Natural rates of sediment containment of PAH, PCB and metal inventories in Sydney Harbour, Nova Scotia

Science of the Total Environment 407 (2009) 4858–4869

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Science of the Total Environment j o u r n a l h o m e p a g e : w w w. e l s ev i e r. c o m / l o c a t e / s c i t o t e n v

Natural rates of sediment containment of PAH, PCB and metal inventories in Sydney Harbour, Nova Scotia J.N. Smith a,⁎, K. Lee a, C. Gobeil b,c, R.W. Macdonald d a

Bedford Institute of Oceanography, Fisheries and Oceans Canada, Dartmouth, NS, Canada B2Y 4A2 Institut Maurice Lamontagne, Fisheries and Oceans Canada, Mont Joli, Québec, Canada G5H 3Z4 Université du Québec, INRS-ETE, Québec, QC, Canada G1K 9A9 d Institute of Ocean Sciences, Fisheries and Oceans Canada, Sidney, BC, Canada V8L 4B2 b c

a r t i c l e

i n f o

Article history: Received 26 November 2008 Received in revised form 8 May 2009 Accepted 14 May 2009 Available online 11 June 2009 Keywords: PAH PCB Pb Contaminant Sediment Geochronology

a b s t r a c t Analyses of metal and organic contaminants were carried out on 41 sediment cores, dated using 210Pb and 137 Cs, from the heavily industrialized region of Sydney Harbour, N.S. to evaluate the history of contamination and to predict the rates of natural containment of the harbour by sediment burial. Geochronologies for metals (eg. Pb, As) and polycyclic aromatic hydrocarbons (PAHs) are correlated with the development of the steel and coke industries in the Sydney region while polychlorinated biphenyl (PCB) geochronologies reflect the disposal of electrical equipment used in the steel mill/coking operations. Pb was derived mainly from atmospheric emissions and its concentration has declined exponentially with time in harbour sediments since the closure of the steel mill/coke ovens in the 1980s with a time constant of about 15 years. This represents the time scale for the circulation of this particle-associated contaminant in transient catchment basins prior to permanent deposition in the sediments. PAH and PCB sediment concentrations have also declined exponentially with time since the 1980s, but with a smaller time constant of 10 years owing to the fact that they enter the harbour directly with steel mill and coke oven effluent rather than through atmospheric pathways. Since the time dependence for the burial of metal and organic inventories can be modeled by first order processes, future contaminant levels can be predicted for surface sediments in Sydney Harbour. Mean sediment concentrations of metal and organic contaminants in the upper 5 cm throughout most of the harbour are predicted to decline to levels below the effects range-medium (above which organisms are very likely to be negatively affected by the presence of a contaminant) by 2030. Crown Copyright © 2009 Published by Elsevier B.V. All rights reserved.

1. Introduction Sydney Harbour, Nova Scotia like many other urban marine inlets, has long been used as a convenient and inexpensive waste disposal area for many types of materials. In particular, it has been subject to atmospheric and effluent inputs of contaminants, including metals and polycyclic aromatic hydrocarbons (PAHs), from a large coking and steel manufacturing facility (Furinsky, 2002; Lambert et al., 2006) since the 1890s. Much of the effluent was in the form of coal tar residues that were discharged from the coking ovens into Coke Ovens Brook and subsequently transported into the “tar pond” (Fig. 1). The tar pond is an enlarged portion of a small tidal tributary, Muggah Creek located in the South Arm of Sydney Harbour. PAH and metal releases from Muggah Creek are transported directly into Sydney estuary (Fig. 1). Since many PAHs and metals are extremely particlereactive, a major component of these inputs accumulates in the sediments of the estuary.

⁎ Corresponding author. E-mail address: [email protected] (J.N. Smith).

The tar pond that drains into Muggah Creek has been estimated to hold 700,000 tonnes of coal tar contaminated sediment, representing 3500 tonnes of PAHs as coal tar, plus heavy metals (Lambert et al., 2006). It also contains 3.6 tonnes of polychlorinated biphenyls (PCBs) derived from the usage of electrical equipment in the steel plant. Within the estuary, the highest PAH concentrations have been reported for the South Arm with concentrations decreasing towards the outer harbour (Furinsky, 2002). In 1986 elevated PAH concentrations were reported (Uthe and Musial, 1986) in digestive glands of American lobster (Homarus americanus) taken from the estuary while more recent analyses indicate continued contamination of lobster digestive glands with PAHs (King et al., 1993). In 1980, the lobster fishery in the South Arm was closed and the steel plant and coke ovens were finally shut down in the late 1980s. Remediation of the tar pond using in situ stabilization techniques is presently underway, but environmental issues regarding the large contaminant inventories in the sediments remain unresolved. One concern is the extent to which surface sediments are undergoing natural containment by the deposition of sediments less contaminated than those deposited during peak industrial activities in the past. If natural containment is sufficiently rapid and effective, then it

0048-9697/$ – see front matter. Crown Copyright © 2009 Published by Elsevier B.V. All rights reserved. doi:10.1016/j.scitotenv.2009.05.029

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Fig. 1. Sediment cores were collected at stations in Sydney Harbour, N.S. with four cores being collected in Muggah Creek (inset). Steel mill and coking plants discharged wastes into Sydney Harbour via Coke Oven Brook, the tar ponds and Muggah Creek.

may not be necessary to undertake large investments in capital and resources to clean harbour sediments, but simply encourage the natural recovery of the sediment regime. The purpose of this paper is to use radionuclide tracers (210Pb and 137Cs) to determine the time scales for the historical accumulation of contaminants in the sediments of Sydney Harbour and to predict future harbour remediation rates. 2. Methods and environmental setting Sydney Harbour, located in eastern Cape Breton, Nova Scotia opens onto Sydney Bight at the eastern end of the Scotian Shelf. The harbour is “Y” shaped with the seaward arm dividing into a northwest and south arm (Fig. 1) with the city of Sydney and the Sydney steel plant and coke ovens sites located on the eastern bank of the south arm. Fine-grained clays and silts cover the bottom of the south arm and these grade seaward into coarser-grained sands that cover the bottom of the harbour entrance. Sydney River empties into the harbour at its head while Muggah Creek empties into the harbour on its eastern flank, resulting in two-layered estuarine circulation with fresh water outflow at the surface and deep, near bottom inflow. Forty one Lehigh gravity and box cores were collected at locations indicated in Fig. 1 between 1999 and 2001. Sediment cores were subsampled at 1–2 cm intervals, homogenized and divided into aliquots for radionuclide, metal and organic contaminant analyses. 210 Pb was measured on freeze-dried sediment samples by alpha counting of 210Po electrodeposited onto nickel discs while the 226Ra supported 210Pb levels were determined by Ge gamma ray detection methods (Smith and Walton, 1980). 137Cs was measured on dried sediment samples using a hyper-pure Ge, gamma ray detector having a 1 cm diameter well. Metals were analyzed following acid digestion using multi-element ICP-MS techniques (Smith and Ellis, 1982). Detailed porosity profiles were determined for each core and used to determine the integrated mass of dry sediment (mass–depth; g cm− 2) above each depth (cm) horizon.

Fig. 2. (Upper panel). Sedimentation rate determined from least squares, exponential fits to 210Pb distributions for Stas. 1b, 4b and 19a were used to estimate core sedimentation rates ex given in Table 1. (Lower panel). 137Cs geochronologies for cores 1b, 4b and 19a were calculated using the 210Pb sedimentation rates (Table 1), the atmospheric input function (inset) introduced into Eq. (1) and a “systems time averaging” residence time of 15 years.

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Sediment samples for organic contaminant measurements were air-dried and sieved, mixed with anhydrous sodium sulphate and copper, placed in a cellulose extraction thimble, and spiked with surrogate internal standards. The sediments were soxhlet extracted with dichloromethane and purified using a Solid-Phase extraction column packed with silica gel. PAHs were eluted with hexane: dichloromethane and analyzed using high resolution gas chromatography coupled to a mass selective detector in the selective ion monitoring mode (SIM) as outlined previously (King and Chou, 2003). A duplicate, certified reference material (NIST 1944) and operational blank was routinely performed with each batch of 10 samples. Total PAH levels (hereafter referred to simply as “PAH levels”) are the sum of 16 PAHs, including benzo(a)pyrene, anthracene, pyrene, etc. (King and Chou, 2003). Total PCBs (hereafter referred to as “PCB levels”) are the sum of 159 congeners. 3. Results and discussion 3.1. Core sedimentation rates and geochronologies Sedimentation rates and geochronologies were determined from Pb and 137Cs sediment depth distributions using previously outlined methods (Smith and Walton, 1980; Smith and Ellis, 1982). The sedimentation rate was calculated from the slope of a least squares exponential fit of excess 210Pb (total 210Pb minus 226Ra-supported 210 Pb) as a function of sediment depth and mass–depth (g/cm2), as illustrated for three cores (1b, 4b 19a) from the central part of Sydney Harbour in Fig. 2. The geochronology was then estimated from the 210 Pb-determined sedimentation rate. 137Cs was plotted on the geochronological scale determined from the 210Pb sedimentation rate (Fig. 2). Generally, the 137Cs threshold horizon and maximum should be in agreement with 210Pb dates of approximately 1952 and 210

1964 corresponding to the 137Cs fallout record (insert; Fig. 2). The absence of this agreement may indicate that the core has undergone mixing by processes such as bioturbation or be otherwise undatable (Smith, 2001). However, deviations of the 137Cs core profiles from the fallout record can also result from 137Cs transport and retention in different phases of the environment (Smith et al., 1987). The transport of 137Cs and other airborne contaminants into Sydney Harbour occurs both directly, by atmospheric inputs to the water column and indirectly, by runoff from the surrounding watershed. The retention of 137Cs in the water column and soils of the drainage basin introduces delays in 137Cs transport prior to permanent deposition in the sediments. These delays can produce a 137 Cs maximum at a later 210Pb date compared to the year (1963–64) of actual maximum delivery of fallout 137Cs to the Earth's surface and a “tail” in the 137Cs distribution skewed upward towards the sediment surface (Smith and Schafer, 1999). Several investigators (Dominik et al., 1987; Robbins et al., 2000) have employed one component box models in which 137Cs is transported through a well-mixed reservoir prior to deposition in lake and marine sediments described by; dFs = dt = λs Fa −½λs + λFs

ð1Þ

where Fa is the atmospheric flux of 137Cs, Fs is the flux from the reservoir, λ = ln 2/t1/2 (where t1/2 is the 30.2 years half life of 137Cs) and λs = 1/Tswhere Ts is the residence time of 137Cs in the reservoir. Robbins et al. (2000) used a value of Ts = 15 years to simulate 137Cs distributions in sediments from a central Florida bay. The 137Cs distribution calculated using this model for Ts = 15 years is compared in Fig. 2 with the 137Cs geochronology for cores 1b, 4b and 19a showing reasonable agreement with the slopes of the “tail” of the experimental distribution. Robbins et al. (2000) refer to this process as “system time averaging”, because the reservoir in which the tracer

Fig. 3. Upper left panel: Distribution of sedimentation rates (ω) for Sydney Harbour shows highest values in Muggah Creek and central harbour regions; Upper right panel: Sediment inventory map for Pb exhibits similar features to sedimentation rate map. Bottom left panel: Map of PAH inventories shows highest values in Muggah Creek and the lowest at harbour entrance. Bottom left panel: PCB inventory map is similar to PAH inventory map.

J.N. Smith et al. / Science of the Total Environment 407 (2009) 4858–4869 Table 1 Sedimentological data and contaminant inventories for Sydney Harbour cores. Longitude ω

210

Pb

PAHs

PCB

°N

°W

cm/ year

dpm/ cm2

mBq/ cm2

µg/ cm2

µg/ cm2

µg/ cm2

46.14645 46.15675 46.15827 46.15753 46.18257 46.18223 46.15315 46.12883 46.13560 46.22238 46.17642 46.18470 46.15095 46.15137 46.16635 46.18467 46.14645 46.15000 46.1531 46.14083 46.15827 46.15095 46.15550 46.16865 46.17567 46.14383 46.14883 46.13850 46.13217 46.19433 46.17550 46.15483 46.14450 46.14467 46.15117 46.15167 46.15200 46.15117 46.19083 46.17150 46.16433

60.20748 60.20840 60.20977 60.21403 60.20842 60.21437 60.21143 60.20622 60.20075 60.21583 60.27735 60.22228 60.21468 60.21068 60.20608 60.22017 60.20748 60.20783 60.2074 60.20183 60.20977 60.21468 60.21500 60.21522 60.21167 60.20583 60.21367 60.20033 60.20267 60.21700 60.20417 60.21000 60.20733 60.21217 60.20050 60.20233 60.20333 60.20067 60.21750 60.20617 60.21217

0.631 0.382 0.45 0.515 0.427 0.423 0.244 0.207 0.403 0.209 0.109 0.486 0.556 0.282 0.417 0.475 0.628 1.023 0.444 0.303 0.582 0.637 0.734 0.506 0.542 0.637 1.08 0.662 0.299 0.556 0.417 0.244 0.426 0.514 1.82 1.98 4.55 1.072 0.353 0.545 0.479

61 41 64.9 78.2 55.5 51.9 20.9 19.3 44.2 34.3 36.6 61.5 62.5 27.4 49.4 90.5 68.8 133 12.6 23.2 61.9 101 97.2 120 84.4 83 149 54.1 16.6 62.5 49.4 20.9 38.3 51.7 123.4 180.8 280.3 67.5 40.6 45.8 43.5

226 170 131 129.6 39.5 66.8 37.8 25 121.0 48.9 46.9 68.6 76.8 44.2 61.5 102.3 356.8 390.9 60 40.5 69.7 126.9 168.8 127.3 86.8 313.2 254.9 180.5 29.9 69.2 59.7 92 227.8 137.2 511.8 421 375.3 411.1 66.2 29.7 75.4

5298 3277 2458 2817 1662 726 1598 1804 3249 289 703 1362 2956 2643 2621 1868 3529 4507 3784 2656 2336 2740 2951 2098 1656 3324 2708 6681 2031 995 1707 2200 4422 3954 4725 7773 6432 8876 1357 2710 2426

5529 – 2551 – – – – 881 2516 1214 164 – – 2536 2023 857 – 9905 2202 2019 – – 1609 662 469 6211 3533 – – – – – – – 225877 141477 195844 145937 – – –

19.70 – 7.72 3.59 – – 1.95 0.95 8.48 0.20 0.18 – – – 3.59 2.59 – 31.20 – 1.83 – 9.78

Station Latitude

1a 3a 4a 5a 6a 7a 17a 18a 19a 23a 26a 28a 29a 30a 31a 843 1b 4b 7b 10b 12b 18b 20b 21b 26b 31b 35b 5c 6c 14c 16c 46c 6d 7d 13d 14d 15d 16d 23d 24d 36d

Pb

137

Cs

4861

simulated by a systems time averaging model and a time constant of 15 years. In the present study, “system time averaging” probably reflects mean particle circulation times within the water column of Sydney Harbour and the erodable soils of the drainage basin where particles can undergo multiple cycles of deposition and resuspension before undergoing permanent burial in the sediments (Smith and Schafer, 1999). In a study of particle size measurements in surficial sediments (Stewart et al., 2001), sediment regimes were identified along the western shoreline of Sydney Harbour that are particularly subject to sediment resuspension and which may constitute a transient repository zone for particle-reactive contaminants. The distribution of sedimentation rates for Sydney Harbour (Fig. 3; Table 1) shows elevated values (N0.5 cm/year) in the deeper central channel. This is consistent with the results of a benthic boundary layer transport model driven by in-situ, current observations indicating that transport of resuspended bottom particles towards the head of the harbour occurs over distances of 5–10 km (Petrie et al., 2001). The highest sedimentation rates in Sydney Harbour (N5 cm/year) were measured in Muggah Creek which receives direct outflow from the Tar Pond, while the lowest were measured in the seaward approaches to the harbour where the sediment cover becomes increasingly coarsegrained. 3.2. Pb geochronology

4.32 – – – 8.40 – 0.52 – – – – – 151.20 – – – – –

Uncertainties in sedimentation rates and radionuclide inventories average ± 8%.

undergoes delays is usually poorly defined. However, these lags in 137 Cs transport are a common feature of most coastal marine systems. Most of the 41 cores collected in Sydney Harbour exhibited good agreement between 210Pb geochronologies and 137Cs distributions

A Pb geochronology for a typical core from the central channel (core 1a), illustrated in Fig. 4, shows an increase from background levels of 30 µg/g in the early 1900s to a broad maximum during the 1960s–1980s characterized by Pb concentrations in excess of 400 µg/g. The signal then decreases by almost a factor of 2 through the 1990s to the core surface. This is notably different compared to the record of atmospheric delivery of Pb from the combustion of leaded gasoline which peaked in 1973 with the change to unleaded gasoline (Nriagu, 1989). As exhibits a similar geochronology (not shown), increasing from background levels of about 10 µg/g to maximum levels of 70–100 µg/g during the 1960s–1980s. These metal geochronologies are in contrast to the Ag geochronology (Fig. 4) which exhibits relatively constant concentrations in recently deposited sediments following large increases in the 1950s. Ag is a tracer of sewage discharges and its geochronology reflects the history of population growth in the region while Pb and As geochronologies are consistent with the history of local coke production. Coking operations began in the early 1900s and increased their production rates to maximum values during the 1970s and early 1980s before closure in 1988 (Barlow and May, 2000). During the coking operation, the temperature of the coal is increased to about

Fig. 4. Sediment geochronologies for Pb, PAHs and PCBs in core 1a from the central harbour all exhibit maxima in the 1970s–1980s associated with discharges from the steel and coke industries. In contrast, Ag is associated mainly with sewage discharges and tends to track urbanization and population growth.

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1000 °C under low oxygen conditions, at which point much of the arsenic and lead is evaporated as species such as As2, AsN, Pb and PbCl2 (Furinsky, 2000). Upon cooling these species adhere to fine particulates that are released to the atmosphere with coal oven gases. The detection of elevated levels of Pb and As in soils and house dust in Sydney indicates that significant stack discharges of these contaminants have occurred from the coke ovens (Lambert and Lane, 2004). An inventory map for anthropogenic Pb (background inventories removed) is illustrated for Sydney Harbour in Fig. 3. Pb inventories are greatest in the central region of the harbour, close to the main sewage outfalls that drain the watershed for the steel plant and coke ovens (Fig. 1). A total Pb inventory for the Harbour of 284 tonnes (Table 2) was calculated from the Pb inventory map using linear interpolation to estimate contaminant concentration gradients between different stations. 3.3. Future projected Pb concentrations The post-1983 rates of decline in Pb inputs to the harbour following the closing of the coke ovens are illustrated for a suite of six cores on a semi-log scale in Fig. 5 that are representative of the range of sedimentation rates and contaminant inventories (Table 1) measured in Sydney Harbour. Since most of the Pb contamination occurs via local atmospheric transport, Eq. (1) provides an appropriate model for characterizing Pb uptake in the sediments. The decline of Pb concentrations following the cessation of atmospheric inputs in the mid-1980s can be approximated in most cores by an exponential curve having a time constant of about 15 years. Since this is the same time constant employed for 137Cs, it can be assumed that the same type of “systems time averaging” occurs for particle-bound Pb. This time constant represents a residence time for Pb in erodable or leachable soil and/or sediment phases from which it is subject to release according to first order processes followed by permanent deposition in harbour sediments. Robbins et al. (2000) similarly found that they could explain the uptake of atmospherically-derived Pb in corals and sediments in marine embayments in Florida using first order processes and the same residence time of 15 years employed for 137 Cs. The utility of this model is that future surface Pb concentrations can be predicted by assuming that Pb concentrations in each core will continue to decrease exponentially from their 1990 values with a time constant of 15 years. This is accomplished by curve fitting an exponential function to the Pb concentration profile for each core at the 1990 time horizon. For some cores, future Pb projections are based on as few as three post-1985 data points, but it is assumed that the large number of cores employed in this study will average out inconsistencies in individual cores. The range of future projected Pb concentrations is indicated by the cross hatched area in Fig. 5.

Fig. 5. Pb geochronologies in a suite of representative Sydney Harbour cores exhibit exponentially decreasing concentrations with time following closure of the coke ovens in the 1980s with a time constant of about 15 years. Hatched region represents predicted range of future Pb surface sediment concentrations.

for dates not directly corresponding to an actual Pb measurement. These results can be used to identify periods when Pb concentrations in specific regions of the Harbour exceeded environmental quality guidelines. The National Oceanographic and Atmospheric Administration (NOAA) has proposed different criteria including the effects range-low (ER-L) corresponding to background concentrations and below which the presence of contaminants has little chronic or acute effect on benthic organisms and the effects range-medium (ER-M) above which organisms are very likely to be negatively affected by the presence of a contaminant (Jones et al., 1997). The ER-L and ER-M for Pb are 46.7 µg/g and 218 µg/g, respectively which correspond to the blue and yellow contours, respectively in Fig. 5. Pb levels are observed in Fig. 5 to exceed the ER-L coincident with the construction of the steel plant in the early 1900s. They increased above the ER-M throughout most of the harbour following the expansion of the steel and coking facilities in the 1950s and 1960s. Although Pb levels declined during the 1980s there were still many locations where Pb levels in surface sediments exceeded the effects range-medium (ER-M) by 2000. The projected Pb concentration map for 2020, based on the systems time averaging model and a residence time of 15 years for each core indicates that by 2020 surface concentrations of Pb will have fallen below the effects range-medium (ER-M) level at all core locations in the Harbour. 3.5. PAH geochronologies

3.4. Pb concentration mapping The spatial history of Pb contamination of Sydney Harbour is illustrated by surface Pb concentration maps in Fig. 6. Pb geochronologies for each core were used to estimate the sediment surface concentration of Pb at decadal intervals between 1880 and 2000. Linear interpolation was used within each core to estimate Pb values

PAH geochronologies for the same suite of cores used to characterize the declines in Pb concentrations are illustrated in Figs. 4 and 7. Total PAH levels increased from relatively low values in the early 1900s to maximum values extending into the early 1980s. The geochronologies for PAHs are similar to those for Pb (Fig. 4) with most cores exhibiting intra-core correlations between Pb and PAH

Table 2 Total sediment loads (tonnes) of contaminants in Muggah Creek and Sydney Harbour.

Sydney Harbour Muggah Creek Total

As (tonnes)

Ag (tonnes)

Cu (tonnes)

Pb (tonnes)

Zn (tonnes)

Aliphatic PAHs (tonnes)

Aromatic PAHs (tonnes)

PAHs (tonnes)

PCBs (tonnes)

32.0 0.24 31.2

1.20 0.04 1.24

88.7 1.89 90.6

281 2.5 284

376 5.6 382

60 41 101

114 41 155

234 134 368

0.454 0.070 0.524

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Fig. 6. Reconstruction of historical Pb surface concentration maps is based on Pb geochronologies for sediment cores. Extrapolation of Pb concentrations to 2020 was estimated from the model using a Pb time constant of 15 years.

concentrations characterized by least squares, regression coefficients of r2 N 0.9. A high correlation between Pb and PAHs has also been measured in soils and airborne dust particles in the Sydney region indicating that PAH contamination is also derived from the steel mill/ coking operations (Lambert and Lane, 2004). The PAH assemblages are dominated by higher molecular weight, four to six ring structures

which tend to be the most toxic and persistent. The anthracene/ phenanthrene ratio always exceeded a value of 0.2 which is a clear sign of the presence of combustion compared to petroleum products (Yunker et al., 2002). The spatial distribution of PAH inventories in Sydney Harbour (Table 1) is illustrated in Fig. 3. The highest inventories were measured

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leakage of Pb contamination from the Tar Pond into Sydney Harbour is considered to be much less significant. The inventory of PAHs in Muggah Creek (ca. 0.035 km2) based on the mean value for inventories measured in cores 13d, 14d, 15d, and 16d (ca. 177 mg cm− 2) is about 63 tonnes (Table 2). However, these cores were not sufficiently long to penetrate the pre-1900 region of background PAH levels that generally lie below 90 cm in Muggah Creek sediments. If a standard PAH curve is estimated from mean PAH geochronologies for Sydney Harbour sediments, then the additional PAH loadings in deeper sediments can be calculated for Muggah Creek cores. It is estimated that the PAH inventories are 136 tonnes in Muggah Creek and 234 tonnes in Sydney Harbour giving a total of about 370 tonnes (Table 2). This is approximately 10% of the 3500 tonnes of PAHs estimated to reside in the Tar Pond and its surroundings. 3.6. PAH concentration maps

Fig. 7. PAH geochronologies exhibit exponentially decreasing concentrations with time following closure of the coke ovens in the 1980s with a time constant of about 10 years. Hatched region represents predicted range of future PAH surface sediment concentrations.

in Muggah Creek where they exceed values of 100,000 µg/cm2. PAH sediment inventories decrease by several orders of magnitude with increasing distance from the head of Muggah Creek as illustrated in a log–log plot in Fig. 8. Also shown are log–log plots for the sedimentation rate (ω) and the Pb inventory. Despite the high intra-core correlation with Pb, PAH sediment inventories decrease much more rapidly with increasing distance from Muggah Creek compared to those for both ω and Pb. This suggests that, although the time dependent input functions for both contaminants in Sydney Harbour are similar, different mechanisms govern the transport of the two contaminants through the water column. The similar slope for Pb inventories and ω as a function of distance in Sydney Harbour is consistent with an atmospheric Pb source distributed uniformly across the harbour and indicates that the variance in Pb inventories is governed mainly by sedimentation and particle dynamics rather than the spatial configuration of the input function. In contrast, the extremely high levels of PAHs in Muggah Creek relative to other regions of the Harbour suggest that the dumping of coke wastes into the Tar Ponds followed by leakage into Muggah Creek constitutes an additional and probably larger source for PAHs in Sydney Harbour. Since maximum Pb levels in Muggah Creek are about the same or lower than those at proximal stations (eg. 1a, 4b) in the central harbour for the same time-stratigraphic horizons,

PAH levels measured in Muggah Creek were all in excess of 1000 µg/g and values of 10,000 µg/g were measured at Station 15d. PAH levels in the 1960s–1980s stratigraphic intervals of the sediment cores decrease with increasing distance from Muggah Creek to levels of the order of 200–500 µg/g in the central part of the harbour (eg. Sta. 1a) and then to levels of about 10 µg/g at Station 26a in the North Arm. These can be compared to sediment PAH levels of 200 µg/g measured in the Saguenay Fjord, Quebec (Smith and Levy, 1990), levels of 169 µg/g and 87 µg/g in Baltimore and Boston Harbours, respectively (Fostor and Wright, 1988; Laflamme and Hites, 1978), levels of 20–27 µg/g in Halifax Harbour (Gearing et al., 1991), levels of 12 µg/g in San Francisco Bay (Chapman et al., 1987) and background levels b1 µg/g measured in the Gulf of Maine (Laflamme and Hites, 1978). PAH surface concentration maps in Fig. 9 were determined from PAH geochronologies for each core following the same protocols as applied to the Pb geochronologies. For cores having incomplete PAH geochronologies (eg. some box cores do not contain pre-1940 sediments), the PAH distribution was determined using the PAH/Pb regression and the Pb geochronology for that core. The ER-L and ER-M for PAHs (4.02 and 44.8 µg/g, respectively; [Jones et al., 1997]) correspond to the blue and yellow contours, respectively in Fig. 9. It can be seen that by 1960, PAH concentrations exceed the ER-M in most cores from Sydney Harbour. 3.7. Future projected PAH concentrations PAH concentrations in most cores (Fig. 7) decreased by about a factor of 2 following the termination of coking activities in the 1980s.

Fig. 8. Log–log plots of core Pb inventories as a function of distance from head of Muggah Creek exhibit a similar slope to those of the sedimentation rates (ω), while the slope for PAH and PCB inventories is much steeper indicating that Muggah Creek represents a significant source for the latter contaminants.

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Fig. 9. Reconstruction of historical PAH surface concentration maps is based on PAH sediment geochronologies. Extrapolation of PAH concentrations to 2020 was estimated from the model using a time constant of 10 years.

This relative decline in PAH concentrations with time is better simulated by an exponential curve having a time constant of 10 years (Fig. 7), compared to the time constant of 15 years used for 137Cs and Pb. The smaller time constant for PAH decreases reflects the fact that PAH transport into the harbour occurs mainly by direct discharge into Muggah Creek rather than from an atmospheric pathway. The 10 year

time constant for PAHs reflects “systems time averaging” through multiple cycles of resuspension and deposition solely within Sydney Harbour itself and does not include delays associated with the transport of atmospherically deposited contaminants through the drainage basin. Future surface PAH concentrations were predicted for each core assuming that PAH concentrations in each core will

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continue to decrease exponentially from their 1990 values with a time constant of 10 years. The range of future projected PAH concentrations is indicated by the cross hatched area in Fig. 7. The projected PAH surface concentration map (Fig. 9) indicates that by 2020 surface concentrations will still exceed the effects range-medium (ER-M) at core locations in the central part of the harbour including Muggah Creek. 3.8. PCB geochronologies Polychlorinated biphenyls (PCBs) were routinely used in electrical systems in the steel mill/coking operations in Sydney and subsequently discharged in sewers and runoff into the harbour. About 85% of the PCB congeners measured in the sediments were hexa or hepta chlorinated species which are more chemically inert and particle reactive compared to their lighter counterparts and tend to follow particle transport pathways through the environment. The major PCBs can be represented as 153, 138, 187, 170 and 180 based on literature comparison to Aroclor 1260 (the main source in sediments) and are considered among the most toxic in the environment. PCB geochronologies for the same suite of cores used to characterize the recent declines in 137Cs, Pb and PAH concentrations are illustrated in Fig. 10. PCB levels increased from low values (b0.005 µg/g) in sediments deposited prior to their introduction in industrial compounds in the 1940s to maximum values (N1 µg/g) during the 1970s–1980s and then declined rapidly in sediments deposited following a ban on the usage of PCBs in 1977 (Valette-Silver, 1993). PCB levels in excess of 1 µg/g have also been observed in other highly contaminated sediment regimes such as the Sangamo–Weston/ Twelvemile Creek/Lake Hartwell Superfund site in South Carolina (Brenner et al., 2004). However, maximum PCB levels of about 0.04 µg/g measured in sediment cores from San Francisco Bay (Venkatesan et al., 1999) and Tampa and Galveston Bays (Santschi et al., 2001) are more typical of sediments in which PCB inputs are related to the more general urbanization/industrialization of the region rather than to inputs associated with a specific industrial source. The geographical distribution of PCB inventories in Sydney Harbour is illustrated in Fig. 3. The highest inventories were measured in Muggah Creek where they exceeded values of 100 µg/cm2. As with PAHs, PCB sediment inventories decreased by several orders of

magnitude between Muggah Creek and the outer harbour (Fig. 8) indicating that their direct discharge into the tar ponds and Muggah Creek constitutes the most significant source of PCBs in Sydney Harbour. The PCB inventory in Muggah Creek estimated from core 14d is 0.054 tonnes, which increases to 0.070 tonnes after extrapolating from the measured distributions to background levels (Table 2). The PCB inventory in harbour sediments outside Muggah Creek is 0.454 tonnes giving a total of 0.52 tonnes of PCBs in Sydney Harbour. This is approximately 14% of the PCB inventory of 3.6 tonnes estimated to reside in the tar pond. 3.9. PCB concentration mapping PCB surface concentration maps (Fig. 11) were determined from sediment geochronologies using the same protocols applied to Pb and PAH geochronologies. The ER-L and ER-M for PCBs (0.023 and 0.180 µg/g, respectively; (Jones et al., 1997)) correspond to the blue and yellow shadings, respectively in Fig. 11. PCB concentrations increased to levels exceeding the ER-M throughout the central harbour by 1960 and had only fallen below the ER-M at stations on the western side of the harbour by 2000. Similarly to PAHs, the post-1980s decline in PCB concentrations with time can be approximated by an exponential curve having a time constant of 10 years (Fig. 10). Since PCBs were dumped directly into Muggah Creek and the tar ponds together with PAHs, it can be assumed that the 10 year time constant for PCBs reflects similar processes governing particle cycling through transient depositional regimes. Future surface PCB concentrations were predicted for each core assuming that concentrations will continue to decrease exponentially from their 1990 values with a time constant of 10 years. The range of future projected PCB concentrations is indicated by the cross hatched area in Fig. 10. The combined results for Sydney Harbour (Fig. 11) indicate that by 2020 surface sediment concentrations will still remain above the ER-M in the central harbour and Muggah Creek. Brenner et al. (2004) also modeled PCB concentration changes in sediments from a highly contaminated “Superfund” site in South Carolina where PCB levels were in the same range as those measured in Sydney Harbour. They fitted the measured PCB decreases with sediment depth with exponential curves that correspond to time constants for the natural recovery of the sediments of about 6–8 years, comparable to the time constant of 10 years used for PCB declines in Sydney Harbour. They predicted that PCB levels would decline below a given level of 0.05 µg/g at various core locations between 2009 and 2041. 3.10. Natural containment of contaminant inventories

Fig. 10. PCB geochronologies for a suite of representative Sydney Harbour cores exhibit exponentially decreasing concentrations with time since the 1980s with a time constant of about 10 years. Hatched region represents predicted range of future PCB surface sediment concentrations.

Sydney Harbour contaminant inventories are undergoing natural containment by the continuous deposition of less contaminated sediments. Contaminants such as Pb, PAHs and PCBs, discharged prior to the closure of the steel facilities, are still cycling through depositional regimes in the harbour on time scales of 10–15 years, but their concentrations in recently deposited sediments are gradually decreasing. As a result, the main inventory of Pb, PAHs and PCBs is being buried at a rate of 0.2–2 cm/year throughout the harbour. In the regions closest to Muggah Creek, where sedimentation rates tend to be the highest, the contaminant inventories are presently separated from the sediment surface by about 5–15 cm of less contaminated sediments and in some cores, the maximum PCB and PAH levels occur at depths of N50 cm. Using the modeling approach outlined above, future mean contaminant concentrations have been predicted for the upper 5 cm of the sediments at each station in Sydney Harbour. The choice of a 5 cm depth interval is dictated by the fact that this represents the principal habitat for sediment infauna whose health is of concern.

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Fig. 11. Reconstruction of historical PCB surface concentration maps is based on PCB sediment geochronologies. Extrapolation of PCB concentrations to 2020 was calculated assuming a residence time of 10 years.

The results for Pb are illustrated in Fig. 12 (upper) which map the year that Pb levels will decrease below the ER-L (46.7 µg/g) and the ER-M (212 µg/g) using a time constant of 15 years. These results show that throughout the entire harbour Pb levels will have fallen below the ER-M and the ER-L by 2020 and 2050, respectively. PAH and PCB concentrations were modeled using a shorter time constant of 10 y. Fig. 12 (middle) shows that PAH concentrations will fall below the ER-M (44.8 µg/g) in the central region of the harbour by 2030, but will continue to remain above the ER-M in Muggah Creek until 2060. PAH levels will only decline below the ER-L (4.02 µg/g) in the central part of the Harbour by 2060 and will remain above the ER-L in Muggah Creek until 2090. PCB levels (Fig. 12; bottom) will also fall below the ER-M (0.18 µg/g) in the central parts of the harbour and in Muggah Creek by about 2030 and 2060, respectively. These results suggest that Pb contamination of Sydney Harbour sediments does not represent a long term threat and that natural containment will reduce Pb levels below those predicted to have a significant impact on organisms during the next 10–20 years. PAH and PCB levels will have declined below the ER-M by about 2030

in almost all parts of the harbour. The sole exception is Muggah Creek where PAH and PCB levels will remain above the ER-M until 2060. Anthropogenic remediation issues should be considered in the context of these results. Future dredging of harbour sediments intended to accommodate larger vessels must be evaluated in the context of the magnitude of contaminant resuspension and enhanced bioavailability that will invariably result from this type of operation. Present remediation activities in the tar ponds may also result in future releases of PCBs and PAHs that could contaminate the layer of cleaner surficial sediments that has presently “capped” the main inventory of contaminants deposited during the 1960–1980s. One additional caveat is that as the harbour sediments become less contaminated, they may become more hospitable environments for benthic organisms. This could result in the development of the larger and more diverse benthic communities that are typical of the less contaminated, outer reaches of Sydney Harbour (P. Stewart, pers. comm.). Enhanced concentrations of infauna could result in higher rates of bioturbation and the return of contaminated sediments more

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Fig. 12. Maps of year in which Pb (top figures), PAH (middle figures) and PCB (bottom figures) concentrations in surface sediments decrease below the effects ranges, ER-L and ER-M which are given in the text.

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