«Contaminant Mass Balance for Sinclair and Dyes Inlets, Puget Sound, WA Eric Crecelius1, Robert K. Johnston2*, Jim Leather2, Joel Guerrero2, Martin ...»
Contaminant Mass Balance for Sinclair and Dyes Inlets, Puget Sound, WA
Eric Crecelius1, Robert K. Johnston2*, Jim Leather2, Joel Guerrero2, Martin Miller1, and Jill Brandenberger1
Battelle Marine Sciences Laboratory Space and Naval Warfare Systems Center
1529 West Sequim Bay Rd 53475 Strothe Road Code 2362
Sequim, WA 98382 San Diego, CA 92152-6310
*Author to whom correspondence should be addressed: email@example.com Abstract Sinclair Inlet and Dyes Inlets have historically received contaminates from military installations, industrial activities, municipal outfalls, and other nonpoint sources. For the purpose of determining a “total maximum daily load” (TMDL) of contaminants for the Inlets, a contaminant mass balance for the sediments is being developed. Sediment cores and traps were collected from depositional areas of the Inlets and surface sediment grabs were collected from fluvial deposits associated with major drainage areas into the Inlets. All sediment samples were screened using X-Ray fluorescence (XRF) for metals, and immunoassay for PAHs and PCBs. A subset of split-samples was analyzed using inductively coupled mass spectrometry (ICP/MS) for metals and gas chromatography mass spectrometry (GC/MS) for phthalates and PAHs, and GC electron capture detector (GC/ECD) for PCBs. Sediment cores were age-dated using radionuclides to determine the sedimentation rate and the history of sediment contamination. Streams and storm water outfalls were sampled in both the wet and dry seasons to assess loading from the watershed.
Seawater samples collected from the marine waters of the Inlets and boundary passages to central Puget Sound were used to estimate the exchange of contaminates with central Puget Sound. The historical trends from the cores indicate that contamination was at a maximum in the mid 1900s and decreased significantly by the late 1990s. The thickness of the contaminated sediment is in the range of 15 to 45 cm. The initial drafting of the mass balance model for copper indicates that approximately 4,000 Kg/yr enters the Inlets and 64% of this loading is from the leaching of boat and ship hulls. The removal of copper from the Inlets is primarily a function of sedimentation (68%) and seawater advection into Central Puget Sound (32%).
Introduction Sinclair and Dyes Inlets, located in the Puget Sound west of Seattle, Washington (Figure 1), have historically received contaminants from military installations, industrial activities, municipal outfalls, and other non-point sources. Both inlets host Navy facilities, with the largest being the Puget Sound Naval Shipyard (PSNS) and Naval Station on Sinclair Inlet. A little over hundred years ago, Sinclair Inlet and the Kitsap Peninsula were relatively undeveloped. The Navy established a base in Sinclair Inlet in 1891, in 1896 the Shipyard began operation, and in 1901 thetown of Bremerton was founded (The Sun 2001).
Rapid development in Bremerton and a boom in the population of Kitsap County followed major expansions at the Shipyard during World War I and World War II. At the height of World War II the population of Bremerton peaked at more than 80,000 people and industrial operations poured out goods for the war effort. Following the end of World War II, work at the Shipyard was reduced, but the Shipyard’s workload remained high throughout the cold war and into the 80s and 90s. In 1975 the Submarine Base at Bangor was established (Horn 1999) and since the late 70s Kitsap County has experienced rapid growth in population, infrastructure, and development of open space. Currently, about a quarter of a million people live in Kitsap County (U.S. Census Bureau 2001). In 1998, the Washington State Department of Ecology listed some of the sediments in both Inlets under Section 303(d) of the federal Clean Water Act because concentrations of mercury (Hg), copper (Cu), other metals, and organic chemicals exceeded sediment quality standards. To address environmental issues in the Inlets and surrounding watershed, PSNS Project ENVironmental InVEStment (ENVVEST, PSNS Project ENVVEST 2002) was initiated as a cooperative effort among PSNS, regulatory agencies, and local stakeholders to evaluate sources, assess the mass balance, and develop “total maximum daily loads” (TMDLs) of contaminants for the Inlets.
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12. DISTRIBUTION/AVAILABILITY STATEMENTApproved for public release; distribution unlimited
13. SUPPLEMENTARY NOTES Georgia Basin Puget Sound Research Conference, Vancouver, BC, Canada April 2, 2003
14. ABSTRACT see report
15. SUBJECT TERMS
In this paper we present the methods and results used to characterize historical loading and estimate the current mass balance of contaminants in the Inlets. The results obtained for Cu and Hg are used to illustrate the historical profile of contaminant loading in the estuary and the initial mass balance calculations for Cu are presented to demonstrate the technical approach. The complete set of findings and raw data will be published in a separate technical report (Crecelius et al. in prep).
Methods In April 2002, sediment cores were collected from depositional zones within the Inlets (Figure 1). These were age-dated to estimate the rate of removal of contaminants from the water column and to examine the historical trends in sediment contamination. Water and sediment samples were collected from streams entering the Inlets and were analyzed to determine loadings from different land use areas. Seawater samples from the Inlets were analyzed to estimate transport of contaminants between the Inlets and into the Central Puget Sound. Because the Inlets are hydrodynamicly connected, contaminants discharged into one are shared with the other through tidal mixing; therefore, both Inlets must be treated as one water body in terms of understanding environmental quality.
Sediment coring Sediment cores were collected at eight locations (Figure 1) using a piston corer supported in a frame and operated by SCUBA divers. This device recovered sediment cores with minimal disturbance or shortening compared with that caused by gravity coring. The core barrel had an inside diameter of 8.9 cm, a total length of 120 cm, and the length of recovered cores was about 100 cm each. The cores were kept upright during transportation and storage in a refrigerator. They were sectioned by extruding the sediment out the top of the core barrel using a piston inserted in the bottom of the core barrel. The top of the core (0-5 cm) was sectioned into two 2.5 cm sections and the remainder (5-100 cm) was sectioned into 5 cm sections.
Stream and Seawater Sampling Water samples were collected during two seasons (wet – March 2002, and dry – September 2002) from eight streams draining into the Inlets, and at eight marine water stations located within the Inlets and passages, which exchanges with Central Puget Sound (Figure 1). Water samples were collected for trace metals analysis in pre-cleaned Teflon bottles using ultra-clean sampling techniques, following the U.S.
Environmental Protection Agency (EPA) Method 1669 (EPA 1996a). All water samples were placed on ice in coolers and maintained at 4ºC while being transported to the laboratory. Half of each raw water sample was filtered through a 0.45-micron pre-cleaned filter in a Class 100 clean bench to produce the “dissolved” fraction.
Sediment Chemistry Screening A split of approximately 50 gm of each sediment sample was screened using XRF for Cu, lead (Pb), iron (Fe), and zinc (Zn); and immunoassay for PAHs and PCBs (Kirtay and Apitz 2000). For metals analysis 5 g of wet sediment were exposed to X-ray energy and the fluorescing spectrum was analyzed to quantify the metal ions present in the sample (U.S. EPA 1998). Approximately 10% of the samples analyzed using XRF were also analyzed using ICP-MS to validate the screening results.
Sediment Analyses The sediment samples were homogenized and then split for metals and organics analysis. The metal sample split was freeze-dried and the ratio of wet-to-dry sediment was determined as the fraction of solids for each sample. Each freeze-dried sample was homogenized by milling, acid-digested, and then analyzed for Cu by ICP-MS and for Hg by cold vapor atomic absorption spectroscopy (CVAA). Sediment splits for organics were solvent-extracted and analyzed by GC/MS for phthalates, PAHs, and GC/ECD for PCBs (data not shown). Sediment digestion, extraction, and analysis methods for metals and organic compounds are provided in Lauenstein and Cantillo (1998).
Radionuclides Analysis To determine sedimentation rates and the chronology of sediment contamination, subsamples of sediment core sections were analyzed for lead-210 ( 210Pb) and cesium-137 (137Cs). To prepare the sediment for alpha counting, about 3 g dry sediment was acid-digested, and polonium-210 (210Po) was plated on a silver disk. The activity of 210Pb was determined by counting the granddaughter 210Po alpha particle on a silicon barrier diode detector, similar to the method of Koide et al. (1973). The activity of 137Cs was determined by gamma-counting approximately 50 g dry sediment, using a germanium diode detector.
Sedimentation Rate Sedimentation (cm/yr) and deposition (gm/cm2/yr) rates were estimated using a steady-state 210Pb dating technique (Lavelle et al.1986). This method assumes that 1) sedimentation rate is constant; 2) loss of 210Po occurs only by radioactive decay; and 3) mixing is confined to the surface mixed layer. The year of deposition for the midpoint of each core section was calculated. The relationship between sediment wet density and sediment dry weight was used to calculate the deposition rate. The location of the subsurface Cs peak in each core was used to evaluate the accuracy of the 210Pb core dating technique.
Water Analyses Ultra-clean trace metal sampling and analysis techniques were followed for metals in water samples (U.S.
EPA 1996a, b, c). All water samples were analyzed for Hg using cold vapor atomic fluorescence spectrometry (CVAF) in accordance with EPA Method 1631, Revision E (EPA 2002). Freshwater samples for all other metals (e.g., Cu for this paper) were analyzed using ICP-MS, in accordance with EPA Method 1638 (EPA 1996b). Seawater samples were preconcentrated using Fe and palladium (Pd) prereduction, in accordance with the Battelle Marine Sciences Laboratory (BMSL) standard operating procedure (SOP) for seawater preconcentration, which is derived from modifications to EPA Method 1640 (EPA 1996c). The preconcentrated samples were then analyzed by ICP-MS.
Results Sediment Screening Results The analytical results obtained for Cu, Pb, and Zn showed good agreement between XRF and ICP-MS (Figure 2). The rapid estimates of contaminant concentrations from the screening analysis allowed the more expensive, low-detection-level chemistry analysis by ICP-MS to be focused on selected samples. This was particularly useful for the core subsections, because preindustrial background values for Pb were identified in the screening methods, and subsections deeper than those showing stabilization of Pb at background levels were not submitted for ICP-MS analysis.
Figure 2. The relationship between Cu, Pb, and Zn measured by XRF and ICP-MS obtained for sediment samples.
Temporal Trends from Dated Cores The historical trends from the cores indicate an increase in the contaminant levels of Hg, Cu, PAHs, and PCB beginning around 1900. The significant increase in these contaminants peaked between 1940 and
1960. The results obtained for Cu and Hg are used to illustrate the historical profile determined from the data (Figures 3 and 4). The thickness of the layer of contaminated sediment ranged from 15 cm to 50 cm.
These trends, or peaks and declines, of contaminants in the Sinclair and Dyes Inlet cores accurately reflect the known history of both the uses of these contaminants and environmental regulations. In the 1970s, Congress created the Environment Protection Agency (EPA) and enacted environmental regulations to control pollution sources. The enforcement of environmental laws is reflected in the core profiles, which shows a significant decline in the later part of the 20th Century. The average sedimentation rate for the eight core profiles was estimated to be 0.115 ± 0.051 g/cm2/y; with the average thickness of the annually deposited sediment estimated to be 0.25 ± 0.12 cm/yr for the Inlets.
Historical uses of Hg date far earlier than the 1930s; therefore, the initial increase of Hg is more diffuse in the cores than the profiles for the other chemicals. The peak in Hg for both Sinclair and Dyes Inlets occurred in the 1940s (Figure 3), corresponding to the escalation of industrial activities associated with World War II. However, the reduction of Hg used by the Navy and the passage of regulations designed to minimize Hg discharges are also reflected in the core profiles, in which Hg levels decrease dramatically after the peak in the 1940s. Present-day sediment quality is improving; however, Hg is still slightly above the Washington State sediment quality standard (SQS) of 0.41 µg/g.