Chemical speciation in tropical waters — A cautionary tale

The Science of the Total Environment, 58 (1986) 9-35 Elsevier Science Publishers B.V., Amsterdam - - Printed in The Netherlands

CHEMICAL SPECIATION CAUTIONARY TALE

IN TROPICAL

WATERS

9

-- A

M. WHITFIELD and D. R. TURNER

Marine Biological Association of the U.K., The Laboratory, Citadel Hill, Plymouth PL1 2PB (United Kingdom)

ABSTRACT Speciation studies in natural waters have revealed a number of shortcomings in the application of the traditional equilibrium approach. Conventionally most work has been concentrated at 25°C and 1 atmosphere pressure - - conditions typical of tropical surface waters. However, even where complete equilibrium can be assumed and only dissolved inorganic species are involved, considerable uncertainties remain in many of the stability constants for complexation, hydrolysis and protonation equilibria. Despite such difficulties the tendency has been for a proliferation of modelling excercises, with species distributions often quoted to a few tenths of a percent, instead of a systematic comparison of model predictions with experimental measurements. Realistic equilibrium calculations must also take account of the interactions of dissolved species with colloidal material and with (frequently uncharacterised) organic ligands. Problems concerned with distributions of complexing sites with varying properties must be recognised in assessing the importance of such equilibria. Finally, many chemical systems are pushed far from equilibrium by biological and photochemical processes so that account must be taken of the presence of metastable components and of the rate with which such species attain their equilibrium state. Examples are provided by the presence of the reduced forms of many elements in air-saturated waters. To overcome some of these problems, numerous 'operational' procedures have been developed which divide the total element concentrations into various fractions by techniques such as filtration, ion-exchange and UV irradiation. Since the total concentrations of many elements may not exceed 10-gM, such procedures are prone to experimental error. In addition, their relationship to the molecular concept of speciation is not always clear. A re-assessment is required of our approach to chemical speciation studies that will provide a much closer link between the processes we intend to study and the procedures we use for elucidating the chemistry of the elements in natural waters.

INTRODUCTION

Studies of topics as diverse as the solubility of minerals in sea water and the influence of trace metals on the viability of marine organisms indicate t hat the involvement of the chemical elements in geological and biological processes can n o t be interpreted simply in terms of their total concentrations. To provide useful, predictive descriptions of such processes we must subdivide the total element c o n c e n t r a t i o n s in a m a n n e r t h a t is both mechanistically meaningful and experimentally tractable. Since t here are potentially as many ways of car r y in g out such a subdivision as t he re are marine chemists interested in the 0048-9697/86/$03.50

© 1986 Elsevier Science Publishers B.V.

10 problem, it is necessary to agree on some simple definitions. The most useful approach is t h a t adopted by Burton (1979) who suggested three basic subdivisions: (i) Species are those entities which can be described in terms of a welldefined chemical stoichiometry (e.g. ions, molecules, complexes); (ii) A chemical form is defined as a chemically coherent group of chemical species (e.g. all inorganic complexes) or of less well-defined entities (e.g. colloids, particulate matter); (iii) A chemical fraction is a group of chemical forms resolved by, and thus operationally defined by, particular analytical procedures (e.g. dissolved fraction). Chemical fractions are not mutually exclusive and they specify the behaviour rather than the identity of the chemical species involved. Modelling studies of chemical speciation have been biased towards 25°C and 1 atmosphere pressure by the physical chemists' predilection for simple experiments and comfortable working conditions. Although this adherence to standard conditions often restricts the applicability of speciation models in temperate and high latitude waters, the physical chemist should find himself in familiar surroundings when discussing chemical processes in tropical waters where temperatures of 25°C are commonplace. Consequently, the intention of this paper is to discuss critically the various approaches to the estimation, interpretation and utilisation of data on chemical speciation in sea water. Although the discussion should strictly be confined to chemical species, the pressure of practical problems and the desire to provide a broader description of the state of constituents in natural waters have conspired to produce a rapidly growing body of literature describing various attempts to measure chemical forms and fractions in sea water (Florence and Batley, 1980; Florence, 1982a). As we shall see, measurements of chemical forms and chemical fractions are usually more readily accessible t h a n detailed speciation pictures. Consequently, we will also consider the problems of providing useful interpretations of these empirical measurements in terms t h a t are relevant to geologically and biologically important processes.

CALCULATIONSOF CHEMICALSPECIATION PATTERNS The most direct method for assessing chemical speciation in sea water is to calculate the equilibrium distribution of the various species using a thermodynamic model (Fig. 1). Such an approach has been used to good effect to elucidate the general rules relating chemical periodicity to the speciation of the elements in natural waters (Nieboer and Richardson, 1980; Turner et al., 1981). However, when we move from this broad canvas to consider in detail the speciation of particular elements the picture becomes less well-defined. A review of chemical speciation for a well studied element such as lead (Table 1) indicates wide differences in species distribution between the various models. Wide variations are also found even when the solution composition and

11

START Fix:- Interactions to be included, conc of o.lt I significo.nf components, T, P, pH, pE, I . ~

Select a complex stability constant o.vailab[e7

YES 1 conditional ~ _ ~ on other interactions?

~

No

published vo.[ues in agreement ?

YES I~ constan t at correct ionic strength7

f

NO

__•check

implicit1 assumptions j

]

use crifical NO :,It ompilotions or reassess original dato. f NO reliable activity coeffients avo.ilable ? j

YES

Iconstantq II requ red I

correct

consto.nt at correct T and P 7 YES $< NO

de fermine or use Lineo.r correlations

constant to T & P required

NO inferpoto.te available data or use o.n electrostatic equo.t on

I

have all complexes been consider ed ? solve equilibrium I equations for 0.[I complexes

E chec mo e,

I

against experimental I data

Fig. 1. Procedure for speciation calculations in

seawater.

environmental parameters are defined at the outset (Nordstrom et al., 1979; Whitfield and Turner, 1980; Zuehlke and Byrne, 1984). Undoubtedly, the greatest uncertainties arise in the selection and correction of the conditional stability constants (Fig. 1). The first stage in the protocol for selecting stability constants for the sea water model (Fig. 1) is to see what data are available in simple media. The

12

TABLE 1 Speciation models for lead in s e a w a t e r Species

Choridedominated models a

Pb 2÷ PbCl ~, PbC1 °, PbCla PbCO~, Pb(CO3 C1) Pb(CO3) ~ PbOH +, Pb(OH) ° Pb(OHC1) ° PbSO °

b

c

Carbonatedominated models

Carbonate/ hydroxide d o m i n a t e d models

d

f

e

g

1

5

9

1

2

2

2

98

74

68

28

25

26

21

18

68

67

40

47

5

1 2

6 1

32

31

2

19 1

(a) M o r e l and M o r g a n (1972); (b) Dyrssen and Wedborg (1975); (c) Long and A n g i n o (1977); (d) Florence and Batley (1976); (e) Whitfield and T u r n e r (1980); (f) S t u m m and B r a u n e r (1975); (g) Sipos et al. (1980a).

compilations of Sill~n and Martell (1964, 1971), which are most widely used, provide an uncritical listing of conditional stability constants determined in a variety of media over a range of ionic strengths using different experimental methods. Stability constants at infinite dilution are frequently presented, although details of the media and extrapolation procedures used must be obtained by reference to the original literature. The selection of appropriate stability constants from such compilations can be highly subjective, and in consequence collections of critically selected stability constants have been prepared (Smith and Martell, 1975, 1976; Martell and Smith, 1974, 1977, 1982). Although the use of such compilations will reduce the variability of speciation models by restricting the choice of stability constants, they must be approached with caution since space does not permit the compilers to indicate how the critical values have been selected, or in some instances recalculated, from the original literature. In addition, the media used in the experimental determinations are rarely identified. For hydrolysis constants only, a more detailed critical compilation has been prepared by Baes and Mesmer (1976), who give full details of the procedures by which they arrive at their selected constants, and also provide equations allowing the constants to be calculated at different ionic strengths. Extension of this approach to other ligands would, however, be an enormous task. In many instances the constants required are simply not available in the literature. Examples are provided by the carbonate complexes of the lanthanide elements, the borate complexes of most trace metals and the complexes formed by major cations with trace anions such as chromate, arsenate and selenate. In such cases it is necessary to estimate the stability constants using a plausible analogy (Langmuir, 1979). For example log-log correlations

13 between the stability constants of carbonate and oxalate complexes (Langmuir, 1979; Turner et al., 1981) have been used for this purpose although they are rarely reliable to better t h a n an order of magnitude. For example, the constant estimated for the formation of YCO~ at infinite dilution (log K = 6.94) (Turner et al., 1981) may be compared with the value recently obtained experimentally (log K = 8.2 calculated from the data given by Spahiu (1985)). It is usually necessary to adjust the experimental stability constants to the ionic strength of seawater. A variety of approaches have been employed, ranging from the use of simple electrostatic equations (usually the Davies equation, Whitfield and Turner, 1980) to the use of a non-linear least squares fit to the available experimental data. The latter approach, although the most direct, is not without its problems. In a survey of the speciation of 58 elements involving 528 complexes (Turner et al., 1981), the critical stability constants presented by Smith and Martell (1976) were interpolated over the available ionic strength range at 25°C and 1 atmosphere using the equation log flj = log flo + SAz211/2/(1 + BIll2) + CI

(1)

Sufficient data for a non-linear least squares fit were available for only 7.8% of the complexes considered. The ionic strength variability in the majority of cases was consequently obtained by analogy with complexes for which adequate data were available. In 29% of the non-linear least squares fits the F-ratios indicated that the experimental data were too scattered for a significant fit to be obtained. The complexes concerned included such familiar species as AgC1° and ZnC1 ÷ (Fig. 2). One major problem is that stability constants obtained at the same ionic stength in different media can often vary because of interactions between the medium ions and the equilibrium components. For example, anions resulting from the ionisation of EDTA form complexes with Na ÷ and K + ions which cause the stability constants obtained in sodium and potassium chloride media (Martell and Smith, 1982) to deviate significantly from those measured in tetramethyl ammonium chloride media (Martell and Smith, 1982; Arena et al., 1983). Millero and Byrne (1984) indicate how such specific medium effects can be treated by ionic solution theory. By comparison with the selection and correction of stability constants, the procedures used to incorporate the stability constants into thermodynamic models, though diverse, are subject to relatively small errors. Thus models based on the ion-pair concept (Dickson and Whitfield, 1981; Millero and Schreiber, 1982) and the specific interaction concept (Whitfield, 1978; Millero, 1983) are able to provide adequate descriptions of the major protolytic equilibria in sea water where the stability constants are well defined. The most promising model for trace metal speciation is the hybrid model which is being systematically developed by Millero (1982). This model is based on the concept of using the Pitzer equations to specify the weaker interactions involving the major sea salt components and to use the activity coefficients defined in this way as the basis for an ion-pair model for the more strongly interacting species (Whitfield, 1975).

14 m. []

g-

....i

, ~0.0

0.5

i 1.0

i 1.5

,

2.0

2.5

2.0

2.5

11121MI12

q co

o1:

v~. o _l

[]

o. ~0.0

O.S

1.0

1.5

11/2/i,41/2

Fig. 2. Critical stability constants as a function of ionic strength for AgC1° and ZnC1+ . Open symbols - - Smith and Martell (1976); closed symbols - - Martell and Smith (1982). The solid lines show non-linear least squares fittings to Eqn (1). I n a d d i t i o n to t h e s e l e c t i o n of s t a b i l i t y c o n s t a n t s , t h e s e l e c t i o n of c h e m i c a l species c a n also give rise to d i s c r e p a n c i e s b e t w e e n s p e c i a t i o n models (Nords t r o m et al., 1979; Whitfield a n d T u r n e r , 1980). T h e m o s t o b v i o u s e x a m p l e is p r o v i d e d by m i x e d l i g a n d complexes. S t a t i s t i c a l l y s p e a k i n g , if a m e t a l is able, for i n s t a n c e , to f o r m c o m p l e x e s MC12 a n d M(OH)~ w i t h s t a b i l i t y c o n s t a n t s fl2(C1) a n d fl2(OH), t h e n a m i x e d l i g a n d c o m p l e x of t h e f o r m MOHC1 s h o u l d also be f o r m e d w i t h a n effective s t a b i l i t y c o n s t a n t defined by fl(OH, C1) = 2 [fl2(C1)fl2(OH)] 1/2. A l t h o u g h s u c h c o m p l e x e s h a v e b e e n c o n s i d e r e d in a n u m b e r of sea w a t e r models ( D y r s s e n a n d W e d b o r g , 1975; Whitfield a n d T u r n e r , 1980) w h e r e t h e y u s u a l l y r e p r e s e n t less t h a n 20% of t h e t o t a l dissolved metal, little d i r e c t e v i d e n c e h a s b e e n p r e s e n t e d for t h e i r e x i s t e n c e in sea water. H o w e v e r , m i x e d l i g a n d c o m p l e x e s c o u l d b e c o m e i m p o r t a n t in p o r e w a t e r s w h e r e add i t i o n a l i n o r g a n i c ligands (e.g. H S - , S 2-, S~-, $20~ , SO 2-) are f o r m e d a n d q u i t e h i g h c o n c e n t r a t i o n s of a m i n o a n d c a r b o x y l i c acids a n d thiols m a y also be e n c o u n t e r e d (Byrne, 1983). T h e e s t i m a t i o n of s t a b i l i t y c o n t a n t s m i g h t t h e n

15 become more t han a simple statistical excercise as the larger ligands might introduce steric effects which hinder the formation of some mixed complexes. The problems cited so far arise from the incomplete state of our knowledge concerning complexation in simple systems. If we are to select the most accurate speciation models for trace metals from our u n c e r t a i n data base it will be necessary to test the models whenever possible against experimental data. The models describing protolytic equilibria have already been subjected to this kind of scrutiny and have, in general, worked extremely well (Whitfield, 1978; Dickson and Whitfield, 1981; Millero and Schreiber, 1982; Millero, 1983). When we consider applying speciation models to nat ural waters a wider range of problems is e nc ount e r ed which relates to the role of the biota in geochemical cycling. These problems are less tractable than those of a purely physicochemical origin and their solution calls for a closer integration of modelling studies and experimental measurement.

THE IMPACT OF THE BIOTA ON CHEMICALSPECIATION The active and persistent presence of the biota causes marked changes in the speciation of the elements which affect the applicability of equilibrium models. Although it is interesting to consider the consequences of this i nt ervent i on on the evolution of the elemental requirements of the biota (Whitfield, 1981; Williams, 1981; Morel and Hudson, 1985) we will c o n c e n t r a t e here solely on its impact on chemical speciation. The biota are able to harness the sun's energy to pull chemical processes away from equilibrium and hence to produce a bewildering a r r a y of chemical species. For example, given the present atmospheric oxygen cont ent (itself a biological artefact) an equilibrium model would suggest tha t selenium should be present solely as SeO~ in sea water. In fact a wide range of selenium compounds is found which includes SeO~ , SeO~ , (CH3)2Se, Se 2 and a range of seleno-amino acids. Similarly, complex inventories can be prepared for other non-metals and metalloids involved in biological cycling. The influence of the biota on the speciation of trace metals, wh eth er or not they are bioessential, is not quite so complicated but it does impose limitations on the use of thermodynamic modelling. The influence of the biota on chemical speciation can be considered in terms of (i) the production of particulate phases, (ii) the release of dissolved organic mater and (iii) the control of redox processes. Particulate matter

P h o t o s y n t h e t i c processes result in a massive flux of particulate m at t er from the surface layers to the deep oceans. This material consists approximately of 25% by weight of particulate organic m at t er and 75% of biologically produced mineral phases (Whitfield and Watson, 1983). The particles themselves can also provide a site for the deposition of hydrous oxides of iron and manganese. From our present perspective the primary effect of this particulate flux on speciation

16 will result from adsorption/desorption reactions occurring at the large, active surface areas produced. Such processes are most readily incorporated into chemical speciation models if adsorption is considered to result from reaction between the dissolved components and chemical species on the particle surface (Morel et al., 1981; Sposito, 1983). In establishing such models it is necessary to define (i) the nature of the surface species, (ii) the stoichiometry of the interaction and (iii) the electrostatic (coulombic) contribution to the free energy of the interaction between the trace metal ion and the charged surface of the particle. Although these requirements are entirely analogous to those which must be fulfilled in the establishment of a speciation model in solution, the introduction of particle surfaces produces significant complications by enabling many equally plausible options to be proposed for each characteristic. As a consequence, many different models can be constructed based on the concept of surface complexation which might give equally plausible fittings to the experimental data while invoking quite different mechanisms (Morel et al., 1981). The surface species selected must be defined with sufficient care to allow for a proper characterisation of the influence of the most important environmental parameters on the exchange process. To account for the pH dependence of adsorption~iesorption reactions, protonated or hydrolysed surface species of the type =-S-OH °, - S - O , - S OH2+ are usually postulated. Analogous species involving the exchange of major cations and anions must also be considered when modelling the adsorption characteristics of surfaces in sea water in relation to measurements made in other media. Each of these interactions must be characterised by an intrinsic stability constant for a surface reaction and a site concentration (usually expressed in terms of the surface area of the particles). The surface species proposed for interaction with the trace metal cations or complexes can be chosen with some flexibility depending on the extent to which the influence of pH and medium ions have been characterised by the foregoing intrinsic stability constants. The stoichiometry selected for the adsorption/desorption equilibria is not critical since the sites identified in the model do not necessarily have any physical reality. A 1:1 stoichiometry is usually implied. The remaining contribution towards the free energy of the adsorption/desorption reaction arises from the coulombic interactions resulting from the surface charge on the particle. The surface potential (~) cannot be measured directly and it is usually estimated by fitting the experimental data to an empirical relationship between ~ and the surface charge. Major differences in the treatment of the surface charge arise from the geometric description of the surface sites. For example, if all of the sites invoked are assumed to lie on a single plane, the surface charge can be estimated simply by summing the charges associated with the individual sites. However, in one of the most useful models (Davis et al., 1978; Davis and Leckie, 1978, 1979) two planes are invoked, the inner plane containing the protonated sites and the outer plane the sites associated with the medium and trace metal ions. A more complex algorithm is consequently required for calculating the nett charge. The intrinsic constants themselves should ideally be estimated at zero ionic

17 strength and zero surface charge; a procedure t h a t requires a careful double extrapolation of the experimental data (Davis and Leckie, 1979). The treatment of adsorption~iesorption reactions by the surface complexation model is therefore quite a complicated business involving the experimental determination of half a dozen or more fitting parameters, so simpler empirical treatments have also been employed. Probably the most useful are those which dispense with detailed considerations of the particle surface and consider the use of adsorption isotherms to describe the distribution of adsorbed species over a range of sites with different free energies (Sposito, 1981). It is only necessary to define the nature of the isotherm, the maximum adsorption capacity and the stoichiometry of the hydrogen-cation exchange in a particular medium. This approach can be used to summarise the adsorption characteristics of a surface over a wide range of pH and metal concentrations for a particular medium while recognising explicitly the heterogeneity of the surface (Kinniburgh et al., 1983). A number of less rigorous surface complexation models have been applied to natural particle populations in sea water (Schindler, 1975; Balistrieri et al., 1981; Li, 1981; Bourg and Mouvet, 1984) and, on occasion, explicit account has been taken of surface coatings of organic matter (Balistrieri et al., 1981; Hunter, 1983). Morel and his co-workers (Morel and Morel-Laurens, 1983; Morel and Hudson, 1985) have used such models to show that the flux associated with adsorbed species is insufficient to account for the depletion of trace metals (e.g. Fe, Mn, Zn, Cu, Ni, Cd) from the surface layers of the ocean, the bulk of the transport being associated with trace elements incorporated directly in the organic matrix. The adsorption process is probably more active in the deep ocean, and 'scavenging' onto detrital particulate matter has been postulated as a removal mechanism for copper, lead and thorium (Brewer and Hao, 1979). However, it is only in the case of thorium that an adequate description of the adsorption/desorption process has been provided for natural particle populations (Bacon and Anderson, 1982). Dissolved organic matter The processes of photosynthesis and respiration release a diversity of organic components into solution and these are augmented by the waste products from the feeding and metabolism of zooplankton and other grazers and predators. The dissolved and particulate organic material produced in this way is actively reworked by bacteria and by a host of protozoa. As a consequence, sea water contains a considerable number of potential ligands ranging from polydisperse and refractory organic compounds with no unequivocal structural description (e.g. humic and fulvic acids), through molecularly welldefined weak complexing agents (e.g. carboxylic and amino acids) to very strong and structurally specific chelating agents as exemplified by the formation of chlorophyll, vitamin B12 and the iron-specific siderophores (Gagosian and Lee, 1981; Mantoura, 1981; Morel and Morel-Laurens, 1983). The humic and fulvic acids are difficult to characterise since they are operationally

18

defined fractions rather than distinct chemical species and they incorporate a range of molecular types each of which might contain an array of functional groups. The assessment of stability constants for the interactions of trace metals with such heterogeneous mixtures is consequently a matter of selecting the best statistical fit to a range of equally feasible models from rather featureless titration curves (Cabaniss et al., 1984; Fish and Morel, 1984; Turner et al., 1986). Models which assume the presence of a limited number of discrete sites are readily incorporated into chemical speciation models and, by good fortune, they also seem to provide a reasonable fit to the experimental data over limited ranges of pH, metal concentration and ligand concentration. However, the models used are best considered as statistically reasonable summaries of the data rather than expressions of structural reality. Consequently, the polyfunctional character of the components of the humic and fulvic acid fractions can cause problems when attempts are made to apply the fitting models over a wider range of conditions (Buffie, 1984; Buffie el al., 1984). The stability constants and apparent site concentrations are found to be dependent on the metal/ligand ratios, and continuous equilibrium functions rather t h a n discrete equilibrium constants appear to be more appropriate (Buffie et al., 1984). Fortunately, the complexation reactions involved are not very strong and they exert a significant influence on the complexation of only a handful of metals (Cu 2÷, Fe 3÷, Hg 2÷ and, marginally, Pb 2÷ ). Only sparse data are available for the stability constants of metal interactions with the molecularly well-defined ligands that have been characterised in natural waters (e.g. amino acids, carboxylic acids, sugars). It would appear, however, that complexation by such ligands is likely to be most significant in interstitial waters or in biologically active areas where sufficiently high ligand concentrations can be generated (Byrne, 1983; Valenta et al., 1984). In such circumstances, mixed ligand complexes might assume a particular importance. This possibility deserves further consideration as does the complexation o f trace metals [particularly b-type metals such as Cd, Hg, Cu(I)] by thiols and sulphur-containing amino acids such as cystine. Some groups of algae produce compounds known as siderophores which are highly specific complexing agents for iron (pK values > 30). It is unlikely that such complexing agents will affect the speciation of other trace metals since the mass action effect enables magnesium (pK -~ 16) to compete very effectively for the available sites (Morel and Morel-Laurens, 1983). However, Florence (1982b) has indicated that a substantial fraction of the copper, lead, cadmium and zinc in natural waters is associated with highly stable complexes with organic ligands. Such strongly complexed metal is unable to exchange with other metal species in solution, and should be considered as a chemical form distinct from the free metal, inorganic complexes and weaker organic complexes such as those with fulvic acids.

19

Redox potential One of the first parameters that must be fixed when establishing a speciation model is the redox potential of the environment. This indicates the stable oxidation states of the various elements present and therefore controls the nature of the species that are available for complexation. In air-equilibrated waters the redox potential should be set by the oxygen tension via the O2/H20 couple. The redox potential calculated thermodynamically for this couple (1.299V at 25°C, 0.21atm. oxygen) suggests that only F-, C1- and B r - should be present as elemental anions and that most other elements should be present in their highest oxidation states. However, the O2/H20 reaction is kinetically hindered and does not exert a controlling influence in practice and the effective redox condition of air-equilibrated waters is equivocal. Metabolic processes produce a variety of reduced species [e.g. I , As(III), Se(IV), Mn(II)] which persist in the water column for long periods of time because of their slow reaction with the dominant oxidising agent, molecular oxygen. The biologically generated reduced species are augmented by the involvement of inorganic components [notably Fe(III), Mn(IV), Cu(II)] as electron acceptors in the photo-oxidation of organic matter in the surface layer of the ocean (Zafiriou et al., 1984). The redox chemistry of oxygenated waters is consequently complex and poorly coupled so that direct measurement, rather than speciation modelling, is required to unravel the chemistry of the elements involved. In stagnant systems where the available oxygen is rapdily exhausted, a variety of electron acceptors will be called in to play a role in the oxidative breakdown of detrital organic matter by microorganisms. The sequence o f acceptors is dictated largely by thermodynamic considerations (see, for example, Stumm, 1978; Stumm and Morgan, 1981; Turner et al., 1981). Although in general the various redox couples are more efficiently interlinked in anoxic systems than in oxygenated systems, anomalies still exist. For example, molecular hydrogen is found at all depths in some anoxic systems, frequently with the highest concentrations just below the oxic/anoxic boundary (Scranton et al., 1984), although the use of water as an electron acceptor would not be predicted thermodynamically at these levels. Thermodynamically unstable methylated compounds, notably of mercury, germanium, arsenic, selenium and lead can also be produced in anoxic sediments and contribute significantly to the geochemical cycling of these elements (Andreae, 1983). In the case of germanium the methylated compound is suggested to be the dominant form in sea water. A final example of disequilibrium might be provided by arsenic in anoxic sediments which persists as As(V) to a significant degree even in the presence of hydrogen sulphide (see Knox et al., 1984 and references therein). The presence of life in the oceans therefore complicates the application of chemical speciation models both by introducing considerable uncertainty into the characterisation of complexation reactions (particularly with organic ligands and particle surfaces) and by using the sun's energy to move the system

20

away from equilibrium, thereby restricting the applicability of thermodynamic modelling. As a consequence a number of attempts have been made to take a more direct look at chemical speciation in natural waters and its implications for biological and geochemical processes. We will now consider these experimental approaches to speciation.

DIRECT MEASUREMENTS ON NATURAL WATERS

The difficulties encountered in constructing reliable speciation models and the problems surrounding the application of the thermodynamic equilibrium concept in natural waters have resulted in the development of a range of pragmatic methods for studying trace element speciation (Fig. 3; Florence and Batley, 1980; Florence, 1982b; Turner, 1984, Kramer and Duinker, 1984). Ideally, the procedures employed should be selective and sensitive and should have a clearly defined response with respect to the chemical speciation of the element concerned. They should preferably be non-destructive or at least require a minimal manipulation of the sample. Needless to say such an ideal is rarely, if ever, achieved in practice. Physical separation and chemical transformation techniques (Fig. 3) invariably result in the destruction of the sample. Physical separation techniques have little chemical specificity since the fractions obtained are defined solely by the physical properties of the components. Chemical transformation techniques are slightly more selective since they yield fractions that result from the response of the solution components to a particular chemical operation. How-

PFIYS I CRL SEPRRRT I ON

~

F i l t r a t ion, s i z e ? r a c t ionat ion I;el f i l t r a t i o n Outgass ing o f vol a t i les O ial ys is x SolvenL e x t r a c L i o n

/Particle

leaching

CHEMI CRL TRRNSFORMRT ION ~ - - P h o L o o x idaL ion ~'Oer i v a t isaL ion /ASV

~1

-Sur£ace ~ b

CHEMICRL

SENS I NC

~ " B u l k

u

,,,,~-ion exchange x ioassa~ I +

I SE=÷ s p e c t r o s c o p y ~+

Fig. 3. Direct measurement techniques. The techniques are listed in order of decreasing destructive effect and increasing chemical specificity. (*) Non-destructive techniques; ( + ) equilibrium techniques; ( x ) destructive techniques which can be used to study equilibria in the perturbed sample.

21

ever, they are in general sensitive to chemical form rather than individual chemical species. Procedures included in these two categories enable the total element concentrations to be subdivided into various operational fractions but they do not in themselves provide information about chemical speciation. However, if the transformations between the various chemical species are slow such procedures can be used as the precursors to a chemically selective analytical method. For example, hydride generation followed by chromatographic separation and analysis by atomic absorption spectroscopy can be used to determine alkyl compounds and inorganic forms of As, Sb, Ge and Sn, including differentiation of inorganic As(III)/As(V) and Sb(III)/Sb(V) (Andreae, 1983). Derivatisation followed by chromatographic detection can also be used to determine the redox speciation [Se(IV)/Se(VI)] of selenium (Measures et al., 1983). Taken individually these destructive procedures can also provide useful circumstantial evidence concerning the chemical forms of the elements in solution. For example, photo-oxidation procedures can be used to assess the relative concentrations of organic and inorganic nitrogen and phosphorus compounds (Butler et al., 1979) and to give a crude measure of the quantity of a particular metal t h a t is bound up in organic complexes (Florence, 1982a). Furthermore, size fractionation techniques can provide an indication of the nature of metal-organic complexes in solution (Smith, 1976). In summary, physical fractionation and chemical derivatisation techniques are of little interest in themselves so far as chemical speciation studies are concerned. Their main value is as precursors to chemically specific analytical procedures or as sources of circumstantial information concerning the chemical forms of the elements. However, these procedures have been used as the basis for a number of elaborate schemes for the fractionation of the total dissolved component into a number of 'compartments' (Bartley and Gardner, 1978). Such schemes inevitably involve much painstaking work under clean conditions so t h a t it is seldom possible to process more t h a n one sample per day. Despite such care, the relevance of the various 'compartments' measured to the chemical speciation of the elements remains obscure since the compartments themselves are chemically ill-defined and they are not mutually exclusive. The most direct evidence concerning chemical speciation is provided by the chemical sensing methods (Fig. 3) since the forms or species determined are clearly defined. These techniques may be divided into surface techniques and bulk techniques. Surface techniques involve the introduction of an active surface (e.g. ion-exchanger, electrode) into the solution and the measurement of an appropriate response (amount exchanged, current flow, potential shift). The electrochemical techniques which fall into this category are nondestructive since only a small proportion of the total metal concentration is taken up by the interaction with the surface. The fraction sensed by the surface techniques (with the exception of the ion-selective electrode) is liable to be kinetically controlled, and thus depends on the extent to which the solution equilibria can re-adjust in the time that is allowed for the interaction between the solution and the surface. The bulk or spectroscopic techniques in their simplest form involve no modification of the sample and no direct contact

22

between the sample and the measuring device. However, such techniques are not very sensitive and for direct speciation of trace metals it is necessary to increase the trace metal concentration to the micromolar level (Brown and Kester, 1980; Byrne, 1981). In considering the applications of chemical sensing techniques it is necessary to take account of the relationship between the kinetics of the reaction controlling speciation and the timescale of response of the analytical method (Fig. 4). If the speciation reactions are slow then physical or chemical transformation techniques can be used to separate or preconcentrate the various chemical forms prior to species-sensitive analysis. If concentrations are high enough it is also possible to determine the individual chemical species directly in the sample. For example, the various components of the nitrogen system can be determined by chemical transformation followed by direct spectrophotometry (NO~, NO~, N H : ) or by volatilisation followed by direct chromatography (N20). However, if the equilibria between the individual species are rapidly attained, special care must be taken to minimise the disturbance during sampling and analyses. This is where a judicious combination of direct analysis and chemical modelling is most appropriate. For example, no suitably sensitive, direct analytical techniques are available for the determination of the individual concentrations of HCO~ and CO~- in natural waters. However, the use of an appropriate chemical model in conjunction with direct determinations of total carbon dioxide and pH or alkalinity enable their concentrations to be determined with some accuracy. Models for trace metal speciation in sea water are not yet as reliable as those for acid-base equilibria for a similar approach to be feasible. Consequently, the chemical sensing technique must be used where possible to test the predictions of the available chemical models or to provide information on chemical form

TECHNIQUES Spec fropholomefry (UVIVIsl ~|S[

DP Polorogrophy

Derivo f,sohon MnOz Equil Ion exchange Diotysis

E[ectrochem Ptoflng

m~nute

I

I

I

I

-8

-6

-/+

-2

Pb(l]) [d(II} Co([I) Cu(ll) Zn(II) Fe~II)

L 0 Log f/seE

Fe[~TI) AI{III)

hour

tl

Co(Ill)

d(ly week

I I

2

4

11

6

CdEDTA Fe(]l)

year

~ II

8

Cr(ED As(II])

Mn(II)

Ligand exchange II-I) OxidQfion

Se(M

PROCESSES Fig. 4. Timescales of chemical processes and m e a s u r e m e n t techniques. The ligand e x c h a n g e rates for metal ions are water e x c h a n g e rates ( M a r g e r u m et al. (1978)), which give an indication of the rates of metal complexation reactions. The timescale of the m e a s u r e m e n t techniques indicate the time available for chemical reactions as a result of non-equilibrium conditions imposed by each technique.

23 that is of direct relevance to particular environmental problems. To illustrate the use of such techniques we will consider their application to three problems of increasing complexity, (i) the verification of speciation models for lead, (ii) the determination of the organic complexation capacity of sea water and (iii) the determination of the biologically available fraction of trace metals.

Verification of speciation models for lead Chemical sensing techniques have been used to good effect to check the reliability of chemical models describing the status of protolytic equilibria in sea water and in estuarine waters (Whitfield, 1978; Dickson and Whitfield, 1981; Millero, 1983) but their use in verifying speciation models for trace metals have so far received little attention. The most coherent example of this approach is provided by studies of the speciation of lead in sea water. Reviews of the available models (Whitfield and Turner, 1980; Byrne, 1981) revealed that the predicted speciation patterns fall into three groups (Table 1): (i) domination by chloro-species, (ii) domination by carbonate complexes (notably PbCO °) and (iii) distribution between chloro, hydroxy and carbonate species with some importance given to hydroxy complexes. Sipos et al. (1980b) tried to differentiate between the various models by undertaking a careful study of the half-wave shifts observed in the pseudo polarograms obtained by the anodic stripping voltammetry of lead in a variety of media. These experiments entailed, inter alia, the redetermination of the stability constant for PbCO ° in perchlorate media. These constants were used in the formulation of a speciation model for lead which suggested the predominance of hydroxy and carbonate complexes. Furthermore, they compared the half-wave potential shifts observed in an artificial sea water medium over the pH range 5-9 with values calculated from the sea water model using the Lingane equation

AE1/2 -

n----F

i=1

j=l

The predicted and observed values are in good agreement, thus discounting those models that suggested a predominance of chloro complexes. Some discrepancy persisted however between the remaining models concerning the importance of hydroxy complexes in lead speciation. This problem was considered by Byrne (1981) who studied the UV absorption spectra of lead (enriched to 7 pM) in natural sea water. In particular, he looked at the effect o f p H on the absorption at 250 nm and was able to show that the effect depended on the total carbon dioxide content (CT) of the water. Furthermore, the results from a series of experiments at different CT values could be brought onto a single plot if the molar absorptivities were plotted against - log [CO~ ] rather than pH. An analysis of the absorbance data at five wavelengths suggested that the dominant form in sea water at pH 7.83 was PbCO~ and that species such as PbOH ÷, PbHCO~ and Pb(CO3)~- were of only minor importance. Taken together, the experimental evidence suggests that lead speciation in sea water is

24 dominated by PbCO ° (Florence and Batley, 1976; Whitfield and Turner, 1980). A systematic application of experimental measurements of this kind would do much to dispel the uncertainty t h a t still surrounds the speciation of many trace elements in sea water.

Complexation capacity Attempts have been made to circumvent some of the problems associated with estimating the influence of organic complexes on trace metal speciation by using chemical sensing techniques to determine directly the overall complexation capacity of natural waters with respect to certain metals (Hart, 1981; Florence, 1982a; Neubecker and Allen, 1983; Kramer and Duinker, 1984). In general, the sample is titrated with a trace metal and the free metal ion concentration (or a quantity assumed to be proportional to it) is determined via a chemical sensing technique (Fig. 3). Copper is normally selected for such titrations since it forms strong complexes with organic ligands and it is also close to its toxicity limit in many natural waters (Morel and Hudson, 1985). The resulting titration curve usually exhibits an inflection point which is interpreted as the point at which the concentration of metal ion added is equivalent to the concentration of complexing ligands in the sample. The slope of the titration curve beyond the inflection point should correspond to that which would be obtained in the absence of the complexing ligand. A suitable data analysis procedure (preferably based on the non-linear least squares approach) can then be used to estimate the apparent ligand concentration (or complexation capacity [L]) and a conditional equilibrium constant (K*). [L] is not a measure of the degree of complexation experienced by the t i t r a n t in the natural water. In reality most of the complexation sites identified are likely to be occupied by Ca 2÷ and Mg 2÷ and the value of [L] obtained is a measure of the potential for such sites to be occupied by the trace metal under consideration provided a sufficient excess of free metal ion is added. It will include neither the proportion of ligands initially occupied by the added metal nor the proportion of ligands occupied by metals which form stronger complexes. Similarly, the K* values obtained are unlikely to represent stability constants for the complexation of the trace metal with a single ligand since many competing ligands may be present in the solution and some of the complexation is likely to occur via ligand displacement reactions rather t h a n by direct complexation with the free ligand (Neubecker and Allen, 1983). The values of [L] and K* measured by the various experimental techniques will depend largely on (i) the time scale of the measurement technique, (ii) the time of equilibration of the added trace metal with the sample prior to analysis, and (iii) the range of metal concentrations employed. The timescale of the measurement technique is particularly important since this defines the 'window' of organic complexes t h a t are accessible to the method. For electrochemical techniques, Rfizid (1984) has provided a useful treatment of the kinetics associated with interactions at the electrode surface and those associated with the slow equilibration of the added metal spiked with the organic ligands.

Procedure

Single sample: add successive aliquots of Cu, allow equilibration time, measure labile Cu

Multiple samples: add ion-exchanger, add Cu spike, equilibrate, filter, measure dissolved Cu

Multiple samples: add Cu spike, equilibrate, add microorganisms, equilibrate, measure growth characteristic

Single sample: add successive aliquots of Cu allow equilibration time and measure [Cu 2 + ]

Sensing technique

ASV

Ion exchange

Bioassay

ISE Direct measurement of [Cu 2+ ] with few complications

Direct measurement of the effect of complexation on bioavailability

Only total Cu needs to be measured, so a variety of sensitive techniques can be used. Relatively free of organic artefacts

Sensitive analytical techniques (to 1 0 - l ° M )

Advantages

Sensing techniques for complexation capacity measurement

TABLE 2

Restricted to total Cu > 10-6M, problems arise with Cu ISE in chloride media

Result depends on organism type, physiological state, and growth characteristic measured. Other factors can affect viability

Procedure is time-consuming and prone to contamination, requires separate calibration

Labile Cu is an ill-defined fraction including inorganic and some organic complexes, so t h a t response depends on the nature of the ligand

Problems

Sunda and Hanson, 1979 Varney et al., 1984 J a r d i m and Allen, 1984

Gillespie and Vaccaro, 1978 Sunda and Gerguson, 1983

Figura and McDuffie, 1979 Jardim and Allen, 1984 van den Berg, 1984

Chau et al., 1974 Tuschall and Brezonik, 1981 J a r d i m and Allen, 1984 Varney et al., 1984

References

26 Although a variety of techniques have been used to assess the complexation capacity of natural waters (Hart, 1981; Neubecker and Allen, 1983), the most widely used methods correspond to the chemical sensing techniques discussed earlier (Fig. 3), and are summarised in Table 2. Despite the varying response t h a t one might expect from one experimental procedure to the next, relatively few studies have been undertaken to compare the [L] and [K*] values obtained by applying different techniques to the same sample (Tuschall and Brezonik, 1981; Jardim and Allen, 1984; Varney et al., 1984). A comparison of values obtained with secondary sewage effluents, algal exudates and humic acids using ion exchange (MnO2), ISE and ASV techniques was presented by Jardim and Allen (1984). The ISE measurements invariably gave the highest values of [L]. Interestingly, the MnO2 technique, which is also supposedly an equilibrium method, gave complexation capacities which were only one-half to one-sixth of the ISE values. Models implying the presence of two complexing sites invariably gave a better fit to the ISE data, whereas the data obtained from the other techniques indicated the presence of only a single site. Similar findings were reported by Varney et al. (1984) who used copper and lead titrations to determine the complexation capacity of a single fulvic acid extract using ISE, differential pulse polarography (DPP), linear scan anodic stripping voltammetry (LSASV) and differential pulse anodic stripping voltammetry (DPASV) techniques. With copper as the t i t r a n t the measured complexation capacities decreased in the sequence ISE

> DPP

> LSASV -~ DPASV

corresponding to the decreasing timescale of the measurement techniques. The equilibrium ISE measurements were again best described by a two site model, whereas the data obtained from the kinetic electrochemical methods could be adequately described by a single site model. Indeed, attempts to fit the kinetic electrochemical data to a two site model failed to converge. The stability constants obtained from single site fittings to all the data sets were very similar. When lead was used as a t i t r a n t no indication of an end point was obtained with the kinetic electrochemical techniques, suggesting either the absence of lead-fulvic acid complexes or t h a t such complexes are kinetically labile over the timescale of the experiments. The ISE data for lead indicated a complexation capacity which was actually greater than t h a t measured for copper, but which resulted from the formation of substantially weaker complexes. The interpretation of the kinetic electrochemical data was further complicated by the adsorption of organic matter at the electrode surface as indicated by the observed shifts in peak potential. These effects increase in the order DPP

< LSASV < DPASV

with the ISE measurements being unaffected. The sensitivities of the techniques increase in the order ISE

< DPP

< LSASV < DPASV

27 This suggests that, using electrochemical techniques, one can have sensitive measurements or one can have unequivocal measurements but one cannot have both. Since there are significant variations in the complexation capacities measured by the various techniques, it is important to establish the relevance of the different values to studies of the impact of chemical speciation on biological and geochemical processes. In this context the key comparison would be between the complexation capacities estimated by the various chemical methods and those obtained by bioassay techniques since the bioassay methods provide the clearest link between complexation capacity and biological utilisation (Sunda et al., 1984). Preliminary analysis of the results of such a study indicate the bioassay estimates to be well correlated with those from ion-exchange measurements, while complexation capacities estimated from ASV titrations correlate with neither of the former (D. H. Plummer, personal communication, 1985). The introduction of cathodic stripping voltammetry (van den Berg, 1984) in conjunction with ion-exchange techniques promises to extend the measurements to a wider range of trace metals and hence provide a broader basis for considering the utility of the complexation capacity technique.

Biologically available chemical species As work progresses it is becoming apparent that there is no single, simple mechanism for the uptake of trace metals by aquatic organisms. Consequently, a varied array of dissolved chemical species have been identified as being biologically available, including free metal ions, non-polar organic molecules, cationic and anionic hydrolysis products and metal complexes (Turner, 1984). The uptake of these components from solution appears to be restricted by transport through the biological membranes rather than by uptake onto the cell surface so t h a t there is no contribution from solution kinetics to the biological response. However, there is a significant kinetic component to the practical methods used to assess speciation in solution (Figs. 3 and 4). If the kinetics of the interactions in solution are slow (e.g. the oxidation of reduced forms of As, Se and I, hydrolysis of metal alkyls) then any suitably selective and sensitive analytical methods can be applied. Consequently, it is possible to show t h a t iodine is assimilated as iodate via the same route as nitrate ions and t h a t arsenic is readily assimilated as arsenate by the route developed for phosphate uptake. Similarly, the ready assimilation and toxicity of the alkyls of mercury, lead and tin have been investigated (Smith and Smith, 1975; Marchetti, 1978; Zuckerman et al., 1978; Sterritt and Lester, 1980; Wong et al., 1981; Topping and Davies, 1981) and related to their lipid solubility. If the kinetics of the solution interactions are rapid it is essential either to use techniques t h a t do not disturb the equilibria involved or to ensure t h a t the timescale of the analytical technique is adequately characterised. For a number of trace metals exhibiting mobile complexation equilibria, the free ion

28 appears to be the dominant biologically available species. This has been confirmed experimentally for copper (Anderson and Morel, 1978; Hodson et al., 1979; Sunda and Gillespie, 1979; Petersen, 1982) and, in a more limited fashion, for cadmium (Sunda et al., 1978), lead (Schulz-Baldes and Lewin, 1976; Merlini and Pozzi, 1977), zinc (Anderson et al., 1978; Allen et al., 1980; Peterson, 1982) nickel (Spencer and Nichols, 1983) and iron (Morel and Hudson, 1985). The free metal ion concentration in these studies was measured directly only in the case of cadmium, using an ion-selective electrode. In the other cases the free metal concentration was calculated from the solution composition using a speciation model, thus providing some circumstantial corroboration of the models employed. However, as has been shown for lead (Whitfield and Turner, 1980; Table 1), a variety of chemical models showing quite different distributions of the inorganic complexes, can often give closely similar free metal levels. Since the ISE technique for measuring free metal ion concentration is restricted to high total metal concentrations (> 10-6M) the application of less direct procedures must be considered. The bacterial bioassay techniques appear to be suitable for estimating free copper levels when in experienced hands (Sunda et al., 1984). An alternative approach might be to try and circumvent some of the problems associated with the ASV technique. For example, kinetic effects could be reduced by using a very thin diffusion layer (Magjer and Branica, 1977) and organic interferences could b e minimised by stripping into different media (Ariel et al., 1964; Anderson et al., 1982) or by analysis of the deposited metal using non-electrochemical techniques (Batley and Matousek, 1977). On occasion, individual metal complexes have also been shown to be directly assimilated by organisms. There is circumstantial evidence for the uptake of copper hydroxo complexes (Andrew et al., 1977; Dodge and Theis, 1979; Magnusson et al., 1979), although this is by no means unequivocal. Indications that mercury is available in an inorganic form in sea water (Davies, 1976) suggest that it can be assimilated as the chloro complex. There is also evidence for the assimilation of H2VO~ (Stendahl and Sprague, 1982) and Al(OH)~ (Helliwell et al., 1983). In the case of aluminium the conclusions are based on speciation calculations using a measured pH and experimental estimates of monomeric aluminium concentrations from ion exchange measurements (Driscoll et al., 1980). The toxicity of copper is also enhanced by the presence of organic ligands suggesting that Cu-citrate and Cu-ethylenediamine complexes (Guy and Kean, 1980), and Cu-DDC and Cu-oxine complexes (Poldoski, 1979; Florence et al., 1983) can be assimilated directly. The uptake of iron by algae is also enhanced by the formation of highly specific complexes with hydroxamate siderophores (Murphy et al., 1976). It is possible that further refinement of such experiments, possibly along the lines followed by Sunda et al., (1984), might provide useful practical tests for chemical speciation models. SUMMARY AND CONCLUSIONS A proper understanding of the speciation of the chemical elements dissolved in sea water is an essential prerequisite for providing a quantitative de-

29

scription of the processes controlling its composition and of the way in which this composition influences and is influenced by the vigorous cycle of life in the oceans. The establishment of reliable equilibrium speciation models for the trace components is hampered by the lack of reliable equilibrium constants at the appropriate ionic strength, even at 25°C and 1 atmosphere. The application of thermodynamic equilibrium models to the determination of speciation in natural waters is further hindered by the actions of the biota which ensure, (i) a significant flux of particles with active surfaces, (ii) the presence of a variety of organic ligands and (iii) the formation of significant concentrations of metastable redox states which persist for long periods of time. In the face of such uncertainties and in response to the pressure of practical problems related to the nature of the dissolved components in sea water, a range of pragmatic techniques have been developed to subdivide the total trace element concentrations into a variety of fractions with varying degrees of chemical significance. Examples were provided by discussion of techniques developed to measure the complexation capacity of natural waters and of the biological availability of trace metals. The efficacy of the pragmatic techniques depends in part on the rate of equilibration of the various species in the natural sample and in part on the chemical sensitivity and selectivity that they impart to the analysis. Where the rate of equilibration is slow a wide range of destructive techniques followed by appropriate chemical work-up can be used to estimate the concentrations of the various dominant species. Such analyses have provided us with our knowledge of the presence of metastable lower oxidation states in seawater. Where the rate of equilibration is rapid it is necessary to employ chemical sensing techniques. The ideal technique would not disturb the sample in any way and would provide an equilibrium measurement of selected species. Unfortunately, the techniques which approach this ideal most closely (ISE, spectroscopy) are the least sensitive and can only be used where the natural concentrations are artificially enhanced. The other chemical sensing techniques (bioassay, ion exchange, electrochemical methods) introduce some degree of disturbance to the sample and it is necessary to take due account of the timescale of their response when interpreting the results of measurements. Ironically, then, the pragmatic techniques meet their greatest problems when they are applied to those systems to which the thermodynamic equilibrium approach is most appropriate. However, this irony can be turned to good effect if the thermodynamic models themselves can be placed on a firmer footing by a careful assessment of the relevant stability constants and by a comparison of the results of the modelling exercises with direct measurements in natural waters.

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