Environ. Sci. Technol. 1996, 30, 1645-1651
Effects of High pH on Arsenic Mobility in a Shallow Sandy Aquifer and on Aquifer Permeability along the Adjacent Shoreline, Commencement Bay Superfund Site, Tacoma, Washington PAUL E. MARINER,* FRED J. HOLZMER, RICHARD E. JACKSON, AND HANS W. MEINARDUS INTERA Inc., 6850 Austin Center Boulevard, Suite 300, Austin, Texas 78731
FREDERICK G. WOLF Elf Atochem North America Inc., Health, Environment and Safety Department, 2901 Taylor Way, Tacoma, Washington 98421
A groundwater arsenic plume, derived from arsenite wastes disposed at a chemical plant in Tacoma, WA, extends to the shore of the Hylebos Waterway. The plume is characterized by high-pH, high-silica concentrations generated by past disposal of highpH brines on site. Aquifer Kd values for arsenic decrease at least 10-fold as the pH increases from 8.5 to 11. Near the shore, aquifer sands are cemented, predominantly by opal quartz. Cementation reduces porosity to about 19%; however, very little pore space is interconnected. Along the shore face, a massive amorphous precipitate, high in Si and Mg, is found. SOLMINEQ calculations show that mixing highpH, high-silica groundwater with seawater causes initial supersaturation of brucite [Mg(OH)2] and magnesium hydroxysilicates. The cementation has likely considerably reduced the cumulative discharge of arsenic to the waterway.
Introduction Commencement Bay, located in southeastern Puget Sound, accommodates seven industrial waterways within the city limits of Tacoma, WA. In 1985, the Commencement Bay Nearshore/Tideflats Superfund Site Remedial Investigation found arsenic contamination in the sediments of the Hylebos Waterway (1). Among the potentially responsible parties is Elf Atochem North America Inc., which owns a chemical plant on the waterway. * Corresponding author telephone: (512) 346-2000, ext. 221; fax: (512) 346-9436; e-mail address:
[email protected].
0013-936X/96/0930-1645$12.00/0
1996 American Chemical Society
From 1940 to 1971, Pennwalt Corporation, the former owner of the Elf Atochem chemical plant, produced sodium arsenite (NaAsO2), a pesticide/herbicide called Penite. The disposal of Penite wastes in unlined pits on the property generated a groundwater arsenic plume that today extends to the shore of the Hylebos Waterway (Figure 1). An advanced pump-and-treat system has been operating since 1991 to contain and remediate the plume. In aquatic systems, inorganic arsenic occurs primarily in two oxidation states, As(V) and As(III) (2). In general, As(V) predominates in oxidizing waters, while As(III) predominates under reducing conditions. The reduction of arsenate to arsenite is very slow in natural aquatic systems, which explains why arsenate is found in reducing environments (3). Alternately, arsenite can be found in oxidizing environments (4, 5). In the pH range of 7-12, the pH range pertinent to this study, the predominant aqueous arsenate species are HAsO42- and AsO43-. The predominant arsenite species in the absence of sulfide are H3AsO30, H2AsO3-, and HAsO32-; otherwise arsenic sulfide species predominate (6). Organic forms of arsenic have not been detected at the site. The aqueous solubility of arsenic depends on the arsenic oxidation state and the concentrations of other dissolved elements. Certain arsenate minerals have been shown to limit the aqueous solubility to the 1 000 mg/L range (7). For arsenite, the presence of sulfide reduces the aqueous solubility to around 1 mg/L (6). However, waters can be oversaturated with respect to arsenic sulfide minerals due to slow mineral formation (4). Under oxidizing and mildly reducing conditions, groundwater arsenic concentrations are usually controlled by adsorption, not mineral precipitation (5). Arsenic is strongly adsorbed to iron minerals (5, 7, 8). Hydrous iron oxide, a ubiquitous secondary mineral in aquifer sediments, has a very high specific surface area (around 600 m2/g) and thus a very high adsorption capacity (9). The extent of arsenic adsorption under equilibrium conditions is described by the value of the distribution coefficient, Kd (L/kg). The distribution coefficient measures the equilibrium partitioning ratio of adsorbed to dissolved contaminant, i.e., the adsorbed concentration (mg of As adsorbed/kg of aquifer sediment) divided by the aqueous concentration (mg of As dissolved/L of groundwater). The value of Kd is conditional because it depends strongly on the pH of the water, the arsenic oxidation state, and the temperature (4, 5, 7, 8). In acidic and neutral pH waters, As(V) is extensively adsorbed while As(III) is relatively weakly adsorbed. In high-pH waters, Kd values are considerably lower for both oxidation states. Site Hydrogeology. The Elf Atochem chemical plant is located on a former tidal marsh of the Puyallup River delta. During the 1920s, the tidal flats were dredged to make waterways capable of transporting ocean-going vessels. The land surface of the surrounding marsh was raised above flood level with the dredged sediments to provide industrial sites. The Elf Atochem plant is on the south shore of the Hylebos Waterway. There are three distinct shallow sand aquifers beneath the plant. They are referred to as the upper, intermediate, and lower aquifers and are predominantly composed of
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FIGURE 1. Map of the arsenic plume in upper aquifer, October 1994. Prior to cementation of the sediments near the shore and on the shore face, the general direction of groundwater flow was likely toward the waterway. The sheet pile barrier was installed in 1990. The pumpand-treat system has controlled groundwater flow since 1991.
sand and silty sand. The upper aquifer consists of the sands dredged from the waterway. The aquifers are separated by units known as the first and second aquitards, comprised mainly of silty clay and clayey silt. The first aquitard, originally the surface of the tide flats, is generally continuous except in some regions near the shore. The second aquitard is continuous across the site. While high tides cause some water from the Hylebos to flow into the aquifers, there is a net flux of water from the aquifers to the waterway. Past Pennwalt operations at the Elf Atochem site greatly altered the hydrogeochemistry of the upper and intermediate aquifers. The activity of most significance was the disposal of “brine treatment mud”, a slurry of calcium carbonate, magnesium hydroxide, and caustic wastes into disposal ponds immediately to the southeast of the Penite pits (Figure 1). The disposal began in 1967 and ended in 1990. Leaching of these wastes generated a large plume of high-pH water in the upper unconfined aquifer, which dissolved large amounts of silica from the aquifer sands. Figure 1 shows pH contours in October 1994. It was suspected that precipitation reactions had occurred near the shore due to the mixing of high-pH groundwater with seawater. These reactions were thought to have reduced aquifer permeability near the shore soon after the disposal ponds were installed, thereby limiting the cumulative historical discharge of arsenic-contaminated groundwater to the waterway. Indeed, a zone of low
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permeability was indicated by historically high piezometric elevations in the upper and intermediate aquifers near the shore. Objectives. The objectives of this study were 2-fold. The first objective was to identify the processes and minerals responsible for the retardation of arsenic in the groundwater plume. The second objective was to determine the cause, extent, and effect of the low-permeability zone in the upper and intermediate aquifers near the shore.
Experimental Section Aquifer core samples were collected from the borehole locations shown in Figure 1 using a cable-tool drill rig and a 5-ft Solinst cohesionless sand sampling tool. After recovery, the cores and PVC liners were cut into sections, capped, and sealed for shipment. For core R, five 10-cm sections from the upper and intermediate aquifers were shipped to Battelle Marine Sciences Laboratory, Sequim, WA, for atomic adsorption (AA) analysis of arsenic in pore water and X-ray fluorescence (XRF) analysis of solid-phase arsenic concentrations. For cores S and T, ten 10-cm sections from the intermediate aquifer were shipped to Battelle for detailed analysis. The samples were centrifuged to separate the pore water from the sediments. The pore water was analyzed for arsenic (AA) and pH, and the centrifuged sediments were analyzed for solid-phase arsenic (XRF). The sediments were then composited for grain-size
fractionation analysis. There were five size fractions for each core (1000 µm). The fractions were analyzed for Al, Cl, Fe, Si, As, Cu, Mn, Pb, Sb, and Zn by XRF and subsequently subjected to acid extractions to determine the importance of secondary minerals on the adsorption of arsenic in the sands. The acid extractions were designed to selectively dissolve the secondary minerals, particularly aluminum, iron, and manganese oxides, and any metals adsorbed to them. The extractions were performed on 1-g samples from each size fraction using the procedures outlined in Jackson and Inch (10). The procedure involved adding 15 mL of 0.1 M hydroxylamine hydrochloride, an acidified reducing agent, to 1-g samples in 0.1 M potassium tetroxalate (pH 1.5) and analyzing for extracted metals. Boreholes C-1, C-2, and C-3 were drilled specifically to investigate the cemented sands in the upper and intermediate aquifers near the shore. These borehole sediments were sampled continuously from 3 ft below ground surface to the second aquitard. An extensive study of the shore face at low tide provided additional samples of cemented sands and mineral precipitates. The study was initiated at low tide when the seepage face of the upper aquifer and most of the first aquitard were exposed. A series of photographs was taken to document the study, and shoreface samples were collected for analysis. A geology rock hammer was used to probe for cemented shore-face sediments. In all, six shore-face samples, one sample of precipitate from a nearby extraction well, and 12 samples from boreholes C-1, C-2, and C-3 were analyzed by XRF and X-ray diffraction (XRD) for chemical and mineral composition. Several samples of the cemented sands were also analyzed by energy dispersive X-ray spectrometry (EDS), scanning electron microscopy (SEM), and thinsection microscopy. For the thin-section microscopy, a thin-section slide was prepared to compare the ratio of solid (mineral) area to voids, as a measure of porosity. In addition, the thin section was vacuum impregnated with blue epoxy to aid in the identification of interconnected pore spaces.
Results and Discussion Groundwater Geochemistry. Much of the groundwater in the upper aquifer at the site has a pH of 11 or higher (Figure 1). Figure 2 shows groundwater silica concentrations at the site versus pH. Below pH 9, silica concentrations generally fall below 100 mg/L. Above pH 11, they exceed 10 000 mg/L. The high concentrations are likely caused by the dissolution of silica along the flow path. In the noncemented aquifer sand samples, Si is primarily found in plagioclase feldspar [(Na,Ca)Al(Si,Al)3O8], quartz [SiO2], and potassium feldspar [KAlSi3O8]. According to XRD analyses on aquifer sediment samples, these minerals, make up approximately 55%, 20%, and 10% of the mass, respectively. The groundwater data in Figure 2 closely match a plot of the solubility of amorphous silica (amSiO2) versus pH, suggesting that am-SiO2 is the mineral controlling groundwater silica concentrations. Table 1 gives a representative groundwater sample from the plume. SOLMINEQ, a chemical equilibrium model developed by the U.S. Geological Survey (11), was used to interpret the data. Adjusting the water composition for saturation with respect to calcite and am-SiO2 required (1) an increase in the dissolved silica concentration from 680 to 1020 mg/L and (2) a minor pH adjustment from 10.76
FIGURE 2. Plot of dissolved silica concentration versus pH in groundwater samples from site monitor wells in 1989 (18). The solubility lines of am-SiO2 and quartz are plotted for comparison.
to 10.70. The calculated composition at saturation is given in Table 1, column 2. The dissolved silica and arsenic species in this high-pH water account for the vast majority of the total measured alkalinity. Na+ and Cl- account for nearly all the positive and negative charge in solution; thus, the -15% charge balance error is likely due to error in the measurement of one or both of these ions. The upper and intermediate aquifer materials are comprised predominantly of sand-sized grains. According to the results of bulk XRF analyses, they have a high Si and Al content (ca. 300 and 75 g/kg, respectively). Fe concentrations range from about 32 to 62 g/kg. Mn content ranges from 0.55 to 1.22 g/kg. The results of the grain-size fractionation analysis and acid extractions are presented in Table 2. They indicate that the aquifer sands contain a large concentration of iron and aluminum oxide secondary minerals, especially in the largest and smallest size fractions. The highest concentration of arsenic is found in the smallest size fraction. The calculated percentage of arsenic extracted from core S is 115%. It is impossible to extract more than 100% of the arsenic; thus, the overall analytical error in this calculation (primarily due to errors in the measurement of arsenic concentrations) is at least 15%. The calculated percentage of arsenic extracted from core T is 88%. These calculations suggest that nearly all of the solid-phase arsenic is extractable. A large fraction of the solid-phase arsenic is likely adsorbed to iron and aluminum oxides (5). About 20% of the total iron and 5% of the total aluminum was extracted in the process. The fractions not extracted are believed to be incorporated in primary minerals. Field Kd values were calculated from the core samples, as shown in Table 3. These values are plotted versus pore water pH in Figure 3. The data show that Kd decreases at least 10-fold as the pH increases from 8.5 to 11. This trend is consistent with the hypothesis that the majority of the arsenic is adsorbed (4, 7). The variation in Kd data at a given pH may be due to variations in redox conditions. The distribution of arsenic between the +3 and +5 oxidation states were not measured in these samples. Arsenic is more extensively adsorbed under anaerobic conditions (5). Because As(V) concentrations were not measured, the saturation with respect to Ca3(AsO4)2 and Mg3(AsO4)2 could not be estimated. Previous geochemical modeling indicated these minerals could limit As(V) solubility at high
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TABLE 1
Measured and Modeled Composition of Groundwater, Hylebos Water, and a 1:1 Mixture Prior to Precipitation unit
groundwater from 6D17-1a
model groundwaterb
model Hylebos waterc
1:1 mixtured
mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L as CaCO3 °C mg/L
10.76 8
