Environ. Sci. Technol. 2003, 37, 2743-2749
Strong Hg(II) Complexation in Municipal Wastewater Effluent and Surface Waters HEILEEN HSU AND DAVID L. SEDLAK* Department of Civil & Environmental Engineering, University of California, Berkeley, California 94720
The speciation of mercury(II) in the aquatic environment is greatly affected by the presence of ligands capable of forming extremely strong complexes with Hg(II). In this study, a novel competitive ligand exchange (CLE) technique was used to characterize Hg(II)-complexing ligands in samples collected from three municipal wastewater treatment plants, a eutrophic lake, a creek located downstream of an abandoned mercury mine, and a model water containing dissolved Suwannee River humic acid. These samples contained 3.3-15.9 mg/L dissolved organic carbon and were amended with 1.0-1.7 nM Hg(II) for CLE analysis. Results indicated that all samples contained labile Hg(II)complexing ligands with conditional stability constants similar to those of reduced sulfur-containing ligands. Two wastewater effluent samples also contained approximately 0.5 nM of ligands that formed extremely strong Hg(II) complexes that did not dissociate in the presence of competing ligands. The conditional stability constant of these extremely strong or nonlabile complexes (i.e., cKHgL) were estimated to be greater than 1030, for the reaction Hg2+ + L′ ) HgL. The third wastewater sample and the eutrophic lake sample contained lower concentrations, 0.070.09 nM, of nonlabile Hg(II)-complexing ligands. The results suggested that these extremely strong Hg(II)complexing ligands should account for most of the dissolved Hg(II) species in municipal wastewater effluent and may dominate Hg(II) speciation in effluent-receiving waters.
Introduction The complexation of Hg(II) plays an important role in the transport and methylation of mercury in the aquatic environment (1-7). Previous research suggests that the mobilization of Hg(II) in watersheds is enhanced by complexation with dissolved organic matter (1, 2) and by association with colloidal material (5, 7-9). In addition, the uptake of mercury by microorganisms depends on the concentration of bioavailable mercury species (6, 7, 10, 11). In particular, complexation of Hg(II) by natural organic matter (NOM) can reduce the concentration of these bioavailable species and thereby inhibit the biological uptake of both Hg(II) and monomethylmercury (MeHg) (10, 11). Hg(II) complexation is also important to mercury methylation because uncharged, inorganic Hg(II) species are thought to affect the net rate of MeHg production by sulfate-reducing bacteria (7). Results from equilibrium speciation models indicate that sulfide and thiols (as low-molecular weight compounds or * Corresponding author phone: (510)643-0256; fax: (510)642-7483; e-mail:
[email protected]. 10.1021/es026438b CCC: $25.00 Published on Web 05/14/2003
2003 American Chemical Society
as functional groups on NOM) form extremely strong complexes with Hg(II) (12, 13). In the presence of oxygen, both types of reduced sulfur-containing ligands are unstable. As a result, Hg(II) speciation in oxic waters should be controlled by weaker complexes with other functional groups on NOM. However, the importance of such complexes is unclear because nanomolar concentrations of sulfide and thiols have been detected in oxic surface waters (14-19). Therefore, reduced sulfur-containing ligands also may be important in oxic waters. The concentration of strong Hg(II)-complexing ligands in natural waters probably depends on the fraction of the water that originated from reducing environments. Anaerobic sediments may be an important source of reduced sulfurcontaining ligands in surface waters because sediment porewaters typically contain high concentrations of sulfides and polysulfides. Additionally, nucleophilic substitution of sulfide and polysulfides into organic matter results in the formation of thiol functional groups in NOM (20). Reduced sulfur-containing ligands can diffuse from sediment pore waters to oxic surface waters (16) where they persist through stabilization induced by complexation with metals such as Cu(II) and Zn(II) (17, 21). Municipal wastewater effluent also may be a source of reduced sulfur-containing ligands in surface waters (17, 19). Wastewater effluent contains relatively high concentrations of organic matter, and processes such as anaerobic digestion could mimic some of the processes that occur in anaerobic sediments. The purpose of this study was to characterize Hg(II)complexing ligands originating from different sources. Although previous studies have indicated the presence of strong Hg(II)-complexing ligands on isolated NOM (22, 23), it has been difficult to extend the available speciation techniques to natural waters because the methods lack the needed sensitivity. To measure metal speciation under conditions close to those encountered in surface waters, Hg(II) speciation was measured by competitive ligand exchange (CLE) followed by C18 solid-phase extraction (SPE). The CLESPE technique was used to compare the relative binding strength of Hg(II)-complexing ligands in a model water containing dissolved Suwannee River humic acid, and in samples from three municipal wastewater treatment plants, a eutrophic lake, and a creek contaminated by runoff from an abandoned mercury mine. Analysis of the titration data provided insight into the sources and concentration of strong Hg(II)-complexing ligands in the aquatic environment.
Material and Methods Materials. All chemicals were purchased from Fisher Scientific at the highest available purity unless otherwise noted. Stock solutions were prepared by dissolving chemicals in deionized water (Barnstead Nanopure II). Trace-metal grade acids were used for all acid solutions. Stock solutions of 3.0 mM diethylammonium diethyldithiocarbamate (Sigma-Aldrich) and 5.0 mM glutathione (Sigma-Aldrich) were prepared daily for use in CLE experiments. Stocks solutions of 20% SnCl2‚H2O in 1N HCl, 30% NH2OH‚HCl, and bromine monochloride (BrCl) were prepared according to procedures outlined by EPA Method 1631 (24). A stock solution of 5.0 µM Hg(II) was prepared by diluting a Sigma-Aldrich ICP/AA Hg(II) standard (1 g of Hg(II) per liter in 10% HNO3) in 0.5% (v/v) BrCl. A stock solution of 2.2′-dithiobis(5-nitropyridine) (DTNP) was prepared by dissolving 2.0 mM DTNP (SigmaAldrich) in HPLC-grade acetonitrile. All bottles used for sample collection and storage were made of Teflon or glass. Sample filters and tubing were made VOL. 37, NO. 12, 2003 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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of polyethylene. All equipment was cleaned in Micro detergent (Cole Parmer), followed by a 24-h soak in 1 N HCl. The SPE column consisted of a 1-cm diameter glass column with polyethylene fittings (BioRad) packed with 0.5 g end-capped C18 resin (Supelco ENVI-18 resin). The resin columns were cleaned prior to extraction by first suspending the resin in HPLC-grade methanol (CH3OH) and pumping a solution of 1 N HCl in 10% CH3OH through the resin for approximately 10 min at 4 mL/min. The resin was rinsed by pumping approximately 30 mL of deionized water through the column. CLE-SPE Analysis. The CLE-SPE technique utilizes two competing ligands, diethyldithiocarbamate (DEDC-) and glutathione (GSH3-), followed by solid-phase extraction to separate Hg(II) complexes with the competing ligand from Hg(II) complexes with naturally occurring ligands. DEDC forms hydrophobic Hg(DEDC)20 complexes that are retained by C18 SPE, while GSH forms hydrophilic HgHn(GSH)2n-4 complexes (where n ) 1,2 at pH 6-9 (25)) that are not retained by the C18 resin. This technique builds upon previous CLE approaches (26, 27) and extends them to Hg(II). The concentrations of hydrophobic and hydrophilic naturally occurring ligands were initially defined by measuring Hg(II) retained by C18 SPE in sample aliquots to which no competing ligands were added. To characterize the naturally occurring ligands, the competing ligands were added in separate titration steps. The formation of Hg(DEDC)20 complexes was quantified by the concentration of Hg(II) retained during C18-SPE minus the initial concentration of hydrophobic Hg(II) (determined from the aliquot with no competing ligand added). The formation of the hydrophilic HgHn(GSH)2n-4 complexes was quantified by subtracting the concentration of hydrophilic Hg(II) initially present in the sample from the total hydrophilic Hg(II) concentration. The strength of the Hg(II)-ligand complexes was measured by titrating a sample with a range of competing ligand concentrations. CLE-SPE was performed on grab samples that were collected in glass bottles and filtered within 2 h of collection using a peristaltic pump fitted with an in-line 0.45-µm cartridge filter (Gelman Aquaprep 600) and polyethylene tubing. Separate aliquots of filtered samples were collected for pH and dissolved organic carbon (DOC) measurements. Separate aliquots of filtered and unfiltered samples also were analyzed for total Hg(II). Equipment blanks, which consisted of deionized water pumped through the sample collection and filtration equipment, resulted in an average increase of 0.0041 nM Hg ((0.0013 nM, n ) 4) in the water. Immediately after filtering, 1.1-1.3 nM Hg(II) was added to the samples undergoing speciation analysis. After equilibrating the Hg(II)-amended samples at room temperature for at least 2 h, the competing ligand was added and equilibrated for 45-65 min prior to C18 SPE. These equilibration times were chosen to prevent replication problems that may occur with shorter times (i.e., less than 10 min) and to avoid changes in speciation and oxidation of the competing ligands by biological activity that may occur with longer equilibration times (i.e., greater than 24 h). Approximately 280 mL of sample was pumped through the acid-cleaned resin column at 4-5 mL/minute using a peristaltic pump fitted with 1/8′′ i.d. Teflon tubing and Cole Parmer C-Flex tubing in the pump head. Hg(II) retained on the C18 resin was eluted by rewetting the resin with 2 mL of CH3OH and pumping approximately 80 mL of 1 N HCl through the column. The C18-extracted sample and the C18 eluate were preserved with 0.5-1.0% (v/v) bromine monochloride and later analyzed for total Hg(II). Precautions were taken to minimize artifacts caused by the presence of reactive gaseous Hg in laboratory air. During the SPE pumping, the 250 mL bottles that contained the 2744
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sample reservoir and the C18-extracted sample were stored in 2 L glass bottles. The headspace of these larger glass bottles was continuously purged at approximately 300 mL/minute with ultrahigh purity N2 that was further purified using goldcoated sand traps. The extent of Hg contamination during the extraction procedure was assessed by pumping deionized water through the C18 resin columns. The average increase of [Hg(II)] in deionized water that passed through the column was 0.0052 nM ((0.005 nM, n ) 3). In the HCl eluate, the average increase in [Hg(II)] was 0.0486 nM ((0.0061 nM, n ) 3). Because all samples used in this study contained 0.5-1.7 nM Hg(II), these results indicated that Hg contamination during the SPE step was not significant. The average recovery (( standard deviation) of Hg(II), quantified by the sum of Hg(II) in the HCl eluate and in the C18-extracted sample, was 88.6% (( 8.8%, n ) 47). Sample Analysis. Total Hg(II) concentration was measured by cold vapor atomic fluorescence spectroscopy (CVAFS) using SnCl2 reduction followed by dual-stage gold amalgamation (28, 29). The average Hg(II) detection limit, based on three times the standard deviation of the bubbler blank, was 0.002 nM. Sample aliquots for DOC quantification were stored in amber glass vials and frozen until analysis. DOC was measured by combustion catalytic oxidation followed by nondispersive infrared detection (Shimadzu TOC-5000A). The detection limit for DOC was 0.5 mg/L. GSH may be unstable in aerobic waters due to catalytic oxidation (30). Therefore, the concentration of GSH was measured in samples being analyzed for Hg(II) speciation by derivatization with DTNP followed by quantification by high performance liquid chromatography (31). In samples collected after the 45-65 min GSH equilibration step, the average recovery of GSH from the concentration initially added was 88.6% ((7.0%, n ) 10). The result indicated that most of the GSH was stable in these samples during the equilibration step. Therefore, the concentrations of added GSH or measured GSH (where available) were reported in the results of this study. CLE-SPE Method Validation. Although published stability constants for Hg(DEDC)20 and HgHn(GSH)2n-4 are available (see Table 1), the stability constants of these complexes were estimated for our experimental conditions by conducting CLE titrations in model water solutions. The stability constant for Hg(DEDC)20 was determined by conducting the CLE titration on a HgEDTA2- solution. The solution consisted of 1.1 mM K2H2EDTA and 0.5 nM Hg(II) dissolved in deionized water. The HgEDTA2- solution was buffered by 4.0 mM 3-Nmorpholino propansulfonic acid (MOPS, Sigma Aldrich) and adjusted to pH 7.2 with concentrated NH4OH (trace-metal grade). In this HgEDTA2- solution, all of the Hg(II) is hydrophilic, as indicated by a recovery experiment in which 104% of the initial Hg(II) was measured after the sample was passed through the C18 resin column. Therefore, the stability constant for the Hg(DEDC)20 complex was measured by adding varying concentrations of DEDC to aliquots of the HgEDTA2- solution and measuring the amount of Hg(DEDC)20 retained on the C18 resin column. The stability constant of Hg-GSH complexes was determined separately by conducting a GSH titration of a Hg(DEDC)20 solution. This solution consisted of 0.5 nM Hg(II) and 0.09-0.11 µM DEDC, buffered at pH 7.4 by 1.5 mM MOPS. GSH was varied from 0.74 to 10.5 µM. The solution was equilibrated at least 45 min prior to SPE. Because Hg(DEDC)20 is retained during C18 SPE, the concentration of Hg-GSH complexes was determined by measuring the concentration of Hg(II) passing through the C18 resin column. Study Sites. Hg(II)-complexing ligands were evaluated in a model water containing dissolved humic acid isolates and
TABLE 1. Published and Measured Stability Constants Used for Modeling of CLE Titration Data (25 °C, I ) 0) log K H+
ref
+ S EDTA4- + 2H+ S H2EDTA24+ EDTA + 3H S H3EDTAEDTA4- + 4H+ S H4EDTA
10.948 17.221 20.340 22.500
(42) (42) (42) (42)
Hg2+ + EDTA4- S HgEDTA2Hg2+ + H+ + EDTA4- S HgHEDTA-
23.106 26.706
(42) (42)
DEDC- + H+ S HDEDC0
3.38
(36)
Hg2+ + 2DEDC- S Hg(DEDC)20
38.48a 39.94a 33.35b
(37) (38) measured
GSH3- + H+ S HGSH2GSH3- + 2H+ S H2GSHGSH3- + 3H+ S H3GSH0
8.88 17.12 20.40
(35) (35) (35)
Hg2+ + 2GSH* S Hg(GSH)2* c
30.66b 30.76b
(25) measured
Hg2+ + L′ S HgL
log cKHgL
EDTA4-
HEDTA3-
Two-phase solvent extraction constants for (Hg2+)aq + 2(DEDC-)aq S (Hg(DEDC)20)solv. b pH 7.4 c [GSH*] ) [GSH3-] + [HGSH2-] + [H2GSH-] + [H3GSH] ≈ [H2GSH-] at pH 6-8; [Hg(GSH)2* ] ) [Hg(GSH)24-] + [HgH(GSH)23-] + [HgH2(GSH)22-] ≈ [HgH2(GSH)22-] at pH 6-8. a
in freshly collected field samples. The field sites consisted of three municipal wastewater treatment plants, a eutrophic lake, and a creek contaminated by an abandoned mercury mine. The humic acid solution was prepared by dissolving Suwannee River humic acid (International Humic Substance Society) at a concentration of 20.2 mg/L in 3.0 mM KNO3. The humic acid solution was buffered by 1.4 mM MOPS, and the pH was adjusted to 7.1 with concentrated NH4OH. Hg(II) was added to a total concentration of 1.7 nM. Municipal wastewater effluent samples were collected from three different municipal wastewater treatment plants (referred to as WWTP 1, WWTP 2, and WWTP 3). WWTP 1 is a 165 MGD (7.23 m3/s) treatment plant that employs oxygen-activated sludge during secondary treatment. Samples were collected after the secondary clarifiers, prior to the chlorination/dechlorination process. WWTP 2 is a 35 MGD (1.5 m3/s) treatment plant that also employs oxygen-activated sludge for secondary treatment. These samples were collected before and after chlorination/dechlorination (to test the effects of chlorination on Hg(II)-complexing ligands). Finally, WWTP 3 is a 2.1 MGD (0.092 m3/s) plant that employs trickling biofilters for secondary treatment followed by ultraviolet (UV) disinfection. Samples were collected at the discharge point (i.e., after UV disinfection). Surface water samples were collected from two sites. The first site was at the outlet of Lake Anza, a shallow eutrophic lake located in Tilden Regional Park, Berkeley, CA. Inputs to the lake are predominantly precipitation and runoff, with nutrient loadings derived mainly from the surrounding recreational area, golf course, and botanical gardens. Samples from the lake were collected on three different dates in January and February 2002. The second surface water site was Alamitos Creek, which is located downstream of the abandoned New Almaden Mercury Mine in Santa Clara County, CA. Samples from the creek were collected in September 2002.
Results The CLE-SPE method was calibrated by a DEDC titration of a HgEDTA2- solution and a GSH titration of a Hg(DEDC)20
FIGURE 1. CLE-SPE results for (a) DEDC titration of 0.5 nM Hg(II), 1.1 mM EDTA, pH 7.4; solid line represents the model plot of % Hg(II) as HgEDTA2-, using the FITEQL-optimized KHg(DEDC)2; (b) GSH titration of 0.5 nM Hg(II), 0.11 µM DEDC, pH 7.4. Solid line represents the model plot of % Hg(II) as Hg(GSH)2*, based on KHg(DEDC)2 estimated in part (a) and using the FITEQL-optimized stability constant cKHg(GSH)2 for Hg(GSH)2*. solution. In the solution containing HgEDTA2-, the percentage of Hg(II) passing through the C-18 resin (i.e., hydrophilic Hg(II)) decreased as [DEDC] increased (Figure 1a), indicating the formation of Hg(DEDC)20. The stability constant for Hg(DEDC)20 was estimated from the titration data using FITEQL, a nonlinear least-squares-fitting program for chemical equilibrium data (32). The estimate used the previously published stability constant for HgEDTA2- and EDTA4- protonation constants summarized in Table 1. The FITEQL estimate of KHg(DEDC)2 is provided in Table 1, and results of the model fit are depicted with the solid line in Figure 1a. The stability constant for the Hg-GSH complex was estimated with a similar procedure. As the concentration of GSH increased in the Hg(DEDC)20 solution (i.e., as [DEDC]T/ [GSH]T decreased), the concentration of hydrophilic Hg(II) complexes increased indicating that Hg-GSH complexes were formed (Figure 1b). Using the Hg(DEDC)20 stability constant from the HgEDTA2- titration, the stability constant for the Hg-GSH complex was estimated with FITEQL. The results of the model fit are indicated with the solid line in Figure 1b. Samples analyzed as part of this study included environments where biological activity and anaerobic conditions were likely to produce strong Hg(II)-complexing ligands. The three WWTP samples, which are representative of processes commonly employed for secondary treatment, contained between 10 and 14 mg/L DOC (Table 2). The sample from Lake Anza also contained relatively high DOC, while the sample from Alamitos Creek contained relatively low DOC. The concentration of DOC used in the Suwannee River humic acid solution was chosen to resemble a high-DOC surface water. Analysis of total Hg(II) concentrations in unfiltered, and filtered samples (Table 2) indicated that all of the field samples VOL. 37, NO. 12, 2003 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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TABLE 2. Dissolved Organic Carbon (DOC), pH, Unfiltered and Filtered (
