Historical trends in trace metal and sediment accumulation in intertidal sediments of Moreton Bay, southeast Queensland, Australia

Chemical Geology 300-301 (2012) 152–164

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Research paper

Historical trends in trace metal and sediment accumulation in intertidal sediments of Moreton Bay, southeast Queensland, Australia Guia Morelli a,⁎, Massimo Gasparon a, b, Daniela Fierro c, Wan-Ping Hu a, Atun Zawadzki c a b c

School of Earth Sciences, The University of Queensland, St Lucia Qld., 4072, Australia National Centre for Groundwater Research and Training, Australia ANSTO, Institute for Environmental Research, New Illawarra Rd, Lucas Heights, 2232, NSW, Australia

a r t i c l e

i n f o

Article history: Received 18 August 2011 Received in revised form 19 January 2012 Accepted 23 January 2012 Available online 1 February 2012 Editor: J.D. Blum Keywords: 210 Pb 137 Cs Intertidal sediments Heavy metals Moreton Bay Australia

a b s t r a c t Temporal trends in heavy metal pollutants were reconstructed from the analysis of four sediment cores collected from intertidal areas in Moreton Bay, southeast Queensland, Australia. The geochronology of the past ~ 150 years was established using short-lived radionuclides 210Pb and 137Cs. The 210Pb-derived sedimentation rates varied from 0.16 ± 0.01 g/cm 2/y to 0.71 ± 0.30 g/cm2/y, indicating that sediment deposition is spatially highly variable across the bay. Increases in sedimentation rates over the past years are in agreement with the period of major development in the area and land use intensification after European settlement. Geochemical pre-European trace metal backgrounds in the bay's sediments could be established from the integration of geochronological data, down-core heavy metal concentrations, sediment properties, and the known historical events. Crustal elements such as Al, Fe, Ti, and rare earths were used as proxies to identify changes in sediment sources. The increasing temporal trends in Pb, Zn, Cd, and Ni correlated to the major development of Moreton Bay catchments (deforestation for agriculture, industrialization, and urban expansion) in the last century. Metal concentrations typically increased from about one to two orders of magnitude since about 1920 compared to background levels. Results show that sedimentation changes and enrichment in metal contaminants since European arrival are preserved in the sedimentary record. Significant spatial and temporal variability, however, indicate that the correct assessment of contamination in sediment cores in complex estuarine-marine embayments requires the careful integration of different proxies. © 2012 Elsevier B.V. All rights reserved.

1. Introduction Sediments constitute important carriers for pollutants in coastal environments. Estuarine embayments may become important reservoirs for contaminants, as they tend to act as sinks for fine, contaminant-reactive sediments (e.g. Morrisey et al., 2000; Swales et al., 2002; Liaghati et al., 2003). Heavy metals are used in all kinds of industrial activities, which have increased metal inputs in coastal areas, and their inherent toxicity and non-degradability is the basis for monitoring ecologic impact (Lacerda et al., 1992). Reworking of old sediments through natural, physical, biological, and human activities, such as dredging or mixing, can easily release into the environment anthropogenic metals associated with old sediments, with negative impact on the ecosystem (Morrisey et al., 2000; Swales et al., 2002). Sediment cores are environmental archives of past anthropogenic pollution, and the study of heavy metal profiles in undisturbed sediments can provide a complete record of temporal contamination history (e.g. Nriagu, 1979; Weiss

⁎ Corresponding author. E-mail address: [email protected] (G. Morelli). 0009-2541/$ – see front matter © 2012 Elsevier B.V. All rights reserved. doi:10.1016/j.chemgeo.2012.01.023

et al., 1999; Smith, 2001; Cundy et al., 2003). Most of our current knowledge on the history of metal pollution is based on studies of sediment cores from lakes and coastal areas, snow, and ice deposits. Metal contaminants in the Greenland ice sheet have been traced to Roman mines and smelters (Hong et al., 1996; Rosman et al., 1997; Ruppert and Deicke, 2006), and historical metal fluxes have been recorded in lakes and coastal intertidal sediments both in the northern hemisphere (Greenland, North America and Europe; Shotyk et al., 1996; Weiss, et al., 1999; Swales et al., 2002; Cundy et al., 2003) and in the southern hemisphere (Connor and Thomas, 2003; Gasparon and Matschullat, 2006a; Roychoudhury, 2007). Australia is a very good location for fieldwork to provide baseline data on anthropogenic impacts on environmental systems, thanks to its relatively short and well constrained history of industrialization and urbanization, compared to longer histories elsewhere. In the last two centuries, European settlement spread across the continent developing new practices that rapidly displaced those used by indigenous people. In a relatively short time-frame (from early 1800 until now) industrialization, agriculture, forestry, and urban expansion caused fast rates of change on pre-existing dynamics in the environmental systems (Dodson and Mooney, 2002). Besides, the geographic and climatic characteristics

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of Australia have concentrated urban expansion, with around 90% of Australian population living around the coastline (Dennison and Abal, 1999). Intense ecosystem degradation in estuaries and closed embayments is associated with large population centers. Thus, sediment cores obtained from those coastal areas provide a good chronologic record of contamination since European settlement. Temporal and spatial variations in contaminant levels, however, can be induced by natural sediment variability and pollutant supply and, also depend on environmental conditions at the water–sediment interface, as these control the overall chemical behavior and stability of the different species. The geologic substrate is the main natural source of inorganic chemical species into the ecosystem (e.g. Gasparon and Matschullat, 2006b), and different factors, such as catchment topography and hydrology, climate, and geographic location, determine the release of inorganic chemicals into the environment. Potential anthropogenic contamination is evaluated only in comparison with these natural background values (Matschullat et al., 2000; Dodson and Mooney, 2002; Reimann and de Caritat, 2005; Reimann et al., 2005; Gasparon et al., 2007). To establish when contamination initially occurred absolute age is determined, and then pre-contamination values are used to calculate baseline levels and enrichment factors (natural or anthropogenic). 210 Pb (t1/2 = 22.26 years) has been used to date sediments deposited during the last 150 years (Koide et al., 1972; Appleby et al., 1979; Nittrouer et al., 1983/1984; Zuo et al., 1991; Appleby and Oldfield, 1992; Pfitzner et al., 2004; Ruiz-Fernandez and Hillaire-Marcel, 2009). Independent dating evidence, however, is needed as a validation of the 210Pb derived age/depth relationship (Nittrouer et al., 1983/1984; Lynch et al., 1989; Appleby and Oldfield, 1992; Smith, 2001; Simms et al., 2008), because variation in sediment accumulation rates, disturbed sediment record, mixing, and bioturbation can lead to a misinterpretation of 210Pb results, and thus incorrect interpretation of temporal pollution trends. 210Pb accumulation rates can be tested by comparison with the record of artificial fallout isotopes such as 137Cs (t1/2 = 30.14 years), which provides relatively reliable mass accumulation rates for sediments deposited in the last 50 years (Nittrouer et al., 1983/1984; Lynch et al., 1989; Zuo et al., 1991; Smith, 2001; Pfitzner et al., 2004; Pedersen et al., 2007). The interval covered by the decay of 210Pb corresponds to the onset of Australia's major urban and industrial development. Within this time frame, Moreton Bay in southeast Queensland is an ideal location to assess the impact of European settlement because of its ecologic significance, and because it has been one of the fastest growing regions of Australia during the past decades (Australian Bureau of Statistics, 2009). Previous investigations in Moreton Bay focused on nutrient distribution and their impact on the ecosystem (Dennison and Abal, 1999) and metal distribution in recent sediments (Clark, 1998; Preda and Cox, 2002; Cox and Preda, 2005). Changes in coral species in the last 200 years have been related to human impact (Lybolt et al., 2010), however to date no work has reconstructed the chronology and spatial extent of human impact from the depositional record in sediments. In this study we aimed for the first time to assess historical human impact as preserved in the marine sedimentary record since European arrival in Moreton Bay in 1824 (Steel, 1972; Capelin et al., 1998; Neil, 1998). A total of thirty-seven sediment cores were collected from various sites across the bay. This article focuses only on four cores that yielded an undisturbed 210Pb chronology. 210Pb and 137Cs activities depth profiles were determined, and sediment accumulation rates were obtained from 210Pb-calculated chronologies. Geochronological data, integrated with down-core trace element distributions were used to establish pre-European trace metal backgrounds in the bay's sediments, and to assess the extent of human impact and the main activities and timeframes associated with it.

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Predominantly crustal elements such as Al, Fe, Ti, and rare earths were used as proxies to identify changes in the sediment provenance and geochemistry. Enrichment factors for contaminants (Pb, Zn, Cd, Ni, and Cu) were calculated to evaluate post-European metal pollutant levels. 2. Material and methods 2.1. Study area Moreton Bay in southeast Queensland (27°S, 153°E) is a semienclosed area protected from the open ocean to the east by sand barrier islands Moreton, North and South Stradbroke (Fig. 1). The bay is one of Australia's largest estuarine systems, and comprises a total catchment region of 21,220 km 2, with five main rivers discharging in the bay area of only 1523 km 2 (Dennison and Abal, 1999). The catchment geology is formed by basement rocks made of Palaeozoic (Devonian–Carboniferous) metamorphosed sandstone, greywacke and minor shale and chert (Day et al., 1983). Mesozoic sedimentary rocks overlie the Palaeozoic basement, with Triassic and Jurassic sequences of sandstone and siltstone, and minor rhyolite and tuff. Younger basalts of Tertiary age and laterite form coastal plains and terraces in the Redcliffe Peninsula and in the Redland Area (Cox and Preda, 2005). The present marine embayment of Moreton Bay is the result of sea level oscillations that controlled sedimentary deposition since 6500 BP, when sea level reached the present height (Maxwell, 1970; Jones et al., 1978; Jones and Stephens, 1981; Flood, 1983; Stephens, 1992). The bay includes wetlands, intertidal zones bordered with mangroves, coral communities around the central islands, dune barrier islands, and swamps. For this particular physiography the Moreton Bay Marine Park was established in 1993 with sections recognized as “Wetlands of International Importance” (Ramsar Convention, 1971). In 2009 the protected “no take” zones were extended to cover 16% of the bay area, in order to prevent future degradation of the natural environment and its many animal species (birds, dugongs, turtles, and whales). The wellknown history of urban development and the well-documented natural events since European arrival make Moreton Bay an ideal location to quantify human impact. Since European settlement in 1824 (Steel, 1972; Capelin et al., 1998; Neil, 1998) a rapid destructive exploitation of the region's land resources started. After only ten years, all available timber had been cut (Johnston, 1988; Neil, 1998), although impact from large-scale industrial activities and urban settlements remained relatively low until the end of the 19th century. The degradation of the landscape reached its maximum at the beginning of the 20th century, when the area of cropland and pastures increased rapidly. Intensive land use along the coast and in the floodplains, together with wide-scale catchment modifications led to increasing sediments yields into Moreton Bay by a factor of four or five (Neil, 1998) together with nutrients and contaminants supply (Dennison and Abal, 1999). Human impact on the bay's sediment quality continued into the 21st century, following the rapid increase in urbanization and development of major infrastructures and industrial activities. Over the past decade this area has been one of the most rapidly growing regions in Australia with population higher than 2.7 million and expected to double by 2026 (Australian Bureau of Statistics, 2009). 2.2. Sampling Sediment cores were collected in four intertidal areas in Moreton Bay (Fig. 1), to cover the wide range of ecological habitat present (spatial variability) and to be representative of the different development and land use in each catchment. Core G6 (27°17′ S; 153°1′ E) was collected in May 2007 in the Pine River between two mangrove islands about 2 km from the estuary; core G31 (27°37′ S; 153°20′ E)

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Fig. 1. Map of Moreton Bay and location of the sampling sites.

in October 2008 on the west coast of Russell Island; while cores G37 (27°9′ S; 153°2′ E) and G34 (27°11′ S; 153°3′ E) were collected in April 2009, in Deception Bay at the mouth of Burpengary Creek (G37) and in front of the mangroves bordering south Deception Bay (G34). Sediment cores were obtained by gently pushing into the sediment 150–300 cm long aluminum pipes by hand at first, and then using a sledgehammer. The tubes were filled with seawater, and a test plug was inserted at the top, tightened, and topped up with seawater to ensure a tight vacuum. The cores were then pulled vertically out of the ground using a purpose-built steel lever. Care was taken to minimize disturbance and keep the cores in a vertical position during transport from the field to the laboratory. This sampling technique has been used successfully in a range of sediment types and environments and it yields undisturbed samples of unconsolidated sediments (Lynch et al., 1989; Moura et al., 2004). Compression of the sediments during cores extraction was minimized by using wide diameter (7.5 cm) and thin walled (~1 mm) aluminum tubes (Lynch et al., 1989). Sediment compaction was less than 10 cm for the muddy core G6, and 0–5 cm for the sandy G37, G34, and G31 cores. Cores were cut longitudinally, split in two with a nylon string to minimize

contamination, sliced at 1 cm intervals, and stored frozen until analysis (Loring and Rantala, 1992). The sample fraction used for metal analyses was selected from the inner part of the core, to avoid any contact with the aluminum pipe. 2.3. Analytical methods Selected sediment intervals were dated using the ANSTO (Australian Nuclear Science and Organization) facilities. The activities of radionuclides 210Pb and 137Cs were determined by Compton suppression gamma spectrometry using an active NaI(Tl) suppression annulus, a NaI(Tl) plug detector, and a reverse electrode germanium (REGe) detector all housed within an inert lead shield. The 210Pb activity was determined using the 46.5 keV photopeak and the 226Ra activity was estimated using 214Pb and 214Bi photopeaks at 351.9 keV and 609.3 keV, respectively. Unsupported 210Pb activity was calculated by subtracting 226Ra activity from 210Pb activity. The 137Cs activity was determined using the 661.6 keV photopeak after subtraction of the 214Bi photopeak interference (at 665.5 keV). The analyses were conducted on 30–40 g of sample. The sediment

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core chronology was determined using the Constant Rate of Supply (CRS) model (Appleby and Oldfield, 1983). Samples G6 0–6, G6 17–22, and G6 39–44 were analyzed using a similar setup in the VKTA — Nuclear Engineering and Analytics Inc., Dresden in Germany. Grain size was determined using a Malvern laser particle size analyser (located in the School of Chemical Engineering of The University of Queensland). Organic matter (OM) was estimated using the LOI method (Heiri et al., 2001) by heating 0.2 g of dry sample for 4 h at 550 °C. For elemental analysis sediments were oven dried at 60 C and grinded to a fine powder. Samples were digested under clean room conditions (Class 100). Complete dissolution was obtained by acid attack in screw-top Teflon beakers on a hot plate using a combination of ultrapure acids (HNO3 + HF, HNO3, HCl + HNO3) until complete dissolution was obtained. Trace element concentrations ( 60Ni, 65Cu, 66Zn, 114Cd, 208Pb) were determined by Inductively Coupled Plasma Mass Spectrometry (ICP-MS) using a Thermo X7 ICP-MS. Major elements (Al, Fe, and Mn) were analyzed by Inductively Coupled Plasma Optical Emission Spectroscopy (ICP-OES) using a Perkin-Elmer Optima 3300 DV ICP-OES. Both instruments are located in the School of Earth Sciences (The University of Queensland). Laboratory quality control consisted of analyses of blanks and international reference materials (AGV-2, MESS-3) in replicates for each digestion batch. Accuracy was within 2% for all metals for AGV-2 and less than 3% for MESS-3 (with the exception of Zn, 10% and Cd, 12%). Reproducibility for AGV-2 and MESS-3 was good (RSD b 5%) for all the elements of interest (except Zn, 9%, and Cd, 8% in AGV-2). Detection limits were several orders of magnitude lower than the

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elemental concentrations found in the samples (detection limits in pg/g are 15.5 for Ni, 19.3 for Cu, 27.9 for Zn, 1.1 Cd, 0.4 for Pb). 3. Results and discussion 3.1. Radioisotope dating The 210Pb and 137Cs depth profiles are shown in Fig. 2. Uncertainties are based on 1sigma counting errors and unsupported 210Pb and 137Cs activities are corrected to the reference age (date of collection). 3.1.1. 210Pb activities The in-situ production of supported 210Pb was obtained indirectly by measuring the activity of 226Ra using gamma spectrometry. The activity of total 210Pb was determined by measuring 210Pb directly using gamma spectrometry. Unsupported 210Pb was inferred by subtracting the activity of supported 210Pb from the activity of total 210 Pb. Total 210Pb activity was detected in the top 100–150 cm of all cores (Fig. 2a). Activities reached their maxima at or near the surface, ranging from a maximum of 56.7 ± 5.7 Bq/kg (G34) to a minimum of 23.9 ± 2.7 Bq/kg (G37). Unsupported 210Pb activities in the surface sediment vary from 34.5 ± 4.8 Bq/kg in core G34 to a minimum of 11.7 ± 1.7 Bq/kg in core G6, and show a decreasing trend with increasing depth in the top 50 cm. Cores G34, G6, and G31, however, have their highest unsupported 210Pb activities (42.3 ± 5.8; 16.2 ± 2.7; 37.4 ± 6.1 Bq/kg) in the sub-surface sediments. In core G6 unsupported 210Pb varies from 16.2 ± 2.7 Bq/kg in the upper portion of the

Fig. 2. 210Pb and 137Cs profiles in cores G37, G34, G6, and G31: a) Total and unsupported 210Pb activities (Bq/kg) versus depth (cm); b) 137Cs activities (Bq/kg) versus depth (cm). Horizontal error bars represent 1σ counting errors. Vertical error bars represent depth interval.

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Bay by Hancock (2001). The finer sediments (b20 μm fraction) analyzed by Hancock (2001) and the greater solubility of 137Cs in seawater (He and Walling, 1996; Ruiz-Fernandez and Hillaire-Marcel, 2009), as opposed to coarser sediments and slightly less saline estuarine water considered in this study, account for the differences.

sediment profile to activities lower than 8.5 ± 3.1 Bq/kg below 25 cm, with an almost exponential decrease in the 15–60 cm interval. In core G31 unsupported 210Pb activities in the top sediments are relatively high (~ 32.2 ± 5.8–37.4 ± 6.1 Bq/kg) and below 40 cm no longer show a decay profile, with values close to background levels. Overall, total 210Pb activities at the surface are considerably low, compared to activities found in other Pacific coastal zones (a summary of 210Pb activities along the coast of the Pacific Ocean is presented in Ruiz-Fernandez and Hillaire-Marcel, 2009), but are similar to activities found in sediments from north-eastern Australia (Pfitzner et al., 2004) or from Tasmania (Seen et al., 2004). Total and unsupported 210 Pb activities measured are consistent with activities obtained in three cores by Hancock (2001), collected in the central mud deposition area in Moreton Bay (total 210Pb activity ranging between 16.9 and 57 Bq/kg and unsupported 210Pb activity ranging between 15 and 43 Bq/kg). The high total 210Pb activities found in the lower sediments of cores G37, G34 and, G31 are likely related to the high fraction of fine particles in the sediments. Grain size is known to affect the adsorption of 210Pb in sediments, with finer sediments typically retaining higher amounts of 210Pb due to the higher total surface area (He and Walling, 1996).

3.1.3.

210

Pb geochronology and accumulation rates Pb-derived sediment geochronology was established using the Constant Rate of Supply (CRS) model in which a constant rate of 210Pb supply with time in the same location is assumed, regardless of the rate of sediment accretion (Goldberg, 1963; Appleby and Oldfield, 1978; Robbins, 1978). The use of this model was considered the most suitable for this study because in a complex depositional estuarine/marine environment, like Moreton Bay, atmospheric and river discharge are the two main sources of unsupported 210Pb (Farmer, 1991; Appleby and Oldfield, 1992) and sedimentation rates may have high variability in space and in time. Sediment ages and mass accumulation rates for cores G37, G34, G6, and G31 (Table 1) were calculated using unsupported 210Pb activities down to where the activity reached a minimum level. CRS age models are shown in Fig. 3. The two cores collected in Deception Bay (G37 and G34) have similar mass accumulation rates. The top 17 cm of G37 provide a geochronology of the past 85 ± 6 years. Mass accumulation rates vary between 0.19 ± 0.01 and 0.27 ± 0.15 g/cm 2/y in the younger sediments, with calculated sedimentation rates between 0.13 and 0.27 cm/y. The 137Cs activity profile shows peaks at 5 and 10 cm (1.4 ± 0.3 and 1.2 ± 0.4 Bq/kg) constraining the top 11 cm to post 1963. This data are consistent with the 210Pb age derived for this depth between 1961 and 1971. Compared to core G37, G34 has higher unsupported 210Pb activities in the top 9 cm, but below 12 cm activities are relatively low. Only the top 12 cm of the core provide a geochronology for the last 70 ± 3 years. Mass accumulation rates are similar to those of core G37, varying between 0.16 ± 0.01 and 0.24 ± 0.03 g/cm 2/y in the top sediments. G34's calculated sedimentation rates vary between 0.12 and 0.25 cm/y. Significant 137Cs activity is detected between 3 and 9 cm in core G34. However, no marked peak can be identified and the vertical profile shows a decreasing trend against depth. According to the 210Pb-derived age, 210

3.1.2. 137Cs activities In the Austral-Pacific region, 137Cs concentrations in sediments are five times lower than in the Northern Hemisphere (UNSCEAR, 2000; Pfitzner et al., 2004), and the first measurable 137Cs activities are detected in sediments deposited in 1954, with peak concentrations during 1963 corresponding to the Pacific ground nuclear weapon tests (Longmore et al., 1983; Wasson et al., 1987; Wallbrink et al., 2003; Wallbrink, 2004). In Moreton Bay fallout patterns of the 137Cs radionuclide show a peak in core G37, G6, and G31. Core G34's activity decreases with depth with a peak around 5 cm, however this is not statistically significant (see error bars in Fig. 2b). 137Cs peak activity ranges from a maximum of 3.2 ± 0.5 (Bq/kg) in core G31 (at 11.5 cm), to 2.4 ± 0.3 (Bq/kg) in core G6 (at 25.5 cm), 2.2 ± 1.1 (Bq/kg) in core G34 (at 5.5 cm), and a minimum of 1.2 ± 0.4 (Bq/kg) for core G37 (at 10.5 cm). Although activities are quite low, they are similar to those measured in marine sediments in central Moreton

Table 1 210 Pb derived accumulation and sedimentation rates. Sedimentation rates in cm/y are calculated from the age of each sediment intervals. ANSTO ID

Core

L439 L996 L997 L440 L998 L999 L433 L991 L434 L992 L993

G37 G37 G37 G37 G37 G37 G34 G34 G34 G34 G34 G6 G6 G6 G6 G6 G6 G6 G31 G31 G31 G31 G31 G31 G31

K897 K898 K899 K901 L113 L427 L428 L429 L430 L114 L432

Depth (cm)

Calculated CRS ages (years)

Years

0–3 4–6 7–8 10–11 13–14 16–17 0–3 3–5 5–6 7–9 11–12 0–6 9–10 12–13 17–22 25–26 33–34 39–44 0–3 5–6 11–12 19–20 25–26 29–30 39–40

5±3 18 ± 4 28 ± 5 42 ± 5 62 ± 7 85 ± 6 7±2 17 ± 2 26 ± 3 46 ± 3 70 ± 3 4±2 14 ± 3 20 ± 3 36 ± 4 55 ± 5 75 ± 4 101 ± 5 2±2 8±3 19 ± 4 30 ± 5 38 ± 6 43 ± 7 59 ± 6

2003 1990 1980 1966 1946 1923 2002 1992 1983 1963 1939 2003 1993 1987 1971 1952 1932 1906 2006 2000 1989 1978 1970 1965 1949

Time period

CRS model

2000–2005 1986–1994 1975–1984 1961–1972 1940–1953 1917–1929 2001–2004 1990–1995 1981–1986 1960–1966 1935–1942 2001–2005 1990–1996 1983–1990 1967–1975 1947–1956 1927–1936 1901–1911 2004–2008 1997–2003 1985–1993 1972–1983 1964–1976 1958–1971 1943–1955

0.27 ± 0.15 0.27 ± 0.06 0.26 ± 0.04 0.25 ± 0.03 0.22 ± 0.02 0.19 ± 0.01 0.23 ± 0.06 0.24 ± 0.03 0.21 ± 0.02 0.17 ± 0.01 0.16 ± 0.01 0.71 ± 0.30 0.66 ± 0.14 0.61 ± 0.10 0.54 ± 0.06 0.46 ± 0.04 0.44 ± 0.02 0.40 ± 0.02 0.72 ± 0.62 0.66 ± 0.25 0.61 ± 0.14 0.65 ± 0.11 0.68 ± 0.11 0.68 ± 0.10 0.67 ± 0.07

Mass accum. rates (g/cm2/y)

Sedimentation rate (cm/y)

0.27 0.27 0.25 0.15 0.20 0.13 0.23 0.25 0.17 0.12 0.15 0.71 0.65 0.50 0.44 0.31 0.40 0.31 0.72 0.64 0.58 0.70 0.80 0.74 0.63

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Table 2 Relative fractionation of LREE, MREE, and HREE expressed by the ratio of PAAS normalized La, Lu, and Gd in Moreton Bay sediment cores. The ratios represent the average of all sediments, the average of post-, and of pre-European sediments in each core following 210 Pb geochronology results.

G37 average G37 post-European G37 pre-European G34 average G34 post-European G34 pre-European G6 average G6 post-European G6 pre-European G31 average G31 post-1949 G31 pre-European

Fig. 3. a) Ages (years) against Cumulative Dry Mass (g/cm2) for cores G37, G34, G6, and G31. Average sedimentation rates calculated for each core are shown in cm/y. Error bars are based on 1σ error; b) depth (cm) against CRS-model derived dates. The slope of the curves shows different sedimentation rates for the four cores. Vertical error bars represent the interval of year as calculated from CRS model. Horizontal error bars are the dated sediment intervals.

sediments around 8 cm, being deposited between 1960 and 1966 should show a peak in 137Cs activity. The highest value of 137Cs is found at 4 cm (2.4 ± 0.6 Bq/kg) but sediments between 5 and 9 cm

Fig. 4. PAAS normalized REE patterns for the cores collected in Moreton Bay. Pre- and post-European sediments are distinguished based on 210Pb geochronology results.

La/Lu(PAAS)

Gd/Lu(PAAS)

La/Gd(PAAS)

0.82 0.81 0.71 0.77 0.78 0.75 0.49 0.81 0.21 0.58 0.55 0.75

1.39 1.36 1.37 1.36 1.25 1.38 0.72 0.69 0.82 0.83 0.83 0.85

0.59 0.60 0.52 0.57 0.80 0.54 1.24 1.24 1.22 1.42 1.45 1.32

have 137Cs activities still within the error range of the sediments above, consistent with the 210Pb calculated ages. The core collected along the Pine River (G6) provides the most complete geochronology (~100 years) in the top 60 cm. The relatively low surface activity (11.7 ± 1.7 Bq/kg) indicates a possible surface-mixing layer, where mass transfer at the water/sediment interface may affect the 210Pb values (Heijnis et al., 1987; Pfitzner et al., 2004). Calculated ages attribute the deposition of the top 44 cm to the last 100 ± 5 years (Tab. 1). Mass accumulation rates vary from 0.46 ± 0.04 and 0.71 ± 0.30 (g/ cm 2/y) in the top 26 cm, to 0.4 ± 0.02 (g/cm 2/y) below. Calculated sedimentation rates range from 0.31 cm/y to 0.71 cm/y in the younger sediments. The 137Cs activity depth profile shows the typical post bomb pattern, with activities varying from 1.4 ± 0.2 Bq/kg in the top sediment to a maximum peak of 2.4 ± 0.3 Bq/kg at 25 cm. The peak constrains the deposition of the top 25 cm of the core during or post nuclear bomb testing started in 1953 and peaked in 1963. Ages derived from the 210Pb model are in agreement with 137Cs results, indicating a sediment age of 55 ± 5 years at 25 cm, corresponding to the years 1947–1956. Core G31 collected in South Moreton Bay provides a geochronology for the top 40 cm (~59 ± 6 years). Below 55 cm the presence of large bivalve shells indicates a change in the depositional environment (storm deposit), therefore only the top 40 cm were considered to calculate the age model. Mass accumulation rates vary between 0.72± 0.62 g/cm2/y in the top 3 cm, to a minimum of 0.61 ±0.14 g/cm 2/y in the sediments below. Calculated sedimentation rates range between 0.63 and 0.72 cm/y. The high total 210 Pb activity below 80 cm correlates with a higher proportion of clay and silt (around 80–90% clay+ silt as compared to ~70% in the upper part of the core). 137Cs activity is detected in the top 30 cm with a peak at 11 cm (3.2±0.5 Bq/kg), and decreases to minimum values at 25 cm (2.0 ± 1.4 Bq/kg). As these two values are within the error range it is not possible to accurately relate the peak at 11 cm to the

Fig. 5. Vertical variability of Ti/Al, Fe/Al ratios, and Terrigenous index for the four cores collected in Moreton Bay.

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time of the expected 137Cs peak. According to the 210Pb age model, sediments at 25 cm were deposited in the years 1958–1971. The presence of higher 137Cs activity in the same interval is used to constrain the deposition of the top 30 cm to post-bomb period (from 1959). The uncertainty in the correlation between 137Cs peak (~1963) and the age determined from 210Pb in cores G34 and G31 is explained by the error in 137Cs detection, the low resolution represented by the sediment interval, and the sampling resolution. Furthermore, from 210Pb age results, 137Cs activity is expected to peak in the top 10 cm, where sediment/water interaction can possibly cause mobility of 137Cs. 137Cs is known to be relatively mobile in seawater (Stanners and Aston, 1981; Johnson-Pyrtle and Scott, 2001). The “ideal” profile of 137Cs could have been therefore disturbed and shifted in the sediments. 137Cs in the southern hemisphere has had a limited use as geochronological marker, due to lack of a significant yearly deposition maxima, to the lower concentration in sediments (Pfitzner et al., 2004), and for the absence of significant increase in activities following the Chernobyl accident in 1986 (Schuller et al., 2002). Longmore et al. (1983) and Hancock (2000) suggested that 137Cs in the Australian environment is only useful to provide a depth, indeed a fixed date, for the first detectable quantities of the radionuclide. Overall, 137Cs activities measured in this study are consistent with 210Pb CRS geochronology, and the 210Pb data indicate an undisturbed depositional environment.

sedimentary source and show that sedimentation is spatially and temporally consistent. Similar degrees of variability are identified in the Ti/Al, Fe/Al, and the Terrigenous index (Terrigenous index = Al/ Al +Fe + Mn) vertical trends (Fig. 5), indicating that sediments sources since pre-European settlement have remained similar. Minor variability observed of conservative elements, as well as in the REE, are attributed to local changes in the relative abundance of primary rock forming minerals (such as feldspars, plagioclase or mafic minerals) caused by minor changes in sediments supply, associated to storms and flooding events. For example in core G37 the marked peak at 40 cm of Fe/Al, and Ti/Al values is related to a flood occurred in 1931 followed by an erosion surface over which younger sediment have been deposited.

3.2. Anthropogenic vs. natural sediment sources

3.2.2.1. Deception Bay. Total heavy metals, Al, and fine fraction content display consistent trends with depth. Cores G37 and G34 have high and relatively constant metal concentrations below 90–100 cm, associated to high percentage of mud (86–98% mud), as shown by the lower Al-normalized profiles. In the middle section of the cores, Cu, Ni, Zn, Cd, and Pb decrease similarly, and then increase towards the top. In G37 metals increase in the top 20 cm, while in G34 higher concentrations are found in the top 10 cm. In both cores the increasing trends in the top 40–50 cm of the metal/Al profiles suggest that metal enrichment is not only influenced by higher mud content. In core G34 the normalized profiles show similar peaks at 12 cm for Cu and Pb, and Zn increases toward the surface with a peak at 4 cm.

Elements such as Al, Fe, Ti, and, rare earth elements are commonly used as proxies to define sediment sources, and therefore identify changes in clastic sediment supply (Gasparon et al., 2007; de Carvalho Gomes et al., 2009). The rare earth elements (REE) have a coherent behavior during weathering, erosion, and fluvial transportation (McLennan, 1992). The REE composition and abundance in sediments depend on the original source rock mineralogy, and reflect the depositional processes responsible for the formation of the sediments. Therefore, specific patterns in the REE profiles are used here in combination with 210Pb geochronology results, to trace sediment provenance, and to trace temporal changes in sediment supply (Munksgaard et al., 2003; Kamber et al., 2005; de Carvalho Gomes et al., 2009). 3.2.1. REE and conservative elements All REE concentrations were normalized to Post-Archean Average Shale (PAAS; see Taylor and McLennan, 1985). Pre- and postEuropean PAAS-normalized REE patterns for the four cores are shown in Fig. 4 and the enrichment in different sediments intervals is shown in Table 2. In Deception Bay post-European (G37 0–50 cm; G34 0–30 cm) and pre-European sediments (G37 100–179 cm and G34 99–174 cm) have sedimentologic and geochemical similarities with similar REE patterns, positive Eu anomalies, and a slight enrichment in heavy REE-HREE (La/Lu(PAAS) ratio ranging between 0.77 and 0.81). In core G37 sediments between 40 and 44 cm have strong light REE-LREE and middle REE-MREE enrichment (La/Lu(PAAS) = 1.83; Gd/ Lu(PAAS) = 2.04) and negative Eu anomaly (Eu ~ 0.17–0.29 μg/g). In core G6 sediments deposited after 1906 (top 40 cm) have small positive Eu anomalies, and are enriched in HREE (La/Lu = 0.71), compared to pre-European sediments (below 60 cm), which are characterized by a more marked Eu positive anomaly and lower enrichment in HREE (La/Lu = 0.83). The lower concentrations of REE in the bottom sediments are consistent with the typical low REE contents found in sandy sediments (Taylor and McLennan, 1985). In contrast G31 sediments have positive Eu anomalies and a little enrichment in HREE. Sediments deposited since 1949 differ from sediments deposited in pre-European times (below 170 cm) only for a slight Eu and HREE enrichment (La/Lu(PAAS) = 0.55). The strong similarities among REE patterns found in sediments from different parts of Moreton Bay suggest a common mixed

3.2.2. Vertical distribution of trace and major elements To reduce variability in metal concentration due to differences in grain size and clay content, heavy metal concentrations were normalized against Al (Fig. 6). Aluminum was preferred to other elements because of its conservative nature in marine environment. Aluminum is a good proxy for clay mineral content, since it is associated with the alumino-silicate fraction (Loring and Rantala, 1992; Roychoudhury, 2007; Badr et al., 2009) and it has negligible anthropogenic input in the study area.

3.2.2.2. Pine River. Three distinct major changes are identified in core G6. The coarse quartzose sand (mud b 12%) below 60 cm represents a very different depositional environment from the top sediments. Metals show little variability and very low concentrations (3.9– 0.04 μg/g Ni, 2.9–6.23 μg/g Cu, 15–23 μg/g Zn, 0.028–0.13 μg/g Cd, and 5.3–6.47 μg/g Pb). Between 60 cm and 50 cm metals increase (for example Pb from ~ 6 μg/g to 11 μg/g) while the top 50 cm are characterized by higher metal concentrations (Pb ~ 20 μg/g). The down-core variability of Al and fine fraction in the top 50 cm shows that physical sedimentation conditions have not changed significantly. The same consistency is shown in the trend of REE and of conservative elements. Normalized patterns suggest that metal concentrations are not mainly controlled by grain size, which is relatively constant (mud ~ 86–98%). Between 37 and 23 cm Pb, Zn, Ni, and Cu have higher values, decrease until 9 cm and, then increase again towards the surface. 3.2.2.3. South Moreton Bay. In core G31 all metals increase above 50 cm. Lead, Cd, and Zn peak at 10 cm (37.7, 83.23 and 0.3 μg/g, respectively). Copper follows the same trend of Pb. Al-normalized trends show that the high metal concentrations in the lower part of the core are associated with a large fraction of fine sediments and probably to the high amount of OM below 100 cm (17–15%), which is able to accumulate high concentrations of metals (Lin and Chen, 1998; Marchand et al., 2006). In the upper sediments, normalized Pb, Zn, and Cd peak at the surface. Metal increase towards the surface is higher than the natural metal/Al background for core G31.

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Fig. 6. Trends of grain size, LOI, total and Al-normalized metals vs. depth for the four cores. Ages on the left represent the oldest dated sediments in each core (dated with 210Pb). The gray field is the sediment interval over the full-inferred 210Pb age range derived from the exponential decrease of 210Pbuns (≥150 years from the date of collection → ≥1850). Vertical arrows distinguish pre- and post-European sediments.

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Fig. 6. (continued)

G. Morelli et al. / Chemical Geology 300-301 (2012) 152–164 Table 3 Metal/Al background values in each sediment core calculated as the average of the two lowest sediments based on 210Pb dating. The Moreton Bay average metal/Al values are calculated from the average background of the four cores. Location

Core

Deception Bay

G37 150–180 G34 150–175 G6 125–130 G31 140–170 Average Standard deviation 95% confidence interval

Pine River South Moreton Bay Moreton Bay

cm

Pb/Al

Cd/Al

Ni/Al

Cu/Al

Zn/Al

2.1 1.9 1.8 1.5 1.8 0.2

0.018 0.011 0.015 0.015 0.015 0.0027

4.0 4.6 1.7 4.2 3.6 1.3

3.0 2.9 1.3 2.1 2.3 0.8

7.0 8.7 5.6 6.7 7.0 1.3

0.2

0.0026

1.3

0.8

1.3

3.3. Assessment of human impact 3.3.1. Definition of geochemical background The definition of pre-industrial geochemical backgrounds is the first step in the assessment of metal enrichment in surface sediments and therefore in the quantification of human impact in an ecosystem (Reimann and de Caritat, 2005; Reimann et al., 2005; Mil-Homens et al., 2006). The local pre-European metal background values are defined following 210Pb geochronology results. In each core the average concentrations of the deepest two samples are taken as representative of natural geochemical sediment composition. The choice of those samples as representative of the pre-European background is justified by the close similarity in metal/Al values in the background sediments across the bay, based on the small standard deviation among the different sites, as shown in Table 3. The Al-normalized profiles (Fig. 6) show that increasing metal concentrations in recent times are independent on grain size. Metal/Al values in recent sediments are higher than background, and the increasing trends towards the top indicate that modern sediments (post ~ 1900) contain anthropogenic metals. The complexity of this estuarine-marine embayment, however suggests that it would be inappropriate to establish a common metal background value representative of the pre-European sediments in the four cores and to extend it to the entire Moreton Bay. 3.3.2. Estimation of anthropogenic metal enrichment Estimating “Enrichment Factor” based on Al-normalized values allows plotting geochemical trends across large areas, which have different proportions of sand and mud (such as Moreton Bay), allowing comparison of metal enrichment in sediments at a spatial scale

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(Reimann et al., 2009). EF is defined as the ratio between the normalized concentrations of a metal in the sediments divided by the normalized concentration of the same metal in the background sediments (EF = (Met/Al)sample/(Met/Al)background). According to many authors, EF b 1 indicates no enrichment, EF = 1–3 is minor enrichment, EF = 3–5 is moderate to severe enrichment, EF = 10–25 is severe, EF = 25–50 very severe, and EF > 50 is extremely severe (Vreca and Dolenec, 2005; Wang et al., 2007). This classification is appropriate for heavily contaminated sites, as stated by Reimann and de Caritat (2005), and although EF is a very useful indicator of impact close to contamination sources. Indeed, EF is strongly influenced by other processes (e.g. biogeochemical processes) that can redistribute chemical elements in the environment (Reimann and de Caritat, 2000; 2005) and the significance of EF strongly depends on the chosen background values (Loring and Rantala, 1992; Reimann and de Caritat, 2005; Vreca and Dolenec, 2005). Consequently, for cores G37, G34, and G31, the average concentrations of the deepest two samples in each core are here chosen as representative of the preEuropean natural background values. In contrast, for G6 the EF is calculated using the average metal background of the 4 cores. This choice is made because of the marked change in grain size and quartz abundance between the bottom and the top core, in an attempt to minimize EF variability due to local lithological variations. Fig. 7 shows EF in the post-European period in each core. In Deception Bay, sediments deposited after European settlement are enriched 1× or 2× times over background levels, indicating little or no impact until the end of the nineteenth century. G37 sediments deposited after 1923 are enriched 1.5× over background levels for Zn, Pb, and Cd, while Ni decreases after 1966. Higher enrichment for Pb, between 37 and 60 cm is correlated to changes in mineral composition (higher proportion of Pb- rich Paleozoic shales compared with the rest of the core), as shown by the REE patterns. These values are excluded from the enrichment trend (dotted lines in Fig. 7) to emphasize that enrichment of trace metal in sediments does not necessarily reflect anthropogenic influence, but may have a diagenetic or lithological origin (Zwolsman et al., 1993; Reimann and de Caritat, 2005). Core G34 shows similar little enrichment over background levels, with increasing trends since about 1897. The maximum enrichment in Pb (2.6×) and Cu (1.6 ×) correspond to 1963. The Pine River sediments deposited since the beginning of the century are also little enriched in Zn, Pb, Cd, and Cu (EF = 1.5–2 ×). EFs in core G6 increase since the beginning of the century (1906) and peak around 1950–1960. Core G31 is also slightly enriched in Zn, Cr, and Cu (~1.5× levels), and Cd and Pb (~2×) in the top 50 cm (since 1949). Lead and Cd are 3× background values at 10 cm, corresponding to 1989.

Fig. 7. Enrichment factors (E.F.) for Pb, Zn, and Cu of post-European settlement sediments. Sediments deposited between ~ 1920 and ~ 1980 are highlighted in gray. The bold dashed black line indicates the Moreton Bay discovery (in 1824) inferred assuming a constant average sedimentation rate of 0.2 cm/y in G37; 0.18 cm/y in G34; 0.31 cm/y in G6; in core G31 a shell layer between 50 and 70 cm does not allow extending the same sedimentation rate of the sediments above. By inferring a constant sedimentation rate of 0.34 cm/y (average for the 4 cores) or of 0.69 cm/y (average for G31) sediments below 100 are older than 1850.

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3.3.3. Correlation between historical pollution events and metals record in sediments Chronological trends of metal concentrations follow the pattern of population growth and major development in the Moreton Bay catchment (Fig. 8). The trends of Pb and Zn in core G6 show this correlation, and two main periods are identified: 1) a pre-European period (pre-1824), when little or no industrial activity disrupted the natural variability of sediment supply and metal contribution to Moreton Bay; and 2) a post-European settlement period (post-1824), characterized by a phase of relatively low impact until 1930, and a phase of major development between 1930 and 1990. After 1842 Moreton Bay was opened to free settlement and population started to grow. Between 1824 and 1920/1930 increased developments in the bay caused rapid catchment modification (deforestation, cropping, and agriculture development), which impacted the sediment supply to the bay, leading to increased sedimentation rates (for example in core G6) and increased metal fluxes. After the 1930s, the Moreton Bay area became further industrialized and urbanized, and the composition and abundance of sediment run-off was influenced by increasing population, increased land clearing, road and infrastructure construction, dredging, mining, introduction of leaded gasoline, and higher coal consumption (Neil, 1998). This fast development is evident in the sedimentary record in Moreton Bay characterized by low metal concentrations in pre-European time, and increase and peak in the late 20th century. However, metal enrichment is not dramatic compared to background, as it is lower than 5 times background values. Lead levels in particular reflect the temporal trends of gasoline consumption and composition. Leaded petrol marketed by the Commonwealth Oil Refineries was being sold in Australia by August 1932 (Cook and Gale, 2005), and the relatively high fuel consumption and substantial exhaust emissions typical of vehicles of the time injected large quantities of lead into the environment. The 1970s were the period of maximum leaded gasoline usage in Australia (Cook and Gale, 2005), and the increase in Pb levels in cores G31 and G6 well corresponds to the period between the 1950s and the 1970s. In the Pine River core, which provides the longest temporal resolution, metals reach maximum concentrations around the years 1932–1952, when maximum industrial development occurred, coupled with little concern for environmental pollution. For instance,

Fig. 8. a) Mass accumulation rates (g/cm2/y) (left y-axis), E.F. for Pb and Zn in core G6 (right y-axis) since 1850, and increase in population growth in Moreton Bay catchment (upper x-axis); horizontal arrows indicate (a) relatively low impact period, and (b) major development period. Population data are taken from Queensland Treasury (2009).

in this period higher concentrations of Cu are associated to increased use of fertilizers. Similarly, the increase in Zn and Cd correlates with the high use of Pb and smelting emissions in the same period. Zn is associated with fossil fuel combustion, its large use in metallurgy and industrial materials, such as a pigment in paints, and as one of the main emissions from waste incineration. Although statistically insignificant, the slight decrease in metal levels in the surface sediments is consistent with a decrease in emissions and improvement of environmental conditions after the decline of Pb usage in most of human activities (painting, manufacturing, batteries). In 1985–1996 unleaded gasoline was introduced, and leaded gasoline was gradually phased out until it was completely banned in 2002 (NICNAS, 2003; Cook and Gale, 2005). The banning of backyard incinerators in 1987, together with the closure of two local coal-fired power stations in 1986, resulted in a decrease of atmospheric pollution levels around Brisbane. It is possible to recognize the Pb decrease in the last few years in cores G6 and G31, because they provide a long temporal resolution record. In contrast, the low sedimentation rates in Deception Bay make it difficult to correlate specific contamination events to the historical pollution trends, and to establish if the surface decrease in Pb is coincident with the decrease in leaded gasoline fuel after 1990. Similarly the decrease in Ni could be associated to a decrease in the use of fertilizers due to changes in land use in the surrounding area. Compared with other estuarine embayments in Australia, such as for example Port Jackson in New South Wales (Taylor et al., 2004), and around the world (e.g. Venice Lagoon — Bellucci et al., 2002), heavy metal concentrations in Moreton Bay intertidal sediments are low. Absolute metal loads and relative enrichment have remained low since the beginning of industrialization compared to background levels. Metal concentrations in all the cores are well below the low and upper sediment quality limits (50–220 mg/kg for Pb, 200–410 mg/kg for Zn, and 65–270 mg/kg for Cu) defined in the guidelines for contaminated sediments (ANZECC and ARMCANZ, 2000).

4. Conclusions 210 Pb chronology in association with 137Cs chronology provides the first detailed record of sediment deposition over the last ~100 years in Moreton Bay intertidal areas. Temporal trends in anthropogenic pollutants and changes in sediment supply since European settlement (1824) are preserved in four sediment cores. Accelerated sedimentation rates over the past ~100 years are in agreement with the period of major development and are linked with enhanced erosion. In Deception Bay, since ~1923, mass accumulation rates are about 0.20 g/cm 2/y. The core collected in the Pine River gives the most complete record of sediment deposition, going back to ~1906, with mass accumulation rates of about 0.54 g/cm 2/y. In South Moreton Bay mass accumulation rates are about 0.69 g/ cm 2/y. The different accretion rates found in each area are evidence for the high spatial variability of the depositional environments in the bay. Geochemical and sedimentologic analyses show relatively homogeneous characteristics for the sediments, and similar rare earth patterns and conservative element concentrations reveal that no major changes occurred in the sediment sources in the last centuries. A local natural geochemical background is calculated to represent the pre-European sediment composition of each site. A consistent increase in Pb, Zn, Cd, Cu, and Ni is observed in most of the cores compared to background levels. Metal concentrations increase about 1 to 2 orders of times during the post-1923 period with maximum anthropogenic impact recorded during the 1950s–1970s. Results indicate that changes in sedimentary patterns and metal contamination correspond with the arrival of Europeans in Moreton Bay (post c.a. 1824)

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and are closely matched with documented historic events in the catchment. This work provides the first investigation of temporal and spatial variability of sedimentary and geochemical patterns in Moreton Bay, and it is the first assessment of human induced changes to the natural sediments deposition since European arrival. Our observations suggest that the estimation of past human impact recorded in sediment cores from similar intertidal estuarine embayments is possible, but only after combining different proxies ( 210Pb and 137Cs geochronology, heavy metals, and sediment properties) with the known historic events in the area. Further investigations could focus on distinguishing sediment provenance using additional geochemical fingerprints, and a correlation with the current historic reconstruction; modeling the current heavy metals load from the catchments to quantify metals deposition from each catchment area and in each part of the bay; and a quantification of the amount of metals trapped in mangroves roots or metals concentrations at increasing distance from mangroves in the intertidal areas. Acknowledgments We would like to thank ANSTO (Australian Nuclear Science and Organization) for the 210Pb dating (AINSE — Australian Institute of Nuclear Science and Engineering — grants AINGRA08094 and AINGRA0908). Thanks to Henk Heijnis for his assistance with the interpretation of the 210Pb data. Thanks go also to Mery Malandrino and Edoardo Mentasti (Università degli Studi di Torino), Maurizio Aceto and Matteo Oddone (Università del Piemonte Orientale) for making their lab facilities available for interlaboratory data validation. Thanks to Peter Colls and Roshni Narayan for their assistance in the field. This work was carried out while G. Morelli was a PhD student at The University of Queensland, with the financial support of a University of Queensland International Postgraduate Research Scholarship (IPRS). This project was supported by the Australian National Centre for Groundwater Research and Training. References ANZECC and ARMCANZ, 2000. Australian and New Zealand guidelines for fresh and marine water quality. Volume 1. Australian and New Zealand Environment and Conservation Council and Agriculture and Resource Management Council of Australia and New Zealand: National Water Management Strategy. Paper No. 4. The Guidelines, Environment Australia. Department of Environment and Heritage, Canberra. Appleby, P.G., Oldfield, F., 1978. The calculation of 210Pb dates assuming a constant rate of supply of unsupported 210Pb. Catena 5, 1–8. Appleby, P.G., Oldfield, F., 1983. The assessment of 210Pb data from sites with varying sediment accumulation rates. Hydrobiologia 103, 29–35. Appleby, P.G., Oldfield, F., 1992. Application of lead-210 to sedimentation studies. In: Ivanovich, M., Harmon, R.S. (Eds.), Uranium-series Disequilibrium Applications to Earth, Marine and Environmental Sciences. Clarendon Press, Oxford, pp. 731–778. Appleby, P.G., Oldfield, F., Thompson, R., Huttunen, P., Tolonen, K., 1979. 210Pb dating of annually laminated lake sediments from Finland. Nature 280 (5717), 53–55. Australian Bureau of Statistics, 2009. Regional population growth 2007–2008. Australian Bureau of Statistics, Brisbanehttp://www.abs.gov.au. Badr, N.B.E., El-Fiki, A.A., Mostafa, A.R., Al-Mur, B.A., 2009. Metal pollution records in core sediments of some Red Sea coastal areas, Kingdom of Saudi Arabia. Environmental Monitoring and Assessment 155, 509–526. Bellucci, L.G., Frignani, M., Paolucci, D., Ravanelli, M., 2002. Distribution of heavy metals in sediments of the Venice Lagoon: the role of the industrial area. Science of the Total Environment 295 (1–3), 35–49. Capelin, M., Kohn, P., Hoffenberg, P., 1998. Land use, land cover and land degradation in the catchment of Moreton Bay. In: Tibbets, I.R., Hall, N.J., Dennison, W.C. (Eds.), Moreton Bay and Catchment. School of Marine Science, The University of Queensland, Brisbane, pp. 55–66. Clark, M.W., 1998. Management implications of metals transfer pathways from a refuse tip to mangrove sediments. Science of the Total Environment 222, 17–34. Connor, S.E., Thomas, I., 2003. Sediments as archives of industrialization: evidence of atmospheric pollution in coastal wetlands of southern Sydney, Australia. Water, Air, and Soil Pollution 149, 189–210. Cook, D.E., Gale, S.J., 2005. The curious case of the date of introduction of leaded fuel to Australia: implications for the history of Southern Hemisphere atmospheric lead pollution. Atmospheric Environment 39 (14), 2553–2557.

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