The effect of pH on the complexation of Cd, Ni and Zn by dissolved organic carbon from leachate-polluted groundwater

PII: S0043-1354(00)00127-5

Wat. Res. Vol. 34, No. 15, pp. 3743±3754, 2000 7 2000 Elsevier Science Ltd. All rights reserved Printed in Great Britain 0043-1354/00/$ - see front matter

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THE EFFECT OF pH ON THE COMPLEXATION OF Cd, Ni AND Zn BY DISSOLVED ORGANIC CARBON FROM LEACHATE-POLLUTED GROUNDWATER JETTE B. CHRISTENSEN and THOMAS H. CHRISTENSEN* Department of Environmental Science and Engineering/Groundwater Research Centre, Building 115, Technical University of Denmark, DK-2800, Lyngby, Denmark (First received 1 April 1999; accepted in revised form 1 January 2000) AbstractÐComplexation of cadmium (Cd), nickel (Ni) and zinc (Zn) by dissolved organic carbon (DOC) in leachate-polluted groundwater was measured using a resin equilibrium method. Metal-DOC complexation was measured at di€erent DOC concentrations over a range of pH values . The results were compared to simulations made by two speciation models (WHAM and MINTEQA2). Of these models, WHAM came closest to simulating the experimental observations although it systematically overestimated the pH dependence of metal-DOC complexation. Accepting a variation in the free metal ion activity of a factor of 3±4 the WHAM model provided useful predictions of the complexation of Cd and Zn by DOC in the pH range 5±8, and of Ni in the pH range 5±7. At pH 8, however, the model overestimates the extent of Ni-DOC complexation to an unacceptable degree. The MINTEQA2 model predicts virtually no pH dependence for DOC complexation of Cd, Ni and Zn and is thus in very poor agreement with the experimental results. As an alternative approach, relations between the conditional complex formation constant (log Kc) and pH were estimated for each metal. Using these relations for estimating the complexation of Cd, Ni and Zn by DOC a deviation in the free metal ion activity up to a factor of 2 can be expected. 7 2000 Elsevier Science Ltd. All rights reserved Key wordsÐcomplexation, cadmium, nickel, zinc, DOC, leachate-polluted groundwater, WHAM, MINTEQA2, conditional complex formation constant

INTRODUCTION

The speciation and mobility of heavy metals in groundwater may be signi®cantly a€ected by the presence of dissolved organic carbon (DOC) due to the formation of dissolved complexes between DOC and the heavy metals. In particular in leachate-polluted groundwater, the concentrations of DOC can be high (up to 500 mg C lÿ1, Christensen et al., 1994), and signi®cant complexation has been demonstrated in leachate-polluted groundwater with respect to Cd, Ni and Zn (e.g. Christensen and Christensen, 1999) and Cu and Pb (Christensen et al., 1999). The complexing ability of DOC is primarily related to its content of functional groups (e.g. carboxylic and phenolic groups) (Livens, 1991). Because these groups exhibit acid-base behavior, the complexation capacity of the DOC with respect to metals depends on the pH of the system. These acid-base properties have been demonstrated for *Author to whom all correspondence should be addressed. Tel.: +45-4525-1603; fax: +45-4593-2850; e-mail: [email protected]

humic and fulvic acids extracted and puri®ed from groundwater (Higgo et al., 1993, Christensen et al., 1998b), but cannot technically be demonstrated for DOC in its original matrix. Two approaches have been used to quantify the metal complexation by DOC for environments with varying pH-values: (1) introduction of the proton titration characteristics of the organic carbon into computer-based speciation models, thereby predicting the amount of metal complexation sites available as a function of pH, and (2) the use of conditional complex formation constants regressed against pH for a range of experiments bracketing the pH-values of interest. Speciation models, like WHAM and MINTEQA2, have incorporated parameters accounting for the acid-base properties of the organic ligand. The ligand properties are derived from proton titrations of isolated and puri®ed organic matter, assuming that their acid-base properties resemble the acid-base properties of the DOC in its original matrix. The metal binding constants are intrinsic and constant for all pH-values. In the MINTEQA2 model the pH-e€ect on complexation is dealt with solely in terms of the number of sites available on the ligand, in competition with H+, at any pH-

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Jette B. Christensen and Thomas H. Christensen

value. In the WHAM model the electrostatic e€ect arising from variations in pH is taken into account by a correction factor for calculating the speci®c cation binding and by consideration of non-speci®c binding by accumulation of counter ions. Conditional complex formation constants over a range of pH-values have been experimentally determined for several metals: Cd: Saar and Weber (1979), John et al. (1988), Ephraim and Xu (1989) and Pettersson et al. (1993); Zn: Ephraim (1992a) and Pettersson et al. (1993); Cu: Ephraim and Allard (1994a); Ca: Ephraim and Allard (1994b); and Fe: (Ephraim, 1992b). These experiments were conducted using di€erent types of dissolved organic matter and di€erent methods for determination of the log Kc. The conditional nature of both constants and derived regression equations means that the approach has limited value outside the dataset from which they were obtained. These approaches may be reasonable depending on their purpose, but for both approaches very few comparisons exist with independent measurements representing DOC in its original matrix at a range of pH-values. This lack of validation limits our knowledge about how much DOC-complexation of heavy metals varies with pH in real samples and how well we can predict this variation from the existing knowledge. Christensen and Christensen (1999) and Christensen et al. (1999) showed that existing models correctly predicted increasing complexation with increasing ligand concentration at a single pH-value, but that signi®cant deviations were seen for speci®c metals and models. The pH dependence, however, of DOC-metal complexing in natural groundwater or analogues thereof remains an area that has received very little attention. The purpose of this paper is to provide laboratory measurements of the complexation of Cd, Ni and Zn as a function of DOC concentrations in leachate-polluted groundwater for a range of pHvalues and to compare these measurements with predictions on the basis of existing pH-dependent conditional complex formation constants and of existing speciation models WHAM and MINTEAQ2. MATERIAL AND METHODS

Approach of the resin equilibration method The method makes use of the fact that DOC-complexed metal does not sorb onto a cation-exchange resin. This issue was addressed by Christensen and Christensen (1999) who showed that the concentration of DOC in the experiments was slightly higher after exposure to the resin. The increase in DOC content was shown not to in¯uence the distribution of Cd. Thus, the e€ect of metal-DOC complexation ((M+MDOC)/M; where MDOC are metalDOC complexes and M is dissolved metal including M2+ and any inorganic metal complexes) can be determined by comparing the metal distribution in two solutions equilibrated with a resin: A sample solution that contains DOC in its original matrix, and a reference solution that is

DOC-free but is otherwise identical to the DOC-containing sample. By performing the resin equilibration experiments with DOC-containing groundwater mixed in various ratios with the reference solution, the presence of DOC-complexed metal can be determined for a range of DOC concentrations. By performing the determinations at a range of ®xed pH-values, the pH dependence of complexation can be investigated. Leachate-polluted groundwater samples Two leachate-polluted groundwater samples (S1 and S2) were taken from the anaerobic part of the leachate pollution plume at the Vejen Land®ll, Denmark. The samples were obtained from the methanogenic zone of the plume 7 m downgradient of the land®ll and approximately 2 m (S1) and 5 m (S2) below the groundwater table. Sampling procedures are described in detail elsewhere (Christensen et al., 1996). Previously, the wells were sampled and a characterization of the dissolved organic matter was performed in terms of humic acid, fulvic acid and a hydrophilic fraction (Christensen et al., 1998a). The samples were aerated for one week to remove traces of volatile organic carbon, inorganic carbon and any remaining sul®de, ®ltered to remove particulate matter (>0.45 mm, cellulose nitrate ®lter from Sartorius, GoÈttingen, Germany) and pH adjusted with concentrated HNO3 to the natural pH (6.60 2 0.03). The samples were stored in the dark at 48C. Characteristics of the aerated and ®ltered leachate-polluted-groundwater samples are shown in Table 1. Synthetic, inorganic solutions (reference samples) In order to be able to determine the sorption of heavy metals onto the resin in the absence of DOC, synthetic inorganic solutions (R1 and R2) were made up of Ca2+, Mg2+, Na+, K+ and NH‡ 4 salts of chloride, sulfate and nitrate to mimic the composition of the leachate-polluted groundwater samples with respect to cations, chloride, sulfate and ionic strength. Speciation calculations concerning the inorganic elements in the leachate-polluted groundwater showed that chloride (Clÿ) and carbonate …CO32ÿ † would form inorganic complexes with Cd and Ni, respectively. Clÿ complexes constituted up to 45% of Cd not bound in DOC complexes whereas NiCO3 complexes constituted a signi®cant fraction (more than 5% of the Ni not bound in DOC complexes) only at high pH-values (e.g. pH 8). Because of the presence of these inorganic complexes, it was important to ensure that the Clÿ and the CO32ÿ concentrations of the references were identical with the concentrations in the leachate-polluted groundwater. The pretreated original samples contained (prior to exposure to resin) low concentrations of Al, Fe and Mn, but no attempts were made to mimic these metals in the reference solution since the species of these metals were not known. Furthermore it would not be possible to prevent Al, Fe and Mn from precipitation in an aerobic reference solution at pH 6.6. As reported by Christensen and Christensen (1999), model simulations involving these metals showed that the neglect of these metals only marginally a€ected the model estimates of the complexation of Cd, Ni and Zn. Resin preparation 1

Serdolit CS 1C cation exchange resin (polystyrene with sulfonic acid groups, 200±400 mesh, Na+ form) was available from Serva (Heidelberg, Germany). The resin was converted from the Na+-form by exposure to the synthetic inorganic solutions (R1 and R2, respectively). The procedure used is described in detail by Christensen and Christensen (1999).

E€ects of pH on the complexation of Cd, Ni and Zn by dissolved organic carbon Resin equilibration experiments The resin equilibration experiments were carried out at room temperature (approximately 208C) in 50 ml polyethylene bottles containing 25 ml of solution and 150 mg of the corresponding resin (R1 or R2 saturated). One metal was added to each bottle at a predetermined concentration. The bottles were equilibrated for 7 days (the equilibration time was based on preliminary tests). During equilibration, pH was adjusted by addition of small amounts of HNO3 or NaOH to maintain constant and identical pH in the experiments at predetermined values in the range pH 5±8. After equilibration, solutions were separated from the resin by centrifugation (3 min at 3000 rpm) and acidi®ed to pH < 1.5 (conc HNO3). The resin equilibrium experiments involved 4±7 series. Each series was conducted at a speci®c pH value in the range pH 5±8 and consisted of 7±8 bottles containing di€erent DOC concentrations obtained by mixing leachate-polluted groundwater and reference solution in di€erent ratios. Identical amounts of metal were added to all bottles. The equilibrium solution concentrations in the experiments ranged as follows: Cd: 4.5  10ÿ8Ð11  10ÿ8 M, Ni: 1.7  10ÿ6 Ð4.2  10ÿ6 M and Zn: 3.1  10ÿ6Ð7.6  10ÿ6 M. The concentration levels were chosen to match concentration levels that could be encountered in polluted aquifers. The distribution of heavy metal between the solution and the resin in the absence of DOC (Kd,R) was determined in triplicate and the mean value used in the calculations. The relative standard deviation on the Kd,R values was in the range of 1±8%. The distributions of heavy metals in the presence of DOC were based on single values (one for each DOC-concentration). Instrumental analyses Heavy metals in solution were determined by graphite furnace atomic absorption spectrophotometry (PerkinElmer 5000, deuterium background correction, HGA 400 graphite furnace, AS-1 automatic sample injection system) after solvent extraction by 1.0% Na diethyldithiocarbamate trihydrate in 4-methylpentan-2-one. All samples were acidi®ed to pH < 1:5 (conc HNO3) before solvent extraction. Ca, Mg, K, Na, Fe, Mn, Al were determined by ¯ame atomic absorption spectrophotometry (Perkin-Elmer by the 2standard autoana370). Cl and NH4 were analyzed 2 lyser routine (Technicon Autoanalyser II). DOC analysis was performed with a TOC-analyser (O-I-Analytical

3745

Model 700). pH was measured by a pH Meter (Hanna 1 pH elecInstruments DP 7916R ) using a combination 1 trode (Radiometer pH 106007-3,5-NC ). General laboratory practice All chemicals used were analytical grade (Merck, pro analysis). All plastic and glassware were cleaned and soaked in 2 M HNO3, for at least 12 h, then rinsed with deionized distilled water and dried at 608C in a convection oven. Data treatment For each equilibration experiment the ratio between metal on the resin and in solution was expressed as a distribution coecient, Kd. The distribution coecient is considered to be constant within the metal concentration range used (Holm et al., 1995). The DOC-complexed metal does not associate with the resin (shown for DOCCd complexes by Christensen and Christensen, 1999) and the DOC-containing samples were identical to the reference samples except for the content of DOC. For these reasons the di€erence in distribution coecient can be related to the e€ect of DOC complexation at a ®xed concentration of metal on the resin as shown in Fig. 1. The e€ect of metal±DOC complexation calculated from the metal distribution measured after equilibration with the resin is approximately equal to the e€ect of metal± DOC complexation in the original sample, because the e€ect of complexation is independent of the metal concentration when the metal concentration is much smaller than the DOC concentration (Christensen and Lun, 1989). These conditions are easily met in the present study because the total dissolved metal concentrations were within the range 10ÿ7 to 10ÿ6 M while the DOC concentration was between 10ÿ4 to 10ÿ3 moles of sites lÿ1. The estimated molar site concentrations were based on characteristics of the DOC samples as described by Christensen et al. (1998a).

MODELS

Basic concept of WHAM The WHAM model (Tipping, 1994) is a combination of several sub-models. These include models

Table 1. Characteristic of the aerated leachate-polluted groundwater samples (S1 and S2)

pH Speci®c conducivity (mS cmÿ1) Ionic strengtha DOC (mg C lÿ1) Average molecular weight of DOCb (g moleÿ1) TALc (meq lÿ1) Ca2+ (mmole lÿ1) Mg2+ (mmole lÿ1) Mn2+ (mmole lÿ1) Fe2+ (mmole lÿ1) Al3+ (mmole lÿ1) Na+ (mmole lÿ1) K+ (mmole lÿ1) ÿ1 NH+ 4 (mmole l ) ÿ1 SO2ÿ 4 (mmole l ) Clÿ (mmole lÿ1) NO3ÿ (mmole lÿ1) a

Leachate-polluted groundwater S1

Leachate-polluted groundwater S2

6.60 4.95 0.056 187 2085 0.02 3.07 1.92 0.034 0.016 0.075 14.30 4.94 13.50 b.d.d 18.62 30.26

6.60 2.08 0.023 79 1942 2.02 2.56 1.10 0.016 0.014 0.010 5.09 1.95 2.50 0.52 4.51 11.42

Estimated by speciation program MINTEQA2 (Allison et al., 1991). Obtained by size exclusion chromatography (Christensen et al., 1998a). Calculated total alkalinity related to the content of HCO3ÿ and CO32ÿ : The calculated alkalinity of the samples is very low due to the aeration of the samples. d Below detection limit 0.02 mmole lÿ1. b c

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Jette B. Christensen and Thomas H. Christensen

for inorganic solution speciation and a humic metal ion binding model called Model V as described in detail by Tipping (1993) and Tipping and Hurley (1992). Model V describes the binding of metal ions to humic substances by a discrete binding sites model in which binding is modi®ed by electrostatic interactions. There is an empirical relation between net humic charge and an electrostatic interaction factor. The model also takes into account the accumulation of counter ions in the di€use layer using a Donnan-type expression. The discrete binding sites are represented by two types of site (Type A and B) and within each type of site there are four di€erent sites present in equal amounts. The two types of site are described by intrinsic proton binding constants ( pKA and pKB) and spreads of the values (DpKA and DpKB) within each type of site. There are nA (meq kgÿ1) type-A sites (associated with carboxylic-type groups) and nB=nA/2 (meq kgÿ1) type-B of sites (often associated with phenolic-type groups). Metal binding occurs at single proton binding sites or by bidentate complexation between pairs of sites depending on a proximity factor that de®nes whether pairs of proton binding groups are close enough to form bidentate complexes. The two types of sites (A and B) have separate intrinsic exchange constants ( pKMHA and pKMHB). It is implicit in Model V that the variation in the binding anities of the proton and metal ions on monodentate sites are perfectly correlated, e.g., a high proton anity site also has a high anity for all metal ions. However, the introduction of bidentate sites reduces the

correlation between the proton anity distribution and the metal ion anity distribution. The default model parameters contained in the database originate from published data describing proton and metal binding to isolated humic and fulvic acids. The parameter values in the default database are ``best average'' parameters determined on a basis of 1±8 data sets (Tipping, 1993, Tipping and Hurley, 1992). Basic concepts of MINTEQA2 MINTEQA2 (version 3.11) includes a sub-model for estimations of the complexation of metals with dissolved organic matter. This is a composite ligand model with a Gaussian anity distribution (Dobbs et al., 1989, Susetyo et al., 1991 and Allison and Perdue, 1994). The model assumes that the composite ligand consists of a population of discrete binding sites in which the probability of occurrence of a binding site is normally distributed with respect to its log K value for proton or metal binding. The non-speci®c binding of cations due to electrostatic interactions is not taken into account in this model. The binding sites in MINTEQA2 are represented only by one type of site (``carboxylic'') characterized by a mean binding constant (m ) and a spread (s ) of the log K value around this mean value. In MINTEQA2, it is assumed that only monodentate binding occurs and that the ratio of metal to proton binding constants is the same for all sites in the distribution. This means that the standard deviation (s ) is the same for binding protons as for metal ions. The database available for proton and metal interaction is based on results reported by Susetyo et al. (1991) and includes only data for binding to the carboxylic sites. The humic substance used by Susetyo et al. (1991) was dissolved organic carbon from the Suwannee River sampled by reverse osmosis (Serkiz and Perdue, 1990). Describing the composite ligand model in MINTEQA2, Allison and Perdue (1994) pointed out that the model has not yet been satisfactorily veri®ed. They also suggested that including a second type of binding site would improve the model considerably. Binding properties

Fig. 1. Illustration of the approach used to calculate the e€ect of complexation from the measured distribution coecients (Kd).

The proton binding properties of isolated and puri®ed fulvic acids from the land®ll-leachate-polluted groundwater samples (Christensen et al., 1998b) agreed well with the default proton binding parameters contained in the WHAM model database and consequently these default parameters have been adopted in the present study. Furthermore Christensen and Christensen (1999) showed that the default metal binding parameters for Cd and Ni in the WHAM model database can be usefully applied to modelling Cd and Ni binding by DOC and these parameters are also adopted in the

E€ects of pH on the complexation of Cd, Ni and Zn by dissolved organic carbon

present study. The default values for binding of Zn, however, were considered to be too large and an improved binding constant was suggested (Table 2). For the MINTEQA2 model it was recognized by Allison and Perdue (1994) that the DOC molar site concentration of 1 meq gÿ1 was far too low. Thus the molar site concentration used in modelling the experimental data by MINTEQA2 was based on results from a simple titration of the fulvic, humic and hydrophilic fractions of the leachate-polluted groundwater as reported by Christensen et al. (1998a). The molar site concentrations were estimated as weighted average values taking into account the weight fraction of humic acid, fulvic acid and the hydrophilic fraction (Table 2). The total acidity measured in the three fractions was used. The phenolic sites were assumed to behave the same as carboxylic sites with respect to metal binding. This was necessary as no database for metal binding to phenolic sites exists in MINTEQA2. Default metal binding parameters in the MINTEQA2 model were used for modelling Cd, Ni and Zn binding by DOC in leachate-polluted groundwater, except for Ni in sample S1where an adjusted value was used according to Christensen and Christensen (1999) (Table 2). Modelling experimental data A characterization of the DOC performed by Christensen et al. (1998a) showed that DOC in the two leachate-polluted groundwater samples consisted of about 8% humic acids, 48% fulvic acids and 24% hydrophilic fraction. About 20% of the DOC was lost during the isolation. The hydrophilic fraction is usually also lost in the traditional puri®cation of the fulvic and humic acid fractions and only recently has the hydrophilic fraction gained attention (e.g. Aiken et al., 1992, Peuravuori and Pihlaja, 1998). A previous characterization of the DOC in terms of its acid dissociation properties and elemental composition indicated that the humic acids as well as the hydrophilic fraction in leachate polluted groundwater had characteristics resembling

3747

fulvic acids (Christensen et al., 1998a). Thus, in this study it was assumed that the humic acids and the hydrophilic fraction of the DOC samples had proton and metal binding properties similar to the fulvic acid fraction. In the WHAM model 100% of the DOC was entered as fulvic acids. The fulvic acid concentrations were entered as mg C lÿ1. Proton and metal binding constants in MINTEQA2 were derived from unfractionated DOC, and the DOC concentrations are given as molar site concentration (eq lÿ1) using the molar site concentrations (meq gÿ1) reported in Table 2 and assuming 50% carbon content. The model speciation calculations were performed at 208C. RESULTS AND DISCUSSION

The complex formation of Cd, Ni and Zn by DOC in leachate-polluted groundwater is shown in Fig. 2 at di€erent pH values expressed as the e€ect of complexation (M+MDOC)/M; where MDOC is metal complexed by DOC and M is the concentration of metal in solution not bound in DOCcomplexes (M thus includes inorganic metal complexes in addition to the free metal ion). In the absence of metal-DOC complexes the e€ect of complexation, (M+MDOC)/M, will be unity. Figure 2 shows increasing complex formation at increasing pH values for all combinations of metals and leachate-polluted groundwater samples, varying for all three metals between 1.9±16.6 in sample S1 (180 mg C lÿ1) and 1.4±4.3 in sample S2 (80 mg C lÿ1). For all three metals, the e€ect of complexation is slightly higher in S1 than in S2 at identical DOC concentration. This might be due to variations in the complexing ability of the DOC and the fact that the acid capacity (number of sites) of the DOC in S1 is higher than of the DOC in S2 (Table 2). Comparison of experimental results and WHAM model predictions The ability of the WHAM model to predict the complexation of Cd, Ni and Zn as a function of

Table 2. Default and improved metal binding parameters for the WHAM model and MINTEQA2 model. Improved metal binding parameters were estimated for sample S1 and S2 by Christensen and Christensen (1999) S1

S2

WHAM pKMHA (Cd) pKMHA (Ni) pKMHA (Zn)

±a ±a 1.7b

±a ±a 1.7b

``best average'' 1.5 1.4 1.3

MINTEQA2 Molar site concentration (meq gÿ1)b Mean log KCd (m ) Mean log KNi (m ) Mean log KZn (m )

7.44c ±a 3.9 ±a

6.92c ±a ±a ±a

Suwannee River DOMd 1.00 3.3 3.3 3.5

a

The default metal binding parameter was used in the model predictions. Improved metal binding constants from Christensen and Christensen (1999). Speci®c molar site concentrations obtained on similar samples by Christensen et al. (1998a). d Dissolved organic matter isolated from Suwannee River. b c

Default

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Jette B. Christensen and Thomas H. Christensen

DOC-concentration at di€erent pH values for S1 and S2 is shown in Fig. 3. Of primary concern in environmental studies is the free metal ion activity in natural waters and thus the usefulness of these models should be evaluated in terms of their ability to predict this parameter. The WHAM model generally overestimated the pH dependence of metal±DOC complexation relative to the experimental results. However, accounting for metal±DOC complexation, the model was able to reproduce the free metal ion activity of Cd, Ni, and Zn in the pH range 5±7 within a factor of 3. The deviations of the free metal ion activity were

highest for Ni and Zn in sample S2 at pH 7. At pH 8 the WHAM model produced good predictions of Cd and Zn binding by DOC in sample S1. However, larger di€erences between observed and predicted free metal ion activities were seen for Cd and Zn in sample S2 at pH 8 (a di€erence of 3.5) and for Ni in sample S1 at pH 8 (a di€erence of 5±7). Comparison of experimental results and MINTEQA2 model predictions The MINTEQA2 model predictions of the complexation of Cd, Ni and Zn by DOC in leachatepolluted groundwater are shown in Fig. 4. Clearly,

Fig. 2. Experimental results expressed as (M+MDOC)/M for Cd, Ni and Zn (log scale) vs DOC concentrations at pH values ranging from 5±8. M is the concentration of metals not bound in DOC complexes and MDOC is the concentration of metals bound in DOC complexes.

E€ects of pH on the complexation of Cd, Ni and Zn by dissolved organic carbon

the MINTEQA2 model predicts only very weak pH dependence of metal-DOC complexation. The fact that MINTEQA2 fails to reproduce the signi®cant pH e€ect observed in the experiments presented in this paper is closely related to the high content of cations (e.g. Ca2+ and Mg2+) in leachate polluted groundwater. These cations strongly in¯uence the protonation of DOC in the model calculations. At pH 5 less than 5% of the DOC was protonated and at pH 6 no more protonated DOC occurred in the model predictions. Thus, the model will only show

3749

very little e€ect on the metal-DOC complexation of varying pH in the range of 5±8 and most of the variations can be explained by complexation of metals by inorganic ligands at di€erent pH values. The high content of cations did not saturate the DOC as more than 60% of the DOC was presented as free DOC according to the model predictions. The MINTEQA2 model predictions of the pHe€ect would probably improve considerably by including a second type of binding site with a pKavalue at 8±10 (``phenolic type sites'').

Fig. 3. Comparison of experimental results and WHAM model predictions of (M+MDOC)/M for Cd, Ni and Zn (log scale) as a function of DOC at varying pH-values. M is the concentration of metals not bound in DOC complexes and MDOC is the concentration of metals bound in DOC complexes.

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Jette B. Christensen and Thomas H. Christensen

Estimation of the conditional complex formation constant (log Kc) as a function of pH As an alternative to the computer speciation models, conditional complex formation constant can be expressed as a function of the pH of the experimental determinations. The log Kc for each metal and pH level has been determined by the equations given in Appendix A. The stoichiometric coecient i (moles of ligands binding one mole of metal) varied in the range of 0.80±1.60, showing values below as well as above unity within each series of metal and leachate-polluted groundwater.

Thus, in the estimation of the complex formation constant, i is assumed to be equal to one. Figure 5 shows the log Kc for Cd, Ni and Zn in sample S1 and S2 estimated for each DOC concentration level and each pH value. Average values and standard deviations of the log Kc at each pHvalue are given in Table 3. Figure 5 shows clearly a linear relation between log Kc and pH. The equations of the regression lines are given in Table 4 for each metal in both leachate-polluted groundwater samples. The leachate-polluted groundwater samples S1

Fig. 4. Comparison of experimental results and MINTEQA2 model predictions of (M+MDOC)/M for Cd, Ni and Zn (log scale) as a function of DOC at varying pH-values. M is the concentration of metals not bound in DOC complexes and MDOC is the concentration of metals bound in DOC complexes.

E€ects of pH on the complexation of Cd, Ni and Zn by dissolved organic carbon

and S2 showed di€erent complexing properties mainly related to the di€erences in the DOC (acid capacity), di€erences in ionic strength and ion composition of the matrix. However, if a deviation on the free metal ion activity of a factor of 2 is acceptable, the complexation of Cd, Ni and Zn by DOC in leachate-polluted groundwater can be described by one relation between log Kc and pH for each metal (Table 4), assuming that ionic strength and cation composition are within a range not signi®cantly exceeding the leachate-polluted groundwater samples used in these experiments. Figure 6 compares the log Kc obtained in this

Fig. 5. Complex formation constants for the complexation of Cd, Ni and Zn by DOC in leachate-polluted groundwater as a function of pH.

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work at varying pH values to data in the literature. In order to be able to compare the results obtained by Ephraim and Xu (1989), Ephraim (1992a) and Pettersson et al. (1993) to the log Kc values determined in this work, complex formation constants obtained by these authors were corrected by a pH dependent factor according to Ephraim et al. (1989). For Cd the level of the log Kc values determined by ion exchange method by Ephraim and Xu

Fig. 6. Comparison of conditional complex formation constants determined in these experiments and reported in the literature for Cd and Zn at varying pH values. No values for Ni were found in the literature. (a)±(c) John et al. (1988) determined the conditional stability constants for di€erent concentrations of humic substances; (a) 5  10ÿ6 M, (b) 1.25  10ÿ5 M, (c) 2.5  10ÿ5 M (d)±(f) Ephraim and Xu (1989) used three di€erent methods for determination of the conditional stability constants, (d) ion exchange method, (e) ultra ®ltration, (f) ion selective electrode.

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Jette B. Christensen and Thomas H. Christensen

Table 3. Conditional complex formation constants (log Kc) for Cd, Ni and Zn in two leachate-polluted groundwater samples (S1 and S2). The values are average values of the constants determined at each pH for a range of DOC concentrations as can be seen in Fig. 5 S1

Cd pH pH pH pH Ni pH pH pH pH pH pH pH Zn pH pH pH pH pH pH pH

S2

log Kc (mole lÿ1)

log Kc (gC lÿ1)

Standard deviation (mole lÿ1)

log Kc (mole lÿ1)

log Kc (gC lÿ1)

Standard deviation (mole lÿ1)

5 6 7 8

4.28 4.40 4.75 5.10

1.27 1.48 1.73 2.08

0.033 0.074 0.036 0.053

± 4.43 4.58 4.78

± 1.41 1.56 1.76

± 0.066 0.027 0.024

5.0 5.5 6.0 6.5 7.0 7.5 8.0

4.35 4.48 4.59 4.72 4.82 4.91 5.14

1.33 1.46 1.57 1.70 1.80 1.89 2.13

0.048 0.068 0.029 0.012 0.036 0.055 0.034

± 4.18 4.33 4.49 4.64 4.75 4.96

± 1.05 1.31 1.47 1.62 1.73 1.94

± 0.034 0.043 0.027 0.031 0.020 0.033

5.0 5.5 6.0 6.5 7.0 7.5 8.0

4.03 4.18 4.26 4.42 4.59 4.76 5.08

1.01 1.13 1.24 1.40 1.57 1.74 2.06

0.074 0.069 0.056 0.067 0.090 0.077 0.187

± 4.12 4.13 4.26 4.37 4.54 4.65

± 0.98 1.12 1.24 1.36 1.52 1.63

± 0.071 0.030 0.035 0.037 0.131 0.144

(1989) correspond to the observations reported in this work, whereas the log Kc values obtained from Saar and Weber (1979) and Ephraim and Xu (1989) by ultra ®ltration and ion selective electrode are almost one log unit below the observations for Cd in sample S1 and S2. The observations obtained by Pettersson et al. (1993) are at low pH values below the observations from this work but at higher pH the values are at the same level. For Zn the relation between log Kc and pH obtained by Ephraim and Xu (1989) is several log units higher than observed for Zn in sample S1 and S2 whereas the observations obtained by Pettersson et al. (1993) are, with the exception of two data points, at least one log unit lower than observed for Zn in sample S1 and S2. The literature data on conditional complex formation constants determined at di€erent pH-values vary by several log units for both Cd and Zn. However, the existing data are not suciently consistent to determine the reason of this: some variation may be due to experimental methods as shown by Ephraim and Xu (1989) and to di€erences in preparing the DOC. CONCLUSIONS

In this work it was shown that the complex for-

mation of Cd, Ni and Zn by DOC from land®ll-leachate-polluted groundwater kept in the original matrix increased with increasing pH. The experimental results were compared to model prediction using the WHAM model and the MINTEQA2 model. Accepting a deviation in the free metal ion activity of a factor of 3±4 the WHAM model was able to provide good estimates of the extent of complexation of Cd and Zn by DOC in leachate-polluted groundwater in a pH range of 5±8 and of Ni in the pH range of 5±7. The MINTEQA2 model estimated only very little e€ect of pH and thus failed to reproduce the experimental data. This failure is related to the high content of cations in the leachate-polluted groundwater. Although the cations did not saturate the DOC the high content of cations strongly in¯uenced the protonation of the DOC and at pH 5, MINTEQA2 predicted that less than 5% of the DOC was protonated. Thus, the WHAM model used with its default database (binding constants for Zn adjusted) was found to be the most appropriate speciation model at the moment in order to estimate the complexation of Cd, Ni and Zn by DOC in leachate-polluted groundwater at varying pH values. It was shown that a linear relation exists between the conditional complex formation constant and pH yielding a set of regression equations to describe

Table 4. Relations between the complex formation constant (log Kc, moleÿ1) and pH for Cd, Ni and Zn in two leachate-polluted groundwater samples (S1 and S2) S1 Cd Ni Zn

S2 2

log Kc=0.272.pH+2.887 R =0.962 log Kc=0.245.pH+3.120 R 2=0.963 log Kc=0.330.pH+2.316 R 2=0.912

S1+S2 2

log Kc=0.160.pH+3.484 R =0.952 log Kc=0.330.pH+2.317 R 2=0.978 log Kc=0.221.pH+2.847 R 2=0.912

log Kc=0.221.pH+3.157 R 2=0.845 log Kc=0.288.pH+2.719 R 2=0.809 log Kc=0.276.pH+2.581 R 2=0.799

E€ects of pH on the complexation of Cd, Ni and Zn by dissolved organic carbon

this dependence. These relations are based on the two leachate-polluted groundwater samples used in this work and the relations were able to describe the experimental results obtained for these samples within a factor of 2 on the free metal ion activity. However, conditional complex formation constants reported in the literature vary highly and use of conditional complex formation constants for environmental conditions di€erent from those used for their measurements cannot be recommended. AcknowledgementsÐWe would like to thank Torben Dolin for preparing the ®gures, Jerry Allison for useful discussions and Thomas Astrup for technical assistance.

REFERENCES

Aiken G. R., McKnight D. M., Thorn K. A. and Thurman E. M. (1992) Isolation of hydrophilic organic acids from water using nonionic macroporous resins. Org. Geochem. 18, 567±573. Allison J. D., Brown D. S. and Novo-Gradac K. J. (1991) MINTEQA2/PRODEFA2, A Geochemical Assessment Model for Environmental Systems: Version 3.0 Users Manual. US. Environmental Protection Agency, Athens, Georgia, EPA/600/3±91/021. Allison J. D. and Perdue E. M. (1994) Modelling metal± humic interaction with MinteqA2. In Humic Substances in the Global Environment and Implications on Human Health, eds N. Senesi and T. M. Miano, pp. 927±942. Elsevier, Amsterdam. Christensen J. B., Botma J. J. and Christensen T. H. (1999) Complexation of Cu and Pb by DOC in polluted groundwater: a comparison of experimental data and predictions by computer speciation models (WHAM and MINTEQA2). Wat. Res. 33, 3231±3238. Christensen J. B. and Christensen T. H. (1999) Complexation of Cd, Ni and Zn by DOC from polluted groundwater: a comparison of approaches using resin exchange, aquifer material sorption, and computer speciation models (WHAM and MINTEQA2). Environ. Sci. Technol. 33, 3231±3238. Christensen J. B., Jensen D. L. and Christensen T. H. (1996) E€ect of dissolved organic carbon on the mobility of cadmium, nickel and zinc in leachate-polluted groundwater. Wat. Res. 30, 3037±3049. Christensen J. B., Jensen D. L., Filip Z., Grùn C. and Christensen T. H. (1998a) Characterization of the dissolved organic carbon in land®ll polluted groundwater. Wat. Res. 32, 125±135. Christensen J. B., Tipping E., Kinniburgh D. G., Grùn C. and Christensen T. H. (1998b) Proton binding by groundwater fulvic acids of di€erent age, origin and structural characteristics modeled by the Model V and Nica-Donnan model. Environ. Sci. Technol. 32, 3346± 3355. Christensen T. H., Kjeldsen P., Albrechtsen H.-J., Heron G., Nielsen P. H., Bjerg P. and Holm P. E. (1994) Attenuation of land®ll leachate pollutants in aquifers. Crit. Rev. Environ. Sci. & Technol. 24, 119±202. Christensen T. H. and Lun X. Z. (1989) A method for determination of cadmium species in solid waste leachates. Wat. Res. 23, 73±80. Dobbs J. C., Susetyo L. A., Knight F. E., Carreira L. A. and Azarraga V. (1989) Characterization of metal binding sites in fulvic acids by lanthanide ion probe spectroscopy. Anal. Chem. 61, 483±488. Ephraim J. H. (1992a) Heterogeneity as a concept in the interpretation of metal ion binding by humic substances.

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The binding of zinc by an aquatic fulvic acid. Analytica Chimica Acta 267, 39±45. Ephraim J. H. (1992b) Iron interaction with a soil fulvic acid: studies via potentiometric titrations, ultre®ltration and dialysis techniques. Ghana J. Chem. 1, 300±312. Ephraim J. H. and Allard B. (1994a) Copper binding by an aquatic fulvic acid: heterogeneity considerations. Environ. Int.ÐJ. Sci., Technol., Health, Monit. Pol. 20, 89±96. Ephraim J. H. and Allard B. (1994b) Calcium binding by fulvic acids studied by an ion selective electrode and an ultra®ltration method. In Humic Substances in the Global Environment and Implications on Human Health, eds N. Senesi and T. M. Miano, pp. 1125±1130. Elsevier, Amsterdam. Ephraim J., BoreÂn H., Pettersson C., Arsenie I. and Allard B. (1989) A novel description of the acid-base properties of an aquatic fulvic acid. Environ. Sci. Tehnol. 23, 356± 362. Ephraim J. H. and Xu H. (1989) The binding of cadmium by an aquatic fulvic acid: a comparison of ultra®ltration with ion-exchange distribution and ion-selective electrode techniques. Sci. Total. Environ. 81/82, 625±634. Higgo J. J., Kinniburgh D., Smith B. and Tipping E. and Ca2+ (1993) Complexation of Co2+, Ni2+, UO2+ 2 by humic substances in groundwaters. Radiochimica Acta 61, 91±103. Holm P. E., Christensen T. H., Tjell J. C. and McGrath S. P. (1995) Speciation of dissolved cadmium: interpretation of dialysis, ion exchange and computer (GEOCHEM) methods. J. Environ. Qual. 24, 183±190. John J., Salbu B., Gjessing E. T. and Bjùrnstad H. E. (1988) E€ect of pH, humus concentration and molecular weight on conditional stability constants of cadmium. Wat. Res. 22, 1381±1388. Livens F. R. (1991) Chemical reactions of metals with humic material. Environ. Poll. 70, 183±208. Pettersson C., HaÊkansson K., Karlsson S. and Allard B. (1993) Metal speciation in a polluted humic surface water system, Bersbo, Sweden. Wat. Res. 27, 863±871. Peuravuori J. and Pihlaja K. (1998) Multi-method characterization of lake aquatic humic matter isolated with two di€erent sorbing solids. Analyt. Chim. Acta 363, 235±247. Saar R. A. and Weber J. H. (1979) Complexation of cadmium(II) with water- and soil-derived fulvic acids: e€ect of pH and fulvic acid concentration. Can. J. Chem. 57, 1263±1268. Serkiz S. M. and Perdue E. M. (1990) Isolation of dissolved organic matter from the Suwannee river using reverse osmosis. Wat. Res. 24, 911±916. Susetyo W., Carreira L. A., Azarraga L. V. and Grimm D. M. (1991) Fate, distribution and metabolism of inorganic pollutants, speciation. Fluorescence techniques for metal-humic interactions. Fresenius J. Anal. Chem. 339, 624±635. Tipping E. (1993) Modelling ion binding by humic acids. Coll Surf. 73, 117±131. Tipping E. (1994) WHAMÐa chemical equilibrium model and computer code for waters, sediments and soils incorporating a discrete site/electrostatic model of ionbinding by humic substances. Computers & Geosci. 20, 973±1023. Tipping E. and Hurley M. A. (1992) A unifying model of cation binding by humic substances. Geochim. et Cosmochim. Acta 56, 3627±3641. APPENDIX A The reaction for complex formation can be written as: M 2‡ ‡ iDOC $ MDOCi

…1†

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Jette B. Christensen and Thomas H. Christensen

where M2+ is the metal ion, DOC is dissolved organic carbon, MDOCi is the metal-DOC complex and i is a stoichiometric coecient (number of moles of DOC that combines with one mole of heavy metal). The complex formation constant is at equilibrium de®ned as: ‰MCOCi Š Kc ˆ ‰M 2‡ Š
‰DOCŠi
g

…2†

where [MDOCi] is the concentration of metal-DOC complex, [M2+] is the concentration of divalent metal ion, [DOC] is the concentration of DOC not bound in complexes and g is the activity coecient of the divalent metal ion. The activity coecient of the DOC and MDOCi-complex is conventionally assumed to be unity. When [M]T<<[DOC]T, [DOC] is approximately equal to [DOC]T. This assumption is ful®lled, since an acid/base titration of similar DOC samples revealed a complexation capacity at least exceeding the actual metal concentration by a factor of 10 (see Christensen et al., 1998a). Assuming that metal±DOC complex is the only complex present in solution the e€ect of complexation (EC) is de®ned as: ‰M 2‡ Š ‡ ‰MDOCi Š Kd, R EC ˆ ˆ Kd, DOC ‰M 2‡ Š

…EC ÿ 1† g
‰DOCŠi

EC ˆ

‰Cd 2‡ Š ‡ ‰CdDOCi Š ‡ ‰CdCl‡ Š Kd, R ˆ Kd, DOC ‰Cd 2‡ Š ‡ ‰CdCl‡ Š

…5†

Combining equations (2) and (5) allows a calculation of the log Kc for cadmium in the presence of monochloro complexes Kc ˆ

…EC ÿ 1†…1 ‡ Kcl
‰Cl ÿ Š
g† g
‰DOCŠi

…6†

At pH 8 nickel-carbonate complexes (NiCO3) becomes a signi®cant species and corrections EC ˆ

‰Ni 2‡ Š ‡ ‰NiDOCi Š ‡ ‰NiCO3 Š Kd, R ˆ Kd, DOC ‰Ni 2‡ Š ‡ ‰NiCO3 Š

…7†

…3†

Combining equations (2) and (3) allows calculation of the conditional complex formation constant from the e€ect of complexation (EC): Kc ˆ

For cadmium in leachate-polluted groundwater the monochloro complexes will constitute a signi®cant species and corrections must be made in the estimation of the Kc for Cd:

…4†

Combining equations (2) and (7) allows calculation of the Kc for nickel in the presence of carbonate complexes (relevant at pH 8): Kc ˆ

…EC ÿ 1†
…1 ‡ KCO3
‰CO32ÿ Š
g 2 † g
‰DOCŠi

…8†