Chemical gradients in sediment cores from an EPA reference site off the farallon islands — Assessing chemical indicators of dredged material disposal in the deep sea

Pergamon PII: S0025-326X (98)00003-4

Marine Pollu#on Bulletin, Vol. 36, No. 6, pp. 443-457, 1998 Published by Elsevier Science Ltd. Printed in Great Britain 0025-326X/98 $19.00+0.00

Chemical Gradients in Sediment Cores from an EPA Reference Site off the Farallon Islands Assessing Chemical Indicators of Dredged Material Disposal in the Deep Sea M. H. BOTHNER*t, P. W. GILL?, W. S. BOOTHMAN~, B. B. TAYLOR? and H. A. KARL§ tUS Geological Survey, Woods Hole, MA 02543, USA +,.USEnvironmental Protection Agency, Narragansett, RI 02882, USA §US Geological Survey, Menlo Park, CA 94025, USA Heavy metal and organic contaminants have been determined in undisturbed sediment cores from the US Environmental Protection Agency reference site for dredged material on the continental slope off San Francisco. As expected, the concentrations are significantly lower than toxic effects guidelines, but concentrations of PCBs, PAHs, Hg, Pb, and Clostridium perfringens (a bacterium spore found in sewage) were nearly two or more times greater in the surface sediments than in intervals deeper in the cores. These observations indicate the usefulness of measuring concentration gradients in sediments at the San Francisco deep ocean disposal site (SF-DODS) where a thin (0.5 cm thick) layer of dredged material has been observed beyond the boundary. This thin layer has not been chemically characterized by the common practice of homogenizing over the top 10 cm. An estimated 300 million cubic yards of dredged material from San Francisco Bay are expected to be discharged at the SF-DODS site during the next 50 years. Detailed depth analysis of sediment cores would add significant new information about the fate and effects of dredged material in the deep sea. Published by Elsevier Science Ltd. Keywords: contaminants; sediment cores; Farallon Islands; dredged material; reference site; continental slope.

San Francisco Bay is one of the many industrialized estuaries around the USA that requires periodic maintenance dredging of navigable waterways to sustain recreational, commercial and military shipping. An annual economy of US$5.4 billion, including 35 000 jobs, is tied to the present marine activities (Ogden et *Corresponding author. E-mail: [email protected].

al., 1988; US EPA, 1993). In order to maintain and expand shipping channels in San Francisco Bay over the next 50 years, an estimated 300 million cubic yards of sediment are expected to be dredged from the bay and discharged on the continental slope (US EPA, 1993). Ocean dumping regulations specify that adequate scientific information shall be obtained to assess the environmental impact of this discharge. As part of the processes of obtaining such scientific information, the US Geological Survey (USGS) conducted a geological, geophysical and geotechnical study of a 1000 square nautical mile (3400 km 2) area west of the Farallon Islands for the US Environmental Protection Agency (Karl, 1992). These data were used in combination with other physical, chemical and biological oceanographic data in designating a deep-ocean site for the disposal of dredged material generated from San Francisco Bay. The San Francisco deep ocean disposal site (SF-DODS) was designated by Final Rule (59 Fed. Reg. 41243) on August 11, 1994. The Final Rule stipulated various requirements for site use, bioassays of proposed dredged sediment and a comparison of similar bioassays performed on sediments from the reference site as well as annual site monitoring. Sediments proposed for ocean disposal must be evaluated for suitability for ocean disposal in accordance with EPA Ocean Dumping Regulations (40CFR Part 227). The required bioassays provide a conservative (i.e. environmentally protective) determination of suitability for the marine environment by using appropriately sensitive marine organisms. Only sediments determined to contain levels of chemical constituents that are non-toxic to marine organisms are allowed to be disposed in the ocean. Previous testing of sediment from this reference site revealed potential problems related to elevated chemical contaminants 443

Marine Pollution Bulletin and higher mortality during bioassays than expected. One goal of this study was to provide EPA Region 9 with more detailed geochemical characterization of the sediments at this reference site. This report summarizes the sediment characteristics at the reference site, which is located 35 km south-west of the Farallon Islands off San Francisco in water depths of 800-1500m. A complete compilation of chemical data is presented in Bothner et al. (1997). The study area (Fig. 1) is about 10 km south of both the Gulf of the Farallones National Marine Sanctuary boundary and the boundary for the Farallon Island radioactive waste dump area. The geologic setting and results of acoustic mapping surveys that include the reference site are presented in Karl et al. (1994) and Karl (1992). Additional geochemical data on the slope close to this study area and on the adjacent continental shelf are contained in a report by Dean and Gardner (1995). Long-term rates of sediment accumulation in

the vicinity of the reference site are estimated to average about 11 cm kyr -1 (about 8 g cm -2 kyr - t ) over the last 10000 years (Gardner et al., 1997). The designated SF-DODS is located about 30km north-west of the reference site in water depths ranging from 2700 to 3100 m. The site is in a region historically used for ocean disposal (SAIC, 1996). It is located in the south-west corner of a region used as a chemical munitions dumping area by the US Navy (approximately from 1958 until the late 1960s) and about 15 km WNW of a disposal site used for containers of radioactive wastes (1951-1954). In 1993, approximately 1.2 million cubic yards of dredged material from the Alameda Naval Base were discharged at this site, and in 1995 approximately 232000 additional cubic yards of material from Oakland Harbour were discharged. Postdischarge studies were conducted to map the dredged material footprint and determine the sediment chemistry inside and beyond the disposal boundaries

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Volume 36/Number 6/June 1998 (PRC, 1995; SAIC, 1996). The footprint was found to be displaced to the north-west of the site perimeter, in general agreement with the transport direction expected from regional circulation (Noble et al., 1992) and a particle tracking model developed by Hamilton and Ota (1993). Adverse effects of dredged material disposal on biota on the sea floor may occur from toxicity or from smothering. Given the pre-disposal testing requirements, which disqualify toxic dredged material for ocean disposal, smothering is expected to be the most deleterious. A 5 cm thickness threshold, a trigger set in the Final Rule for future management actions at the SF-DODS site, was not exceeded outside of the boundaries by the past dumping. However, a thin ( < 0.5 cm) layer of dredged material was observed beyond the boundaries of the SF-DODS using sediment profile photography (SAIC, 1996). Changes in benthic ecology were also observed. The pre- and post-dumping assessment of toxic compounds within and outside of the SF-DODS was based on analysis of the top 10 cm of sediment subsampled from a box core and homogenized (SAIC, 1991; Blake et al., 1992; SAIC, 1992a,b, 1996). The concentrations of toxic compounds present in a layer of dredged material thinner than 10 cm would be significantly diluted and underestimated during the homoge-

nization or compositing step in sample collection. For example, contaminants that were selectively transported beyond the dump site boundaries would not be well characterized in a homogenized section. Sampling and analysis of discrete depth intervals from undisturbed sediment cores would provide a more sensitive way to assess the level of toxic compounds in any sediment layer within this geographic area.

Sampling Methods In this study, sediment cores at the EPA reference site were taken with a hydraulically damped gravity corer, an instrument designed to sample the sediment with minimal disturbance of the water-sediment interface (Pamatmat, 1971; Bothner et al., ]997). Collecting undisturbed surface sediment is important because anthropogenic influences on sediment chemistry in the marine environment are most clearly seen by comparing the surficial sediment with sediment deeper in the core. The top few millimetres can be disturbed or lost by other types of samplers. That which is most recently deposited is often most easily remobifized. Seven cores were collected between October 31 and November 1, 1994 from water depths ranging from 950 to 1300m (Fig. 2). Recovered core lengths ranged from 42 to 76 cm. As a precaution against contamina-

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site south of the Farallon Islands (see Fig. 1). Numerical values indicate mud content (%) and pyrene concentration (llg kg i) in the surfaceof sedimentcores. 445

Marine Pollution Bulletin tion, all core barrels, sampling utensils, and sample containers were carefully cleaned prior to use. All core barrels and caps were acid washed with 5% nitric acid and rinsed with methanol and distilled water. Subsamples from the cores were taken with spatulas, knives and spoons that were custom-made from high purity titanium, a material with low contamination potential for both organics and trace metals. Subsamples were placed in non-breakable, non-contaminating 500 ml Teflon screw-cap containers for storage and transport. Containers were prepared for both metal and organic analyses by pre-cleaning with 5% nitric acid and rinsed with methanol, acetone, toluene, methylene chloride and hexane. Cores were sectioned on board ship by inserting a cleaned piston into the bottom of the core barrel and clamping the barrel in a vertical rack. The sediment core was extruded from the top of the barrel by pushing the piston upward using a long-throw hydraulic jack. Sediment in direct contact with the core barrel was trimmed off and discarded. A 0-1 cm sample was collected from each core. Two cores were selected for sampling at closely spaced depth intervals. These cores were sectioned at 1 cm intervals to a depth of 10 cm and at 2 cm intervals for the remaining length of the core. All samples were refrigerated while at sea. At the conclusion of the cruise, core sections were transported to USGS laboratories in Palo Alto, CA, where they were subsampled for trace metals, Clostridium perfringens and grain size analysis. The remainder was frozen until analysed for organic compounds.

Analysis Methods Details of the analytical methods are described in Bothner et al., 1997. Below is a brief summary of each technique.

Sediment texture Sediment texture was determined using standard sieving techniques for the sand fraction and Coulter counter methods for the fine fraction. Samples were analysed wet to avoid the formation of 'bricks' which typically result from drying bulk sediment by any method (Barbanti and Bothner, 1993). Results are reported as weight percent sand, silt and clay. Corrections for foraminifera or other biogenic carbonate particles were not made because these particles typically make up less than 0.2 wt% of the sample and do not significantly influence the sand size fraction. Clostridium perfn'ngens Analysis of Clostridium perfringens was performed at Biological Analytical Laboratories, North Kingstown, RI. Spore concentrations were measured using the direct enumeration method, which involves extraction of spores by sonification followed by settling and 446

membrane filtration (Bisson and Cabelli, 1979; Emerson and Cabelli, 1982). The standard deviation among four replicates was 24%.

Trace metals Analyses were conducted by Quanterra Environmental Services, Arvada, CO using EPA methods. The sediments were leached with hot concentrated nitric and hydrochloric acid to solubilize contaminant metals. Determinations (Ag, As, Cd, Cr, Cu, Hg, Ni, Pb, Se and Zn) were made using inductively coupled plasma atomic emission spectroscopy, graphite furnace atomic absorption, or cold vapour atomic absorption (Hg) as appropriate. Analytical error was estimated by analysis of spiked duplicates and averaged 2.6 + 3.7%. Canadian sediment standards MESS-l, PACS-1 and BCSS-1 were included with the analyses. Chemical yields using the hot concentrated acid leach method were typically > 70% of the published total metal concentration for these standards. Exceptions include yields of 34-53% for Cr, suggesting that some metals can be bound in a mineral phase resistant to the leaching procedure. Organic compounds Samples for analysis of organic contaminants were stored frozen in tightly sealed thick-walled Teflon containers to minimize volatile loss, decomposition and contamination during the 20 month storage period. Analyses were conducted by Quanterra Environmental Services using both approved and provisional EPA methods referenced in Bothner et aI., 1997. Briefly, polychlorinated biphenyls (PCBs) were extracted with toluene or methylene chloride and analysed using a high resolution mass spectrometer technique. Isotope dilution, utilizing 13C-labelled analogues, was used to determine individual PCB congeners. The estimated precision of the PCB analysis is _+20-50% for concentrations near the reporting limit and 10-20% at concentrations greater than three times the reporting limit. Polycyclic aromatic hydrocarbons (PAHs) were extracted with methylene chloride and measured on a gas chromatograph/mass spectrometer in the selectedion monitoring mode. The concentrations of PAHs were typically below the normal limit specified for the method, and the estimated error is about a factor of 2. These values are flagged in the data tables (see the respective table footnotes). The good agreement in trends between individual PAH compounds and other sediment parameters such as Clostridium perfringens suggests that the PAH values are, at minimum, a good relative measure. Organic carbon was determined by difference after coulometric measurement of total carbon and carbonate carbon. Acid volatile sulphides (A VSs) and simultaneously extracted metals (SEMs) AVSs and SEMs were determined on selected samples using the EPA methods (Allen et al., 1991;

Volume 36/Number6/June 1998 Boothman and Helmstetter, 1992). The samples were stored frozen until analysis in April, 1996 at the EPA laboratory in Narragansett, RI. The long storage (18 months, frozen in a thick-walled Teflon jar) of these samples introduces some uncertainty to the results. In storage tests on other samples, no significant changes were observed in AVS and SEM measurements on marine sediments after 6 weeks of frozen storage (Boothman, 1992); however, the effect of longer frozen storage is, to our knowledge, untested.

concentrations at or near the analytical detection limit were observed. The concentration gradients are not significantly altered when the data is normalized to account for sediment parameters which can influence contaminant concentrations, such as organic carbon content or sediment grain size. The average concentration of total PCBs (sum of mono- through deca-CB compounds) among the four surface sediments analysed is 2.9_+ 1.0 ppb which is in the range observed at a reference site at 2500 m water depth off New Jersey: ~ 3 p p b (Takada et al., 1997) and 7.5ppb (Lamoreaux et al., 1996). The average value of PCBs at the reference site is a factor of 8 lower than the ERL (effects range - - low) toxicity guideline and a factor of 62 lower than the ERM (effects range - - medium; Table 1; Long et al., 1995). The ERL and ERM are guideline concentrations above which toxic effects in marine organisms were observed in approximately 10% and 50% respectively of reviewed studies. These guideline concentrations are used by EPA Region 9 for initial screening and to evaluate the results of toxicity tests for sediments proposed for ocean disposal. The ERL and ERM values are determined from a collection of studies involving test organisms collected from shallower marine environments. These values may not

Results and Discussion Organic compounds The depth profiles of PCBs provide evidence that the surface sediments at this reference site have already received measurable quantities of these anthropogenic compounds (Fig. 3 and Table 1). PCBs were commercially manufactured and marketed in the USA from 1929 to 1977. They were used widely in industry because of their chemical stability, low boiling point, low solubility and high dielectric constant (Weaver, 1984)• In both cores analysed, the concentrations of mono- through deca-chlorobiphenyls in surface sediments (0-3 cm and 0-2 cm) are elevated compared with sediments from 20-30cm in the core, where 0

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be directly applicable to the deep sea environment where, to date, little information on toxicity to these benthic organisms is available. However, these values are relevant for the pre-disposal testing because, for any proposed dredging project, the relative response of test organisms is measured for exposure to both the sediments proposed for ocean disposal and the sediments fromthe reference site. The depth profiles of PAHs also indicate an anthropogenic input to this site. The concentrations of pyrene (Fig. 4), fluoranthene, phenanthrene, benzo(b)fluoranthene, and benzo(g, h, i)perylene (Table 2) decrease by about a factor of 2 between the surface and the 9-11 cm depth interval in cores from stations 3 and 8. The consistency of the concentration decrease with depth displayed by each of the compounds helps to verify the trends in spite of the relatively high analytical uncertainty at these low concentrations. These compounds are introduced to this environment primarily by combustion of wood (forest or prairie fires), coal, or petroleum products, with increased amounts added to the environment during industrial development. A possible additional source of these compounds in trace quantities could result from deposition of unburned petroleum products such as gasoline and kerosene. The depth profiles are similar to those measured on the continental slope off the state of Washington, where the enrichments in PAHs are attributed to growth in anthropogenic activity within the Columbia River drainage area rather than to post-depositional degradation of non-anthropogenic compounds introduced at a constant rate (Prahl and Carpenter, 1984). The concentrations of pyrene and fluoranthene measured at the reference site are about a factor of 3 higher than those measured in surface sediments from the Hatteras Abyssal Plain in the North Atlantic Ocean (4 ppb dry weight) and about a factor of 10 lower than found in sediments from Buzzards Bay, Massachusetts and the Gulf of Maine (100-130 ppb (LaFlamme and Hites, 1978)). They are comparable with the values measured ( ~ 10 ppb) at a reference site near the 106-mile dump site off New Jersey at 2500 m water depth (Takada, 1997). The high correlation (R2=0.83) between pyrene and mud content in surface sediments (Fig. 2) suggests that the PAH compounds are transported and deposited with the fine fraction of marine sediment. The depth gradient for PAils is less steep than observed for PCBs, probably because there are natural pre-industrial sources of PAHs, such as forest fires, and, therefore, a higher background concentration.

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Spores of the bacterium Clostridium perfringens have been used in marine sediments as a tracer of human sewage (Hill et al., 1993). In coastal Massachusetts, for example, sewage discharge accounts for > 100000sporesg-1 in harbour sediments (Durell,

Volume 36/Number 6/June 1998

1995). It is somewhat surprising that their concentrations in the surface sediments of this rather remote offshore site average 780_+ 380 spores g-1, significantly above detection limits ( < 4 0 spores g-l, Table 3) and significantly higher that concentrations measured at the reference site for the 106-mile dump site off New Jersey at 2500 m water depth, 43-89 spores g-1 (Hill et al., 1993). The concentrations decrease with sediment depth in the two cores analysed (Fig. 4). Elevated concentrations persist to about 7 cm at the deep station 8 (1166 m) and to at least 21 cm at station 3 (980 m). Although material dredged from any urban harbour and barged offshore would be a mechanism for introducing Clostridium perfringens to the deep sea, we do not know of any dumping in the immediate vicinity of the reference site. Natural transport processes carrying continental shelf sediment with an urban component offshore provide a potential source. In the summer of 1988, an anticyclonic eddy remained over the continental shelf north of the Farallon Islands and transported an estimated 105 t of suspended sediment from the shelf to the deep ocean (Washburn et al., 1993).

Marine mammals, known to migrate through the reference area, may also provide a source of Clostridium perfringens to the environment, but the magnitude of their faecal contribution has not been estimated.

Trace metals Trace metal profiles in sediment cores from offshore marine areas can also provide evidence for contaminant inputs from coastal ouffalls or the atmosphere, but this signal is weak at the EPA reference site. Within the suite of heavy metals that were analysed in cores, only Hg and Pb depth profiles have a small (about 2 x ) enrichment in surface sediments relative to sediments at 10 cm depth and below (Fig. 4, Table 3). An enrichment of Hg and Pb within the top 10 cm of somewhat lower magnitude is observed w h e n e a c h metal is normalized by computing the ratio of metal to organic carbon or mud content. The concentrations of Pb in surface sediments of the reference site (4.2+0.5ppm) are lower than observed on the East Coast continental shelf or slope, 15-30 ppm (Bothner et al., 1981, 1987). Prevailing winds at this reference

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0.26 0.31

Ag (gtgg-1)

Composite sample from two depth intervals. b N R C Canada, 1987. c C. perfringens analysis of this sample was done with four replicates; all others used two replicates. .~ a Reporting limit raised due to the matrix of the sample. t.~ e H g values have been converted from gg g - 1 wet. e Result is detected below reporting limit or is an estimated concentration.

Spiked duplicates 3-1 47-49 3-1 47-49 8-1 41-43 8-1 41-43

PACS-1

B CSS - 1

Standards MESS-1

this work published b std dev. this work published b std dev. this work published b std dev.

53.47 48.89 52.95 54.28 67.43

34.91 37.73 31.06 7.14 11.57

47.67 44.80 54.66

46.81 48.07

36.51 35.83

3-1 0 - 1 3-1 2 - 3 3-1 0-1, 3-1 2 - 3 a 3-1 5 - 7 3-1 9-11 3-1 19-21 3-1 29-31 3-1 47-49

(cm)

Silt (%)

Sand (%)

Core ID and depth interval

0.68 0.67 1.2 1.2

0.71 0.59 0.1 0.42 0.25 0.4 2.6 2.38 0.2

0.56 d 0.7 d

0.680 0.69 d 0.68 d 0.99 d 0.84 d

0.48 d 0.56 d

0,65 0.5 a

0.31 0.28 0.35

0.43 0.42 0.53 0.64 0.31

0.38 0.53

Cd (ggg-~)

53.6 53.2 61.4 61.4

24.3 71 11 46.5 123 14 52.8 113 8

59.6 64

71.6 63 64.8 57.5 52.8

61.8 65.6

63 63.2

46.6 54.6 50

47.4 50.2 74 58.6 50.2

43.4 57.2

Cr (ggg-~)

4.57 0.16

0.129 0.012

0.171 0.014

0.116 0.134

0.134 0.24 0.07 0.07 0.046 t" 0.065

0.109

0.105 0.092 0.115 f

0.125 0.074 0.06 0.065 0.078 f 0.069

Hg (gtgg-~)~

Concentration of metals, C. perfringens and sediment texture (dry weight basis).

TABLE 3

50.7 50.7 42.7 41.5

20.3 25.1 3.8 13.7 18.5 2.7 450 452 16

15.2 16.5

20.2 18.9 17.4 18.1 16.3

19.1 20.3

17.6 22.4

16.9 16.8 19.8

15.3 14.3 33 20.8 25.9

14.5 15.4

Cu ( g g g 1)

112 113 104 98.7

19.8 29.5 2.7 43.1 55.3 3.6 31 44.1 2

42.8 53.9

58.3 56.9 51.6 55.7 51.8

55.5 58

51.4 60.1

53.1 51.1 54.9

49.4 47.8 80.9 60.9 66.3

47 50.2

Ni ( g g g ,)

108 109 109 103

147 191 17 86.3 119 12 750 824 22

44.3 55.5

64 61.9 56.3 60.2 55.8

59.9 61.7

55 65.7

57.1 58.1 61.6

52 50 74.4 59.3 63.1

49.4 52.3

Zn ( g g g 1)

9.4 8.8 5.4 5.4

32 34.0 6.1 24.5 22.7 3.4 388 404 20

4.1 3.3 d

4.5 3.1 d 3.4 d 2.8 d 2.6

4.5 5

4.6 4.4

4.4 4.6 4.2

3.1 3 9.7 4.9 6.3

5.3 4

Pb ( g g g 1)

2.7 2.6 3.4 3.9

< 1.0 0.34 0.06 < 1.0 0.43 0.06 0.58 f 1.09 0.11

0.72 a'f 0.76 d'f

<2.5 d 0.88 a'f 0.82 a'f Id 0.92 d'f

0.6 d'f 0.98 ax

0.98 ax < 1.0d

0.98 0.98 1.3

0.68 0.78 2 1.2 0.47

0.64 0.88

Se (ggg-~)

3.5

2.13

2.88 2.9

1.62 1.42

1.78 1.6 1.48 1.25 1.05

1.85 1.93

1.91 1.83

1.55 1.46 1.94

1.18 1.05 1.49 0.99 0.81

1.22 1.22

(%)

Total organic carbon

~D

cr

e-,

<

Marine Pollution Bulletin

site off the west coast are from the ocean toward the continent (Dorman and Winant, 1995). Much greater distances from upwind continental sources in the Pacific than the Atlantic sampling sites may explain the relatively lower signal of lead and other atmospheric pollutants in the sediments of this reference site. In Table 4, the concentrations of trace metals in surface sediments of the reference site are compared with values in uncontaminated sediments from other areas as well as to toxicity guidelines. The average concentration of Cd, Se and Ag are between 1.6 and 3.8 times higher than reported for world average shales. These three elements, plus Ni, have higher concentrations at the reference site than at offshore locations off Hawaii and New Jersey or at nearshore background stations off the US East Coast (Table 4). There is no evidence from analysis of cores at stations 3 and 8 that the surface concentrations of these four metals are higher than concentrations at the bottom of the cores in sediments deposited prior to industrial influence, so the elevated concentrations relative to crustal material and shales are probably due to the natural mineral assemblage rather than to anthropogenic additions. The average concentration of each metal in surface sediments from this reference site is lower than the ERM toxic effects guideline described by Long et al., 1995 except for nickel. The average concentration of nickel is 50.8ppm, just below the ERM value (51.8 ppm). However, the ERM value is lower than the

concentration in average crustal material (75 ppm) and average shale (95 ppm, Table 4) and is established with less confidence than for other metals (Long et al., 1995). Average values of Cr and Pb are a factor of 3 lower than those measured in the same general area by previous investigators. The low Cr results may be due to the low chemical yields for this element by the EPA leaching method that were quantified by analysis of sediment standards. A similar explanation for the low Pb values is not as likely because our results on Pb in standard sediments were within one standard deviation of accepted values.

A VSs and SEMs

In recent years, laboratory and field experiments have shown that the presence of AVS has a significant impact in controlling the bioavailability of heavy metals in marine sediments (e.g. see the journal issue containing Berry et al. (1996) for a compilation of 15 papers delineating the relationships between AVS, metals and benthic biological response, and their application to derivation of sediment quality criteria). The bioavailability is reduced when precipitation of metals as sulphides occurs in pore water and effectively removes the metals from biological utilization. In sediments where the molar concentration of AVS exceeds that of SEM, dissolved metals in pore waters are low and toxic effects attributable to metals are not observed.

TABLE 4 Average concentrations of metals in surface sediment (0-1 cm, n = 7) from the Farallon Island. Reference site compared with ERM ~ values and values in uncontaminated sediments from other areas.

Se

Ag

Total organic carbon (%)

4.2 0.5

0.95 0.23

0.38 0.07

1.67 0.24

-71

13.8 --

---

0.52 --

-1.01

410 70 80

218 12.5 20

-0.05 0.6

3.7 0.07 0.1

----

Average concentration (ppm dry weight) As

Cd

Cr

Cu

H~

Ni

Zn

This study av. conc. std dev.

2.2 0.3

0.51 0.18

58.2 5.9

17.2 1.7

0.11 0.02

50.8 4.8

54.9 6.5

Previous studyc Previous studyd

---

0.37 --

168 140

24.3 20

0.08 --

66.3 70

ERM values ~ Av. crust e Av. shale e

70 1.8 6

9.6 0.2 0.3

370 100 100

0.71 0.08 0.4

51.8 75 95

Hawaii offshoref

1.8

--

23.7

8.7

0.03

13.4

25.3

2.3

--

New Jersey continental riseg

--

0.2

69.9

22.8

--

37.1

86.4

19.1

--

0.036

--

US East Coast coastal sediment" bkg. mean std. dev.

5.7 3.3

0.13 0.09

38 24

12 l0

0.04 0.03

15 11

58 38

22 14

0.3 0.2

0.05 0.04

---

270 55 57

Pb

Long et al., 1995. b Hg values for stations 3 and 8 were determined on sample blends from depths of 0-3 cm and 0-2 cm respectively. SAIC (1992b); area near the reference site, n --- 5. Dean and Gardner (1995), average of surface samples from cores G27 and G28 from in or near the reference site. Krauskopf, 1967. Torresan (1995), US Geological Survey, reference station 102. g Bothner (1996), US Geological Survey, 106-mile dumpsite reference station. Alvin Dive 3075. h Strobel et aL (1995), average among background samples from EMAP, Virginia province; N = 181 to 476 samples.

452

<0.05

0.31

Volume 36/Number 6/June 1998

No significant AVS was found in the reference site. AVS concentrations in the reference site sediment core samples ranged from <0.005 to 0.11amolg - j dry sediment (Table 5), which are quite low compared with AVS concentrations found in most coastal marine sediments. Uncontaminated muddy coastal marine sediments generally have AVS concentrations ranging from 0.5 to 31amolg J in oxidized sediments to 15-500 lamolg -~ in anoxic sediment. Assuming no losses during the period of frozen storage, the absence of AVS in these sediments may be attributed to unfavourable conditions for its formation in this deepwater environment. AVS production is driven by microbial sulphate reduction and requires: reactive iron and sulphate ion in pore water; organic carbon to metabolize; sulphate-reducing bacterial activity. The sediments were 62-84% unconsolidated aluminosilicate mud and should not be limited by iron or sulphate availability. Organic carbon concentrations were moderate in the core sediments (0.8-1.9%). However, in sediment from 1000 m water depth, and deeper, the organic carbon may have already been degraded relative to organic carbon in sediments from shallower water. Finally, references have shown microbial sulphate-reducing activity to be temperature dependent (Jorgensen and Sorensen, 1985). The annual mean temperature of bottom water in this area is between 3 and 4°C (Levitus et al., 1994) and so may not have sufficient sulphate-reducing activity to generate measurable AVS. A possible consequence of the low AVS is that these sediments would have little capacity to sequester, adsorb, or otherwise retain additional metal loading as solid sulphides if the metals were added in labile chemical forms. The levels of SEMs show a slight

excess of metals above AVS (0.2-0.6jamolg -1, Table 5), primarily due to the extremely low AVS rather than elevated metal (SEM) concentrations. These bioavailable 'free' metals are at very low concentration. In EPA laboratory experiments it has been shown that at least 2 - 3 btmol g J of 'excess' metals (i.e. concentration [SEM]-concentration [AVS]) are required for any toxic effects to be observed (Berry et al., 1996; Hansen et al., 1996) because other sediment components (e.g. organic matter and iron and manganese oxides in oxic sediments) also contribute to solid-phase binding of the metals.

Characterizing Dredged Material Deposited Outside the S F - D O D S The post-dumping survey of the SF-DODS in 1995 (SAIC, 1996), which used samples composited over the top 10 cm for chemical analyses, concluded that the sediments at and beyond the perimeter of this dump site appear to be uncontaminated by the disposal of dredged material from Oakland Harbor. However, a change in the infaunal successional stage, from larger head-down deposit feeders (Stage III) to small surface feeding or filtering organisms (Stage I), was noted outside the dumpsite at the time of the post-dumping survey of 1996 relative to an earlier survey (SAIC, 1996). The processes offered as possible explanations for this biological change included: accumulation of dredged material (burial or smothering); enhanced inputs of labile organic detritus (a potential food source for deposit feeders); erosion; other physical disturbances (SAIC, 1996). Another possibility, not easily identified with a 10 cm composited sample, is that a chemical anomaly (i.e.

TABLE 5

AVS and SEMs. Chem ID

Core

Depth (cm)

Rep

AVS (gmol g i)

SEM (gmolg- l)

SEM (~tgg - ~) Zn

Cu

Ni

Pb

Cd

Cr

Mn

Fe

<0.27 <0.22 0.74 <0.25 <0.24 0.30 0.29 0.34 0.67

< 1.16 <0.94 37.02 4.19 6.43 10.79 10.96 12.2 23.39

<7.74 <6.32 <2.92 <3.65 < 2.38 1.47 <0.89 1.41 4.91

<5.03 <4.11 6.43 6.02 5.23 6.37 6.04 7.27 13

<8.13 <6.64 <3.07 <3.83 < 2.5 1.24 <0.93 1.25 t.82

<1.16 <0.94 <0.43 <0.54 < 0.35 0.16 <0.13 0.12 0.36

<3.87 <3.16 <1.46 < 1.82 2.26 4.7 4.16 6.38 8.45

6.19 8.22 13.61 13.5 15.6 13.06 13.2 16.2 26.23

2881 2838 2931 3145 3049 2433 2712 2980 4090

0.37 0.42 0.43 0.48 0.44 0.41

14.04 14.78 15.12 16.66 15.76 15.17

1.24 2.23 1.66 2.68 2.9 2.59

7.72 8.68 9.52 10.3 8.32 7.81

< 1.08 2.55 2.11 2.53 1.95 0.9

<0.15 <0.10 <0.14 <0.10 0.13 0.12

1.81 3.47 4.86 5.57 4.6 3.94

16.22 16.77 18.27 19.13 18.24 17.27

2798 2949 3145 3222 2825 2680

16 91

10 84

7 84

17 92

15 87

6 85

8 71

Sediment samples

25342 25343

2-3

25344

5-7

25345

9-11

25346

19-21

1 2 l 2 1 2 1 2 1

0-1 1-2 2-3 3-4 5-7 9-11

1 1 1 1 1 1

25347 25348 25349 25350 25351 25352

3-1

8-1

0-1

<0.04 <0.03 <0.01 <0.02 0.10 0.08 0.04 0.06 0.03 <0.0052 0.0054 <0.0049 0.02 0.01 0.06

QA summary

Calibration verification (%) Spike recovery (%)

± 15 72-77

9 96

453

Marine Pollution Bulletin

5 cm

B

Fig. 5 (A) SVPS photograph showing the two depositional events recorded at the eastern edge of the SF-DODS (Fig. 1) in September 1993. An arrow shows the boundary between the first and second events, which is marked by a thin (2-3 ram) oxidized layer above dark sediment. This oxidation took place when the dark mud clasts were exposed to the overlying water between events. The second layer was deposited following the oxidation of the first layer and, in turn, is already becoming oxidized. The rough boundary is typical of recently deposited dredge material. No infauna are observed (station 14 C-2 from PRC (1995)). (B) SVPS photograph showing a continuous layer (0.5 cm thick) of dredge material mud clasts draped over the water-sediment interface. The location is 2.9 km east of the designated dredge site perimeter. Note that the mud chips are both light and dark in colour. The dark chips could be high in organic carbon and contaminants such as typically found in recently deposited harbour sediments. The light chips may represent over-consolidated clay from deeper in the dredging channel which may have been deposited well before anthropogenic influence and contain low organic carbon and contaminant concentrations. The arrows point to partially filled voids formed by benthic organisms. (Station 16 A-2, September 1993, PRC (1995).) (C) Plan view of the sea floor 2 km west of the dump site perimeter in September 1993. Scale of long dimension is approximately 1 m. The brown muddy surface is littered with high (white) and low (blue) reflectance dredge material mud chips. The surface is heavily tracked by epibenthic organisms. Also evident are spiral faecal castings (F), tips of ophiuroid arms (O), anemones (A), sea pens (P) and burrows. (Station 11 Rep D, September 1993, PRC (1995).)

454

Volume 36/Number6/June 1998 enriched organic carbon and/or contaminants) in surface sediments has influenced the infaunal successional stages. The uppermost sediments were photographed with a sediment vertical profile system (SVPS) (Rhoads and Germano, 1982) during the post-dumping survey. The photographs (Fig. 5(B)) indicate that a surface layer, _<0.5 cm thick, of thin sand-sized mud chips was deposited to the NW of the dumpsite boundary. A similar distribution was observed following the dumping in 1993. Rhoads (1997) has suggested that these fine particles have spalled off larger clay clumps as they descend through the water column and are sorted and differentially transported according to grain size, shape, and hydraulic settling velocity. Photographs obtained using the SVPS in 1993 illustrate the blocky dredged material deposited within the SF-DODS (Fig. 5(A)) and sand-sized mud chips outside the dump site border (Fig. 5(B)). Note that the mud chips are bimodal with respect to colour and reflectivity (Fig. 5(B)). The light grey chips may be over-consolidated clay ( > 10000years old) that was dredged from the more deeply buried harbour sediment where contamination levels are low. In contrast, the black mud chips could be sulphide-rich material from recent deposits in the harbour that would potentially have higher contaminant and/or organic carbon concentrations (Rhoads, 1997). Collection and chemical characterization of this layer is needed. A photograph of the sea floor (Fig. 5(C)), taken 2 km west of the dump site perimeter reveals the presence of mud chips on the sediment surface and various benthic organisms. These three images were collected after the 1993 dumping of dredged material, but similar features and distributions were observed following the 1995 discharge (SAIC, 1996). The mud chip layer beyond the perimeter of the dump site might have higher contaminant and/or organic carbon concentrations than the bulk dredged material deposited inside the dump site because contaminants and organic carbon are typically enriched in the finest sediment fraction relative to the coarse fraction or bulk sample. The process of spalling and selective transport of the fine fraction by currents may transport particles with relatively high contaminant and organic carbon concentrations. Coarser material, including disarticulated shells of shallow water bivalves having lower contaminant concentrations, settle quickly in the central area of the dump site. The presence or absence of a chemical anomaly of this mud chip layer could be readily determined using coring and subsampling techniques which allow characterization of thin sediment layers, such as those employed at the reference site.

Conclusions and Recommendations We have found that PCBs, PAHs, and Clostridium are slightly enriched in the surface

perfringens

sediments of the reference site compared with deeper horizons in sediment cores. The magnitude of enrichment is not considered to be of toxicological concern since contaminant concentrations are well below the ERM and ERL guidelines. Since the reference site is outside of areas exposed to direct dumping, the observed enrichments probably indicate deposition of contaminants from the atmosphere and/or coastal waters to this general area of the continental slope. The concentrations of AVSs and SEMs are quite low compared with concentrations found in coastal marine sediments and indicate both that there is little bioavailable 'free' metals and that these slope sediments have little capacity to bind additional metals should any be added in labile chemical form. The recommendation following the results of this study is that the analysis strategy for evaluating the potential toxicity of dump site sediments include detailed chemical profiles in undisturbed sediment cores. This technique reveals chemical gradients and surface enrichments that are not revealed by analyses of sediments homogenized over the top 10cm. One immediate application of the more detailed sampling and analysis plan would be to determine the potential chemical and ecological impact of a sand-sized mud chip layer that dispersed beyond the border of the SF-DODS following dumping in 1993 and 1996 and which may have induced the changes in infaunal species succession that were observed. The additional data would also contribute to our understanding of fundamental bio-geochemical processes in the deep sea. For example, analysis of sediment cores will help identify whether any chemical constituents added with dredged material are mixed, buried, or transported out of the SF-DODS by natural processes. Cores collected at different times from the same locations would reveal the influence of these processes on the long-term changes in chemical inventories. This information is necessary for evaluating the long-term fate and effect of dredged material in the deep sea. The authors wish to thank Marilyn Buchholtz ten Brink, Allan Y. Ota, Donald C. Rhoads, Hideshige Takada, Bruce W. Tripp and Page C. Valentine for helpful discussions and constructivereviews. This research is a componentof an InteragencyAgreementbetween the US EnvironmentalProtection Agency and the US Geological Survey. Allen, H. E. et al. (1991) Determinationof acid volatile sulfideand selected simultaneously extracted metals in sediment. US Environmental Protection AgencyReport 821/12-91/100, US EPA Office of Water, Office of Science and Technology,Health and EcologicalCriteriaDiv.,Washington,DC. Barbanti, A. and Bothner,M. H. (1993) A procedure for partitioning bulk sediments into distinct grain-size fractions for geochemical analysis.Environmental Geology and Water Science 21, 3-13. Berry, W. J., Hansen,D. J., Mahoney,J. D., Robson, D. L., Di Toro, D. M., Shipley,B. P., Rogers, B., Corbin,J. M. and Boothman,W. S. (1996) Predicting the toxicity of metal-spiked laboratory sediments using acid volatile sulfide and interstitial water normalizations. Environmental Toxicology and Chemistry 15(12), 2067-2079. Bisson, J. W. and Cabelli, J. J. (1979) Membrane filter enumeration method for Clostridium perfringens. Applied and Environmental Microbiology 37(1), 55-66. 455

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