Evaluation of Environmental Effects on Metal Transport from Capped

Previous studies conducted in our laboratories have shown that submarine groundwater discharge (SGD) can significantly increase metal fluxes from capp...
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Environ. Sci. Technol. 2001, 35, 4549-4555

Evaluation of Environmental Effects on Metal Transport from Capped Contaminated Sediment under Conditions of Submarine Groundwater Discharge C H U N H U A L I U , * ,‡ J E N N Y A . J A Y , † A N D TIMOTHY E. FORD† Environmental Science and Engineering Program, Department of Environmental Health, Harvard School of Public Health, 655 Huntington Avenue, Boston, Massachusetts 02115, and Gradient Corporation, 238 Main Street, Cambridge, Massachusetts 02142

Previous studies conducted in our laboratories have shown that submarine groundwater discharge (SGD) can significantly increase metal fluxes from capped contaminated sediment to the overlying water. Five columns were set up in the laboratory to evaluate the effects of environmental factors such as groundwater pH, sediment depth, and groundwater flow rate on metal transport from capped contaminated sediment under conditions of SGD. Acidified groundwater discharge was shown to enhance the mobility of all tested metals except Mo. Although much of the released metal was adsorbed by the capping material, significant increases of initial or steady-state fluxes to the overlying water were observed for Ni, Cu, Zn, Cd, Pb, and Mn. Additional sediment depth enhanced steadystate fluxes for all tested metals except Mo, Cd, and Pb. Increased SGD rates did not significantly change the average metal concentration in the outflow to the overlying water for most metals; however, all metal releases were higher due to the greater flow at increased SGD rates. The residence time and the redox conditions may be important in evaluating environmental effects on capping efficiency.

Introduction Contaminants released to estuarine and coastal waters are likely to be scavenged by particles and removed to the sediment. Toxic contaminants can subsequently migrate from sediments and exert adverse ecological effects (1). Capping, or subaqueous containment of sediment, is receiving considerable attention as a remediation option for highly contaminated sediment in low hydrodynamic energy environments (2). While nearshore areas are often sites of contaminated sediments and candidate sites for capping, these areas are also likely to be subjected to submarine groundwater discharge (SGD), which may cause continuous advective flow through sediment and capping material. SGD and its role in metal transport have been examined in recent laboratory studies and results suggested that SGD decreased capping efficiency for all tested metals (3). Further, capping * Corresponding author phone: (781)401-3200; fax: (781)4012575; e-mail: [email protected]. † Harvard School of Public Health. ‡ Parsons Engineering Science, Inc. 10.1021/es001763p CCC: $20.00 Published on Web 10/18/2001

 2001 American Chemical Society

enhanced release of redox sensitive metals such as Mn, Fe, and Mo and other metals which may co-transport with Mn and Fe; this phenomenon was more pronounced in the presence of SGD. Accumulation of Mn at the surface of the capping material was also observed under conditions of SGD (3). Metal transport and capping efficiency under conditions of SGD may be affected by a variety of environmental factors including characteristics of SGD, sediment and capping material, hydrogeochemical conditions at the site, and biotic and human activities. Increased metal mobility with increased groundwater acidity has been shown to be an important mechanism for H+ buffering of near-shore sediments (4-7). The most pronounced effects on metal release from acidified soil and dredged sediment have been reported for Zn, Cd, and Mn (8, 9). Studying metal release at lower pH is especially important due to the predominance of free ionic species, potentially available for biological uptake (5). The depth of contaminated sediment under a cap determines the total mass of metals and the residence time of groundwater through the contaminated sediment layer. In addition, depth of sediment may affect the redox conditions in the sediment and capping layers, which, in turn, may affect release of redox sensitive metals. Increasing SGD rates should enhance the pore water transport velocity, while decreasing the contact time between groundwater and sediment. SGD flow rate may also affect the redox profile in the sediment and capping layers, with potentially significant effects on release of redox sensitive metals. Accurately predicting capping efficiency at environmentally different sites requires an understanding of environmental effects on metal transport from capped sediment. This study was designed to evaluate effects of groundwater pH, the depth of capped sediment, and the groundwater flow rate on metal release and capping efficiency under conditions of simulated groundwater inflow.

Methods and Materials Material Collection. Details of material collection were presented previously (3). In brief, sediment was collected from a nearshore site in New Bedford Harbor, MA (depth ≈ 1 m) in October 1998. Particles with diameters greater than 2-3 mm were carefully removed prior to the experiment (e.g. worms, crabs, snails, pebbles, rocks, sticks). Sand was collected from Ossipee Beach, NH and presoaked with groundwater. Table 1 presents characterization of the sediment and the sand. Groundwater was collected from the Brigham Street Well located in Northboro, MA, on September 25, October 9, November 2, and December 8, 1998. The groundwater was deoxygenated by bubbling with pure nitrogen to a dissolved oxygen concentration less than 0.3 ppm. The deoxygenated groundwater was then stored in a nitrogen-filled bag during its use as an inflow source. Experimental Method. The experiment was conducted by setting up five columns in the laboratory: a control column to simulate neutral (pH 7) submarine groundwater discharge through 5 cm sediment capped by 15 cm capping material at a flow rate of 4.4 × 10-4 m/h specific discharge, an acidified column with similar settings as the control column to simulate acidified groundwater (pH 3) discharge, an increased-depth column with similar settings as the control column except the sediment depth was 8 cm, a high-flow column to simulate elevated neutral groundwater flow (1.1 × 10-3 m/h specific discharge) through capped sediment with similar settings as the control column, and a combination column testing VOL. 35, NO. 22, 2001 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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TABLE 1. Characterization of Sand and Sediment Used in the Experiment

material

Ni

Cr

Cu

Zn

Mo

Cd

Pb

Mn

particle density (kg/m3)

sediment sand

55 3.4

251 4.44

454 1.67

514 27

9.0 0.42

4.40 0.054

156 13.2

311 148

2323 2632

metal concentration (µg/g dry sediment)

combined effects of acidified groundwater (pH 3), increased sediment depth (8 cm), and elevated groundwater flow rate (1.1 × 10-3 m/h). The diameter of each polycarbonate column is 17 cm. A layer of sand placed at the bottom of each column evenly distributed the influent. Well mixed sediment was carefully placed on the top of this sand layer. Approximately 7080 g groundwater-soaked sand was used to form a 15 cm capping layer in each column. Groundwater was then added to the top to form a 2.5 cm water layer. Columns were allowed to settle for 24-48 h between additions of different layers (sediment, sand, and water) or before the initiation of the experiment. A six channel peristaltic pump (Scientific Industries, Inc., Bohemia, NY) was used to pump water through the columns. Groundwater was acidified to pH 3 using pure HCl (Fisher, Pittsburgh, PA) before it was used as an inflow water source for the acidified column and the combination column. The overlying water, the capping sand, and the bottom sand were kept at pH 3 for these two columns. For all the other columns, groundwater at the original pH (approximately 7) was used as the inflow source and the initial overlying water. The columns were set up in the laboratory at room temperature (∼20 °C). The following metals were studied: Ni, Cr, Cu, Zn, Mo, Cd, Pb, and Mn. Sampling methods were presented previously (3). In brief, the effluent was collected in cylinders connected to the column outlet, and the volumes were recorded. The collected water sample was filtered through a 0.4 µm polycarbonate filter and stored at pH < 2 prior to analysis for dissolved metal concentration. At the end of the experiment, core samples that included both the capping layer and the sediment layer were taken with modified 200-ml cylinders. One core from each column was sectioned at 1 cm interval, and the samples were stored frozen in 50 mL polyethylene tubes before they were digested and analyzed. Chemical Analyses. Details of chemical analyses including trace metal analysis, sediment digestion, and salinity measurement were presented previously (3). In brief, salinity and pH were measured for the water samples immediately after sampling using a YSI Model 30 salinity meter (YSI incorporated, Yellow Springs, OH) and a pH meter (VWR Scientific, Bridgeport, NJ). The sediment samples were microwave-digested. Briefly, approximately 0.2 g of sediment sample was microwavedigested with concentrated hydrofluoric and nitric acids in a microwave digestion bomb (Parr, Moline, IL). The samples were then diluted with 1.5% boric acid. Metal concentrations in the sediment digestates and water samples were determined by Inductively Coupled PlasmaMass Spectrometry (ICP-MS, Perkin-Elmer Elan 5000, Norwalk, CT). The detection limit was determined according to the USEPA approach (10). Values smaller than the detection limit were assigned to zero instead of the detection limits to prevent overestimation of metal release. Dissolved oxygen in the overlying water was measured using a Dissolved Oxygen Meter (VWR, West Chester, PA) several times throughout the experiment. Quality Control. All handling and sample processing complied with the strict procedures for trace metal analysis. Deviation of observed values from the certified values for a reference marine sediment, BCSS-1, was less than 10% for all tested metals. Precision was estimated from the replicate 4550

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porosity 0.58 0.29

analysis of 10% of the samples (randomly chosen) and the coefficient of variation was less than 10% for all tested metals. For ICP-MS analysis of dissolved metal concentrations, deviation of observed reference solution NIST 1643d values from certified values was never more than 10%. Data Analysis. Mass balance calculations and determination of dissolved metal fluxes have been previously described (3). Briefly, the cumulative dissolved metal release at a certain time point t was calculated by summing the metal release each day prior to t, derived from the dissolved metal concentration and the collected effluent volume. Metal concentrations in the groundwater in this study were in the range of metal concentrations in natural groundwater (11) and were therefore not subtracted from the metal release. The cumulative dissolved metal release was then plotted against the cumulative effluent volume, and the curve was modeled using linear regression or using a nonlinear regression SAS program (NLIN) for a segmented model (12). The regression slope of this cumulative metal release (µmol) against volume (m3) provided an estimate of the amount of metal released per volume of groundwater outflow (µmol/ m3). This value multiplied by the flow rate (4.41 × 10-4 m/h specific discharge for the control column, the acidified column, and the increased-depth column, and 1.05 × 10-3 m/h specific discharge for the high-flow and combination columns) and a conversion factor (24 h/day) provided a flux of metal release to the overlying water in µmol/m2/day.

Results Figure 1 shows salt release versus cumulative flow volume. The salt flux along with the initial several liters of the effluent (referred to as initial flux hereafter) was significantly higher than the following flux (referred to as steady-state flux hereafter). The acidified column and the control column had similar curves. Compared to the control and acidified columns, the high-flow column had lower initial salt flux but similar steady-state flux; the increased-depth column had elevated steady-state flux. Figure 2 plots cumulative metal release versus cumulative flow volume. The slope of the regression line represents the average metal concentration in the outflow discharge to the overlying water. Similar to the salt curve, most metals had elevated initial fluxes compared to the steady-state fluxes. Both the acidified column and the combination column had enhanced average metal concentrations in the initial 10 L of groundwater outflow (where the experiment ended for the combination column) compared to the other columns for all metals except Cr, Cu, and Mo. The average concentration of outflow during the steady-state period was very similar for the control column and the high-flow column for most of the metals. The pore volumes of the capped sediment were approximately 0.7 L and 1.1 L for the columns with 5-cm and 8.2-cm sediment, respectively. The figure shows that metal release to the overlying water continued after over 30 times of the sediment pore volume of groundwater discharged through the sediment. Table 2 presents metal flux to the overlying water. Acidified groundwater significantly enhanced initial fluxes of Cu, Mn < Ni, Cd < Zn < Pb and steady-state fluxes of Ni < Zn < Cu compared to the fluxes from the control column. Increased

FIGURE 1. Cumulative salt release vs cumulative flow volume.

TABLE 2. Dissolved Metal Flux to the Overlying Water (µmol/m2/day)a Ni Cr Cu Zn Mo Cd Pb Mn a

initial steady-state initial steady-state initial steady-state initial steady-state initial steady-state initial steady-state initial steady-state initial steady-state

control

acidified

2.0 ( 0.6 0.17 ( 0.02 0.44 ( 0.02 0.22 ( 0.01 1.6 ( 0.2 0.21 ( 0.03 3.2 ( 0.4 0.18 ( 0.01 1.08 ( 0.05 1.08 ( 0.05 0.26 ( 0.07 0.009 ( 0.002 0.001 ( 0.001 0.0062 ( 0.0003 299 ( 5 22 ( 2

6.2 ( 0.6 0.27 ( 0.03 0.45 ( 0.01 0.23 ( 0.01 3.7 ( 0.3 0.47 ( 0.05 25 ( 1 0.33 ( 0.07 0.76 ( 0.05 0.76 ( 0.05 0.70 ( 0.06 0.008 ( 0.004 0.066 ( 0.004 0.0014 ( 0.0002 564 ( 27 41 ( 69

increased-depth 2.0 ( 0.3 0.30 ( 0.01 0.38 ( 0.01 3.7 ( 0.8 0.37 ( 0.07 3.2 ( 0.9 0.26 ( 0.04 0.68 ( 0.01 0.68 ( 0.01 0.26 ( 0.02 0.002 ( 0.001 0.0007 ( 0.0001 0.0054 ( 0.0001 534 ( 16 55 ( 6

high-flow

combination

3.7 ( 0.8 0.40 ( 0.06 0.79 ( 0.04 0.49 ( 0.01 6.3 ( 0.6 0.45 ( 0.06 5.1 ( 0.9 0.5 ( 0.1 1.77 ( 0.04 1.77 ( 0.04 0.7 ( 0.1 0.015 ( 0.003 0.005 ( 0.001 0.020 ( 0.001 712 ( 33 141 ( 473

5.9 ( 0.7 1.60 ( 0.37 0.90 ( 0.01 7.8 ( 0.6 2.5 ( 0.7 36 ( 3 10 ( 5 1.19 ( 0.03 1.2 ( 0.1 0.04 ( 0.02 0.13 ( 0.02 0.0031 ( 0.0006 1302 ( 197

The number following ( gives the 95% confidence interval.

sediment depth enhanced initial fluxes of Cu and Mn and increased steady-state fluxes for all metals with the exception of Mo, Cd, and Pb. Increased SGD rate increased initial and steady-state fluxes for all metals except that the increase was not statistically significant for steady-state Mn flux. The combination column had elevated initial and steady-state fluxes for all metals except that steady-state Pb flux was decreased compared to the control column. Brown material was observed at the surface of the capping material in all columns except the combination column. Table 3 provides a comparison of metal concentrations in this material and in the original sand material. Surface accumulation of Mn was significant in the collected surface material for all of the columns. Cu in the increased-depth column, Zn in the acidified and increased-depth columns, and Cd in the increased-depth and high-flow columns were found to accumulate significantly at the surface of the capping material.

Metal concentrations in the original material and average metal concentrations in the sediment and capping material by the end of the experiment are presented in Table 4. The bottom 2-3 cm of the sediment was not recoverable for all of the cores, and only capping material was recovered for the combination column. Metal concentrations in the sediment tended to decrease in the acidified column with the exception of Mo and Pb. However, Cr and Zn were the only metals that decreased significantly. Average Mo concentrations in the sediment by the end of the experiment for all the columns were close to the Mo concentration in the original sediment. Significant changes in metal concentration in the capping material were as follows: (1) enhanced Zn and Cd concentrations were detected for both the acidified column and the combination column, (2) elevated Pb concentration was detected for the combination column, and (3) Cr concentration in the acidified column was elevated in the 2-cm section above the sediment-capping interface. VOL. 35, NO. 22, 2001 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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TABLE 3. Metal Concentration in the Original Sand and Solid Material Accumulated at the Capping Surfacea metal concentration (µg/g dry sediment)

original sand control acidified increased-depth high-flow a

Ni

Cr

Cu

Zn

Mo

Cd

Pb

Mn

3.4 ( 0.6 2.5 ( 0.3 2.6 ( 0.3 3.9 ( 0.3 3.3 ( 0.3

4.44 ( 0.07 3(2 3(2 5(2 3(2

1.67 ( 0.09 2.1 ( 0.4 1.6 ( 0.4 3.0 ( 0.4 1.7 ( 0.4

27 ( 1 25 ( 3 77 ( 3 33 ( 3 23 ( 3

0.42 ( 0.03 0.36 ( 0.05 0.34 ( 0.05 0.47 ( 0.05 0.42 ( 0.05

0.054 ( 0.0007 0.044 ( 0.011 0.07 ( 0.01 0.10 ( 0.01 0.09 ( 0.01

13.2 ( 0.6 15 ( 3 13 ( 3 13 ( 3 12 ( 3

148 ( 6 554 ( 61 291 ( 61 252 ( 61 299 ( 61

Data for the original sand and the control column are from Liu et al. (3). The number following ( gives the 95% confidence interval.

FIGURE 2. Cumulative metal release vs cumulative flow volume: 1, control; 2, acidified; 3, increased-depth; 4, high-flow; and 5, combination.

Discussion Groundwater pH Effects. The observed increases in both initial and steady-state Ni, Cu, and Zn fluxes and initial Cd, Pb, and Mn fluxes from the acidified column may be caused by the elevated solubility of these metals in acidified water. 4552

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This observation is supported by previous studies (13, 14) and the measured decrease of sediment metal concentrations in this experiment. However, the acidified groundwater had no apparent effects on Cr fluxes and steady-state Cd and Mn fluxes. This may be related to the adsorption of the metals by the capping layer.

TABLE 4. Average Metal Concentrations in the Original Material and the Sediment and Capping Material by the End of the Experimenta metal concentration (µg/g dry sediment) Ni original sand control acidified increased-depth high-flow combination original sediment control acidified increased-depth high-flow

3.4 ( 0.6 3.2 ( 1.0 3.7 ( 1.3 3.5 ( 0.7 4(1 5(1 55 ( 2 58 ( 2 49 ( 8 52 ( 6 56 ( 2

Cr

Cu

Zn

Mo

Average Metal Concentration in Capping Material 4.44 ( 0.07 1.67 ( 0.09 27 ( 1 0.42 ( 0.03 5(1 1.7 ( 0.4 27 ( 4 0.39 ( 0.05 4.4 ( 0.8b 1.7 ( 0.2 40 ( 10 0.41 ( 0.06 4(1 2(2 28 ( 4 0.41 ( 0.08 5(1 1.8 ( 0.3 28 ( 7 0.4 ( 0.1 5(2 1.7 ( 0.2 33 ( 3 0.40 ( 0.09 251 ( 8 250 ( 4 191 ( 48 237 ( 22 241 ( 15

Average Metal Concentration in Sediment 454 ( 28 514 ( 27 9.0 ( 0.4 459 ( 13 509 ( 18 8.8 ( 0.1 379 ( 121 391 ( 83 9 ( 1 441 ( 74 470 ( 69 8.7 ( 0.5 470 ( 74 541 ( 56 8.4 ( 0.9

a The number following ( gives the 95% confidence interval. included in the average calculation.

b

Cd

Pb

Mn

0.054 ( 0.0007 0.05 ( 0.02 0.07 ( 0.01 0.059 ( 0.005 0.059 ( 0.004 0.072 ( 0.008

13.2 ( 0.6 12 ( 2 14.2 ( 0.7 15 ( 2 16 ( 2 15.5 ( 0.9

148 ( 6 179 ( 40 149 ( 17 138 ( 16 170 ( 32 171 ( 24

4.40 ( 0.06 4.6 ( 0.5 3.7 ( 0.8 4.1 ( 0.5 4.2 ( 0.3

156 ( 42 157 ( 2 168 ( 5 156 ( 27 134 ( 19

311 ( 11 338 ( 16 280 ( 39 316 ( 57 332 ( 51

The segments within 2 cm distance from the cap-sediment interface were not

FIGURE 3. pH change of the effluent samples. Decreased Mo fluxes by acidified groundwater may be related to the redox chemistry of Mo. In anoxic sediments, Mo occurs mainly as Mo(IV) complexes with sulfide (15). However, all of the Mo species in the lower states are progressively unstable and tend to revert to Mo (VI). The critical p values for redox transformation of Mo(VI)/Mo(IV) should shift from -3.9 (EH ) -231 mV) at pH ) 8 to 0.5 (EH ) 30 mV) at pH 5 (13). The p value of -3 to -1 (EH ) -200 to -50 mV) in sediment close to the sediment-capping interface should therefore favor formation of Mo(VI) at pH 8, while the reduced form is dominant at pH 5. MoO42- is soluble and the predominant species at pH 7-8 (16), while Mo(IV) has low solubility compared to Mo(VI). Therefore, acidified groundwater decreased Mo release from sediment. Acidified groundwater increased the initial Pb flux while decreasing the steady-state flux. Dissolved Pb concentrations increase rapidly between p ) -5 to -4 (EH ) -296 to -237 mV) at pH 8 and between p ) 0-1 (EH ) 0-59 mV) at pH 3 (17). The sediment EH values for all columns were generally around -350 to -200 mV, close to the critical p range at pH 8 and much lower than the critical p range at pH 3. This provides an explanation for the lower steady-state Pb flux at lower pH. Although the groundwater was deoxygenated before being used as inflow source, it contained a certain amount of dissolved oxygen. Therefore, the initial sediment EH value may be lower than the steady-state EH value.

Dissolved Pb concentration may increase as marine systems become more reduced due to the formation of soluble Pb-S complexes, as observed by Brugmann (18). This may explain the elevated initial Pb flux from the acidified column. All metal concentrations except Pb appeared to decrease in the sediment during the experiment, while the differences were statistically significant only for Cr and Zn. Conversely, metal concentrations appeared to increase in the capping material, with statistically significant differences for Cr, Zn, and Cd. The metal loss from the sediment and the metal increase in the capping material were much greater than the measured metal release to the overlying water, suggesting that acidified groundwater increased metal mobility in the sediment, while much of the released metal was adsorbed by the capping material. The metal mobilities from sediment in the acidified column were generally much lower than the metal mobilities observed in another study (19). In that study, 17-98% of total metal release was reported in laboratory bench experiments using Hamburg harbor sediments acidified with sulfurous acid to pH 4.0: Cd and Co, 98%; Mn, 91%; Cu, 84%; Ni, 66%; Cr, 45%; Fe, 27%; and Pb, 17%. However, that study did not attempt to simulate capping or SGD and the sediment was acidified to pH 4. Neutralization of acidified groundwater by the sediment observed in our experiment (Figure 3) resulted in >99% reduction in H+ concentration for both the VOL. 35, NO. 22, 2001 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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acidified and the combination columns, in agreement with a similar study which tested groundwater neutralization by uncapped sediment (7). The pH of acidified groundwater used in this study was 3, which has been used in several acid leaching studies (e.g. ref 20). Although this is not representative of natural conditions (Groundwater pH as low as 4.0 was observed (8).), experimental results indicated the significance of pH in metal mobilization from capped contaminated sediment. A decrease in pH will affect specific metals differently. A sharp increase in adsorption with increasing pH through a restricted 1-2 unit pH range, referred to as a pH-adsorption edge, is characteristic for each metal on any chosen adsorbent (21). The effects of acidified groundwater on metal fluxes to the overlying water are dependent on both the mobility of individual metals and the adsorption capacity and kinetics of the capping material. Capped Sediment Depth Effects. Steady-state salt flux from the increased-depth column was significantly greater than flux from the control column (Figure 1). The increase in steady-state fluxes for Ni, Cr, Cu, Zn, and Mn may therefore be simply due to increased amounts of metals in the sediment. Elevated steady-state fluxes may also be caused by longer residence time of groundwater in the sediment (4.4 days for the increased-depth column compared to 2.7 days for the control column, estimated from flow rate and sediment porosity). Increased sediment depth may also decrease sediment EH, in turn contributing to the increase of initial and steady-state Mn fluxes by favoring soluble Mn2+ formation. Almost identical initial fluxes of Ni and Zn from the control and the increased-depth columns agree with the initial salt flux data (Figure 1), suggesting no significant effects from increased sediment depth. Cotransport with Mn may contribute to the higher initial Cu flux and steady-state fluxes of Ni, Cr, Cu, and Zn. Increased sediment depth decreased initial Cr flux, possibly due to lowered EH values resulting in decreased solubility of Cr(III) hydroxide, oxide, and halides, and metal chromites. The EH difference decreased as groundwater with the same dissolved oxygen concentration flowed through the columns. Therefore, increased steadystate Cr flux was observed. The presumably more reducing environment in the capped sediment in the increased-depth column would be expected to decrease Mo release, due to the low solubility of Mo in the lower valence state. Almost identical initial fluxes of Cd from the control column and the increased-depth column agree with the initial salt flux data. Cd generally occurs as Cd(II) in natural aquatic environments and, thus, is not directly affected by the fluctuations in the redox conditions of a system. However, Cd is sulfophilic (with weaker binding with sulfide compared to other metals), and oxidation of sulfide may enhance Cd release. Previous studies (5, 22) suggest that small changes from reduced to oxidized conditions can enhance Cd release. The presumed lower EH values in the sediment and capping material did not favor sulfide oxidation; therefore, the steadystate Cd flux from the increased-depth column decreased. In general, metal release may increase with increased sediment depth. Field sediment depth is usually much larger than under the experimental conditions; therefore, fluxes of most of the metals are expected to be higher in the field than in the laboratory. For Pb, it is possible that field Pb fluxes may be higher when field EH values become lower than -300 mV at pH 8 due to formation of the dominant soluble Pb-S complexes. Groundwater Flow Rate Effects. Minimal effects on average metal concentration in the steady-state outflow may be due to negligible solubility changes between residence times of 1.2 and 2.7 days for the high-flow column and the control column, respectively. 4554

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Decreased average Mo concentration in the steady-state outflow from the high-flow column can be explained by redox speciation change of Mo. The sediment-capping interface for the control column had a EH around 0 mV, while the high-flow column had a EH around -250 mV. In the highflow column, the appropriate p value for Mo oxidation occurred in the capping layer, while in the control column it occurred at the sediment-capping interface, where the source of Mo was located. Thus, the environment in the control column favored transformation of lower valence Mo(IV) to high valence MoO42-. In general, increased SGD rates may have little effect on average metal concentrations in the outflow; however, they may greatly enhance metal flux due to greater flow at increased SGD rates. Interactions of Environmental Factors. Metal release under different combinations of environmental effects appears to be controlled by the most significant effect. For example, increased fluxes due to acidified SGD were also observed in the combination column. The residence time and redox condition change caused by combined effects are two important parameters to evaluate metal behavior.

Acknowledgments This publication was made possible by grant number 2 P42 ES-05947 from the National Institute of Environmental Health Sciences, NIH, with funding provided by EPA. Its contents are solely the responsibility of the authors and do not necessarily represent the official views of the NIEHS, NIH, or EPA. The authors would like to thank MIT Sea Grant for financial support. We are grateful to two anonymous reviewers for helpful comments. The authors would also express their thanks to James Shine, Nick Lupoli, Raveendra Ika, and Robert Weker for laboratory support and Paul Catalano for assistance with statistical analysis.

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(15) Green, M. L. H. In Molybdenum: an Outline of its Chemistry and Uses; Braithwaite, E. R., Harber, J., Eds.; Elsevier Science B. V.: Amsterdam, 1994; pp 94-145. (16) Braithwaite, E. R. In Molybdenum: An Outline of its Chemistry and Uses; Braithwaite, E. R., Haber, J., Eds.; Elsevier: Amsterdam, 1994; p 1. (17) Sadiq, M. In Toxic Metal Chemistry in Marine Environments; Sadiq, M., Eds.; Marcel Dekker: New York, 1992; pp 304-355. (18) Brugmann, L. Mar. Chem. 1988, 23, 425-440. (19) Calmano, W.; Ahlf, W.; Fo¨rstner, U. In Proceedings International Conference Heavy Metals in the Environment, September, 1983, Heidelberg; CEP Consultants Ltd.: Edinburgh, 1983; pp 952955.

(20) Korte, N. E.; Skopp, J.; Niebla, E. E.; Fuller W. H. Water, Air, Soil Pollut 1975, 5, 149-156. (21) Bourg, A. C. M. Cont. Shelf Res. 1987, 7, 1319-1332. (22) Gambrell, R. P.; Khalid, R. A.; Verloo, M. G.; Patrick, W. H., Jr. Dredged Material Research Program Report D-77-4; U.S. Army Corps of Engineers: Vicksburg, MS, 1977; p 309.

Received for review October 12, 2000. Revised manuscript received July 17, 2001. Accepted August 14, 2001. ES001763P

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