Sedimentwater systems

A multitude of speciation methods is now available to determine typical metal species, the relative distribution of metals among various sediment fractions, or the kind and kinetics of transformation between different metal forms. In particular, sequential chemical extractions (like the 'Tessier Method') provide valuable information about the strength of binding between a metal and main sediment fractions, and hence on its mobility and bioavailability. However, these approaches do not measure discrete metal species but, at best, a group of species, or a particular metal-binding sediment fraction defined by the used extraction. For this reason, bioavailability of sediment-associated metals is not directly related to the prevailing pattern of operationally defined metal binding forms, in contrast to water-borne metal species. Other approaches, as a complement to chemical extraction, include the measurement of sediment characteristics, like its capacity to buffer protons and electrons (e. g. sulphide-binding, acid neutralizing capacity, or redox potential), thermodynamic and kinetic solubility calculations, and more recently spectroscopic analysis or the determination of the AVS (acid-volatile sulphides). General overview

As pointed out by Reuther (1999) in his review on trace metal speciation in aquatic sediments, the concept of chemical speciation has been established as a practical tool during the past 50 years, to assess the behaviour of trace metals in both natural and polluted sediment/water systems. A multitude of more or less sophisticated analytical procedures is now available and can be applied in an attempt to determine typical metal species (such as the form of binding between the metal and a solid phase), the relative distribution of the metal among the various, most common, species, and the kinds and kinetics of transformation between different, typical species.

Classical chemical extraction is still today the most applied method to obtain information about the speciation, mobility, bioavailability or potential toxicity of a metal. It originated from leaching procedures used first in soil science to assess the plant-available portion of trace nutrients (Jackson, 1958). Similar methods have also been developed by exploration geochemists, including a wide range of extracting agents to trace ore mineralization sites in surface soils or sediments. Later, environmental geochemists adapted the methodology to elucidate current and past metal contamination events in fluvial and lake sediments, and to get an estimate of the mobilization potential (Forstner, 1995). With the introduction of sequential extraction procedures (Tessier et al. 1979; Tessier and Turner, 1996) and by improving their analytical selectivity and accuracy, successive leaching of metals from sediments made it possible to provide information about the relative strength of major sediment metal-solid associations. From this, it remained clear that there is no single extraction or analytical scheme that is good for trace metal speciation in general. Instead, any approach for speciation of metals in sediments must be chosen in relation to a carefully defined problem.

However, many of the currently used chemical extraction techniques are based on the so-called "Tessier method" (Tessier et al., 1979, see Table 5.5) and differ only in practical details, such as sample-weight / extractant-volume ratio, extraction time, washing between individual extraction steps, etc.

Table 5.5. Examples of sequential extraction procedures for trace metals in sediments (from Reuther, 1999).

Fraction Moore et al., 1988 Tessier et al., 1979

Exchangeable ions 1 M NaOAc, pH 8.2

Carbonate-bound metals 1 M NH2OH HCl in

Easily reducible phase 0.25 M NH2OH HCl 0.1 M NH2OH HCl

Moderatelyreduciblephase 0.02 M (NH4)2C2O4 in

Organic / sulphidic phase 0.1 M Na4P2O7 30% H2O2, pH 2, HNO3

(organic material) reextr. in NH4OAc, pH 2

2x KClO4 + HCl (resist. sulphides)

Residual, mineral phase HNO3 / HClO4 / HF aqua regia

As discussed by Reuther (1999), there is a plethora of methods available today for the physico-chemical speciation of metals in sediments, each of them adapted to a specific problem. These include a variety of more or less sophisticated approaches, from direct measurements using ion-selective electrodes or expensive equipment, like electron microprobe or scanning electron microscope coupled to an energy-dispersive system, over indirect physico-chemical methods to biological methods involving different kinds of laboratory tests or bioassays (see also sections 5.4.4, 5.4.5 and chapter 7 below). Total concentration approach

Among the indirect physico-chemical methods, perhaps the most widely used is the destruction of the sediment matrix by strong acid digestion, followed by determining the "total metal content" in the final solution by means of spectrophotometry analysis, like AES, AAS, or ICP-

MS. However, these total sediment metal estimates do not really reflect the totality of the metals present and it is supposed that only the harshest digestion method is most reliable for assessing residual or true total metal concentrations in sediments (and soils). Obtained total concentrations are usually normalised to sediment dry weight (or in some few cases to the content of sediment organic matter). However, it is usually not well defined how large a fraction of the total amount of a certain metal is really extracted and available for analysis, and how the size of this fraction may differ from one metal to another. Moreover, and most relevant in this context, these methods do not take into account what kinds of metal species occur in the sediment and, consequently, no information is obtained about the potential future behaviour of the metal, such as its mobility or its chemical and biological activity (toxicity). Partly theoretical approaches to metal speciation

In order to get a preliminary idea of the behaviour of a metal in a specific sediment, it is possible to use various measures of the sediment's capacity to buffer protons and electrons, like its sulphide-binding capacity, acid neutralizing capacity, or redox potential. The theoretical background as well as the measuring methods are summarised by Reuther (1999). In this context it should be mentioned that the rather new AVS (AcidVolatileSulphide) concept has attracted particular interest, especially in the US, because of its documented ability to assess the bio-available fraction of certain metals in anoxic sediments. The benefits and drawbacks in using the AVS concept, as well as its theoretical justification, will be further discussed in later sections of this report (see especially section 5.4.3 below; and Brumbaugh et al., 1994).

Thermodynamic calculations of the solubility behaviour of principal solid-metal phases based on pH-Eh diagrams may also contribute to the understanding of how a specific sediment type will "handle" a certain concentration of a metal under well defined conditions. For example, the distribution of metals between various complexing inorganic and organic ligands in sediment pore water can be calculated based on the knowledge of the complex-forming stability constants, the ligand concentrations, the pH and the ionic strength of the pore water. However, the distribution of metals among complexes formed with various ligands involves equilibrium as well as kinetic factors, which means that calculated data do not always correspond with experimental or field measurements. Chemical equilibrium models used to predict metal solubility of solutions in equilibrium with pure minerals do not account for the kinetics of precipitation-dissolution of solids or for the slow reaction rate of certain trace metal ligand exchanges. Therefore, in many occasions, it is today considered more efficient to make direct determinations of some of the critical parameters, such as the concentration offree metal ions in the porewater, e. g. by using ion-selective electrodes, anodic stripping voltametry or by separating the low-molecular solutes by means of dialysis prior to chemical analysis. More empirical approaches to metal speciation (chemical extraction)

Sequential chemical extraction of trace metals from sediments is intended - in each specific step - to release metals associated with a specific sediment phase (for more details see section 5.4.4 below). As reactions between solute metals and particle surfaces are heterogeneous (i.e. adsorption, electron and proton transfers), the extraction efficiency is more kinetically controlled and strongly depends on the availability of specific surface areas and the type of reactive sites (i.e. high- and low-energy sites). In general, results of the chemical extraction depend on (i) the extraction time, (ii) the liquid/solid ratio, and (iii) effects based on the pH and buffer/carbonate status of the sample (Reuther, 1999). Faced with the selection of a chemical extraction procedure, a great variety of options have been described, covering various leaching tests, single-step extraction, using different types of buffer solutions, chelating agents or weak acids, and multistep sequential extraction.

Single-step extraction is mostly used to estimate the availability and potential uptake of trace elements by plants. However, there is no clear relationship between extractability (by any of the proposed buffers or weak acids) and plant availability of trace metals (Reuther, 1999). These methods may, at best, provide some relative data or indication on the potential mobility or availability of metals when comparing different sediments or soils.

Similar to single-step extractions, sequential extraction procedures (SEPs) measure no discrete or stoichiometrically defined metal species, but rather a group of species, or a fraction of the total metal content as defined by the method used (operationally defined). The methods usually identify 37 different metal fractions, which are dissolved by the use of progressively stronger reagents, and by minimising overlaps between the reagents. The main binding forms and mechanisms - and thus, the metal fractions in a sediment or soil that can be separated by sequential extraction - can be described in the following, very simplified way (after Jenne & Luoma, 1977; Salomons & Forstner, 1980):

1. Physical sorption of cations in surface atomic layers and pores, e.g. by relatively weak 'van der Waals' forces.

2. Chemical sorption of ions or molecules with surface ligands of Fe or Mn oxy-hydrates, carbonates, sulphides and phosphates, or by hydrolytic adsorption.

3. Ion exchange by compensation of charges in the mineral lattice with exchangeable cations (e.g. of negatively charged OH-groups in clays and Fe hydroxides, or carboxyl and phenolic groups in organic substances).

4. Precipitation of dissolved compounds.

5. Complexation with dissolved or solid organic matter.

6. Fixation in inert lattice positions of minerals.

Many of the attempts to determine metal speciation in aquatic sediments have as an ultimate goal to assess the bioavailability of the metal under the prevailing conditions or under conditions that are likely to occur in a near future. However, the troublesome fact is that there is no simple or direct relationship between the distribution of particular metal species in a sediment and its bioavailability. While the bioavailability of waterborne metals may be directly related to the prevailing pattern of species for this metal, the same is not true for sediment-associated metals (Reuther, 1999). However, in spite of this general difficulty, some useful "rules of thumb" may be noted. For example, "acid-volatile sulphide" (AVS) is still considered to be a key binding phase for metals in anaerobic marine and freshwater sediments, controlling pore water concentrations and the bioavailability of certain metals to benthic organisms.

High concentration of metals in the pore water often correlates with a low AVS content in the sediment (Brumbaugh et al., 1994). It has been suggested that saturation of the sediment's binding capacity is indicated at an SEM/AVS ratio >1, a situation which may result in a flux of trace metals into the pore water. However, the sediment-binding capacity for metals can be greater than indicated by AVS, suggesting the presence of other important binding sites, such as Fe/Mn oxides, organic matter, etc. In fact, there are also other labile sulphides present in many sediments, beside AVS. Three forms of labile sulphides with complex reaction equilibria were measured by Brouwer & Murphy (1995) in sediments: molecular hydrogen sulphide, AVS and heat-volatile sulphides. Thus, the exclusive use of the parameters AVS and SEM when characterising the bio-available fraction of trace metals in sediments does not seem to be sufficient for all types of sediment. In addition, it certainly is useful to determine total organic carbon in the sediment as well as the metal concentration in pore water, and - if feasible -to determine also the acid-producing potential (APP) and the acid-

consuming capacity (ACC) of the particular sediment (Reuther, 1999) (see section 5.4.3 below).

Nonetheless, there is a preponderance of evidence showing that the SEM/AVS model is applicable in dynamic, bioturbated and oxidising field conditions due to the enhanced stability of sulphide complexes of copper, cadmium, zinc, nickel and lead relative to the stability of the iron and manganese mono-sulphide complexes. FeS and MnS therefore act as a buffer for the oxidation of the other metal sulphides. When finally the less soluble metal sulphides are oxidised, freshly formed iron and manganese oxides together with the organic carbon coating on sediment particles will act as new reactive surfaces having a high affinity for free metal ions. As such, the concern of remobilisation under oxidised conditions is minimal. See also detailed comments on section 5.4.6 bioturbation, resuspension and bioirrigation.

A pertinent example was given by Zhuang et al (1994), who investigated the effect of aeration on cadmium bioavailability in sediments in a series of laboratory aeration experiments in batch reactors during periods of approximately one month. During aeration the concentrations of metals associated with AVS and with pyrite decreased. At the same time there were increases in the concentrations of hydrous iron and manganese oxides and these materials became increasingly more important in the binding of cadmium. Following the aeration more than 50 % of the cadmium was associated with the extractable iron and manganese components and approximately 2 % of the cadmium released during the oxidation of AVS entered into the liquid phase.

Although we know that different sediments, with a similar quantity of metals, behave toxicologically differently, due to different types of metal binding and sediment characteristics, chemical speciation alone may not be sufficient to assess or predict the biological effect potential of trace metals in sediments (Reuther, 1999). As indicated above, the toxicity of many divalent metals in anaerobic sediments, as an example, is controlled by variable amounts of AVS, but by mainly Fe and Mn oxides in aerobic sediments, in addition to organic substances.

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