Environ. Sci. Technol. 1998, 32, 620-625
Effect of Short-Term Resuspension Events on Trace Metal Speciation in Polluted Anoxic Sediments STUART L. SIMPSON,* SIMON C. APTE, AND GRAEME E. BATLEY Centre for Advanced Analytical Chemistry, CSIRO Division of Coal & Energy Technology, PMB 7, Bangor, NSW 2234, Australia
The effects of the short-term resuspension of a contaminated anoxic estuarine sediment on solid-phase metal speciation have been studied. Preliminary experiments investigated the oxidation rates of model metal sulfide phases to provide mechanistic information for interpreting the observations on the natural sediment. FeS and MnS model phases were particularly labile and oxidized rapidly in aerated waters. In contrast, CdS, CuS, PbS, and ZnS model phases were kinetically stable over periods of several hours. The oxidation rate of free sulfide (HS-) was significantly slower than that of FeS and MnS. Upon sediment resuspension, the rapid decrease in acid volatile sulfide (AVS) could be accounted for by the oxidation of iron monosulfide phases. Over prolonged resuspension periods (>300 min), AVS decreased to values lower than the simultaneously extracted metals [SEM ) ∑Cd, Cu, Ni, Pb, Zn (1 M HCl extraction, 30 min)] concentration, indicating that a significant fraction of trace metal sulfide phases may be oxidized during resuspension events. During 8-h sediment resuspension experiments, SEM(Cu) increased from 0.1 to ca. 2 µmol/g while the SEM measured for the other metals remained constant. The increase in SEM(Cu) was shown to be an artifact of the AVS/SEM analytical procedure. Fe(OH)3, formed through the oxidation of FeS, dissolved upon acidification to produce Fe3+(aq), which subsequently oxidized acid-insoluble copper sulfide mineral phases in the sediment. The implications of these observations for natural systems and for the assessment of metal toxicity using AVS/SEM procedures are discussed.
Introduction Equilibrium partitioning approaches have been used to predict the porewater concentrations of heavy metals in contaminated oxic sediments and the potential fluxes to overlying waters. The oxic fraction in silty sediments usually extends to depths of 2-5 mm. More important, therefore, is the need to understand the behavior of anoxic sediments. While the equilibrium solubility of metals in anoxic sediments may in principle govern dissolved metal concentrations, of much greater significance is the potential for oxidative release that might occur during the resuspension induced by physical disturbances such as dredging or strong currents or by the introduction of oxic waters at depth by burrowing organisms. Bioirrigation rates as high as 750 mL/h have been observed * Corresponding author phone: +61 9710 6807; fax: +61 97106837; e-mail:
[email protected]. 620
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 32, NO. 5, 1998
(1). It is likely, especially in estuarine sediments, that the oxidation of sulfide phases represents the major source of metals to oxic overlying waters. In most anoxic sediments, sulfides are considered the predominant solid phases controlling the concentrations of the metals Cu, Cd, Fe, Mn, Ni, Pb, and Zn (2, 3). The low solubility of metal sulfides results in low porewater concentrations. Upon resuspension of a sediment into an oxic overlying water, these metal sulfides may oxidize, thereby producing oxidized sulfur species (i.e., SO42-, S0), and release the associated metal to the water column (4). The released metal may in turn be quickly scavenged by or coprecipitated with iron and manganese hydroxides or complexed by organic matter (5). Following the 6-h resuspension of a polluted anoxic sediment into seawater, Hirst and Aston (6) observed significant release of Fe and Mn but not of Cu or Zn. Upon resuspension of a contaminated dredge spoil into an estuarine water, Prause et al. (7) observed that neither Pb or Cd were released within 10 h and only Cd was observed to be released over 50 days. Calmano et al. (8) investigated the long-term (25 days) mobilization and scavenging of heavy metals following resuspension of an anoxic river sediment. Total releases of metals were small [Cd (5%) > Zn > Cu > Pb (0.7%)], and calculated scavenging of metals was in the reverse order [86%-30%]. Cu and Zn released correlated well with sulfate concentration, indicating that sulfide oxidation was the predominant release process. The research to date has investigated the effect of sediment resuspension on the majority of potentially toxic metals. However, despite the many comprehensive studies, there is very little consensus on the release and long-term mobility of metals. The combined effects of redox potential, Eh, together with pH, salinity, and organic matter contents of both sediment and overlying water greatly influence research conclusions. Although much studied, these interacting parameters are not currently understood to the extent that the mobilization or scavenging potential of a sediment may be predicted (9, 10). pH is most commonly identified as the master variable. Recent efforts to develop quality criteria for aquatic sediments in both freshwater and marine environments have targeted the sulfide phases during sediment appraisal. In particular, the measurement of acid volatile sulfide (AVS) and simultaneously extracted metal (SEM) contents of sediments have been applied to assess bioavailability and predict toxicological effects of the metals Cd, Cu, Ni, Pb, and Zn (11-13). Short-term changes in the distribution of metals associated with the sulfide fraction of sediments has strong implications for the assessment of sediment quality. The present investigation reports the effects of the shortterm resuspension of a contaminated anoxic estuarine sediment into an aerated overlying water on solid-state metal speciation. Experiments were also conducted with model metal sulfide phases to elucidate the mechanisms by which these changes occur. This paper forms part of a larger investigation on metal release from sediments in polluted systems.
Experimental Section All glass and plasticware was cleaned by soaking in 10% HNO3 (v/v) for 48 h followed by 12-h soaking and repeat rinsing with deionized water (Milli-Q). Nalgene bottles (LDPE) were used in all experiments. All chemicals were ACS reagent grade or equivalent analytical purity. S0013-936X(97)00568-3 CCC: $15.00
1998 American Chemical Society Published on Web 01/22/1998
Analyses. pH measurements were made using a combination pH probe (AEP321, Activon) calibrated against standard buffers. Dissolved oxygen (DO) was measured using a YSI 58 dissolved oxygen meter, calibrated by the manufacturer-recommended procedures. The moisture content of the sediment was determined following drying at 110 °C. Total organic carbon was determined by ‘loss on ignition’ at 450 °C. The procedure used for acid volatile sulfide (AVS) and simultaneously extracted metal (SEM) determinations was that described by Allen et al. (14). Total reducible sulfide (TRS) was determined according to the method described by Canfield et al. (15). Total particulate metals (TPM) were measured following an aqua regia digestion (20 min, 450 W domestic microwave at 10% power). As a check on analytical quality, a standard reference material (PACS-1, National Research Council Canada) was analyzed in each batch of samples. All metal ion determinations were performed using ICP-AES calibrated with commercially available standards. Sediments and Waters. A single sediment sample (30 kg, 10-30 cm depth) was collected from an intertidal location in the Cook’s River, NSW, Australia. The entire sediment was sieved (4 mm plastic sieve) to remove larger debris, thoroughly homogenized, and stored in a plastic container under 20 cm of overlying estuarine water (21‰). The sediment was then not disturbed for 5 weeks. Seawater was collected from Cronulla, NSW, while freshwater was collected from the Woronora River, NSW. The waters were stored at room temperature and were discarded upon observation of pH changes or biological activity (usually after 6 weeks). Model Metal Sulfide Oxidation Studies. Amorphous sulfides of Cd(II), Cu(II), Fe(III), Mn(II), Ni(II), Pb(II), and Zn(II) were prepared by spiking 5 mL of 50 mM Na2S solution with 2.5 mL of the respective metal chloride (100 mM, in deoxygenated water) to precipitate 250 µmol of metal sulfide. Copper(I) sulfide (125 µmol of Cu(I)2S) was prepared similarly by spiking sulfide solution with Cu(I)Cl (prepared as a suspension). All syntheses were performed in a nitrogenfilled glovebox with utmost care to prevent any oxidation. The metal sulfides, which precipitated immediately, were stored for 16 h in this environment until oxidation experiments were initiated. For oxidation studies, the metal sulfide slurry (7.5 mL) was quantitatively transferred to 92.5 mL of Milli-Q water (housed in a 250-mL Nalgene bottle). This bottle was placed on a purpose-built roller running at 20 rpm for the desired resuspension period (0-8 h). In further experiments, both copper sulfide phases were resuspended with (i) 250 µmol of Fe(III) for 0-8 h and (ii) 0-1000 µmol of Fe(III) for 6 h present in the Milli-Q water. The resulting pH of all metal sulfide suspensions was 8.0 ( 0.5. AVS and SEM determinations were made at the end of the resuspension periods following the quantitative transfer of the entire contents of the bottle to the AVS/SEM apparatus. For samples not resuspended (t ) 0), experiments were performed in both the absence and the presence of oxygen to ascertain the degree of sample oxidation occurring upon immediate contact with the oxygenated water. All experiments were at least duplicated to check the precision of the measurements. Sediment Resuspension Studies. Sediment cores were taken from the collected sediment and extruded (minus the top 5 cm) into a clean plastic container in a nitrogen-filled glovebox. This sediment was further homogenized using a glass rod and stored in a closed container to prevent moisture loss. For resuspension studies, ca. 2.5 g of wet sediment (ca. 1 g dry wt) was transferred from the glovebox on a piece of Parafilm, weighed, and immediately deposited in 100 mL of the respective natural water and resuspended for 0-8 h as described for the metal sulfide oxidation experiments.
FIGURE 1. AVS measured for separate experiments on sulfide (HS-) and model metal sulfide phases following suspension for 0-8 h. The initial sulfide concentration was 250 µmol in all experiments. Separate but identical resuspension experiments were undertaken for measurements of (i) pH, DO, and TPM; (ii) AVS and SEM; and (iii) TRS. Measurements of pH and DO were made immediately following the resuspension period, and TPM was determined on residual sediment. AVS/SEM and TRS determinations were made following quantitative transfer of the entire contents of the bottle to the respective apparatus. In further experiments, the effect of Fe(III) and Fe(II) on SEM extracted from the sediment was investigated (no resuspension). For these experiments, the sediment samples were transferred directly to the AVS/SEM apparatus, which contained 0-400 µmol of (i) Fe(III) (FeCl3‚6H2O) and (ii) Fe(II) (FeSO4‚7H2O) in 100 mL of deoxygenated Milli-Q water. SEM (but not AVS) was subsequently determined.
Results and Discussion Model Metal Sulfide Oxidation Studies. AVS and SEM analyses are useful tools for interpreting changes occurring to metal sulfide phases as a result of oxidation processes. The AVS procedure used (1 M HCl for 30 min (14)) extracts H2S, amorphous FeS, mackinawite (FeS(1-x)), and some greigite (Fe3S4) but no pyritic sulfides (FeS2) from sediments. The metals that dissolve during this extraction are termed SEM. It is assumed that the sulfide phases extracted are predominantly monosulfide phases (e.g., FeS, MnS, ZnS) and that for every mole of sulfide measured as AVS there will be a corresponding mole of metal measured in the SEM fraction. Also assumed is that nonsulfidic binding phases (e.g., hydroxide/carbonate phases) are insignificant in anoxic, sulfidic sediments. For the model metal sulfide phases, the observation that AVS measured equals SEM measured indicates that no oxidation has occurred. In contrast, if AVS is less than SEM, then some oxidation of the metal sulfide phases may have taken place. During the AVS/SEM analyses (1 M HCl, 30 min), the model metal sulfide phases CdS, FeS, MnS, and ZnS were observed to dissolve rapidly, while PbS dissolved less rapidly. In contrast, the CuS and NiS model phases were not observed to dissolve. No AVS was measured for either CuS or NiS suspensions (Figure 1) because the sulfide ion binds these metals with such strength that they are stable under high acidities. The findings for Cu are consistent with the observation by Allen et al. (14). The SEM(Ni) initially observed upon acidification (ca. 60% of the total Ni, t ) 0, Figure 2) was attributed to incomplete formation of NiS and/ or very fast oxidation of the NiS model phase rather than dissolution of NiS by the HCl. The changes to AVS and SEM upon suspension of the model metal sulfide phases are shown in Figures 1 and 2 along with the changes occurring to 250 µmol of sulfide (added as Na2S). From Figure 1 it is apparent that, upon suspension in aerated waters, both FeS and MnS oxidize readily. For MnS, however, manganese (hydr)oxide phases VOL. 32, NO. 5, 1998 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
621
FIGURE 2. SEM extracted from model metal sulfide phases (250 µmol total metal) following suspension for 0-8 h. formed (e.g., MnO2, MnOOH) that were not completely soluble, and a small amount of dark brown precipitate remained at the end of the AVS extraction process. This resulted in a decrease in SEM(Mn) for longer suspension times, as observed in Figure 2. For both CdS and ZnS, the recoveries of AVS indicated that 300 min), the observation that AVS dropped below SEM indicated that almost half of the ZnS would have oxidized. Given the observed stability of model metal sulfide phases in oxygenated waters, we believe that the contrasting behavior of the natural sediment indicates that a significant fraction of the sedimentary trace metal sulfide phases were not present as monosulfide phases, e.g., metals associated with pyritic phases (4, 22). Although thermodynamics may predict the formation of trace metal monosulfide phases, the complexity of sediments, in particular the sulfur cycle (23), results in many sediments being far from equilibrium systems with a significant number of metal solid phases coexisting for each trace metal present. Reaction kinetics may be very important in determining metal binding sites. In polluted sediments, more recently added metals may be expected to be enriched in the outer layers or absorbed to surfaces of particles rather than occluded deep within the mineral matrixes. Trace metal pyritization is an important sink for many metals in anoxic marine sediments (4). Upon resuspension of sediments in oxic seawater, Morse (4) observed that the oxidative release of heavy metals from the authigenic pyrite fraction was greater than the extent of pyrite iron oxidation. Trace metals adsorbed to the surface of pyritic phases may be acid soluble and thus observed in the SEM fraction; however, no sulfide would be released. Furthermore, the binding of metals to particulate organosulfur species in sediments is very poorly understood. Such phases may also play an important role in the binding of trace metals. Another plausible explanation for the discrepancy between the observations for the natural sediment and those for model metal sulfide phases is that microbial-catalyzed oxidation may be occurring (24, 25). Although the exact nature of the trace metal sulfide phases in the studied sediment has not been accurately elucidated in these studies, very important implications can still be drawn from this work. In particular the discussion above questions the fundamental assumption of the AVS/SEM theory, that the sulfide phases extracted are predominantly monosulfide phases (e.g., FeS, MnS, ZnS) and that for every mole of sulfide measured as AVS there will be a corresponding mole of metal measured in the SEM fraction. The feasibility of using AVS/SEM measurements to evaluate sediment quality, assess the bioavailability, or predict toxicological effects of the metals Cd, Cu, Ni, Pb, and Zn has received much emphasis recently (26). In consideration of the observation that CuS and NiS were not soluble in 1 M HCl, there is little theoretical basis for the interpretation of sediment toxicity from AVS/SEM relationships for these metal ions. Furthermore, the observations that (i) Fe(III) oxidizes CuS during the SEM extraction resulting in greater SEM(Cu) measured but no corresponding measurement of AVS and
that (ii) copper sulfide phases of different stoichiometry (i.e., Cu(II)S and Cu(I)2S) are extracted to different extents result in the interpretation of SEM/AVS relationships for Cu being quite ambiguous. The attempts by several researchers (27, 28) to use AVS/SEM measurements as a tool to investigate changes in the bioavailability of the metals Cu and Ni were unfortunately flawed by the observations discussed above. The observation by Besser et al. (28), that in sediments incubated with oxic overlying waters, SEM(Cu) concentrations increased while SEM(Zn) concentrations remained constant relative to sediments incubated under anoxic conditions can also be interpreted as merely artifacts of the AVS/SEM method. Simpson et al. (29) provide a fuller discussion of this subject. During bioturbation, filter feeders inject oxygenated water into lower sediments while surface dwellers regularly resuspend the top layers of the sediment. Sediment disturbances as a result of bioturbation have been proposed as processes that may significantly disrupt the metal sulfide binding in sediments (30). In particular, there has been concern about the release of the heavy metals Cd, Cu, Ni, Pb, and Zn from their sulfide phases. The results of the current investigation indicate that, during bioturbation, FeS and MnS phases (usually present in large excess to other metal sulfides) may buffer the effects of bioirrigation and that trace metal sulfide phases may remain predominantly unoxidized for some time. However, any trace metals occluded in or coprecipitated with iron sulfide phases may be particularly prone to oxidative release. Further experiments are required to accurately measure and relate the release of dissolved trace metals from anoxic sediments to the oxidation of metal sulfide phases occurring upon short-term exposure to oxic waters.
Literature Cited (1) Foster-Smith, R. L. J. Exp. Mar. Biol. Ecol. 1978, 32, 73-95. (2) Emerson, S.; Jacobs, L.; Tebo, B. In Trace Metals in Sea Water; Wong, C. S., et al., Eds.; Plenum Press: New York, 1983; pp 579-608. (3) Dyrssen, D.; Kremling, K. Mar. Chem. 1990, 30, 193-204. (4) Morse, J. W. In Metal Contaminated Aquatic Sediments; Allen, H. E., Ed.; Ann Arbor Press: Ann Arbor, MI, 1995; Chapter 8, pp 187-199. (5) Balistrieri, L. S.; Brewer, P. G.; Murray, J. W. Deep-Sea Res. 1981, 28A, 101-121. (6) Hirst, J. M.; Aston, S. R. Estuarine Coastal Shelf Sci. 1983, 16, 549-558. (7) Prause, B.; Rehm, E.; Schulz-Baldes, M. Environ. Technol. Lett. 1985, 6, 261-266. (8) Calmano, W.; Fo¨rstner, U.; Hong, J. In Environmental Geochemistry of Sulfide Oxidation; Alpers, C. N., Blowes, D. W., Eds.; American Chemical Society: Washington, DC, 1994; pp 298321.
(9) Calmano, W.; Ahlf, W.; Baabe, H.; Fo¨rstner, U. In Heavy Metals in the Hydrological Cycle; Astruc, M., Lester, J. N., Eds.; Selper Ltd.: London, 1988; pp 501-506. (10) Fo¨rstner, U.; Ahlf, W.; Calmano, W. Mar. Chem. 1989, 28, 145158. (11) Di Toro, D. M.; Mahony, J. D.; Hansen, D. J.; et al. Environ. Sci. Technol. 1992, 26, 96-101. (12) Berry, W. J.; Hansen, D. J.; Marhony, J. D.; et al. Environ. Toxicol. Chem. 1996, 15, 2067-2079. (13) Hansen, D. J.; Berry, W. J.; Mahony, J. D.; et al. Environ. Toxicol. Chem. 1996, 15, 2080-2094. (14) Allen, H. E.; Fu, G.; Deng, B. Environ. Toxicol. Chem. 1993, 12, 1441-1453. (15) Canfield, D. E.; Raiswell, R.; Westrich, J. T.; et al. Chem. Geol. 1986, 54, 149-156. (16) Di Toro, D. M.; Mahony, J. D.; Hansen, D. J.; Berry, W. J. Environ. Toxicol. Chem. 1996, 15, 2168-2186. (17) Zhang, J.-Z.; Millero, F. J. In Environmental Geochemistry of Sulfide Oxidation; Alpers, C. N., Blowes, D. W., Eds.; American Chemical Society: Washington, DC, 1994; pp 393-409. (18) Yao, W.; Millero, F. J. In Geochemical Transformations of Sedimentary Sulfur; Vairavamurphy, M. A., Schoonen, M. A. A., Eds.; ACS Symposium Series 612; American Chemical Society: Washington, DC, 1995; pp 260-279. (19) Dutrizac, J. E. Hydrometallurgy 1990, 23, 153-176. (20) Di Toro, D. M.; Mahony, J. D.; Gonzalez, A. M. Environ. Toxicol. Chem. 1996, 15, 2156-2167. (21) Davison, W.; De Vitre, R. In Environmental Particles Vol. 1; Buffle, J., van Leeuwen, H. P., Eds.; Lewis Publichsers: London, 1992; pp 315-355. (22) Zhuang, Y.; Allen, H. E.; Fu, G. Environ. Toxicol. Chem. 1994, 13, 717-724. (23) Luther, G. W., III; Church, T. M. In Sulfur Cycling on the Continents; Howarth, R. W., Stewart, J. W. B., Ivanov, M. V., Eds.; Wiley: Chichester, 1992; pp 125-142. (24) Suzuki, I.; Chan, C. W.; Takeuchi, T. L. In Environmental Geochemistry of Sulfide Oxidation; Alpers, C. N., Blowes, D. W., Eds.; American Chemical Society: Washington, DC, 1994; Chapter 5, pp 60-67. (25) Petersen, W.; Willer, E.; Willamowski, C. Water Air Soil Pollut. 1997, 99, 515-522. (26) Ankley, G. T.; Di Toro, D. M.; Hansen, D. J.; Berry, W. J. Environ. Toxicol. Chem. 1996, 15, 2056-2066. (27) Slotton, D. G.; Reuter, J. E. Mar. Freshwater Res. 1995, 46, 257265. (28) Besser, J. M.; Ingersoll, C. G.; Giesy, J. P. Environ. Toxicol. Chem. 1996, 15, 286-293. (29) Simpson, S. L.; Apte, S. C.; Batley, G. E. SETAC NEWS 1997, July, 18-19. (30) Peterson, G. S.; Ankley, G. T.; Leonard, E. N. Environ. Toxicol. Chem. 1996, 15, 2147-2155.
Received for review July 1, 1997. Revised manuscript received November 14, 1997. Accepted November 20, 1997. ES970568G
VOL. 32, NO. 5, 1998 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
625
