Environ. Sci. Technol. 2009, 43, 837–842
Kinetic Interactions of EDDS with Soils. 2. Metal-EDDS Complexes in Uncontaminated and Metal-Contaminated Soils D A N I E L C . W . T S A N G , * ,† THEO C. M. YIP,‡ AND IRENE M. C. LO‡ Department of Civil and Natural Resources Engineering, University of Canterbury, New Zealand, and Department of Civil and Environmental Engineering, The Hong Kong University of Science and Technology, Hong Kong, China
Received January 16, 2008. Revised manuscript received November 10, 2008. Accepted November 18, 2008.
The effectiveness of biodegradable EDDS has aroused substantial interest over the past few years, yet there has been little information on the fate of metal-EDDS complexes that, prior to biodegradation, stay in contact with the soils. This study conducted 7-day batch experiments to investigate the kinetic interactions of CuEDDS2-, ZnEDDS2-, PbEDDS2-, and AlEDDS-, which are newly formed during EDDS application, with uncontaminated and metal-contaminated soils at pH 5.5 and 8. In uncontaminated soils, metal-EDDS complexes were adsorbed and induced mineral dissolution. In contaminated soils, on the contrary, significant metal exchange with sorbed metals on the soil surfaces (i.e., Cu, Zn, and Pb) resulted in a greater extent of metal resorption of the metal-EDDS complexes. The interactions of metal-EDDS complexes, moreover, are influenced by the characteristics of the metal center. Compared with ZnEDDS2- and PbEDDS2-, CuEDDS2- was least adsorbed or exchanged, which may be attributed to higher ionic potential and the electron configuration of Cu. In addition, AlEDDS- was partially exchanged on the soil surfaces at low pH while entirely dissociated in solution at high pH. Therefore, the fate of individual metal-EDDS complexes in the subsurface depends on the metal center, other sorbed metals and mineral cations on soils, and solution pH.
Introduction EDDS ([S,S]-ethylenediaminedisuccinic acid, a structural isomer of EDTA) has recently emerged as a promising chelating agent for enhancing in situ (e.g., flushing, phytoextraction) and ex situ (e.g., washing, heap leaching) remediation of heavy-metal-contaminated soils (1-9). EDDS could efficiently extract heavy metals (1, 4, 5, 8) and is less toxic than EDTA to plants, fungi, and microorganisms (3). It has been shown that, in addition to extraction of target metals, mineral cations such as Al, Fe, Ca, and Mn are dissolved by uncomplexed (i.e., free) EDDS, especially with increasing reaction time (4, 5, 8, 10-12). The reactivity of metal-EDDS complexes has to be considered because the reactions of EDDS in the environment occur predominantly in the form * Corresponding author e-mail:
[email protected]; fax: 64 3 364 2758; tel: 64 3 364 2394. † University of Canterbury. ‡ The Hong Kong University of Science and Technology. 10.1021/es8020292 CCC: $40.75
Published on Web 01/06/2009
2009 American Chemical Society
of metal-EDDS complexes (13). A negligible amount of EDDS, if not in excess, would remain uncomplexed according to speciation calculation (5, 12, 14). Although most metal-EDDS complexes were biodegradable during wastewater treatment (15, 16) and in soils (11, 14) and residual problems were insignificant in the long run (7), a lag phase, in the range of 7-32 days, was necessary for the population growth of appropriate microbes in soils without sludge amendment (6, 11, 14). The length of lag phase may vary with the soil type and extent of metal contamination (11). Therefore, biodegradation is minimal in the course of soil washing and flushing that has a reaction time or residence time in the order of days. The newly formed metal-EDDS complexes would stay in contact with contaminated soils until solution extraction; some may travel in the subsurface to react with uncontaminated soils in the case of incomplete recovery of residual EDDS solution (e.g., Figure 1, Supporting Information). The interactions of metal-EDDS complexes with soils remain uninvestigated. In the literature, adsorption of uncomplexed EDTA or metal-EDTA complexes on Fe and Al oxides increased with decreasing pH (17, 18) and increasing quantity of oxides (19). The metal-EDTA complexes were mobile in the subsurface (20, 21). The weaker (i.e., less stable) EDTA complexes of Ca and Mg were dissociated at low pH and uncomplexed EDTA was adsorbed, whereas stronger metal-EDTA complexes were adsorbed as ternary surface complexes (17). Complexes of the same structure in solution displayed similar adsorption behavior, while complexes of the same charge but different structure had completely different adsorption behavior (13, 17, 18). The adsorbed complexes, if mononuclear, can destabilize the M-O bonds at mineral surfaces, promoting the detachment of mineral cations into the solution phase (i.e., ligandpromoted dissolution) (22-24). Dissolution rates increased with the number of ligand functional groups coordinated to the surfaces (25). Both uncomplexed EDTA and metal-EDTA complexes could enhance mineral dissolution; but the dissolution kinetics of metal-EDTA complexes varied with the metal center and was one or two orders of magnitude slower than that of uncomplexed EDTA (26). In addition, organic matter that is closely bound to minerals was simultaneously mobilized (19, 27, 28). Mineral and organic matter dissolution would reduce shear strength and aggregate stability of soils and influence the metal speciation in solution. In consideration of these findings, the fate of metalEDDS complexes probably depends on solution pH, soil surface characteristics, and type of metal-EDDS complexes. Thisstudyinvestigatedthe7-daykineticinteractionsofmetalEDDS complexes with uncontaminated and artificially contaminated soils. The adsorption of metal-EDDS complexes, extraction of target metals, metal resorption from metal-EDDS complexes, and mineral dissolution, and metal and EDDS speciation were evaluated to investigate the fate of metal-EDDS complexes in uncontaminated and contaminated regions.
Experimental Section Soils and Metal-EDDS Solutions. The soil characteristics, artificial contamination, and the resulting Cu, Zn, and Pb distribution were reported in the companion paper (12). Artificially contaminated soils were used so that the fate of metal-EDDS complexes in uncontaminated soils could be directly compared with that in contaminated soils. Competitive sorption resulted in higher concentration of Cu, which was preferentially sorbed, than Zn and Pb in the soils (29). VOL. 43, NO. 3, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
837
FIGURE 1. Adsorption kinetics of 3 mM metal-EDDS complexes on (a) uncontaminated soils; and (b) contaminated soils (2, ∆ CuEDDS2-; 9, 0 ZnEDDS2-; b, O PbEDDS2-; closed and open symbols represent data at pH 5.5 and pH 8, respectively). It is noted that a large fraction of weakly bound metals was present in the artificially contaminated soils (12), whereas metals are in general more strongly bound in fieldcontaminated soils due to the aging effect (1, 28). Metal speciation in soils (e.g., Zn phyllosilicate) also plays an important role in determining the metal extraction efficiency by EDDS (30). The results of the companion paper (12) indicated that target metals (Cu, Zn, and Pb) were effectively extracted by and fully complexed with EDDS. The major dissolved mineral cation was Al, which was complexed with EDDS to varying extent depending on pH. Dissolution of Fe was minor under EDDS deficiency; contribution of FeEDDS- to total EDDS was important only under EDDS excess at pH 8. Therefore, CuEDDS2-, ZnEDDS2-, PbEDDS2-, and AlEDDS- were of primary concern in this study. Yet, FeEDDS- may be more important in some soils (10), depending on the type of oxides and respective EDDS-promoted dissolution kinetics (13, 26). Metal-EDDS solutions were prepared by dissolving equimolar amounts (3 mM) of respective metal-nitrate salts (analytical-grade Al(NO3)3, Cu(NO3)2, Zn(NO3)2, or Pb(NO3)2, Fisher Scientific) and [S,S]-EDDS (30% Na3EDDS solution, Innospec Ltd., HK) in background solution (0.01 M NaNO3 in >18.0 MΩ ultrapure water) and heating at 90 °C for 3 h to ensure a complete formation of metal-EDDS complexes (18, 31). The solutions were buffered at pH 5.5 with 0.01 M MES (2-morpholinoethane-sulfonic acid) buffer, and pH 8 with 0.01 M HEPES [4-(2-hydroxyethyl)1-piperazine-ethane-sulfonic acid], respectively. The solution pH varied by 0.3 at most after batch experiments. To avoid biodegradation and photodegradation of EDDS, metal-EDDS solutions contained sodium azide (0.2 g L-1 NaN3) and were kept in dark. The AlEDDS solution was used right after preparation. Batch Kinetic Experiments. All experiments were run in duplicate or triplicate. The uncontaminated and contaminated soils (0.5 g) were mixed with different metal-EDDS solutions at a soil/solution ratio of 50 g L-1 by end-overend shaking at 26 rpm for 16 different reaction times (1 min to 168 h) at room temperature. The molar ratio of 838
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 43, NO. 3, 2009
FIGURE 2. Kinetics of heavy metal extraction from contaminated soils by 3 mM metal-EDDS complexes and uncomplexed (free) EDDS: (a) Cu; (b) Zn; and (c) Pb (2, ∆ CuEDDS2-; 9, 0 ZnEDDS2-; b, O PbEDDS2-; 1, 3 AlEDDS-; [, ] uncomplexed EDDS; closed and open symbols represent data at pH 5.5 and pH 8, respectively; error bars represent the standard deviations).
EDDS in metal-EDDS solution to target metals in contaminated soils was approximately equal to one. Sequential extractions of soils (temporal change of Cu, Zn, and Pb distribution), solution analyses (concentrations of EDDS, target metals, mineral cations, and dissolved organic matter), and metal and EDDS speciation calculations using Visual MINTEQ (32), which is based on MINTEQA2 (33), were performed as described in the companion paper (12). Adsorption of CuEDDS2-, ZnEDDS2-, and PbEDDS2-, which do not dissociate in initial solution, was calculated by the difference of EDDS concentrations in solution after various reaction times. In the absence of EDDS, metal extraction (except at pH 5.5 where 20% of Zn was extracted) and mineral dissolution of contaminated soils were negligible (12), whereas dissolution of Al (119 mg kg-1) and Ca (11 mg kg-1) of uncontaminated soils was observed. Thus, further extraction or dissolution, if observed, was induced by metal-EDDS complexes.
FIGURE 3. EDDS speciation of 3 mM metal-EDDS solution in contact with contaminated soils at pH 5.5: (a) CuEDDS2-; (b) ZnEDDS2-; (c) PbEDDS2-; and (d) AlEDDS- (initial EDDS speciation was 100% CuEDDS2-, 99.9% ZnEDDS2-, 99.9% PbEDDS2-, and 37.0% AlEDDS-).
Results and Discussion Adsorption of Metal-EDDS Complexes. The metal-EDDS adsorption was more significant at pH 5.5 than at pH 8 (Figure 1), which is well recognized as ligand-like adsorption behavior, because hydroxyl groups are more protonated and the surfaces become more positive at lower pH. Metal-ligand complexes can be adsorbed as ternary surface complexes on oxide hydroxyl groups through inner-sphere complexation (13, 17, 18), outer-sphere complexation (i.e., electrostatic interaction), and hydrogen bonding (34-37). It is uncertain about the dominant adsorption mechanisms (13), which may vary with solution conditions: from type B (i.e., bridged by ligand) to type A (i.e., bridged by metal) surface complex with increasing pH (17, 38); from inner-sphere to outer-sphere with increasing metal-to-ligand ratio (39, 40); and increasing significance of hydrogen bonding with increasing complex concentration and pH (36, 37). Adsorbed metal-EDDS complexes on uncontaminated soils were about 11-22 mmol kg-1 (18-37%) at pH 5.5 (Figure 1a). Metal-EDDS complexes are expected to be initially adsorbed on soils and then, in part, released back into solution with time due to EDDS-promoted dissolution if the new metal-EDDS complexes containing mineral cations are less adsorbed (17, 18). The decrease of EDDS adsorption at the end of 168 h was probably a sign of slow detachment of adsorbed metal-EDDS complexes from crystalline oxides (e.g., rate coefficients of goethite dissolution by metal-EDTA complexes were ∼10-7 s-1 (26)) and minor adsorption of the new metal-EDDS complexes. Adsorbed metal-EDDS complexes on contaminated soils, on the other hand, were about 3-10 mmol kg-1 (5-17%) at pH 5.5 (Figure 1b). The lesser extent of metal-EDDS adsorption was probably because, in contaminated soils, metal-EDDS complexes preferentially undergo metal exchange with sorbed metals (subsequent section) rather than adsorption on surface hydroxyl groups. It should be noted that PbEDDS2- was more significantly adsorbed than CuEDDS2- and ZnEDDS2- at low pH (Figure 1). Adsorbed ZnEDDS2- and PbEDDS2- (12 and 22 mmol kg-1) on uncontaminated soils (Figure 1a) corresponded well with 10 and 23 mmol kg-1 of sorbed Zn and Pb (Figure S2a,
Supporting Information), respectively. Speciation calculations also revealed minimal dissociation of metal-EDDS complexes (except AlEDDS-) in initial solution prior to interaction with soils. The observed difference in metal-EDDS adsorption, therefore, is unlikely to derive from dissociation of weaker complexes in solution at low pH. As a structural isomer of EDTA, EDDS is likely to complex with Cu, Zn, and Pb as a quinquedentate structure in solution (5-coordination with metal, where one uncomplexed carboxyl group on the ligand and one water molecule coordinated to the metal center are free for coordination to the surfaces). The complexes of the same structure are supposed to display the same adsorption affinity (13, 17, 18), which was not observed in this study. It is probable that the characteristics of the metal center also influence the metal-EDDS adsorption. The ionic potential (z2/r) of Cu is similar to Zn and much larger than Pb, because the divalent cation with smaller ionic radius (Cu (0.072 nm) ∼ Zn (0.074 nm) < Pb (0.12 nm)) possesses a larger electric field and less covalent character (41, 42). For the metal with larger ionic potential, there is stronger attractive force between the ligand and the metal center (43); CuEDDS2- and ZnEDDS2-, thus, have a compact molecular structure that is more stable in solution and less favorable for adsorption as ternary surface complexes. These characteristics of the metal center account for the difference in empirical stability constants of metal-EDDS complexes (44). The determination of AlEDDS- adsorption, however, was unsuccessful, as a large amount of visible Al precipitates formed when Al was displaced from EDDS complexes by excessive CuSO4 prior to colorimetric analysis. It has been shown that uncomplexed Al could precipitate as Al(OH)3 in solution at pH 5 (45). Although it is possible to directly measure EDDS concentration at a low concentration range (31), the determination of AlEDDS- adsorption is difficult in view of AlEDDS- dissociation in solution (subsequent section). Metal Exchange with Sorbed Metals. In addition to the adsorption shown in Figure 1, some adsorbed metal-EDDS complexes may induce metal exchange on the soil surfaces. VOL. 43, NO. 3, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
839
FIGURE 4. Mineral dissolution kinetics of uncontaminated soils by 3 mM metal-EDDS complexes: (a) CuEDDS2-; (b) ZnEDDS2-; (c) PbEDDS2-; and (d) AlEDDS- (b, O Al; 1, 3 Fe; 9, 0 Ca; 2, ∆ Mn-; closed and open symbols represent data at pH 5.5 and pH 8, respectively; error bars represent the standard deviations). The metal center is displaced from the complex and possibly resorbed on soils (i.e., immobilization) while a new metalligand complex of greater thermodynamic stability is formed and eventually detached from the surfaces (22, 24, 26). Metal exchange with mineral cations in the oxide structures has been investigated in previous studies (26, 39); yet metal exchange with sorbed metals on the soil surfaces (i.e., Cu, Zn, and Pb) could be crucial as well. Sorbed Cu, Zn, and Pb in contaminated soils were extracted by metal-EDDS complexes, despite being slower and less significant than extraction by uncomplexed EDDS (Figure 2). A fraction of metal extraction took place instantaneously (within 1 min) and the rest continued slowly with time, with the first-order rate coefficients of 10-6 to 10-5 s-1 (Table S1). The occurrence of metal exchange was confirmed by the temporal change of EDDS speciation. The proportion of ZnEDDS2- decreased along with an increase of CuEDDS2(Figures 3b and S3b); the proportion of PbEDDS2- decreased along with an increase of CuEDDS2- and ZnEDDS2- (Figures 3c and S3c). During metal exchange, the dissociative tendency of CuEDDS2-, ZnEDDS2-, and PbEDDS2- is important because the metal center (Cu, Zn, or Pb) of adsorbed metal-EDDS complex is first displaced from the intermediate surface complex. The dissociation rate of metal complex is shown to be a function of the ionic potential and ligand field stabilization energy (depending on the electron configuration in d-orbital of the metal center) (42, 46). A complex of Pb, which has low ionic potential, is more easily dissociated, whereas a complex of Cu, which has high ionic potential and four strong equatorial bonds due to Jahn-Teller distortion, is inert to dissociation. Besides, high sorption strength of Pb may facilitate its complex dissociation for metal exchange (47). Therefore, PbEDDS2- resulted in the most significant metal exchange and extraction of sorbed metals, ZnEDDS2the second, and CuEDDS2- the least (Figures 2 and 3). In 840
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 43, NO. 3, 2009
addition, metal exchange at pH 8 (Figure S3), and the resulting extraction of sorbed metals (Figure 2), was less significant probably due to less metal-EDDS adsorption on the surfaces at high pH. Moreover, complex dissociation and metal exchange may become more favorable with decreasing complex concentration and increasing sorbed metal concentrations on the soils (47). The metal distribution of soils (Figure S4) revealed that PbEDDS2- extracted exchangeable, carbonate, and oxide fractions of sorbed Zn and carbonate and oxide fractions of sorbed Cu; and that ZnEDDS2- extracted carbonate and oxide fractions of sorbed Cu. It appears that a portion of weakly and strongly sorbed Cu and Zn on soil minerals could be exchanged, although exchange with sorbed Cu in the oxide fraction took longer (Figure S4). In line with metal exchange, these extracted metals (Cu by ZnEDDS2- and PbEDDS2-; Zn by PbEDDS2-) were almost entirely complexed with EDDS (Figures S5 and S6). On the other hand, Zn extraction by CuEDDS2- was 22% at pH 5.5, which was close to that by background solution (20%) at the same pH (12). Yet, Pb extraction by ZnEDDS2- (18-23%) and by CuEDDS2- (1-3%) (Figure 2) may result from mobilization of organic matter as the extracted Pb was predominantly complexed with DOM rather than with EDDS (Figures S5 and S6). In consequence of fast metal exchange on contaminated soils, resorption of Zn and Pb (Figure S2b) by far exceeded the corresponding metal-EDDS adsorption (Figure 1b). Compared with the initial contamination, weakly sorbed fractions (i.e., exchangeable and carbonate) of Pb were greater whereas those of Cu and Zn were similar (data not shown), suggesting that the majority of resorbed Pb were weakly bound and susceptible to leaching out afterward. In addition, there were smaller exchangeable but larger carbonate fractions of resorbed metals at pH 8 than at pH 5.5. It was found that AlEDDS- extracted Cu, Zn, and Pb (Figure 2), mainly from exchangeable, carbonate, and oxide fractions
(Figure S4), together with Al resorption on contaminated soils (Figure S2). The rate coefficients of metal extraction were about 10-7 and 10-6 s-1 at pH 5.5 and 8, respectively (Table S1). Despite having a sexidentate structure (6coordination with metal, where all six coordinative positions are occupied), AlEDDS- readily dissociates due to strong hydrolysis of Al, analogous to AlEDTA- (18, 26). At equilibrium, only about 36% of Al was complexed with EDDS at pH 5.5 and negligible amount at pH 8 (Figures S5d and S6d). Thus, metal extraction by AlEDDS- (Figure 2) was largely facilitated by uncomplexed EDDS and, to a lesser extent, by metal exchange of AlEDDS-, as indicated by corresponding decrease in the proportions of uncomplexed EDDS and AlEDDS- (Figures 3d and S3d). Metal Exchange with Mineral Cations. Adsorbed metalEDDS complexes may also exchange with mineral cations, resulting in soil dissolution. In uncontaminated soils, Al dissolution was greatest, while Fe dissolution was noteworthy as well; Ca and Mn dissolution proved to be minor (Figure 4). The rate coefficients of mineral dissolution (10-10 to 10-6 s-1, Table S1) followed the order of PbEDDS2- > ZnEDDS2> CuEDDS2-. Nevertheless, as reflected by the presence/ absence of AlEDDS-, metal exchange partially contributed to Al dissolution by PbEDDS2- at pH 5.5 but it was negligible in the other cases (Figures S7 and S8). In view of the significant amount of Al (119 mg kg-1) dissolved by background solution, cation exchange of readily available Al sources (e.g., electrostatically bound) (48, 49) appears to primarily result in Al dissolution, which was more substantial in metal-EDDS solution because cation exchange was enhanced by varying extent of Cu, Zn, and Pb resorption (Figure S2a) and additional 9 mM of Na from Na3EDDS. Dissolved Al resulting from cation exchange may form colloidal precipitates that pass through 0.2 µm filters. Moreover, Al dissolution may be facilitated by organic matter dissolution and colloid mobilization, which were greater at pH 8. Thus, the predominant forms of Al in solution were colloidal Al(OH)3, Al-DOM, and hydrolyzed Al species (Figures S7 and S8). On the other hand, dissolved Fe was predominantly bound with DOM during the initial stage, while the importance of FeEDDS- increased with time (Figures S7 and S8). Significant DOM binding to Fe in the presence of EDDS was corroborated by recent findings (14); colloidal Fe(OH)3 was present together with FeEDDS- at pH 8 during the later stage probably due to the instability of FeEDDS- (16). Metal exchange with Fe was fastest and most significant for PbEDDS2-, followed by ZnEDDS2-, and CuEDDS2- the least, which was consistent with the strength of metal-EDDS complexes as discussed before. It is also noted that Fe dissolution (Figure 4) and the significance of FeEDDS- (Figures S7 and S8) were greater at pH 8 than at pH 5.5 despite less metal-EDDS adsorption. A change of adsorbed form from binuclear to mononuclear ternary surface complex, which promotes mineral dissolution, was less likely for quiquedentate metal-EDDS complexes (17, 26). Nevertheless, FeEDDS- that is newly formed after metal exchange may have a higher tendency to detach from the mineral surfaces at pH 8 than at pH 5.5, because EDDS complexation with Fe is stronger (more stable) at pH 8 and thus more capable of destabilizing the Fe-O bonds in the oxide structures so as to maximize the bonding between Fe and EDDS (24, 50). Moreover, initial increase of Fe dissolution was probably attributed to enhanced organic matter dissolution and colloid mobilization at higher pH. However, in contaminated soils (Figure S9), where readily available Al sources had largely been depleted during the artificial contamination process, notably less Al dissolution was observed; constant Ca dissolution at low pH became relatively important. Dissolution of Fe was negligible although metal exchange with Fe is more thermodynamically favorable than that with other metals. Adsorbed metal-EDDS com-
plexes appear to primarily exchange with sorbed metals that are readily accessible on the surfaces rather than with mineral cations that reside in the oxide structures, in agreement with the findings of uncomplexed EDDS and EDTA studies (12, 21). Engineering Implications. During application of uncomplexed EDDS, metal-EDDS complexes are newly formed and, prior to the onset of biodegradation, undergo adsorption and metal exchange on the soil surfaces. These interactions result in metal resorption that reduces the effectiveness of EDDS application. In contaminated soils metal-EDDS complexes principally exchange with sorbed metals on the soil surfaces and lead to a greater metal resorption, of which the majority are weakly bound. However, if residual metal-EDDS solution is leaked to uncontaminated soils, metal-EDDS complexes are adsorbed and result in mineral dissolution. The influences of metal-EDDS complexes are distinct in contaminated and uncontaminated regions. The significance of the interactions varies among metal-EDDS complexes, depending on the characteristics of the metal center. Newly formed CuEDDS2- is less likely to be adsorbed and exchanged and, in turn, cause less metal resorption and mineral dissolution, while extraction of Pb, which imposes greater risk to human health than Cu, is impeded by significant metal exchange of PbEDDS2- with other sorbed metals and mineral cations. It is, therefore, important to extract the residual metal-EDDS solution prior to their further interactions with soils.
Acknowledgments We thank the Research Grant Council of Hong Kong for providing financial support under General Research Fund with project account 616608 for this research study.
Note Added after ASAP Publication Due to a production error, this paper published ASAP January 6, 2009 with an incorrect version of the Supporting Information; the corrected version published ASAP January 29, 2009.
Supporting Information Available First-order rate coefficients (Table S1); schematic diagram (Figure S1); metal resorption (Figure S2); EDDS speciation (Figure S3); metal distribution in soils (Figure S4); metal speciation in solution (Figure S5–S8); mineral dissolution of contaminated soils (Figure S9). This material is available free of charge via the Internet at http://pubs.acs.org.
Literature Cited (1) Vandevivere, P.; Hammes, F.; Verstraete, W.; Feijtel, T.; Schowanek, D. Metal decontamination of soil, sediment, and sewage sludge by means of transition metal chelant [S, S]-EDDS. J. Environ. Eng. 2001, 127, 802–811. (2) Kos, B.; Lestan, D. Induced phytoextraction/soil washing of lead using biodegradable chelate and permeable barriers. Environ. Sci. Technol. 2003, 37, 624–629. (3) Grcman, H.; Vodnik, D.; Velikonja-Bolta, S.; Lestan, D. Ethylenediaminedissuccinate as a new chelate for environmentally safe enhanced lead phytoextraction. J. Environ. Qual. 2003, 32, 500–506. (4) Tandy, S.; Bossart, K.; Mueller, R.; Ritschel, J.; Hauser, L.; Schulin, R.; Nowack, B. Extraction of heavy metals from soils using biodegradable chelating agents. Environ. Sci. Technol. 2004, 38, 937–944. (5) Hauser, L.; Tandy, S.; Schulin, R.; Nowack, B. Column extraction of heavy metals from soils using the biodegradable chelating agent EDDS. Environ. Sci. Technol. 2005, 39, 6819–6824. (6) Finzgar, N.; Zumer, A.; Lestan, D. Heap leaching of Cu contaminated soil with [S,S]-EDDS in a closed process loop. J. Hazard Mater. 2006, B135, 418–422. (7) Luo, C. L.; Shen, Z. G.; Li, X. D.; Lou, L. Q. EDDS and EDTAenhanced phytoextraction of metals from artificially contaminated soil and residual effects of chelant compounds. Environ. Pollut. 2006, 144, 862–871. VOL. 43, NO. 3, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
841
(8) Polettini, A.; Pomi, R.; Rolle, E.; Ceremigna, D.; De Propris, L.; Gabellini, M.; Tornato, A. A kinetic study of chelant-assisted remediation of contaminated dredged sediment. J. Hazard. Mater. 2006, B137, 1458–1465. (9) Tandy, S.; Schulin, R.; Nowack, B. Uptake of metals during chelant-assisted phytoextraction with EDDS related to the solubilized metal concentration. Environ. Sci. Technol. 2006, 40, 2753–2758. (10) Koopmans, G. F.; Schenkeveld, W. D. C.; Song, J.; Luo, Y.; Japenga, J.; Temminghoff, E. J. M. Influence of EDDS on metal speciation in soil extracts: Measurement and mechanistic multicomponent modeling. Environ. Sci. Technol. 2008, 42, 1123–1130. (11) Meers, E.; Tack, F. M. G.; Verloo, M. G. Degradability of ethylenediaminedisuccinic acid (EDDS) in metal contaminated soils: Implications for its use soil remediation. Chemosphere 2008, 70, 358–363. (12) Yip, T. C. M.; Tsng, D. C. W.; Ng, K. T. W.; Lo, I. M. C. Kinetic interactions of EDDS with soils: 1. Metal re-sorption and competition under EDDS deficiency. Environ. Sci. Technol. 2009, 43, 831–836. (13) Nowack, B. Environmental chemistry of aminopolycarboxylate chelating agents. Environ. Sci. Technol. 2002, 36, 4009–4016. (14) Tandy, S.; Ammann, A.; Schulin, R.; Nowack, B. Biodegradation and speciation of residual SS-ethylenediaminedisuccinic acid (EDDS) in soil solution left after soil washing. Environ. Pollut. 2006, 142, 191–199. (15) Schowanek, D.; Feijtel, T. C. J.; Perkins, C. M.; Hartman, F. A.; Federle, T. W.; Larson, R. J. Biodegradation of [S,S], [R,R] and mixed stereoisomers of ethylene-diamine-disuccinic acid (EDDS), a transition metal chelator. Chemosphere 1997, 34, 2375–2391. (16) Vandevivere, P. C.; Saveyn, H.; Verstraete, W.; Feijtel, T. C. J.; Schowanek, D. R. Biodegradation of metal-[S,S]-EDDS complexes. Environ. Sci. Technol. 2001, 35, 1765–1770. (17) Nowack, B.; Sigg, L. Adsorption of EDTA and metal-EDTA complexes onto goethite. J. Colloid Interface Sci. 1996, 177, 106– 121. (18) Nowack, B.; Lutzenkirchen, J.; Behra, P.; Sigg, L. Modeling the adsorption of metal-EDTA complexes onto oxides. Environ. Sci. Technol. 1996, 30, 2397–2405. (19) Vulava, V. M.; Seaman, J. C. Mobilization of lead from highly weathered porous material by extracting agents. Environ. Sci. Technol. 2000, 34, 4828–4834. (20) Friedly, J. C.; Kent, D. B.; Davis, J. A. Simulation of the mobility of metal-EDTA complexes in groundwater: the influence of contaminant metals. Environ. Sci. Technol. 2002, 36, 355–363. (21) Tsang, D. C. W.; Lo, I. M. C.; Chan, K. L. Modeling the transport of metals with rate-limited EDTA-promoted extraction and dissolution during EDTA-flushing of copper-contaminated soils. Environ. Sci. Technol. 2007, 41, 3660–3667. (22) Stumm, W. Chemistry of the Soil-Water Interface; John Wiley & Sons: New York, 1992. (23) Blesa, M. A.; Morando, P. J.; Regazzoni, A. E. Chemical Dissolution of Metal Oxides; CRC Press: Boca Raton, FL, 1994. (24) Sposito, G. The Surface Chemistry of Natural Particles; Oxford University Press: New York, 2004. (25) Ludwig, C.; Casey, W. H.; Rock, P. A. Prediction of ligandpromoted dissolution rates from the reactivities of aqueous complexes. Nature 1995, 375, 44–47. (26) Nowack, B.; Sigg, L. Dissolution of Fe(III) (hydr)oxides by metalEDTA complexes. Geochim. Cosmochim. Acta 1997, 61, 951– 963. (27) Yang, Y.; Ratte, D.; Smets, B. F.; Pignatello, J. J.; Grasso, D. Mobilization of soil organic matter by complexing agents and implications for polycyclic aromatic hydrocarbon desorption. Chemosphere 2001, 43, 1013–1021. (28) Tsang, D. C. W.; Zhang, W.; Lo, I. M. C. Copper extraction effectiveness and soil dissolution issues of EDTA-flushing of artificially contaminated soils. Chemosphere 2007, 68, 234–243. (29) Tsang, D. C. W.; Lo, I. M. C. Competitive Cu and Cd sorption and transport in soils: A combined batch kinetics, column and sequential extraction study. Environ. Sci. Technol. 2006, 40, 6655–6661. (30) Kirpichtchikova, T. A.; Manceau, A.; Spadini, L.; Panfili, F.; Marcus, M. A.; Jacquet, T. Speciation and solubility of heavy
842
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 43, NO. 3, 2009
(31)
(32) (33)
(34)
(35)
(36)
(37)
(38)
(39)
(40)
(41) (42)
(43) (44)
(45) (46) (47)
(48)
(49)
(50)
metals in contaminated soil using X-ray microfluorescence, EXAFS spectroscopy, chemical extraction, and thermodynamic modeling. Geochim. Cosmochim. Acta 2006, 70, 2163–2190. Tandy, S.; Schulin, R.; Suter, M. J. F.; Nowack, B. Determination of [S,S’]-ethylenediamine-disuccinic acid (EDDS) by high performance liquid chromatography after derivatization with FMOC. J. Chromatogr., A 2005, 1077, 37–43. Gustafsson, J. P. Visual MINTEQ Version 2.53;; Department of Land and Water Resources Engineering, Royal Institute of Technology: Stockholm, Sweden, 2008. Allison J. D.; Brown D. S; Novo-Gradac K. J. MINTEQA2/ PRODEFA2, A Geochemical Assessment Model for Environmental Systems, Version 3.0; Environmental Research Laboratory, Office of Research and Development, U.S. Environmental Protection Agency: Athens, GA, 1991. Bargar, J. R.; Persson, P.; Brown, G. E., Jr. Outer-sphere adsorption of Pb(II)EDTA on goethite. Geochim. Cosmochim. Acta 1999, 63, 2957–2969. Burnett, M. G.; Hardacre, C.; Mallon, J. M.; Mawhinney, H. J.; Ormerod, R. M. The enhanced adsorption of cadmium on hydrous aluminium(III) hydroxide by ethylenediaminetetraacetate. Phys. Chem. Chem. Phys. 2000, 2, 1273–1279. Guclu, K.; Apak, R. Modeling of copper(II), cadmium(II), and lead(II) adsorption on red mud from metal-EDTA mixture solutions. J. Colloid Interface Sci. 2000, 228, 238–252. Guclu, K.; Apak, R. Modeling the adsorption of free and heavy metal complex-bound EDTA onto red mud by a nonelectrostatic surface complexation model. J. Colloid Interface Sci. 2003, 260, 280–290. Fitts, J. P.; Persson, P.; Brown, G. E., Jr.; Parks, G. A. Structure and bonding of Cu(II)-glutamate complexes at the γ-Al2O3-water interface. J. Colloid Interface Sci. 1999, 220, 133–147. Ballesteros, M. C.; Rueda, E. H.; Blesa, M. A. The influence of iron (II) and (III) on the kinetics of goethite dissolution by EDTA. J. Colloid Interface Sci. 1998, 201, 13–19. Alacio, T. E.; Hesterberg, D.; Chou, J. W.; Martin, J. D.; Beauchemin, S.; Sayers, D. E. Molecular scale characteristics of Cu(II) bonding in goethite-humate complexes. Geochim. Cosmochim. Acta 2001, 65, 1355–1366. Sposito, G. Chemical Equilibria and Kinetics in Soils; Oxford University Press: New York, 1994. Fasfous, I. I.; Yapici, T.; Murimboh, J.; Hassan, N. M.; Chakrabarti, C. L.; Back, M. H.; Lean, D. R. S.; Gregoire, D. C. Kinetics of trace metal competition in the freshwater environment: some fundamental characteristics. Environ. Sci. Technol. 2004, 38, 4979–4986. Furukawa, K.; Takahashi, Y.; Sato, H. Effect of the formation of EDTA complexes on the diffusion of metal ions in water. Geochim. Cosmochim. Acta 2007, 71, 4416–4424. Martell, A. E.; Smith, R. M.; Motekaitis, R. J. NIST Critically Selected Stability Constants of Metal Complexes, Version 8.0; National Institute of Standards and Technology: Gaithersburg, MD, 2004. Lu, X.; Chen, Z.; Yang, X. Spectroscopic study of aluminium speciation in removing humic substances by Al coagulation. Water Res. 1999, 33, 3271–3280. Martell, A. E.; Hancock, R. D. Metal Complexes in Aqueous Solutions; Plenum Press: New York, 1996. Kent, D. B.; Davis, J. A.; Joye, J. L.; Curtis, G. P. Influence of variable chemical conditions on EDTA-enhanced transport of metal ions in mildly acidic groundwater. Environ. Pollut. 2008, 153, 44–52. Jardine, P. M.; Taylor, D. L. Fate and transport of ethylenediaminetetraacetate chelated contaminants in subsurface environments. Geoderma 1995, 67, 125–140. Mayes, M. A.; Jardine, P. M.; Larsen, I. L.; Brooks, S. C.; Fendorf, S. E. Multispecies transport of metal-EDTA complexes and chromate through undisturbed columns of weathered fractured saprolite. J. Contam. Hydrol. 2000, 45, 243–265. McBride, M. B. Environmental Chemistry of Soils; Oxford University Press: New York, 1994.
ES8020292
