Bioavailability of trace metals to aquatic microorganisms: importance of chemical, biological and physical processes on biouptake

Biochimie 88 (2006) 1721–1731 www.elsevier.com/locate/biochi

Bioavailability of trace metals to aquatic microorganisms: importance of chemical, biological and physical processes on biouptake I. Wormsa, D.F. Simona,b, C.S. Hasslera,1, K.J. Wilkinsonb,* a

CABE (Analytical and Biophysical Environmental Chemistry), University of Geneva, 30, quai Ernest Ansermet, 1211 Geneva 4, Switzerland b Department of Chemistry, University of Montreal, CP 6128, succursale Centre-ville, Montreal, Canada H3C 3J7 Received 9 February 2006; accepted 7 September 2006 Available online 28 September 2006

Abstract An important challenge in environmental biogeochemistry is the determination of the bioavailability of toxic and essential trace compounds in natural media. For trace metals, it is now clear that chemical speciation must be taken into account when predicting bioavailability. Over the past 20 years, equilibrium models (free ion activity model (FIAM), biotic ligand model (BLM)) have been increasingly developed to describe metal bioavailability in environmental systems, despite the fact that environmental systems are always dynamic and rarely at equilibrium. In these simple (relatively successful) models, any reduction in the available, reactive species of the metal due to competition, complexation or other reactions will reduce metal bioaccumulation and thus biological effects. Recently, it has become clear that biological, physical and chemical reactions occurring in the immediate proximity of the biological surface also play an important role in controlling trace metal bioavailability through shifts in the limiting biouptake fluxes. Indeed, for microorganisms, examples of biological (transport across membrane), chemical (dissociation kinetics of metal complexes) and physical (diffusion) limitation can be demonstrated. Furthermore, the organism can employ a number of biological internalization strategies to get around limitations that are imposed on it by the physicochemistry of the medium. The use of a single transport site by several metals or the use of several transport sites by a single metal further complicates the prediction of uptake or effects using the simple chemical models. Finally, once inside the microorganism the cell is able to employ a large number of strategies including complexation, compartmentalization, efflux or the production of extracellular ligands to minimize or optimize the reactivity of the metal. The prediction of trace metal bioavailability will thus require multidisciplinary advances in our understanding of the reactions occurring at and near the biological interface. By taking into account medium constraints and biological adaptability, future bioavailability modeling will certainly become more robust. © 2006 Elsevier Masson SAS. All rights reserved. Keywords: Bioaccumulation; Uptake; Internalization; Microbiology; Trace metal; Diffusion; Complex; Regulation

1. Introduction In natural waters, the bioavailability of trace metals, including their toxicity, is thought to be related to their ability to cross biological barriers (e.g. plasma membrane) and it is

* Corresponding

author. E-mail address: [email protected] (K.J. Wilkinson). 1 Present address: CSIRO Atmospheric and Marine Research, Castray Esplanade, Hobart 7000, TAS, Australia 0300-9084/$ - see front matter © 2006 Elsevier Masson SAS. All rights reserved. doi:10.1016/j.biochi.2006.09.008

most often predicted by the concentration [1] or flux [2] of internalized metal. The biouptake process depends not only on the internalization pathways and their specificity but also on the physicochemistry of the medium and the size and nature of the organism [3]. Equilibrium models that have been developed to predict the role of chemical speciation on metal bioavailability are often (qualitatively) successful in predicting a reduction in trace metal effects due to complexation (inorganic and organic ligands, [4]) or competition (e.g. H+, Ca2+, etc., [5]); both processes result in a decreased interaction of the metal with uptake sites on the surface of the organism. None-

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theless, a fundamental understanding of the biouptake process is currently lacking, especially for the conditions that are the most relevant to the natural environment (i.e. presence of multiple stressors, ligand heterogeneity and polydispersity, nonequilibrium conditions, etc.). Such a quantitative understanding requires insights into: ● the behavior of metal species during their transport from the bulk solution (i.e. >few microns from the biological surface) to the biological interface; ● the transfer of the chemical across the biological membrane and; ● the role of the organism in modifying the chemistry and biology of the uptake process (Fig. 1; [2,6–10]). This mini-review examines the influence of these processes on biouptake and is concluded with two examples where the importance of chemical, biological and physical processes is demonstrated. 2. Role of the physicochemistry of the bulk solution A consensus exists in the literature with respect to the key fluxes that define the interaction of trace metals with aquatic organisms (Fig. 1). Trace metals, and their complexes, must first diffuse from the external medium to the surface of the organism (mass transport). Metal complexes are often dynamic, able to dissociate and reassociate (complexation/dis-

sociation) in the time that it takes to diffuse to the biological surface. To have an effect, the metal must react with a sensitive site on the biological membrane (adsorption/desorption), often but not necessarily followed by biological transport (internalization). Any of these metal fluxes (e.g. diffusive: Jdiff; chemical dissociation: Jkin; internalization: Jint, etc.) may be ratelimiting for the overall biouptake process. Their relative magnitudes will vary according to the chemical nature of the compounds being accumulated, the size and type of the organism and the physicochemical conditions in the immediate vicinity of the organism (e.g. pH, metal or ligand concentrations, membrane potentials) [2,3,11]. Furthermore, certain biological responses to the metals (e.g. efflux, pH changes at the biological interface, modifications to the surface charge) will influence the physicochemistry of the metals in solution and at the interface and thus modify the internalization flux or the concentration of accumulated metal. Since for several trace metals in natural waters (e.g. Cu, Al, Cr, Hg), complexes can typically account for > 90% of the total metal concentration [6], the identification of the rate-limiting flux is not simply academic: it will determine whether metal complexes can contribute fully, partially or not at all to biological internalization. This implies that a large number of parameters will determine the relative magnitude of the fluxes with important implications on the overall biouptake process and the subsequent biological effects. In many cases, trace metal transport across the biological membrane is rate-limiting (Jint << Jdiff, Jkin) [12] and the overall process can be simplified to a thermodynamic equilibrium

Fig. 1. Conceptual model of some of the important physicochemical processes leading to and following the uptake of a trace metal by an aquatic microorganism. L corresponds to a ligand that can complex the metal, M (subscripts L: lipophilic; h: hydrophilic; bio: biological; int: internalized).

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among metal species in the bulk solution and those bound to sensitive sites on the biological surface (with a number of limiting assumptions, see [12,13]). Although the chemical equilibrium assumption is perhaps unrealistic in dynamic natural media, equilibrium models such as the free ion activity model (FIAM, [6,13,14]) and the biotic ligand model (BLM, [15–17]) have been overall very successful in predicting bioavailability [13,15]. Under these pseudo-equilibrium conditions, metal concentrations and speciation are the same in the bulk solution and at the membrane–water interface. Furthermore, the dynamics of the metal complexes (their mobility due to size, chemical lability, etc.) is a non-issue since equilibrium is maintained throughout the solution. Biouptake and toxicological effects are then best predicted by the concentration of any of the equilibrium species in the bulk media, most often the free ion (FIAM) or surface bound metal (BLM). In such a case, constants describing the uptake process may be determined by analyzing the data as a steady-state process with saturable uptake kinetics (Eq. 1). ka

k int

M þ Rs Ð M
Rs
! M int kd

with K M
Rs ¼

ka kd

(1)

where the first step describes the rapid adsorption of a metal (charges omitted for simplicity) to a biological transport site, Rs, KM–Rs is the stability constant for the adsorption and ka and kd are the association and dissociation rate constants. The ratelimiting step corresponds to the reorganization of the transport site and can be characterized by the internalization rate constant, kint. Under this condition of pseudo-equilibrium, it is possible to relate the internalization flux (Jint) to the concentration of metal in solution (Michaelis–Menten equation, Eq. 2): J int ¼

J max ½M  K m þ ½M

with K m ¼

k d þ k int ka

(2)

where Jmax, the maximum internalization flux, is attained when the transport sites are saturated by the metal of interest. Furthermore, since kint << kd, the affinity constant for the adsorption of metal to the transport site, KM–Rs, can be determined from the reciprocal of the Michaelis–Menten constant, Km. This constant2 indicates the selectivity of the transporter for the metal (see following section). If, on the other hand, mass transport is limiting (Jdiff << Jint), it will set an upper limit to the uptake flux that can be accessed by the organism. A rate-limiting diffusion will be most relevant for the case of a relatively rapid internalization flux or a relatively slow diffusive flux. For example, a diffusive limitation is likely to be observed for colloidal or particulate metal species; for metal uptake by an organism in a biofilm, sediments, etc. or for rapidly accumulating metals such as Ag (I) [18] or Zn(II) [19]. In such cases, metal complexes will 2

Note that most affinity constants in the literature have been determined in complex (culture) media without taking trace metal speciation into account. Therefore, they are often underestimated due to an overestimation of the actual metal species (presumably free aquo ion) that are actually bound to the transport sites.

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contribute to biouptake according to their chemical dynamics (reflected by their chemical lability and physical mobility [3, 11]). This also implies that the movement of the microorganisms or of the surrounding water will enhance advective solute supply [25–27] by decreasing the thickness of the diffusion boundary layer [28]. Due to an imperfect capacity to absorb nutrients brought to the surface of the organism by mass transfer, many intermediate cases also exist for which the organism can only partially exploit mass transport increases [29,30]. For various reasons, few studies have rigorously examined the uptake of trace metals in the context of models other than the simple thermodynamic ones. Due to this lack of experimental verification, it is currently unclear to what extent external factors such as complex lability, mass transport and surface properties (e.g. charges, presence of slime layer, etc.) may influence solute fluxes. As will be shown in the following sections, internal factors such as the nature of the internalization strategy or the presence of biological regulation can also greatly influence the magnitude of the biouptake fluxes. 3. Metal internalization Internalization is the key step in the overall biouptake process. As opposed to the processes in the bulk solution, the plasma membrane is biologically active and often able to control the magnitude of the internalization fluxes according to the requirements of the organism. Due to the overall hydrophobic nature of the biological membrane, only neutral or non-polar molecules cross into the cytosol by passive diffusion (based upon the concentration gradient between external and internal compartments and a partition coefficient between the aqueous and lipid phases). For metals, this mode of transport has been shown to be relevant for dithiocarbamate metal complexes (e.g. pesticides), organomercury and neutral silver chloride complexes [20–24]. The vast majority of environmentally relevant metal species are hydrophilic and their transport through the biological membrane is mediated by specific proteins. In a few cases, trace metal internalization has been shown to occur via anionic transporters [15]. In that case, metals are piggy-backed across transporters meant for low molecular weight ligands such as citrate [25–27], thiosulfate [28] or phosphate [29]. This means of transport will depend greatly on metal speciation but also upon the cell nutrition [30]. Other highly specific ligands such as siderophores may be produced by the cell specifically to facilitate the transport of essential cations that are present at low concentrations in the environment (e.g. Fe, Co) [31,32]. The majority of trace metals are thought to be transported as the free aquo ion. While the transport sites often demonstrate a high affinity for required metals (low Km or high KM–Rs, see above), they do not always have a high selectivity. Indeed, one common mechanism of toxicity is for a toxic metal to bind to the site of an essential metal with a similar ionic radius or coordination geometry [33,34]. For example, for both bacteria and green algae, Ca2+ has been shown to reduce the internalization of Pb2+, Cd2+ and Zn2+ [35–39]. Similarly, the toxicity

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of Tl is attenuated in the presence of K+ because it is transported via a Na+/K+-ATPase [40,41]. In prokaryotes, the CorA Mg transporter has been shown to be involved in Ni, Co and Mn toxicity (and internalization) while not being affected by Zn and Fe [42,43]. In yeast, the Alr1 protein, showing a large similarity with CorA, appears to be involved with both Mg internalization and homeostasis [44] and Al toxicity [44,45]. At a molecular level, three major classes of transition metal transporters have been identified: P-type ATPases [46,47]; zinc regulated transporter/iron-regulated transporter (ZRT/IRT) and the natural resistance associated macrophage proteins (Nramp) proton symporter. P-type ATPases appear to be involved with both the internalization and detoxification of trace metals [48– 51]. While the ZRT/IRT1 proteins (ZIP) [52] were first shown to be involved with Zn and Fe uptake (and homeostasis) [53, 54], they also allow the entry of Mn and Cd. The Nramp family of proteins have been shown to transport Fe(II), Zn and Mn and to promote the uptake of Co, Cd, Cu, Ni and Pb [55]. Despite such enormous differences in molecular structure and function, most trace metal transport, except that involving the ATPases, can be described by simple Michaelis–Menten uptake kinetics and equilibrium considerations. The toxicity and uptake of trace metals depends upon the concentration of major cations or competing trace metals in the water of interest [56]. Under equilibrium conditions, a reduction in binding to the transport sites (and subsequent reduction in Jint) is predicted in the presence of a competing ion. In that case, the specificity of the trace metal transporters can be evaluated from measured internalization fluxes and the inhibition constants, Ki (1/KC–Rs), corresponding to the concentration of competiting cation required to decrease the metal uptake flux by 50% (Eq. 3): J int ¼

J max ½MK M
Rs K M
Rs ½M  þ K C1
Rs ½C 1  þ … þ K Cn
Rs ½C n  þ 1

(3)

Protons may also compete for the uptake of trace metals: generally, a decrease in pH should lead to a decrease in uptake due to the (equilibrium) protonation of the metal binding sites [5]. In addition, for a cation/H symport such as Nramp, noncompetitive interactions (i.e. protons directly involved in the transport mechanism) can be observed via Jmax variations. For example, for both Escherichia coli and S. thyphimurium, an increase in pH from 5.5 to 8.2 resulted in a threefold decrease in the maximum Mn uptake flux, without any change in the Km [57]. For the uptake of a divalent metal by multiple transport pathways [58,59], a simple hyperbolic representation of the uptake fluxes as a function of the metal concentration, as observed for a single transport system, is no longer observed [60–62]. For example, for essential trace metals, internalization may be linked to both a low and a high affinity transporter (e.g. [63]). If the maximum internalization fluxes are sufficiently different (most often by several orders of magnitude, i.e. Jmax1 >> Jmax2), then biouptake can be distinguished by two distinct, saturable processes, as has been reported for Mn

Fig. 2. Implication of multitransport pathways and regulation of transporter expression on the biouptake process. a. Internalization is via two distinct transporters, one of low affinity (Km1 = 10–5.3 M, Jmax1 = 6.6 × 10–13 mol cm –2 s –1 ) and the other of high affinity (K m2 = 10 –8.5 M, Jmax2 = 5.1 × 10–15 cm–2 s–1). In this case, the overall biouptake fluxes (dashed line) are not simply related to the [MZ+] in solution. Such an effect is not apparent on a linear scale (Inset). b. Up-regulation of a high affinity transport system (Km2 = 10–7.5, Jmax2 = 5.1 × 10–16 cm–2 s–1). The increased expression of a trace metal transporter can progressively lead to a non linearity between the [M2+] in solution and Jint as observed by the overall uptake fluxes, presented as the two dashed lines in b. In certain cases, Jint increased to the point where it attained the maximum flux that could be allowed due to diffusive constraints (dotted line: Jdiff = f([MZ+])). In this case, internalization would no longer be rate-limiting, such that the dissociation of the trace metal complexes becomes a potential source of trace metal uptake (i.e. metal complexes become bioavailable under these conditions). While the figures presented here are meant to be generic, similar uptake curves have been observed for a number of trace metals including Ag+, Zn2+, Pb2+, Al3+, Co2+, Ni2+, Cd2+.

biouptake by yeast [64] or Zn uptake by marine diatoms [63] (Fig. 2A). On the other hand, the uptake of (essential) trace metals is often confounded due to their regulation. In some cases, microorganisms adapt to low metal concentrations by: ● increasing the production of transporters (e.g. Mn biouptake by Thalassiosira pseudonana [65]) or ; ● increasing the affinity of the transporters (e.g. Zn biouptake by T. pseudonana [63]). 4. Impact of cellular regulation on metal bioavailability Organisms have a number of transport systems that are sensitive to their external surroundings [66]. For example, in yeast, the Zrt1 (ZIP family) transporter may be induced at

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low Zn concentrations and rapidly degraded via the ubiquitin pathway if the Zn concentration increases. Transporter degradation is likely stimulated, but with less efficiency, in the presence of Cd or Co [67]. Such feedback has also been demonstrated in marine phytoplankton for Mn, Cd and Zn [68]. Regulation of the transport pathways will directly affect the metal internalization fluxes (Fig. 2B). If Jint is changed significantly, the overall rate-limiting step of the biouptake process may be modified with important consequences on metal bioavailability, including the biological dissociation of the trace metal complexes. For example, under conditions of Zn limitation, Zn uptake has been shown to be increased to the point where diffusion becomes limiting and Zn–NTA complexes become bioavailable (see example below) [19]. The transition from biological (thermodynamic) to chemical (dynamic) limitations on the uptake process is more likely to be observed for essential than for toxic metals. In addition to modifying cellular uptake, metal ions can induce a variety of cellular responses that may directly or indirectly influence bioavailability. The bioavailability of a metal ion inside the cell is determined by its capacity to interact with proteins producing toxicity. For instance, toxic effects may be observed due to: the inactivation of proteins (e.g. enzymes) as a result of their interaction with metals [69–71]; a competitive interference on the uptake of one metal by another [72] or by the induction of an oxidative stress with a subsequent cell damage [51,73]. Microorganisms have thus developed several protective mechanisms including: ● intracellular binding or sequestration by metal-complexing agents [72]; ● compartmentalization and transport of metals to subcellular compartments (e.g. vacuoles) [51,74–76]; ● efflux [77–79] and ● extracellular sequestration [80–82] (Fig. 1). The intracellular sequestration of trace metals is an effective mechanism used by microorganisms (mainly eukaryotes) to render toxic metal ions less bioavailable. Several classes of intracellular metal chelators such as glutathione, amino acids, phytochelatins [83], metallothioneins, organic acids [84–86] and thioredoxins (TRX) [87] are involved in intracellular complexation and can participate in the cell protection mechanism. For example, the tripeptide glutathione (GSH) is generally maintained in very high ratios (500:1) of reduced to oxidized form [88] and might be the first barrier against toxic damage; metal complexation may lead to increases in the level of oxidized GSH (i.e. GSSG), resulting in the activation of further detoxification mechanisms [89,90]. Enhanced proline levels can protect Chlamydomonas reinhardtii by maintaining a more reduced environment (and thus higher glutathione levels) [69] and may sequester metal ions prior to the synthesis of metallothioneins or phytochelatins [91–93]. Highly complexing phytochelatins (PC) have been detected in algae [94], yeast [95] and bacteria [96,97]. Their induction depends on the concentration and nature of the metal (mainly Cd, Cu,

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Ag, Ni) [98] and the organism [77,99,100]. Metallothioneins (MTs) are ubiquitous, low molecular weight cysteine rich proteins that bind metal ions in thiolate clusters [72]. They have several physiological functions such as storage, transport of essential metal ions (Zn, Cu) and detoxification of non essential metal ions (Cd, Hg). Amino acids (e.g. proline, cysteine, histidine) can complex toxic trace metals in the cytosol [84, 101,102] and organic acids such as citrate, malate and oxalate have also been implicated in metal tolerance, metal transport and vacuolar metal sequestration. TRX are small ubiquitous proteins with highly reactive exposed disulfide sites that have been implicated as an electron donor in a wide range of biochemical pathways [88]. In many cells, reduced TRX regulates the activities of enzymes by reducing disulfide to thiol [87,88]. The effects of excess metal ions in the cytosol may also be reduced by efflux or by compartmentalization [103,104]. Toxic metals that are complexed by phytochelatins, polyphosphate bodies and organic acids [85,100] may be stored in internal compartments including vacuoles [105] and chloroplasts [74]. The transport of chelated metal ion from the cytosol into the vacuole is mediated by transporter proteins. For example, in fission yeast, HMT1, a protein with a similarity to the ABCtype transporter, is localized in the vacuolar membrane where it mediates the Mg-ATP energized transport of Cd–PC complexes and Apo-PCs [76]. Polyphosphates, found in the cells of all organisms, are also effective chelators of metal ions [106, 107]. Indeed, non-membranous electron dense deposits of Cd and phosphates have been observed in the vacuoles of Chlamydomonas acidophila following exposure to Cd, Zn and Cu and cells recovering from a Cd stress have been shown to increase cytoplasmic and sugar phosphates [75]. Microorganisms (mainly prokaryotes) have effective efflux mechanisms for metals [76,108,109]. For example, Cupriviadus metallidurans is able to activate the synthesis of a nickel efflux pump, CrnCBA, once nickel has entered into the periplasm [110]. Chlorella vulgaris excludes copper [111] while diatoms, coccolithophores and cyanobacteria have been shown to exclude metal-complexes [79,80]. In addition to decreasing toxicity, efflux will reduce the net metal flux and may modify the chemical speciation of the metal through the expulsion of inert trace metal complexes. Microorganisms can also excrete compounds that will complex metal ions in the extracellular medium in order to reduce their bioavailability [112–115]. For example, Cd bioaccumulation by Rhodospirillum rubrum (gram negative bacteria) was greatly decreased by the production of an unidentified extracellular ligand [35]. Algae have been observed to produce extracellular complexing agents including polysaccharides, proteins, peptides and small organic acids that are able to decrease the concentration of bioavailable metal in the immediate vicinity of the cell [116]. For example, in marine systems, copper, cadmium and zinc may be strongly complexed by the thiol rich exudates and amino acids that are produced by the coccolithophorid, Emiliania huxleyi [117–119]. Finally, the screening of Cd responsive genes in A. thaliana has resulted in the identification of a gene that encodes for a putative metal binding protein that is exclusively located at the cell membrane and which

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upon introduction into yeast cells confers an increased tolerance to cadmium exposure [120]. The secretion of metal chelators reduces bioavailability, uptake and toxicity which in turn, results in a decreased production of metal binding ligand. In summary, even once a trace metal has succeeded in entering a cell, its effects may be reduced by complexation inside the cytosol; compartmentalization; efflux or by modification of the extracellular trace metal speciation. All these processes are expected to influence trace metal bioavailability, biouptake and toxicity in a complex feedback loop that will affect the biology, chemistry and physics (e.g. diffusion) at the biological–water interface. 5. Examples of interactions among chemistry, biology and physics: metals at the biological interface Zn and Fe are essential trace metals that are required for the development of microorganisms. Their complex chemical speciation, highly specialized biological uptake and intracellular homeostasis have been extensively studied to the point that they are good examples of the complexity of trace metal bioavailability (Fig. 3). Indeed, under some conditions, the bioavailability of either metal is likely limited by its diffusive flux or by ligand exchange kinetics [19,121–123] such that, in addition to the free aquo ion, labile metal complexes can contribute to bioavailability [3]. Due to the low solubility of its oxyhydroxides, Fe3+ is predicted to be present at concentrations (< 10–17 M at pH 7) that are not sufficient to allow for the optimal growth of microorganisms [124]. In fact, most bioavailable iron in the natural environment is strongly associated to poorly characterized organic ligands [125–127], some of which are related to specific Fe(III) binding ligands (siderophores) with large stability constants (1025–1050 M–1; [125,128,129]). Contrary to what

would be predicted by the steady state models presented above, several studies have indicated that iron complexes [123,130–132] and colloidal iron [133,134] are accessible to phytoplankton and that bioavailability is enhanced by both photochemical and biological mechanisms [135–137]. Iron internalization strategies are highly specific and species-dependent (Fig. 3; e.g. [138]). Under iron deficiency, some prokaryotic phytoplankton [139] and bacteria [124] excrete strongly chelating siderophores in order to increase Fe solubility and bioavailability (Fig. 3). One strategy involves recognition of the Fe(III) siderophore complex by a specific membrane receptor, transport inside the microorganism and reduction of the strong Fe(III) complex to a weak Fe(II) complex in order to allow its dissociation (e.g. [124,138]). In addition, both ABC and low affinity transport systems have been reported in various bacteria [124] and prokaryotes can also express additional receptors that recognize siderophores produced by other organisms (e.g. E. coli, [48]). Some eukaryotes are known to possess an Fe-transporter that is coupled to a surface reductase (e.g. [131]). In this case, iron chelates are dissociated prior to their internalization [131,140]. Even though most eukaryotes do not secrete siderophores, many have transport systems for Fe(III) siderophore complexes [130,131]. With such complex interactions, competition is likely to occur among microorganisms for the complexed iron in natural waters and an increased siderophore production may not directly reflect Fe bioavailability for a given microorganism. Iron storage and detoxification is mostly mediated by ferritin and ferritin-like compounds. Intracellular iron(II) is oxidized at the surface of the ferritin; each molecule can store up to 2000–4000 Fe(III) atoms [124,141]. Reactive oxygen species (ROS) are produced [142] during an aerobic metabolism process in which 23–24 atoms of Fe are required [143]. ROS react with iron via the Fenton reaction to produce highly reac-

Fig. 3. Schematic representation of major processes influencing iron bioavailability to microorganisms. In this simplified diagram, Fe(II) and Fe(III) ions are considered as the two chemical species that are bioavailable. Iron bound within organic complexes are potentially bioavailable, depending upon the concentrations, affinity constant, lability and existing biological uptake strategies. In this figure, the excretion, recognition and transport of the siderophores leads to a speciesspecific iron bioavailability. Reduction of organic Fe(III) complexes facilitates their dissociation in order to allow iron biouptake. The width of the arrows indicates the relative instability of the Fe(III) complexes. Precipitation of Fe(III) hydroxides and formation of colloidal iron are omitted for simplicity. Inside the microorganism, iron regulation is shown with dotted arrows. Regulation of ferritin and the coupled regulation of iron and ROS are not shown for clarity. A second microorganism able to compete for the bioavailable iron using similar siderophores and/or surface reductases is also illustrated.

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tive and toxic hydroxy radicals. Furthermore, iron may be a cofactor of ROS detoxification enzymes (e.g. [124,144]). For this reason, iron regulation is strongly coupled with oxidative stress mechanisms for bacteria, cyanobacteria and microalgae [124,143–145]. The ferric uptake regulator (Fur) protein is the major prokaryotic iron repressor requiring Fe(II) binding [146]. For deficient iron levels and in response to a ROS stress, Fur is inactivated and the genes involved in iron acquisition (transporters, siderophores, specific receptor) are no longer repressed [124,143,144,147]. The inactivation of Fur is also coupled to the repression of small RNAs in bacteria which stimulate the degradation of mRNA corresponding to Fe containing proteins. Such a process results in the strict regulation of iron requirements and iron biouptake. Although little is known about the genes involved in the intracellular homeostasis of iron in eukaryotes, a similar control to what has been observed in bacteria and cyanobacteria has been reported in yeast [147]. Finally, physiological changes to iron limitation such as a decrease in cell size (to enhance diffusion) [121,122]; a replacement of iron-requiring enzymes (e.g. ferredoxin substitution by flavodoxin, [148]) or a slower growth rate have been observed in some species of phytoplankton and may decrease iron requirements by a factor of 6 to 8 [122]. Nonetheless, the biological cost of such strategies is apparent through decreases in photosynthetic activity, nitrate fixation and chlorophyll or protein production rates [139,148,149]. Zn uptake is another good example of the interplay occurring among the physical, chemical and biological fluxes that are present in and around the biological interface. In natural waters, Zn is found as the free ion or in relatively weak complexes [150,151]. Zn uptake is often tightly regulated: Zn internalization fluxes and intracellular Zn contents can be maintained relatively constant despite large variations in bulk solution [Zn2+] (10–8.5–10–11 M) [19,63]. At low Zn concentrations, Zn internalization can be upregulated to the point where Zn diffusion becomes limiting and thermodynamically strong Zn complexes become bioavailable [19,152] to the organism. The role of Zn homeostasis has been demonstrated clearly in Saccharomyces cerevisiae where both high (ZRT1) and low (ZRT2) affinity transport systems are overexpressed in Zn depleted cells [52–54]. When the cells have sufficient Zn, the high affinity system (ZRT1) is repressed and Zn is eliminated to the vacuoles [53,54,59,153]. Such transcriptional and posttranscriptional regulation of the transport system has resulted in a Zn transport capacity that is highly dependent upon the Zn biological status [59,153] but quite independent on the Zn solution chemistry. Intracellular GSH levels might also play a role in regulating Zn export [46]. Zn is often [98,99], but not always [154], a weak inducer of PCs, probably due to its low binding affinity [155,156], especially when compared to Cd. In addition, other ligands such as citrate, malate, oxalate and unknown thiolates [58,157,158] have been shown to be involved in intracellular Zn chelation while polyphosphates have been shown to play an important role in Zn removal in Chlorophyta and Bacillariophyta [159]. Finally, zinc efflux is used to regulate intracellular metal contents for both bacteria (e.g. [58]) and microalgae [152,160].

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In summary, trace metal bioavailability depends on physical, biological and chemical factors that are highly complex and interdependent. While physicochemical factors may limit trace metal internalization, organisms can adapt resulting in modifications to the rate-limiting flux and thus to the chemical availability of metal complexes. Essential trace metals are often highly regulated in order to avoid both situations of trace metal deficiency and overload. Strategies are varied, depending on the organism, the metal and the physicochemistry of the metal in solution. None of the processes are entirely independent as each process can lead to modifications in the different fluxes that control trace metal bioavailability in organisms. A better understanding of trace metal bioavailability in the natural environment will depend upon further advances to understanding the biological, chemical and physical processes occurring at the biological interface. Acknowledgements The authors thank the Swiss National Science Foundation, the National Science and Engineering Research Council of Canada and the ECODIS project (European Commission’s 6th framework program, subpriority 6.3 “Global Change and Ecosystems”, contract 518043) for funding contributing to this work. References [1]

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