WRS Rapport Metalen Hoofdstuk 9

9. Sorption kinetics of dissolved zinc and cadmium on harbor sediment suspended in oxic seawater; a laboratory simulation.

Abstract

The results of sorption experiments of trace metals in suspensions of harbour sediment with seawater are reported. Anoxic sediment from the Rotterdam harbour was suspended in oxic seawater to simulate the sorption kinetics of Cd and Zn when harbour sediment is dumped at sea. During the simulated suspension the dissolved concentrations of Zn, Cd and major redox-sensitive element (Fe, Mn) were measured with fast separation techniques. The redox potential (Eh) measurement showed that oxidation of the anoxic sediment suspension can be very fast when oxygen was added. This redox change was also reflected in a rapid decrease of dissolved Fe, P and Mn in the sediment slurry. The dissolved Cd concentration showed a rapid decrease (0-10 minutes) during suspension. This was attributed to: (1) sorption onto sulphides, (2) coprecipitation with Fe- oxyhydroxides and/or (3) adsorption onto Fe-hydroxides. After 1 day Cd was released from the suspended matter which could only be attributed to the oxidation of metal-sulphides. The dissolved concentration of Zn in suspension showed similar initial adsorption, but Zn showed no release after several days. These experiments show the importance of scavenging when anoxic harbour sediments, contaminated with heavy metals, are dumped in oxic seawater.

9.1. Introduction

The port of Rotterdam (The Netherlands) is situated in the upper estuary of the river Rhine and Meuse. Suspended material carried by the rivers is deposited in this harbour area. To maintain access to the harbours by bulk carriers and supertankers the channels have to be dredged. About 23.106 m3 harbour sludge have to be removed each year. This sludge is contaminated with organic and inorganic pollutants and the major part is deposited at a designated area in the North Sea, approximately 10 km off the coast of The Hague. The liberation of pollutants may affect the quality of marine life at the dumping site or in the coastal zone when fine particles are carried with the water (VISSER et al., 1991). These problems have been addressed in studies by GUSTAFSON, 1971, HALL, 1989, STERN & GRANT, 1981, SNITZ et al., 1979, HIRST & ASTON, 1983, TRAMONTANO & BOHLEN, 1984 and PRAUSE et al., 1985. Speciation under anoxic environment. Sediments deposited in the harbour have a high organic carbon content. Organic carbon in the sediment will be degraded. A first step in the degradation sequence is the consumption of oxygen (STUMM & MORGAN, 1981). Oxygen is consumed rapidly in the toplayer of the sediment, the underlying sediment will become oxygen depleted. Changes between oxic and anoxic environment in sediment cores occurs within the range of millimetres (DAVISON et al., 1991, DAVISON et al., 1994). The succession of reactions in the sediment is mainly reflected in the vertical distribution of components (WILSON ET AL., 1986).
When anoxic harbour sludge is dumped into oxygenated seawater the redox conditions will change. In general the behaviour of heavy metals in aquatic systems is primarily controlled by Eh (redox potential) and the pH (DAVIS & LECKIE, 1978, BENJAMIN & LECKIE, 1981, CHAROENCHAMRATCHEEP et al., 1987, CALMANO, 1993). The sorption behaviour of heavy metals depends on liquid speciation and on available sorption sites. Usually trace metals are for a major part adsorbed on active surface sites of inorganic and organic particles. Under anoxic conditions sulphides can effectively bind Zn and Cd (LU & CHEN, 1977). The reaction constants show that Cd is more effectively bound by sulphide than Zn (STUMM & MORGAN, 1981):

In pore water studies of anoxic sediment the dissolved Zn and Cd concentration is almost zero which indicates that effective adsorption sites are formed with a high affinity for heavy metals (DAVIS-COLLEY et al., 1985, GIBLIN et al., 1986, HUERTA-DIAZ & MORSE, 1992). WALLMANN (1990) calculated for pore water systems that the concentration of non-sulphidic complexes and minerals were not important for heavy metal binding compared to sulphidic species. In marine and estuarine sediments sulphides are not limited because seawater contains sulphate which will diffuse into the sediment where it can be reduced to sulphides. In anoxic sediments the redox-sensitive elements like Fe and Mn will react to the thermodynamically most favourable state, Fe(III)-species will change to Fe(II) and Mn(IV) to Mn(II) species. In estuarine sediments the organic carbon content is high enough to maintain anoxic conditions.
Speciation under oxic conditions. In an oxic environment the speciation is different compared to the speciation under anoxic conditions. In oxic seawater at pH of 7.8 - 8.2 cadmium chlorocomplexes are dominant species of dissolved Cd for zinc the free ion and Zn-hydroxides are dominant (LU & CHEN, 1977, COMANS & VAN DIJK, 1988, HEGEMAN et al., 1992). Trace metals will adsorb to specific surface metaloxides. With constant Eh and pH values almost linear adsorption isotherms between the solid and liquid trace metal concentrations can be expected (MOREL, 1983). Solid sulphide and dissolved sulphide species are not stable, and will be oxidized in oxic seawater.
Suspension of anoxic sediments. When anoxic harbour sediment is dumped into the oxic water the chemical equilibria are disturbed. Divalent Fe in anoxic sediment will be oxidized to trivalent Fe in oxic seawater and precipitated as Fe(OH)3 (STUMM & MORGAN, 1981, ROEKENS & VAN GRIEKEN, 1983). Fe(III)- hydroxide has a high affinity for Cd and Zn. These freshly formed amorph Fe-hydroxides have an enormous sorption capacity (JENNNE, 1968) compared to the crystallized or aged ironhydroxides (FULLER et al., 1993). In natural systems there is a complete range of different Fe-precipitates (KUMA et al., 1991) and the iron oxyhydroxides will age once they are precipitated (BENJAMIN & LECKIE, 1981). The cations of Cd and Zn can also coprecipitate with the Fe-hydroxides. Sorption and coprecipitation will decrease the dissolved Cd and Zn concentration. (SHIGEMATSU et al., 1975, FRANCIS & DODGE, 1990, NYFFELER et al., 1986). BOURG (1988) stated that in the transition of anoxic to oxic conditions the following consecutive reactions can be distinguished; (1) oxidation of metal sulphides, followed by (2) the formation under oxic conditions of new species such as dissolved chlorocomplexes in marine and estuarine environments. Sequential extraction studies of sediments showed that heavy metals associated with the reduced phase (sulphides) in anoxic conditions were transformed to the carbonaceous phase upon oxidation of the sediment (STENEKER et al., 1987, RECKE & FÖRSTNER, 1986). The Fe(III)-formation can be retarded by the formation of Fe(II)- organic complexes where Fe(II) is an intermediate in the oxidation of organic carbon in the oxic water (STUMM & MORGAN, 1981). Whereas a re-oxidation of Mn(II) is accomplished by complex catalytic surface oxidation of Mn by ironoxides (SUNG & MORGAN, 1981, VAN DER WEIJDEN, 1975).
Toxicity. In sediments the toxicity is dependent on the availability of the metals. Mostly the free metal ion activity is correlated with biological effects. DI TORO et al., (1992) demonstrated the toxicity of cadmium when the ratio [SEM]/[AVS] > 1. AVS (Acid Volatile Sulphide) represents a sulphide fraction extracted with cold hydrochloric acid and SEM represents the Simultaneously Extracted Metal concentration. If the reactive pool of sulphide can no longer keep the trace metals from solution the metals will become available in solution. In general the key question is: will the specie become available for biotic life? These suspension experiments gave us some unexpected results in a non- equilibrium system.
Sorption kinetics of heavy metals from anoxic to oxic conditions are not understood completely. Therefore we investigated the kinetic behaviour of dissolved Zn and Cd during simulated dumping of anoxic harbour sediment in oxic water. Simple laboratory experiments with fast separation techniques showed the sorption processes during suspension of anoxic harbour sediment in oxic seawater.

9.2. Material and methods

Sediment: Harbour sediment was collected from the Rotterdam harbour area, located at the entrance of the Rhine river (Hook of Holland). With a large boxcorer undisturbed sediment samples could be obtained. The sediment was stored at 4 øC until use.
Seawater: North Atlantic seawater (Salinity: 35 g.kg-1 or þ), collected in as subsurface layer near the central Doggerbank area, was diluted (1:6) with Milli-Q water (Millipore) to obtain a salinity of 30 þ which is equal to the salinity at the coastal dumping site.
Material: All handling material was treated with 0.1 M HCl for several days and rinsed several times with Milli-Q water. In sorption experiments with trace metals PFA (PolyFluorAlkoxy polymer) vessels exhibited the least adsorption to the vessel wall. The filter set (polyethylene and polyacetate) was tested on adsorption of trace metals, no significant adsorption was measured for Zn and Cd.
Redox potential measurements: A suspension of harbour sediment was incubated in a glass three-neck round bottom flask for 4 weeks. The suspension was stirred with a magnetic stirrer in the dark. The flask was sealed with air-tight septum caps. A metal needle connected to a N2-flask which was pushed through the septum kept an overpressure of nitrogen above the suspension. Therefore oxygen could not penetrate the suspension and experimental handling through the other openings was possible. A Pt electrode with calomel reference electrode was pushed gently into the stirred suspension. The pH and the redox potential (Eh), calibrated with a Zobell solution, were recorded continuously during 350 minutes. At t=310 minutes the nitrogen purge was ended, the stopper was removed, and air could diffuse into the slurry. Thirty minutes later (t=340 minutes) we started to bubble air through the suspension in order to maximize the contact between oxygen and the suspension.
Suspension experiments: Undisturbed harbour sediment was used to simulate the suspension of harbour sediment at sea. Small subsamples of the sediment were taken with a half-open polyethylene 11 cm3 syringe. The open end of the syringe, with the piston downward, was placed gently on the sediment surface. The piston was held and the syringe was pushed into the sediment until the syringe was completely filled with the wet sludge. The syringe was pulled out of the sediment and the contents was immediately pushed into a well-stirred PFA vessel filled with 700 cm3 seawater (S=30). From this suspension subsamples were taken in order to determine the dissolved elemental concentration. The sampling scheme is illustrated in figure 1. It was essential to use a very quick sampling technique to obtain a high resolution in time. Therefore filtration with a high capacity filter (diameter: 47 mm, 0.45 æm, polyacetate, Schleicher & Schöll) was used. The filtrate was acidified to pH=1 by addition of suprapur HNO3 (Merck) and analyzed by ICP-AES (Induced Couple Plasma-Atomic Emission Spectroscopy) for dissolved Fe, Mn, and P. High resolution measurements of the trace elements Cd and Zn could not be made by ICP-AES due to low concentrations and matrix effects. The suspension concentration (m) was determined after the subsamples were taken. To determine m a subsample of the sediment suspension was washed three times with water, centrifuged, dried at 105 øC and weighed.
Sorption of heavy metals during suspension: Sorption of Cd and Zn during suspension was followed by the addition of the metals spiked with radionuclides in diluted seawater (S=30 promille at 20 øC. To avoid isotopic exchange reactions (ATKINSON et al., 1971), solutions with a high concentration of Zn (1.0 mg/L) and Cd (1.0 mg/L) were prepared in a PFA vessels (total volume: 700 cm3) and these were spiked with a high specific activity 65Zn or 109Cd radionuclide. Before the suspension experiment started, the gamma-activity of the solution was determined in fivefold using 1 mL aliquots. The anoxic harbour sediment was added in a similar fashion as described in the previous section and an identical filtration scheme was applied (fig. 1).

Figure 1. Schematic representation of the quick sampling technique. The experimental manipulations 1 - 3 were performed within 30 seconds.
(1) The suspended sediment is sucked through a small piece of tubing into the syringe. (2) The tube is quickly removed and replaced by a large filter holder (diameter: 47 mm) with Luer- Lock connection. (3) The piston presses the suspension through the filter until approximately 5 mL solution is collected in the sample vessel.

In parallel experiments pre-oxidized sediments were used. The same volume (11 cm3) of the anoxic sediment was suspended in a 50 mL seawater and clean air was bubbled through the suspension for 24 hours. This pre-oxidized sediment was added to the spiked seawater and the dissolved concentration of Zn and Cd was measured in time. The pH was measured continuously. Zn and Cd released from the sediment were less than 5 % of the added concentration.
The gamma-activities (A) of the samples, corrected for background activity, were determined by a multichannel counter (LKW-Wallac) with NaI-crystal. The channel window included the photopeak of the radionuclide. The fraction on the initial concentration (F) in the dissolved phase at time t was calculated according to the equation:

F(t) = A(t) * A(t=0)-1 .

9.3. Results and discussion

9.3.1. Redox potential during sediment oxidation

In fig. 2 the Eh is shown in a anoxic incubated sediment suspension.

Figure 2. Redox potential of incubated harbour sludge after implementation of a Pt electrode. The sediment slurry in the vessel was purged with nitrogen (t= 0-310 minutes), at t=310 nitrogen purging was inactivated and the stoppers were opened and at t=340 minutes aeration with pressurized air was started. Sediment/water= 1:10, pH = 8.5 ñ 0.1, Salinity = 30 þ.

The slow decrease of the redox potential from t=0 to 200 minutes is a result of a slow adaptation of the electrodes to the anoxic environment. When oxygen was allowed to contact the surface of the slurry at t= 310 minutes (fig. 2), the Eh increases almost immediately. The redox potential increases much faster when air is bubbled through the slurry. Although the measured Eh cannot be used to calculated the actual thermodynamic state of the species present in the suspension (MOREL, 1993), it is clear that reduced species can be oxidized quickly when oxygen is present. This experiment shows also that redox-related processes during the oxidation of reduced sediments can be very fast. Therefore, rapid separation techniques should be used to monitor dissolved trace metals during the oxidation process. Similar Eh changes can be expected when anoxic sediments are dumped in oxic seawater. The fast initial sorption processes can only be monitored when the time scale is within minutes. In our experiment the colour of the suspension changes from black (anoxic) to light-brown (oxic) within 24 hours. But the colour of a suspension did not represent the Eh during oxidation, the colour was still black when the redox potential was at maximum.

9.3.2. Sorption kinetics of redox sensitive elements

In fig. 3 dissolved Fe, Mn and P concentrations are shown after suspension of the harbour sludge in oxic water.

Figure 3. Dissolved concentrations of iron, manganese and phosphorous versus suspension time. The concentration is expressed as a fraction of the concentration at t = 0.5 minutes. At t=0 harbour sediment was added and kept in suspension. Initial concentration (t=0.5 min.): Fe 0.74 mg/L, P 2.1 mg/L and Mn 0.56 mg/L. Suspension concentration: 5 g/L, pH= 7.5 - 7.7, Salinity = 30 promille.

The redox sensitive elements (Fe, Mn) disappear within minutes, similar to the change of the redox potential in the suspension (fig. 2). It is obvious that we would have missed the rapid decrease in concentrations when we had not used a fast separation technique. Instead we would have observed a constant concentrations.
The quick decrease of dissolved Fe represents the precipitation of Fe-hydroxide. Dissolved phosphorous (ortho- phosphate) has the same pattern as dissolved iron because phosphate is strongly sorbed on reactive ironhydroxides. Freshly precipitated ironhydroxides have a high surface area and a high sorption capacity (ATKINSON et al., 1967, JENNE, 1968, BALISTRIERI & MURRAY, 1979). Once formed in the oxic water iron hydroxide is relative stable and the dissolution rate is low at pH values between 7.0 and 8.2 (KUMA et al., 1993, PATRICK & HENDERSON, 1981). The decrease of dissolved Mn can be attributed to the oxidation of Mn2+ and precipitation as MnOx (2ó x ó4). During the suspension sulphides, present in the anoxic estuarine sediment, can be oxidized and protons are produced. In our studies a decrease of the pH was not observed. The suspension has a high Acid Neutralization Capacity (ANC) because the Rotterdam harbour sediment has a high CaCO3 content and the pH will therefore remain constant. CALMANO et al., (1993) showed the importance of the Acid Neutralization Capacity (ANC) of suspended anoxic sediment under oxic conditions. A decrease of the pH from 8 to pH=3-4 will have major consequence for the behaviour of trace metals (DAVIS & LECKIE, 1978).
The observed decrease of the dissolved concentrations (fig 3.) with suspended harbour sediment has not been shown before with a resolution time of 0.5 minutes. The redox-sensitive elements decrease in the same time interval as the increase of the redox potential. Similar fast oxidation processes will occur when anoxic estuarine sediment is dumped in the oxic water or the anoxic bottom sediment is resuspended during storm conditions. Induced by redox processes new sorption sites of iron and manganese oxides are formed.

9.3.3. Trace element kinetics during suspension

Cadmium. Figure 4 shows the decline of the dissolved Cd concentration when anoxic and pre-oxidized sediment was suspended in oxic seawater.

Figure 4. Adsorption of cadmium expressed as the initial concentration of Cd in solution versus contact time in sediment-seawater suspension. Anoxic sediment ( , initial concentration: 1 mg/L, m = 3.2 g/L ) and oxic sediment (open square, initial Cd concentration: 1 mg/L, m= 4.2 g/L, pH= 7.9 ñ 0.1), Salinity = 30 promille.

The dissolved Cd concentration shows a rapid decrease when anoxic sediment was suspended. After one day partial release of Cd from the sediment to the dissolved phase is observed. The pre-oxidized sediment showed a different sorption behaviour. Although the sediment composition was identical the sorption kinetics depend on the oxidation state of the sediment. The sorption kinetics of the oxidized sediment was similar to the behaviour observed for Cd on suspended matter from the River Rhine (COMANS & VAN DIJK, 1988). A rapid initial adsorption within a few hours is followed by a slower adsorption during several days. The sorption on the anoxic sediment showed a very strong adsorption followed by a partial desorption after a few days. When the sorption experiment is followed for several days the anoxic sediment showed pseudo-equilibrium; a similar distribution as the pre-oxidized sediment. This is shown in fig. 4 when the suspension time was higher than 4 days. The observed kinetics in the anoxic sediment can be explained by; (1) adsorption to non-oxidized sulphides (2) coprecipitation with freshly formed iron(hydr)oxides and (3) adsorption on the iron(hydr)oxides. The solid sulphides species have a very high affinity for heavy metals (DAVIES-COLLEY et al., 1985, HUERTA- DIAZ & MORSE, 1992).
The oxidation of sulphides to sulphate occurs slower than the sorption of Cd on the sulphide-sites. With increasing contact time of the sulphide minerals and oxygenated water these minerals will be oxidized and the adsorbed trace metals released. The liberated Cd will then be adsorbed on other existing or newly formed sorption sites on the suspended matter. Freshly formed Fe and Mn oxyhydroxides have a high specific surface area and can adsorb a large amount of dissolved trace metals. With increasing amounts of host phases the dissolved trace-metal concentrations will decrease. The freshly formed Fe and Mn oxyhydroxides are not stable and will start to age. Therefore, adsorbed trace metals may become partly released.
Dissolved heavy metals can also be scavenged from solution by coprecipitation with Fe or Mn oxyhydroxides. In that case the heavy metals will be more tightly held by the substrate and will, upon aging, more readily become incorporated in the lattice of the host phase.
The experiments (Fig. 4) show that the initial sorption capacity of the anoxic suspended matter is higher than the capacity of the pre-oxidized suspended matter. After about 24 hours the capacities of the two suspensions start to become equal which is completed after about a week. We ascribed that to the above mentioned sequence of sorption on sulphide phases. Subsequent release and sorption on freshly formed Fe and Mn oxyhydroxides and further release upon aging of these host phases.
In the experiments with Cd, sulphidic species have the major adsorption capacity and scavenge the heavy metals from solution. When oxidized Cd will be release, the amount released is approximately equal the amount which was scavenged by sulphides. This can be conclude from the pre-oxidized sediment which has the same concentration when time was higher than 5 days. Zinc. Figure 5 shows the fast decrease in dissolved concentrations of Zn after addition of the suspended harbour sludge.

Figure 5. Adsorption of zinc expressed as the initial concentration of Zn in solution versus the contact time in sediment-seawater suspension. Anoxic sediment ( , initial concentration: 1 mg/L, m = 1.6 g/L ) and oxic sediment (open square, initial concentration: 1 mg/L, m= 0.92 g/L, pH= 8.0 ñ 0.2), Salinity = 30 promille.

A pseudo-equilibrium is reached in 15 minutes similar as for Cd. In the experiment where anoxic harbour sludge was used the sorption was higher than in the pre-oxidized suspension. Although the suspension concentration of the pre- oxidized suspension was denser (1.6 g/L) the sorption was less than in the experiments were anoxic added sediments (0.92 g/L). Therefore, the anoxic added suspension must have a higher sorption capacity. This indicates that the anoxic sediment had very reactive sorption sites with a high affinity for zinc. We attribute this affinity due to sulphides which were not immediately oxidized when anoxic sediment was dumped. In contrast to Cd, Zn was not released after 24 hour. This can be explained by a high affinity of Zn for ironhydroxides which are formed during oxidation of the sediment. A second reason is that Zn-sulphides have a very high oxidizing rate (LU & CHEN, 1977) and therefore will not be released easily. Zn has a higher affinity for sorption sites on solid ironoxides than Cd in seawater suspensions (VAN DER WEIJDEN, 1975). Suspended anoxic sediments have a high scavenging capacity of heavy metals. These results show that separation must be fast in order to show the kinetic effect of heavy metals during simulated in oxic sediment. Pseudo-equilibrium for the sorption of Cd and Zn on suspended sediment is attained within 15 minutes.
Anoxic harbour sediments which are dumped in seawater can immediately scavenge heavy metals from the dissolved phase. When harbour sediment is immediately redeposited on the bottom no harm will be done to the aquatic environment because anoxic conditions will be re-established and the sediment will be covered by an oxic blanket containing sorption sites in the form of ironhydroxides (THOMSON et al., 1993). If the suspended sediment is kept in suspension heavy metals can be partially released when the sulphide phase is oxidized. In the marine environment, the sediment suspension will be diluted and establish a new equilibrium with the seawater which is low in trace-metals. The heavy metals on the suspended sediment will be released continuously to the water phase where it becomes available to biotic life.

9.4. Conclusions

With simple laboratory experiments the sorption of cadmium and zinc was simulated when anoxic sediment is suspended in oxic seawater. Parallel studies of the redox potential, the sorption behaviour of Fe, Mn, Cd and Zn during suspension gave evidence for the chemical processes during dumping. The redox potential increased rapidly when oxygen could enter the anoxic suspension. The dissolved concentration of the redox sensitive elements Fe and Mn and of P decreased rapidly when anoxic sediment was suspended. Dissolved Cd and Zn were scavenged from solution when anoxic sediment was suspended. The scavenging of Cd was attributed to a combined interaction of: (1) sorption to non-oxidized sulphides, (2) (co)precipitation of Fe2+, (3) adsorption to freshly precipitated Fe(oxy)hydroxides. The sorption to mineral sulphides is considered as the main contribution of the rapid decrease in cadmium. For Zn strong adsorption was observed on anoxic sediment and this was attributed to adsorption to sulphides and precipitated ironhydroxides. If the anoxic sediments can settle in coastal water, Zn and Cd are captured by the sediment. These trace metals can be released to the water phase when the sediment is kept in suspension in the oxic seawater.

Registry No. Cd, 7440-43-9; Cd2+ ,22537-48-0, CdS, 1306-23-6; Zn, 7440-66-6, Zn2+, 23713-49-7, ZnS, 1314-98-3; Fe, 7439-89- 6, Fe2+, 15438-31-0, Fe3+, 20074-52-6, Mn, 7439-96-5, Mn2+, 16397-91-4, Mn4+, 19768-33-3, S2-, 18496-25-8.

9.5. REFERENCES

ATKINSON R., POSNER, A. & J. QUIRK, 1967. Adsorption of potential determining ions at the ferric oxide aqueous electrolyte surface. J. Phys. Chem. 71, 550-558.

ATKINSON, R.J., POSNER, A.M. & J.P. QUIRK, 1971. Kinetic of heterogeneous isotopic exchange reactions: derivation of an Elovich equation. Proc. R. Soc. Lond. A. 323, 247-256.

BALISTRIERI L. & J.W. MURRY, 1979. Surface of Goethite (àFeOOH) in seawater. In: E.A. Jenne: Chemical modelling in aqueous systems. ACS pp 275-298.

BENJAMIN, M.M. & J.O. LECKIE, 1981. Multiple-site adsorption of Cd, Cu, Zn and Pb on amorphous iron oxyhydroxide. J. Colloid Interface Sci. 79: 209-221.

BOURG A.C.M. 1988. Metals in aquatic and terrestrial systems: sorption, speciation, and mobilization. p3-32 In: Salomons W. and Förstner U.(eds) Chemistry and biology of solid waste. Dredged material and mine tailings. Springer-Verlag 1988 pp 305.

COMANS, R.N.J. & C.P.J. VAN DIJK, 1988. Role of complexation processes in cadmium mobilization during estuarine mixing. Nature, 336: 151-154.

CALMANO, W., HONG, J., U. FöRSTNER, 1993. Binding and mobilization of heavy metals in contaminated sediments affected by pH and redox potential. In: IAWQ, First International Specialized Conference on contaminated Aquatic Systems: Historical Records, Environmental Impact and Remediation. Milwaukee, Wisconsin, USA, June 14-16, 1993 pp 306-316.

CHAROENCHAMRATCHEEP, C., SMITH, C.J., SATAWATHANANONT, S., PATRICK, W.H., JR 1987. Reduction and oxidation of acid sulphate soils of Thailand. Soil Sci. Soc. Am. J. 51, 630-634.

DAVIES-COLLEY, R.J., NELSON, P.O., WILLIAMSON, J., 1985. Sulphide control of cadmium and copper concentrations in anaerobic estuarine sediments. Mar. Chem. 16, 173-186.

DAVIS, J.A., LECKIE, J.E., 1978. Surface ionization and complexation at the oxide/water interface. II. Surface properties of amorphous iron oxyhydroxide and adsorption of metal ions. J. Colloid Interface Sci. 67, 90-107.

DAVISON, W., GRIME, G.W., MORGAN, J.A.W., CLARKE, K. 1991. Distribution of dissolved iron in sediment pore water at submillimetre resolution. Nature 352, 323-324.

DAVISON, W., ZHANG, H., GRIME, G.W. Performance characteristics of gel probes used for measuring the chemistry of pore waters. Environ. Sci. Technol. 28, 1623-1632.

DI TORO, D.M., MAHONY, J.D., HANSEN D.J., SCOTT, K.J., CARLSON, A.R. & ANKLEY, G.T., 1992. Acid volatile sulphide predicts the acute toxicity of cadmium and nickel in sediments. Environ. Sci. Technol. 26, 96-101.

FEELY, R.A., FREFRY, J.H., MASSOTH, G.J., METZ, S., 1991. A comparison of the scavenging of phosphorus and arsenic from seawater by hydrothermal iron oxyhydroxides in the Atlantic and Pacific Oceans. Deep-Sea Res. 38, 617-623.

FRANCIS, A.J. DODGE .C.J. 1990. Anaerobic microbial remobilization of toxic metals coprecipitated with ironoxide. Environm. Sci. Technol. 24, 373-378

FULLER, C.C., DAVIS, J.A., WAYCHUNAS, G.A., 1993. Surface chemistry of ferrihydrate: Part 2. Kinetics of arsenate adsorption and coprecipitation. Geochim. Cosmochim. Acta. 57, 2271-2282.

GIBLIN, A.E., LUTHER, G.W. & VALIELA, I., 1986. Trace metal solubility in salt marsh sediments contaminated with sewage sludge. Estuar. Coast. Shelf Sci., 23, 477-498.

GUSTAFSON, J.F., 1972. Beneficial effects of dredging turbidity. World Dredging Mar. Const. 9, 44-72.

HALL, L.A. 1989. The effects of dredging and reclamation on metal levels in water and sediments from an estuarine environment off Trinidad, West Indies. Environ. Poll. 56, 189-207.

HEGEMAN, W.J.M., C.H. VAN DER WEIJDEN, ZWOLSMAN, J.J.G., 1992 Sorption of zinc on suspended particles along a salinity gradient: a laboratory study using illite and suspended matter from the river Rhine. Neth. J. Sea Res. 26, 285-292.

HIRST J.M. & S.R. ASTON, 1983. Behaviour of copper, zinc, iron and manganese during experimental resuspension and reoxidation of polluted anoxic sediments. Estuarine, Coastal and Shelf Science 16, 549-558.

HUERTA-DIAZ, M.A. & J.W. MORSE, 1992. Pyritization of trace metals in anoxic marine sediments. Geochim. Cosmochim. Act. 56, 2681-2702.

JENNE, E.A., 1968. Controls on Mn, Fe, Co, Ni, Cu and Zn concentration in soils and water: The significant role of hydrous Mn and Fe oxides. Am. Chem. Soc. Symp. Ser. 93: 337- 387

KUMA, K., SUZUKI, Y., MATSUNAGA, K., (1993). Solubility and dissolution rate of colloidal gamma-FeOOH in seawater. Wat. Res. 27, 651-657.

LU, C.S.J. & K.Y. CHEN, 1977. Migration of trace metals in interfaces of seawater and polluted surficial sediments. Environ. Sci. Technol. 11 1977 174-182

MOREL, F.M.M. 1983. Principles of aquatic Chemistry. John Wiley & Sons, New York, pp 446.

NYFFELER, .U.P., SANTSCHI .P.H.LI .Y.H. 1986. The relevance of scavenging kinetics to modeling of sediment-water interactions in natural waters. Limnol. Oceanogr. 31, 277-292.

PATRICK W.H. JR, WILLIAMS B.G. & J.T. MORAGHAN, 1973. A simple system for controlling redox potential and pH in soil suspensions. Soil Sci. Soc. Amer. Proc. 37, 331-332.

PATRICK W.H. JR, & KHALID R.A., 1974. Phosphate release and sorption by soils and sediments: effect of aerobic and anaerobic conditions. Science, 186, 53-55.

PATRICK, W.H. & R.E. HENDERSON, 1981. Reduction and reoxidation cycles of manganese and iron in flooded soil and in water solution. Soil. Sci. Soc. Am. J., 45, 855-859.

PRAUSE B., REHM E., SCHULZ-BALDES M., 1985. The remobilization of Pb and Cd from contaminated dredge spoil after dumping in the marine environment. Environmental Technology Letters. 6, 261-266.

RECKE, M., & U. FÖRSTNER, 1986. Oxidation increases mobility of toxic metals in dredged sludges. Chem. Environ., Proc. Int. Conf.London, Selper, 809-818.

ROEKENS, .E.J., VAN GRIEKEN .R.E. 1983. Kinetics of iron(II)oxidation in seawater of various pH. Mar. Chem. 13, 195-202.

SHIGEMATSU T., OMORI T., AOKI T. & M. MATSUI 1975. Coprecipitation behavior of zinc in the oxygenation process of ferrous iron. Bull. Int. Chem. Res., Kyoto Univ. 53, 435-433.

SNITZ, F.L., WEBER, W.J. JR., BARNEY, J.L., POSNER, J.C., 1979. Effect of calcium and sediment concentrations on the release of metals and nutrients from dredge spoil dispersions. Aquat. Toxicol. Boek, 322-341.

STENEKER R.C.H., VAN DER SLOOT H.A., DAS H.A. 1987, Leaching studies on dredged material in oxidized and reduced stat. Report: ECN-87-092 Netherlands Energy Research Foundation, P.O. BOX 1, 1755 ZG Petten, NL.

STERN, D.A. & GRANT .C.L. 1981. A laboratory investigation of heavy metal adsorption on man dredge spills. Bull.Environ. Contamin. Toxicol. 26, 213-218.

TRAMONTANO, J.M. & W.F. BOHLEN, 1984. The nutrient and trace metal geochemistry of a dredge plume. Estuarine, Coastal an Shelf Science. 18, 385-401.

STUMM. W. & J.J. MORGAN, 1981. Oxidation and Reduction In: STUMM, W. & J.J. MORGAN. Aquatic Chemistry. John Wiley & Sons, Inc., New York: p 418-503 Chpt. 7.

SUNG, W. & MORGAN .J.J., 1981. Oxidation removal of MnII from solution catalysed by the gamma-FeOOH(dipidocrocite)surface. Geochim. Cosmochim. Acta 45, 2377-2383.

THOMSON, J., HIGGS, N.C., CROUDACE, W., COLLEY, S. & D.J. HYDES, 1993. Redox zonation of elements at an oxic/post-oxic boundary in deep-sea sediments. Geochim. Cosmochim. Acta 57, 579-595.

TURNER, D.R., WHITFIELD, M. & A.G. DICKSON, 1981. The equilibrium speciation of dissolved components in freshwater and seawater at 25oC and 1 atm pressure. Geochim. Cosmochim. Acta, 45: 855-881.

VAN DER WEIJDEN, C.H. 1975. Sorption experiments relevant to the geochemistry of Manganese nodules. Ph.D Thesis, University of Utrecht. 154 pp.

VISSER, M., DE RUIJTER, W.P.M., POSTMA, L. (1991). The distribution of suspended matter in the Dutch coastal zone. Neth. J. Sea Res. 27, 127-143.

WALLMANN K. 1990. Die Frühdiagenes und ihr Einfluss auf die Mobilität der Spurenelemente As, Cd, Co, Cu, Ni, Pb und Zn in Sediment- und Schwebstoff-Suspensionen. Ph.D Thesis, Hamburg.

WILSON, T.S.R., THOMSON, J., HYDES, D.J., COLLEY, S., CULKIN, F., SORENSEN, J., 1986. Oxidation fronts in pelagic sediments: diagenetic formation of metal-rich layers. Science 232, 972- 975.

Titel rapport: