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Use of Algae for Removing Heavy Metal Ions From Wastewater: Progress and Prospects

October 23, 2005

By Mehta, S K; Gaur, J P

ABSTRACT Many algae have immense capability to sorb metals, and there is considerable potential for using them to treat wastewaters. Metal sorption involves binding on the cell surface and to intracellular ligands. The adsorbed metal is several times greater than intracellular metal. Carboxyl group is most important for metal binding. Concentration of metal and biomass in solution, pH, temperature, cations, anions and metabolic stage of the organism affect metal sorption. Algae can effectively remove metals from multi-metal solutions. Dead cells sorb more metal than live cells. Various pretreatments enhance metal sorption capacity of algae. CaCl^sub 2^ pretreatment is the most suitable and economic method for activation of algal biomass. Algal periphyton has great potential for removing metals from wastewaters. An immobilized or granulated biomass-filled column can be used for several sorption/ desorption cycles with unaltered or slightly decreased metal removal. Langmuir and Freundlich models, commonly used for fitting sorption data, cannot precisely describe metal sorption since they ignore the effect of pH, biomass concentration, etc. For commercial application of algal technology for metal removal from wastewaters, emphasis should be given to: (i) selection of strains with high metal sorption capacity, (ii) adequate understanding of sorption mechanisms, (iii) development of low-cost methods for cell immobilization, (iv) development of better models for predicting metal sorption, (v) genetic manipulation of algae for increased number of surface groups or over expression of metal binding proteins, and (vi) economic feasibility.

KEYWORDS algae, biosorption, heavy metal, immobilization, isotherm, ion exchange, seaweed, periphyton

I. INTRODUCTION

Heavy metal pollution of waterbodies due to indiscriminate disposal of industrial and domestic wastes threatens all kinds of inhabiting organisms (De Filippis and Pallaghy, 1994). Therefore, it is necessary to alleviate heavy metal burden of wastewaters before discharging them into waterways. A number of physicochemical methods, such as chemical precipitation, adsorption, solvent extraction, ion exchange, membrane separation, etc., have been commonly employed for stripping toxic metals from wastewaters (Eccles, 1999). However, these methods have several disadvantages, such as incomplete metal removal, expensive equipment and monitoring system requirements, high reagent or energy requirements and generation of toxic sludge or other waste products that require disposal. Further, they may be ineffective or extremely expensive when metal concentration in wastewater is in the range 10-100 mg l^sup -1^.

The use of biological processes for the treatment of metal enriched wastewaters can overcome some of the limitations of physical and chemical treatments and provide a means for cost- effective removal of metals. A great deal of interest has recently been generated using different kinds of inexpensive biomass for adsorbing and removing heavy metals from wastewater (Volesky and Holan, 1995). In this context, accumulation of metals by microorganisms, including algae, has been known for a few decades, but has received increased attention only in recent years because of its potential for application in environmental protection or recovery of precious or strategic metals (Tsezos, 1985, 1986; Volesky, 1987; Malik, 2004). Metal accumulation capacity of algal biomass is either comparable or sometimes higher than chemical sorbents (see Tables 1 and 2). Therefore, algal biomass may serve as an economically feasible and efficient alternative to the existing physicochemical methods of metal removal and recovery from wastewaters. The term “biosorption” has been more frequently used instead of accumulation; however, it should be used for the adsorption of metal ions on the dead biomass, and it includes metal ion binding on extracellular as well as intracellular ligands (Volesky and Holan, 1995; Aksu, 1998). However, the term “biosorption” has been rather loosely used for describing metal accumulation by live biomass as well.

This review explores the prospects of using algae for removing heavy metal ions from wastewaters. After a brief account of accumulation of metals by algae and the mechanisms involved, the paper discusses at length metal removal by algae and possibilities of using this technology on a commercial scale.

II. ACCUMULATION OF METALS BY ALGAE

Algae accumulate high concentrations of metals depending on their concentration in the external environment. The concentration factor for heavy metals varies greatly in different algal species, but it increases as the metal concentration in the water decreases (Kelly, 1988; Sharma and Azeez, 1988). Since the amount of metal accumulated by algae is related with the concentration of metal in water, it may be possible to use metal content of indigenous algae for biomonitoring of metal pollution in a waterbody (De Filippis and Pallaghy, 1994). Furthermore, the inherent metal accumulation capability of algae could be used to alleviate the burden of toxic metal load, and to recover precious metals (e.g., gold and silver) from wastewaters.

The accumulation of heavy metals in algae involves two processes: an initial rapid (passive) uptake followed by a much slower (active) uptake (Bates et al., 1982). During the passive uptake, metal ions adsorb onto the cell surface within a relatively short span of time (few seconds or minutes), and the process is metabolismindependent. Active uptake is metabolism-dependent, causing the transport of metal ions across the cell membrane into the cytoplasm. In some instances, the transport of metal ions may also occur through passive diffusion owing to metal-induced increase in permeability of the cell membrane (Gadd, 1988).

Bates et al. (1982) described a simple method for distinguishing adsorbed from intracellular metal in algal cells. The method involves washing cells with 2 mM EDTA, which removes surface-bound metal, leaving behind the intracellular fractions. The intracellular metal can be quantified after digesting the EDTA-washed cells. The surface adsorbed metal can be quantified by subtracting intracellular metal from the total metal accumulated by the cell. The method has been commonly used by several researchers to distinguish adsorbed from intracellular metal in spite of inevitable removal of a small fraction of intracellular metal during EDTA washing (Bates et al., 1982). The relative importance of metal adsorption and uptake may vary with algal species and metal ions (Trevors, Stratton, and Gadd, 1986). Generally, adsorption contributes much more, even >80% (Mehta, Singh, and Gaur, 2002; Mehta, Tripathi, and Gaur, 2000), than uptake to total metal accumulation by algal cells. However, there are a few reports suggesting a greater contribution of uptake than adsorption or almost similar contribution of both the processes to total metal accumulation (Avery, Codd, and Gadd, 1998). After the first minute of exposure to Cu, >90% of total metal content was found adsorbed on the surface of Scenedesmus subspicatus (Knauer, Behra, and Sigg, 1997). With the passage of time, the relative contribution of the surface-bound Cu declined with a concomitant increase in intracellular Cu, thereby suggesting the transport of adsorbed Cu into the cells. Interestingly, surface adsorption has been invoked as an important mechanism in algae for tolerating elevated levels of heavy metals (Lombardi, Vieira, and Sartori, 2002).

TABLE 1 Recent reports on metal sorption capacity of some algae.

TABLE 1 Recent reports on metal sorption capacity of some algae.

TABLE 1 Recent reports on metal sorption capacity of some algae.

Figure 1 shows the probable sites of an algal cell for the binding of metal ions. Accumulation of metals is due to adsorption onto the cell surface (wall, membrane or external polysaccharides) and binding to cytoplasmic ligands, phytochelatins and metallothioneins, and other intracellular molecules. Localization of metal ions on algal cell has been carried out by electron microscopy and X-ray energy dispersive analysis studies. Spectroscopy has been used for determining the oxidation state of bound metal on algal cell. Transmission electron microscopy has shown cell wall as the most likely location of Cd adsorption by Ectocarpus siliculosus (Winter, Winter, and Pohl, 1994). Scanning electron microscopy, in combination with X-ray microanalysis, has clearly revealed that most of the sites for metal sorption are present on the surface of algal cells (Klimmek et al., 2001).

The algal cell wall has many functional groups, such as, hydroxyl (-OH), phosphoryl (-PO^sub 3^O^sub 2^), amino (-NH^sub 2^), carboxyl (-COOH), sulphydryl (-SH), etc., which confer negative charge to the cell surface. Since metal ions in water are generally in the cationic form, they are adsorbed onto the cell surface (Grist et al., 1981; Xue, Stumm, and Sigg, 1988; Grist, Martin, and Grist, 1991; Romero-Gonzalez, Williams, and Gardiner, 2001; Skowronski and Ska, 2000). Each functional group has a specific pKa (dissociation constant) (Eccles, 1999; Niu and Volesky, 2000), and it dissociates into corresponding anion and proton at a specific pH. These functional groups are found associated with various cell wall components, e.g., peptidoglycan, teichouronicacid, teichoic acids, polysaccharides and proteins. Because distribution and abundance of cell wall components vary among different algal groups, the number and kinds of functional group also vary in different algal groups. Among different cell wall constituents, polysaccharides and proteins have most of the metal binding sites (Kuyucak and Volesky, 1989a).

TABLE 2 Metal sorption capacity of some chemical sorbents.

Peptidoglycan, consisting of linear chains of N- acetylglucosamine and β 1, 4-N-acetylmuramic acid with peptide chain, is the major component of cell wall of cyanobacteria (blue- green algae), providing mostly carboxylic groups for metal binding (Chojnacka, Chojnacki, and Gorecka, 2005). Lipopolysaccharides, lipids and membrane proteins are also important in metal binding by cyanobacteria. Phosphoryl groups are mainly associated with lipopolysaccharides, lipids and peptidoglycans of cell wall. Amino groups are associated with membrane proteins and peptide component of peptidoglycan. It has been suggested that carboxyl groups on the cyanobacterial cell wall are the dominant active sites for the binding of metal ions (Chojnacka, Chojnacki, and Gorecka, 2005). Many strains of cyanobacteria have an outer sheath or capsule made up of polysaccharides. Capsule of cyanobacteria is anionic in nature, owing to presence of uronic acids and/or of other charged groups (De Philippis and Vincenzini, 1998). Due to its anionic nature cyanobacterial capsule usually shows very high affinity towards metal ions and may offer a promising chelating agent for removal of heavy metals from wastewaters (De Philippis et al., 2001). A wide difference in metal binding capacity of capsulated and uncapsulated cyanobacterial strains has been demonstrated (Pradhan and Rai, 2000). A cyanobacterial strain with thick capsule showed higher metal binding capacity than strains devoid of or with a thin layer of capsule (Singh, Pradhan, and Rai, 1998; De Philippis et al., 2003). However, presence of gelatinous capsule around cyanobacterial cell may slow down the diffusion rate of metal ions into the chelating matrix of cell wall (De Philippis et al., 2003).

FIGURE 1 Metal binding sites of a typical algal cell. The alphabet M represents the metal species (independent of its oxidation state).

Cell wall of green algae contains heteropolysaccharides, which offer carboxyl and sulfate groups (Lee, 1980) for sequestration of heavy metal ions. The extracted polysaccharides (12% of dry weight) from Ulva sp. contained 16% sulfate and 15-19% uronic acid (McKinnel and Percival, 1962). Protein content of cell wall of green algae ranges from 10 to 70% (Siegel and Siegel, 1973). Aspartic and glutamic acid account for 12% of protein content of cell wall, which correspond to 0.15 meq g^sup -1^ of carboxylic groups per dry weight. In the cell wall of green algae, lysine and arginine make up 13% of protein (~0.08 meq g^sup -1^ of amino groups) (Chapman, 1980; Lee, 1980; Guing and Blunden, 1991).

Marine algae have been the focus of numerous biosorption studies and their excellent metal binding capacity has been widely acknowledged. The main constituents of the cell wall of brown algae are cellulose, as the fibrous skeleton, and alginate and fucoidan that constitute the amorphous matrix, and extracellular mucilage (Lee, 1980). It is fairly well recognized that alginate is mainly involved in metal accumulation by brown algae (Kuyucak and Volesky, 1989a). Alginate is defined as the ammonium or alkali salt of alginic acid. Following a mild acid hydrolysis of alginic acid, three kinds of segments are demonstrated as building blocks of the polymer. These are D-mannuronic acid and L-guluronic acid units, and alternating D-mannuronic acid and L-guluronic acid residues (Davis, Volesky, and Mucci, 2003). The carboxyl groups of each polymer segment may play an important role as the site for cation binding. Polyguluronic acid shows a high specificity for divalent metal ions (Puranik, Modak, and Paknikar, 1999). Haug (1967) showed that the affinity of alginates for divalent cations such as Pb^sup 2+^, Cu^sup 2+^, Cd^sup 2+^, Zn^sup 2+^, Ca^sup 2+^, etc. increased with guluronic acid content. Alginate content of brown algae is 10 to 40% of dry weight (Percival and McDowell, 1967). Alginate concentration in Sargassum fluitans is 45% of its dry weight, corresponding to 2.25 mmol of carboxyl groups g^sup -1^ of biomass (Fourest and Volesky, 1996). Calpomenia has 5-14% alginate (~0.25-0.7 meq g^sup – 1^ carboxyl groups) (Kalimuthu, Kaliaperumal, and Ramalingam, 1991). Brown algae also contain about 5 to 20% sulfated polysaccharide fucoidan (Chapman, 1980), about 40% of which is sulfate esters. Fucoidan is a branched polysaccharide sulfate ester with L-fucose bulling blocks, which are predominantly α(1 [arrow right] 2) linked. Uptake of trivalent cations has been attributed mainly to the presence of this sulfated polysaccharide (Figueira, Volesky, and Mathieu, 1999). About 0.27 0.03 meq g^sup -1^ sulfonate groups were reported in Sargassum fluitans (Fourest and Volesky, 1996). These latter authors showed that the contribution of sulfonate to heavy metal binding is generally small, but may be significant at low pH. Protein (10% of biomass) contributes ~0.17 meq/g carboxyl groups to Sargassum biomass. Lysine and arginine (10% of protein) offer ~0.07 meq/g amine group. In brown algae, the carboxyl group of alginate is more abundant than either carboxyl or amine groups of protein.

Functional groups involved in metal sorption by algae have been identified, but not fully, by FTIR spectroscopy, pH-titration, potentiometric and conductimetric titration techniques, and also after blocking of functional groups with certain chemical agents. FTIR analysis led Ting, Teo, and Soh (1995) to suggest the involvement of carboxyl groups in binding of Au (III) on Chlorella vulgaris. Participation of carboxyl groups in adsorption of Cu, Cd and Pb on cyanobacterial cell wall has also been demonstrated (Yee et al., 2004). Seki and Suzuki (1998) showed that biosorption of bivalent metal ions by brown algae, namely, Macrocystis pyrifem, Kjellmaniella crassiforia and Undaria pinnatifida, was due to binding to carboxyl groups of alginic acid. On the basis of potentiometric titration, ^sup 13^C-nuclear magnetic resonance (NMR), and chemical analysis, Fourest and Volesky (1997) have demonstrated that heavy metal binding capacities of four different brown seaweeds (S.fluitans, Ascophyllum nodosum, Fucus vesiculosus, and Laminaria japonica) were directly proportional to the number of carboxyl groups. Romero-Gonzalez, Williams, and Gardiner (2001) found up to 95% decrease in Cd biosorption by dealginated seaweed waste after esterification of biomass, and thus concluded that carboxyl groups are largely responsible for Cd biosorption. Furthermore, using FTIR analysis of non-living S. fluitans, Fourest and Volesky (1996) showed binding of Cd and Pb largely on the carboxyl and to a small extent on sulfonate groups. Mehta, Singh, and Gaur (2002) demonstrated the involvement of carboxyl groups in adsorption of Cu onto C. vulgaris. Seki, Suzuki and Iburi (2000) suggested that sorption of bivalent metal ions by a marine microalga Heterosigma akashiwo involves monodentate binding on carboxylic- and phosphatictype sites. A significant role of carboxyl groups in sorption of heavy metals has also been very well demonstrated in fungi as well as in higher plants (Kapoor and Viraraghavan, 1997; Aksu and Dnmez, 2001).

Participation of sulfonate, amino and hydroxyl groups in adsorption of various metal ions has been demonstrated, but to a lesser extent than that of carboxyl groups. On the basis of IR spectroscopy of nonliving Sargassum natans, Kuyucak and Volesky (1989a) demonstrated the binding of Co on carbonyl groups. Potentiometric titration study on Cblamydomonas reinhardtti revealed the presence of amino and thiol groups along with carboxyl groups. According to Greene et al. (1986), amino group plays an important role in binding of Au to C. vulgaris as pretreatment of the alga with succinic anhydride, that is known to react with amino groups making them unavailable for metal binding, resulted in a significant (50%) decrease in Au binding. On the basis of an IR data, Gong et al. (2005) concluded that amino and hydroxyl groups play a predominant role (at high pH) in binding of Pb on Spirulina maxima. Ting, Teo and Soh (1995) studied the mechanism of Au (III) sorption by C. vulgaris using X-ray photoelectron spectroscopy (XPS) analysis. They found elemental gold on the cell surface, and this led them to suggest the involvement of a reduction process in Au sorption. Greene et al. (1986) suggested that sulphydryl groups are not involved in Au binding by Chlorella vulgaris.

The above discussion leads to the generalization that carboxyl groups of cell wall polysaccharides play a predominant role in heavy metal sorption by algae and cyanobacteria. The other functional groups, like sulfonate and amino, play a relatively minor role in metal sorption. Thiol group plays an important role in sorption of metals like Cd at lower pH (>2) (Sheng et al., 2004a). Although various potential fuctional groups have been demonstrated in algae, their mere presence does not guarantee participation in biosorption of metals. Steric hinderance, conformational changes or cross- linking, all of which change with environmental conditions (pH, ionic strength, competing cations or ligands), may prevent some surface functional groups from binding to metal (Adhiya et al., 2002). For example, surface charge may be neutralized either by binding with cations or by cross-linking between oppositely charged surface groups.

Recent studies reveal that the process of heavy metal biosorption involves mechanisms, such as ion exchange, complexation, electrostatic attraction and microprecipitation (Volesky and Holan, \1995). Ion exchange has been shown to be the most important mechanism for the biosorption of metal ions by algal biomass (Grist et al, 1990). Greene, McPherson, and Darnall (1987) observed that the binding of Cu^sup 2+^, Pb^sup 2+^, Zn^sup 2+^, Ni^sup 2+^ and Cr^sup 3+^ to Spimlina platensis was accompanied by the liberation of protons from the biomass. Release of Mg^sup 2+ ^after binding of Cu^sup 2+^ and Zn^sup 2+^ on Osdttatoria anguistissima has been reported by Ahuja, Gupta and Saxena (1999). Williams and Edyvean (1997) observed a concomitant release of Ca^sup 2+^ during biosorption of Ni^sup 2+^ by the brown alga Ecklonia maxima. All these studies suggest that heavy metal sorption involves exchange of metal ions with surface bound protons or cations. Ion exchange capacity is highly variable among different algal species. Pirszel, Pawlik, and Skowronski (1995) reported the maximum ion exchange capacity (in term of Na+/H+ exchange) of five algal species ranging from 41-825 eq g^sup -1^ dry weight. Large variability in ion exchange capacity of different algal forms may be due to variation in cell wall composition. Unicellular algae generally show a higher ion exchange capacity than filamentous forms due to higher surface/ volume ratio (Pirszel, Pawlik, and Skowronski, 1995). Ion exchange capacity and consequently metal sorption capacity, of algal biomass increases with an increase in pH (Pirszel, Pawlik, and Skowronski, 1995; Mehta, Singh and Gaur, 2002). This has been explained as an effect of decreasing competition with protons for the same binding sites (Crist et al., 1994). Some researchers believe that other mechanisms, like complexation (or coordination), are important in metal sorption by algae (Davis, Volesky, and Mucci, 2003). Raiz, Argaman, and Yannai (2004) studied the mechanism of metal binding by Sargassum vulgaris and reported that Cd binding involves chelation, whereas binding of Pb is mediated by a combination of ion exchange and chelation. Adhiya et al. (2002) reported that Cd biosorption to CHamydomonas reinhardtti involves complexation with carboxylic groups. Electrostatic attraction and covalent binding, respectively, mediate Ni and Zn adsorption on Chaetopbora elegans (Andrade, Rollemberg, and Nobrega, 2005). In view of complexicity of the composition of the algal surfaces, it is possible that various mechanisms operate simultaneously to varying degrees depending on algal species and the environmental conditions (Sheng et al., 2004a). Aluminum sorption onto algal cells involves a different kind of mechanism. Al ions bind to biomass in the form of polynuclear Al species such as [Al^sub 6^(OH)^sub 12^(H^sub 2^O)^sub 12^)^sup 6+^ (Hsu and Bates (1964) and Al^sub 13^(OH)^sup 7+^^sub 32^ (Bottero et al., 1980). These polymerized Al ions prevented other heavy metal ions from accessing to binding site.

Once metals are inside the cell, they may bind to intracellular components or precipitate (Gadd, 1988). Biological macromolecules and enzymes with appropriate functional groups or metal co-factors are impacted by metal activity. In Ankistrodesmus falcatus, Wong et al. (1984) have demonstrated that Sn was about 85% in the cellular polysaccharide fraction, 15% in the protein fraction, and 0.2% in lipid and low molecular weight fractions. Metals may be detoxified by accumulation in polyphosphate bodies and in intracellular metal- binding proteins of cyanobacteria and eukaryotic algae (Zhang and Majidi, 1994), and within the vacuoles of some eukaryotic algae (Gadd, 1988; Garnham, Codd, and Gadd, 1992). The presence of one metal may change the distribution of others among cellular components. Okamura and Aoyama (1994) have shown that when present in a mixture, Cd and Cr (VI) influence each other’s concentration and distribution among membrane, cell wall, and soluble and miscellaneous fractions of Chlorella ellipsoidea.

Algal cells have considerable potential in adsorbing anionic species of certain metals. For instance, algal cells can also adsorb Cr (VI) with considerable ease at low pH values (

III. METAL REMOVAL BY ALGAL BIOSORPTION

Limited efforts have been made to use algal biomass for removing metal ions from aqueous solutions. Although thousands of algal species are known, only a few of them have been investigated for their metal sorption ability and subsequent use in wastewater treatment. Metal biosorption experiments have been conducted with freshwater green algae (e.g., Chlorella, spp., Cladophom spp., Scenedesmus spp., Chlamydomonas reinhardtii), brown algae (e.g., Sargassum natans, Fucus vesiculosus, Ascophyllum nodosum, Laminaria japonica), and blue-green algae (like Microcystis aeruginosa and Oscillatoria). Microalgae are easy to grow in culture and some algal species are being grown commercially in large quantity. Metal sorption ability of algae varies greatly from species to species and even among strains of a single species for any metal (Table 1), although this variation may also be due to variable experimental conditions in different studies. A suggestion has also been made that cells grown under different conditions vary with regard to composition of their cell wall, and hence in biosorption characteristics (Chojnacka, Chojnacki, and Gorecka, 2005). Some algae show a high affinity for sorbing a particular metal ion, whereas others do not show such specificity and may sorb several metal ions.

The affinity of various algal species for binding of metal ions shows different hierarchies. In general, metal ions with greater electronegativity and smaller ionic radii are preferentially sorbed by algal biomass. The available literature suggests that Pb is sorbed maximally compared to other metals in a majority of algal species (Tiem, 2002; Davis, Volesky, and Mucci, 2003). Brown algae like Ascophyllum, Sargassum, etc., sorb more metal than other algae due to their high alginate content. In general, metal sorption capacity of algae is the least for Ni, although M. aeruginosa possesses a very high Ni sorption capability (Pradhan et al., 1998). Lee et al. (2000) ? screened 48 species of red, brown and green marine algae for their Cr (VI) adsorbing potential, and found extraordinarily high selectivity of Cr (VI) sorption in Pachymeniopsis sp. (Rhodophyta). This alga was found to be poor in sorbing other heavy metals from water. Metal sorption potential of brown algae has been more thoroughly investigated compared to that of other marine algae and freshwater filamentous algae. Freshwater filamentous diatoms and sheath-forming filamentous blue-green algae, that might bind heavy metals with great efficiency, have somehow failed to receive adequate attention of the researchers.

The metal removal ability of algae is comparable and even at times higher than that of other sorbents (Table 3). Metal removal efficiency of the commonly used ion exchangers and resins is very low at or below 10 mg l^sup -1^ metal concentration in the solution. Nevertheless, algae may almost completely remove metal ions from solutions having low metal concentrations (Mehta and Gaur 2001a; Mehta, Tripathi, and Gaur, 2002). Axtell, Sternberg and Claussen (2003) found 97% removal of Pb by the macroalga Macrospom from a solution having 39.4 mg l^sup -1^ initial concentration of Pb. Many times algae have been found to outperform other biosorbents in removing metals from solutions. Algal biomass could be used in the form of biotraps for the removal of heavy metals from industrial effluents. AlgaSORB, a commercial product, consisting of gel encapsulated algal cell wall, has a remarkable affinity for Hg, Pb, Cd, Cr, Cu, Zn, Ni, Ag, Au, etc. (Darnall, 1989). An important characteristic of AlgaSORB is that high concentrations of common ions do not interfere with sorption of metal ions. Where synthetic resins are inefficient in removing heavy metals from wastewaters, having high concentrations of total dissolved solids (TDS), AlgaSORB could be used successfully to remove toxic metals (Darnall, 1991). Chitoplex, an insoluble cross-linked chitosan, is another commercial biotrap for detoxification of heavy metal-containing industrial wastewaters (Crusberg, Weathers, and Baker, 1991).

A. Factors Affecting Metal Sorption by Algae

Biosorption of heavy metals by algae may be affected by several factors, including concentration of metal and biomass, pH, temperature, presence of competing ions and the metabolic stage of the organism.

TABLE 3 Recent reports on metal removal efficiency of algae and some other materials.

TABLE 3 Recent reports on metal removal efficiency of algae and some other materials.

1. Initial Metal Ion Concentration

Sorption and removal of heavy metals by alg\al biosorbents largely depend on the initial concentration of metals in the solution. Metal sorption initially increases with increase in metal concentration in the solution, and then becoming saturated after a certain concentration of metal (Da Costa and Leite, 1991; Aloysius, Karim, and Arif, 1999; Mehta and Gaur, 2001a,b,c; Mehta, Singh, and Gaur, 2002; Mehta, Tripathi and Gaur, 2002). Algal cell surface has several kinds of functional groups with varying affinity for an ionic species. Low and high affinity functional groups are involved in sorption of metal ions at high and low concentrations of metal ions, respectively. In contrast to sorption, the removal of a metal generally decreases with its increasing concentration in the solution (Mehta and Gaur, 2001a,c). The latter authors found that Chlorella. vulgaris could remove 70 and 80% Ni and Cu, respectively, from their 2.5 mg l^sup -1^ solutions; however, only 37 and 42% Ni and Cu, respectively, were removed from their 10 mg l^sup -1^ solutions.

2. pH

Most of the studies have shown that sorption of metal ions in batch as well as in continuous system is a function of pH of the solution. Therefore, efforts have been made to find out optimum pH for maximizing metal removal by algae. The sorption of Cr (VI) and Cd on Padina sp. and Sargassum sp. (Sheng et al., 2004b), and Cs sorption on Padina australis (Jalali-Rad et al., 2004) was optimal at pH 2. Yu and Kaewsarn (1999) found very little sorption of Cu by Durvillaea potatorum at pH below 2, but it increased with a rise in pH. They found maximum Cu sorption between pH 3 and 4, and the plateau was reached at around pH 5. There are numerous studies showing increased metal sorption with increasing pH of the solution. Zhou, Huang, and Lin (1998) suggested that the optimum pH for Cu and Cd sorption by Laminaria japonica and Sargassum kjellmanianum lies between 4 and 5. Cossich, Tavares and Ravagnani (2002) found maximal Cr (III) sorption capacity of Sargassum sp. at pH 4. Some studies show selective sorption of specific metals due to widely distinct pH optima for their sorption. zer et al. (1994) showed optimum sorption of Pb and Cr (IV) by Cladophora crispata at pH 5.0 and 1.0, respectively. The different optimum pH for the above two metals could be due to variable nature of their chemical interaction with the algal cell. At pH 5.0, cells have a net negative charge on the surface that favors the binding of Pb (II) to surface ligands. At pH below iosoelectric point, cells have a positive charge that inhibits the approach of positively charged Pb (II) ions. Because Cr (IV) is anionic in nature, it does not bind at high pH when the overall algal surface charge is negative. Cr (IV) and other anions bind to algal surface at low pH when algal surface is positively charged. In general, acidic pH (3-5) is most favorable for the sorption of metal ions (Table 4).

Since a majority of the metal binding groups of algae are acidic (e.g., carboxyl), their availability is pH-dependent. These groups generate a negatively charged surface at acidic pH, and electrostatic interactions between cationic species and the cell surface is responsible for metal biosorption. A decreased metal sorption by algae has been frequently observed at extremely acidic pH (

3. Metal Speciation

A large influence of pH on metal sorption can also be related to the availability of free metal ions in test solution. Metals in wastewaters occur in a variety of chemical forms, namely, free aquo ions, complexed with inorganic and organic ligands, and adsorbed on particulate phases. Indeed, toxic effects of metal ions to organisms depend on the concentration of free aquo ions (Sunda and Guillard, 1976; Anderson and Morel, 1982; Sunda and Ferguson, 1983). Likewise, sorption of metals is also a function of concentration of free metal ions in the solution (Volesky and Holan, 1995). Availability of metal ions for binding onto algae depends on chemical speciation, which in turn is determined by pH of the solution.

TABLE 4 Optimum pH for sorption of heavy metals by algae.

TABLE 4 Optimum pH for sorption of heavy metals by algae.

4. Biomass Concentration

The amount of a metal ion recovered from a solution is affected by biomass concentration. Roy, Greenlaw, and Shane (1993) showed that increasing biomass of dried and pulverized cells of Chlorella sp. resulted in decreased Cd binding per unit cell mass. They found 91% decrease in Cd binding by C. minutissima after a 12-fold increase in biomass concentration. Mehta and Gaur (2001a) tested the effect of biomass concentration on Cu and Ni sorption by C. vulgaris at different metal concentrations. Their results show that sorption (metal sorbed per unit biomass) of Cu and Ni was maximal at the lowest tested biomass concentration. The above authors have shown that increase in biomass concentration from 5 mg l^sup -1^ to 1 g l^sup -1^ decreased the Langmuir constant Q^sub max^ (maximum metal sorption capacity) for Cu and Ni by 2.6- and 3-fold, respectively. Mehta and Gaur (2001a) determined the effect of biomass concentration on the affinity of biomass for metal binding. They showed a decrease in affinity of C. vulgaris for Cu and Ni with increase in biomass concentration. Gong et al. (2005) also found a decreased Pb sorption capacity of Spirulina maxima with increasing biomass concentration. They observed that increasing biomass concentration from 0.1 to 20 g l^sup -1^ led to a marked reduction in Pb sorption capacity of Spirulina maxima from 121 to 21 mg g^sup – 1^. Similarly, Hamdy (2000) has noticed a decreased sorption of Cr, Co, Ni, Cu and Cd by four different algae with increasing biomass concentration. Increase in biomass concentration reduces metal sorption per gram of biomass in a number of other systems (Fourest and Roux, 1992; Brady and Duncan, 1994; Singleton and Simmons, 1996; Fogarty et al., 1999), although this is generally attributed to a shift in the sorption equilibrium. Other probable explanations for such a relationship between biomass concentration and sorption may be limited availability of metal, increased electrostatic interactions, interference between binding sites and reduced mixing at higher biomass concentrations (Meikle, Gadd, and Reed, 1990; Fourest, Canal, and Roux, 1994). Itoh, Yuasa, and Kobayashi (1975) suggested that electrostatic interactions between cells could be significant for the biomass-dependent metal sorption, with a larger quantity of metal ions being adsorbed when distance between cells is greater. While an increased biomass concentration has a negative effect on the sorption capacity (amount of metal sorbed per unit biomass) of a biosorbent, the total metal removed (% of initial concentration) by a biosorbent is higher at higher biomass concentrations. However, there is no straightforward relationship between biomass concentration and metal removal. Mehta and Gaur (2001c) have shown that increasing the biomass concentration by 100- fold enhanced Ni and Cu removal from 13 to 65% and 10 to 85%, respectively, from solutions containing 5 mg l^sup -1^ Ni or Cu. Increased metal removal at higher biomass concentration is simply due to greater availability of metal binding sites (Metha and Gaur, 2001c).

5. Temperature

Contrasting results have been obtained regarding the effect of temperature on sorption of heavy metals by algae. Tsezos and Volesky (1981), Kuyucak and Volesky (1989c), and Aksu and Kutsal (1991) reported a slight increase in cation sorption by powdered seaweed biomass with increase in temperature from 4 to 55C. Similarly, Aksu (2002) recorded increased Ni^sup 2+^ biosorption by dried biomass of Chlorella vulgaris with enhancement of temperature from 15C (48.1 mg/ g) to 45C (60.2 mg/g). Increased metal sorption with increase in temperature suggests that metal biosorption by algae is an endothermic process. On the contrary, some studies indicate exothermic nature of metal sorption by algae. For instance, Cd^sup 2+^ sorption by Sargassum sp. biomass slightly decreased with an increase in ambient temperature (Cruz et al., 2004). Some other studies arrive at a similar conclusion (Aksu, 2001; Benguell and Benaissa, 2002). Increased biosorption of heavy metals with increasing temperature has been ascribed to bond rupture that perhaps enhances the number of active sites involved in metal sorption or higher affinity of sites for metal. Park and Gamberoni (1997) suggested that at higher temperatures the ions can be adsorbed more actively on adsorption sites. Interestingly, there are reports showing no effect of temperature on metal sorption (Norris and Kelly, 1979; Zhao, Ha\o, and Ramelow, 1994). Cossich, Tavares, and Ravagnani (2002) studied the effect of temperature on Cr sorption by Sargassum sp. at different pH. They have suggested that the effect of temperature was not as pronounced as was that of pH. Norris and Kelly (1979) and de Rome and Gadd (1987) have shown that sorption of heavy metals is relatively unaffected over a moderate range of temperature. Likewise, Mehta, and Gaur (2001a) and Mehta, Singh, and Gaur (2002) could not observe a pronounced effect of temperature on extracellular binding of Cu and Ni by Chlorella vulgaris.

Owing to dependence on metabolism, metal uptake by live cells is considerably affected by variations in temperature. There are reports showing altered metal uptake by live organisms with change in temperature regime (Skowronski, 1986; Mehta and Gaur, 2001a; Mehta, Singh, and Gaur, 2002), with maximum uptake occurring at specific temperature optima.

6. Presence of Anions and Cations

Metal sorption by an algal biosorbent is also affected by the presence of anions in the medium. Lowered accumulation of U, Co and Cu by green microalgae in the presence of carbonate, orthophosphate, sulphate, nitrate, EDTA and chloride ions has been observed (Rai, Gaur, and Kumar, 1981). However, increased accumulation of ionic species like nitrate and ammonium occurred in the presence of Cu and Fe in Anabaena doliolum (Rai and Mallick, 1992).

The presence of other cations, including metal ions, also significantly affects metal sorption by algae (Mehta and Gaur, 2001a,b; Mehta, Singh and Gaur, 2002; Mehta, Tripathi, and Gaur, 2002). Mehta and Gaur (2001a,b) demonstrated mutual interference in sorption of Cu and Ni by Chlorellavulgaris. However, Axtell, Sternberg, and Claussen (2003) did not find any interactive effect on removal of Pb and Ni by Macrospom from a multi-metal mixture. Actual industrial wastewaters contain different kinds of impurities, which may significantly affect metal biosorption (Ho and McKay, 2000). Among such impurities, light ions exist in most industrial effluents and they greatly affect metal sorption potential of biosorbents (Lee and Volesky, 1997; Chen and Yiacoumi, 1997; Low, Lee, and Liew, 2000). However, Adhiya et al. (2002) did not find inhibitory effect of Ca^sup 2+^ and K+ on Cd sorption by lyophilized Chlamydomonas reingardtii. Reduced metal uptake in the presence of light metals is attributed to competition for cellular binding sites, or precipitation or complexation by carbonates, bicarbonates or hydroxides of Ca and Mg (Rai, Gaur, and Kumar, 1981). Other compounds that could be considered as impurities in metal removal process are surfactants and some chelating agents. The nature of impurities differs depending on the type of effluent to be treated. High concentrations of salts like NaCl in solution also decrease the rate of metal sorption by algae (Cho et al., 1994). Garnham, Codd, and Gadd (1993a) and Corder and Reeves (1994) have shown that Na decreases the accumulation of Ni, Co and Cs by some algae and cyanobacteria. Yun, Niu, and Volesky (2001) also noted suppression of sorption of Cr (VI) and Va by Na and Cl. High concentrations of monovalent cations Na+ and K+, increase the ionic strength of a wastewater. The increased ionic strength of a wastewater often decreases metal biosorption capacity of the biomass (Greene, McPherson, and Darnall, 1987; Ramelow, Fralick, and Zhao, 1992). The inhibitory effect of Na is more pronounced with weakly bound metals such as Zn or Ni. It is important to note that Na+ and K+, being monovalent cations, do not compete directly with covalent binding of heavy metals by the biosorbent. Whereas most of the studies report inhibitory effect of light metal ions on sorption of heavy metals by biosorbent, a few of them show no effect. For example, Pawlik and Skowronski (1994) could not observe any change in Cd transport in Synechocystis in the presence of K. Likewise, Cs sorption capacity of Padina australis did not change in the presence of Na or K in the solution (Jalali-Rad et al., 2004).

7. Other Factors

Nutrient level, growth rate and illumination greatly influence metal sorption by living algae. Hall, Healey, and Robinson (1989) noticed enhanced uptake of Cu by Chlorello, at limiting PO^sup – 3^^sub 4^ concentration. Uptake of Cu, Cd and Zn by Aphanocapso, increased with increase in NO^sup -^^sub 3^ concentration in the medium (Subramanian, Sivasubramanian, and Gowrinathan, 1994). Algal growth increases with increasing light intensity to saturation level, but the influence of light intensity on metal sorption remains largely unknown. There are reports showing inhibition of metal sorption in the dark (Garnham, Codd, and Gadd, 1992; Pawlik and Skowroski, 1994). However, Subramanian, Sivasubramanian, and Gowrinathan (1994) have shown higher Zn uptake in the dark. Metal sorption is also influenced by the growth phase of algal culture. The adsorption of Ni on the surface of C. vulgaris was higher for cultures in the stationary and decline phase than those in the exponential phase (Mehta, Tripathi, and Gaur, 2000). This may have resulted due to better exposure of metal binding sites or creation of additional sites on the cell surface during these phases (Mehta, Tripathi, and Gaur, 2000). Variations in growth conditions possibly bring about changes in composition of the cell surface thereby affecting metal biosorption characteristics of the biomass.

B. Live vs. Inactivated Biomass for Metal Sorption

A majority of studies on metal accumulation by algae have been performed with living organisms for environmental, toxicological and pharmaceutical purposes rather than with industrial application in mind. Of late, the attention has shifted to non-living algae and other microorganisms for metal removal and/or recovery (Volesky and Holan, 1995; Aksu, 1998). In comparison to live cells, the metal sorption capacity of dead cells may be greater, equivalent or less (zer, zer, and Dursun, 2000). The process of using dead cells can be of great interest, because of the large variety and low cost of these biological materials. In general, inactivated cells display greater metal binding capacity than live cells. They do not require a nutrient supply, and therefore can be used for multiple sorption- desorption cycles. However, vacuum drying decreased uranium sorption by Scenedesmus obliquus from 13 to 0.6 mg g^sup -1^ dry weight (Zhang et al., 1997). Decreased metal uptake after vacuum drying of the biomass may be due to disappearance or closure of the effective sites, capillaries or caves for captivating metal ions on the cell wall as water evaporates from the cells under vacuum condition (Zhang et al., 1997). Conversely, it is widely reported that freeze- dried biomass has a higher metal sorption potential than the live biomass (Winter, Winter, and Pohl, 1994; Bengtsson et al., 1995). Neide, Carrihlo, and Gilbert (2000) observed no change in metal sorbing potential of freeze- and oven-dried biomass, when compared with live biomass. Burdin and Bird (1994) studied metal sorption by living as well as lyophilized Gracilaria tikvahiae, Gelidium pusillum, Agardhiella subulata and Chondrus crispus. They found greater accumulation of Pb in living thalli of all the four species compared to lyophilized thalli. However, they noticed greater accumulation of Ni, Cu and Zn by lyophilized than living thalli of Gracilaria, Gelidium and Chondrus. The above authors found no difference between lyophilized and living thalli of the selected seaweeds for accumulating Cd. While contrasting effect of lyophilization has been found, Adhiya et al. (2002) reported that lyophilized cells and living cells of Chlamydomonas reinhardtti showed similar ATR-FTIR spectra. It suggests that lyophilization does not change chemical composition of the cell surface, including cell wall. Metal sorption capacity of dead cells depends on their pretreatment and any subsequent change in the structure of their cell wall (Somers, 1963; Duddridge and Wainright, 1980). The method of inactivation of cells also influences their metal sorption capacity. Whereas heat killing of biomass may enhance sorption of metals, chemical killing may at times decrease the metal sorption capacity of the biomass (Tobin, Cooper, and Neufeld, 1990). Pretreatment of Chlorella vulgaris with dilute HCl considerably increased Ni and Cu sorption from aqueous solution (Mehta and Gaur, 2001a; Mehta, Tripathi, and Gaur, 2002).

Whereas the use of inactivated biomass has been preferred, some disadvantages also deserve mention. Dead cells cannot be used where biological alteration in valency of a metal is sought. Moreover, degradation of organometallic species is not possible with dead biomass. Another important drawback associated with dead biomass is that there is no scope for biosorption improvement through mutant isolation. The use of live biomass has several demerits; potential for desorptive metal recovery is limited since metal may be intracellularly bound, and metabolic extracellular products may form complexes with metals to retain them in solution.

C. Metal Sorption by Pretreated Biomass

Although many biological materials bind heavy metals, only those with a sufficiently high metal binding capacity are suitable in full- scale biosorption process. Various treatments can be given to increase the metal sorption capacity of the biomass (see Table 5). In comparison to fungal and other biosorbents, fewer searches have been conducted for suitable agents and conditions to increase the inherent metal sorption ability of algal biosorbents. Several studies show that CaCl^sub 2^ pretreatment is the most suitable and economic method for activation of algal biomass. CaCl^sub 2^ pretreatment increased Pb sorption capacity of Spirulina maxima by 84-92%. Feng and Aldrich (2004) have also used CaCl^sub 2^ for activation of Ecklonia maxima for Cd, Cu and Pb sorption. Yu \and Kaewsarn (1999) have shown an increased Cu sorption following CaCl^sub 2^ treatment of the marine alga Durvillaeapotatorum. Kaewsarn and Yu (2001) pretreated Padina sp. with 0.2 M CaCl^sub 2^, and noted that the pretreated biomass bound 0.5 mmol g^sup -1^ Cd, and in a column removed 98% of Cd within 35 min. Gong et al. (2005) observed 84-92% increase in Pb sorption capacity of Spirulina maxima after CaCl^sub 2^ pretreatment of the biomass. Next to CaCl^sub 2^, mineral acids (e.g., HCl and HNO^sub 3^) have been found highly effective in increasing metal sorption potential of some algae. On the basis of electron micrograph of pretreated Durvillaea potatorum, Yu, Kaewsarn and Duong (2000) have reported that CaCl^sub 2^ pretreatment followed by thermal treatment provides most uniform cylinder or fiber-like shapes for the biosorbent. Mehta and Gaur (2001a) used several chemicals to enhance metal removal capacity of Chloretta vulgaris. Pretreatment with dilute HCl considerably increased sorption of Cu and Ni by C.vulgaris. Other agents like methanol, acetic acid, NaOH and hot water nevertheless decreased metal sorption. HCl pretreatment of biomass was effective in enhancing metal sorption from single as well as binary metal solutions (Mehta, Tripathi, and Gaur, 2002). Kalyani, Rao, and Krishniah (2004) demonstrated enhanced Ni sorption capacity of Sargassum sp. from 181.2 to 250 mg g^sup -1^ following acid pretreatment. Zhao, Hao, and Ramelow (1994) studied effects of different pretreatments on subsequent sorption of several metal ions from solution by six strains of seaweeds. All the pretreatments increased the ability of the biomass to bind metals (Pb, Cu, Zn, Cd, Cr, Mn, Ni, Co, etc.). Zhang et al. (1997) showed decreased uranium sorption by Scenedesmus obliquus following pretreatment with dilute HCl, NaOH, NaCl and ethanol. They assumed that functional groups taking part in uranium sorption were on the cell surface and also that pretreatment of live cells by alcohol, NaOH, NaCl, or even water perhaps caused contraction of cells, thereby making sites less accessible to metal ions. Decrease in U uptake due to HCl pretreatment of the biomass was probably due to penetration of protons (with smaller radius than uranium) somewhere into the surface and this probably increased rigidity of macromolecules that might otherwise have been involved in metal binding (Zhang et al., 1997). Mehta, Tripathi, and Gaur (2002) demonstrated that acid pretreatment not only increased metal sorption but also was able to alleviate the inhibitory effect of other metal ions on sorption of the metal of interest.

Xanthanation of biomass has been recently introduced as a novel approach to increase biosorption of Pb by algal biomass. Xanthanation of biomass increased sulfur content of the Undaria pinnatifida biomass from 0.15% to 13.7%, and xanthanated biomass sorbed three times more Pb than the native biomass (Kim et al., 1999). Kim et al. (1999) carried out xanthanation of U. pinnatifida in two steps. In the first, hardening of cell wall was carried out by cross-linking using epichlorohyrin. In the second step, xanthanate group was introduced by chemical reaction of the biomass with carbon disulfide. However, Kim et al. (1999) have shown that acid treatment decreased sulfur content of the xanthanated biomass indicating its instability in acidic condition. Due to instability of xanthate in acidic conditions, HCl or HNO^sub 3^ should not be used for desorption of bound metal. Klimmek et al. (2001) described a new method for increasing metal sorption capacity of algal biomass. They introduced phosphate groups in Lyngbya taylori in order to increase the metal removal capacity of the biomass. The additional insertion of phosphate groups into the cell wall was achieved by esterification of hydroxyl groups of the polysaccharides using phosphoric acid in a urea solution. They showed that phosphorylation increased phosphate content of the biomass from 0.6 to 4.4 mmol g^sup -1^ of biomass, thereby enhancing Cd^sup 2+^ sorption capacity of L. taylori sevenfold.

TABLE 5 Effect of biomass pretreatment on its metal sorption efficiency.

D. Use of Immobilized Algae in Sorption of Metals

Algal biomass cannot be used directly in a standard sorption process. It is generally of small particle size and low strength and density, which can limit the choice of a suitable reactor and make biomass or effluent separation difficult (Tsezos, 1986). The use of native or even powdered algal biomass is not desirable as it may clog the column. However, biomass immobilized within or on an inert matrix has the inherent advantage that high flow rate can be achieved with minimal clogging of the column, and also that the control of size of sorbent particle and high biomass loading of the column reactor are feasible (Shumate et al., 1980; Tsezos, 1986; Yakubu and Dudeney, 1986). Efforts have been made to test the efficacy of immobilized algae and cyanobacteria for the removal of heavy metals from aqueous solution. Immobilization is a general term that describes many different forms of cell attachment or entrapment (Lopez, Lazora, and Marques, 1997). Various techniques, such as flocculation, adsorption on surfaces, covalent binding to carriers, crosslinking of cells, encapsulation in polymer gel and entrapment in polymeric matrix, are used for cell or biomass immobilization.

For immobilization of biomass, the supporting material is either natural, including agar, alginate and carrageenan, or synthetic, such as silica gel, polyacrylamide and polyurethanes. Natural polymers are better than synthetic polymers due to the latter’s toxicity to biomass. Alginate has been most extensively used for immobilization of algal as well as other kinds of biomass. Alginate is extracted from algae as water-soluble sodium salt. When calcium replaces sodium, ionic cross-linking between carboxylic acid groups occurs giving a gelatinous substance. It may be noted that alginate beads are vulnerable to low as well as high pH, and are stable when the pH of the external medium is between 5-9. At pH below 3 and above 9, there is a considerable loss of alginate. At pH 11, alginate beads most often break. Treatment of highly acidic effluent by alginate-immobilized biomass may result in shrinkage of biomass- loaded beads. Furthermore, when alginate is used as immobilization support one cannot use acid or alkali during the desorption process. Leakage of biomass is more likely when alginate is used as matrix. Biomass leakage from alginate beads occurs when orthophosphate or Na+ contacts with the alginate immobilized system. This system is therefore not good for the treatment of wastewaters that contain high levels of Na+ or orthophosphate. Among synthetic polymers, polyacrylamide has been most extensively used (Darnall et al., 1986; Robinson and Wilkinson, 1994). Polyacrylamide immobilization is not prone to damage by cation replacement or chelation unlike the commonly used calcium alginate system; however, high cost and toxicity to living cells restricts its application in immobilization process. It has been shown that mechanical strength of organic gel materials, such as alginate and polyacrylamide, is not high enough to permit column operation on a large scale because they may be crushed due to column bed pressure (Hu and Reeves, 1997). Silica gel has also been used for entrapment of algal cells or biomass (Rangsayatorn et al., 2004). Silica gel is prepared by decreasing the pH of alkali silicate to less than 10. The solubility of the silica is then reduced to form the gel. As silica begins to gel, cells in silica are trapped in porous gel that is three-dimensional SiO^sub 2^ network, of which 60% is water.

Akhtar, Iqbal, and Iqbal (2003) developed a new biosorbent by immobilizing Chlorella sorokiniana within luffa sponge discs and showed that microalgal-luffa sponge disc system has good biosorption properties with respect to Ni. Microalgal-luffa sponge immobilized discs removed Ni very rapidly, with 97% of equilibrium loading being reached in 5 min. The regenerated microalgal-luffa sponge immobilized disc retained 92.9% of the initial binding capacity for Ni up to five cycles of reuse in fixed bed column reactor.

Immobilization generally tends to increase metal accumulation by biomass (Darnall et al., 1986; Aksu, 1998; Guo et al., 2000). Immobilized cells accumulate more metals than free cells due inter alia to (i) enhanced photosynthetic capacity (Khummongkol, Canterford, and Freyer, 1982), and (ii) increased cell wall permeability (Brouers et al., 1989). Immobilization of living biomass also provides protection to cells from metal toxicity (Bozeman, Koopman, and Bitton, 1989). On the contrary, some reports show a higher metal sorbing efficiency of free cells compared to immobilized cells (Wong and Pak, 1992). For example Rangsayatorn et al. (2004) showed that free cells of Spirulina platensis accumulated 98 mg g^sup -1^ while alginate and silica immobilized cells accumulated 71 and 37 mg g^sup -1^ Cd, respectively. Bag, Lale, and Turker (1999) obtained similar results. Change in the structure of cell wall during immobilization process is perhaps responsible for decreased metal sorption capacity of the immobilized biomass (Rangsayatorn et al. (2004)). It is also probable that part of cell surface might be shielded by the gel matrix making sites for metal binding unavailable. Size of immobilized bead is a crucial factor for use of immobilized biomass in biosorption process. Beads should not be extremely small as the column filled with them might quickly choke. Due to low pressure drop in column and poor intraparticle mass transfer, bigger beads are also not recommended (Volesky, 2001). It is recommended that beads should be in the size range between 0.7 and 1.5 mm, corresponding to the size of commercial resins meant for removing metal ions (Volesky, 2001).

Since i\mmobilization of a biosorbent is costly per se (Sag and Kutsal, 1998), a few researchers have used dried biomass or converted it into granules by certain chemical treatments (Holan, Volesky, and Prasetyo, 1993; Volesky and Prasetyo, 1994). A few reports are available on cross-linking of cells with various chemical compounds. Several cross-linking materials could be found in literature (Holan, Volesky, and Prasetyo, 1993): aldehydes, polysaccharides, sulfones, vinylketones, epoxides, etc. Each such material has a specific group capable of cross-linking with the cell wall of the biomass. Formaldehyde forms a chemical cross-linking between adjacent hydroxyl groups of sugars of the cell wall. Glutaraldehyde cross-links with amino groups of cell wall due to the presence of prolonged carbon chain. In cross linking with polyethylene imine, biomass is first embedded in polyethylene imine, and free amino groups of the matrix are further cross-linked with glutaraldehyde, which gives further stability to the immobilized system. Glutaraldehyde and polyethylene imine treatments of biomass introduce amino and some other chemical groups thereby altering metal sorption (Leusch, Holan, and Volesky, 1995). However, chemical treatment for cross-linking of biomass may decrease metal sorption (Jalai-Rad et al., 2004).

The immobilized system should be stable, and light should be able to pass through easily if live algal biomass is used. It should not discharge the biomass during treatment. For an immobilized cell system to be effective, the process of immobilization must not cause irreversible, structural, physiological and metabolic damage to the cell. Immobilized biomass in the form of beads should not expand or swell during metal removal. Some algae exist naturally in an immobilized state either as films on surface or entrapped within gel or slime of their own synthesis (Brouers et al., 1989). Such algae can be exploited for removing metals from wastewaters. Although immobilized and granulated algae have great potential in removing toxic metal ions from wastewaters, there is a need to develop an economically feasible method for producing and using them.

E. Metal Sorption From Binary and Multi-Metal Solutions

Many industrial wastewaters contain high levels of more than one metal. For instance, the wastewater arising from fur cleaning and dyeing industry has been shown to contain 7.04, 20.14, 0.17, 1.73 and 0.12 mg l^sup -1^ of Cu, Cr, Ni, Zn and Cd, respectively (Klein et al., 1974). Similarly, Cr and Cu are frequently encountered together in industrial wastewaters, e.g., from mining, metal cleaning, plating, electroplating, metal processing, dyeing and petroleum industries. In metal cleaning, plating and metal processing industries, Cu and Cr concentrations may approach 100- 120 mg 1 and 10-270 g l^sup -1^, respectively (Sag and Kutsal, 1996). Effluent coming out from mining operations have Cu, Zn, Pb, Cr, As and Se together (Volesky, 2001). Cr, Ni, Cd and Zn are reported to occur together in effluents generated during electroplating operations (Volesky, 2001). Most of the industrial effluents have high concentrations of Al along with other metal ions. Although Al is not a major environmental problem, its ubiquitous presence in solution interferes with sorption of many other metals (Lee et al., 2004). While the accumulation of single species of heavy metal ions by algal biomass has been extensively studied, very little attention has been given to the study of multi- metal system. The presence of a multiplicity of metals leads to interactive effects on physiological and biochemical processes, e.g., growth, metal uptake, etc., of various organisms. (Ting, Lawson, and Prince, 1991). Studies on multi-metal systems have shown competitive interactions amongst metals for binding onto sorption sites. Cd and Fe interfere sorption of each other by Sargassum sp from the binary metal solution (Figueira, Volesky, and Ciminelli, 1997). Metals differ with regard to their ability to interfere with sorption of other metals. For example, 1.5 mM Fe caused 24% reduction in Cd biosorption, but at a similar concentration Cd caused up to 45% reduction in sorption of Fe by Sargassum from a binary metal solution (Figueira, Volesky, and Ciminelli, 1997). Lee et al. (2004) reported that 1 mM concentrations of Pb and Al at pH 4.5 mutually reduced Al and Pb uptake by Laminaria japonica to 22.3 and 83%, respectively. Most of the time, total adsorption capacity of biosorbent is lower for binary metal system than the single component due to inhibitory effect of one metal on binding of other metal. Mehta, Tripathi, and Gaur (2000) have shown mutual interference in sorption of Cu and Ni by Chlorella vulgaris from the binary metal solution. Inhibition in Ni sorption due to Cu was stronger than inhibition in Cu sorption due to Ni. Despite the fact that one metal ion present in a solution may decrease sorption of the other metal ion and vice versa, overall metal sorption does not necessarily decrease (Figueira, Volesky, and Ciminelli, 1997). Apiratikul et al. (2004) found that overall adsorption capacity of Caulerpa lentifollifera for Cd and Cu was higher than adsorption of Cd alone, but was lower than adsorption of Cu in single component system. This indicates that binding sites for Cd are not the same as that for Cu. According to them, Caulerpa lentifollifera has common binding sites for Cu and Pb, and Cu and Pb showed competitive inhibition for sorption of each other in the binary system. Further, competitive interaction between two metals depends on the concentration of algal biomass in the solution. Terry and Stone (2002) determined the effect of biomass concentration on sorption of Cd and Cu from binary metal solution by Scenedesmusabundans. They have shown that at the lowest tested biomass concentration, competitive effect was observed at Cu and Cd concentrations above 4 mg l^sup -1^ each. At the highest algal concentration considered, no competitive effect occurred at Cu and Cd concentration in the ra




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