Environ. Sci. Technol. 2000, 34, 3665-3673
Metal Speciation and Bioavailability in Contaminated Estuary Sediments, Alameda Naval Air Station, California P E G G Y A . O ’ D A Y , * ,† S U S A N A . C A R R O L L , ‡ SIMON RANDALL,§ ROGER E. MARTINELLI,‡ S U S A N L . A N D E R S O N , #,| JOHN JELINSKI,# AND JOHN P. KNEZOVICH‡ Geology Department, Arizona State University, Tempe, Arizona 85287-1404, Earth and Environmental Sciences Directorate, Lawrence Livermore National Laboratory, Livermore, California 94550, Department of Earth Sciences, University of Bristol, Bristol, BS8 1RJ England, Energy and Environment Division, Lawrence Berkeley National Laboratory, Berkeley, California 94720, and Bodega Marine Laboratory, University of California Davis, P.O. Box 247, Bodega Bay, California 94923
Measurements of simultaneously extracted metals (SEM), acid volatile sulfide (AVS), and invertebrate toxicity were combined with X-ray absorption spectroscopy (XAS) to evaluate metal speciation and ecological hazard of contaminated sediments from the Seaplane Lagoon, Naval Air Station Alameda (CA). This site is characterized by moderate to low toxicity in surface sediments and by metal concentrations in sediments and porewaters that increase with depth. Standard 1-h ΣSEM/AVS measurements for surface sediments were compared with time-series (0.2524 h) measurements of metal and sulfide release from sediments at 30 cm. Results show that AVS is rapidly and completely evolved after 1 h, but metal extraction continues with time and is not complete after 24 h. Sedimentwater interface tests of invertebrate toxicity using sand dollar embryos (D. excentricus) and adult amphipods (E. estuarius) exposed to intact cores showed no to low toxicity in surface sediments. In sediments from 30- and 60-cm depth, high toxicity in several replicates was attributed to factors other than metal concentrations, such as high dissolved ammonia or low dissolved oxygen concentrations. Metal speciation and bonding determined from XAS show that cadmium (100%), zinc (≈80%), and manganese (≈50-70%) are associated with monosulfide phases in the sediments. The remaining fraction of zinc and manganese and all of the chromium and lead are ligated by oxygen atoms, indicating association with oxide, carbonate, or silicate minerals. Iron is present in the sediments in two fractions, as Fe(II) in the sulfide phase pyrite and as oxygenligated octahedral iron, probably associated with clay minerals. Bulk chemical measurements of porewaters and sediments, and speciation information from XAS, suggest that AVS could be accounted for by volatilization of porewater sulfide. Our results indicate that metals are removed from porewaters by formation of monosulfide phases only for cadmium and partially for zinc and manganese but not for lead or chromium, even though these are reduced, 10.1021/es9913030 CCC: $19.00 Published on Web 07/27/2000
2000 American Chemical Society
anoxic sediments typical of a restricted marine estuary environment. Comparison of geochemical, spectroscopic, and biological data provides new insight for the interpretation of ΣSEM/AVS measurements and points out the need for synergistic biological/geochemical tests for determining potential ecological hazard.
Introduction The estuary sediments of the East Outfall Site of the Seaplane Lagoon, at the former Naval Air Station (NAS) Alameda located on an island in San Francisco Bay (CA), contain elevated metal concentrations that are potentially hazardous to aquatic biota. The most abundant metals in the sediments are cadmium, lead, chromium, zinc, copper, and nickel. Concentrations of these metal contaminants above background levels in San Francisco Bay result from a 57-year history of military and industrial activity at this site. From 1940 to 1975, the Seaplane Lagoon received about 300 million gallons of wastewater from industrial and storm-sewer outfalls from NAS Alameda (1). The naval station was decommissioned in 1997 under the Base Closure and Realignment Act of 1988 and its subsequent amendments enacted by Congress to reduce the number of active military bases and to return sites to the public and private sectors (2). This legislation also mandated remediation of any hazardous materials and contamination at these sites by the Department of Defense. Owing to its prime development location in San Francisco Bay, NAS Alameda is the subject of ongoing contamination and toxicology studies, hazard assessment, and remediation by the Department of Navy. This process has been accelerated by the “Fast-Track Cleanup Program” introduced in 1993 to expedite base cleanup and property transfer to the community. Key elements of the Fast-Track Program include innovative approaches to rapid assessment of environmental risks, novel cleanup technologies, and substantial community involvement in order to coordinate intended site reuse with the required level of remediation for a particular land use designation (such as residential, industrial, recreational, or a combination) (2). The Seaplane Lagoon and its surrounding area have been designated as a mixed-use commercial marina site in the City of Alameda land use plan (1). Thus, remediation plans must be appropriate for this designated use. The need for rapid sediment contamination and toxicity assessments at sites such as the Seaplane Lagoon underscores recent efforts by the U.S. Environmental Protection Agency (EPA) to establish standard testing methods and quality criteria for metals in sediments (3, 4). One measure of bioavailability proposed by the EPA for five metals (cadmium, copper, lead, nickel, and zinc) is comparison of SEM (simultaneously extracted metal) to AVS (acid volatile sulfide). This is the sum of the molar concentrations of these five metals that are simultaneously extracted from sediment during a cold-acid extraction compared to the molar concentration of sulfide that is volatilized during the same extraction (by either ratio or difference). Measurement of SEM and AVS were adapted from sequential extraction * Corresponding author phone: (480)965-4581; fax: (480)965-8102; e-mail:
[email protected]. † Arizona State University. ‡ Lawrence Livermore National Laboratory. § University of Bristol. # Lawrence Berkeley National Laboratory. | University of California Davis. VOL. 34, NO. 17, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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techniques that were originally designed to separate different forms of iron sulfide (i.e. “reactive” FeS(s) vs FeS2 (pyrite)) within bulk sediments (see ref 5 for a summary of methods). A commonly used procedure for SEM/AVS measurement is sediment digestion in either 1.0 or 6.0 M HCl at room temperature for 30 min or 1 h (6-9). The acid solution is then sparged with N2, and the evolved, volatile sulfide fraction is trapped in basic solution and measured as molar AVS. The concentrations of metals simultaneously extracted during the acid digestion are measured and summed (ΣSEM). The following assumptions regarding these determinations are often made. First, it is assumed that these five metals were removed primarily from porewaters by precipitation as metal monosulfide phases and that the SEM extraction measures the metal concentrations associated with sulfide (7, 10). Second, it is assumed that the primary contributors to AVS are iron sulfide solid phases, which are thought to include mackinawite (Fe1+xS), greigite (Fe3S4), amorphous FeS (6, 7, 11), and perhaps minor quantities of other solid metal sulfides such as ZnS, PbS, NiS, and CdS (12, 13). Third, it is assumed that pyrite (FeS2) does not dissolve and contribute sulfide to AVS (5, 14). Ratios of ΣSEM/AVS > 1 (or differences of ΣSEM-AVS > 0) are interpreted as potentially toxic because the extractable sediment metal concentration has exceeded the amount of “reactive” sulfide present (6, 7, 15). This criterion stems from the assumption that metals dissolved into solution whose monosulfide solid phases are less soluble than FeS(s) would react with FeS(s) and precipitate via the exchange reaction
Me2+(aq) + FeS(s) ) MeS(s) + Fe2+(aq)
(1)
where Me2+(aq) is an aqueous divalent metal and MeS(s) is the corresponding metal monosulfide phase (6, 9). When the molar ratio of ΣSEM to FeS(s), as measured by AVS, exceeds 1, there is potential metal bioavailability because this ratio is interpreted as insufficient FeS(s) to precipitate metals from solution as given by eq 1. Previous studies have discussed some of the problems with the foregoing assumptions about SEM and AVS measurements. A number of studies have pointed out that ΣSEM/AVS predictions may overestimate metal availability if a significant quantity of metals were bound in phases other than monosulfides and that the acid extraction will leach metals from other nonsulfide phases (8, 13, 15, 16). Cooper and Morse (9) suggested that metals may be present in sulfide phases that are poorly soluble in cold HCl and whose solubility is subject to surface area effects. Thus, metal toxicity could be underestimated if these sulfides were oxidized and released to solution. Furthermore, it has been recognized that this approach is applicable only to reduced sediments with measurable AVS and that other measures of metal toxicity are needed for oxidized sediments (10). Although the EPA has acknowledged limitations of the ΣSEM/AVS methodology, it has been endorsed as the best technology developed to date for assessing the bioavailability of these five metals, partly because the method is thought to have a strong theoretical foundation (3, 4). In this paper, we compare selected SEM and AVS measurements and biological toxicity tests for metalcontaminated sediments from one site at the Seaplane Lagoon to the speciation and local molecular bonding of metals in sediment solid phases as determined directly by synchrotron radiation X-ray absorption spectroscopy (XAS). This method is a unique molecular probe for complex materials because it is element specific, has relatively high sensitivity, does not require a vacuum, and is nondestructive. Thus, spectra for a number of elements can be sequentially measured on a bulk, untreated sediment samples with porewaters present. Standard ΣSEM/AVS measurements for sediment samples were made for five metals (cadmium, 3666
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copper, lead, nickel, and zinc). In addition, we examined the extraction of these and other metals (chromium, iron, and manganese) and AVS production as a function of time and compared this to total metal concentrations from sediment digestion. The ΣSEM/AVS and XAS results were compared with sediment-water interface toxicity tests to assess geochemical factors that affect organism toxicity and the ability of these different tests to predict potential ecological hazards at this site. The data presented here are representative of a larger set of geochemical and biological measurements taken over a 16-month period at this site, which is typical of a temperate-latitude, restricted marine estuary with reducing conditions below the sediment-water interface. By combining geochemical, biological, and spectroscopic data, we use a molecular chemical basis to evaluate metal speciation and potential bioavailability in these reduced estuary sediments.
Methods Sample Collection and Analysis. Between July, 1997, and November, 1998, a series of sediment cores (2-m gravity cores and 0.5-m push cores) and grab samples were collected from the East Outfall Site, Seaplane Lagoon, for geochemistry and toxicity studies (see Figure SI-1, Supporting Information, for site location). For the geochemical and spectroscopic results given here, we focus on two of the 0.5-m Plexiglas push cores (GH-CC-SC4 and GH-CC-SC9; referred to here as SC4 and SC9). Plexiglas cores were sectioned in ambient laboratory conditions under argon and subsamples taken at 3-cm intervals, with sampling from only the center portion of the core to ensure no sediment oxidation. Porewaters were separated by centrifugation (5000 rpm for 60 min in argonfilled centrifuge tubes), filtered (using 0.22 µm polycarbonate filters), and analyzed for major and trace elements by either ICP-MS or ICP-AES. Chloride concentrations were determined using an ion specific probe. Sulfide, sulfate, sulfite, and thiosulfate were determined by ion chromatography using a Beckman 421A controller, a LDC Milton Roy Conducto Monitor III conductivity detector, and a Waters 4.6 × 150 mm IC-Pak Anion HC column. Sediment mineralogy was determined by X-ray diffraction (XRD) on freeze-dried and ground samples (Scintag PAD V instrument using a Cu K-R source at 45 keV and 35 mA from 2 to 92° 2θ in 0.02° steps). Detection limits for crystalline phases by XRD is ≈2 wt %. For total sample element analyses, sediments were acid digested in a microwave (with UA2 acid mixtures; Unisolv, Inc.) and analyzed by ICP-AES (17). Analysis protocol adhered to EPA SW-846 Method 6010A. For XAS, subsamples (≈100 mg) from core SC4 were loaded into a Teflon sample holder, sealed with Kapton tape, and frozen until data collection, which varied from several weeks to several months. To ensure that freezing did not alter the metal speciation, a fresh core (SC9) was collected in November, 1997, and XAS spectra for selected metals were measured within 1-4 days of core recovery. Complete details are given in Carroll et al. (17). SEM/AVS Measurements. Sediment samples taken from cores at 0-5 and 30 cm (collected in November, 1997, and April, 1998) were measured for AVS and SEM as a function of time from 0.25 to 24 h. For AVS, the sample was placed in a round-bottom reaction flask with two serial Pyrex traps containing 80 mL of 0.5 M sodium hydroxide to precipitate volatile sulfide. The sample (15.6 g wet weight) was sparged for 10 min with nitrogen gas. Twenty milliliters of 6 M HCl was then introduced into the reaction flask using a glass syringe, the sample was sparged with nitrogen for the allotted time, and the volatile sulfide fraction was collected in the traps. The sodium hydroxide was decanted into analytically clean glass containers (a separate container for each trap) and analyzed for sulfide (methylene blue method) spectrophotometrically (Shimadzu UV1201 UV-vis spectropho-
tometer) at a wavelength of 670 nm. Calibration curves, matrix spikes, apparatus blanks, and standard recoveries were employed in the sample analysis. Seven aliquots of acid were sampled from the reaction flask over time, filtered, and analyzed by flame or graphic furnace atomic adsorption (Perkin-Elmer 5100 Series Atomic Absorption Spectrophotometer) for SEM concentrations. Toxicity Testing. Four replicate intact cores from the East Outfall site (collected in July, 1997) were used for sedimentwater interface toxicity tests (as opposed to standard sediment-based or water-only tests) at three sediment depths. Sediment was extruded from cores and 2-cm subsections were removed at depths of 0, 30, and 60 cm. The core sections were placed in a sediment-water interface exposure chamber (18), covered with 500 mL of clean 0.45-µm filtered reference seawater, and equilibrated for 24 h. Organisms were then introduced for either a 3-day development test using the sand dollar Dendraster excentricus (modified from ref 19) or for a 10-day survival test using the amphipod Eohaustorius estuarius (20, 21). Ten individual amphipods and approximately 5000 sand dollar embryos per chamber were used, and tests were conducted at 16 °C. For all tests, both seawater and site water controls were run using four laboratory replicates each with aeration and no sediment present. Dissolved ammonia (Orion model 420a meter and ammonia electrode) and oxygen (Orion model 810 meter and dissolved oxygen electrode) were measured at the beginning and the end of the exposures. Un-ionized ammonia was calculated according to Bower and Bidwell (22). Statistical analyses followed EPA guidelines (20). XAS Data Collection and Analysis. X-ray absorption spectroscopy (XAS), including X-ray absorption near-edge structure (XANES) and extended X-ray absorption fine structure (EXAFS) analysis, were used to characterize the speciation and bonding of metals in sediments from cores (collected in July, 1997 and April, 1998). Fresh and freshfrozen sediment samples were examined with porewater present. Fluorescence spectral data were collected at Stanford Synchrotron Radiation Laboratory (SSRL) on wiggler beamlines 4-1 and 4-3 using either a 13-element Ge array detector, a four-element Ge array detector, or a Lytle detector. Spectra were collected at ambient temperature with the sample in a He atmosphere to prevent oxidation of sensitive elements. For a given element, energy was calibrated using a reference foil spectrum. EXAFS spectra were quantitatively analyzed using EXAFSPAK (23) and FEFF (24, 25) according to methods described in O’Day et al. (26). Absorption spectra for crystalline reference compounds and fresh precipitates for each element were collected and analyzed for comparison to sediment spectra. Details are given in Carroll et al. (17).
Results Sediment and Porewater Chemistry. Representative chemical composition and mineralogy of bulk sediments and composition of extracted porewaters from two Seaplane Lagoon cores (SC4 at 1.5- and 34.5-cm depth; SC-9 at 30-cm depth) are shown in Table 1. Trace metal concentrations in the sediments are 3-to-6 orders-of-magnitude higher than those measured in the extracted porewaters. Microelectrode measurements of intact cores showed removal of dissolved oxygen within the first 2-10 mm below the sediment-water interface (17). Dissolved sulfide is found in porewaters below about 30-40 mm, and concentrations measured in two cores (SC2 and SC4) vary from 23.9 to 43.5 mmol L-1 HS- ((17); representative analyses at 1.5 and 34.5 cm shown in Table 1). There were no seasonal variations in the depth range of suboxic and anoxic zones noted among cores taken in July, 1997, November, 1997, and April, 1998 (17). The estuary sediments are fine-grained and composed primarily of quartz (80-90%), with about 10-15 wt % in the clay-sized fraction
TABLE 1. Representative Chemical Composition and Mineralogy of Bulk Sediments and Element Concentrations in Extracted Porewaters at Two Depths in Seaplane Lagoon Core SC4 Collected on July 10, 1997 and One Depth in Core SC9 Collected on April 8, 1998 (Porewaters Were Not Extracted from Core SC9) from Carroll et al. (17) core depth (cm)
SC4-1
SC4-12
1.5
34.5
Sediment Concentrationsa (µmol g-1 dry wt)
Cd Cr Cu Fe Mn Ni Pb Zn
0.131 ( 0.016 4.67 ( 0.44 1.70 ( 0.12 788 ( 70 7.81 ( 0.73 2.04 ( 0.19 1.03 ( 0.19 4.39 ( 0.80
pH
7.35
1.71 ( 0.16 14.7 ( 0.9 3.30 ( 0.22 958 ( 69 7.68 ( 0.55 2.64 ( 0.19 6.13 ( 1.09 7.88 ( 1.29
SC9-12 30.0
0.543 ( 0.030 9.83 ( 0.39 6.64 ( 0.07 728 ( 7 5.02 ( 0.20 2.32 ( 0.27 5.66 ( 0.06 13.6 ( 0.1
Porewater Concentrations 7.39
(mol L-1) ClHS-
0.482 0.028
Cd Cr Cu Fe Mn Ni Pb Zn
0.0017 ( 0.0002 0.0102 ( 0.0012 0.0082 ( 0.0014 54 ( 1 7.19 ( 0.07 0.034 ( 0.004 0.0009 ( 0.0001 0.063 ( 0.0107
0.776 0.042
(mmol L-1)
0.0056 ( 0.0007 0.2958 ( 0.0365 0.0148 ( 0.0025
