Final Report: A Modeling and Experimental Investigation of Metal Release from Contaminated Sediments The Effects of Metal Sulfide Oxidation and ResuspensionEPA Grant Number: R825277
Title: A Modeling and Experimental Investigation of Metal Release from Contaminated Sediments The Effects of Metal Sulfide Oxidation and Resuspension
Investigators: Di Toro, Dominic M. , Mahony, John D.
Institution: Manhattan College
EPA Project Officer: Hahn, Intaek
Project Period: November 15, 1996 through November 14, 1999 (Extended to November 14, 2000)
Project Amount: $544,463
RFA: Risk-Based Decisions for Contaminated Sediments (1996) RFA Text | Recipients Lists
Research Category: Hazardous Waste/Remediation , Land and Waste Management
Objective:The objective of this research project was to construct and validate a mechanistically realistic model of the release of potentially toxic metals from contaminated sediments to the overlying water. These are cadmium, copper, nickel, lead, and zinc, all of which form metal sulfides more insoluble than iron sulfide. The model is intended to be used in conjunction with water column fate and transport models. The model is intended to be compatible with sediment quality criteria based on simultaneously extracted metals (SEM) and acid volatile sulfide (AVS) concentrations in sediments as well as interstitial water toxic units (Ankley, et al., 1996). The previously available model formulations for computing the metal flux from sediments are incomplete because they take no account of the critical and central importance of metal sulfide formation, dissolution, and oxidation. They are based on partition coefficients. Consequently, they provide only a rough approximation of what is actually controlling the release of metals from sediments. The purpose of this project is to remedy this situation by producing a model that calculates the flux of metals from the sediment to the overlying water. The model can be incorporated into presently available water quality models for metals.
Summary/Accomplishments (Outputs/Outcomes):The project has been successfully completed. A numerical one-dimensional transport and reactive sediment model has been developed to predict AVS and SEM profiles for sediments exposed to molecular oxygen through contact with an overlying water column. The flux to the overlying water also is computed. The model divides the sediment into multiple thin vertical layers in which the reactions take place. The model considers the vertical profiles of dissolved oxygen, particulate organic carbon, dissolved sulfide, iron sulfide, iron oxyhydroxide, iron(II), dissolved and sorbed metal (to particulate organic carbon and iron oxyhydroxide), and metal sulfides. The sediment mass balance equation takes the following form:
where [c(z)] is the concentration of chemical per unit bulk volume of the sediment as a function of the depth z, fp is the fraction of the chemical that is in the particulate from, fd is the fraction of the chemical in the dissolved form, Dp(z,t) is the diffusion coefficient for particulate phase mixing (bioturbation), Dd is the diffusion coefficient for aqueous phase mixing, and is the sediment porosity. The term represents the sum of all sources or sinks of the chemical that are first order with respect to the chemical concentration [c(z)]. The term represents the sum of all sources or sinks of the chemical that are zero order with respect to the chemical concentration.
Figure 1. Structure of the sediment flux model.
The structure of the model is presented in Figure 1. The features incorporated in the model are:
? The oxidation of particulate organic carbon (POC) in the aerobic layer using oxygen(a) and in the anaerobic layer using sulfate (c)
? The formation (d) and oxidation (b) of iron sulfide (FeS(s))
? The formation (g) and oxidation (e) of metal sulfides (MS(s)) as a source and sink of dissolved metals respectively, partitioning of metals (f) to iron oxyhydroxide (FeOOH) and particulate organic carbon
? Particulate (bioturbation) and diffusive (pore water) mixing.
Experiments were carried out to support the model development. The oxidation kinetics of the metal sulfides were determined for pure phase metal sulfides and from naturally contaminated sediments. A series of experiments was performed to measure the metal fluxes from spiked and intact field collected cores and to determine the vertical profiles of SEM and AVS.
The sediment metal flux model has been generalized from its predecessor (Di Toro, et al., 1996) in the following ways. The model computes metal fluxes to the overlying water and the flux computations have been compared to observations. The quantity of iron sulfide is now computed from the kinetics and solubility of iron sulfide rather than the empirical partition coefficient method used previously. The metals considered have been extended from only cadmium in the previous model to nickel, lead, and zinc.
The results for nickel, cadmium, and lead for spiked sediment experiments are shown in Figure 2. The results for the nickel (Figure 2a-c) display an interesting feature: the decrease in metal concentrations after 15 days. The model reproduces this behavior as a consequence of the creation of new sorption sites via the oxidation of iron(II) to iron oxyhydroxide (see reaction b in Figure 1 or Figure 1b). The vertical profile of AVS (Figure 2a) results from the oxidation of iron sulfide (Figure 1b) as oxygen penetrates from the overlying water to the sediment over a period of 150 days. At the start of the experiment, AVS is a constant value throughout the sediment column. As time elapses, sulfide is produced in the anaerobic portion of the sediment (Figure 1c,d), while sulfide is oxidized in the aerobic portion (Figure 1b). The depth of oxygen penetration is indicated by the decrease in sulfide at a depth of approximately 45 mm. At time zero, the SEM also is uniformly distributed as a function of depth. After 150 days, nickel SEM is increasing toward the sediment-water interface. As time elapses, oxygen penetrates the sediment core, and nickel sulfide is oxidized (Figure 1e). Once this oxidation takes place, the nickel is free to either sorb to solid phase organic carbon or iron oxyhydroxide (Figure 1g), or remain in the aqueous phase and diffuse to the overlying water, or to the anaerobic portion of the sediment. The majority of nickel is in the oxic portion of the core after the 150 day period of oxidation. This is a result of the affinity of nickel for iron oxyhydroxide.
The parameters for the iron system that were determined via calibration with the nickel data set are used in the calibration of the other data sets. Published data are available for the partition coefficients and sorption capacities to iron and particulate organic carbon. In the final calibration, these parameters are close to those values found in the literature.
For cadmium (Figure 2d-f), the vertical AVS profile shows less of an oxidative loss of sulfide as a function of depth (Figure 2d). This is due to a slower oxidation rate of cadmium sulfide. The model predicts a uniform profile of SEM in the anaerobic portion, and an increase in SEM at the oxic-anoxic interface (Figure 2e). The overlying water cadmium profile shows a gradual release of cadmium from the sediment over the first 20 days of the experiment. This is due to cadmium sulfide oxidation. The data then show a decrease and then an eventual plateau of cadmium in the overlying water. The flux of cadmium into the sediment is due to the production of iron sorption sites that accompanies iron(II) oxidation.
Figure 2. Experimental (symbols) and model calculated (lines) vertical profiles of AVS (A, D, G), SEM (B, E, H), and overlying water dissolved metal concentration as a function of time (C, F, I) for nickel (A-C), cadmium (D-F), and lead (G-I).
The leveling of the cadmium concentration indicates that cadmium sulfide oxidation has slowed as a result of dissolved oxygen mass transfer limitation, and that the cadmium resulting from cadmium sulfide oxidation has been trapped within the core by sorption or reformation of cadmium sulfide.
For the lead data, the slope of the oxidative front suggests that the oxidation rate of lead sulfide is slower than the value used in the calibration (Figure 2g). A faster oxidation rate was used to fit the initial release of metal to the overlying water (Figure 2i), and to agree with the experimental determined rate of oxidation. Because the slope of the oxidative front is controlled by the metal sulfide oxidation rate, fast oxidation rates yield flat slopes for the oxic interface. The model SEM profile has a slight increase of SEM above 25 mm. The experimentally determined profile for SEM is relatively constant in depth. This feature in the model predictions is due to lead sorption to iron oxyhdroxide.
The model calibration for the zinc spiked data set was performed concurrently with the calibration with the naturally contaminated data set (Figure 3). The naturally contaminated data set contained mostly iron sulfide, with some zinc sulfide present in the sediment upon collection. Suitable model calibrations were achieved for the vertical profile of AVS and overlying water zinc. The zinc overlying water is similar to that found for the nickel spiked data set. The high initial values of zinc concentration in the water column are due to equilibration of the overlying water with the sediment pore water. This resulted in a rapid flux of zinc to the overlying water.
Figure 3. Experimental (symbols) and model calculated (lines) vertical profiles of AVS (A, D), SEM (B, E), and overlying water dissolved metal concentration as a function of time (C, F) for zinc in spiked (A-C), and a field collected (D-F) sediment.
The model predicted decrease in zinc overlying water is due to the production of additional sorption sites by the oxidation of Fe(II) as described for the nickel spiked data set. The model calibration was able to reproduce the naturally contaminated sediment data, while using the same metal-specific parameters used for the zinc-spiked experiment. The major difference is in the diffusive transport rate.
As can be seen from the results shown in Figures 2 and 3, the model is capable of simulating the flux of metals to the overlying water from both spiked and naturally contaminated sediments. The parameters used for the simulations and the code for the computer program are contained in the final report of the project. Theses that report the results have been prepared and presented at conferences. Final publications are in progress.
Ankley GT, Di Toro DM, Hansen DJ, Berry WJ. Technical basis and proposal for deriving sediment quality criteria for metals. Environmental Toxicology and Chemistry 1996;15(12):2056-2066.
Di Toro DM, Mahony JD, Hansen DJ, Berry WJ. A model of the oxidation of iron and cadmium sulfide in sediments. Environmental Toxicology and Chemistry 1996;15(12):2168-2186.