the role of phytoplankton cells on the control of heavy metal concentration in seawater

ELSEVIER Marine Chemistry 48 (1995) 215-236

The role of phytoplankton cells on the control of heavy metal concentration in seawater

Melchor GonzAlez-Dkila

Deparfamento de Quimica, Facultad de Ciencias del Mar, Universidad de Las Palmas de Gran Canaria, 35017 Las Palmas de Gran Canaria. Spain

Revision accepted 21 July 1994

Abstract

An overview is presented of the ability of phytoplankton to passively adsorb and actively assimilate heavy metals from their aqueous environment and to release into the environment organic ligands capable of complexing metals. The uptake of all necessary trace metals by phytoplankton occurs via binding to a surface ligand and subsequent transfer across the cell membrane. This sorptive process can be explained by using surface complex formation equilibria; due to the heterogeneity of the algal surface, multi-site binding models must be developed. The production of extracellular organic matter with metal complexing properties plays an important role in decreasing the concentration of free metal ions and thus mitigating the potential toxic effects on organisms. However, while much research has been carried out on the uptake of single species of metal ions, little attention seems to have been given to the study of multimetal ion systems. Synergistic and antagonistic interactions between multiple trace metals are expected and could be very important in the oceans. These types of behaviours are extremely complex, influencing uptake and release of natural ligands and limiting plankton production and plankton species composition in the oceans. Future investigations should be carried out in order to gain understanding how the combinations of metal ions affect the physiological, biochemical and ecological processes of phytoplankton in seawater.

1. Introduction

The control of trace metal concentrations in the oceans has been a subject of considerable dispute ever since Krauskopf (1956) demonstrated that these concentrations were all below the limits set by the least soluble compound to the expected in a medium with the composition of seawater. There- fore, the scavenging of trace metals by particulate matter appears to be the most important process in regulating the concentration and distribution of metals in the oceans (Whitfield and Turner, 1987). Moreover, it is now widely accepted that

adsorption and complexation of trace metals by natural dissolved and particulate organic materials is a critical factor that influences both bio-availability and toxicity to organism (Coale and Bruland, 1988; Goncalves and Lopes da Conceicao, 1989), and thereby regulating the residual concentration of dissolved metal ions in lakes, rivers and oceans (Morel and Hudson, 1985; Sigg, 1987; Gonzblez-DCvila and Millero, 1990; GonzBlez-DBvila et al., 1990). Numerous species of algae, both macro- and microalgae, are capable of sequestering significant quantities of either nutrient or toxic heavy metal ions from

0304-4203/95/$09.50 0 1995 Elsevier Science Publishers B.V. All rights reserved SSDZ 0304-4203(94)00045-X

216 M. Gonzdez-D&da/Marine Chemistry 48 (1995) 215-236

aqueous solutions. The algal metal sequestering process occurs by different mechanisms (Wood and Wang, 1983; Folsom et al., 1986; Greene et al., 1986; Watkins et al., 1987; Greene and Darnall, 1988; Crist et al., 1990, 1992), depending on the algae, the metal ion species, the solution conditions, and whether the algal cells are living or non-living. The following mechanisms of algal resistance to toxic elements have been observed (Sunda and Guillard, 1976; Wood and Wang, 1983; Greene et al., 1986; Xue et al., 1988; Maeda and Sakaguchi, 1990):

(1) The development of energy-driven efflux pumps that keep toxic element levels low in the interior of the cell.

(2)

(3)

(4)

Oxidation state change by which a more toxic form of a metal can enzymatically and intracellularly be converted to a less toxic form. Precipitation of insoluble metal complexes on the cell surface.

(5)

(6)

(7)

Complexing of metal ions with excreted metabolites (extracellular products), which can extracellularly mask a toxic metal. Vaporization and elimination by means of con- verting a toxic metal to the volatile chemical species. Binding of metal ions with protein or poly- saccharides in the interior of the cell, which may deactivate the metal ion’s toxicity. Methylation of the element, which can enzymatically and intracellularly prevent a toxic element from reacting with a -SH group.

Phytoplankton affects trace metal chemistry in natural and oceanic waters not only by surface reactions, but also by metal uptake and by production of extracellular organic matter with metal complexing properties. The release of extracellular organic matter from marine phytoplankton appears to be a major source of labile substrate to the dissolved organic matter (DOM) pool in the open ocean (Duursma, 1961; Wangersky, 1978; Zhou. and Wangersky, 1989a). Both the direct extracellular exudation products and the secondary products after biochemical modifica- tion have been demonstrated to have the ability to complex trace metals (Fogg and Westlake, 1955; Swallow et al., 1978; van den Berg, 1979;

Mantoura, 1981; Fisher and Fabris, 1982; Imber et al., 1985; Zhou and Wangersky, 1985, 1989b; Seritti et al., 1986). A large fraction of this organic matter is surface active and represents the main part of surfactant activity in the sea (Zutic et al., 198 1). Extracellular production of organic ligands by phytoplankton may be the most important factor in controlling the effects of biological activity on the trace metal-organic interactions. This production may therefore be the key to understanding the dynamics of these compounds in the oceans. The production of extracellular organic compounds by phytoplankton depends on the physiological state of the cells as well as on environmental factor such as temperature, salinity, nutrient concentrations, light intensity, grazing by zooplankton, etc., and on the presence of toxic compounds in the medium (Jensen, 1984; Zhou and Wangersky, 1989a). In this sense, there has been much speculation about the influence of trace metals on primary productivity in the marine environment. A deficiency of the bioactive trace metals Mn, Fe, Co, Ni, Cu and Zn, may limit oceanic plankton production (Brand et al., 1983) and an excess of certain of these same metals may inhibit plankton growth (Brand et al., 1983; Sunda, 1988/1989). Based on the results of bioassay experiments (Sunda and Guillard, 1976; Anderson and Morel, 1978,1982; Sunda and Ferguson, 1983; Folsom et al., 1986) and theoretical models (Jackson and Morgan, 1978; Hirose and Sugimura, 1983; Wood and Wang, 1985) it is assumed that the chemical parameters controlling metal-organism interactions are the free metal ion activities and not the total metal concentration. Thus, chemical speciation plays an important role in the bioactive metals in oceanic chemistry, which controls plankton production and influences plankton species composition in the ocean (Morel and Hudson, 1985). The oceanic chemistries of the bioactive trace metals and their potential influences on oceanic plankton production and species composition have been the subject of recent discussions and reviews (Martin and Fitzwater, 1988; Martin and Gordon, 1988; Sunda, 1988/ 1989; Martin et al., 1989, 1990, 1991; Bruland et al., 1991).

In this paper an overview is presented of the role

M. Gonzdez-DrivilalMarine Chemistry 48 (1995) 215-236 217

of phytoplankton-trace metal interactions in regulating the ion concentrations. Recent evidence is described regarding the binding of metals to the surface of the algae as well as to the ligands produced and excreted by algae. The methodology of the determination and characterization of the different concentrations of the species involved in such processes are concerned with the following aspects:

first, the kinetics of uptake, which is generally described as a biphasic mechanism; and

second, the binding of heavy metals to the algal surface and to the exudate describing different models of surface complex formation that are applicable to data interpretation and that permit a generalization on the affinity of algal surfaces as a function of pH and solution variables.

To account for these studies, new data of experiments with Dunaliella tertiolecta and trace metals in natural seawater are given throughtout this paper (Gonzalez-Davila et al., 1994; Santana- Casiano et al., 1994) together with data from previous work in these areas.

2. Kinetics of uptake

When dealing with binding of metals on algae it is useful to understand the nature of their interaction. Ion-exchange phenomena based on electro- static interaction occur with some metals, e.g. alkali, alkaline-earth metals and transition metals (Crist et al., 1988, 1990, 1992). The release of protons, when other metals, e.g. Cu, are adsorbed demonstrates additional covalent-type bonding for this transition metal (Van Cutsem and Gillet, 1981; Van Cutsem et al., 1984; Crist et al., 1988). Redox reactions with certain noble metals (Au) also occur (Greene et al., 1986; Watkins et al., 1987). To account for these processes, some kinetics studies have been carried out following the proton and metal uptake on organic surfaces.

Biological surfaces acquire their charge through the dissociation of H+ from the active functional groups. The charge is governed by a dissociation constant of each functional group and, therefore, is dependent on the H+ concentration in solution. Phytoplankton cells contain various functional

groups, such as carboxylic, amino, thio, hydroxo and hydroxo-carboxylic, that can interact with metals ions. The number of functional groups on the biosurface is constant, which yields a certain maximum negative surface charge, i.e. maximum proton adsorption capacity. When the pH of an algal suspension is decreased, a proton uptake process occurs and metals are released. The rate of this proton uptake is characterized as a two phase process, a very fast “surface” uptake and a slow diffusion of protons into the dense phase of the cell wall which follows a first-order kinetics. Values of kH can be calculated using Eq. (1):

ln(Co)/(Ct) = krrt (1)

Crist et al. (1992) have shown: (1) the effect of pH on the first-order rate constant is much higher for Na+ than for Ca2+ exchange; (2) proton uptake occurs at a considerably faster rate for marine algae compared to Vaucheria, which is representative of freshwater algae; and (3) the rate of Na-exchange is more susceptible to pH changes. All these results demonstrate that the higher Nat content at interior sites in marine relative to freshwater algae controls the process and provides a metal that can diffuse faster than the divalent cations Ca2+, Mg2+. Furthermore, differences in algal microstructure, such as porosity and chemical composition of cell walls, may also account for these faster rates (Kaufman et al., 1980; Kahl et al., 1986). The transport of charge- compensating counterions through the cell wall to the solution controls the process. In accordance with Crist et al. (1992), it seems reasonable to assume that the driving force for proton uptake results from an activity difference:

This difference produces a chemical potential for proton uptake which can occur, however, only with the compensating transfer of positive ions (Na+, Ca2+, Mg*+, etc.) out of the alga. In basic solutions, these activities are reversed, and protons move out of the algae, presumably with an in-flow of, amongst others, Na+, Ca2+ and Mg2+.

When metals are added to seawater containing algal cells, it is well documented (Khummongkol et al., 1982; Gadd, 1988; Xue et al., 1988; Harris

218 M. Gonzcilez-D&la/Marine Chemistry 48 (1995) 215-236

and Ramelow, 1990; Huang et al., 1990; Xue and a relatively fast adsorption step (pseudo- Sigg, 1990; Garnham et al., 1992) that the uptake equilibrium) and then approaches equilibrium kinetics consist of two phases (Fig. 1). First, a rapid at a much reduced rate. This second step step of uptake to a surface ligand. This process has increases with the square root of time been found to be independent of light, and [cu(II&orbed 0: 0.106.10-’ Jt, thus confirming unaffected by a temperature reduction or the that this secondary slower uptake is a diffusion- presence of the uncoupler CCCP (carbonyl controlled process (Crank, 1975). It is important cyanide 3-chlorophenylhydrazone) and defined as to observe that the initial uptake uses inorgan- the biosorptive phase. Second, a relatively slower ically complexed copper from the solution (the phase due to active uptake or intracellular uptake labile copper decreases while the organically by passive diffusion. This second phase is reduced complexed copper is not affected by the incubation by lowering the temperature and by incubation time). The slower step decreases the amount of either in the dark or in the presence of CCCP in organically complexed copper showing that the light. The intracellular-metal locations can further metal uptake by the lie cells is regulated account for the binding to cellular and organellar by the intracellular metabolic processes. These membranes, polyphosphates (Jensen et al., 1982) different interactions and locations indicate a and/or metal binding peptides (phytochelatins) mechanism to regulate the free metal ion concen- known to be induced in algae in response to tration inside the cell as well as for detoxification heavy metal exposure (Robinson, 1988). Fig. 1 (Jones and Gadd, 1990) possibly by an inflow of shows the changes in the inorganic and total protons from the vacuolar membrane to cytosol. dissolved copper as a function of the time for the This multiphasic uptake kinetics may indicate the uptake of copper (6.35. 10e7 M) in an algal- existence of parallel uptake mechanisms. Different seawater suspension of 2.42. lo7 cells/l of mechanisms can be sited on different plasma Dunaliella tertiolecta (Gonzalez-Davila et al., membranes, such as the plasmalemma and 1994). The initial uptake of copper (5-10 min) is tonoplast (Laties, 1969) and thus, the multi-

0 50 100 150 200 250

TIME (mm)

Fig. 1. Kinetics of Cu(I1) binding and uptake by Dunaliella tertiolecta at a cell density of 2.42. lo7 cells/l in seawater. The kinetic process is described in two phases. Adsorption on the cell surface groups takes place in a few minutes. In the second phase, a slow decrease during 3 or 4 h is observed, due to the diffusion controlled process into the inside of the cell.

M. Gonzdez-DrivilalMarine Chemistry 48 (1995) 215-236 219

affinity uptake system for each metal depends on the external concentration that is present (Epstein, 1976; Kotyk, 1989). Quantitative models that address this basic mechanism have been developed by Khummongkol et al. (1982) and refined by Ting et al. (1989, 1991). They assumed a linear relationship between the metal concentration at the cell surface and in the solution, which is valid in very diluted solutions, and did not take into account the competitive effect between the surface groups on the cell wall and the ligand excreted into the solution.

3. Models of binding of metals to algal surfaces

The interaction between an ion and a solid surface is generally described by the electrical double layer theory. Several surface complex formation models have been derived and successfully applied to systems of metal adsorption on solid oxides. Sigg (1987) has indicated that adsorption equilibria in biological system are, to a great extent, similar to those of an oxide surface. However, there are some differences between these two particle groups. For example, most biological surfaces are more hydrophilic than oxides. As a result, water molecules are much more tightly bound to the biological surface than to the oxide surface.

Attempts to describe and predict the adsorption of metal ions by aqueous particles have treated surface sites as homogenous groups with a single stability constant using simple adsorption isotherms (Freundlich or Langmuir isotherms) or with a single intrinsic constant using a surface complex model. Various attempts have been made to model adsorption behaviour of metal ions onto heterogeneous solids using modified adsorption isotherms.

The interaction of metal ions with proteins on biological surfaces is quite similar to the hydrogen equilibrium (Steinhart and Reynolds, 1969), since protein binding sites are almost the same for metal ions and hydrogen. Metals ions become coordinated to the algal surface, that is, to complex forming groups on this surface. In natural seawater, due to the high ambient metal

concentration most of the sites will be bonded with proton at low pH or with calcium, magnesium and sodium at high pH. When metal ions like Cu2+, Mn2+, Zn2+, Ni2+, Cd2+, Fe3+. Pb2+, etc. are present or added to this solution, metal ions are sorbed on the alga and protons and metals are released. Thus, equilibrium constants determined in the adsorption process (the assimilation or active uptake of trace metals is not considered in this review) for the various metals are essentially adsorption at specific sites and ion exchange constants for Ca, Mg and protons bonded in the several kinds of ligands to be found on the cell surface. However, in the presence of anionic complexing ligands (exudates and natural organic materials), the adsorption characteristics of metals can change significantly. Some metal-ligand complexes are strongly bound by surfaces while others form non-adsorbing complexes in solution, in which case the ligands compete with the surface for coordination of metal ions. In addition to occupying surface sites, and possibly also blocking adjacent sites, the adsorption of ligands onto the surface alters the surface charge. Moreover, metal ions may form complexes with the surface either directly or via previously bound ligands or may form complexes in solution and can cause desorption of metals from the surface. Thus, adsorption or complexation on specific sites or through ion exchange reactions are going to be involved, both on the cell surface groups and with dissolved ligands. In such a condition, it is totally impossible split the contribution of each compartment to the change in the ambient metal concentration and in the pH (seawater shows a high buffer capacity) in order to establish the stochiometry of the ion exchange process. Crist et al. (1990, 1992) have shown that, in natural waters, the sorption process (adsorption plus assimilation) is indeed an ion exchange process with the same amount of metal assimilated as the sum of Na+, Ca2+, Mg2+ and H’ released. In seawater, the adsorption process may be described by surface complexation formation or adsorption equilibria where metal ions and protons are exchanged and where the master variable is pH. The following reactions for the surface reaction of metals may be generalized according to the

220 M. Gonzdez-DcivilalMarine Chemistry 48 (1995) 215-236

concept of surface complexation (Schindler and Stumm, 1987) in terms of metal-proton exchanges:

5 SOH + M2+ %j_ SOM+ + H+ (2)

5 SOH; + M2+ %$= SOM+ + 2H+ (3)

and:

= SO- + M2+ w$= SOM+ (4)

The degree of surface protonation depends on the acid-base equilibria:

5 SOH; HZ SOH + H+; Ks,i (5)

= SOH ++= SO- + H+; Ki2 (6)

For the reaction of biological substances with metal, the equations may be considered similar to that of metal oxide except algal cell surface can be regarded as a polyfunctional macromolecule with multiple binding sites. Amino, phosphate, sulfhydryl, carboxyl o hydroxy groups, and the functional side chain of amino-acid residues, such as histidine, cysteine, aspartic acid and glutamic acid are major potential adsorption sites of phytoplankton cell surface for metal ions (Sigg, 1987; Mohl et al., 1988; Huang et al., 1990). Each functional group exhibits its specificity toward metal ions. For simplification, the reaction of the weakly acidic sites, rS-H, of the algal surface with a divalent metal is expressed in a general equation in which charge on the surface is not considered:

- S,-Hj + M2+ @= Si-M +jH+ (7)

where i represents the type of site (algal surface may be approximated similar to a surface with polyfunctmnal macromolecules), and j is macroscopic proton coefficient. Applying mass action law to this exchange reaction:

Kii = i E Si-M) WV {G S,-Hj} [M2’]

the the

(8)

where {_ Si-M} is th e concentration of the ith surface site occupied by metal (mol/g, algal dry wt); {Z Si-Hj} is th e concentration of the free ith site (mol/g); [H+] is the equivalent concentration of hydrogen ion in the aqueous phase (M); and

[M2’] is the concentration of metal ion in the aqueous phase (M).

The reaction is considered as a surface complex formation process, in which Kij represents an “adsorption equilibrium coefficient” which is independent of pH.

An equilibrium quotient, such as that given in Eq. (8) has been shown (Stumm and Morgan, 1981; Sposito, 1983; Buffle, 1988) to depend some- what on the charge of the surface which in turn depends on the extent of surface binding of metal ions and protons. However, as a first and useful approximation the surface may be treated as if it was uncharged (the potential of the surface sites, QJ~, equal zero) and use intrinsic equilibrium constants. Kg values obtained by this way are a semi-empirical expression for the average effect of metal ion binding to a polyfunctional surface, valid for a range of relative surface coverage, 0.

The surface sites may be considered a homogeneous system, i = 1, with the assumption that the maximum adsorbed metal concentration Imax, is constant and is independent of pH. The biological binding site may be considered a metal-occupied site, {= S-M}, proton-occupied sites, {= S-H}, {Z S-H;}, and unoccupied sites, { = S}. In solution, this proton-occupied and proton-unoccupied sites can be present as Ca2+, Mg2+, Na+-occupied sites which are exchanged with the trace metals to yield = S-M with the consequence release of protons and metals. In other words, ion exchange and complex formation may occur on the biological surface.

r max = {- S-}+{z S-H}+{- S-H:!}+{= S-M}

(9)

When the amount of metal added is less than that required for saturation of a site, the adsorption of metal can be expressed in terms of the free metal ion concentration and the hydrogen ion concentration.

I’={=S-M}= Imax P2+l 1 ; P+l I W’12

(10)

Ko 4 x + w2+1

If the site can exist in j states of protonation, = S- H, = S-H;? . . . and S-Hi, the total adsorption on

M. Gonzdez-D&la/Marine Chemistry 48 (1995) 215-236 221

the algal surface according to a single-site adsorption model is given by:

(11)

Using a double reciprocal plot (Langmuir plot), i.e. l/{=S-M} vs. 1/[M2+], Eq. (11) may be arranged to:

(12)

For any pH we can define Eq. (8) in term of a conditional constant, Kh,j, that is a function of the proton concentration:

{z S-M} KH.j = {= S_Hj}[M2+] (13)

Using Eq. (13) the well known reciprocal Lang- muir plot valid for a given pH is found:

(14)

By a mass law plot or Scatchard plot, i.e.

{= S-M}

w2+1 us. {- S-M}

Eq. (11) can be rearranged as:

{= S-M}

W2’l = Kh,jrmax - K&{T= S-M} (15)

which is valid for a given pH. The adsorption constants determined by these methods represent macroscopical adsorption equilibrium constants involving various adsorption mechanisms. The speciation of the metal in solution regulates the free metal ion concentration [M2+] and influences the resulting bound metal concentration { = S-M}.

The biological surface consists of several different types of adsorption sites. The algal surface groups consist in abundant major proton- dissociating groups and the presence of small numbers of strong binding sites containing nitrogen and sulphur atoms, in a similar way as those considered in determining the natural

complexing capacity of seawater (Buffle, 1988). When a metal ion is introduced into an algal system, adsorption may occur at any site. Those sites with greater affinity for the metal ions will be occupied first. In this case, Eqs. (14) and (15) may not give a linear relationship; rather more than one linear or even curvilinear relationship may occur.

Many researchers have used the multiple site concept and have developed the idea of split Langmuir and Freundlich isotherms for metal adsorption onto hydrous solids (Sposito, 1984; Stroes-Gascoyne et al., 1987). With the assumption that each specific site has a binding constant Ki,j, given by Eq. (8), for which corresponding capacity is rmax,i, and that the maximum adsorption on the algae is sum of all the maximum sites:

n

r max = c r max,i (16) i=l

The adsorption of a metal onto the ith site havingj values for the macroscopic proton coefficient is given by Eq. (17):

(17)

This is a multi-site binding model in which the total adsorption, l?, is a function of both hydrogen ion and metal ion concentration. This model can be solved by applying a non-linear optimization procedure (Wilkinson, 1988) to the experimental pH-dependent adsorption data with [H+] and [M2+] as the independent variables and Ima+ Ki,j, and j as adjustable parameters. For most cases, a two-site Langmuir isotherm, which classifies various sites into two groups, i.e. high affinity and low affinity, has been applied successfully. Eq. (17) is now expressed as:

If we use conditional stability constants Kn,i valid for a given pH, the two-site Langmuir isotherm

222 M. Gonzrilez-DrivilaiMarine Chemistry 48 (1995) 215-236

may be expressed as:

I’={=$M}= &ax,1 [M2+l GJ’ + [M*+l

Lx,2 [M*+l + R-IJ1 + w2+1

(19)

Kn,i values are equilibrium quotients capable of expressing the average effect of metal ion binding to a mixture of algal surface ligands. Fig. 2 shows the sequential binding of proton by the various functional groups of the algal surface, carried out by alkalimetric (acidimetric) titrations. pK values around 4-5, 8-9.5 and sometimes lower than 3 have been found, showing the presence of groups that protolyze around these pH values such as carboxylate anion systems, proton transfer of amino groups and protonation of sulfide groups, respectively. Using the potentiometric data, the total number of protons bound to both the suspension and the supernatant with no discrimination between the various kinds of acidic groups with which the proton may be associated and the average proton adsorption constant, can be computed.

12

10

8

z a

6

4

2-

From the viewpoint of an ion exchange reaction, the proton release is stoichiometrically related to the amount of metal ion adsorbed on the biological surface. Generally, the ratio released proton/adsorbed cation increases with pH (Kinnburgh et al., 1983) and more than one proton is released to the solution per metal ion adsorbed (Benjamin and Leckie, 1981). However, in most of the studies on biological surfaces, the ratio is lower than one (Huang et al., 1991). This may be explained by considering that the divalent metal can replace protons from the proton- occupied sites; thereafter, one or two protons will be released. Sometimes, the trace metal can bind with unoccupied sites or metal-occupied sites; therefore, there are no protons released. It confirms that ion exchange and complex formation may occur on this biological surface. This same conclusion has been pointed out by Parks (1982) who has indicated that usually fewer hydrogen ions are released than would be expected in the process of metal adsorption. This low exchange ratio may be also due to the presence of phosphate groups on the surface. Within the

-3 -2 -1 0 1 2 3 4

HCl 0 1 M added ml NaOH 0.1 M added ml

Fig. 2. Alkalimetric (acidimetric) titrations of Dunaliella tertiolecta in 0.7 M NaCl used to characterize the proton-binding capacity of

the algae.

M. Gonzdez-DlivilalMarine Chemistry 48 (1995) 215-236 223

normal pH range, the phosphate groups are ionogenic sites which bind metals such as Cu(I1) without releasing any protons (Barker, 1971).

Experimental data obtained from the titration of algal suspensions at constant pH with increments of metal ions (Fig. 3) and of algal suspensions in presence of metal ions with increments of acid (Fig. 4) can be used to evaluate the affinity of metal ions to the functional groups of the algal surface. The different metal concentrations involved have been determined by using polarographic (DPP) and voltammetric techniques, both DPASV and DPCSV together with a speciation model, ion selective electrodes carefully pretreated and preconditioned, flameless atomic absorption spectroscopy, X-ray microanalysis and scintilla- tion counter for radionuclides determination.

When an initial metal concentration (in this case we shall considered copper) is added to a seawater solution, which contains algal cells at a fixed pH and it is kept in touch until a pseudo-equilibrium time is achieved, the metal ion is adsorbed on the cell walls and is bound to the ligands produced and

1 2

1 0

0.8

t

excreted by the alga. The total concentration of copper added in a function of bound copper by algal surface (= S,-Cu) and by exudates (L,-Cu) is given by:

TCuadd = [Cu(II)les c {= s,-Cu} + c [L,-Cu] i n

(20) where the Cu(I1) equilibrium concentration, that is the labile copper is:

[Cu(Il)]eq = [Cu2+] + c [Cu(OH),] + [CuC03]

+/Cu(HC03)21 + [cu(co3);1

+ [CuC12]

and

TCudd = o-l [CU2+] +C{E Si+ZU}

(21)

+ c Lb-Cul (22)

in which o 1 [CU~+]/[CU(II)],~; the ionization

/_j_

calibratmn

/’ w//411

I 4

Fig. 3. Titration curves of algae (2.22-10’ cells/l) with copper in seawater showing the changes of the complexing capacity of seawater as

a consequence of the addition of algae. Curve I represents the calibration curve in microwave acidified seawater. Curve 2 is the natural

complexing capacity of seawater. Curves 3 and 4 represent the total dissolved and labile concentration of copper in the presence of the

algae. From the data, the amount of Cu(II) adsorbed may be determined. The concentrations of free Cu are obtained by correcting for

speciation (Eq. 21).

224

20 c

M. Gonzdez-D&da/Marine Chemistry 48 (1995) 215-236

I / /

0 adsorbed

v total dissolved

v . labile

. v

. labile without

PH

Fig. 4. Titration curves of Cu(II) (4.323.10-’ M) as a function of pH in seawater and in the presence of 1.72-10’ cells/l Dunaliella terriolecta. The difference between the total dissolved copper and labile copper gives the Cu complexed by exudates.

fraction coefficient is constant because the concentration of the inorganic ligands are not significantly affected by variations in the copper concentration. It must be noted that in most studies carried out on the adsorption of trace metals by biological surfaces, the metal bound to exudates has not been considered in the total mass balance or have been included in Eq. (21). The copper free ion concentration can be determined using an ion-selective electrode (in natural water) or using DPASV and DPCSV; a-values can theoretically be determined using speciation models (Miller0 and Hawke, 1992). It must be pointed out that labile copper determined by DPASV also includes quasi-labile organic- complexed copper. Analyzing the peak potential for the different metal ions with and without microorganisms at different pH values, the peak potential shifts gives us information of the presence or absence of labile metal organic complexes in the solution. If the potential shift does not correspond with the changes in the metal speciation, the contribution of the labile metal organic complexes can be evaluated using the DeFord-Hume’s method (Goncalves

et al., 1987):

a&+ 1ogW + P1 IL’]) (23)

p1 [L’] being the product of the conditional stability constant and the labile ligand concentration not bound to copper. In this case, the free metal concentration is determined by Eq. (24):

[CUllabile = [CU*‘](QI~~ + /A[L’l) (24)

In seawater this technique, in conjunction with DPCSV, can be successfully applied because it is possible to determine both labile copper at natural pH in both filtered and non-filtered algal cultures, and total dissolved copper, that is complexed plus labile, in acidic and UV or microwave treated seawater. Using Eqs. (20) to (24) all of copper species can be determined. The free copper concentrations (pCu) found in algal cultures (Dunaliella tertiolecta) in seawater, after 10 min equilibration time with copper concentrations between 50 nM to PM, ranged from 10 to 8, which is slightly higher than that found in natural seawater (Gonzilez- DLvila et al., 1994). However, if we consider the kinetic studies shown in Fig. 1, the increase in the

M. Gonzblez-DbvilalMarine Chemistry 48 (1995) 215-236 225

equilibrium time allows an intracellular metal- similar as if we tried to express the interaction of diffusion process to occur. As a consequence these a metal ion with a variety of soluble complex formers concentrations will decrease and will fall in the in terms of only one constant. Nevertheless, such a range of seawater concentration. It also shows constant is an average equilibrium quotient, obtained the kinetic control of both the excreted by applying the law of mass action to a mixture of compounds and the binding to algal surfaces, functional groups; it permits a generalization for a which have never been considered before. Metal given range of coverage of the surface by metal buffer techniques have been used in order to ions. Thus, phytoplankton cells appear to be an obtain copper free concentrations in the range important substrate for metal binding in natural lo-l2 to lo-” (Sunda and Guillard, 1976; Xue samples. After microalgae have accumulated metals and Sigg, 1990). Competitive behaviour between they may sink, thus transporting metals to deeper the complex metal buffer and the algal culture waters and sediments and/or be ingested by con- follows, as we will show below. sumers in surface waters (Fisher et al., 1983).

Table 1 shows the conditional equilibrium constants, K& considering homogeneous sites, obtained using Langmuir and Scatchard plots such as those shown in Fig. 5 (Langmuir plot; Eq. 14) and Fig. 6 (Scatchard plot; Eq. 15). Obviously, the data cannot be fitted readily; at high coverage of the surface with metal or at high metal concentrations, considerable deviation from a linear relationship is generally observed. This indicate that metal ions become bound first to the highest-affinity surface ligands and subsequently to those of lesser activity. Thus, the equilibrium constants Ksn obtained from the experimental data show a tendency to decrease with increased surface coverage. Constants evaluated from low surface coverage are of more general interest with regard to seawater. It must be kept in mind, however, that the representation of the extent of metal interaction with a cell surface in terms of one single equilibrium constant is a construct,

The experimental data can be fitted much better by both introducing factors that diminish the binding tendency with increasing surface coverage and by considering that adsorption data can be quantified by different classes of sites known to exist on the biological surface. According to the proposed two-site model (Eqs. 17-19) high- affinity parameters and low-affinity parameters have been estimated from Scatchard plots (Fig. 6) in accordance with Sposito (1984) and Stroes- Gascoyne et al. (1986) using non-linear regressions with computer iteration (Huang et al., 1991; Gonzalez-Davila et al., 1994) or using the MICROQL - UCR program (Xue et al., 1988). Some values for the four parameters derived from the two-site model are presented in Table 2. The site heterogeneity appears to be the major factor for the decreased adsorption energy when the adsorption reaction was in progress. It is apparent that the two-site model with four parameters is an

Table 1

Binding constants and binding capacities to biological surfaces

Algae Metal Media rmax (molig) log Ki Reference

C. rheinhardii CU (0.01 M KNOX, pH = 7)

C. rheinhardii cu (0.01 M KNO,,NTA, pH = 6.5)

D. tertiolecta cu (SW, 360/w, pH = 8.2)

Yeast S. cerevisiaed Cu (0.05 M NaC104, pH = 4.0)

Yeast A. oryzaea cu (0.05 M NaC104, pH = 5.0)

Yeast R. oryzaea cu (0.05 M NaC104, pH = 5.0)

D. tertiolecta Pb (SW, 36%0, pH = 8.2) C. vulgaris Cd (DW stand. culture, pH = 6.8) C. vulgaris Zn (DW stand. culture, pH = 6.8)

C. rheinhardii Cd (0.01 M KNO,, pH = 7)

’ Values for yeast are introduced for comparative purposes.

5.0. 1o-5

1.9.10-6

1.53.10-5

3.0.10~5

2.1.10-4

1.4.10-4

l.3.10-5 1.1.10m5

8.0. 1O-6

1.6

10.0

8.85

5.93

4.83

4.61

8.22

5.8

Xue et al. (1988)

Xue and Sigg (1990)

Gonzilez-DBvila et al. (1994)

Huang et al. (1990)

Huang et al. (1991)

Huang et al. (1991)

Gonzilez-DBvila et al. (1994)

Ting et al. (1991)

Ting et al. (1991)

Xue et al. (1988)

226 M. Gonzdez-D&la/Marine Chemistry 48 (1995) 215-236

log KHS = 8.85 * 0 04

-15 -I r max = 22.88 * 3.44 10 mol cell

J

0 2 4 6 ‘3

[CL@)]-’ 1O-7 M-’

Fig. 5. Langmuir plot (Eq. 14) for the adsorption of Cu(I1) on 2.22.10’ cells/l Dunaliella tertiolecta in seawater. At high metal

concentrations considerable deviation from a linear relationship is observed.

50 c

40 t

30 c

20 1 10 1

__ 2.5 -

‘5 ” 2.0 -

j 1.5 -

8

z 2

3

*-‘II

z 0.5

0.0 0 1 2 3 4 5 6

[C”(II)] lo7 M

0 .

0.

. .

\ .

-I

1 4

01 / I

0 5 10 15 20 25

IS-Cu{ 1ol5 (ml cell -1

)

Fig. 6. Scatchard plot (Eq. 15) for the data presented in Figs. 3 and 5. The presence of at least two ligands with different affinity for the

surface sites of the algae is demonstrated. Insert: Two-site Langmuir isotherm for Cu(II)-binding to algae. The curve is calculated

according to Eq. (19).

Table 2

M. Gonzilez-DbvilajMarine Chemistry 48 (1995) 215-236 221

Binding constants to the surface of algae according to a two-site Langmuir isotherm

Algae Metal log K; log K; I- max,l bok) r max.2 @Wg) Reference

C. rheinhardii CU 8.0 6.5 0.7.10-5 4.1.10-5 Xue et al. (1988)

D. tertiolecta CU 9.5 8.4 0.5.10~5 1.6.10-5 Gonz6lez-Divila et al. (1994)

Yeast S. cerevisiae Cu 7.3 5.7 1.1.10-5 1.9.10-5 Huang et al. (1990)

Yeast A. oryae CU 5.3 4.8 1.1.10-4 1.1.10-4 Huang et al. (1991)

Yeast R. oryzae cu 5.2 4.4 5.0.10-5 1.5.10-4 Huang et al. (1991)

D. tertiolecta Pb 8.4 7.2 0.4. 1o-5 1.3.10-5 Gonzblez-Dkila et al. (1994)

adequate method to predict the metal adsorption isotherm by algae, even though the algal surface is characterized by a multiple-site adsorption scheme. Taking into account that the total electronic energy of the 1: 1 type Cu(II)-complex is higher than that of the 1:2 type Cu(II)-complex (Takahashi and Tannaka, 1986) based on the two-site model, Cu(I1) adsorption sites can be divided in two major types: a monodentate (1:l) and a bidentate (1:2) type. The high-affinity sites can be assumed to be monodentate binding; the low-affinity sites for metal ions involved two protons sites, i.e. it is bidentate. This can be assumed if the number of maximum proton adsorption sites is in agreement with the maximum Cu(I1) adsorption sites, i.e.

However when metals such as Cu2+, Pb2+, Cd’+, A13+ are adsorbed on algal surface groups, not only protons but metals, Na+, Ca2+ and Mg2+ and other trace metals are released. Thus, the growth of phytoplankton in marine environments is also controlled by the activities of free metal ions other than copper, i.e. Co*+, Mn2+, Zn2+, Fe3+, etc. (Anderson and Morel, 1978; Jackson and Morgan, 1978; Hirose and Sugimura, 1983, 1985). Additions of copper result in an increase in the activities of toxic metal ions other than copper as a result of metal exchange reactions. A probable explanation for this release is the complex formation of Cu2+ with adjacent algal ammonium or carboxylic acid groups, which can be represented by the reactions:

Alg-(NH&+ + Cu2+ i [Alg-(NH2)2Cu]2+ + 2H+

Alg-(C02H2), + Cu2+ --f [Ah-(C02)$u] + 2H+

(26)

Some combinations of these reactions and other

reactions involving sulfhydryl groups are also possible.

The conditional adsorption constants estimated from the homogeneous and heterogeneous model are a function of both proton concentration and the apparent proton adsorption. Eq. (18) allows us to predict the metal adsorption constant from the conditional state to the intrinsic state, considering the proton competition with metal ions for the same sites. These intrinsic values are independent of pH. If the intrinsic adsorption constants are higher than the apparent adsorption constant of the proton, it indicates that the effect of the proton on the metal adsorption is insignificant. In contrast, if it is close to the apparent proton adsorption constant, the proton effect on metal adsorption is significant. Only in the study of Huang et al. (1991) the intrinsic adsorption constant has been taken into account.

Most of the studies of the interactions between phytoplankton and trace metals have focused on the effects of a single metal. The uptake of nutrient and toxic metal ions is a function of the concentration of other metals. In general, a mixture of metals can produce three possible types of behaviour: synergism, antagonism, and non- interaction. Different response parameters have been used as useful test criteria. These include cell number, cell growth (considered mostly), cell volume, 14C02 uptake, nitrogenase activity, dissolved oxygen, and metal uptake. Several reasons have been advanced for these synergistic or antagonistic actions. The synergistic behaviour showed for the combined effect of copper and nickel on the cell population of Chlorella vulgaris

(Ting et al., 1991) was postulated to be the result of an increase in membrane permeability of the algal cells. In contrast, antagonistic interaction was

228 M. Gonnilez-DbvilalMarine Chemistry 48 (1995) 215-236

reported by these same authors in the case of cadmium and selenium for the same alga. It was suggested that screening or competition for the binding sites on the cellular surfaces had resulted in the metal ions mutually ameliorating their individual toxic effects. In most cases, it is not known whether competition occurs at the cell- surface uptake sites or at intracellular metabolic sites. However, as metal chelating functional groups found in both membrane transport proteins and intracellular enzymes are never completely specific for one metal (Sunda, 1988/ 1989), competition at both sites is likely to occur.

Few researchers have employed metal uptake as the test criterion in studying algal/metal interactions. Copper uptake is inhibited by a number of metals including cobalt, nickel, cadmium and manganese (Mierle, 1983). Morel et al. (1991) have shown that there may be physiological tradeoffs with respect to phytoplankton trace element requirements. They showed that under certain circumstances, Zn*+ deficiency can be compensated by Cd2’. Braek et al. (1980) reported both antagonistic and synergistic actions for zinc and cadmium uptake by different algae. Ting et al. (1991) showed that these same metals do not compete for the binding sites on the surface components of the cell wall structure of the alga Chlorella vulgaris over the range of metal concentration studied by them (lower than 80 PM). However, the intracellular uptake of zinc is impeded by the presence of cadmium. The presence of cadmium promotes the reverse carrier reaction in the membrane transport of zinc. The difference in the behaviour of these two metals is likely to be due to the essential and non-essential nature of zinc and cadmium ions, respectively, for cellular metabolism. Other metal antagonisms observed from laboratory studies (except Sunda, 1986) following the effect over the cell number have been reported for Cu2+ and Mn2+ (Sunda et al., 1981; Murphy et al., 1984; Sunda, 1986), for Cu*+ and Fe3+ (Murphy et al., 1984; Harrison and Morel, 1983), and for M2+ and Zn2+ (Sunda, 1986). These antagonistic effects have been hypothesized by Sunda and coworkers (Sunda et al., 1981; Sunda, 1986; Sunda and Gessner, 1989) to inhibit the phytoplankton growth often

observed when deep water is brought to the surface. These synergistic and antagonistic interactions make it difficult to attribute community responses to any one perturbation. These inter- active effects of a mixture of metals on an aquatic organism are extremely complex and appear as an important field for future research. The change in metal ion ratios may produce unfavourable ratios for the growth rate of phytoplankton cells. In this same context, as has been pointed out by Murphy et al. (1984) and Bruland et al. (1991), some cases of primary production limitation apparently solely due to iron (Martin et al., 1989; Brand, 1991; Coale, 1991; Morel et al., 1991) could be due to antagonistic effects of other trace metals and iron. The growth stimulation observed in the Fe3+ addition experiments by Martin et al. (1989) and Brand (1991) may increase the low [Fe3+] : [Cu*+] ratios in seawater increasing the growth rates of phytoplankton cells. Alternatively, it has been suggested that added Cu2+ may have displaced nutritionally limiting iron from organic ligands, thereby making it more available to phytoplankton. In such conditions, one of the possible mechanisms to maintain free metal ratios within favourable boundaries is that phytoplankton may reduce metal concentrations by producing strong metal-complexing organic ligands. The study of metal-organic interactions in seawater is therefore important for establishing the significance of such interactions to geochemical cycling and biological availability of metal ions in the natural water and seawater systems.

4. Effect of complexing ligands released by marine phytoplankton

The ability of microorganisms to release into the environment organic metabolites capable of complexing metals has been known for some time (Fogg and Westlake, 1955; Neilands, 1967, 1973). This complexation is important since it can enhance the bioavailability of metals required for growth or, conversely, reduce the ionic concentration of deleterious metals such as copper (Gnassia-Barelli et al., 1978; Butler et al., 1980; Sueur et al., 1982; Gordon and Millero, 1984,

M. Gonzblez-DcivilajMarine Chemistry 48 (1995) 215-236 229

1987; Hirose and Sugimura, 1985; Wangersky, 1986; Schreiber et al., 1990; Xue and Sigg, 1990). In other words, the action of the metal-organic complexes in marine environments may result in a metal buffering capacity, which means that the activities of the free metal ions are kept constant under seawater conditions. Because of the great abundance of phytoplankton species and their role as the first link of the marine food chain, studies of the products of their metabolism and of their interaction with other constituents in natural aquatic systems are of particular interest in the complex investigations of biogeochemical processes in the natural environment. Three basic types of chemical interactions have been noted:

(1) the oxygen-metal-oxygen bonding, found in hydroxamic acids, cathechols and sideromy- tins (Byers and Arceneaux, 1977; Snow, 1970);

(2) the oxygen-metal-nitrogen bonding found in mycobactins (Neilands, 1967) and

(3) the oxygen-metal-sulfur bonding found in fluopsins.

Moreover, the discovery that at least certain microorganisms produce and release siderophores (Simpson and Neilands, 1976; Bailey and Taub, 1980) is very important in studies on metal toxicity because these compounds are capable of affecting speciation and thus toxicity when complexing metals such as copper.

The production of extracellular organic compounds by phytoplankton algae depends on the physiological state of the cell as well as on environmental factors and the presence of toxic compounds in the medium. Jardim and Pearson (1984) found that, when stressed with copper, P. horyanum and A. cylindrica produce more complexing material than under usual growth conditions. In batch cultures of the diatom Phaedactylum tricornutum, Zhou and Wangersky (1985) found that organic material was released to the culture medium at all growth stages. They also recorded diurnal fluctuations in both the amount and copper complexing capacity values. However, when dealing with the study of the complexing capacity of natural seawater and with culture media, analytical problems in the determination of organic matter arise, mainly

because this organic matter represents a complex mixture of different classes of compounds. Methods using analytical techniques, such as HPLC coupled with FTIR and MS acting as detectors, must be developed in order to investigate the different individual extracellular compounds. Electrochemical methods can be generally applied for the determination of surface active substances on the basis of measurements of adsorption effects on the electrode surface. The adsorption effects measured at electrodes represent the gross effect of a complex mixture of surfactants present in a natural sample. Thus, the surfactant activities of marine and freshwater samples and of phytoplankton have been successfully determined on the basis of capacity current measurements using phase-sensitive ac. polarography (Cosovic and Vojvodic, 1982, 1987; Kozorac et al., 1989). As expected, cultures of alga grown on different media had different surfactant activity. According to the shape of the capacity current-potential curves, not only the quantity but also the composition and type of surface- active material vary. In stress conditions, phytoplankton cells excrete a completely different material from that in normal conditions where there are no limitations of essential trace metals or chelators. Taking into account the studies of Cosovic and Vojvodic (1982) for capacity current- potential curves of some selected surface-active substances representative of different types of surface-active materials such as lipids, proteins, carbohydrates and humic material, similar to natural constituents of seawater, it can be shown whether the predominant compounds in the non- filter samples are composed of more hydrophilic surfactants, e.g. humic and fulvic acids, or correspond mainly to fatty material such as, for example, unsaturated fatty acids like oleic, linoleic, etc. Zhou and Wangersky (1985) and Kozorac et al. (1989) have found changes in material released during phytoplankton growth attributed to real changes in the quantity and quality of organic material. The application of this fast technique to systems where different metal ions are being added to natural seawater samples could determine the effects in the composition of the excreted ligands by the phytoplankton.

230 M. Gonzdez-D&la/Marine Chemistry 48 (1995) 215-236

The conditional stability constants, as well as metal complexing capacity, have been understood as physicochemical parameters for the roles of metal-organic complexes in natural waters. These values give some idea as to how the ionic species of the metal would be affected by micro- to nano- molar concentrations of complexing material. The complexing capacity of a sample is a measure of the metal-buffering capacity and is of fundamental importance for a quantitative assessment of the fate of polluting metals in natural waters. Temporal variability in complexing capacity may influence the succession of phytoplankton species in natural waters.

Taking into account that seawater and plankton cultures may contain a variety of organic ligands with different complexing strengths and concentrations, different analytical methods will determine metal-complexing organic ligands of different strengths and concentrations because the different methods have different detection windows. Thus, determination of organic complexation at any single detection window may give only partial information on the true concentration of organically complexed and other forms of metals in algal cultures. Since copper forms stable complexes and is also an essential metal it is one of the ions of choice for the determination of the complexing capacity. Table 3 shows the Cu-complexing parameters of

seawater and exudates excreted by phytoplankton cells. Studies with other metals are lacking and should also be considered in future work. Only a complete study of the nickel(I1) complexing capacity in Narraganset Bay (Muller and Kester, 1991) has been carried out and included in Table 3. Different techniques, such as fixed-potential amperometry (Waite and Morel, 1983), bacterial bioassay (Sunda and Ferguson, 1983; Sunda et al., 1984; Hering et al., 1987) DPASV (Plavsic et al., 1982; Kramer and Duinker, 1984; van den Berg, 1984; Gonzalez-Davila and Millero, 1990; Gonzalez Davila et al., 1990) DPCSV of Cu- catechol complexes and Cu-tropolene complexes (van den Berg, 1984; Donat and van den Berg, 1992), have been used to determine the total ligand concentrations, conditional stability constants and ambient pCu in seawater. As can be seen in Table 3, the concentrations and copper- complexing strengths of natural ligands vary, not only with the places where the natural and seawater were taken up but also with the method used. Thus, for the same sample of North Sea seawater determined by tropolene and catechol DPCSV methods, tropolene appears to determine copper- complexing ligands that exist at higher concentrations but that form weaker complexes than those detected using catechol. The different results on the determination of organically-complexed copper and other metals are important because

Table 3

Parameters controlling the free metal-complexing capacity in both natural estuarine and oceanic waters and in algal cultures

Media (referencea)

Atlantic (1)

North Pacific (2)

Lower Newport River estuary (3)

Narragansett Bay (4)

North Sea (5)

North Sea (5)

D. tertiolecta f/2 culture SW (6) C. rheinhardii (0.01 M KNOs , NTA) (7)

Macroalgae (SW) (8) D. tertiolecta Gran Canaria SW (6)

Narragansett Bay values for Zn2+ (9)

CL,1 log Kcond(M),l log &md(M).Z 1% K;

@Ml

4-36 23-99 10.2-12.4 8.3-9.2

1.8 1.6 12.9 9.5510.6

110~300 8.779.6

50 100 12.4 10.0

16 12.4

8 5.6

950 8.5

540669540 0.9-1.9.10@ 8.6611.0 8.4-10

270%1010 10.2-9.8

255 1.53.10-5 9.30 8.85

446 40-100 2.20. 1O-7 M 9.3-9.4 7.5-8.3 6.9667.90

a (1) = Buckley and van den Berg, 1986; (2) = Coale and Bruland, 1988; = (3) Sunda et al., 1984; (4) = Sunda and Hanson, 1987;

(5) = Donat and van den Berg, 1992; (6) = Gonzalez-Davila et al., 1994; (7) = Xue and Sigg, 1990; (8) = Sueur et al., 1982;

(9) = Muller and Kester, 1991.

M. Gonzdez-D&da/Marine Chemistry 48 (1995) 215-236 231

their implications for trace metal speciation measurements in freshwater, coastal and oceanic samples. Measurements reported by Coale and Bruland (1988, 1990) and Hirose (1998) indicate that 95-99% of the total dissolved Cu(I1) in surface seawater is bound by strong organic complexes, primarily by the stronger (L,) of the two Cu-complexing ligands. These strong complex-formers together with metal sorption processes lower the free Cu2+ concentration in oceanic surface waters to levels that would not inhibit their growth. In deeper seawater, where the photic zone disappear and amount of Li copper-complexing ligand is decreased (Coale and Bruland, 1990) the free copper concentration increase about 2000 fold.

With regard to complexing characteristic of algal exudates, Table 3 shows that various amounts of complexing ligands are released by different algae, both macro- and microalgae, in seawater. On a dry weight basis and over a range of free copper concentration of lo-” to 10e9 M, similar results are found for macro- and microalgae in seawater. It must be pointed out that complexation of trace metals by natural organic ligands produced and excreted by phytoplankton are not the sole sources of metal complexing ligands. There is evidence that some marine bacteria (Schreiber et al., 1990) and marine fungi (Sunda and Gessner, 1989; Huang et al., 1990, 1991) also adsorb and produce ligands that strongly complex Cu2+, although the relative importance of these groups in affecting metal speciation in the open ocean is as yet unknown. Differences in the conditional complexation constants for algal exudates is related to the different complexing agents released by algae. Also, the increase of the conditional constants for copper complexation with exudates as a function of pH is related to the acid-base properties of the exudate functional groups. The range of 8.5 to 11 found for the complexation constants is very high, and similar for the log K’ of EDTA in seawater and slightly higher than the one reported by Mantoura et al. (1978) for humic substances (9.3) in seawater. However, only Moffet et al. (1990) provided the first direct laboratory evidence for production of a strong Cu-binding ligand by marine photoauto- troph (the cyanobacterial genus Synechococcus).

This ligand has a Cu-complexing strength identi- cal to that of the stronger ligands, Li, shown in Table 3. In other studies, ligands with complex strength lo* like L2 have been found. Taking into account the different complexing ligands described, further work on the characterization of the exudates in terms of molecular weight, chemical properties and metal complexing behaviour must be carried out. The release of complexing ligands is a mechanism for the removal of metal from the cells and may act as a detoxifying mechanism of the algae that prevent the accumulation of metal in the cells. In this sense, we have found that the addition of copper stimulates the production of extracellular ligands and that a maximum is reached at a certain level of added copper in accordance with Xue and Sigg (1990). However, other factors, such as culture medium characteristics, existence of stressed conditions such as those found during the mixing of fresh and saline water in estuarine waters (Kozorac et al., 1989), the lack of trace nutrient metals, and the physiological state and photosynthetic activity of the algae (Kozorac et al., 1989; Zhou and Wangersky, 1989b), influence production and release of exudates. Such factors determining the release and characteristics of exudates require further evaluation and will be the object of future work on metal-binding on algae both to the surface and to the exudates.

5. Conclusions

Biological production in the oceans can strongly influence the oceanic chemistry of trace metals. In turn, bioactive trace metals may limit oceanic plankton production and plankton growth. As a result of the advances in our knowledge of the oceanic concentrations, distributions and cycles of trace metals we conclude that both the binding to surface and to extracellular ligands must be considered in studying the interactions of trace metals with algae. Phytoplankton influences the residual concentrations of metal ions both by metal binding to algal surfaces and subsequent intracellular transport and by the production of strong ligands that affect the free concentration of metal ions. The adsorption of metals is readily

232 M. Gonzblez-D&da/Marine Chemistry 48 (1995) 215-236

interpreted in terms of surface complex formation equilibria in multi-metal multi-ligand speciation calculations. In assimilation studies, the counter flow of cations in the ion exchange mechanism makes the reverse reaction feasible. This second effect acts as a mechanism to regulate the free metal ion concentration inside the cell. The factors determining the importance of these phenomena require further evaluation. The specific extracellular products of aquatic organism which can form complexes with trace metals have not been well characterized. When they have been identified they have not generally been quantified in a manner such as the molar concentrations, required for thermodynamic speciation models. Also, the kinetics of the binding to algal surface ligands, the intracellular uptake and the excretion of exudates in the oceanic time scales should be considered.

However, the interaction between phytoplankton and metal ions in seawater is not an isolated process as it has been considered in the most laboratory studies. Many trace metals together with different kinds of phytoplanktonic cells (microorganisms in general) and their exudate products are continuously interacting in the oceans. The uptake and metabolism of nutrient metal ions is a function of the concentration of other metals, both nutrient and toxic metal ions, which often act in a competitive or antagonistic manner. Extensive field studies of trace metal interactions accurately describing metal-biota interactions are lacking. These studies should be focused on the understanding of the influence of bioactive trace metals and their speciation on the levels of primary productivity and species composition. However, these studies should consider not only the effects on the phytoplankton growth but also on the sorption and competitive processes and on the metal-induced production of complexing exudates. Metal titrations involving different metals, such as Cu2+, Mn2+, Fe3+, Pb2+, Co2+, Ni2+ and Zn2+, should be carried out in the water column to take into account changes both in the production of organic exudation ligands and in the adsorption and assimilation of metal ions as a function of depth, pH, temperature, salinity and light intensity. Laboratory studies must reproduce the

oceanic conditions and concentrations of both metal ions and biological population. These experiments must consider the effects on the binding and uptake of (1) different individual metal ions on each alga species, (2) different metal combinations, and (3) the effect of temperature, pH and salinity. Only when all of these aspects are being considered, the influence of biological communities on the oceanic chemistry will begin to be understood.

Acknowledgements

I thank Jesus Perez-Pefia and Magdalena Santana-Casiano for their special and essential contributions to this paper. Special thanks are extended to Frank J. Miller0 for critical review and helpful suggestions.

References

Anderson, D.M. and Morel, F.M.M., 1978. Coppper sensitivity of Gonyalaaax tamerensis. Limnol. Oceanogr., 23: 283-295.

Anderson, M.A. and Morel, F.M.M., 1982. The influence of aqueous iron chemistry on the uptake of iron by the coastal diatom Thalassiosira weissjogii. Limnol. Oceanogr., 27: 789-813.

Bailey, K.M. and Taub, F.B., 1980. Effects of hydroxamata siderophores (strong Fe(II1) chelators) on the growth of algae. J. Phycol., 16: 334-339.

Barker, R., 1971. Organic Chemistry of Biological Compounds. Prentice-Hall, Englewood Cliffs, NJ.

Benjamin, M.M. and Leckie, J.O., 1981. Multiple-site adsorption of Cd, Cu, Zn, and Pb on amorphous iron oxyhydr- oxide. J. Colloid Interface Sci., 79: 2099221.

Braek, G.S., Malnes, D. and Jensen, A., 1980. Heavy metal tolerance of marine phytoplankton. IV. Combined effect of zinc and cadmium on growth and uptake in some marine diatoms. J. Exp. Mar. Biol. Ecol., 42: 39-54.

Brand, L.E., 1991. Minimum iron requirements of marine phytoplankton and the implications for the biogeochemical control of new production. Limnol. Oceanogr., 36: 17.56- 1771.

Brand, L.E., Sunda, W.G. and Guillard, R.R.L., 1983. Limita- tion of marine phytoplankton reproductive rates by zinc, manganese, and iron. Limnol. Oceanogr., 28: 1182-l 195.

Bruland, K.W., Donat, J.R. and Hutchins, D.A., 1991. Inter- active influences of bioactive trace metals on biological production in oceanic waters. Limnol. Oceanogr., 36: 1555-1577.

M. Gonzilez-DcivilajMarine Chemistry 48 (199s) 215-236 233

Buckley, P.J.M. and van den Berg, C.M.G., 1986. Copper complexation profiles in the Atlantic ocean. A comparative study using electrochemical and ion exchange techniques. Mar. Chem., 19: 281-296.

Buffle, J., 1988. Complexation Reactions in Aquatic Systems: An Analytical Approach. Ellis-Horwood, Chichester.

Butler, M., Haskew, A.E.J. and Young, M.M., 1980. Copper tolerance in the green alga, Chlorella vulgaris. Plant Cell Environ., 31: 119-126.

Byers, B.R. and Arceneaux, J.E.L., 1977. Microbial transport and utilization of iron. In: E.D. Weinberg (Editor), Micro- organisms and Minerals. Dekker, New York, NY, Micro- biol. Ser., 3, pp. 215-249.

Coale, K.H., 1991. Effects of iron, manganese, copper and zinc enrichments on productivity and biomass in the subarctic Pacific. Limnol. Oceanogr., 36: 1851-1864.

Coale, K.H. and Bruland, K.W., 1988. Copper complexation in the Northeast Pacific. Limnol. Oceanogr., 33: 1084-1101.

Coale, K.H. and Bruland, K.W., 1990. Spatial and temporal variability in copper complexation in the North Pacific. Deep-Sea Res., 37: 317-336.

Cosovic, B. and Vojvodic, V., 1982. The application of ac polarography to the determination of surface active sub- stance in seawater. Limnol. Oceanogr., 27: 361-369.

Cosovic, B. and Vojvodic, V., 1987. Direct determination of surface active substances in natural waters. Mar Chem., 22: 363-375.

Crank, J., 1975. The Mathematics of Diffusion. Oxford Univ. Press, New York, NY.

Crist, R.H., Oberholser, K., Schwartz, D., Marzoff, J., Ryder, D. and Crist, D.R., 1988. Interactions of metals and protons with algae. Environ. Sci. Technol., 22: 755-760.

Crist, R.H., Martin, J.R., Guptill, P.W., Eslinger, J.M. and Crist, D.R.. 1990. Interaction of metals and protons with algae. 2. Ion exchange in adsorption and metal displace- ment by protons. Environ. Sci. Technol., 24: 337-342.

Crist, R.H., Oberholser, K., McGarrity, J., Crist, D.R., Johnson, J.K. and Brittsan, J.M., 1992. Interaction of metals and protons with algae. 3. Marine algae, with emphasis on lead and aluminum. Environ. Sci. Technol., 26: 4966502.

Donat, J.R. and van den Berg, C.M.G., 1992. A new cathodic stripping voltammetry method for determining organic complexation in seawater. Mar. Chem.. 38: 69-90.

Duursma, E.K., 1961. Dissolved organic carbon, nitrogen and phosphorus in the sea. Neth. J. Sea Res., 1: 1-148.

Epstein, W., 1976. Kinetics of ion transport and the carrier concept. In: U. Luttge and M.G. Pitman (Editors), Encyclo- pedia of Plant Physiology, IIB. Springer, Berlin, pp. 93-126.

Fisher, N.S. and Fabris, J.G., 1982. Complexation of Cu, Zn and Cd by metabolites excreted from marine diatoms. Mar. Chem., 11: 245-255.

Fisher, N.S., Bjerregard, P. and Fowler, S.W., 1983. Inter- actions of marine plankton with transuranic elements. 1. Biokinetics of neptunium, plutonium, americium and californium in phytoplankton. Limnol. Oceanogr., 29: 432-440.

Fogg, C.E. and Westlake, D.F., 1955. The importance of extracellular products of algae in freshwater. Proc. Assoc. Theor. Appl. Limnol., 12: 219-232.

Folsom, B.R., Popescu, N.A. and Wang, J.M., 1986. Compara- tive study of aluminum and copper transport and toxicity in an acid-tolerant freshwater green alga. Environ. Sci. Technol., 20: 616-620.

Gadd, G.M., 1988. Accumulation of metals by microorganisms and algae. In: H.J. Rehm (Editor), Biotechnology, 6b. VCH. Weinheim, pp. 401-433.

Garnham, G.W., Codd, G.A. and Gadd, G.M., 1992. Kinetics of uptake and intracellular location of cobalt, manganese and zinc in the estuarine green alga Chloreila salina. Appl. Microbial. Biotechnol., 37: 270-276.

Gnassia-Barelli, M., Romeo, M., Laumond, F. and Pesando, D., 1978. Experimental studies on the relationship between natural copper complexes and their toxicity to phytoplankton. Mar. Biol., 47: 15-19.

Goncalves, M.L.S. and Lopes da Conceicao, A.C., 1989. Metal ion binding of Copper (II), Zinc (II), and Lead (II) by the algae Selenastrum capricornutum Printz. Sci. Total Environ., 38: 1555166.

Goncalves, M.L.S., Sigg, L., Reutlinger, M. and Stumm, W., 1987. Metal ion binding by biological surfaces: voltammetric assessment in the presence of bacteria. Sci. Total Environ., 60: 105-l 19.

Gonzalez-Davila, M. and Millero, F.J., 1990. The adsorption of copper to chitin in seawater. Geochim. Cosmochim. Acta. 54: 761-768.

Gonzalez-Davila, M., Santana-Casiano, J.M. and Millero, F.J., 1990. The adsorption of Cd(I1) and Pb(I1) to chitin in seawater. J. Colloid Interface Sci., 137: 102-l 10.

Gonzalez-Davila, M., Santana-Casiano, J.M., Perez-Pefia, J. and Millero, F.J., 1994. The binding of Cu(I1) to the surface and exudates of the alga Dunalieila tertiolecta in seawater. Environ. Sci. Technol, in press.

Gordon, A.S. and Millero, F.J., 1984. Electrolyte effects on attachment of an estuarine bacterium. Appl. Environ. Microbial. Mar., 47: 495-499.

Gordon, A.S. and Millero, F.J., 1987. Sorption of copper by a marine vibrio in seawater. Va. J. Sci., 38: 194-199.

Greene, B. and Darnall, D.D., 1988. Temperature dependence of metal ion sorption by Spirulina. Biorecovery, 1: 27-41.

Greene, B., Darnall, D.W., Alexander, M.D., Henzl, M.T., Hosea, J.M. and McPherson, R.A., 1986. Interaction of gold(I) and gold(II1) complexes with algal biomass. Environ. Sci. Technol.. 20: 627-632.

Harris, P.O. and Ramelow, G.J., 1990. Binding of metal ions by particulate biomass derived from Chloreila vulgaris and Sce- nedesmus quadricauda. Environ. Sci. Technol., 24: 220-228.

Harrison, G.I. and Morel, F.M.M., 1983. Antagonism between cadmium and iron in the marine diatom Thalassiosira weis- fiogii. J. Phycol., 19: 495-507.

Hering, J.G., Sunda, W.G., Fergunson, R.L. and Morel, F.M.M., 1987. A field comparison of two methods for the determination of copper complexation: bacterial bioassay and fixed-potential amperometry. Mar. Chem., 20: 299-3 12.

234 M. Gonzrilez-DcivilajMarine Chemistry 48 (1995) 215-236

Hirose, K., 1988. Metal-organic ligand interaction in seawater: multimetal complexation model for metal titration. Mar. Chem., 25: 39-48.

Hirose, K. and Sugimura, Y., 1983. Relationship between metallic elements and organic matter in marine environments. Pap. Met. Geophys., 34: 291-306.

Hirose, K. and Sugimura, Y., 1985. Role of metal-organic complexes in the marine environment. Mar. Chem., 16: 239-247.

Huang, C., Huang, C.P. and Morehart, A.L., 1990. The removal of Cu(II) from dilute aqueous solutions by S. serevisiae. Water Res., 24: 433-439.

Huang, C., Huang, C.P. and Morehart, A.L., 1991. Proton competition in Cu(I1) adsorption by fungal mycelia. Water Res., 25: 1365-1375.

Imber, B.E., Robinson, M.G., Ortega, A.M. and Burton, J.D., 1985. Complexation of zinc by exudates from Skeletonema costatum grown in culture. Mar. Chem., 16: 131-139.

Jackson, G.A. and Morgan, J., 1978. The metal-chelator interactions and phytoplankton growth in sea water media: theoretical analysis and comparison with reported observations. Limnol. Oceanogr., 23: 2688282.

Jardim, W.F. and Pearson, H.W., 1984. A study of the copper- complexing compounds released by some species of cyanocabteria. Water Res., 18: 985-989.

Jensen, A., 1984. Excretion of organic carbon as a function of nutrient stress. In: 0. Holm-Hansen, L. Bolis and R. Gilles (Editors), Marine Phytoplankton and Productivity. Springer, Berlin, pp. 61-72.

Jensen, T.E., Baxter, M.M., Rachlin, J.W. and Jani, W., 1982. Uptake of heavy metals by Plectonema boryanum (cyanophyceae) into cellular components, especially poly- phosphate bodies: and X-ray energy dispersive study. Environ. Pollut., 27: 119-127.

Jones, R.P. and Gadd, G.M., 1990. Ionic nutrition of yeast- physiological mechanisms involved and implications for biotechnology. Enzyme Microbial Technol., 12: l-17.

Kahl, J.L., Faulkner, L.R., Dwarakanath, K. and Tachikawa, H., 1986. Reversible oxidation and rereduction of magnesium phthalocyamine electrodes. Electrochemical behavior and in situ Raman spectroscopy. J. Am. Chem. Sot., 108: 5434-5440.

Kaufman, F.B., Schroeder, A.H., Engler, E.M., Kramer, S.R. and Chambers, J.Q., 1980. Ion and electron transport in stable, electroactive tetrathiofulv’alene polymer coated electrodes. J. Am. Chem. Sot., 102: 483-488.

Khummongkol, D., Canterford, G.S. and Fryer, C., 1982. Accumulation of heavy metals in unicellular algae. Bio- technol. Bioeng., 4: 2643-2660.

Kinnburgh, D.G., Barker, J.A. and Whitfield, M., 1983. A comparison of some simple adsorption isotherms for describing divalent cation adsorption by ferrihydrite. J. Colloid Interface Sci., 95: 370-384.

Kotyk, A., 1989. Kinetic studies of transport in yeast. Methods Enzymol., 174: 5677591.

Kozarac, Z., Plavsic, M. and Cosovic, B., 1989. Interaction of cadmium and copper with surface-active organic matter and

complexing ligands release by marine phytoplankton. Mar. Chem., 26: 313-330.

Kramer, C.J.M. and Duinker, J.C., 1984. Complexation capacity and conditional stability constants for copper on sea and estuarine waters, sediment extracts and colloids. In: C.J.M. Kramer and J.C. Duinker (Editors), Com- plexation of Trace Metals in Natural Waters. Nijhoff, The Hague, pp. 217-228.

Krauskopf, K.B., 1956. Factors controlling the concentrations of thirteen rare metals in sea water. Geochim. Cosmochim. Acta, 9: l-33.

Laties, G.G., 1969. Dual mechanism of salt uptake in relation to compartmentation and long-distance transport. Annu. Rev. Plant Physiol., 20: 899116.

Maeda, S. and Sakaguchi, T., 1990. Accumulation and detoxification of toxic metal elements by algae. In: I. Akatsuka (Editor), Introduction to Applied Phycology. SPB, pp. 109-136.

Mantoura, R.F.C., 1981. Organo-metallic interactions in natural waters. In: E.K. Duursma and R. Dawson (Editors), Marine Organic Chemistry. Elsevier, Amster- dam, pp. 179-223.

Mantoura, R.F.C., Dickson, A. and Riley, J.P., 1978. The complexation of metals with humic material in natural waters. Estuarine Coastal Mar. Sci., 6: 387-408.

Martin, J.H. and Fitzwater, S.E., 1988. Iron deficiency limits phytoplankton growth in the north-east Pacific subarctic. Nature, 331: 341-343.

Martin, J.H. and Gordon, R.M., 1988. Northeast Pacific iron distributions in relation to phytoplankton productivity. Deep-Sea Res., 35: 177-196.

Martin, J.H., Gordon, R.M., Fitzwater, S. and Broenkow, W.W., 1989. VERTEX: Phytoplankton/iron studies in the Gulf of Alaska. Deep-Sea Res., 36: 6499680.

Martin, J.H., Gordon, R.M. and Fitzwater, S., 1990. Iron in Antarctic waters. Nature, 345: 156- 158.

Martin, J.H., Gordon, R.M. and Fitzwater, S., 1991. The case for iron. Limnol. Oceanogr., 36: 1793-1802.

Mierle, G., 1983. Studies on the uptake and toxicity of copper and other transition metals ions with a green alga, Scenedesmus acuminatus. Diss. Abstr. Int., 44103: 663B.

Millero, F.J. and Hawke, D.J., 1992. Ionic interaction of divalent metals in natural waters. Mar Chem., 40: 19-48.

Moffett, J.W., Zika, R.G. and Brand, L.E., 1990. Distribution and potential sources and sinks of copper chelators in the Sargasso Sea. Deep-Sea Res., 37: 27-36.

Mohl, W., Motsch, H. and Schweiger, A., 1988. Magnetic studies of Cu(I1) adsorbed on the surface of the bacterium Klebsiella pneumoniae. Langmuir, 4: 580-583.

Morel, F.M.M. and Hudson, R.J.M., 1985. The geobiological cycle of trace elements in aquatic systems: Redfield revisited. In: W. Stumm (Editor), Chemical Processes in Lakes. Wiley, New York, NY, pp. 251-281.

Morel, F.M.M., Hudson, R.J.M. and Price, N.M., 1991. Limitation of productivity by trace metals in the sea. Limnol. Oceanogr., 36: 1742-1755.

M. Gonzdez-D&da/Marine Chemistry 48 (1995) 215-236 235

Muller, F.L.L. and Kester, D.R., 1991. Voltammetric determination of the complexation parameters of zinc in marine and estuarine waters. Mar. Chem., 33: 71-90.

Murphy, L.S., Guillard, R.R.L. and Brown, J.F., 1984. The effects of iron and manganese on copper sensitivity in diatoms: Differences in the responses of closely related neritic oceanic species. Biol. Oceanogr., 3: 187-201.

Neilands, J.B., 1967. Hydroxamic acids in nature. Science, 156: 1443-1447.

Neilands, J.B., 1973. Microbial iron transport compounds (siderochromes). In: G. Eichhorn (Editor), Inorganic Bio- chemistry. Elsevier, Amsterdam, pp. 167-202.

Parks, G.A., 1982. Problems and opportunities in adsorption research. Chem. Aust., 49: 389-395.

Plavsic, M., Krznaric, D. and Branica, M., 1982. Determination of the apparent copper complexing capacity of seawater by anodic stripping voltammetry. Mar. Chem., 11: 17-3 1.

Robinson, N.J., 1988. Algal metallothioneins: secondary metabolites and proteins. J. Appl. Phycol., 1: 5-18.

Santana Casiano. J.M., Gonzblez-DBvila, M., Perez-Pefia, J. and Millero, F.J., 1994. Pb2+ interactions with the marine phytoplankton Dunaliella tertiolecta. Mar. Chem., 48: 115-l 29.

Schindler, P.W. and Stumm, W., 1987. The surface chemistry of oxides, hydroxides, and oxide minerals. In: W. Stumm (Editor), Aquatic Surface Chemistry. Wiley, New York, NY, pp. 83-107.

Schreiber, D.R., Millero, F.J. and Gordon, AS., 1990. Production of an extracellular copper-binding compound by the heterotrophic marine bacterium Vibrio alginolyticus.

Mar. Chem., 28: 275-284. Seritti, A., Pellegrini, D., Morelli, E., Barghiani, C. and Ferrara,

R., 1986. Copper complexing capacity of phytoplanktonic cell exudates. Mar. Chem., 18: 351-357.

Sigg, L., 1987. Surface chemical aspects of the distribution and fate of metal ions in lakes. In: W. Stumm (Editor), Aquatic Surface Chemistry. Wiley, New York, NY, pp. 331-350.

Simpson, F.B. and Neilands, J.B., 1976. Siderochromes in cyanophyceae: isolation and characterization of schizokinen from Anabaena sp. J. Phycol., 12: 44-48.

Snow, G.A., 1970. Mycobactins: iron chelating growth factors from mycobacteria. Bacterial. Rev., 34: 99-125.

Sposito, G., 1983. On the surface complexation model of the oxide-aqueous solution interface. J. Colloid Interface Sci., 91: 329-340.

Sposito, G., 1984. The Surface Chemistry of Soils. Oxford Univ. Press, New York, NY.

Steinhardt, J. and Reynolds, J.A., 1969. Multiple Equilibria in Proteins. Academic, New York, NY, 214 pp.

Stroes-Gascoyne, S., Kramer, J.R. and Snodgrass, W.J., 1987. A new model describing the adsorption of copper on MnO,. Environ. Sci. Technol., 20: 1047-1050.

Stumm, W. and Morgan, J.J., 1981. Aquatic Chemistry. Wiley, New York, NY, 2nd ed.

Sueur, S., van den Berg, C.M.G. and Riley, J.P., 1982. Measure- ment of the metal complexing ability of exudates of marine macroalgae. Limnol. Occanogr., 27(3): 536-543.

Sunda. W.G., 1986. Neritic-oceanic trends in trace metal

toxicity to phytoplankton communities. In: Proc. 4th lnt. Ocean Disposal Symp., pp. 19-29.

Sunda, W.G., 1988/1989. Trace metals interactions with marine phytoplankton. Biol. Oceanogr., 6: 41 l-441

Sunda, W.G. and Ferguson, R.L., 1983. Sensitivity of natural bacterial communities to additions of copper and to cupric ion activity: a bioassay of copper complexation in sea water. In: C.W. Wong, E. Boyle, K. Bruland, J. Burton and E.D. Goldberg (Editors), Trace Metals in Sea Water. Plenum. New York, NY, pp. 871-891.

Sunda, W.G. and Gessner, 1989. The production of extracellular copper-complexing ligands by marine and estuarine fungi. Chem. Speciat. Bioavail., 1: 65-70.

Sunda, W.G. and Guillard, R.R.L., 1976. The relationship between cupric ion activity and the toxicity of copper to phytoplankton. J. Mar. Res., 34: 51 l-529.

Sunda, W.G. and Hanson, A.K., 1987. Measurement of free cupric ion concentration in seawater by a ligand competition technique involving copper sorption onto C,s SEP-PAK cartridges. Limnol. Oceanogr., 32: 537-551.

Sunda, W.G., Barber, R.T. and Huntsman, S.A., 198 1. Phyto- plankton growth in nutrient rich seawater; importance of copper-manganese cellular interactions. J. Mar. Res., 39: 567-586.

Sunda, W.G., Klaveness, D. and Palumbo, A.V., 1984. Bio- assays of cupric ion activity and copper complexation. In: C.J.M. Kramer and J.C. Duinker (Editors), Complexation of Trace Metals in Natural Waters. Nijhoff, The Haque, pp. 339-409.

Swallow, K.C., Westal, J.C., McKnight, D.M., Morel, M.M.L. and Morel, F.M.M., 1978. Potentiometric determination of copper complexation by phytoplankton exudates. Limnol. Oceanogr., 23: 538-542.

Takahashi, K. and Tannaka, K., 1986. Adsorption of copper ions and alkyldithiocarbonate at silica by a molecular orbital method. J. Colloid Interface Sci., 113: 21-31.

Ting, Y.P., Lawson, F. and Prince, I.G., 1989. Uptake of cadmium and zinc by the alga Choreila vulgaris: Part 1. Individual ion species. Biotechnol. Bioeng., 34: 990-999.

Ting, Y.P., Lawson, F. and Prince, LG., 1991. Uptake of cadmium and zinc by the alga Chlorella vulgaris: II. Multi- ion situation. Biotechnol. Bioeng., 37: 445-455.

Van Cutsem, P. and Gillet, C., 1981. A thermodynamic study of Cu++-Zn++ ion exchange in the Nitella jexilis cell wall. Plant Soil, 62: 367-375.

Van Cutsem, P., Mestdagh, M.M., Rouxht, P.G. and Gillet, C., 1984. Preliminary ESR study and relation with ion exchange thermodynamics of copper adsorbed on a biological ion exchanger-the NiteNaJIexilis cell wall. React. Polym. Ion Exch. Sorbents, 2: 3 l-35.

Van den Berg, C.M.G., 1979. Complexation of copper in natural waters. Determination of complexing capacities and conditional stability constants of the copper complexes in natural waters by MnOz and the applications for phytoplankton activity. Ph.D. thesis. McMaster Univ., Hamilton, Ont.

Van den Berg, C.M.G., 1984. Determination of the complexing capacity and conditional stability constants of complexes of

236 M. Gonzcilez-D&la/Marine Chemistry 48 (1995) 215-236

copper(B) with natural organic ligands in seawater by cathodic stripping voltammetry of copper-catechol complex ions. Mar. Chem., 15: 12681274.

Waite, T.D. and Morel, F.M.M., 1983. Characterization of complexing agents in natural waters by copper(II)/copper(I) amperometry. Anal. Chem., 55: 12681274.

Wangersky, P.J., 1978. Production of dissolved organic matter. In: 0. Kinne (Editor), Marine Ecology, IV, Dynamics. Wiley, New York, NY, pp. 1155220.

Wangersky, P.J., 1986. Biological control of trace metal residence time and speciation: a review and synthesis. Mar. Chem., 18: 269-297.

Watkins, J.W., Elder, R.C., Greene, B. and Darnall, D.W., 1987. Determination of gold binding in an algal biomass using EXAFS and XANES spectroscopies. Inorg. Chem., 26: 1147-1151

Whitfield, M. and Turner, D.R., 1987. The role of particles in regulating the composition of seawater. In: W. Stumm (Edi- tor), Aquatic Surface Chemistry. Wiley, New York, NY, pp. 457-493.

Wilkinson, L., 1988. SYSTAT: The system for statistic, SYSTAT Inc., Evanston, IL.

Wood, J.M. and Wang, H.K., 1983. Microbial resistance to heavy metals. Environ. Sci. Technol., 17: 582A-590A.

Wood, J.M. and Wang, H.K., 1985. Strategies for microbial resistance to heavy metals. In W. Stumm (Editor), Chemical Processes in Lakes. Wiley, New York, NY, pp. 81-98.

Xue, H.B. and Sigg, L., 1990. Binding of Cu(I1) to algae in a metal buffer. Water Res., 24: 1129- 1136.

Xue, H.B., Stumm, W. and Sigg, L., 1988. The binding of heavy metals to algal surfaces. Water Res., 22: 917-926.

Zhou, X. and Wangersky, P. J., 1985. Copper complexing capacity in cultures of Phaedactylum tricornutum: diurnal changes. Mar. Chem., 26: 21-40.

Zhou, X. and Wangersky, P. J., 1989a. Study of copper- complexing organic ligands: isolation by a Sep-Pak Cl8 column extraction technique and characterization by Chromarod thin-layer chromatography. Mar. Chem., 26: 21-41.

Zhou, X. and Wangersky, P.J., 1989b. Production of copper- complexing organic ligands by the marine diatom Phaeo- dactylum tricornutum in a cage culture turbidostat. Mar. Chem., 26: 239-259.

Zutic, V., Cosovic, B., Marcenko, E., Bihari, N. and Krsinic, F., 198 1. Surfactant production by marine phytoplankton. Mar. Chem., 10: 505-520.

the role of phytoplankton cells on the control of heavy metal concentration in seawater

Documents