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Metal bioavailability and toxicity through a sediment core
Environmental Pollution 116 (2002) 159–168 www.elsevier.com/locate/envpol
Metal bioavailability and toxicity through a sediment core U. Borgmann*, W.P. Norwood National Water Research Institute, Environment Canada, Burlington, Ontario, Canada L7R 4A6 Received 20 March 2001; accepted 3 July 2001
‘‘Capsule’’: Metal bioaccumulation, relative bioavailability, and toxicity to Hyalella azteca varied with depth and age of deposition of sediment core sections. Abstract Sediment cores from Richard Lake near Sudbury, Ontario, were sectioned and analyzed for total metal content, plus metal bioavailability and toxicity to Hyalella azteca (after equilibration with oxygenated overlying water). Strong and similar sediment profiles were observed for Cd, Co, Cu and Ni in the sediment. However, these differed from metal bioavailability profiles (bioaccumulation by Hyalella and metals in overlying water). Bioavailability profiles for Cu also differed from those for Cd, Co or Ni. The deepest sediment layers, deposited prior to industrial development, were non-toxic. Sediment toxicity was attributed to Ni dissolution into overlying water. Moreover, differential bioavailability of Ni in surface and deeper sediment layers was observed. This can affect the interpretation of toxicity data for sediments collected by different methods (e.g. core vs. grab samples). Based on Pb-210 dating and trends in Ni in the core, chronic toxicity of surface sediments from Richard Lake might approach non-toxic levels in about 15 years. Crown Copyright # 2001 Published by Elsevier Science Ltd. All rights reserved. Keywords: Metals; Toxicity; Bioavailability; Sediment; Core; Profiles
1. Introduction Recent studies on lakes in the Sudbury area revealed sediment toxicity to Hyalella and other species, and identified nickel as the causative agent (Borgmann et al., 2001b). However, Ni contamination from smelting operations has been occurring for over a century (Winterhalder, 1995), and the effects seen today are not necessarily attributable to recent industrial activity. Sediment core profiles reveal low concentrations of metals in deep sediments, increasing concentrations in more recent sediments, and sometimes a decrease in the most recently deposited sediments at the surface (Nriagu and Rao, 1987; Borgmann et al., 2001b). Most sediment toxicity tests are conducted with bulk sediment samples which represent sediments deposited over many years, or even decades. This may allow spatial resolution of sediment contamination and effects, but it does not provide any temporal information. This raises several interesting questions. Is the vertical profile * Corresponding author. Tel.: +1-905-336-6280; fax: +1-905-3366430. E-mail address: [email protected] (U. Borgmann).
of total metals in sediment cores a reliable predictor of metal bioavailability and sediment toxicity? Are the deeper, uncontaminated, sediments really non-toxic? Are contaminated sediments subjected to geochemical alterations that affect their bioavailability and toxicity after many years of burial? If so, how does this affect the interpretation of sediment toxicity tests conducted with bulk sediments? When will reductions in metal emissions in the Sudbury area result in non-toxic sediments? Some of these questions might be answered by examining the toxicity of sediment core sections. Studies on contaminant bioavailability or toxicity in sediments from different depths and ages are relatively rare. Rosiu et al. (1989) measured toxicity of sediment core sections from the Detroit River to chironomids and estimated the volume of toxic sediments present. Stemmer et al. (1990) observed higher mortality of Ceriodaphnia exposed to deeper (10–21 cm) than to surface sediment or grab samples of polynuclear aromatic hydrocarbon (PAH) and metal contaminated sediments. Increased toxicity in subsurface samples was also observed in marine amphipods exposed to sediments contaminated with PAHs, metals and DDT (Swartz et al., 1991), and in chironomids exposed to
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DDT-contaminated sediments (West et al., 1994). Toxicity generally correlated with contaminant concentrations in these profiles. Harkey et al. (1995), however, observed much lower bioaccumulation of PAHs by worms (Lumbriculus) exposed to surface sediments than to deeper and older sediments contaminated with similar total PAH concentrations. All of these authors obtained information about sediment toxicity which was not obtainable from tests with grab samples alone. In order to obtain a better understanding of the biological significance of the metal core profiles in sediments from the Sudbury area, we collected several sediment cores from Richard Lake and measured total metal concentrations, toxicity to Hyalella, and metal bioaccumulation by Hyalella for individual core slices from the surface to 20 cm depth. Three separate cores were analyzed in this way to ensure that the results were not due to collection artifacts associated with any single core. A fourth core was sectioned and subjected to both total metal analysis and Pb-210 dating. These results were then compared to metal bioaccumulation and toxicity studies with grab samples. In addition, an attempt was made to predict the temporal rate of change in surface sediment toxicity and to estimate the date when surface sediments might once again be nontoxic in this lake.
2. Materials and methods Sediment cores were obtained from Richard Lake (46.4378 N, 80.9167 W) on May 18, 1999 by inserting 10-cm-diameter core tubes into a mini box core sample collected at the deepest location (10 m) in the lake. Cores were collected carefully so as to minimize disturbance of the sediment surface. These were sealed with plastic caps and stored upright at 4 C in a fridge and returned to the lab intact. Surface sediments were collected by taking additional core samples, extruding these manually and slicing off the top section (approx. 0.5 cm). Since the sediments were black and apparently anoxic close to the sediment surface, surface ‘‘spoon’’ samples were also taken by extruding sediment cores until the sediment was close to the top of the core tube, and then carefully removing only the lighter colored surface ‘‘oxidized’’ sediment with a stainless steel spoon. These ‘‘surface slice’’ and ‘‘surface spoon’’ samples were stored for 3–7 days in polypropylene bags at 4 C. In addition, three mini-ponar grab samples of sediment were collected and also stored in polypropylene bags at 4 C. Within 3–7 days of sediment collection, the cores were carefully sectioned in the laboratory into 0.5 cm intervals for the first 5 cm, 1 cm intervals for the next 5 cm, and 2 cm intervals for depths from 10 to 20 cm using a
hydraulic extruder (Mudroch and MacKnight, 1994). Subsamples (15 ml) of each sediment section were carefully (so as to minimize mixing of sediments with water) placed into Imhoff settling cones with 1 l of overlying water immediately after sectioning, allowed to settle, and aerated before addition of test animals. Imhoff settling cones were used as test chambers because these allow static tests to be conducted with small sediment and large overlying water volumes, thereby avoiding the need for water renewal (Borgmann and Norwood, 1999a). Overlying water consisted of dechlorinated Burlington city tap water (originating from Lake Ontario, hardness 130 mg L1, alkalinity 90 mg L1, DOC 2.3 mg l1). Measurements of overlying water during the tests produced a mean pH of 8.28 (range 8.02–8.46, n=143), specific conductance of 326 mS cm1 (306–351, n=143), oxygen concentration of 8.1 mg l1 (7.5–9.0, n=22), and non-detectable ammonia ( < 0.5 mg l1, n=144). Three separate experiments were run, each with sediment sections from one core. In addition, with each experiment, one test container contained sediment from the ‘‘surface slice’’ sample, another contained sediment from the ‘‘surface spoon’’ sample, and another contained sediment from one of the three ponar grab samples. Amphipods for bioaccumulation measurements were cultured in dechlorinated Burlington city tap water as described in Borgmann et al. (1989). Cultures and experimental animals were kept in an incubator at 23 C with a 16 h light:8 h dark photoperiod; the same conditions were used for the experiments. Sediments and overlying water were gently aerated and allowed to equilibrate for 7–11 days before test animals were added. Fifteen 0–1 week old Hyalella were added directly into each test chamber and exposed to the test sediments for 28 days. Tetra-Min1 fish food flakes were added as food to each container in the following amounts: 2.5 mg at the start of week 1, 2.5 mg twice in week 2, 2.5 mg three times in week 3, and 5 mg three times in week 4. Seven to nine days after addition of the young amphipods, 15 adult amphipods were placed into cages suspended above the sediments to allow determination of metal bioaccumulation from the overlying water. Cages consisted of 250-ml polypropylene specimen containers with the bottom cut out and replaced with 200-mm mesh nylon screens (Borgmann and Norwood, 1999b). A 55 cm piece of cotton gauze was added to the cage to provide a substrate for the animals. Caged amphipods were exposed for 7 days, then removed from the test containers, rinsed in clean water, and placed in 120-ml plastic specimen containers with 50 mM EDTA and cotton gauze for 24 h to clear their guts. They were then dried at 60 C for 72 h. After the 28-day incubation period for the sediment exposed animals, a sample of overlying water was filtered through a 0.4 micron polycarbonate membrane filter and preserved
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with nitric acid. The surviving amphipods were sorted from the sediment by sieving, rinsed in clean water, placed in specimen containers with EDTA and cotton gauze for 24 h to clear their guts, and dried at 60 C. Metals in overlying water and digested amphipods were analyzed by inductively coupled plasma mass spectrometry (ICP–MS) by the National Laboratory for Environmental Testing (NLET) in Burlington, Ontario. Groups of four dried amphipods (approx. 0.5–3 mg total dry mass) were digested with 70% nitric acid at room temperature for 1 week, after which 30% hydrogen peroxide was added and digestion allowed to continue for another 24 h. Microscopic inspection revealed that tissue samples were completely digested. Each sample was then made up to 5 ml with Milli-Q deionized water for analysis. Spiked water and tissue digests and reference standards were generally withing 10% of expected values, and usually much closer. All concentrations were corrected using appropriate blanks with equivalent acid, peroxide and Milli-Q de-ionized water but no sample. Sediment samples were freeze dried and ground with a mortar and pestle before analysis for total metals. A 0.5 g subsample was digested with concentrated nitric (5 ml) and hydrofluoric (3 ml) acid in Teflon beakers on a hotplate at 95 C and evaporated to dryness. Residues were dissolved in hydrogen peroxide (30%, 1 ml) and nitric acid (0.4 M, 5 ml) and gently heated for 1 h at 95– 105 C. Samples were then cooled, diluted to 50 ml with 0.4 M nitric acid, centrifuged at 5000 rpm for 30 min, and analyzed on a Jy 74 inductively coupled argon plasma optical emission system (ICAP–OES). Cadmium concentrations in the sediment digests, which were below detection limit on the ICAP–OES, were re-analyzed on a Varian SpectraAA 400 graphite furnace atomic absorption spectrophotometer with Zeeman background correction. Recovery of metals from certified reference material (NIST-2704 Buffalo River sediment) was within 10% of the certified values. In order to relate sediment metal concentrations to date of sediment deposition, a fourth core was sectioned into 1-cm slices, analyzed for total metals, and dated by Pb-210 analysis. Dating of freeze-dried sediment samples was performed by Paul Wilkinson, Freshwater Institute, Fisheries and Oceans Canada, Winnipeg. Lead-210 dates were confirmed by Cs-137 analysis (Wilkinson, 1985, and references quoted therein).
3. Results 3.1. Core profiles for metals in sediment, overlying water and caged Hyalella Strong profiles for total metal in the sediment were observed for Cd, Co, Cu and Ni (Fig. 1). Metal
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concentrations were lowest in the deepest part of the core, increased rapidly up to approximately 5 cm depth, and then decreased slightly towards the surface. Profiles for the three core sections used for bioavailability and toxicity testing were quite similar, except that one of the profiles (--, Fig. 1) showed maximum concentrations for all metals at a slightly shallower depth than the other two profiles. The surface slice and surface spoon samples had metal concentrations similar to those near the top of the core. The mini-ponar samples had metal concentrations in the mid range of the concentrations of the top 10 cm of the cores (Fig. 1), as expected since this sampler collects sediments down to about 10 cm depth. Profiles for metals in overlying water, and for metals in Hyalella exposed to overlying water in the cages, differed somewhat from the profiles of total metal in the sediment (Figs. 2 and 3). The profiles of Cd, Co and Ni in Hyalella matched the profiles for these metals in overlying water, but these were more pronounced than the profiles of these metals in the sediment. Metals in Hyalella correlated better with metals in water (r2=0.64, 0.71, 0.80 for Cd, Co and Ni, respectively) than with metals in sediment (r2=0.55, 0.47, 0.55). Metals in water and in Hyalella increased rapidly with increasing depth in the top 5 cm, whereas concentrations of these metals in the sediment increased only moderately with depth over this range (Fig. 1). Concentrations of Cd, Co and Ni in water and Hyalella from test containers with surface sediments, including sections from the top of the core as well as surface slice and surface spoon samples, were similar to those from containers with core sections near 12 cm (Figs. 2 and 3). However, total metal concentrations in the sediment at 12 cm were only about half as high as those at the surface (Fig. 1). Concentrations of Cd and Ni in overlying water and Hyalella from test containers with ponar grab samples were in the mid range of those which had core sections between 0 and 10 cm. However, Co concentrations in water and Hyalella from containers with ponar samples were similar to those with surface and 10 cm core sections, which were substantially lower than samples from intermediate depths. Profiles for Cu in overlying water and in Hyalella differed substantially from the other profiles. Copper in overlying water decreased with increasing depth of the core section in approximately a linear fashion (Fig. 2). Although the data are more variable than those for total metal in sediment, there was no evidence of a maximum near 5 cm with lower concentrations at the surface, as seen for Cu in sediment (Fig. 1). Copper in Hyalella, on the other hand, remained approximately constant, with only a slight trend towards lower concentrations with depth (Fig. 3). This is consistent with the ability of Hyalella to regulate Cu near essential background levels, except at very high environmental concentrations (Borgmann and Norwood, 1995a, b, 1997).
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Fig. 1. Concentrations of Cd, Co, Cu and Ni in three separate sediment cores from Richard Lake (, +, ), in surface slice and surface spoon samples (*) and in three different mini-ponar grab samples (X). Mini-ponar grab samples are plotted at a depth of 5 cm, but the sampler actually collects sediments between 0 and approximately 10 cm depth.
Other metals measured in sediment, overlying water and Hyalella showed less pronounced profiles, or were much lower in concentration, than Cd, Co, Cu or Ni (Table 1). Arsenic and Cr varied little with depth. Manganese was quite variable and sometimes high at the surface, but otherwise showed no consistent trend. Lead showed a strong profile in the sediment, but was moderately low in concentration and much less bioavailable than Cd. Thallium was also low. The profile for Zn was similar in shape to Cu and Ni, but less pronounced, and total Zn in the sediment was much lower than Cu or Ni (Table 1). Zinc concentrations in Hyalella did not vary with depth, and were near the background concentration of this essential metal in Hyalella (Borgmann and Norwood, 1995b). 3.2. Toxicity and Ni in sediment, water and Hyalella Chronic (4-week) toxicity of the sediment core sections to Hyalella produced a profile that matched metal
concentrations. Mortality was almost complete for amphipods exposed to sediments from 3 to 14 cm depths. Survival was high (580%), however, in the deepest sediments (Fig. 4). There was partial survival at the shallowest depths in sections from one of the cores and in two of the tests with the surface spoon sample, coinciding with the slightly lower metal concentrations at the surface. In contrast, survival of adult Hyalella caged in the experimental containers for only one week was 580% regardless of sediment depth. Chronic mortality of young amphipods exposed to sediments correlated reasonably well with Ni bioaccumulation in caged adult Hyalella exposed for one-week in the same test containers (Fig. 4). Survival was compared to bioaccumulation in caged Hyalella because complete mortality in many of the test containers precluded measurement of metal accumulation in sedimentexposed animals. Previous chronic toxicity tests with ponar grab samples from six sites in four lakes in the Sudbury area, including Richard Lake, demonstrated
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Fig. 2. Concentrations of Cd, Co, Cu and Ni in overlying water during toxicity and bioaccumulation tests with sediments from three separate cores from Richard Lake. Symbols as in Fig. 1. Breaks in the lines are due to missing intermediate data values.
that chronic toxicity in caged amphipods was equal to that in sediment exposed animals. This implies that toxicity was caused by a dissolved substance and not the solid-phase sediment (Borgmann et al., 2001b), and suggests that comparison of survival and bioaccumulation in sediment-exposed and caged animals is reasonable. As in the previous study (Borgmann et al., 2001b), Ni was the only metal accumulated in excess of chronic lethal body concentrations. The curve drawn through the survival-bioaccumulation data in Fig. 4 was obtained from a previous study with Ni-spiked sediments conducted in Imhoff settling cones (Borgmann et al., 2001a) and is described by m ¼ 0:0196 þ 44:3 109 Ni2:758 TB
ð1Þ
where m is the weekly instantaneous mortality rate (ln(survival)/t) and NiTB is the total body concentration of Ni in Hyalella. This was obtained by combining the non-linear regression of m against Ni in water (Niw)
m ¼ 0:0196 þ 349 109 Ni2:032 w with bioaccumulation from water (NiTB=2.114 Ni0.737 ) for data in Borgmann et al. (2001a). The fit w between the observed survival and Ni in caged Hyalella with the independently derived survival-body concentration relationship from Ni-spiked sediments (Fig. 4) further supports the hypothesis that toxicity is primarily due to Ni bioaccumulated from dissolved Ni leached from the sediments. Since Ni is the primary cause of sediment toxicity, a closer examination of Ni bioaccumulation from sediment is warranted. Nickel bioaccumulation from sediment appeared to follow a bi-phasic trend (Fig. 5). For data obtained from core sections below 5 cm, Ni bioaccumulated by Hyalella was approximately proportional to Ni in the sediment. Log-log regression of the data gave a slope of 1.18 (95% CI=0.95–1.42). For data obtained from core sections above 5 cm, however, the slope was 3.36 (95% CI=2.58–4.13). The relative Ni
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Fig. 3. Concentrations of Cd, Co, Cu and Ni in Hyalella caged above the sediments for one week during toxicity tests with sediments from three separate cores from Richard Lake. Symbols as in Fig. 1. Breaks in the lines are due to missing intermediate data values.
bioavailability, therefore, decreases rapidly towards the surface. The bi-phasic nature of the bioaccumulation relationship disappears, however, if bioaccumulation is expressed relative to Ni in overlying water (Fig. 5). Bioaccumulation relative to Ni in water was similar to that observed previously with Ni-spiked sediments (Borgmann et al., 2001a). Furthermore, Ni bioaccumulated in Hyalella exposed directly to the sediment, for those treatments with surviving animals, was similar to Ni bioaccumulation by caged Hyalella (Fig. 5). The decreased relative bioavailability of Ni in the top few cm of the sediment is, therefore, due to a reduced relative leaching of Ni into the overlying water. 3.3. Core dating Lead-210 dating of a fourth core sectioned into 1-cm intervals down to 20 cm revealed a constant sedimentation rate on a dry weight basis of 241 g m2 year1. Metal profiles in this core were similar to the
three cores used for bioaccumulation and toxicity testing. In order to adjust for slight differences in sedimentation rates of the different cores, a common reference point was needed. This was obtained by calculating the depth at which the concentration of Cu and Ni, the two most abundant metals, was equal to one-half of the maximum of the three point running mean concentration. For the Pb-210 dated core, this was equal to 8.91 and 10.29 cm for Cu and Ni, respectively. For the three toxicity test cores this ranged from 7.32 to 9.39 and 8.56 to 10.76 cm, respectively. Dividing these depths by the depths obtained for the Pb-210 core and averaging the data for Cu and Ni provided relative depths of the contaminated layer of 1.000, 0.8268, 1.0334 and 1.0498 for the Pb-210 dated and three toxicity cores. Assuming that the depth of the contaminated layer represented an equal deposition time period for each of the cores, these relative depths were used as correction factors to age the three toxicity cores. This allowed plotting of metal concentrations in
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Table 1 Mean metal concentrations (S.D.) measured in sediment core slices and in overlying water and caged Hyalella exposed to those slices for three depth ranges (n=6) Depth range (cm)
As
Cd
Co
Cr
Cu
Mn
Ni
Pb
Tl
Zn
0.036 (0.005) 0.048 (0.004) 0.014 (0.005)
1.25 (0.19) 1.78 (0.19) 0.39 (0.08)
1.37 (0.05) 1.41 (0.09) 1.28 (0.08)
23.3 (0.7) 26.3 (0.8) 2.6 (0.7)
11.9 (5.4) 6.9 (1.4) 6.8 (0.7)
32.1 (5.4) 47.3 (4.1) 10.0 (3.0)
0.41 (0.07) 0.39 (0.08) <0.0125
nm nm nm
4.11 (0.56) 4.69 (0.97) 1.79 (0.20)
Overlying water (nmol l1) 0–1 18 (6) 4–5 34 (8) 16–20 28 (4)
0.60 (0.51) 1.31 (0.20) 0.19 (0.09)
5.5 (8.2) 29.8 (6.2) 0.6 (0.3)
7.4 (0.6) 7.7 (1.9) 6.8 (1.0)
169 (20) 153 (23) 44 (21)
3.9 (3.4) 8.4 (5.0) 1.3 (0.7)
762 (610) 3325 (836) 408 (191)
<0.64 <0.64 <0.64
0.06 (0.03) 0.15 (0.02) 0.08 (0.01)
<150 <150 <150
Hyalella-caged (nmol g1) 0–1 20 (3) 4–5 20 (5) 16–20 23 (4)
13 (5) 19 (2) 6 (3)
8 (6) 24 (2) 3 (1)
25 (3) 27 (6) 25 (4)
1154 (140) 1270 (194) 1052 (189)
83 (20) 81 (9) 95 (28)
346 (199) 1169 (278) 165 (73)
1
Sediment (mol g ) 0–1 nma 4–5 nm 16–20 nm
a
0.6 (0.4) 0.5 (0.2) 0.2 (0.1)
0.4 (0.1) 0.9 (0.1) 0.5 (0.1)
872 (43) 845 (24) 871 (22)
nm, Not measured.
Fig. 4. Survival of young Hyalella during 4-week chronic exposures to sediments from three separate cores from Richard Lake expressed as a function of sediment depth or Ni concentrations in caged Hyalella. Symbols as in Fig. 1 for survival versus depth. Data obtained from < 5 cm (*) and from 55 cm (&) deep sediments are shown in different symbols for survival versus Ni in Hyalella.
each of the cores as a function of predicted date of sediment deposition. This suggests that Ni concentrations in the sediment started increasing around 1890, peaked around 1970, and then declined continuously until the present (Fig. 6).
4. Discussion 4.1. Relative metal bioavailability and sediment depth The relative bioavailability of Cd, Co and Ni differed substantially from that of Cu through the sediment core. Bioaccumulation of Co and Ni followed the trend for total metal in sediment below 5 cm, but dropped rapidly to very low levels for amphipods exposed to surface sediments (Fig. 3). This rapid drop with
decreasing depth in the top 5 cm was not observed for Co and Ni in the sediment (Fig. 1), but did occur in the overlying water (Fig. 2). Concentration trends for Cd in Hyalella, sediment and overlying water generally followed those of Co and Ni, but the data are more variable and the trends less clear, probably because of the much lower concentration of this metal. In contrast, Cu bioaccumulation by Hyalella was almost constant with exposure to sediments from varying depth. In part, this is due to regulation of body Cu by this amphipod (Borgmann and Norwood, 1995a, b, 1997). However, Cu bioavailability to Hyalella in sediments appears to be due to dissolved Cu in the overlying water (Deaver and Rodgers, 1996), and Cu in overlying water followed a very different trend from the other metals, with no noticeable decrease in test chambers with surface sediments (Fig. 2).
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Fig. 5. Nickel concentrations in Hyalella caged above the sediments for one week (*, &) or exposed directly to the sediment for 4 weeks (*, &), as a function of Ni in the sediment or Ni in the overlying water. Data obtained from <5 cm (*, *) and from 55 cm (&, &) deep sediments are shown in different symbols. Lines were obtained from linear regression using the caged animal data. Separate regressions were conducted for the two depth ranges for the sediment concentration data. The dotted line was obtained from previous experiments with Hyalella exposed to Ni-spiked sediments (Borgmann et al., 2001a).
Fig. 6. Nickel in the sediment as a function of date of sediment deposition for the three toxicity/bioaccumulation test cores and the Pb-210 dated core (~). Data after 1970 are shown in expanded time scale in the inset.
Part of the reason for the differences in metal released into overlying water, and the resulting differences in bioavailability, may lie with the chemical nature of the metals. Although Cd, Co, Cu and Ni can all be classified as borderline metals, the relative amount of class A and class B character varies, with class B character increasing in the order Ni < Co < Cd < Cu (Nieboer and Richardson, 1980). Class A metals prefer to bind to oxygen atoms in ligands, followed by nitrogen and then sulfur. The reverse is true for class B metals. With increasing depth through the sediment core, oxygen
gradually becomes depleted, and it might be expected that the greatest availability of oxygen binding sites in both organic and inorganic matter would be found near the surface. Hence surface sediments might be expected to bind Ni more strongly than deeper sediments, whereas depth should have less of an influence on Cu binding. The toxicity of sediment core sections buried below the surface will be controlled by both the concentration of toxic materials incorporated at the time of sediment deposition, and geochemical processes occurring during prolonged in situ incubation. It is, therefore, possible that the relative bioavailability of Ni in the deeper core sections was initially lower than measured here, and similar to the bioavailability of Ni in the surface sediments now. Nevertheless, the deepest sections of the core were non-toxic (Fig. 4). Since Ni bioaccumulation factors (Ni in Hyalella divided by Ni in sediment) increased, rather than decreased, with depth (i.e. the slope of the line in Fig. 5 for < 5 cm depths is greater than 1), this implies that the surface sediments in Richard Lake were probably non-toxic prior to industrial development in the Sudbury area. Toxicity measurements of sediment core sections have provided information on potential changes in sediment toxicity through time, or the total volume of toxic sediments present, in several studies (Rosiu et al., 1989; Stemmer et al., 1990; Swartz et al., 1991; West et al., 1994). In addition, core sections have been used to infer long term changes in contaminant (PAH) bioavailability after burial (Harkey et al., 1995). However, most core sectioning for toxicity or bioavailability has been rather coarse (4–5 cm slice thickness). In the present study slice thicknesses varied from 0.5 cm near the surface to 2 cm deeper in the core. Furthermore, toxicity tests were
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combined with simultaneous estimates of metal bioavailability, allowing direct comparison of sediment contamination and toxic effects. Most of the changes in the relative bioavailability of Ni, the toxic agent in Richard Lake sediments, were observed within the top 5 cm of sediment. Much information would have been lost if sectioning were done only at 5 cm intervals. Core sections for toxicity testing or bioaccumulation measurements in the past have usually been relatively thick in order to provide enough sediment to conduct the tests. However, sediment toxicity and bioaccumulation studies can be conducted in Imhoff settling cones with only 15 ml of sediment (Borgmann and Norwood, 1999a). This makes more detailed analysis of core section toxicity and bioaccumulation possible. 4.2. Implications for sediment collection methodology The variation in relative Ni bioavailability with depth indicates that the results from bioaccumulation and toxicity tests could be affected by the method of sediment collection. To estimate the potential effect this could have on metal impact assessments, toxic sediment concentrations (LC25s) were estimated from Ni bioaccumulation from both the sediment core slices in Richard Lake (Fig. 5) and from previously collected multiple ponar grab samples from 21 sites in 12 lakes in the area between Sudbury and North Bay (Borgmann et al., 2001b). The total range in Ni in ponar grab samples from these lakes, and the variability in Ni bioavailability between these samples, were much greater than the variability observed through the core sections from Richard Lake (Fig. 7). The body concentration which
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caused 25% mortality in chronic tests with Hyalella (the LBC25) was 194 nmol g1 Ni (Borgmann et al., 2001a). Using the regression line for the ponar data in Fig. 7, this is equivalent to a sediment concentration of 19.5 mmol g1 Ni (the ponar-based sediment LC25). On the other hand, the LC25 can also be estimated using only surface sediments from the Richard Lake core. Nickel bioaccumulation in Hyalella exposed to the top centimeter of sediment was 346 nmol g1 (Table 1), which was 1.78-fold greater than the LBC25. If Ni bioavailability is proportional to Ni in the sediment at constant depth in the core, then the LC25 for Ni in the top 1 cm of sediment should be 32.1 mmol g1 (Table 1) divided by 1.78, which equals 18.0 mmol g1 (the surface-sediment LC25). Coincidentally, the more careful core analysis conducted in this study gave an almost equivalent estimate of the LC25 for Sudbury area sediments as the previously collected ponar grab data (18.0 vs. 19.5 mmol g1). Although the relative Ni bioavailability in ponar grab and deeper core samples was higher than Ni bioavailability in surface sediments from Richard Lake, the ponar-based regression line lies below the ponar data from this lake, and closer to the surface sediment samples (Fig. 7). The method of sediment collection does not, therefore, greatly affect the estimation of the LC25 for Ni in sediment in these two Sudbury area studies, but this seems to be more a matter of luck than good experimental design. It resulted from the fact that the regression line in Fig. 7 falls below the mean for the ponar data from Richard Lake. In general, metal concentrations in ponar grab samples collected from Sudbury area lakes are much more variable than surface (0–3 or 0–5 cm) samples. It is, therefore, probably a better practice to use more carefully sectioned surface sediment samples for toxicity testing, rather than sediments collected by ponar or similar samplers, whenever possible. 4.3. Predicted time to recovery from sediment toxicity
Fig. 7. Nickel in Hyalella as a function of Ni in sediment for core slices () or ponar grab samples (X) collected in 1999 from Richard Lake, or from ponar grab samples collected from 12 different lakes in 1998 (*). The line was obtained from linear regression of the 1998 ponar data.
The sediment core Ni chronology in Richard Lake (Fig. 6) is similar to the time trend in Ni production from Sudbury smelters. Smelting started in 1888 and peaked in the 1970s (Winterhalder, 1995). Completion of Inco’s ‘‘Superstack’’ and major emission controls in 1972 (Conroy and Kramer, 1995) probably contributed to the decline in surface metal concentrations in recent years. Interpretation of the details in the sediment core profiles must, however, be done with caution. For example, Ni concentrations decreased more rapidly than did Cu concentrations in the most recent years in Richard Lake profiles (Fig. 1). However, Cu concentrations decreased more rapidly than Ni concentrations in recent years in core profiles from McFarlane Lake, the next lake downstream from Richard Lake and only a few kilometers away (data not shown). Individual lake
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morphometry or geochemistry might be having some influence on core profiles for individual metals. Using the above estimate of Ni toxicity in surface sediments, it is possible to estimate roughly the potential time required for the surface sediment to become non-toxic. Linear regression of Ni concentrations against the estimated date of sediment deposition for data after 1980 (Fig. 6) gave a slope of 0.82 mmol g1 year1 (r2=0.66, n=28). A concentration of 18 mmol g1 (the 4-week LC25) would be reached in 2016, if this trend continues. It is, however, possible that the rate of decrease is increasing, as suggested by data from three of the four cores after 1992 (Fig. 6). Alternatively, the rate of decrease might slow, with Ni concentrations exponentially approaching pre-1890 concentrations of around 10 mmol g1. In either case, Ni concentrations are currently decreasing, and a more precise estimate of when Ni concentrations at the surface will reach the LC25 in Richard Lake should be obtainable in about another decade. Acknowledgements The field work was done by M. Mawhinney and M. Benner of the Technical Operations Section, NWRI, with assistance from the authors. Total metal concentrations in sediment were measured by J. Rajkumar. ICP-MS analysis for metals in water and Hyalella were done by the National Laboratory for Environmental Testing. Lead-210 dating was done by Paul Wilkinson, Freshwater Institute, Fisheries and Oceans Canada, Winnipeg. References Borgmann, U., Norwood, W.P., 1995a. Kinetics of excess (above background) copper and zinc in Hyalella azteca and their relationship to chronic toxicity. Can. J. Fish. Aquat. Sci. 52, 864–874. Borgmann, U., Norwood, W.P., 1995b. EDTA toxicity and background concentrations of copper and zinc in Hyalella azteca. Can. J. Fish. Aquat. Sci. 52, 875–881. Borgmann, U., Norwood, W.P., 1997. Toxicity and accumulation of zinc and copper in Hyalella azteca exposed to metal-spiked sediments. Can. J. Fish. Aquat. Sci. 54, 1046–1054. Borgmann, U., Norwood, W.P., 1999a. Sediment toxicity testing using large water-sediment ratios: an alternative to water renewal. Environmental Pollution 106, 333–339.
Borgmann, U., Norwood, W.P., 1999b. Assessing the toxicity of lead in sediments to Hyalella azteca: the significance of bioaccumulation and dissolved metal. Can. J. Fish. Aquat. Sci. 56, 1494–1503. Borgmann, U., Ralph, K.M., Norwood, W.P., 1989. Toxicity test procedures for Hyalella azteca, and chronic toxicity of cadmium and pentachlorophenol to H. azteca, Gammarus fasciatus, and Daphnia magna. Arch. Environ. Contam. Toxicol. 18, 756–764. Borgmann, U., Ne´ron, R., Norwood, W.P., 2001a. Quantification of bioavailable nickel in sediments and toxic thresholds to Hyalella azteca. Environmental Pollution 111, 189–198. Borgmann, U., Norwood, W.P., Reynoldson, T.B., Rosa, F., 2001b. Identifying cause in sediment assessments: bioavailability and the Sediment Quality Triad. Can. J. Fish. Aquat. Sci. 58, 950–960. Conroy, N., Kramer, J.R., 1995. History of geology, mineral exploration, and environmental damage. In: Gunn, J.M. (Ed.), Restoration and Recovery of an Industrial Region. Springer-Verlag, New York, pp. 3–4. Deaver, E., Rodgers Jr., J.H., 1996. Measuring bioavailable copper using anodic stripping voltammetry. Environ. Toxicol. Chem. 15, 1925–1996. Harkey, G.A., Van Hoof, P.L., Landrum, P.F., 1995. Bioavailability of polycyclic aromatic hydrocarbons from a historically contaminated sediment core. Environ. Toxicol. Chem. 14, 1551–1560. Mudroch, A., MacKnight, S.D., 1994. CRC Handbook of Techniques for Aquatic Sediments Sampling. CRC Press, Boca Ratan. Nieboer, E., Richardson, H.S., 1980. The replacement of the nondescript term ‘heavy metals’ by a biologically and chemically significant classification of metal ions. Environmental Pollution (Series B) 1, 3–26. Nriagu, J.O., Rao, S.S., 1987. Responses of lake sediments to changes in trace metal emissions from the smelters at Sudbury, Ontario. Environmental Pollution 44, 211–218. Rosiu, C.J., Giesy, J.P., Kreis Jr., R.G., 1989. Toxicity of vertical sediments in the Trenton Channel, Detroit River, Michigan, to Chironomus tentans, (Insecta: Chironomidae). J. Great Lakes Res. 15, 570–580. Stemmer, B.L., Burton, G.A., Sasson-Brickson, G., 1990. Effect of sediment spatial variance and collection method on cladoceran toxicity and indigenous microbial activity determinations. Environ. Toxicol. Chem. 9, 1035–1044. Swartz, R.C., Schults, D.W., Lamberson, J.O., Ozretich, R.J., Stull, J.K., 1991. Vertical profiles of toxicity, organic carbon, and chemical contaminants in sediment cores from the Palos Verdes Shelf and Santa Monica Bay, California. Mar. Environ. Res. 31, 215–225. West, C.W., Phipps, G.L., Hoke, R.A., Goldenstein, T.A., VanderMeiden, F.M., Kosian, P.A., Ankley, G.T., 1994. Sediment core versus grab samples: evaluation of contamination and toxicity at a DDT-contaminated site. Ecotoxicol. Environ. Saf. 28, 208–220. Wilkinson, P., 1985. The determination of environmental levels of uranium and thorium series isotopes and Cs-137 in aquatic and terrestrial samples. Can. Spec. Publ. Fish. Aquat. Sci. 78, 51. Winterhalder, K., 1995. Early history of human activities in the Sudbury area and ecological damage to the landscape. In: Gunn, J.M. (Ed.), Restoration and Recovery of an Industrial Region. SpringerVerlag, New York, pp. 17–31.
Metal bioavailability and toxicity through a sediment core U. Borgmann*, W.P. Norwood National Water Research Institute, Environment Canada, Burlington, Ontario, Canada L7R 4A6 Received 20 March 2001; accepted 3 July 2001
‘‘Capsule’’: Metal bioaccumulation, relative bioavailability, and toxicity to Hyalella azteca varied with depth and age of deposition of sediment core sections. Abstract Sediment cores from Richard Lake near Sudbury, Ontario, were sectioned and analyzed for total metal content, plus metal bioavailability and toxicity to Hyalella azteca (after equilibration with oxygenated overlying water). Strong and similar sediment profiles were observed for Cd, Co, Cu and Ni in the sediment. However, these differed from metal bioavailability profiles (bioaccumulation by Hyalella and metals in overlying water). Bioavailability profiles for Cu also differed from those for Cd, Co or Ni. The deepest sediment layers, deposited prior to industrial development, were non-toxic. Sediment toxicity was attributed to Ni dissolution into overlying water. Moreover, differential bioavailability of Ni in surface and deeper sediment layers was observed. This can affect the interpretation of toxicity data for sediments collected by different methods (e.g. core vs. grab samples). Based on Pb-210 dating and trends in Ni in the core, chronic toxicity of surface sediments from Richard Lake might approach non-toxic levels in about 15 years. Crown Copyright # 2001 Published by Elsevier Science Ltd. All rights reserved. Keywords: Metals; Toxicity; Bioavailability; Sediment; Core; Profiles
1. Introduction Recent studies on lakes in the Sudbury area revealed sediment toxicity to Hyalella and other species, and identified nickel as the causative agent (Borgmann et al., 2001b). However, Ni contamination from smelting operations has been occurring for over a century (Winterhalder, 1995), and the effects seen today are not necessarily attributable to recent industrial activity. Sediment core profiles reveal low concentrations of metals in deep sediments, increasing concentrations in more recent sediments, and sometimes a decrease in the most recently deposited sediments at the surface (Nriagu and Rao, 1987; Borgmann et al., 2001b). Most sediment toxicity tests are conducted with bulk sediment samples which represent sediments deposited over many years, or even decades. This may allow spatial resolution of sediment contamination and effects, but it does not provide any temporal information. This raises several interesting questions. Is the vertical profile * Corresponding author. Tel.: +1-905-336-6280; fax: +1-905-3366430. E-mail address: [email protected] (U. Borgmann).
of total metals in sediment cores a reliable predictor of metal bioavailability and sediment toxicity? Are the deeper, uncontaminated, sediments really non-toxic? Are contaminated sediments subjected to geochemical alterations that affect their bioavailability and toxicity after many years of burial? If so, how does this affect the interpretation of sediment toxicity tests conducted with bulk sediments? When will reductions in metal emissions in the Sudbury area result in non-toxic sediments? Some of these questions might be answered by examining the toxicity of sediment core sections. Studies on contaminant bioavailability or toxicity in sediments from different depths and ages are relatively rare. Rosiu et al. (1989) measured toxicity of sediment core sections from the Detroit River to chironomids and estimated the volume of toxic sediments present. Stemmer et al. (1990) observed higher mortality of Ceriodaphnia exposed to deeper (10–21 cm) than to surface sediment or grab samples of polynuclear aromatic hydrocarbon (PAH) and metal contaminated sediments. Increased toxicity in subsurface samples was also observed in marine amphipods exposed to sediments contaminated with PAHs, metals and DDT (Swartz et al., 1991), and in chironomids exposed to
0269-7491/01/$ - see front matter Crown Copyright # 2001 Published by Elsevier Science Ltd. All rights reserved. PII: S0269-7491(01)00190-7
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DDT-contaminated sediments (West et al., 1994). Toxicity generally correlated with contaminant concentrations in these profiles. Harkey et al. (1995), however, observed much lower bioaccumulation of PAHs by worms (Lumbriculus) exposed to surface sediments than to deeper and older sediments contaminated with similar total PAH concentrations. All of these authors obtained information about sediment toxicity which was not obtainable from tests with grab samples alone. In order to obtain a better understanding of the biological significance of the metal core profiles in sediments from the Sudbury area, we collected several sediment cores from Richard Lake and measured total metal concentrations, toxicity to Hyalella, and metal bioaccumulation by Hyalella for individual core slices from the surface to 20 cm depth. Three separate cores were analyzed in this way to ensure that the results were not due to collection artifacts associated with any single core. A fourth core was sectioned and subjected to both total metal analysis and Pb-210 dating. These results were then compared to metal bioaccumulation and toxicity studies with grab samples. In addition, an attempt was made to predict the temporal rate of change in surface sediment toxicity and to estimate the date when surface sediments might once again be nontoxic in this lake.
2. Materials and methods Sediment cores were obtained from Richard Lake (46.4378 N, 80.9167 W) on May 18, 1999 by inserting 10-cm-diameter core tubes into a mini box core sample collected at the deepest location (10 m) in the lake. Cores were collected carefully so as to minimize disturbance of the sediment surface. These were sealed with plastic caps and stored upright at 4 C in a fridge and returned to the lab intact. Surface sediments were collected by taking additional core samples, extruding these manually and slicing off the top section (approx. 0.5 cm). Since the sediments were black and apparently anoxic close to the sediment surface, surface ‘‘spoon’’ samples were also taken by extruding sediment cores until the sediment was close to the top of the core tube, and then carefully removing only the lighter colored surface ‘‘oxidized’’ sediment with a stainless steel spoon. These ‘‘surface slice’’ and ‘‘surface spoon’’ samples were stored for 3–7 days in polypropylene bags at 4 C. In addition, three mini-ponar grab samples of sediment were collected and also stored in polypropylene bags at 4 C. Within 3–7 days of sediment collection, the cores were carefully sectioned in the laboratory into 0.5 cm intervals for the first 5 cm, 1 cm intervals for the next 5 cm, and 2 cm intervals for depths from 10 to 20 cm using a
hydraulic extruder (Mudroch and MacKnight, 1994). Subsamples (15 ml) of each sediment section were carefully (so as to minimize mixing of sediments with water) placed into Imhoff settling cones with 1 l of overlying water immediately after sectioning, allowed to settle, and aerated before addition of test animals. Imhoff settling cones were used as test chambers because these allow static tests to be conducted with small sediment and large overlying water volumes, thereby avoiding the need for water renewal (Borgmann and Norwood, 1999a). Overlying water consisted of dechlorinated Burlington city tap water (originating from Lake Ontario, hardness 130 mg L1, alkalinity 90 mg L1, DOC 2.3 mg l1). Measurements of overlying water during the tests produced a mean pH of 8.28 (range 8.02–8.46, n=143), specific conductance of 326 mS cm1 (306–351, n=143), oxygen concentration of 8.1 mg l1 (7.5–9.0, n=22), and non-detectable ammonia ( < 0.5 mg l1, n=144). Three separate experiments were run, each with sediment sections from one core. In addition, with each experiment, one test container contained sediment from the ‘‘surface slice’’ sample, another contained sediment from the ‘‘surface spoon’’ sample, and another contained sediment from one of the three ponar grab samples. Amphipods for bioaccumulation measurements were cultured in dechlorinated Burlington city tap water as described in Borgmann et al. (1989). Cultures and experimental animals were kept in an incubator at 23 C with a 16 h light:8 h dark photoperiod; the same conditions were used for the experiments. Sediments and overlying water were gently aerated and allowed to equilibrate for 7–11 days before test animals were added. Fifteen 0–1 week old Hyalella were added directly into each test chamber and exposed to the test sediments for 28 days. Tetra-Min1 fish food flakes were added as food to each container in the following amounts: 2.5 mg at the start of week 1, 2.5 mg twice in week 2, 2.5 mg three times in week 3, and 5 mg three times in week 4. Seven to nine days after addition of the young amphipods, 15 adult amphipods were placed into cages suspended above the sediments to allow determination of metal bioaccumulation from the overlying water. Cages consisted of 250-ml polypropylene specimen containers with the bottom cut out and replaced with 200-mm mesh nylon screens (Borgmann and Norwood, 1999b). A 55 cm piece of cotton gauze was added to the cage to provide a substrate for the animals. Caged amphipods were exposed for 7 days, then removed from the test containers, rinsed in clean water, and placed in 120-ml plastic specimen containers with 50 mM EDTA and cotton gauze for 24 h to clear their guts. They were then dried at 60 C for 72 h. After the 28-day incubation period for the sediment exposed animals, a sample of overlying water was filtered through a 0.4 micron polycarbonate membrane filter and preserved
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with nitric acid. The surviving amphipods were sorted from the sediment by sieving, rinsed in clean water, placed in specimen containers with EDTA and cotton gauze for 24 h to clear their guts, and dried at 60 C. Metals in overlying water and digested amphipods were analyzed by inductively coupled plasma mass spectrometry (ICP–MS) by the National Laboratory for Environmental Testing (NLET) in Burlington, Ontario. Groups of four dried amphipods (approx. 0.5–3 mg total dry mass) were digested with 70% nitric acid at room temperature for 1 week, after which 30% hydrogen peroxide was added and digestion allowed to continue for another 24 h. Microscopic inspection revealed that tissue samples were completely digested. Each sample was then made up to 5 ml with Milli-Q deionized water for analysis. Spiked water and tissue digests and reference standards were generally withing 10% of expected values, and usually much closer. All concentrations were corrected using appropriate blanks with equivalent acid, peroxide and Milli-Q de-ionized water but no sample. Sediment samples were freeze dried and ground with a mortar and pestle before analysis for total metals. A 0.5 g subsample was digested with concentrated nitric (5 ml) and hydrofluoric (3 ml) acid in Teflon beakers on a hotplate at 95 C and evaporated to dryness. Residues were dissolved in hydrogen peroxide (30%, 1 ml) and nitric acid (0.4 M, 5 ml) and gently heated for 1 h at 95– 105 C. Samples were then cooled, diluted to 50 ml with 0.4 M nitric acid, centrifuged at 5000 rpm for 30 min, and analyzed on a Jy 74 inductively coupled argon plasma optical emission system (ICAP–OES). Cadmium concentrations in the sediment digests, which were below detection limit on the ICAP–OES, were re-analyzed on a Varian SpectraAA 400 graphite furnace atomic absorption spectrophotometer with Zeeman background correction. Recovery of metals from certified reference material (NIST-2704 Buffalo River sediment) was within 10% of the certified values. In order to relate sediment metal concentrations to date of sediment deposition, a fourth core was sectioned into 1-cm slices, analyzed for total metals, and dated by Pb-210 analysis. Dating of freeze-dried sediment samples was performed by Paul Wilkinson, Freshwater Institute, Fisheries and Oceans Canada, Winnipeg. Lead-210 dates were confirmed by Cs-137 analysis (Wilkinson, 1985, and references quoted therein).
3. Results 3.1. Core profiles for metals in sediment, overlying water and caged Hyalella Strong profiles for total metal in the sediment were observed for Cd, Co, Cu and Ni (Fig. 1). Metal
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concentrations were lowest in the deepest part of the core, increased rapidly up to approximately 5 cm depth, and then decreased slightly towards the surface. Profiles for the three core sections used for bioavailability and toxicity testing were quite similar, except that one of the profiles (--, Fig. 1) showed maximum concentrations for all metals at a slightly shallower depth than the other two profiles. The surface slice and surface spoon samples had metal concentrations similar to those near the top of the core. The mini-ponar samples had metal concentrations in the mid range of the concentrations of the top 10 cm of the cores (Fig. 1), as expected since this sampler collects sediments down to about 10 cm depth. Profiles for metals in overlying water, and for metals in Hyalella exposed to overlying water in the cages, differed somewhat from the profiles of total metal in the sediment (Figs. 2 and 3). The profiles of Cd, Co and Ni in Hyalella matched the profiles for these metals in overlying water, but these were more pronounced than the profiles of these metals in the sediment. Metals in Hyalella correlated better with metals in water (r2=0.64, 0.71, 0.80 for Cd, Co and Ni, respectively) than with metals in sediment (r2=0.55, 0.47, 0.55). Metals in water and in Hyalella increased rapidly with increasing depth in the top 5 cm, whereas concentrations of these metals in the sediment increased only moderately with depth over this range (Fig. 1). Concentrations of Cd, Co and Ni in water and Hyalella from test containers with surface sediments, including sections from the top of the core as well as surface slice and surface spoon samples, were similar to those from containers with core sections near 12 cm (Figs. 2 and 3). However, total metal concentrations in the sediment at 12 cm were only about half as high as those at the surface (Fig. 1). Concentrations of Cd and Ni in overlying water and Hyalella from test containers with ponar grab samples were in the mid range of those which had core sections between 0 and 10 cm. However, Co concentrations in water and Hyalella from containers with ponar samples were similar to those with surface and 10 cm core sections, which were substantially lower than samples from intermediate depths. Profiles for Cu in overlying water and in Hyalella differed substantially from the other profiles. Copper in overlying water decreased with increasing depth of the core section in approximately a linear fashion (Fig. 2). Although the data are more variable than those for total metal in sediment, there was no evidence of a maximum near 5 cm with lower concentrations at the surface, as seen for Cu in sediment (Fig. 1). Copper in Hyalella, on the other hand, remained approximately constant, with only a slight trend towards lower concentrations with depth (Fig. 3). This is consistent with the ability of Hyalella to regulate Cu near essential background levels, except at very high environmental concentrations (Borgmann and Norwood, 1995a, b, 1997).
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Fig. 1. Concentrations of Cd, Co, Cu and Ni in three separate sediment cores from Richard Lake (, +, ), in surface slice and surface spoon samples (*) and in three different mini-ponar grab samples (X). Mini-ponar grab samples are plotted at a depth of 5 cm, but the sampler actually collects sediments between 0 and approximately 10 cm depth.
Other metals measured in sediment, overlying water and Hyalella showed less pronounced profiles, or were much lower in concentration, than Cd, Co, Cu or Ni (Table 1). Arsenic and Cr varied little with depth. Manganese was quite variable and sometimes high at the surface, but otherwise showed no consistent trend. Lead showed a strong profile in the sediment, but was moderately low in concentration and much less bioavailable than Cd. Thallium was also low. The profile for Zn was similar in shape to Cu and Ni, but less pronounced, and total Zn in the sediment was much lower than Cu or Ni (Table 1). Zinc concentrations in Hyalella did not vary with depth, and were near the background concentration of this essential metal in Hyalella (Borgmann and Norwood, 1995b). 3.2. Toxicity and Ni in sediment, water and Hyalella Chronic (4-week) toxicity of the sediment core sections to Hyalella produced a profile that matched metal
concentrations. Mortality was almost complete for amphipods exposed to sediments from 3 to 14 cm depths. Survival was high (580%), however, in the deepest sediments (Fig. 4). There was partial survival at the shallowest depths in sections from one of the cores and in two of the tests with the surface spoon sample, coinciding with the slightly lower metal concentrations at the surface. In contrast, survival of adult Hyalella caged in the experimental containers for only one week was 580% regardless of sediment depth. Chronic mortality of young amphipods exposed to sediments correlated reasonably well with Ni bioaccumulation in caged adult Hyalella exposed for one-week in the same test containers (Fig. 4). Survival was compared to bioaccumulation in caged Hyalella because complete mortality in many of the test containers precluded measurement of metal accumulation in sedimentexposed animals. Previous chronic toxicity tests with ponar grab samples from six sites in four lakes in the Sudbury area, including Richard Lake, demonstrated
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Fig. 2. Concentrations of Cd, Co, Cu and Ni in overlying water during toxicity and bioaccumulation tests with sediments from three separate cores from Richard Lake. Symbols as in Fig. 1. Breaks in the lines are due to missing intermediate data values.
that chronic toxicity in caged amphipods was equal to that in sediment exposed animals. This implies that toxicity was caused by a dissolved substance and not the solid-phase sediment (Borgmann et al., 2001b), and suggests that comparison of survival and bioaccumulation in sediment-exposed and caged animals is reasonable. As in the previous study (Borgmann et al., 2001b), Ni was the only metal accumulated in excess of chronic lethal body concentrations. The curve drawn through the survival-bioaccumulation data in Fig. 4 was obtained from a previous study with Ni-spiked sediments conducted in Imhoff settling cones (Borgmann et al., 2001a) and is described by m ¼ 0:0196 þ 44:3 109 Ni2:758 TB
ð1Þ
where m is the weekly instantaneous mortality rate (ln(survival)/t) and NiTB is the total body concentration of Ni in Hyalella. This was obtained by combining the non-linear regression of m against Ni in water (Niw)
m ¼ 0:0196 þ 349 109 Ni2:032 w with bioaccumulation from water (NiTB=2.114 Ni0.737 ) for data in Borgmann et al. (2001a). The fit w between the observed survival and Ni in caged Hyalella with the independently derived survival-body concentration relationship from Ni-spiked sediments (Fig. 4) further supports the hypothesis that toxicity is primarily due to Ni bioaccumulated from dissolved Ni leached from the sediments. Since Ni is the primary cause of sediment toxicity, a closer examination of Ni bioaccumulation from sediment is warranted. Nickel bioaccumulation from sediment appeared to follow a bi-phasic trend (Fig. 5). For data obtained from core sections below 5 cm, Ni bioaccumulated by Hyalella was approximately proportional to Ni in the sediment. Log-log regression of the data gave a slope of 1.18 (95% CI=0.95–1.42). For data obtained from core sections above 5 cm, however, the slope was 3.36 (95% CI=2.58–4.13). The relative Ni
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Fig. 3. Concentrations of Cd, Co, Cu and Ni in Hyalella caged above the sediments for one week during toxicity tests with sediments from three separate cores from Richard Lake. Symbols as in Fig. 1. Breaks in the lines are due to missing intermediate data values.
bioavailability, therefore, decreases rapidly towards the surface. The bi-phasic nature of the bioaccumulation relationship disappears, however, if bioaccumulation is expressed relative to Ni in overlying water (Fig. 5). Bioaccumulation relative to Ni in water was similar to that observed previously with Ni-spiked sediments (Borgmann et al., 2001a). Furthermore, Ni bioaccumulated in Hyalella exposed directly to the sediment, for those treatments with surviving animals, was similar to Ni bioaccumulation by caged Hyalella (Fig. 5). The decreased relative bioavailability of Ni in the top few cm of the sediment is, therefore, due to a reduced relative leaching of Ni into the overlying water. 3.3. Core dating Lead-210 dating of a fourth core sectioned into 1-cm intervals down to 20 cm revealed a constant sedimentation rate on a dry weight basis of 241 g m2 year1. Metal profiles in this core were similar to the
three cores used for bioaccumulation and toxicity testing. In order to adjust for slight differences in sedimentation rates of the different cores, a common reference point was needed. This was obtained by calculating the depth at which the concentration of Cu and Ni, the two most abundant metals, was equal to one-half of the maximum of the three point running mean concentration. For the Pb-210 dated core, this was equal to 8.91 and 10.29 cm for Cu and Ni, respectively. For the three toxicity test cores this ranged from 7.32 to 9.39 and 8.56 to 10.76 cm, respectively. Dividing these depths by the depths obtained for the Pb-210 core and averaging the data for Cu and Ni provided relative depths of the contaminated layer of 1.000, 0.8268, 1.0334 and 1.0498 for the Pb-210 dated and three toxicity cores. Assuming that the depth of the contaminated layer represented an equal deposition time period for each of the cores, these relative depths were used as correction factors to age the three toxicity cores. This allowed plotting of metal concentrations in
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Table 1 Mean metal concentrations (S.D.) measured in sediment core slices and in overlying water and caged Hyalella exposed to those slices for three depth ranges (n=6) Depth range (cm)
As
Cd
Co
Cr
Cu
Mn
Ni
Pb
Tl
Zn
0.036 (0.005) 0.048 (0.004) 0.014 (0.005)
1.25 (0.19) 1.78 (0.19) 0.39 (0.08)
1.37 (0.05) 1.41 (0.09) 1.28 (0.08)
23.3 (0.7) 26.3 (0.8) 2.6 (0.7)
11.9 (5.4) 6.9 (1.4) 6.8 (0.7)
32.1 (5.4) 47.3 (4.1) 10.0 (3.0)
0.41 (0.07) 0.39 (0.08) <0.0125
nm nm nm
4.11 (0.56) 4.69 (0.97) 1.79 (0.20)
Overlying water (nmol l1) 0–1 18 (6) 4–5 34 (8) 16–20 28 (4)
0.60 (0.51) 1.31 (0.20) 0.19 (0.09)
5.5 (8.2) 29.8 (6.2) 0.6 (0.3)
7.4 (0.6) 7.7 (1.9) 6.8 (1.0)
169 (20) 153 (23) 44 (21)
3.9 (3.4) 8.4 (5.0) 1.3 (0.7)
762 (610) 3325 (836) 408 (191)
<0.64 <0.64 <0.64
0.06 (0.03) 0.15 (0.02) 0.08 (0.01)
<150 <150 <150
Hyalella-caged (nmol g1) 0–1 20 (3) 4–5 20 (5) 16–20 23 (4)
13 (5) 19 (2) 6 (3)
8 (6) 24 (2) 3 (1)
25 (3) 27 (6) 25 (4)
1154 (140) 1270 (194) 1052 (189)
83 (20) 81 (9) 95 (28)
346 (199) 1169 (278) 165 (73)
1
Sediment (mol g ) 0–1 nma 4–5 nm 16–20 nm
a
0.6 (0.4) 0.5 (0.2) 0.2 (0.1)
0.4 (0.1) 0.9 (0.1) 0.5 (0.1)
872 (43) 845 (24) 871 (22)
nm, Not measured.
Fig. 4. Survival of young Hyalella during 4-week chronic exposures to sediments from three separate cores from Richard Lake expressed as a function of sediment depth or Ni concentrations in caged Hyalella. Symbols as in Fig. 1 for survival versus depth. Data obtained from < 5 cm (*) and from 55 cm (&) deep sediments are shown in different symbols for survival versus Ni in Hyalella.
each of the cores as a function of predicted date of sediment deposition. This suggests that Ni concentrations in the sediment started increasing around 1890, peaked around 1970, and then declined continuously until the present (Fig. 6).
4. Discussion 4.1. Relative metal bioavailability and sediment depth The relative bioavailability of Cd, Co and Ni differed substantially from that of Cu through the sediment core. Bioaccumulation of Co and Ni followed the trend for total metal in sediment below 5 cm, but dropped rapidly to very low levels for amphipods exposed to surface sediments (Fig. 3). This rapid drop with
decreasing depth in the top 5 cm was not observed for Co and Ni in the sediment (Fig. 1), but did occur in the overlying water (Fig. 2). Concentration trends for Cd in Hyalella, sediment and overlying water generally followed those of Co and Ni, but the data are more variable and the trends less clear, probably because of the much lower concentration of this metal. In contrast, Cu bioaccumulation by Hyalella was almost constant with exposure to sediments from varying depth. In part, this is due to regulation of body Cu by this amphipod (Borgmann and Norwood, 1995a, b, 1997). However, Cu bioavailability to Hyalella in sediments appears to be due to dissolved Cu in the overlying water (Deaver and Rodgers, 1996), and Cu in overlying water followed a very different trend from the other metals, with no noticeable decrease in test chambers with surface sediments (Fig. 2).
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Fig. 5. Nickel concentrations in Hyalella caged above the sediments for one week (*, &) or exposed directly to the sediment for 4 weeks (*, &), as a function of Ni in the sediment or Ni in the overlying water. Data obtained from <5 cm (*, *) and from 55 cm (&, &) deep sediments are shown in different symbols. Lines were obtained from linear regression using the caged animal data. Separate regressions were conducted for the two depth ranges for the sediment concentration data. The dotted line was obtained from previous experiments with Hyalella exposed to Ni-spiked sediments (Borgmann et al., 2001a).
Fig. 6. Nickel in the sediment as a function of date of sediment deposition for the three toxicity/bioaccumulation test cores and the Pb-210 dated core (~). Data after 1970 are shown in expanded time scale in the inset.
Part of the reason for the differences in metal released into overlying water, and the resulting differences in bioavailability, may lie with the chemical nature of the metals. Although Cd, Co, Cu and Ni can all be classified as borderline metals, the relative amount of class A and class B character varies, with class B character increasing in the order Ni < Co < Cd < Cu (Nieboer and Richardson, 1980). Class A metals prefer to bind to oxygen atoms in ligands, followed by nitrogen and then sulfur. The reverse is true for class B metals. With increasing depth through the sediment core, oxygen
gradually becomes depleted, and it might be expected that the greatest availability of oxygen binding sites in both organic and inorganic matter would be found near the surface. Hence surface sediments might be expected to bind Ni more strongly than deeper sediments, whereas depth should have less of an influence on Cu binding. The toxicity of sediment core sections buried below the surface will be controlled by both the concentration of toxic materials incorporated at the time of sediment deposition, and geochemical processes occurring during prolonged in situ incubation. It is, therefore, possible that the relative bioavailability of Ni in the deeper core sections was initially lower than measured here, and similar to the bioavailability of Ni in the surface sediments now. Nevertheless, the deepest sections of the core were non-toxic (Fig. 4). Since Ni bioaccumulation factors (Ni in Hyalella divided by Ni in sediment) increased, rather than decreased, with depth (i.e. the slope of the line in Fig. 5 for < 5 cm depths is greater than 1), this implies that the surface sediments in Richard Lake were probably non-toxic prior to industrial development in the Sudbury area. Toxicity measurements of sediment core sections have provided information on potential changes in sediment toxicity through time, or the total volume of toxic sediments present, in several studies (Rosiu et al., 1989; Stemmer et al., 1990; Swartz et al., 1991; West et al., 1994). In addition, core sections have been used to infer long term changes in contaminant (PAH) bioavailability after burial (Harkey et al., 1995). However, most core sectioning for toxicity or bioavailability has been rather coarse (4–5 cm slice thickness). In the present study slice thicknesses varied from 0.5 cm near the surface to 2 cm deeper in the core. Furthermore, toxicity tests were
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combined with simultaneous estimates of metal bioavailability, allowing direct comparison of sediment contamination and toxic effects. Most of the changes in the relative bioavailability of Ni, the toxic agent in Richard Lake sediments, were observed within the top 5 cm of sediment. Much information would have been lost if sectioning were done only at 5 cm intervals. Core sections for toxicity testing or bioaccumulation measurements in the past have usually been relatively thick in order to provide enough sediment to conduct the tests. However, sediment toxicity and bioaccumulation studies can be conducted in Imhoff settling cones with only 15 ml of sediment (Borgmann and Norwood, 1999a). This makes more detailed analysis of core section toxicity and bioaccumulation possible. 4.2. Implications for sediment collection methodology The variation in relative Ni bioavailability with depth indicates that the results from bioaccumulation and toxicity tests could be affected by the method of sediment collection. To estimate the potential effect this could have on metal impact assessments, toxic sediment concentrations (LC25s) were estimated from Ni bioaccumulation from both the sediment core slices in Richard Lake (Fig. 5) and from previously collected multiple ponar grab samples from 21 sites in 12 lakes in the area between Sudbury and North Bay (Borgmann et al., 2001b). The total range in Ni in ponar grab samples from these lakes, and the variability in Ni bioavailability between these samples, were much greater than the variability observed through the core sections from Richard Lake (Fig. 7). The body concentration which
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caused 25% mortality in chronic tests with Hyalella (the LBC25) was 194 nmol g1 Ni (Borgmann et al., 2001a). Using the regression line for the ponar data in Fig. 7, this is equivalent to a sediment concentration of 19.5 mmol g1 Ni (the ponar-based sediment LC25). On the other hand, the LC25 can also be estimated using only surface sediments from the Richard Lake core. Nickel bioaccumulation in Hyalella exposed to the top centimeter of sediment was 346 nmol g1 (Table 1), which was 1.78-fold greater than the LBC25. If Ni bioavailability is proportional to Ni in the sediment at constant depth in the core, then the LC25 for Ni in the top 1 cm of sediment should be 32.1 mmol g1 (Table 1) divided by 1.78, which equals 18.0 mmol g1 (the surface-sediment LC25). Coincidentally, the more careful core analysis conducted in this study gave an almost equivalent estimate of the LC25 for Sudbury area sediments as the previously collected ponar grab data (18.0 vs. 19.5 mmol g1). Although the relative Ni bioavailability in ponar grab and deeper core samples was higher than Ni bioavailability in surface sediments from Richard Lake, the ponar-based regression line lies below the ponar data from this lake, and closer to the surface sediment samples (Fig. 7). The method of sediment collection does not, therefore, greatly affect the estimation of the LC25 for Ni in sediment in these two Sudbury area studies, but this seems to be more a matter of luck than good experimental design. It resulted from the fact that the regression line in Fig. 7 falls below the mean for the ponar data from Richard Lake. In general, metal concentrations in ponar grab samples collected from Sudbury area lakes are much more variable than surface (0–3 or 0–5 cm) samples. It is, therefore, probably a better practice to use more carefully sectioned surface sediment samples for toxicity testing, rather than sediments collected by ponar or similar samplers, whenever possible. 4.3. Predicted time to recovery from sediment toxicity
Fig. 7. Nickel in Hyalella as a function of Ni in sediment for core slices () or ponar grab samples (X) collected in 1999 from Richard Lake, or from ponar grab samples collected from 12 different lakes in 1998 (*). The line was obtained from linear regression of the 1998 ponar data.
The sediment core Ni chronology in Richard Lake (Fig. 6) is similar to the time trend in Ni production from Sudbury smelters. Smelting started in 1888 and peaked in the 1970s (Winterhalder, 1995). Completion of Inco’s ‘‘Superstack’’ and major emission controls in 1972 (Conroy and Kramer, 1995) probably contributed to the decline in surface metal concentrations in recent years. Interpretation of the details in the sediment core profiles must, however, be done with caution. For example, Ni concentrations decreased more rapidly than did Cu concentrations in the most recent years in Richard Lake profiles (Fig. 1). However, Cu concentrations decreased more rapidly than Ni concentrations in recent years in core profiles from McFarlane Lake, the next lake downstream from Richard Lake and only a few kilometers away (data not shown). Individual lake
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morphometry or geochemistry might be having some influence on core profiles for individual metals. Using the above estimate of Ni toxicity in surface sediments, it is possible to estimate roughly the potential time required for the surface sediment to become non-toxic. Linear regression of Ni concentrations against the estimated date of sediment deposition for data after 1980 (Fig. 6) gave a slope of 0.82 mmol g1 year1 (r2=0.66, n=28). A concentration of 18 mmol g1 (the 4-week LC25) would be reached in 2016, if this trend continues. It is, however, possible that the rate of decrease is increasing, as suggested by data from three of the four cores after 1992 (Fig. 6). Alternatively, the rate of decrease might slow, with Ni concentrations exponentially approaching pre-1890 concentrations of around 10 mmol g1. In either case, Ni concentrations are currently decreasing, and a more precise estimate of when Ni concentrations at the surface will reach the LC25 in Richard Lake should be obtainable in about another decade. Acknowledgements The field work was done by M. Mawhinney and M. Benner of the Technical Operations Section, NWRI, with assistance from the authors. Total metal concentrations in sediment were measured by J. Rajkumar. ICP-MS analysis for metals in water and Hyalella were done by the National Laboratory for Environmental Testing. Lead-210 dating was done by Paul Wilkinson, Freshwater Institute, Fisheries and Oceans Canada, Winnipeg. References Borgmann, U., Norwood, W.P., 1995a. Kinetics of excess (above background) copper and zinc in Hyalella azteca and their relationship to chronic toxicity. Can. J. Fish. Aquat. Sci. 52, 864–874. Borgmann, U., Norwood, W.P., 1995b. EDTA toxicity and background concentrations of copper and zinc in Hyalella azteca. Can. J. Fish. Aquat. Sci. 52, 875–881. Borgmann, U., Norwood, W.P., 1997. Toxicity and accumulation of zinc and copper in Hyalella azteca exposed to metal-spiked sediments. Can. J. Fish. Aquat. Sci. 54, 1046–1054. Borgmann, U., Norwood, W.P., 1999a. Sediment toxicity testing using large water-sediment ratios: an alternative to water renewal. Environmental Pollution 106, 333–339.
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