Trace metal adsorption modelling and particle-water interactions in estuarine environments

ContinentalShelfResearch.Vol. 7, Nos. 11/12. pp. 1319-1332. l~S7.

0278~343/87 $3.00 + ().iX) (~) 1987PergamonJournals Ltd.

Printed in Great Britain.

Trace metal adsorption modelling and particle-water interactions in estuarine environments ALAIN C. M. BOURG* (Received 5 September 1986; in revisedform 15 May 1987; accepted 18 May 1987) A b s t r a c t - - A brief, critical review of empirical and conceptual metal adsorption models is presented. This covers models conditional to seawater chemistry, as well as models applicable to estuaries or other systems of variable solution chemistry. The conceptual surface complexation approach is used to show why desorption of trace metals does not necessarily occur readily in estuaries, and why in some cases adsorption can even take place. The relevance and weaknesses of this model for the understanding of the mobility and fate of heavy metals in turbid environments, as related to particle dynamics, is discussed using the macrotidal Gironde Estuary, France, as the example.

INTRODUCTION

THE transit, accumulation and, therefore, the fate of particle-reactive trace metals in coastal environments are governed by a combination of solid particle dynamics and chemical transfer processes (MORalS et al., 1982; OLSEN el al., 1982; JOUANNEAUet al., 1983; ELBAZ-POULICHETet al., 1984; DONARDand BOURG, 1986). Hydraulic and sedimentological studies can provide a reasonable picture of suspended solid transport and accumulation in coastal waters and estuaries (MmNoT, 1968; FESTAand HANSEN, 1978). Even so, the range and rapidity of variations in the physico-chemical conditions of estuarine and other coastal systems (e.g. salt marshes) make it difficult to identify the significant solid ~ solution transfer processes by field investigations, especially since they are unfortunately often restricted to either the solid particle or the liquid phase composition. Chemical models are tools which can help in the identification of such processes. In this paper, trace metal adsorption models available for the study of aquatic systems are briefly reviewed. The surface complexation model, a chemical equilibrium approach developed for simple solid hydrous oxides, and successfully used in river water (MouVET and BOURG, 1983), is applied here to estuarine and coastal waters. Its relevance to studies of the mobility of heavy metals as related to particle dynamics is discussed using the macrotidal Gironde Estuary, France, as an example. ADSORPTION

MODELS

The adsorption of particle-reactive chemicals has been described by various types of models. The most commonly used are empirical, although conceptual models have * Department of Water Resources and Quality, National Geological Survey, B . R . G . M . , BP 6009, F-45060 Orl6ans Cedex, France. 1319

A.C.M. BOURG

1320

recently been developed. Some of them are conditional to seawater chemistry, but in estuaries or other systems of variable solution chemistry and turbidity, more elaborate models should be used. One of the most basic geochemical tools for assigning a quantitative value to the chemical element-particle association is the distribution coefficient Kd (OLSEN e t al., 1982). For a given chemical element, this coefficient quantifies its distribution between solid and liquid phases: total concentration in solid phase (per unit mass of solid) Ko = total concentration in solution (per unit volume of liquid) "

(1)

This parameter is quite easy to measure in natural aquatic systems, but is of limited value. Indeed, the numerator does not allow the discrimination between the various forms of solid elements (i.e. in equilibrium with the solution or occluded in the mineral lattice during geological processes) and the denominator does not take into account hydrolysis and dissolved complexation (BouRG, 1982a; BOURG and MOUVET, 1984; BOURG, 1986). Theoretical calculations performed for simple systems involving only uranyl-carbonate-hydroxyl-goethite and cadmium-calcium-carbonate-hydroxyl-EDTAgoethite, showed that KO can vary by several orders of magnitude as the chemistry of the aqueous phase changes (LECKIE and TRIPATHI, 1985). g d is therefore both an empirical and a conditional parameter. It is simply a value applicable only to the particular set of conditions for which it is measured. This distribution coefficient is nevertheless still being widely used. For example, TREFY et al. (1985) measured log (Kd) values of 5.49 _+ 0.18 1 kg-t (n = 15) for lead at the mouth of the Mississippi River. The values of the 15 samples could be validly averaged only because of their relatively ambient constant ionic strength and pH, uniform particle composition and good water mixing. Adsorption isotherms (plots of concentration of adsorbed element per unit mass of solid as a function of concentration of dissolved element per unit volume of liquid) are only a representation of the Kd distribution coefficients as functions of the concentration of dissolved elements. The non-linearity of the adsorption isotherms (parabolic or hyperbolic shape for Freundlich or Langmuir isotherms, respectively) demonstrates that Kd values are most often not constant, even though they might have fixed values for the lowest concentration range of the isotherm. Even then, they remain conditional parameters as is demonstrated, for example, by the pH dependence of adsorption (Fig. 1). Sshaped isotherms can also be observed in the presence of limited quantities of a compound which complexes (i.e. maintains in solution) the element studied. OAKLE¥ et al. (1981) proposed an equilibrium adsorption model to describe the partition of trace metals between the liquid phase and various adsorbents in the marine environment. The adsorption of the dissolved metal ion, M, by a solid S is accounted for by the reaction Keq M + S ~ MS (2) with the corresponding thermodynamic equilibrium constant

Keq

=

([MSI3'M---~/([MI [S]'¢MYX),

(3)

where MS is the surface covered by the metal, S is the remaining surface and ? are the respective activity coefficients (brackets signify concentration in tool 1-1 and overbars

Trace metal adsorption modelling

1321

Kd-ES =

,,"

$

o_ ~

pH=a

/ /

zE I,--

o

Q

OlSSOLVEO [ONEENTRATION ([t) Fig. l.

Adsorption isotherms.

describe surface species). This is a conceptual model. The adsorption reaction is considered equivalent to any solution reaction, such as complex formation. Because of operational difficulties in applying equation (3) (such as the required knowledge of the concentration of all metal-ligand complexes), OAKLEV et al. (1981) made simplifying assumptions and thus derived a conditional equilibrium constant approach. The adsorption reaction is now written as

Mr + ST

K~ MS

(4)

with Ksw = [MSJ/([MT]{ST}),

(5)

where [MS] is the concentration of adsorbed metal in moles per litre, [Mr] is the total concentration of dissolved metal ions in moles per litre and {ST}is the total concentration of the solid adsorbent in grams per litre. Equations (4) and (5) are conditional to seawater, where the concentrations of all major species, pH, pE and ionic strength can be assumed to be constant. Equation (5) is identical to a linear adsorption isotherm where F, the adsorption density of a metal on a solid surface (in mol g-l) is expressed as

r = [MS]/{S~}.

(6)

In this case, the conditional constant Ksw is identical to the distribution coefficient Kd and is equal to the slope of the linear part of the adsorption isotherm (F vs Mr; Fig. 1). This linear adsorption isotherm is obtained only when ~,{ST} >> [MS], where k is the number of moles of surface sites per unit mass of adsorbent (~, is equivalent to the exchange capacity). This means that the adsorption of the trace metal must be negligible compared to the total concentration of available surface sites. OAKLEYet al. (1981) measured Ksw values in the laboratory for typical components or model components of marine sediments (bentonite clay, Fe(OH)3, MnO2 and humic acid) and they applied the model to quantify the relative importance of these four solid phases in the adsorption of Cu and Cd by marine sediments.

1322

A . C . M . BOURG ADSORPTION AS SURFACE COMPLEXATION

The equilibrium adsorption model proposed by OAKLEYet al. (1981) can be developed further. Only a brief outline is given here and the reader interested in more details is referred to the papers or volumes of STUMMet al. (1976), STUMMand MORGAN (1981), ANDERSONand RUBIN (1981) and BOURG (1987). Hydrous oxides as well as all solid particulate matter found in coastal environments contain ionizable functional groups (-OH, -COOH, - N H 2 ) . In the presence of water these surface groups can accept or give out protons, depending on the pH. Surfaces, therefore, behave amphoterically. SOH ~ SO- + H +

(7)

SOH + H + ~ s O H f .

(8)

The adsorption of metal cations ( M z÷) can be interpreted in terms of competition with protons for surface sites, according to the following equation (HonL and STUMM, 1976; SCHINDLERet al., 1976) ( s o n ) , , + M z+ ~ (SO),,M (z-")+ + nH +,

(9)

where, for steric reasons, n is equal to only 1 or 2 (Fig. 2). For natural solids the observed value of n is 1 (MouVET and BOURG, 1983; JOHNSON, 1986). The various surface reactions described by equations (7)-(9) are identical to solution coordination reactions with the restriction that surface sites should be treated as polymers or associations of polymers. A single surface adsorption site is, indeed, not independent of other neighbouring sites. The equilibrium constants (for surface acidity and surface complexation) should be split into an intrinsic chemical term and a coulombic term: p~,,,ac¢

{F~Azt = f~surface Pn(intrinsic)'exp~},

(10)

where • is the electrical potential at the plane of adsorption and Az is the net change in charge number of the surface species due to the adsorption reaction. The value of q~ cannot be measured directly. For acid-base reactions of simple hydrous oxide surfaces, the intrinsic constants can be evaluated by extrapolating the values of the constants measured for various surface charges to conditions of zero surface charge (STur~Met al., 1980). Overall constants are then derived from the pH dependence of the surface charge. For surface complex formation reactions the coulombic term is often taken equal to unity (HOHL and STUUSt, 1976; SCHINDLERet al., 1976) or assumed to be constant at a given ÷

;-OH2 ,-OH

11 Ko2

+

p~

"-0-'~

;-0" Fig. 2. The surface complexation model.

1323

Trace metal adsorption modelling

ionic strength (and is thus included in the value of the intrinsic constants which is therefore a conditional constant). In the ionic strength gradient of estuaries, the electrical potential term can be neglected, in a first approximation, because it will affect all species equally (if they are of the same charge which is the case of the major adsorbing cations, i.e. alkaline earth and transition metals). However, this point definitely needs further investigation. Aquatic and terrestrial solids show the same adsorption pattern as simple hydrous oxides (BouRG, 1987). The adsorption curves present a sharp increase in adsorption with increasing pH through a restricted 1-2 unit pH range. This part of the curve, called the pH-adsorption edge by BENJAMINand LECKIE (1980), is characteristic of each metal on any chosen adsorbent. It is also dependent, although to a much lesser extent, on adsorbate concentration (BENJAMIN and LECKIE, 1981; LECKIE et al., 1980) and ionic strength (BOuRG, 1983a). A P P L I C A T I O N OF A C O N C E P T U A L A D S O R P T I O N M O D E L TO C O A S T A L ENVIRONMENTS

Chemical models of the behaviour of trace metals in the whole salinity gradient of estuaries are scarce. Even fewer are those which consider adsorption processes (BouR6, 1983a). The model results presented here were calculated using the computer code ADSORP (BOuRG, 1982b). The formation constants of the dissolved species have been detailed elsewhere (MouVET and BOURG, 1983). The adsorption properties are taken as those of suspended matter (crdme de vase) of the Gironde Estuary, France, whose adsorption constants were previously published (BouRG and MOUVET, 1984). Unless otherwise noted, the concentration of surface sites is taken as fixed and equal to 10-3 tool 1-1. The generalization of the properties of the cr~me de vase to those of the suspended particles of the Gironde turbidity maximum is justified, as the major characteristics (such as specific surface and particulate organic carbon) of these two particle suites are very similar at a given salinity (BOuRG, 1983b). The end members of the theoretical estuary used in the model calculations are reported in Table 1. The model assumes full adsorption reversibility and instantaneous equilibrium. The total concentration Table 1. Chemical composition (in mole 1 1, unless otherwise noted) of the end members of the theoretical estuary used to model trace metal adsorption (river water of high alkalinity) Constituent

River water

Seawater

Na K Mg Ca CI

0.00030 0.000065 0.00020 0.00057 0.00020 0.00015 0.00136 0.000018 8.43 0.004 0.144

0.47932 0.01045 0.05439 0.01053 0.55862 0.02889 0.00186 0.000275 8.12 0.7 35

SO 4

HCO3total CO3~otaI pH Ionic strength Salinity ( ~ )

From DVRSSENand WEDBORG (1980).

1324

A . C . M . BOURG

I0C ~3 8O Cdt°tat = 10-8-o.001 _M Wm rY o

eo

/

0.1M NaCl

/

4O ~z 2o

/

//

,°o NaNO~M/"~/~/I~'----~"~-~

~0.01

Ill/I/

8O

NoCI+ CaCl2

o

y

NoEl

/ ~

4o

d

2O

/

0.SN NoCI + O.05M MgC{2

Cu total = 10-7-N 6

B

7

9

10

pH

Fig. 3.

Influence of the composition of the aqueous phase on the adsorption of Cd and Cu.

(adsorbed + dissolved) of the metals investigated are 10-8 moll - 1 for Cd, 10 -7 mol 1-1 for Cu, Ni and Pb and 10-6 mol 1-1 for Zn. In the absence of experimental data, the electrical potential • is taken to be zero. The influence of the chemical composition of the dissolved phase on adsorption is demonstrated in Fig. 3. This effect is especially significant for the most weakly adsorbed heavy metals (BOORO, 1983a). Indeed, the Cd and Cu curves in the same medium are quite different. Cd is less strongly adsorbed than Cu and, of the major ions which are adsorption competitors, Mg is more strongly adsorbed. This last observation stresses the importance of the accurate determination of adsorption constants of Mg and Ca.

Importance of turbidity SALOMONS(1980a) conducted an experimental study of the adsorption of Cd and Zn on Rhine River sediments. Calculations with ADSORP reproduced well the observed pattern of decrease of the fraction of Zn adsorbed as a function of chlorinity (BouRG and MOUVET, 1984). Figure 4 presents examples of the adsorbed/dissolved distribution of Cu and Zn in a model estuary (of the macrotidal type) for two positions of the turbidity maximum, and for a constant pH value (pH = 8). The curves result from the combination of the effects of chlorinity and turbidity. Because the turbidity maximum of an estuary is always located at the freshwater-brackishwater interface, one can appreciate the difficulty in identifying the significance of adsorption/desorption processes in the fluvial zone and in the zone of very low salinity. As water enters the turbidity maximum, turbidity increases

Trace metal adsorption modelling

1325

100 80

,?, ~o

cu]_ ....

J

i

4o

20 0

e

s0/ t

i

,

i

i

0

5

10

15

(*/001

cl

20

Fig. 4. Adsorption in a model estuary of the macrotidai type (constant pH of 8 and surface-site concentration of 1 mol kg t).

and the mass action law of Le Chatelier requires that the equilibrium of equation (9) be displaced towards the right. The peak in fraction adsorbed might not always be so sharp because in Fig. 4 the turbidity increase is high (2 orders of magnitude, typical of French Atlantic macrotidal estuaries). The strong decrease of the fraction adsorbed with distance seaward throughout the estuary might also be partly offset by the commonly observed seaward increase in pH (Fig. 5) in estuaries formed with low alkalinity rivers. SALOmONS(1980a) observed a similar trend for Cd in a semi-quantitative profile drawn from adsorption experiments. The effect of turbidity on adsorption in estuaries which are not of the macrotidal type can be summarized by the typical examples shown in Fig. 6. The adsorbed fractions of Zn and Cu follow very closely the turbidity profiles, whatever the shape of the turbidity gradient.

~

higl ' ~tkutinity flyer

"T- m

~-

~

m e d i u m /

~koti.~y

river j

tow 7 ~

/

o tko tini y river

'

,~

~o

CI (%0)

Fig. 5. Typical pH profiles in estuaries (from MOOK and KOENE, 1975).

1326

A.C.M.

5O

i

i

i

i

i

i

i~

i

i

i

i

BOURG

i

i

25 0

. 100

i

i

,

50 0 _ 50

~,

i

r

i

zs

-e i ~ ' ~ ' l

o

10

20

300

10

20 30 0

i~

10 20 30 0

i

10 20

30

Salinity (%01 Fig. 6.

Turbidity and adsorption in model estuaries of types other than macrotidal ( s a m e p H and surface-site concentration as in Fig. 4).

Importance of pH Typical chlorinity-pH relationships in estuaries are schematized in Fig. 5. The pH minimum at intermediate salinities should lead, at constant turbidity, to a decrease in the fraction of metal adsorbed (desorption). Such a decrease is, however, often masked by the presence of the turbidity maximum in this part of the estuary, especially in macrotidal estuaries where the turbidity maximum is very large and spread out. Several authors (e.g. ETCHEBER,1978) have observed a slight increase in the concentration of particulate metal (per unit mass of solid) at the mouth of an estuary which is predictable from the pH distribution.

C3

100

,

,

W

oo r~ 0 80 b~ rm

Surface = 3

x

10"~, mote/l_

rY oo_ o I_J

/,C

2C

8.5

i

i

i

,

,

,

I

5

10

15

20

7.0 0

CL (%o1 Fig. 7.

C u adsorbed in a model estuary (constant turbidity).

1327

Trace metal adsorption modelling

Figure 7 presents the case of Cu for a constant turbidity. The surface-site concentration is substantially smaller than that taken in earlier calculations in order to show more detail in the low chlorinity zone. In this case, there is no increase of the fraction of metal adsorbed at high salinity because of the complexation of dissolved Cu with carbonates; complexation is calculated to be more significant than adsorption on estuarine suspended matter.

Importance of chloride and carbonate complexation Figure 8 presents the fraction of metal adsorbed as a function of chlorinity. The evolution of the adsorbed fraction does not depend only on the adsorption affinity of a metal. The fact that the Cu and Pb curves intersect comes from the different complexation behaviour of these two metals with carbonate and chloride ions in the dissolved phase. The formation of chloro complexes competes with adsorption (VUCETA, 1978; SALOMONS, 1980a, b; BOURG, 1983a). Cd forms stronger chloro complexes than Zn, and it is therefore more affected by increasing chlorinity (Fig. 9). 100

i

J

80 ~

rn ¢.y

i

Surface = 10-3 mote/L

60

0

t'~""'---

5

i-, 15

J

10

20

Ct (%0) Fig. 8.

Adsorption vs chlorinity (constant pH of 8, and constant turbidity with surface-site concentration of 10 ~ m o l l ~).

100

I

I

!

u.l ,,,v

In

rY o t=l 50 _.1

/ ,,/v

-v--

O.S%.I

Cd

19 %o

/ ~ , ~ V .," "d /

0"

N

/

CHLORINITY 0.5 %. I Zn - 0 - - 19 %o I

-v--

m..--..4---@ 8

pH Fig. 9.

Adsorption on Rhine River sediments (from data of

SALOI',~ONS,

1980a).

1328

A . C . M . BOURG

>, 5O

i

i

8

i

"~- 0

I

I

I

7 0

10 20

30

0

10 20 30

SALINITY (%.)

0

- °=°"1

10

20

30

0

-*=

10 20

1,

,

,

30

,

Fig. 10. Effect of pH and turbidity on the adsorption of Cu and Zn (the fractions of metal adsorbed are calculated for the turbidity and pH gradients given in the upper diagrams).

Adsorption geochemical computer codes such as ADSORP allow predictions of the effects of combinations of parameters such as salinity, pH and turbidity. An example for decreasing turbidity and increasing pH is given in Fig. 10. The adsorbed fractions are given in a logarithmic scale to permit a better comparison with pH, which is also of logarithmic nature. These model calculations are far from perfect, as indicated in the next section. They do, however, help to explain the somewhat conflicting evidence on trace metal behaviour during solid-solution exchange in estuaries. The exposure of adsorbed metal to increasing salinity does not necessarily involve desorption. The pH, the dissolved carbonate and the suspended matter concentrations are variables which are at least, if not more, important than salinity. They should, therefore, be included in all estuarine surveys. Relatively strong gradients in pH and turbidity in the very low salinity zone mask salinity effects, and make this poorly studied area of estuaries a very, if not the most, important zone for heavy metal exchange processes. The calculations presented here give only an instantaneous picture of the partitioning of trace metals between adsorbed and dissolved fractions. The retention of metals in estuaries can be strongly affected further by sedimentation and resuspension processes due to local hydrodynamics of the solid phase. MODEL LIMITATIONS

The previous calculations assumed: (1) full and instantaneous adsorption reversibility and (2) constant particle surface properties throughout the salinity gradient. The kinetic and reversibility aspects are discussed here. The regional particle differences will be treated in the next section. Desorption studies (both of amplitude and reaction rate) provide contradictory results. Laboratory experiments showed that only partial desorption took place when chlorinity increased (KHARKAR et al., 1968; SALOMONS, 1980a) or when pH decreased (LION et al., 1982). In other mixing experiments, the release of copper at low salinities, observed in the field, could be reproduced (WINDOMet al., 1983). VAN DERWEIJDEN et al. (1977) observed a greater desorption of metals when suspended

Trace metal adsorption modelling

1329

matter from the Rhine River was suspended in seawater, than when it was mixed with distilled water and artificial seawater made up of only NaNO3. They concluded that complex formation is important in desorption processes during estuarine mixing. PARTICLE-WATER INTERACTIONS AND TRACE METAL BEHAVIOUR IN THE GIRONDE ESTUARY: CHEMICAL AND HYDRODYNAMIC PROCESSES

In the Gironde Estuary, France, 80% of the up-estuary trace metal input is in particulate form, whereas the output to the ocean occurs mainly (80%) in dissolved form (JOUANNEAU, 1982). Moreover, the concentrations of particulate trace metals are smaller in the turbidity maximum than at either end-member of the estuary (ETCHEBER, 1978; ELBAZ-PouLICHETet al., 1984). These observations indicate the existence of a mobilization process affecting drastically the initial riverine metal flux to the ocean. Rapid seaward decreases in the heavy metal content of the suspended matter were observed in the upper reaches of the estuary (ETCHEBER, 1978). They were interpreted in terms of the physical mixing of metal-depleted particulate matter from the mid-estuary, with metalrich fluvial particles (JouANNEAU et al., 1983). Estuaries provide strong gradients in chemical and biological parameters, as well as hydrodynamic conditions, all of which are capable of modifying the nature of particulate matter. It was observed, for example, that in the Gironde Estuary: (1) particles contained 5-15% particulate organic matter and 5-6% Fe and Mn oxyhydroxides in the very low salinity zone, and less than 5% particulate organic matter and 6-7.5% Fe and Mn oxyhydroxides in the middle of the estuary (ETCHEBER et al., 1983); and (2), that particles from the turbidity maximum zone presented specific surface areas twice as large as particles from either end of the estuary (BouRG, 1983b). This variability in the nature of suspended matter makes modelling more difficult in estuaries than in rivers. In a first attempt, a field study was carried out using the computer model ADSORP to investigate the geochemical control mechanisms of dissolved trace metals in the low salinity zone of the Gironde Estuary. Adsorption/desorption for Cd and Zn and precipitation/dissolution (of malachite) for Cu were proposed to explain the field data (BouRG, 1984). Another model was suggested to explain the remobilization of metals in the turbidity maximum (DONARD and BOURG, 1986). This qualitative model recognizes the complementary nature of hydrodynamic processes (related to neap-spring tidal cycles) and of biological and chemical processes taking place in the Gironde Estuary. In this macrotidal estuary the hydrological factors generate a two-step sedimentological cycle. Slightly anoxic conditions which develop in the fluid mud lenses under low energy hydrodynamic conditions induce a trace metal transfer from particulate to dissolved form. Resuspension of the deposited particles, bringing about oxic conditions, then produces a trace metal transfer from dissolved to particulate form. The repetition of this cycle on the iron oxide coatings of particles results in the enhancement of the surface reactivity of the suspended matter. The overall balance of these processes is in favour of a dissolved metal flux to the ocean (JOUANNEAU, 1982), demonstrating the importance of the relative kinetics of the biogeochemical reactions in the context of fluid and particle transport. The trace metals are released from the solid particles during the dissolution of reduced Fe(II) (in slightly reducing bottom water layers) at a faster rate than they are removed from the water

1330

A . C . M . BOUR6

column by adsorption onto, or co-precipitation with, freshly formed Fe(III) hydroxides (in the overlying oxygenated water layers). This model is consistent with the existence of the peak in dissolved trace metals observed just down-estuary of the turbidity maximum in both the Gironde and Seine River macrotidal estuaries (ETCHEBEg, 1978; BOtJST, 1981). The time-scale for this biogeochemical model to develop is set by the fortnightly neap-spring tidal cycle. During periods of decreasing tidal amplitude, at each successive tide, peak current velocities decrease and the duration of slack water increases. During waning tidal ranges, sedimentation engenders pools of fluid mud which accumulate during neap tides. When the tide range increases, the reverse process occurs, with net erosion due to the increasing current velocities at each successive tide. Whichever model is accepted, it involves adsorption/desorption processes as a significant control mechanism for trace metal concentrations. Whether the solid particle dynamics are active in the remobilization process or not, sedimentation/resuspension phenomena are significant for the retention of trace metals in the estuary. CONCLUSIONS

The adsorption model presented here indicates the importance of turbidity (available adsorbing sites), pH and water composition (complexation by chloride and carbonate ions and competition of major cations for adsorption sites) for the regulation of dissolved trace metals in estuaries. With this conclusion in mind, a careful re-evaluation of past field surveys (where pH and turbidity data are available) should be conducted. The identification of the control mechanism for the particulate/dissolved fraction does not limit the significance of solid particle hydrodynamical processes for retention (permanent or temporary trapping) in estuaries. The anoxic conditions resulting from sedimentation render particulate trace metals susceptible to remobilization. Acknowledgements--The work reported here was partly performed at the Institute for Inorganic Chemistry of the University of Bern, Switzerland. This paper was written with the support of the Water-Rock Interactions programme of the French Geological Survey. Many thanks to Olivier Donard (University of Bordeaux, France) for collaboration during part of the Gironde work, to an unknown reviewer for editing the English, and to Anna Kay Bourg for typing the manuscript. REFERENCES ANDERSON M. A. and A. J. RUBIN, editors (1981) Adsorption of inorganics at solid-liquid interfaces, Ann Arbor Science, Ann Arbor, Michigan, 357 pp. BENJAMIN M . M . and J . O . LECKIE 0980) Adsorption of metals at oxide interfaces: effects of the concentrations of adsorbate and competing metals. In: Contaminants and sediments, R. A. BAKER, editor, Ann Arbor Science, Ann Arbor, Michigan, pp. 305-322. BENJAMIN M. M. and J. O. LECKIE (1981) Multiple site adsorption of Cd, Cu, Zn and Pb on amorphous iron oxyhydroxide. Journal of Colloid Interface Science, 79, 209-211. BOURG A. C. M. (1982a) Une mod~:le d'adsorption des m6taux traces par les mati6res en suspension et les s6diments des syst~mes aquatiques naturels. Comptes Rendus Acaddmie des Sciences Paris, 294, 1091-1094. BOURG A. C. M. (1982b) ADSORP, a chemical equilibria computer program accounting for adsorption processes in aquatic systems. Environmental Technology Letters, 3,305-310. BOURG A. C. M. (1983a) Role of fresh water/sea water mixing on trace metal adsorption phenomena. In: Trace metals in sea water, C. S.WONG, J. D. BURTON, E. BOYLE, K. BRULAND and E. D. GOLDBERG, editors, Plenum Press, New York, pp. 195-208. BOURG A. C. M. (1983b) Mod~lisation du comportement des m~taux traces it l'interface solide-liquide darts les systemes aquatiques. Document du B.R.G.M., N°62, Editions du B.R.G.M.. Orl6ans, France, 171 pp.

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133 i

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