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Mercury deposition in a tidal marsh of south San Francisco Bay downstream of the historic New Almaden mining district, California
Marine Chemistry 90 (2004) 175 – 184 www.elsevier.com/locate/marchem
Mercury deposition in a tidal marsh of south San Francisco Bay downstream of the historic New Almaden mining district, California Christopher H. Conaway a,*, Elizabeth B. Watson b, John R. Flanders a, A. Russell Flegal a a
WIGS Laboratory, Department of Environmental Toxicology, University of California at Santa Cruz, 1156 High Street, Santa Cruz, CA 95064, USA b Department of Geography, University of California at Berkeley, 507 McCone Hall, Berkeley, CA 94720, USA Received 18 June 2003; received in revised form 24 February 2004; accepted 24 February 2004 Available online 7 June 2004
Abstract A record of mercury deposition was provided by sediment recovered from piston cores of a south San Francisco Bay tidal marsh that is 30 km downstream of the New Almaden mining district, formerly the largest mercury mining district in North America. Pre-mining sediment mercury concentrations were 0.40 F 0.15 nmol g 1, which are similar to pre-mining concentrations in cores taken from other parts of San Francisco Estuary. Concentrations in the core increase to a maximum of about 6.0 nmol g 1, corresponding to a period in the mid-20th century, nearly 50 years after the peak in mercury production at the mines. The extent of contamination from upstream mining activity appears to reflect the amount of processed ore disposed of at the surface, and also periods when mercury was recovered from reworking these surface ore dumps and open cuts. Transport of this contaminant mercury to the tidal marsh appears to be influenced by hydrologic modifications in the watershed, including dam building and subsidence related to groundwater withdrawal. Although San Francisco Estuary is contaminated with mercury from numerous historic mining sources, including late 19th century hydraulic gold mining in the Sierra Nevada, there is little evidence of pre-mining contamination from natural mercury sources in the southern reach of San Francisco Estuary. D 2004 Elsevier B.V. All rights reserved. Keywords: Mercury; Cores; San Francisco Bay; New Almaden; Tidal marsh
1. Introduction In any study of metal contamination, it is necessary to establish a background concentration in order to be
* Corresponding author. Tel.: +1-831-4595336. E-mail address: [email protected] (C.H. Conaway). 0304-4203/$ - see front matter D 2004 Elsevier B.V. All rights reserved. doi:10.1016/j.marchem.2004.02.023
quantitative. Resolution of that concentration is especially important in highly mineralized areas, because natural weathering of mineral deposits is potentially an important source of contamination. This is a concern for mercury in San Francisco Estuary (Fig. 1), which is not only impacted by historic gold and mercury mining, but also receives water from the highly mercury-mineralized rocks in the surrounding California Coast Range.
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Fig 1. Sample location map.
The California Coast Range hosts the New Idria and New Almaden deposits, as well as many smaller deposits, making it one of the most productive mercury belts in the world (Cargill et al., 1980). The California Coast Range mineral belt is part of the circum-Pacific mercury belt, and the deposits in this larger belt are formed by hydrothermal activity related to active plate tectonic margins. The California Coast Range mercury deposits are characteristically silicacarbonate type hosted in serpentinite that developed in the San Andreas fault system (Rytuba, 2000). These serpentinite bodies served to trap carbon dioxide-rich hydrothermal fluids (Rytuba, 2003), which are known to transport and concentrate mercury in hydrothermal systems (Fein and Williams-Jones, 1997). The primary mercury mineral in these deposits is cinnabar (Rytuba, 2003). The physical and chemical weathering of these cinnabar deposits provides a natural source of elevated mercury fluxes to the estuary. In addition this potential for natural enrichment, San Francisco Estuary is contaminated by mercury
from historic mining activity throughout its watershed. Mercury used in hydraulic mining to extract gold in the Sierra Nevada between 1852 and 1884 resulted in billions of cubic meters of mercury contaminated sediment, much of which was deposited in the northern reach of the estuary (Gilbert, 1917; Hedgpeth, 1979; Hornberger et al., 1999; MarvinDiPasquale et al., 2003). The New Almaden mining district was the largest mercury mining district in North America, and produced 190 Mmol of mercury (Cargill et al., 1980; Bailey et al., 1973). This district is located in the Santa Clara Valley and is only 30 km from San Francisco Estuary (Fig. 1). The mines in this highly mineralized valley drain into Guadalupe River and Coyote Creek, which flow into the southeasterly reach of the estuary. As a result of the historic and contemporary inputs of mercury to the estuary, concentrations often exceed water quality objectives and advisory levels for the consumption of fish (Davis, 1999; Thompson et al., 2000). Numerous studies suggest that mercury has an
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effect on the biota of the estuary, particularly on birds (Hoffman et al., 1998; Hothem et al., 1995, 1998; Hui, 1998; Hui et al., 2001). However, efforts to quantify the impact of anthropogenic fluxes of mercury to the estuary have been hampered by questions of what the natural levels were in this mercuryrich region. Consequently, to study the impact of mining activity and the relative contribution of natural weathering in the estuary, we investigated mercury concentrations in a piston core taken from a tidal marsh near where Coyote Creek and Guadalupe River enter San Francisco Bay. In addition, 3 m deep gravity cores of estuarine sediment were collected and analyzed from several locations in south San Francisco Bay. Mercury data from these samples establish a value for the pre-mining concentration for mercury in the southern reach, and they also show the effect of mercury mining at New Almaden on sediments being transported to the tidal marshes and into the estuary.
2. Methods 2.1. Sampling In December 2000, three 3-m gravity cores were collected from the estuary by the R/V David Johnston at Oyster Point and near the San Mateo Bridge. These cores were split into 1-m subsections on deck, and subsampled for mercury analysis and organic material suitable for radiocarbon dating. In addition, replicate 3-m piston cores were collected in October 2001 and September 2002 from Triangle Marsh in the Don Edwards San Francisco Bay National Wildlife Refuge, which is located in the southern reach of the estuary (Fig. 1). One core was used for chronology, X-radiography, magnetic susceptibility, pollen and X-ray fluorescence analysis (XRF). Twenty-seven samples were analyzed with a Phillips PW 2400 XRF spectrometer to determine bulk elemental composition. A duplicate core was collected in five subsections and immediately refrigerated for later mercury analysis. The core subsections were split and sampled at about 10-cm intervals. For all cores, samples were collected from the radial center to minimize contamination from the core walls and the coring
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process. In order to avoid material disturbed by coring and subsectioning, the ends of the subsections were not sampled. Samples analyzed for mercury were sieved through a 63-Am nylon mesh to minimize grain-size effects, freeze dried and stored in acid-cleaned high-density polyethylene bottles at 20jC until analysis. 2.2. Mercury analysis Sediment total mercury (HgT) analyses were conducted by digesting sediment (about 0.3 g) in boiling concentrated HNO3/H2SO4, followed by 12-h oxidation with bromine monochloride, tin chloride reduction, gold amalgamation and detection by CVAFS (Bloom and Crecelius, 1987; Gill and Fitzgerald, 1987; Mason and Lawrence, 1999). Multiple analyses (n = 6) of PACS-2, a polluted marine sediment certified reference material from the National Research Council of Canada, yielded an average (x F S.D.) concentration of 14.5 F 1.4 nmol g 1 (certified value 15.2 F 1.0 nmol g 1). Analytical detection limits, calculated using three times the standard deviation of the procedural blanks (n = 13), averaged about 0.10 pmol. All sediment digestion blanks were below the detection limit, and therefore the overall method detection limit (MDL) was about 0.50 pmol g 1. The analytical precision of duplicate analyses (n = 3) of a PACS-2 digestate was 3.4 F 3.5%. The precision of duplicate sample analyses (n = 19) was 9.5 F 5.9%. 2.3. Chronology A chronological framework was developed for both the Triangle Marsh and south San Francisco Bay cores using AMS radiocarbon dating. Shell and organic material was collected from the gravity and piston cores and analyzed for radiocarbon at Beta Analytic and Lawrence Livermore National Laboratory. Prior to analysis, shell material was rinsed in deionized water and etched with HCl to eliminate secondary carbonate components. Organic material was sequentially rinsed in HCl, NaOH and HCl to remove carbonates and secondary organic acids. Measured radiocarbon ages for shell material were corrected for a local reservoir age of 625 years (Stuiver and Braziunas, 1993). Although there is a great deal
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of uncertainty in the age of the sediment from the cores in San Francisco Bay, the results of this dating are primarily intended to establish that the sediment is from the pre-mining era, i.e., older than about 250 years. In general, the results were consistent with those described by Ingram et al. (1996) in a core from Oyster Point (location similar to OP-5, this study). In addition to radiocarbon analyses, the chronology of the cores from Triangle Marsh were further established by the first appearance of pollen from non-native taxa Eucalyptus and Plantago lanceolata (narrowleaved plantain). Subsamples for pollen analysis were collected at intervals of 6 –10 cm, and prepared following standard pollen extraction procedures (Faegri and Iversen, 1975). A complete discussion of the pollen dating can be found in Watson (in press).
3. Results and discussion 3.1. Triangle Marsh core general features Mercury concentrations in the Triangle Marsh core vary systemically throughout the length of the core (Fig. 2). Below 150 cm, HgT concentrations are under 1.0 nmol g 1 and show little variation with depth. Between 150 and 70 cm, HgT concentrations increase steadily, from below 1.0 to a peak of over 6.0 nmol g 1. The upper 50 cm are characterized by HgT concentrations between 2.0 and 3.0 nmol g 1. Relating core depth to chronological constraints provided by pollen and radiocarbon dating shows that mercury concentrations began to increase in the later part of the 19th century and reached a maximum in the mid-20th century. The profile of mercury concentrations in this
Fig 2. Results for sediment analyses ( < 63 Am) of Triangle Marsh piston core (location N37 27.323, W121 58.533) total mercury (nmol g 1 F 1 standard deviation) versus depth (cm). Chronology of the Triangle Marsh core is described in detail in Watson (in press).
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core reflects a combination of changes in core lithology, the history of mining in the region and hydrologic factors affecting the transport of contaminant mercury through the upstream watershed.
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suggesting that there has been relatively little diagenetic remobilization. 3.3. Pre-mining mercury concentrations in Triangle Marsh and San Francisco Estuary
3.2. Triangle Marsh core lithology The lowest section of the core, below 250 cm, is characterized by clay-rich sediments that were deposited in an intertidal mudflat environment. The slight increase in mercury concentrations between 200 and 250 cm is coincident with a clayey peat layer that represents a transition from intertidal mudflat to tidal marsh. Sediment between 150 and 200 cm has slightly lower mercury concentrations than the underlying sediments, and represents tidal marsh sediment that was deposited before major anthropological impacts to Triangle Marsh. Sediment between 0 and 150 cm, which includes the large peak in mercury concentration, is characterized by peat and clayey peat, and was deposited during a period of high sedimentation related to local subsidence and increased sediment supply (discussed below). Variations of organic carbon in the sieved sediment reflect the changes in lithology from clay-rich intertidal sediment to peat and have no significant linear correlation to mercury concentration (r2 = 0.01, p = 0.7). The potential for diagenetic remobilization of metals in the Triangle Marsh core was investigated using iron, aluminum, and manganese data from the XRF analysis. Previous studies (Fitzgerald et al., 1998; Varekamp et al., 2003) have used the correlation between these metals as evidence of post-depositional mobility of redox sensitive elements, including mercury. If the redox active metals, iron and manganese, covary independent of aluminum, then diagenetic changes due to redox conditions may have occurred, potentially affecting mercury concentrations. In the Triangle Marsh core, there is a significant linear correlation (r2 = 0.23; p = 0.01) between iron and manganese concentrations, however they are not closely related. Although the range of aluminum concentrations in the core is large (0.1 – 0.6 mol g 1), the linear relationship between aluminum and silica ( p < 0.005, r2 = 0.98) shows that this is due to apparent changes in the abundance of aluminosilicate minerals in the core. Normalizing aluminum to silica results in an aluminum profile similar to that of iron,
The concentrations observed in the pre-mining sediment record throughout the estuary are similar. At Triangle Marsh, the average pre-mining HgT concentration, taken from below a depth of 150 cm, is 0.40 F 0.15 nmol g 1. The average of concentrations observed in depths below 1 m in the San Mateo and Oyster Point cores (Table 1) is 0.35 F 0.05 nmol g 1. These values are similar to concentrations measured by Hornberger et al. (1999) in deep-cores from the northern reach of the estuary (0.30 F 0.05 nmol g 1). In addition, Domagalski (2001) reported concentrations of 0.25 – 0.50 nmol g 1 in stream sediments from the northern Coast Range with no known mercury mineral deposits. This relative homogeneity suggests the contribution from natural weathering of mercury-rich deposits in the Santa Clara Valley to mercury concentrations in San Francisco Estuary is not substantial. Overall, these pre-mining concentrations are within the ranges reported for normal soil and sediment concentrations (Adriano, 2001; Faure, 1998). Consequently, the elevated levels observed in historic and modern sediments in the estuary are predominantly related to anthropogenic activity. The historical background value for San Francisco Estuary shows striking similarity to values obtained Table 1 Core locations, subsample depths, concentration of total mercury (HgT) in nmol g 1 in sieved ( < 63 Am) and freeze-dried sediment, and approximate age Sample ID and location
Depth (cm)
HgT (nmol g 1)
San Mateo SM-4 N37 35.917V W122 15.268V
71 143 242 270 133 280 226 232
0.64 0.37 0.25 0.36 0.30 0.38 0.42 0.43
Oyster Point OP-5 N37 39.014V W122 16.835V Oyster Point OP-6 N37 40.395V W122 21.049V
Approximate age (year) 1500 1600 4000 4300 3600 2200
Approximate ages from AMS 14C dating of shell and organic material by Beta Analytic. Dating of material from these cores is primarily intended to establish that the sediment is from the premining era, i.e., older than about 250 years.
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for background levels in other near shore marine environments. A core collected from the upper estuary of the Tinto River in the Aznalcollar mining area of southwestern Spain revealed background values in bulk sediments between approximately 0.10 and 0.35 nmol g 1 (Leblanc et al., 2000). The background values for bulk sediments in the Gulf of Trieste, which is located downstream of the Idrija mine, have been reported as 0.50 nmol g 1 (Faganeli et al., 1991), although Covelli et al. (2001) estimate a higher background of 0.85 nmol g 1. Background values for bulk sediment from salt marshes in Long Island Sound, and area with no known mercury mining history or mercury mineral deposits range between 0.17 and 0.46 nmol g 1 (Varekamp et al., 2003). Although the considerable differences in core lithologies and background geology prevent any major conclusions from these comparisons, the similarity between the background mercury concentrations in estuaries downstream of heavily mineralized areas and areas with no known mineral deposits is surprising. 3.4. Mining history and downstream concentrations The elevated mercury concentrations in sediments from south San Francisco Bay (Conaway et al., 2003; David et al., 2002) and the gradient from these sediments to the New Almaden mines (Thomas et al., 2002) demonstrate that the mines are the principal source of mercury contamination to the southern reach of the estuary. Large-scale development of the New Almaden district began in about 1845 and continued until 1975, producing over 190 Mmol of mercury (Cargill et al., 1980). The concentrations of HgT in the Triangle Marsh core start to increase shortly before 1870, roughly corresponding to the dramatic increase in processed ore at New Almaden, which reached 500 Mmol moles in 1880 (Cargill et al., 1980). The amount of processed ore decreased substantially between 1910 and 1940, but underwent a brief renaissance in the 1940s, when mercury was primarily recovered from reworking surface ore dumps and open cuts. Previous workers have demonstrated that exposed mercury mine tailings can be the primary source of mercury to downstream watersheds (Ganguli et al., 2000; Rytuba, 2000). The amount of processed ore dumped on the surface and the reworking of surface dumps is therefore expected to have a
large impact on downstream mercury concentrations. The history of mercury deposition in the core appears to follow this prediction, as there appears to be a substantial increase in mercury contamination related to this mid-20th century mining period. Production at New Almaden dropped dramatically in the latter half of the 20th century, and this is reflected in the marsh sediments. During this period, the mercury concentration decreases from its maximum in the core to several nmol g 1, consistent with values for modernday surface sediments in the southern reach, which range from 1.0 to 5.0 nmol g 1 (Conaway et al., 2003; David et al., 2002; Thomas et al., 2002). Because the timing of the increase in mining activity at New Almaden is concurrent with the period of hydraulic mining in the Sierra Nevada, it is necessary to consider the possible contribution of mercury contamination from hydraulic mining sediments discharged into the northern reach being dispersed into the southern reach of the estuary. Some sediment from the Sacramento-San Joaquin river drainage basin becomes deposited in the southern reach (Atwater et al., 1979; Krone, 1979). This proposed transport is supported by analyses of lead isotopes in both surface sediments and waters (Ritson et al., 1999; Steding et al., 2000). Sediment budgets for the estuary, however, suggest that the amount of sediment transported in this manner is relatively small (Krone, 1979). The contribution of contaminated hydraulic mining sediment to the southern reach is consequently considered to be of relatively little consequence to mercury concentrations in that region. 3.5. Hydrologic factors affecting mercury concentrations at Triangle Marsh Although the dominant source of mercury to Triangle Marsh is New Almaden, the transport of this contaminant mercury has been influenced by many factors affecting sediment sources and transport in the watershed. The most obvious evidence of this is the apparent delay between the time of highest mercury production at New Almaden and the peak mercury concentrations in the core (Fig. 3). A similar delay is observed between the 1880s peak of hydraulic mining activity in the Sierra Nevada and the deposition of the mercury-contaminated hydraulic mining sediments in the northern reach of the estuary in the early part of
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Fig 3. Temporal profile of mercury concentrations (nmol g 1) in Triangle Marsh piston core (location N37 27.323, W121 58.533). Treated ore and mercury production figures (Mmol) for New Almaden mines are taken from Cargill et al. (1980). Subsidence data (cm year 1) for the Santa Clara Valley (Benchmark P7, San Jose) are taken from Poland (1984).
the 20th century (Hedgpeth, 1979; Hornberger et al., 1999). Generally, the transport of mercury through the Santa Clara Valley watershed occurs in episodic, peak-flow events (Thomas et al., 2002). The rivers and sloughs in the valley and southern reach have undergone many changes, including both natural and artificial channel realignments—the Guadalupe River in particular has been rerouted several times. In addition, urbanization in the watershed, especially in the city of San Jose, has replaced much of the natural surface with concrete and asphalt, which may favor erosion of contaminated sediment. The building of several major water projects constructed in the upper watershed during the early to mid-20th century, including the Almaden, Andersen, Coyote and Guadalupe Reservoirs, probably also had a major effect on the transport of sediment through the watershed. Another effect of urbanization and water use in the
Santa Clara Valley is local subsidence due to groundwater pumping; between 1934 and 1967 San Jose subsided as much as 3 m (Fig. 3), and Triangle Marsh about 1 m (Poland and Ireland, 1988). All of these changing land use factors likely had a profound impact on the sources and transport of sediment to Triangle Marsh and also on sedimentation at the marsh (Patrick and Delaune, 1990; Watson, in press). 3.6. Mercury accumulation at Triangle Marsh The impact on mercury concentrations by increasing sedimentation rates and supply of contaminated sediment can also be illustrated by estimating the rate of mercury deposition in Triangle Marsh. The mass of the sediment accumulating in g cm 2 year 1 can be calculated as the product of the bulk density of the sediment (g cm 3) and the sediment accumulation
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rate (cm year 1), reported by Watson (in press). Multiplying this result by the molar concentration of HgT (nmol g 1) in the sediment yields the rate of mercury deposition (nmol cm 2 year 1).
mercury to the environment, the subsequent transport of mercury downstream appears to have been predominantly controlled by both the style of mining and changes in the hydrology of the watershed. The relationship between the many hydrological factors mentioned above and the downstream mercury concentrations is complex, and may not be resolvable into one simple relationship. Nevertheless, it is clear that elevated mercury concentrations in south San Francisco Bay are the result of erosion from mercury mine wastes from mercury mines in the watershed, and the contribution of natural weathering in this system is of little importance.
Accumulation Rate ¼ qðsÞHgT where q is equal to sediment bulk density in g cm 3, s is equal to the sediment accumulation rate in cm year 1 and HgT is equal to total mercury concentration in nmol g 1. Although several simplifying strategies were used in this estimation, such as using average values for sedimentation rates and bulk densities, and assuming that the 63 Am fraction makes up the majority of the sediment, the differences shown are on an order of magnitude scale, and should thus be fairly realistic. The results of this calculation (Table 2) show a dramatic increase in mercury accumulation rates during the mining era, particularly the period 1945 – 1975. This temporal increase suggests that the deposition of mercury downstream of the mines is driven principally by mercury mining upstream and changes in hydrology in the watershed, rather than by the amount of mercury produced by the mines.
Acknowledgements The authors are indebted to the pioneering studies of Bill Fitzgerald, who demonstrated the need for rigorous analytical techniques to accurately measure baseline mercury concentrations and to correctly quantify anthropogenic perturbations of mercury’s biogeochemical cycle. We thank Elizabeth Kerin, Kim Bracchi, Miranda Spang, Allison Luengen, Liam Reidy, Brenda Hamilton, Martha Evonuk, Glen Spinelli, Rob Franks and the crew of the R/V David Johnston for their assistance in collecting and analyzing cores. Access to the Don Edwards San Francisco Bay National Wildlife Refuge was approved by Clyde Morris with the U. S. Fish and Wildlife Service. Comments on the research and manuscript were provided by Richard Looker, Tom Grieb, K.E. Abu-Saba, Roberto Anima, John Callaway, P.J. Lechler, Ken Bruland, David Sedlak, R.P. Mason and an anonymous reviewer. This work was funded by a contract with the San Francisco Bay
4. Conclusions This study describes the depositional history of mercury downstream from one of the largest mercury mining districts in the world. Although the watershed is rich in mercury mineral deposits, there appears to be little evidence of contamination downstream of the New Almaden area due to natural weathering. Although the mining era provided an immense source of
Table 2 Mercury accumulation rate calculated for depth intervals in sediment core from Triangle Marsh Depth (cm)
Period
Sedimentation rate (cm year 1)
Bulk density (g cm 3)
HgT (nmol g 1)
Mercury accumulation rate (nmol cm 2 year 1)
0–8 8 – 35 35 – 40 40 – 125 125 – 140 140 – 280
1998 – 2001 1982 – 1998 1975 – 1982 1945 – 1975 1870 – 1945 1570 – 1870
2.7 1.8 0.71 2.8 0.18 0.33
0.8 0.7 0.7 0.6 0.6 0.7
2.3 2.4 2.7 4.2 2.6 0.41
5.0 3.1 1.3 7.0 0.28 0.09
The mercury accumulation rate is the product of sedimentation rate, bulk density of the sediment, and the concentration of mercury (HgT). The sedimentation rate and bulk density values are averages from Watson (in press). Mercury concentrations are averages for each depth interval.
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Steding, D.J., Dunlap, C.E., Flegal, A.R., 2000. New isotopic evidence for chronic lead contamination in the San Francisco Bay estuary system: implications for the persistence of past industrial lead emissions in the biosphere. Proceedings of the National Academy of Sciences of the United States of America 97 (21), 11181 – 11186. Stuiver, M., Braziunas, T.F., 1993. Modeling atmospheric C-14 influences and C-14 ages of marine samples to 10,000 BC. Radiocarbon 35 (1), 137 – 189. Thomas, M.A., Conaway, C.H., Steding, D.J., Marvin-DiPasquale, M., Abu-Saba, K.E., Flegal, A.R., 2002. Mercury contamination from historic mining in water and sediment, Guadalupe River and San Francisco Bay, California. Geochemistry: Exploration, Environment, Analysis 2 (3), 211 – 217. Thompson, B., Hoenicke, R., Davis, J.A., Gunther, A., 2000. An overview of contaminant-related issues identified by monitoring in San Francisco Bay. Environmental Monitoring and Assessment 64 (1), 409 – 419. Varekamp, J.C., Kreulen, B., ten Brink, M.R.B., Mecray, E.L., 2003. Mercury contamination chronologies from Connecticut wetlands and Long Island Sound sediments. Environmental Geology 43 (3), 268 – 282. Watson, E.B., in press. Changes in elevation, accretion, and tidal marsh plant assemblages in a South San Francisco Bay tidal marsh. Estuaries.
Mercury deposition in a tidal marsh of south San Francisco Bay downstream of the historic New Almaden mining district, California Christopher H. Conaway a,*, Elizabeth B. Watson b, John R. Flanders a, A. Russell Flegal a a
WIGS Laboratory, Department of Environmental Toxicology, University of California at Santa Cruz, 1156 High Street, Santa Cruz, CA 95064, USA b Department of Geography, University of California at Berkeley, 507 McCone Hall, Berkeley, CA 94720, USA Received 18 June 2003; received in revised form 24 February 2004; accepted 24 February 2004 Available online 7 June 2004
Abstract A record of mercury deposition was provided by sediment recovered from piston cores of a south San Francisco Bay tidal marsh that is 30 km downstream of the New Almaden mining district, formerly the largest mercury mining district in North America. Pre-mining sediment mercury concentrations were 0.40 F 0.15 nmol g 1, which are similar to pre-mining concentrations in cores taken from other parts of San Francisco Estuary. Concentrations in the core increase to a maximum of about 6.0 nmol g 1, corresponding to a period in the mid-20th century, nearly 50 years after the peak in mercury production at the mines. The extent of contamination from upstream mining activity appears to reflect the amount of processed ore disposed of at the surface, and also periods when mercury was recovered from reworking these surface ore dumps and open cuts. Transport of this contaminant mercury to the tidal marsh appears to be influenced by hydrologic modifications in the watershed, including dam building and subsidence related to groundwater withdrawal. Although San Francisco Estuary is contaminated with mercury from numerous historic mining sources, including late 19th century hydraulic gold mining in the Sierra Nevada, there is little evidence of pre-mining contamination from natural mercury sources in the southern reach of San Francisco Estuary. D 2004 Elsevier B.V. All rights reserved. Keywords: Mercury; Cores; San Francisco Bay; New Almaden; Tidal marsh
1. Introduction In any study of metal contamination, it is necessary to establish a background concentration in order to be
* Corresponding author. Tel.: +1-831-4595336. E-mail address: [email protected] (C.H. Conaway). 0304-4203/$ - see front matter D 2004 Elsevier B.V. All rights reserved. doi:10.1016/j.marchem.2004.02.023
quantitative. Resolution of that concentration is especially important in highly mineralized areas, because natural weathering of mineral deposits is potentially an important source of contamination. This is a concern for mercury in San Francisco Estuary (Fig. 1), which is not only impacted by historic gold and mercury mining, but also receives water from the highly mercury-mineralized rocks in the surrounding California Coast Range.
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Fig 1. Sample location map.
The California Coast Range hosts the New Idria and New Almaden deposits, as well as many smaller deposits, making it one of the most productive mercury belts in the world (Cargill et al., 1980). The California Coast Range mineral belt is part of the circum-Pacific mercury belt, and the deposits in this larger belt are formed by hydrothermal activity related to active plate tectonic margins. The California Coast Range mercury deposits are characteristically silicacarbonate type hosted in serpentinite that developed in the San Andreas fault system (Rytuba, 2000). These serpentinite bodies served to trap carbon dioxide-rich hydrothermal fluids (Rytuba, 2003), which are known to transport and concentrate mercury in hydrothermal systems (Fein and Williams-Jones, 1997). The primary mercury mineral in these deposits is cinnabar (Rytuba, 2003). The physical and chemical weathering of these cinnabar deposits provides a natural source of elevated mercury fluxes to the estuary. In addition this potential for natural enrichment, San Francisco Estuary is contaminated by mercury
from historic mining activity throughout its watershed. Mercury used in hydraulic mining to extract gold in the Sierra Nevada between 1852 and 1884 resulted in billions of cubic meters of mercury contaminated sediment, much of which was deposited in the northern reach of the estuary (Gilbert, 1917; Hedgpeth, 1979; Hornberger et al., 1999; MarvinDiPasquale et al., 2003). The New Almaden mining district was the largest mercury mining district in North America, and produced 190 Mmol of mercury (Cargill et al., 1980; Bailey et al., 1973). This district is located in the Santa Clara Valley and is only 30 km from San Francisco Estuary (Fig. 1). The mines in this highly mineralized valley drain into Guadalupe River and Coyote Creek, which flow into the southeasterly reach of the estuary. As a result of the historic and contemporary inputs of mercury to the estuary, concentrations often exceed water quality objectives and advisory levels for the consumption of fish (Davis, 1999; Thompson et al., 2000). Numerous studies suggest that mercury has an
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effect on the biota of the estuary, particularly on birds (Hoffman et al., 1998; Hothem et al., 1995, 1998; Hui, 1998; Hui et al., 2001). However, efforts to quantify the impact of anthropogenic fluxes of mercury to the estuary have been hampered by questions of what the natural levels were in this mercuryrich region. Consequently, to study the impact of mining activity and the relative contribution of natural weathering in the estuary, we investigated mercury concentrations in a piston core taken from a tidal marsh near where Coyote Creek and Guadalupe River enter San Francisco Bay. In addition, 3 m deep gravity cores of estuarine sediment were collected and analyzed from several locations in south San Francisco Bay. Mercury data from these samples establish a value for the pre-mining concentration for mercury in the southern reach, and they also show the effect of mercury mining at New Almaden on sediments being transported to the tidal marshes and into the estuary.
2. Methods 2.1. Sampling In December 2000, three 3-m gravity cores were collected from the estuary by the R/V David Johnston at Oyster Point and near the San Mateo Bridge. These cores were split into 1-m subsections on deck, and subsampled for mercury analysis and organic material suitable for radiocarbon dating. In addition, replicate 3-m piston cores were collected in October 2001 and September 2002 from Triangle Marsh in the Don Edwards San Francisco Bay National Wildlife Refuge, which is located in the southern reach of the estuary (Fig. 1). One core was used for chronology, X-radiography, magnetic susceptibility, pollen and X-ray fluorescence analysis (XRF). Twenty-seven samples were analyzed with a Phillips PW 2400 XRF spectrometer to determine bulk elemental composition. A duplicate core was collected in five subsections and immediately refrigerated for later mercury analysis. The core subsections were split and sampled at about 10-cm intervals. For all cores, samples were collected from the radial center to minimize contamination from the core walls and the coring
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process. In order to avoid material disturbed by coring and subsectioning, the ends of the subsections were not sampled. Samples analyzed for mercury were sieved through a 63-Am nylon mesh to minimize grain-size effects, freeze dried and stored in acid-cleaned high-density polyethylene bottles at 20jC until analysis. 2.2. Mercury analysis Sediment total mercury (HgT) analyses were conducted by digesting sediment (about 0.3 g) in boiling concentrated HNO3/H2SO4, followed by 12-h oxidation with bromine monochloride, tin chloride reduction, gold amalgamation and detection by CVAFS (Bloom and Crecelius, 1987; Gill and Fitzgerald, 1987; Mason and Lawrence, 1999). Multiple analyses (n = 6) of PACS-2, a polluted marine sediment certified reference material from the National Research Council of Canada, yielded an average (x F S.D.) concentration of 14.5 F 1.4 nmol g 1 (certified value 15.2 F 1.0 nmol g 1). Analytical detection limits, calculated using three times the standard deviation of the procedural blanks (n = 13), averaged about 0.10 pmol. All sediment digestion blanks were below the detection limit, and therefore the overall method detection limit (MDL) was about 0.50 pmol g 1. The analytical precision of duplicate analyses (n = 3) of a PACS-2 digestate was 3.4 F 3.5%. The precision of duplicate sample analyses (n = 19) was 9.5 F 5.9%. 2.3. Chronology A chronological framework was developed for both the Triangle Marsh and south San Francisco Bay cores using AMS radiocarbon dating. Shell and organic material was collected from the gravity and piston cores and analyzed for radiocarbon at Beta Analytic and Lawrence Livermore National Laboratory. Prior to analysis, shell material was rinsed in deionized water and etched with HCl to eliminate secondary carbonate components. Organic material was sequentially rinsed in HCl, NaOH and HCl to remove carbonates and secondary organic acids. Measured radiocarbon ages for shell material were corrected for a local reservoir age of 625 years (Stuiver and Braziunas, 1993). Although there is a great deal
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of uncertainty in the age of the sediment from the cores in San Francisco Bay, the results of this dating are primarily intended to establish that the sediment is from the pre-mining era, i.e., older than about 250 years. In general, the results were consistent with those described by Ingram et al. (1996) in a core from Oyster Point (location similar to OP-5, this study). In addition to radiocarbon analyses, the chronology of the cores from Triangle Marsh were further established by the first appearance of pollen from non-native taxa Eucalyptus and Plantago lanceolata (narrowleaved plantain). Subsamples for pollen analysis were collected at intervals of 6 –10 cm, and prepared following standard pollen extraction procedures (Faegri and Iversen, 1975). A complete discussion of the pollen dating can be found in Watson (in press).
3. Results and discussion 3.1. Triangle Marsh core general features Mercury concentrations in the Triangle Marsh core vary systemically throughout the length of the core (Fig. 2). Below 150 cm, HgT concentrations are under 1.0 nmol g 1 and show little variation with depth. Between 150 and 70 cm, HgT concentrations increase steadily, from below 1.0 to a peak of over 6.0 nmol g 1. The upper 50 cm are characterized by HgT concentrations between 2.0 and 3.0 nmol g 1. Relating core depth to chronological constraints provided by pollen and radiocarbon dating shows that mercury concentrations began to increase in the later part of the 19th century and reached a maximum in the mid-20th century. The profile of mercury concentrations in this
Fig 2. Results for sediment analyses ( < 63 Am) of Triangle Marsh piston core (location N37 27.323, W121 58.533) total mercury (nmol g 1 F 1 standard deviation) versus depth (cm). Chronology of the Triangle Marsh core is described in detail in Watson (in press).
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core reflects a combination of changes in core lithology, the history of mining in the region and hydrologic factors affecting the transport of contaminant mercury through the upstream watershed.
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suggesting that there has been relatively little diagenetic remobilization. 3.3. Pre-mining mercury concentrations in Triangle Marsh and San Francisco Estuary
3.2. Triangle Marsh core lithology The lowest section of the core, below 250 cm, is characterized by clay-rich sediments that were deposited in an intertidal mudflat environment. The slight increase in mercury concentrations between 200 and 250 cm is coincident with a clayey peat layer that represents a transition from intertidal mudflat to tidal marsh. Sediment between 150 and 200 cm has slightly lower mercury concentrations than the underlying sediments, and represents tidal marsh sediment that was deposited before major anthropological impacts to Triangle Marsh. Sediment between 0 and 150 cm, which includes the large peak in mercury concentration, is characterized by peat and clayey peat, and was deposited during a period of high sedimentation related to local subsidence and increased sediment supply (discussed below). Variations of organic carbon in the sieved sediment reflect the changes in lithology from clay-rich intertidal sediment to peat and have no significant linear correlation to mercury concentration (r2 = 0.01, p = 0.7). The potential for diagenetic remobilization of metals in the Triangle Marsh core was investigated using iron, aluminum, and manganese data from the XRF analysis. Previous studies (Fitzgerald et al., 1998; Varekamp et al., 2003) have used the correlation between these metals as evidence of post-depositional mobility of redox sensitive elements, including mercury. If the redox active metals, iron and manganese, covary independent of aluminum, then diagenetic changes due to redox conditions may have occurred, potentially affecting mercury concentrations. In the Triangle Marsh core, there is a significant linear correlation (r2 = 0.23; p = 0.01) between iron and manganese concentrations, however they are not closely related. Although the range of aluminum concentrations in the core is large (0.1 – 0.6 mol g 1), the linear relationship between aluminum and silica ( p < 0.005, r2 = 0.98) shows that this is due to apparent changes in the abundance of aluminosilicate minerals in the core. Normalizing aluminum to silica results in an aluminum profile similar to that of iron,
The concentrations observed in the pre-mining sediment record throughout the estuary are similar. At Triangle Marsh, the average pre-mining HgT concentration, taken from below a depth of 150 cm, is 0.40 F 0.15 nmol g 1. The average of concentrations observed in depths below 1 m in the San Mateo and Oyster Point cores (Table 1) is 0.35 F 0.05 nmol g 1. These values are similar to concentrations measured by Hornberger et al. (1999) in deep-cores from the northern reach of the estuary (0.30 F 0.05 nmol g 1). In addition, Domagalski (2001) reported concentrations of 0.25 – 0.50 nmol g 1 in stream sediments from the northern Coast Range with no known mercury mineral deposits. This relative homogeneity suggests the contribution from natural weathering of mercury-rich deposits in the Santa Clara Valley to mercury concentrations in San Francisco Estuary is not substantial. Overall, these pre-mining concentrations are within the ranges reported for normal soil and sediment concentrations (Adriano, 2001; Faure, 1998). Consequently, the elevated levels observed in historic and modern sediments in the estuary are predominantly related to anthropogenic activity. The historical background value for San Francisco Estuary shows striking similarity to values obtained Table 1 Core locations, subsample depths, concentration of total mercury (HgT) in nmol g 1 in sieved ( < 63 Am) and freeze-dried sediment, and approximate age Sample ID and location
Depth (cm)
HgT (nmol g 1)
San Mateo SM-4 N37 35.917V W122 15.268V
71 143 242 270 133 280 226 232
0.64 0.37 0.25 0.36 0.30 0.38 0.42 0.43
Oyster Point OP-5 N37 39.014V W122 16.835V Oyster Point OP-6 N37 40.395V W122 21.049V
Approximate age (year) 1500 1600 4000 4300 3600 2200
Approximate ages from AMS 14C dating of shell and organic material by Beta Analytic. Dating of material from these cores is primarily intended to establish that the sediment is from the premining era, i.e., older than about 250 years.
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for background levels in other near shore marine environments. A core collected from the upper estuary of the Tinto River in the Aznalcollar mining area of southwestern Spain revealed background values in bulk sediments between approximately 0.10 and 0.35 nmol g 1 (Leblanc et al., 2000). The background values for bulk sediments in the Gulf of Trieste, which is located downstream of the Idrija mine, have been reported as 0.50 nmol g 1 (Faganeli et al., 1991), although Covelli et al. (2001) estimate a higher background of 0.85 nmol g 1. Background values for bulk sediment from salt marshes in Long Island Sound, and area with no known mercury mining history or mercury mineral deposits range between 0.17 and 0.46 nmol g 1 (Varekamp et al., 2003). Although the considerable differences in core lithologies and background geology prevent any major conclusions from these comparisons, the similarity between the background mercury concentrations in estuaries downstream of heavily mineralized areas and areas with no known mineral deposits is surprising. 3.4. Mining history and downstream concentrations The elevated mercury concentrations in sediments from south San Francisco Bay (Conaway et al., 2003; David et al., 2002) and the gradient from these sediments to the New Almaden mines (Thomas et al., 2002) demonstrate that the mines are the principal source of mercury contamination to the southern reach of the estuary. Large-scale development of the New Almaden district began in about 1845 and continued until 1975, producing over 190 Mmol of mercury (Cargill et al., 1980). The concentrations of HgT in the Triangle Marsh core start to increase shortly before 1870, roughly corresponding to the dramatic increase in processed ore at New Almaden, which reached 500 Mmol moles in 1880 (Cargill et al., 1980). The amount of processed ore decreased substantially between 1910 and 1940, but underwent a brief renaissance in the 1940s, when mercury was primarily recovered from reworking surface ore dumps and open cuts. Previous workers have demonstrated that exposed mercury mine tailings can be the primary source of mercury to downstream watersheds (Ganguli et al., 2000; Rytuba, 2000). The amount of processed ore dumped on the surface and the reworking of surface dumps is therefore expected to have a
large impact on downstream mercury concentrations. The history of mercury deposition in the core appears to follow this prediction, as there appears to be a substantial increase in mercury contamination related to this mid-20th century mining period. Production at New Almaden dropped dramatically in the latter half of the 20th century, and this is reflected in the marsh sediments. During this period, the mercury concentration decreases from its maximum in the core to several nmol g 1, consistent with values for modernday surface sediments in the southern reach, which range from 1.0 to 5.0 nmol g 1 (Conaway et al., 2003; David et al., 2002; Thomas et al., 2002). Because the timing of the increase in mining activity at New Almaden is concurrent with the period of hydraulic mining in the Sierra Nevada, it is necessary to consider the possible contribution of mercury contamination from hydraulic mining sediments discharged into the northern reach being dispersed into the southern reach of the estuary. Some sediment from the Sacramento-San Joaquin river drainage basin becomes deposited in the southern reach (Atwater et al., 1979; Krone, 1979). This proposed transport is supported by analyses of lead isotopes in both surface sediments and waters (Ritson et al., 1999; Steding et al., 2000). Sediment budgets for the estuary, however, suggest that the amount of sediment transported in this manner is relatively small (Krone, 1979). The contribution of contaminated hydraulic mining sediment to the southern reach is consequently considered to be of relatively little consequence to mercury concentrations in that region. 3.5. Hydrologic factors affecting mercury concentrations at Triangle Marsh Although the dominant source of mercury to Triangle Marsh is New Almaden, the transport of this contaminant mercury has been influenced by many factors affecting sediment sources and transport in the watershed. The most obvious evidence of this is the apparent delay between the time of highest mercury production at New Almaden and the peak mercury concentrations in the core (Fig. 3). A similar delay is observed between the 1880s peak of hydraulic mining activity in the Sierra Nevada and the deposition of the mercury-contaminated hydraulic mining sediments in the northern reach of the estuary in the early part of
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Fig 3. Temporal profile of mercury concentrations (nmol g 1) in Triangle Marsh piston core (location N37 27.323, W121 58.533). Treated ore and mercury production figures (Mmol) for New Almaden mines are taken from Cargill et al. (1980). Subsidence data (cm year 1) for the Santa Clara Valley (Benchmark P7, San Jose) are taken from Poland (1984).
the 20th century (Hedgpeth, 1979; Hornberger et al., 1999). Generally, the transport of mercury through the Santa Clara Valley watershed occurs in episodic, peak-flow events (Thomas et al., 2002). The rivers and sloughs in the valley and southern reach have undergone many changes, including both natural and artificial channel realignments—the Guadalupe River in particular has been rerouted several times. In addition, urbanization in the watershed, especially in the city of San Jose, has replaced much of the natural surface with concrete and asphalt, which may favor erosion of contaminated sediment. The building of several major water projects constructed in the upper watershed during the early to mid-20th century, including the Almaden, Andersen, Coyote and Guadalupe Reservoirs, probably also had a major effect on the transport of sediment through the watershed. Another effect of urbanization and water use in the
Santa Clara Valley is local subsidence due to groundwater pumping; between 1934 and 1967 San Jose subsided as much as 3 m (Fig. 3), and Triangle Marsh about 1 m (Poland and Ireland, 1988). All of these changing land use factors likely had a profound impact on the sources and transport of sediment to Triangle Marsh and also on sedimentation at the marsh (Patrick and Delaune, 1990; Watson, in press). 3.6. Mercury accumulation at Triangle Marsh The impact on mercury concentrations by increasing sedimentation rates and supply of contaminated sediment can also be illustrated by estimating the rate of mercury deposition in Triangle Marsh. The mass of the sediment accumulating in g cm 2 year 1 can be calculated as the product of the bulk density of the sediment (g cm 3) and the sediment accumulation
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rate (cm year 1), reported by Watson (in press). Multiplying this result by the molar concentration of HgT (nmol g 1) in the sediment yields the rate of mercury deposition (nmol cm 2 year 1).
mercury to the environment, the subsequent transport of mercury downstream appears to have been predominantly controlled by both the style of mining and changes in the hydrology of the watershed. The relationship between the many hydrological factors mentioned above and the downstream mercury concentrations is complex, and may not be resolvable into one simple relationship. Nevertheless, it is clear that elevated mercury concentrations in south San Francisco Bay are the result of erosion from mercury mine wastes from mercury mines in the watershed, and the contribution of natural weathering in this system is of little importance.
Accumulation Rate ¼ qðsÞHgT where q is equal to sediment bulk density in g cm 3, s is equal to the sediment accumulation rate in cm year 1 and HgT is equal to total mercury concentration in nmol g 1. Although several simplifying strategies were used in this estimation, such as using average values for sedimentation rates and bulk densities, and assuming that the 63 Am fraction makes up the majority of the sediment, the differences shown are on an order of magnitude scale, and should thus be fairly realistic. The results of this calculation (Table 2) show a dramatic increase in mercury accumulation rates during the mining era, particularly the period 1945 – 1975. This temporal increase suggests that the deposition of mercury downstream of the mines is driven principally by mercury mining upstream and changes in hydrology in the watershed, rather than by the amount of mercury produced by the mines.
Acknowledgements The authors are indebted to the pioneering studies of Bill Fitzgerald, who demonstrated the need for rigorous analytical techniques to accurately measure baseline mercury concentrations and to correctly quantify anthropogenic perturbations of mercury’s biogeochemical cycle. We thank Elizabeth Kerin, Kim Bracchi, Miranda Spang, Allison Luengen, Liam Reidy, Brenda Hamilton, Martha Evonuk, Glen Spinelli, Rob Franks and the crew of the R/V David Johnston for their assistance in collecting and analyzing cores. Access to the Don Edwards San Francisco Bay National Wildlife Refuge was approved by Clyde Morris with the U. S. Fish and Wildlife Service. Comments on the research and manuscript were provided by Richard Looker, Tom Grieb, K.E. Abu-Saba, Roberto Anima, John Callaway, P.J. Lechler, Ken Bruland, David Sedlak, R.P. Mason and an anonymous reviewer. This work was funded by a contract with the San Francisco Bay
4. Conclusions This study describes the depositional history of mercury downstream from one of the largest mercury mining districts in the world. Although the watershed is rich in mercury mineral deposits, there appears to be little evidence of contamination downstream of the New Almaden area due to natural weathering. Although the mining era provided an immense source of
Table 2 Mercury accumulation rate calculated for depth intervals in sediment core from Triangle Marsh Depth (cm)
Period
Sedimentation rate (cm year 1)
Bulk density (g cm 3)
HgT (nmol g 1)
Mercury accumulation rate (nmol cm 2 year 1)
0–8 8 – 35 35 – 40 40 – 125 125 – 140 140 – 280
1998 – 2001 1982 – 1998 1975 – 1982 1945 – 1975 1870 – 1945 1570 – 1870
2.7 1.8 0.71 2.8 0.18 0.33
0.8 0.7 0.7 0.6 0.6 0.7
2.3 2.4 2.7 4.2 2.6 0.41
5.0 3.1 1.3 7.0 0.28 0.09
The mercury accumulation rate is the product of sedimentation rate, bulk density of the sediment, and the concentration of mercury (HgT). The sedimentation rate and bulk density values are averages from Watson (in press). Mercury concentrations are averages for each depth interval.
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