Han, F. X., and A. Banin. 1997. Long-term transformations and redistribution
of potentially toxic heavy metals in arid-zone soils. I: Under saturated conditions. Water Air Soil Pollut. 95:399-423.
Han, F. X., and A. Banin. 1999. Long-term transformations and redistribution
of potentially toxic heavy metals in arid-zone soils. II: Under field capacity regime. Water Air Soil Pollut. 114:221-250.
Han, F. X., W. L. Kingery, H. M. Selim, and P. Gerald. 2000. Accumulation of heavy metals in a long-term poultry waste-amended soil. Soil Sci. 165:260-268.
Han, F. X., and A. Banin. 2000. Long-term transformations of Cd, Co, Cu, Ni, Zn, V, Mn and Fe in the native arid-zone soils under saturated condition. Commun. Soil Sci. Plant Anal. 31:943-957.
Hazra, G. C., B. Mandal, and L. N. Mandal. 1987. Distribution of zinc
fractions and their transformation in submerged rice soils. Plant Soil 104:175-181.
Gupta, S. K., and K. Y. Chen. 1975. Partitioning of trace metals in selective chemical fractions of near shore sediments. Environ. Lett. 10:129-158.
Kheboian, C., and C. Bauer. 1987. Accuracy of selective extraction procedures for metal speciation in model aquatic sediments. Anal. Chem. 59:1417-1423.
Kim, N. D., and J. E. Fergusson. 1991. Effectiveness of a commonly used sequential extraction techniques in determining the speciation of cadmium in soils. Sci. Total Environ. 105:191-209.
Loganathan, P., R. G. Burau, and D. W. Fuerstenau. 1977. Influence of pH on the sorption of Co2+
by a hydrous manganese oxide. Soil Sci. Soc. Am. J. 41:57-62.
Mandal, L. N., and B. Mandal. 1986. Zinc
fractions in soils in relation to zinc
nutrition of lowland rice. Soil Sci. 142:141-148.
Martin, J. M., P. Nirel, and A. J. Thomas. 1987. Sequential extraction techniques: Promises and problems. Mar. Chem. 22:313-341.
McGrath, S. P., and J. Cegarra. 1992. Chemical extractability of heavy metals during and after long-term applications of sewage sludge to soil. J. Soil Sci. 43:313-321.
McLaren, R. G., and G. S. P. Ritchie. 1993. The long-term fate of copper
fertilizer applied to a lateristic sandy soil in Western Australia. Aust. J. Soil Res. 93:39-50.
Nelson, R. E. 1982. Carbonate and gypsum. In
Methods of Soil Analysis, part 2. Agronomy 9, 2nd Ed. A.L. Page et al. (eds.). ASA, Madison, WI, pp. 181-197.
Patrick, W. H. Jr., and A. Jugsujinda. 1992. Sequential reduction and oxidation of inorganic nitrogen, manganese, and iron in flooded soil. Soil Sci. Soc. Am. J. 56:1071-1073.
Ratner-Zohar, Y., A. Banin, and Y. Chen. 1983. Oven drying as a pretreatment for surface area determination of soils and clays. Soil Sci. Soc. Am. J. 47:1056-1058.
Ribet, I., C. J. Ptacek, D. W. Blowes, and J. L. Jambor. 1995. The potential for metal release by reductive dissolution of weathered mine tailings. J. Contam. Hydrol. 17:239-273.
Schwab, A. P., and W. L. Lindsay. 1983. The effect of redox on the solubility and availability of manganese in a calcareous soil. Soil Sci. Soc. Am. J. 47:217-220.
Shuman, L. M. 1982. Separating soil iron- and manganese-oxide fractions for microelement analysis. Soil Sci. Soc. Am. J. 46:1099-1102.
Shuman, L. M. 1985. Effects of tillage on the distribution of manganese, copper
, iron, and zinc
in soil fractions. Soil Sci. Soc. Am. J. 49:1117-1122.
Silviera, D. J., and L. E. Sommers. 1977. Extractability of copper
, cadmium and lead in soils incubated with sewage sludge. J. Environ. Qual. 6:47-52.
Soon, Y. K. 1994. Changes in forms of soil zinc
after 23 years of cropping following clearing of a boreal forest. Can. J. Soil Sci. 74:179-183.
Sposito, G., L. J. Lund, and A. C. Chang. 1982. Trace metal chemistry in arid-zone field soils amended with sewage sludge: I. Fractionation of Ni, Cu, Zn, Cd, and Pb in solid phases. Soil Sci. Soc. Am. J. 46:260-264.
Sposito, G., C. S. LeVesque, J. P. LeClaire, and A. C. Chang. 1983. Trace elements chemistry in arid-zone field soils amended with sewage sludge: III. Effect of the time on the extraction of trace metals. Soil Sci. Soc. Am. J. 47:898-902.
Tessier, A., P. G. C. Campell, and M. Bisson. 1979. Sequential extraction procedure for the speciation of particulate trace metals. Anal. Chem. 51:844-851.
Tipping, E., N. B. Hetherington, J. Hilton, D. W. Thompson, E. Bowles, and J. Hamilton-Taylor. 1985. Artifacts in the use of selective chemical extraction to determine distributions of metals between oxides of manganese and iron. Anal. Chem. 57:1944-1946.
Zachara, J. M., C. E. Cowan, and C. T. Resch. 1991. Sorption of divalent metals on calcite. Geochim. Cosmochim. Acta 55:1549-1562.
Zachara, J. M., J. A. Kittrick, and J. B. Harsh. 1988. The mechanism of Zn2+
adsorption on calcite. Geochim. Cosmochim. Acta 52:2281-2291.
Zasoski, R. J., and R. G. Burau. 1988. Sorption and sorptive interaction of cadmium and zinc
on hydrous manganese oxide. Soil Sci. Soc. Am. J. 52:81-87.