Environ. Sci. Technol. 1999, 33, 3857-3863
Complexation of Cd, Ni, and Zn by DOC in Polluted Groundwater: A Comparison of Approaches Using Resin Exchange, Aquifer Material Sorption, and Computer Speciation Models (WHAM and MINTEQA2) 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
Complexation of cadmium (Cd), nickel (Ni), and zinc (Zn) by dissolved organic carbon (DOC) in leachate-polluted groundwater was measured using a resin equilibrium method and an aquifer material sorption technique. The first method is commonly used in complexation studies, while the second method better represents aquifer conditions. The two approaches gave similar results. Metal-DOC complexation was measured over a range of DOC concentrations using the resin equilibrium method, and the results were compared to simulations made by two speciation models containing default databases on metalDOC complexes (WHAM and MINTEQA2). The WHAM model gave reasonable estimates of Cd and Ni complexation by DOC for both leachate-polluted groundwater samples. The estimated effect of complexation differed less than 50% from the experimental values corresponding to a deviation on the activity of the free metal ion of a factor of 2.5. The effect of DOC complexation for Zn was largely overestimated by the WHAM model, and it was found that using a binding constant of 1.7 instead of the default value of 1.3 would improve the fit between the simulations and experimental data. The MINTEQA2 model gave reasonable predictions of the complexation of Cd and Zn by DOC, whereas deviations in the estimated activity of the free Ni2+ ion as compared to experimental results are up to a factor of 5.
Introduction Elevated concentrations of dissolved organic carbon (DOC) in polluted groundwater caused, for example, by leachate from solid waste landfills may enhance the mobility of heavy metals in groundwater to an extent that is supposed to be directly proportional to the effect of complexation by the DOC. Speciation models in conjunction with default databases can be useful tools for predicting the significance of metal-DOC complexation. However, these models have not been extensively tested against field-related observations. A major uncertainty in existing speciation models is their ability to represent the poorly defined DOC originating from * Corresponding author telephone: +45 45251603; fax: +45 45932850; e-mail:
[email protected]. 10.1021/es981105t CCC: $18.00 Published on Web 09/18/1999
1999 American Chemical Society
decaying organic waste because the default databases of the existing speciations models are usually based on empirical complexation constants determined on purified humic and fulvic acids of various origins (1, 2). A recent study by Christensen et al. (3) demonstrated the ability of DOC to form complexes with cadmium (Cd), nickel (Ni), and zinc (Zn) in two leachate-polluted groundwater samples (232 and 131 mg of C L-1) by comparing the sorption of the metals onto aquifer material in the presence and absence of DOC. The use of actual aquifer material and the presence of the DOC in its original matrix suggested that the observed complexation by the DOC would be comparable to the complexation encountered in a field situation. However, the metal-DOC complex formation constants estimated from such observations are conditional and affected by pH, the actual metal levels, and the high salt content of the polluted groundwater (4-6). These apparent complex formation constants are thus valid only for the conditions of the experiments. Generalization of such observations on complexation by DOC in leachate-polluted groundwater would therefore require a large data set representing all the conditions of interest (pH, ionic strength, and ion composition) or a model that can account for such variations. The purpose of this paper is 2-fold: (i) To measure the significance of complexation of Cd, Ni, and Zn by the DOC from leachate-polluted groundwater samples by an easyto-operate traditional resin equilibration method and compare the results to those reported by Christensen et al. (3) that involved aquifer material as sorption medium. (ii) To use the resin equilibration method to measure complexation by DOC from leachate-polluted groundwater for a range of DOC concentrations and use these data to evaluate two existing speciation models containing default databases on metal-DOC complexation: the WHAM model (1) and the MINTEQA2 model (7). The WHAM model is an advanced model and was chosen because it represents a commonly used approach in modeling of metal complexation by dissolved organic matter. The MINTEQA2 model is fairly simple with respect to DOC complexation of metals, but it is a well-known speciation model extensively used for calculating the distribution of inorganic aqueous species. Our objective was to establish a significant data set on heavy metal complexation by unpurified DOC in landfill leachate-polluted groundwater, as opposed to existing data sets for purified fulvic and humic acids. Our data set is then used to evaluate how well existing computer speciation models with default parameters can predict the complexation of heavy metals by DOC in strongly polluted groundwater. Such information is valuable for a balanced appreciation of the predictions of heavy metal complexation by general speciation models for a range of environmental conditions.
Materials and Methods Approach of the Resin Equilibration Method. The significance of metal-DOC complexation can be determined by comparing the metal distribution in two solutions that are equilibrated with a cation-exchange resin: a sample solution that contains DOC and a reference solution that is DOC-free but is otherwise identical to the DOC-containing sample. By performing the resin equilibration experiments with DOCcontaining groundwater mixed in various ratios with the reference solution, the presence of DOC-complexed metal can be determined for a range of DOC concentrations. Leachate-Polluted Groundwater Samples. Two leachatepolluted groundwater samples (L1 and L2) were taken from VOL. 33, NO. 21, 1999 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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TABLE 1. Composition of Leachate-Polluted Groundwater and Reference Solutions after Equilibration with a Cation Exchange Resin in the Batch Sorption Experimentsa sample 1
pH specific conductivity (mS cm-1) ionic strengthb DOC (mg of C L-1) av mol wt of DOCd (g mol-1) TALe (mequiv L-1) Ca2+ (mM) Mg2+ (mM) Na+ (mM) K+ (mM) NH4+ (mM) Zn2+ (mM) Cl- (mM)
sample 2
leachate-polluted groundwater L1
reference R1
leachate-polluted groundwater L2
reference R2
6.60 5.78 0.044 244 2085 0.02 3.2 1.5 17.5 5.4 14.4 0.006 19.6
6.60 5.89 0.045 6c
6.60 1.91 0.013 181 1940 0.02 1.3 0.2 7.7 1.1 4.3 f 7.1
6.60 1.97 0.012 9c
0.02 3.3 1.3 15.2 5.0 13.3 0.006 19.8
0.02 1.3 0.2 7.7 1.0 4.1 f 7.4
a 150 mg of resin, 25.00 mL of solution, 7 days of equilibration). b Estimated by the MINTEQA2 speciation program (7). c Organic carbon was not added to the reference solutions but was released from the resin. d Obtained by size exclusion chromatography (10). e Calculated total alkalinity related to the content of HCO3- and CO32- (low due to aeration of the samples). f In sample L2, the concentration of Zn was less than 7.7 × 10-6 mM.
the anaerobic part of the leachate pollution plume at the Vejen Landfill, Denmark. The samples were obtained from the methanogenic zone of the plume 7 m downgradient of the landfill and approximately 2 m (L1) and 5 m (L2) below the groundwater table. Sampling procedures are described in detail elsewhere (3). The samples were aerated in the dark for 1 week to remove volatile organic carbon, inorganic carbon, and traces of sulfide. Any Fe(II) and Mn(II) were oxidized during aeration and removed as particulate matter by filtration (>0.45 µm, cellulose nitrate filter from Sartorius, Go¨ttingen, Germany). pH was adjusted during aeration with concentrated HNO3 to the natural pH (6.60 ( 0.03). The samples were stored in the dark at 1 °C. It was the intention to manipulate the DOC as little as possible and to keep it in its original matrix of leachatepolluted groundwater, but the above-mentioned steps were introduced to gain control of the experimental system when spiked with heavy metals. The removal of volatile organic carbon, for example, xylene and acetic acid, was done to remove easily degradable organics (less than 1 mg of C L-1) that could support growth of biomass during the experiment. Sulfides were removed to avoid the formation of fine colloidal precipitates upon spiking with heavy metals that may have been interpreted as complexed metal. The samples contained low concentrations of Fe(II) (1.0-7.3 mg L-1) and Mn(II) (0.9-2.1 mg L-1), which, if oxidized after the metal spiking, could form iron and manganese oxides that could scavenge part of the heavy metals and interfere with the interpretations of the results. Thus, oxidizable iron and manganese were removed by oxidation and filtration prior to spiking. The release of DOC from the filters as well as the change in DOC during aeration were monitored. The filters released up to 0.8 mg of C L-1, corresponding to a maximum of 0.5% of the DOC content of the samples; DOC measured daily during aeration showed no trends with time (data not shown), suggesting that DOC degradation was less than the uncertainty of the measurements (about 2%). This suggest that no significant addition or removal of DOC took place during the aeration. It cannot be completely ruled out that small-sized DOC may have undergone transformation without loosing carbon. However, it should be pointed out that the DOC in the samples is several years old and the easily oxidizable carbon is a small percent of the total content (supposedly less than 5%). Considering the minor impacts the performed pretreatment of the samples seem to have on the DOC in the leachate-polluted groundwater samples, we believe that the 3858
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experiments performed regarding metal complexation, as close as experimentally possible, deal with DOC as formed in leachate-polluted groundwater. Synthetic, Inorganic Solutions (Reference Samples). 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 NH4+ salts of chloride and nitrate to mimic the composition of the leachate-polluted groundwater samples with respect to cations, chloride, and ionic strength. The ability of the reference samples to mimic the leachate-polluted groundwater samples was excellent, as seen in Table 1, which shows the composition of reference and leachate-polluted groundwater samples after equilibration with the resins. In particular, it was important that the chloride (Cl-) content was identical in reference and leachate-polluted groundwater as Cl- formed complexes with Cd2+. In these experiments, Cl- complexes constituted up to 50% of Cd not bound in DOC complexes. The pretreated original samples contained (prior to exposure to resin) low concentrations of Al (0.3-2.0 mg L-1), Fe (0.8-1.9 mg L-1), and Mn (0.9-1.9 mg L-1), 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, later model simulations involving these metals showed that their presence had very little effect on the model estimates of the complexation of Cd, Ni, and Zn. Resin Preparation. Amberlite-CG 120 I cation-exchange resin (polystyrene with sulfonic acid groups, 100-200 mesh, 4.4 equiv kg-1 total exchange capacity, Na+ form) was available from Serva (Heidelberg, Germany) as Serdolit CG120 I. The resin was converted from the Na+ form by exposure to the synthetic inorganic solutions. The following procedure was used: Approximately 1000 mL of the Na+ form resin was rinsed several times with Milli-Q water and subsequently soaked in 1 M HNO3 for at least 8 h to remove any trace metal impurities. The resin was divided into two parts and exposed to the two synthetic inorganic solutions (R1 and R2, respectively) with careful stirring. The synthetic inorganic solutions were changed every 2 h. This procedure was continued until the concentrations of the cations in the synthetic solutions were identical before and after exposure to the resin. The resins were dried at 45 °C for several days in a convection oven and stored. Resin Equilibration Experiments. The resin equilibration experiments were carried out in 50-mL polyethylene bottles
containing 25.00 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 (based on preliminary test). During equilibration, the pH was adjusted by the addition of small amounts of HNO3 or NaOH to maintain constant and identical pH in the bottles (pH ) 6.60 ( 0.03). After equilibration, the solutions were separated from the resin by centrifugation (5 min at 3000 rpm) and acidified to pH < 1.5 (concentrated HNO3). Each metal was investigated by a series of 15-20 resin equilibration experiments covering a range of DOC concentrations obtained by mixing leachate-polluted groundwater and reference solution in different ratios. The original experimental results for Ni from series L1 had to be discarded due to analytical problems. Consequently only six experiments are reported, which were undertaken using a more limited supply of saved material. Identical amounts of metal were added to all bottles. The equilibrium solution concentrations in the experiments ranged as follows: Cd, 4.4-14.2 × 10-8 M; Ni, 8.5-37.5 × 10-7 M; and Zn, 4.6-12.2 × 10-6 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 was used in the calculations. The distributions of heavy metals in the presence of DOC were based on single values (one for each DOC concentration). Supplementary Aquifer Material Equilibrium Experiments. Sorption experiments using an aquifer material as sorbent were also performed for L1 using the procedure described in the present work. This was done to obtain more experimental data using aquifer material than presented in ref 3 in order to improve the basis for comparison of the two approaches for a range of DOC concentrations. The aquifer material used was identical to the one used in ref 3. Control Experiment. The DOC contents in the batch sorption experiments were found consistently to be slightly higher after equilibration with the resin. A control experiment was therefore undertaken to check if organic carbon released from the resin during equilibration influenced the metal complexation results. Ten batches containing reference solution, resin, and heavy metals were prepared in an identical manner to the sorption experiments. The batches were equilibrated for varying time (up to 7 days, which is the equilibration time used in the complexation experiments) and centrifuged, and the heavy metal concentrations and DOC concentrations were determined. The control experiments showed that organic matter released from the resin during equilibration increased the DOC concentration over time. However, the total heavy metal concentration in solution remained constant with increasing DOC concentrations (data not shown). This indicated that the DOC released from the resin did not bind any of the studied heavy metals. The DOC concentration measured in the reference experiment (5-9 mg of C L-1) was subsequently used to correct the DOC concentrations measured in the dilution series with leachate-polluted groundwater. Instrumental Analyses. Heavy metals in solution were determined by graphite furnace atomic absorption spectrophotometry (Perkin-Elmer 5000, deuterium background correction, HGA 400 graphite furnace, AS-1 automatic sample injection system) after solvent extraction by 1.0% sodium diethyldithiocarbamate trihydrate in 4-methylpentan-2-one. All samples were acidified to pH < 1.5 (concentrated HNO3) before solvent extraction. Ca2+, Mg2+, K+, and Na+ were determined by flame atomic absorption spectrophotometry (Perkin-Elmer 370). Cl- and NH4+ were analyzed by the standard autoanalyzer routine (Technicon Autoanalyzer II).
FIGURE 1. Illustration of the approach used to calculate the effect of complexation from the measured distribution ratios, Kd. DOC analysis was performed with a TOC analyzer (O-IAnalytical Model 700). pH was measured by a pH meter (Hanna Instruments DP 7916R) using a combination pH electrode (radiometer pH106007-3,5-NC). General. 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 and then rinsed with deionized distilled water and dried at 60 °C in a convection oven.
Data Treatment For each equilibration experiment, the ratio between metal retained on the resin (or aquifer material) and in solution was expressed as a distribution coefficient, Kd. The distribution coefficient is considered to be constant within the metal concentration range used (8). The DOC-complexed metal does not associate with the resin, and the DOC-containing samples were identical to the reference samples except for the content of DOC. The distribution coefficient can thus be related to the effect of DOC complexation at a fixed concentration of metal on the resin as shown in Figure 1. The effect of metal-DOC complexation calculated from the metal distribution measured after equilibration with the resin is approximately equal to the effect of metal-DOC complexation in the original sample because the effect of complexation is independent of the metal concentration when the metal concentration is much smaller than the DOC concentration (9). In the reported experiments, the total dissolved metal concentrations were within a range of 10-710-6 M while the DOC concentration was at a level of 10-410-3 mol of sites L-1, suggesting that the concentration of the free DOC was approximately independent of the amount of metal complexed. The estimated molar site concentrations were based on characteristics, described by Christensen et al. (10), for the DOC in the groundwater samples obtained from the leachate plume.
Models Basic Concept of WHAM. WHAM, described by Tipping et al. (1), is a combination of several submodels. These include models for inorganic solution speciation and a humic metal VOL. 33, NO. 21, 1999 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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ion binding model, named Model V as described in detail by Tipping (11) and Tipping and Hurley (12). Model V describes the binding of metal ions to humic substances by a discrete binding sites model in which binding is modified 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 counterions in the diffuse layer using a Donnan-type expression. The discrete binding sites are represented by two types of site (types A and B), and within each type of site there are four different sites present in equal amounts. The two types of sites are described by intrinsic proton binding constants (pKA and pKB) and spreads of the values (∆pKA and ∆pKB) within each type of site. There are nA (mequiv kg-1) type A sites (associated with carboxylic type groups) and nB ) nA/2 (mequiv kg-1) type B 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 defines whether pairs of proton binding groups are close enough to form bidentate complexes. The two types of site (A and B) have separate intrinsic exchange constants (pKMHA and pKMHB). It is implicit in Model V that the variation in the binding affinities of the proton and metal ions on monodentate sites are perfectly correlated, i.e., a high proton affinity site also has a high affinity for all metal ions. However, the introduction of bidentate sites reduces the correlation between the proton affinity distribution and the metal ion affinity 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 (11, 12). Basic Concepts of MINTEQA2. MINTEQA2 (version 3.11) includes a submodel for estimations of the complexation of metals with dissolved organic matter. This is a composite ligand model with a Gaussian affinity distribution (13-15). 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 nonspecific binding of cations due to electrostatic interactions is not taken into account in this model. The discrete binding sites in MINTEQA2 are represented only by one type of site (carboxylic) characterized by a mean binding constant (µ) and a spread (σ) 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 (σ) 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. (14) and includes only data for binding to the carboxylic sites. Describing the composite ligand model in MINTEQA2, Allison and Perdue (15) pointed out that the model has not yet been satisfactorily verified. They also suggested that including a second type of binding site would improve the model considerably. Binding Properties. A detailed description of the proton binding properties of isolated and purified fulvic acids from the landfill leachate-polluted groundwater samples is given by Christensen et al. (16). Site-specific proton binding data for fulvic acids from leachate-polluted groundwater and default proton binding data used in the WHAM model are summarized in Table 2. For comparison, both site-specific and default proton binding parameters were used for modeling the experimental data of Cd, Ni, and Zn complexation to DOC from leachate-polluted groundwater. 3860
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TABLE 2. Specific and Default Proton Binding Parameters and Default Metal Binding Parameters for the WHAM Model and MINTEQA2 Modela WHAM L1-FA
nA nB pKA pKB ∆pKA ∆pKB -P pKMHA (Cd) pKMHA (Ni) pKMHA (Zn)
L2-FA
5.5 2.75b 3.14 8.01 3.11 6.55 77.5
default, best average
6.01 3.01b 3.28 10.6 2.95 6.49 87.0
4.73 b 3.26 9.64 3.34 5.52 103 1.5 1.4 1.3
MINTEQA2 default, Suwannee L1-FA L2-FA River DOMd mean log K (µ) SD (σ) av charge on free ligand (z) av charge on protonated ligand (z) molar site concn (mequiv g-1)c mean log KCd (µ) mean log KNi (µ) mean log KZn (µ)
7.44
6.92
3.87 1.7 -2.8 -1.8 1.00 3.3 3.3 3.5
a Specific proton binding parameters were obtained from a potentiometric titration of fulvic acids isolated and purified from two leachatepolluted groundwater samples (16). b Prefixed at nA/2. c Only the specific molar site concentrations were used in combination with default parameters in modeling DOC-metal interactions. d Dissolved organic matter isolated from Suwannee River.
Table 2 also contains default proton binding parameters for the carboxylic sites given in the MINTEQA2 database. It was recognized by Allison and Perdue (15) that the DOC molar site concentration of 1 mequiv g-1 was far too low. Thus, the molar site concentration used in modeling the experimental data by MINTEQA2 was based on results from a simple titration on the fulvic, humic, and hydrophilic fractions of the leachate-polluted groundwater as reported by Christensen et al. (10). The molar site concentrations used are given in Table 2 and were estimated as weighted average values taking into account the weight fraction of humic acid, fulvic acid, and the hydrophilic fraction. The total acidity measured in the three fractions was used. The phenolic sites were assumed to behave as carboxylic sites with respect to metal binding since no database for metal binding to phenolic sites exists in MINTEQA2. In the actual samples, the phenolic groups were found to constitute 10-15% of the carboxylic groups (in terms of number of sites) as determined by titration (10). Specific metal binding properties for DOC from leachatepolluted groundwater have not been investigated in the present work. Thus, the metal binding parameters for Cd, Ni, and Zn in the models were maintained at the default values presented in Table 2. Modeling Experimental Data. Characterization of the DOC in the two leachate-polluted groundwater samples showed that the DOC consisted of about 8% humic acids, 48% fulvic acids, and 25% hydrophilic fraction (10). The latter fraction is usually lost in the traditional purification of the fulvic and humic acid fractions, and only recently has the hydrophilic fraction gained attention (see, for example, ref 17). A previous characterization of the DOC in terms of its acid dissociation properties and elemental composition indicated that both the humic and the hydrophilic fraction
FIGURE 2. Experimental results obtained by the resin equilibrium method and the aquifer material sorption method (0, from ref 3; O, from this work) expressed as the effect of complexation by DOC for Cd, Ni, and Zn (log scale) vs DOC concentration. Model predictions of the experimental results obtained by the WHAM model and the MINTEQA2 model are also shown. have features resembling fulvic acids (10). Thus, in this study it was assumed that the humic acids and hydrophilic fraction of the DOC samples (all together 40%) 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 of C L-1. The database of proton and metal binding constants in MINTEQA2 (reported in ref 14) was derived from unfractionated DOC. The DOC concentration in MINTEQA2 is given as molar site concentration (equiv L-1) using the molar site concentrations (mequiv g-1) reported in Table 2 and assuming 50% carbon content.
including M2+ and any inorganic metal complexes. For Ni and Zn, M is approximately equivalent to M2+, and the ratio of the DOC-complexed metal to the free metal ion in solution is for Ni and Zn equal to the effect of complexation minus one. For Cd, chloride complexes make up 30-48% of the dissolved Cd not bound in DOC complexes, and in expressing the ratio of DOC-complexed Cd to the free metal ion in solution (Cd2+), the complexes have to be taken into account. For example, Figure 2 shows that, for L2 at a DOC concentration of 100 mg of C L-1, Cd has a DOC complexation effect of 2 and with 30% of Cd bound in chloro complexes the ratio Cd-DOC/Cd2+ is equal to 1.43.
Results and Discussion
The graphs also show the corresponding results from the aquifer material sorption experiments (from ref 3 and supplemented with new data) and the computer predictions by the WHAM model and the MINTEQA2 model as discussed below. It should be noted that the models were evaluated based on experiments performed at pH 6.60 and ionic strengths of 0.045 and 0.013 M, respectively.
Figure 2 shows the results obtained in the experiments by the resin equilibration technique expressed as the effect of complexation by DOC for Cd, Ni, and Zn as a function of DOC concentration. The effect of metal complexation by DOC is defined as (M + MDOC)/M, where MDOC is the metal-DOC complex and M is dissolved metal concentration
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Resin Equilibration Experiments. The results of the resin equilibration experiment showed that the DOC in leachatepolluted groundwater did have the ability to form complexes with Cd, Ni, and Zn. For both leachate-polluted groundwater samples, the most significant effect of complexation was obtained in the undiluted samples (L1, 240 mg of C L-1; L2, 180 mg of C L-1). For Cd, Ni, and Zn in sample L1, 60-80% of the total metal content in solution was bound in DOC complexes at the maximal DOC concentration of 240 mg of C L-1. In sample L2, 70-80% of Cd, Ni, and Zn in solution was bound in DOC complexes at a DOC concentration of 180 mg of C L-1. Differences between the two leachate-polluted groundwater samples in the effect of metal complexation are relatively small and might be influenced by several parameters as differences in ionic strength, cation composition (Ca2+ and Mg2+ and also L1 contains Zn2+ that might compete with Cd2+ for binding sites at the DOC), and complexation capacity of the DOC in the two samples. Effect of DOC Complexation: Comparison of the Two Experimental Approaches. Comparing results from the resin equilibration experiments with results from aquifer material sorption experiments showed good agreement for Cd, Ni, and Zn equilibrated with sample L2 and for Ni for L1 (Figure 2). For Cd in sample L1 (Figure 2), the effect of complexation was slightly less for the aquifer materials than for the resin. The resin equilibrium method was easy to operate in the laboratory, and the results were rather consistent with the data obtained by the aquifer material. This suggests that the resin equilibrium method is an easy and useful approach to measure complexation in leachate-polluted groundwater. Effect of DOC Complexation: Comparison of Model Predictions and Experimental Results. The ability of the two models to predict metal complexation for Cd, Ni, and Zn in the leachate-polluted groundwater samples L1 and L2 varied with respect to both models and metals. The WHAM model gave excellent predictions of Cd complexation to DOC in leachate-polluted groundwater sample L1. Fair predictions were obtained for Ni in sample L1 and for Cd and Ni in sample L2, where observations and predictions were within 50%. The predictions of the effects of complexation by DOC for Zn were 1 order of magnitude higher than experimental results in L2. Considering the uncertainties usually associated with thermodynamic data on complexation reactions, the predictions for Cd and Ni were not unacceptable. For example, at 100 mg of C L-1 in sample L2, the activity of Ni2+ is estimated to be only a factor of 2.5 lower than its observed value. Generally the deviations on the activity of the free metal ion between experimental results and model predictions were within a factor of 2.5. The shapes of the model prediction curves and the experimental curves as a function of DOC concentrations were fairly similar, but for L2, WHAM consistently overestimated the complexation with DOC. The WHAM model was not very sensitive to variations in proton binding parameters as reported by Christensen et al. (16). Using specific proton binding parameters determined for fulvic acids in the two samples of leachate-polluted groundwater did not improve the prediction of metal binding to DOC, suggesting that the deviations are related to the metal binding constants. The fact that the WHAM prediction of Zn binding to DOC from leachate-polluted groundwater was very different from the observations might be due to the fact that the best average parameters for Zn, used in the default database, are based on rather few data and thus may not be reliable. Setting the binding constant of Zn to fulvic acids at 1.7 instead of the default value at 1.3 would provide an excellent prediction of the experimental results. The MINTEQA2 model predictions of DOC binding of Cd were excellent for L1 and L2 and fair for Ni and Zn in L2. The 3862
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prediction of Ni complexation with DOC in L1 was significantly lower than experimental observations. The shape of the MINTEQA2 curves for increasing DOC concentrations was fairly similar to the observations except for Ni in L1. As discussed above, the observed deviations are not considered to be crucial for the use of models. For Ni in L1, where MINTEQA2 gave the worst prediction, the effect of complexation was underestimated corresponding to a deviation in the free metal ion activity up to a factor of 5. Field Relevance of Results. Metal speciation models accounting for complexation with DOC are few and are not validated in the context of strongly polluted groundwater. While some validations exist for extracted and purified humic and fulvic acids, we chose to used DOC in polluted groundwater in order to stay as close as possible to the field conditions regarding the DOC. Minor manipulations were needed to control the experimental system, but we believe on the basis of the performed control measurements and the refractory nature of the DOC that the experiments with respect to DOC were very close to actual field conditions. The manipulation that potentially had the largest effect was the removal of reduced iron and manganese from the samples and the neglect of minor concentrations of Fe, Mn, and Al in the speciation calculations due to uncertainty about the actual species of these metals in the sample. By including Al, Fe(II), and Mn(II) in the WHAM calculations at concentration levels observed in the pretreated samples, estimates of the activity of the free metal ion increased up to 20%. At high concentrations of Fe, Mn, and Al present in an actual plume, these elements should also be included in the speciation modeling. Accepting differences in the effect of complexation between model predictions and experimental results of a factor of 2-3 on the activity of the free metal ion, the WHAM model combined with the default database provided useful predictions of Cd and Ni complexation to DOC from leachatepolluted groundwater. In a leachate-polluted groundwater environment, the MINTEQA2 model gives a useful first approximation of the extent of metal complexation by DOC, accepting uncertainty in the effect of complexation within a factor of 5. However for Zn, MINTEQA2 gave a much better prediction of the complexation by DOC than the WHAM model using the default database.
Acknowledgments Dr. E. Tipping (Institute of Freshwater Ecology, Ambleside, U.K.) and Dr. J. Allison (Allison Geoscience Consultants, Flowery Branch, GA) kindly provided the newest versions of the applied models and commented on the early draft of the manuscript. These contributions are gratefully acknowledged. Anita Hauritz is gratefully acknowledged for technical assistance, and Torben Dolin is acknowledged for making the drawings. This work was financially supported by the Danish Technical Research Council. The study is part of a major research program focusing on the effects of waste disposal on groundwater. The program is funded by the Danish Technical Research Council, The Technical University of Denmark, and the European Commission.
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Received for review October 27, 1998. Revised manuscript received July 19, 1999. Accepted July 28, 1999. ES981105T
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