Bioturbation resuspension and bioirrigation

Oxidation of sediment-associated sulphides during episodes of sediment resuspension, or by introducing oxygen-rich water into non-surficial sediment layers, e.g. by burrowing organisms (bioirrigation), is supposed to be a major mechanism of metal release from sediments to the overlying water. Sediment mixing by benthic animals (bioturbation) can additionally disrupt metal sulphide binding in sediments, while released metals may precipitate again with Fe or Mn hydroxides, or become complexed by organic matter. In fact, new experimental results show that iron monosulphides may act as a buffer for the initial oxidation of associated trace metal sulphides during exposure of anoxic sediments to oxic water, as in the case of bioturbation. New research suggests that biological kinetics (i.e. metal uptake and accumulation processes) are much faster than chemical kinetics (i.e. metal release by desorption from the sediment). Sediment-burrowing organisms tend to accumulate more metals than are initially available in the dissolved phase, indicating that additional metals have to be delivered from various solid phases to the dissolved metal pool, probably due to sulphide oxidation by burrowing, desorption from surfaces or dissociation of precipitated complexes.

One assumption from what has been presented in section 5.4.4 is that sufide minerals may act as a permanent sink for many trace metals. However, Cooper and Morse (1996) (cited in Cooper and Morse, 1998) demonstrated that up to 90 % of estuarine pyrite can oxidize within 1 day of exposure to oxic water and release associated trace metals into the "reactive" HCl-soluble pool indicating that oxidative dissolution of sulphides (e. g. by bioturbation, storm resuspension, or dredging operations) could in contrast act as an important source for metal fluxes into overlying waters.

Bioturbation includes filter feeders injecting oxygenated water into deeper sediment layers, and surface dwellers, which regularly resuspend the top surface layer. The oxidation of sulphide phases in anoxic sediments due to resuspension or by introducing overlying oxic water at depths, e. g. by burrowing organisms (bioirrigation) is supposed to be a major source of metals released to the overlying water, but has not been considered yet as relevant for bioavailability and toxicity studies. Bioirrigation rates as high as 750 ml/h have been reported to occur in natural sediments (Foster-Smith (1978) cited in Simpson et al. 1998). In this context, it has been asked, if bioturbation can significantly disrupt metal sulphide binding in sediments. Or, if FeS and MnS can buffer the effects of bioirrigation at least for a certain time? Oxidized sulfur species (like SO4 and S0) and released metals may subsequently become quickly scavengened again by Fe or Mn hydroxides or complexed by organic ligands in the sediment.

To find an answer to these questions, Simpson et al. (1998) exposed model substances (metal sulfides) to aerated water to simulate the effect of short-term resuspension on solid metal speciation of contaminated anoxic estuarine sediments. As one result, FeS and MnS appeared as labile and rapidly oxidizable phases in contrast to CdS, CuS, PbS and ZnS model phases, which were kinetically stable over periods of several hours and, therefore, not equally affected by resuspension. The observed rapid decrease of AVS (from 119.7 to 20 ^mol/g) upon resuspension observed in natural sediments was consequently related to the oxidation of iron monosulphide, but was slower than observed for the FeS model compound. After prolonged resuspension, AVS dropped to significantly lower levels (10 ^mol/g) than SEM (ZCd, Cu, Ni, Pb, Zn dissolved in 1 N HCl) indicating that also part of the trace metal sulphides have been oxidized during ongoing resuspension. Also total reducible sulphide (TRS), which includes pyritic S, AVS, organic and colloidal S, decreased from 720 to 600 ^mol/g during an 8h-resuspension period, but did not give evidence that pyrite oxidation was significant (see also the study by Otero and Marcias 2002, below). Actually only SEM-Cu increased during an 8h-resuspension from originally 0.1 to about 2 ^mol/g, while all other metal-SEMs remained constant. But it was shown that the observed increasing SEM-Cu concentrations turned out to be an artefact of the used SEM/AVS procedure: during HCl-oxidation, Fe(OH)3 was produced in solution and dissolved again upon acidification to Fe3+, which subsequently oxidized CuS in the sediment. During resuspension, Fe2+ was released through the rapid dissolution of FeS to FeII and SO4. Subsequent oxidation of Fe(II) to Fe(III) occured in turn slowly (minutes to hours at pH 7-8). Summarized, it was shown that dissolved Fe(III) could act as an effective oxidant for Cu sulphide first after acidification of the sample in the SEM/AVS extraction step. Based on their experimental results, Simpson et al. (1998) concluded that in natural sediments, iron monosulphides may act as buffer for the initial oxidation of associated trace metal sulphides during exposure of anoxic sediments to oxic water, for example in the course of bioturbation.

The contradicting behaviour of natural sediment sulphides and model sulphides may point out to the possibility that a significant fraction of the trace metal sulphides are not present as AVS but as pyritic fractions in the sediment. Trace metal pyritization is known to be an important sink for metals in marine anoxic sediments. Morse (ref. given in Simpson et al. 1998) observed that oxidative release of metals from authigenic pyrite was greater than the extent of pyrite oxidation. As a matter of fact, metals adsorbed to the pyrite surface may be acid-soluble and so observed in the SEM-fraction, although no sulphide would be released. In this context, also metal binding on particulate organosulfur phases is still poorly understood. In as far microbial-catalysed oxidation may contribute to the observed discrepancy observed between AVS in natural sediments and model sulphides, remains to demonstrate.

For this reason, Simpson et al. (2000a) question the SEM/AVS theoretical assumptions that (1) only monosulphide phases are extracted and that for every mole sulphide a corresponding mole of metal is measured in the SEM fraction. And (2) the fact that discrete CuS and NiS proved hardly soluble in 1 M HCl renders only little theoretical basis for the interpretation of sediment toxicity from SEM/AVS data at least for these particular metals. The authors further argue that (3) the observation that Fe3+ oxidizes CuS during SEM extraction without a corresponding increase in AVS, which again underlines the rather conservative nature of the AVS approach in assessing the bioavailable part of sediment-bound metals. That stoichiometrically different Cu sulphides may be extracted to different extents seems to make SEM/AVS measurements even more difficult for a proper data interpretation. In fact, the observation that Cu2S may be more oxidized than CuS, can be explained by the stoichiometrically smaller amount of sulphide in Cu2S to be available for oxidation. Another source of error (according to the authors) when measuring AVS (with or without Fe) may be its oxidation to SO4 or S0 rather than due to its dissociation to form H2S, which is finally measured as AVS.

Recent laboratory resuspension experiments with defined metal sulphides showed that ZnS, PbS and CdS were resistant against oxidation in seawater (Simpson et al. 2000a). In contrast, AVS in highly anoxic and metal contaminated field sediments was rapidly oxidized with < 6 % remaining after 24 h, despite high particulate Zn and Pb contents. For this reason, Simpson et al. (2000a) concluded that these metals may occur largely (73-95 %) as non-sulphides. By subsequently adding soluble ionic Zn to the sediment, ZnS was quantitatively formed suggesting a rapid formation of sulphide coatings that protect particles against further sulfidization (see section 5.4.4). Table 5.9 shows that (beside Zn powder) added solid Zn and Pb phases dissolved completely during the AVS extraction with recoveries of 88-107%. Upon resuspension in seawater, the AVS concentration of sediments spiked with ZnCl2 and ZnCO3 reached a plateau corresponding to about 75% of the added solid phase, while other phases were even less (1031%). AVS/SEM analysis 1 and 3 months later did not show any consistent changes, suggesting that a significant portion of metals entering aquatic systems as particulates may become sulfidized for a long time (years). From this, it seems that a rapid formation of insoluble sulphide coatings on the particle surface may occur and protect them against further reaction (sulfidization).

Table 5.9. Reactivity of added solid phase metal constituents (M) in Iron Cove sedimenta and final metal sulphide formation upon sediment resuspension in seawater (from Simpson et al., 2000a)

Solid added

added M

recovered

AVS (|mol/g)

metal sulphide

(|lmol/g)

M (%)b

0 hc

24 h

formation (%)

nothing

0

233 (11)

11 (5)

Zn powder

70

69 (10)

231 (36)

21 (18)

15

ZnO

70

101 (7)

225 (20)

20 (9)

13

ZnCl2

70

100 (5)

242 (35)

64 (8)

76

ZnCO3

120

107 (8)

240 (17)

104 (7)

75

PbO

70

97 (10)

211 (19)

18 (4)

10

PbCl2

70

88 (8)

210 (41)

33 (6)

31

"Standard deviations are given in parentheses, bpercent of added solid-phase metal recovered as SEM (Zn or Pb), "sediment resuspension time

"Standard deviations are given in parentheses, bpercent of added solid-phase metal recovered as SEM (Zn or Pb), "sediment resuspension time

There is increasing evidence today saying that reactive metal sulphides may not play the role in the speciation of metals entering anoxic sediments, as suggested so far by the AVS concept. According to this approach, sediments with an excess of reactive sulphide will have very low (nontoxic) dissolved metal levels in the porewater, due to the low solubility of metal sulphides. The molar excess of acid volatile sulphide (AVS) (mainly FeS) over simultaneously extracted metals (SEM) is consequently used to predict the absence of metal bioavailability and toxicity to benthic biota. It is further supposed that AVS extraction by 1 M HCl (45 min) extracted H2S, amorphous FeS, mackinawite (FeS1-x), some greigite (Fe3S4) but not pyrite. In their experiments, Simpson et al. (2000a) could show that, although CdS, PbS and ZnS were completely extracted, dissolution of NiS and CuS was limited. Significant artefactual oxidation by released Fe3+ was supposed to occur during AVS extraction, most likely due to reactions following acid addition (see above). Unlike for CdS and ZnS, PbS dissolved only slowly during AVS extraction allowing time for Fe3+ to act as an oxidant, which may explain the lower recoveries of PbS in the resuspension experiment with model sulphides, in contrast to CdS and ZnS. Adding Zn2+ or ZnS to the sediment rapidly decreased AVS concentrations upon resuspension indicating that added Zn reacted with sediment AVS and that the formed ZnS was resistant to oxidation. Resuspending the sediment with oxic seawater resulted in AVS values well below SEM. Considering the rapid formation of metal sulphides and their resistance to oxidation (upon resuspension), the authors hypothesize for Pb and Zn that these metals were not present as discrete sulphides in the sediment. The possibility that metals entering sediments in a metallic form, or as oxides or salts, become sulfidized, was confirmed by the performed experiments showing that the formed sulphide coatings can prevent further sulfidization of these solid metal forms. The formation of surface coatings of Fe(II) sulphide on amorphous Fe (III) oxyhydroxide particles was already suggested in previous studies (see Cooper and Morse 1998, Simpson et al. 1998).

Vink (2002) developed a so-called 'Sediment Or Fauna Incubation Experimental system' (SOFIE) to analyse the chemical speciation of metals in porewaters of natural undisturbed sediments at the sediment-water interface, while simultaneously conducting exposure tests with test organisms causing significant bioturbation. It was shown that metals competing for binding opportunities between reactive sulphides and dissolved organic matter explain the observed high dissolved concentrations in sulfidic zones. However immediately after introducing a sediment dweller (here the oligochaete Limnodrilus), equilibrium concentrations at once changed (see Table 5.10). Limnodrilus lives in burrows and irrigates the sediment with overlying oxic surface water. The author observed pronounced effects after colonization of the model sediment by the oligochaete: the redox potential (Eh) rised significantly and pH increased by half a unit, sulphides became oxidized and phosphate released. Also, macronutrients are mobilized and taken up at the same time by benthic organisms for their growth and metabolism. But, these organisms also accumulate and eliminate metals. From this one may conclude that the newly established dynamic state of the sediment-water system confirms that biological kinetics (i. e. uptake and accumulation processes) are much faster than chemical kinetics (i. e. release by desorption from the sediment). A continued steady flux was observed only for As (not shown) and Ni from sediment to porewater. The obtained results in Table 5.10 show that relatively large fractions of free ion activities occur in the overlying oxic surface water. These free metal fractions may increase after some days when organisms are introduced and decline again after increasing uptake of metals by the test organisms. One has to keep in mind that high free metal ion activities occurring in the overlying oxic water do not necessarily mean high total concentrations.

Table 5.10. Free metal ion activities as percentage of total dissolved concentration at equilibrium, and 7 and 21 days after introducing test organisms (Limnodrillus spp.) (from Vink, 2002)

free metal activitiy (% of total dissolved concentration)

Metal

Metal

distance

steady-state

7 days after

21 day af

water-sediment

introducingLimnodrillus

introducing Li

(mm)

+5

29

48

<1

-5

<1

1

40

-10

<1

26

12

-15

<1

<1

29

-25

<1

nd*

17

-30

<1

<1

20

-35

<1

13

<1

-40

nd

<1

<1

+5

4

13

1

-5

8

5

<1

-10

4

10

<1

-15

<1

1

<1

-25

<1

nd

<1

-30

<1

<1

<1

-35

<1

<1

<1

-40

nd

<1

<1

+5

29

26

18

-5

1

23

5

-10

<1

8

1

-15

<1

19

<1

-25

<1

nd

<1

-30

<1

<1

<1

-35

<1

<1

<1

-40

nd

<1

<1

+5

1

34

1

-5

1

1

1

-10

<1

15

1

-15

<1

1

<1

-25

1

nd

<1

-30

5

1

<1

-35

1

2

<1

-40

nd

1

<1

+5

38

44

32

-5

55

35

28

-10

1

47

1

-15

36

1

1

-25

34

nd

5

-30

4

1

1

-35

5

34

1

-40

nd

1

1

*nd = not determined due to failure of the micro-seized ion-exchange columns

Mass calculations showed that Limnodrilus consumed considerably more metals than initially available in the dissolved phase: for Cd 18 times, for Cu 4.4, for Ni 1.3, for Pb 122 and for Zn 10 times more. From this it becomes evident that more metals have to be delivered from various solid phases to the dissolved metal pool, probably due to sulphide oxidation by burrowing, desorption from surfaces or dissociation of precipitated ligands. However body concentrations decreased again for all metals after 11 days (except for As). Statistical analysis of the data showed that the uptake of Cd, Cu, Ni and Zn was indeed related to the free metal ion activity in the overlying water. This was not surprising as burrowing organisms exchange the water in their burrows with overlying water, which was characterized by a higher free metal ion pool than the porewater.

In summary, the studies showed that burrowing organisms will significantly affect metal speciation. Introducing organisms into the sediment will result in the release of metals mainly due to the dissolution of Fe and Mn oxyhydroxides and oxidation of sulphide precipitates. Metals are simultaneously released and taken up again producing antagonistic processes, with rising body concentrations and a decline of the free metal ion activity. In contrast to Cd, Cu, Ni and Zn, As and Pb uptake was best related with total dissolved porewater concentrations. Constructing similar generic models that can transfer chemical information to a biological response should have high future research priority. In this context the FIAM concept, as described by Vink (2002), is a significant step forward (see the following section 5.4.7.1, and chapter 6).

That the resuspension of contaminated sediments may indeed constitute a potential source of toxic compounds, including trace metals, for aquatic ecosystems, was recently also demonstrated by Geffard et al. (2002). The authors examined the release, bioavailability, bioaccumulation and toxicity of PAH and trace metals in resuspended contaminated coastal sediments, and possible effects on the embryogenesis and larval growth of Crassostrea gigas. Test organisms were either directly exposed to elutriates or fed on algae (Isochrysis galbana) contaminated with these elutriates. Both toxic effects and bioaccumulation were more pronounced when the larvae were directly exposed to sediment elutriates. Based on these findings, the authors speculate that a fraction of the studied contaminants (metals and PAHs), adsorbed on and released during resuspension (extraction) from the sediment, was bioavailable and subsequently bioaccumulated by the test larvae.

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