application of ion for the removal of organic pollutants

Process Biochemistry 40 (2005) 997–1026

Review

Application of biosorption for the removalof organic pollutants: a review

Zümriye Aksu∗

Department of Chemical Engineering, Hacettepe University, Beytepe, Ankara 06532, Turkey

Received 30 September 2003; received in revised form 26 March 2004; accepted 4 April 2004

Abstract

In modern society, an increasing number of hazardous organic compounds are being discharged into the environment. Most are degradedor detoxificated by physical, chemical and biological treatments before released into the environment. Although the biological treatments area removal process for some organic compounds, their products of biodegradation may also be hazardous. Moreover, some nondegradablecompounds discharged into the environment along with the treated compounds can cause problems because they usually come back tohumanbeings through the several channels such as bioaccumulation. As a result, organic molecules that are not biodegradable, can still beremoved from the wastewater by the microbial biomass via the process of biosorption. Biosorption is also becoming a promising alternativeto replace or supplement the present removal processes of organic pollutants from wastewaters. Among these pollutants, dyes, phenolics andpesticides have recently been of great concern because of the extreme toxicity and/or persistency in the environment. Biosorption of thesetype of hazardous organics by selected live and dead microoganisms has been investigated by various workers. This review examines a widevariety of microorganisms (fungi, yeasts, bacteria, etc.), which are capable of uptake of organic pollutants, discusses various mechanismsinvolved in biosorption, discusses the effects of various parameters such as pH, temperature, concentrations of organic pollutant, other ions,and biomass in solution, pretreatment method, etc. on biosorption, reports some elution and regeneration methods for biomass; summarizes theequilibrium and kinetic models used in batch and continuous biosorption systems which are important to determine the biosorption capacityof microorganism and to design of treatment processes.© 2004 Elsevier Ltd. All rights reserved.

Keywords:Biosorption; Organic pollutant; Microorganism; Batch system; Continuous system; Equilibrium; Kinetics

1. Introduction

A great number of industry such as textile, paper and pulp,printing, iron-steel, coke, petroleum, pesticide, paint, sol-vent, pharmaceutics, wood preserving chemicals, consumelarge volumes of water, and organic based chemicals. Thesechemicals show a great difference in chemical composition,molecular weight, toxicity, etc. Effluents of these industriesmay also contain undesired quantities of these pollutants andneed to be treated.

Synthetic dyestuffs, one group of organic pollutants,are used extensively in textile, paper, printing industriesand dyehouses. It is reported that there are over 100,000commercially available dyes with a production of over7× 105 metric tonnes per year[1,2]. Dyeing industry efflu-

∗ Tel.: +90-312-2977434; fax:+90-312-2992124.E-mail address:[email protected] (Z. Aksu).

ents constitute one of the most problematic wastewaters tobe treated not only for their high chemical and biologicaloxygen demands, suspended solids and content in toxiccompounds but also for colour, which is the first contami-nant to be recognized by human eye. Dyes may significantlyaffect photosynthetic activity in aquatic life due to reducedlight penetration and may also be toxic to some aquaticlife due to the presence of aromatics, metals, chlorides,etc., in them[1–6]. Dyes usually have a synthetic originand complex aromatic molecular structures which makethem more stable and more difficult to biodegrade. Dyesare classified as follows: anionic—direct, acid and reac-tive dyes; cationic—basic dyes; non-ionic—disperse dyes[3,5]. The chromophores in anionic and non-ionic dyes aremostly azo groups or anthraquinone types. The reductivecleavage of azo linkages is responsible for the formation oftoxic amines in the effluent. Anthraquinone-based dyes aremore resistant to degradation due to their fused aromatic

0032-9592/$ – see front matter © 2004 Elsevier Ltd. All rights reserved.doi:10.1016/j.procbio.2004.04.008

998 Z. Aksu / Process Biochemistry 40 (2005) 997–1026

structures and thus remain coloured for a longer time inthe wastewater. Reactive dyes are typically azo-based chro-mophores combined with different types of reactive groupse.g., vinyl sulfone, chlorotriazine, trichloropyrimidine, di-fluorochloropyrimidine. They differ from all other classesof dyes in that they bind to the textile fibres such as cot-ton to through covalent bonds. They are used extensivelyin textile industries regarding favorable characteristics ofbright colour, water-fast, simple application techniques withlow energy consumption. Water-soluble reactive and aciddyes are the most problematic, as they tend to pass throughconventional treatment systems unaffected. Hence, their re-moval is also of great importance[6–13]. Basic dyes havehigh brilliance and intensity of colours and are highly vis-ible even in a very low concentration[1,4,5,14–16]. Metalcomplex dyes are mostly based on chromium, which iscarcinogenic[1,3,4,17]. Disperse dyes do not ionize in anaqueous medium and some disperse dyes have also beenshown to have a tendency to bioaccumulate[4]. Due tothe chemical stability and low biodegradability of thesedyes, conventional biological wastewater treatment systemsare inefficient in treating dye wastewater. Dye wastewateris usually treated by physical- or chemical-treatment pro-cesses. These include chemical coagulation/flocculation,ozonation, oxidation, ion exchange, irradiation, precipita-tion and adsorption[3–6,8,17–27]. Some of these tech-niques have been shown to be effective, although they havelimitations. Among these are: excess amount of chemicalusage, or accumulation of concentrated sludge with ob-vious disposal problems; expensive plant requirements oroperational costs; lack of effective colour reduction; andsensitivity to a variable wastewater input. In recent years, anumber of studies have focused on some microorganismswhich are able to biodegrade or to bioaccumulate azo dyesin wastewaters. A wide variety of microorganisms includ-ing bacteria, fungi and algae are capable of decolourizing awide range of dyes via anaerobic, aerobic, and sequentialanaerobic–aerobic treatment processes[4–6,28,29]. Cyto-plasmic azo reductases play an important role in the anaer-obic biodegradation of azo dyes to produce colourless aro-matic amines although complete mineralization is difficultand the resulting aromatic amines may be toxic and car-cinogenic. These amines are resistant to further anaerobicmineralization. Fortunately, once the xenobiotic azo com-ponent of the dye molecule has been removed, the resultantamino compounds are good substrates for aerobic biodegra-dation suggesting a choice of a sequential anaerobic–aerobicsystem for wastewater treatment. Although a number ofaerobic biological processes for the removal of dyes fromtextile effluents have been explored, such as decolourizationthrough liquid fermentations by white-rot fungi (amongthese microorganismsPhanerochaete chrysosporium, Tram-etes versicolor, Coriolus versicolor biotransformed ormineralized several azo dyes through catalytic action ofextracellular enzymes, such as lignin peroxidases and man-ganese dependent peroxidase) and bacterial cultures (such

asPseudomonasstrains, mixed bacterial cultures,Bacillussubtilis), yeasts (such asKlyveromyces marxianus, Candidazeylanoides), biochemical oxidation suffers from significantlimitations since more dyestuffs found in the commercialmarket have been intentionally designed to be resistant toaerobic microbial degradation. Reactive azo dyes are elec-tron deficient in nature and this property makes them lesssusceptible to oxidative catabolism. Research has shownthat the efficiency of biological treatment systems is greatlyinfluenced by the operational parameters, the compositionof textile wastewater and the structure and substituents ofdye molecule. The level of aeration, temperature, pH, andredox potential of the system are the variables that shouldbe optimized to produce the maximum rate of dye reduc-tion. To test the ability of microorganisms to reduce dyesfrom a range of dye classes (acidic, basic, direct, disperse,metal-complex, reactive) is also important to determine thetypes of wastewater that can be treated by the system. Thecomposition of textile wastewater is varied and can includeorganics, nutrients, salts, sulfur compounds and toxicantsas well as the colour, so the inhibitory effect of any ofthese compounds on the dye reduction process should beinvestigated[4–6,28,30–35]. The other biological treatmentmethod; bioaccumulation is defined as the accumulation ofpollutants by actively growing cells by metabolism- andtemperature-independent and metabolism-dependent mech-anism steps. Although bioaccumulation of dyes by yeasts[36,37] were accomplished, however, there are significantpractical limitations regarding the inhibition of cell growthat high dye concentrations and requirement of metabolicenergy externally provided. So there is a need to find al-ternative treatment methods that are effective in removingdyes from large volumes of effluents and are low in cost,such as biosorption.

Phenols, another group of organic pollutants are consid-ered as priority pollutants since they are harmful to organ-isms at low concentrations and many of them have beenclassified as hazardous pollutants because of their potentialto harm human health. It should be noted that the contami-nation of drinking water by phenolics at even a concentra-tion of 0.005 mg l−1 could bring about significant taste andodor problems making it unfit for use. Human consumptionof phenol-contaminated water can cause severe pain leadingto damage of the capillaries ultimately causing death. Phe-nol containing water, when chlorinated during disinfectionof water also results in the formation of chlorophenols. Themost important pollution sources containing phenols andphenolic compounds such as nitrophenols, chlorophenols,are the wastewaters from the iron-steel, coke, petroleum,pesticide, paint, solvent, pharmaceutics, wood preservingchemicals, and paper and pulp industries[38–41]. Currentmethods for removing phenolics from wastewater includemicrobial degradation, adsorption on activated carbon,chemical oxidation (using agents such as ozone, hydrogenperoxide or chlorine dioxide), deep-well injection, inciner-ation, solvent extraction and irradiation. Solvent extraction

Z. Aksu / Process Biochemistry 40 (2005) 997–1026 999

methods are expensive, and deep-well injection may leadto contamination of ground water. Adsorption and oxida-tion treatments become exceedingly expensive when loweffluent concentrations must be achieved. Wastewaters con-taining phenol in the range of 5–500 mg l−1 are consideredsuitable for treatment by biological processes. Although bi-ological treatment has shown great promise, a wide varietyof pure and mixed cultures of microorganisms are capableof degrading phenol and phenolics under both aerobic andanaerobic conditions and microbial degradation of thesecompounds is seen as a cost effective method, the biologicaltreatment of phenolics is limited by the intrinsic propertiesof these compounds owing to their toxicity; they are slowto biodegrade, and the degrading microorganism must beexposed to only low concentrations of the substrates. There-fore, alternative technologies have to be explored[42–55].

Advancing increase of production and application of pes-ticides for agriculture as well as for plant protection andanimal health has caused the pollution of soil, ground andsurface water which involves a serious risk to the environ-ment and also to the human health due to direct exposure orthrough residues in food and drinking water. In the world,alarming levels of pesticides have been reported in air, wa-ter, soil as well as in foods and biological materials. Someof these pesticides have been reported to be persistent, toxic,mutagenic, carcinogenic, and tumorogenic. Pesticide con-tamination of water systems has been of major concern in re-cent years. Pesticide residues reach the aquatic environmentthrough manufacturing plants, direct surface run-off, leach-ing, careless disposal of empty containers, equipment wash-ings, etc. Pesticides are divided into many classes, of whichthe most important are organochlorine and organophospho-rous compounds. The chemical stability of organochlorinecompounds is reflected in their resistance to microbial degra-dation. The lipophilic nature, hydrophobicity and low chem-ical and biological degradation rates of organochlorine pes-ticides have led to their accumulation in biological tissuesand subsequent magnification of concentrations in organ-isms progressing up the food chain. Organophosphorouspesticides on the other hand are known to degrade rapidlydepending on their formulation, method of application, cli-mate and the growing stage of the plant. The most importantand common pollutants among organochlorine pesticides aredichlorodiphenyltrichloroethane and its metabolites (DDTs),polychlorinated biphenyls (PCBs), hexachlorocyclohexaneisomers (HCHs), chlordane related compounds (CHLs), hex-achlorobenzene (HCB), cyclodienes, dieldrin, etc. The widerange of pesticides used makes it extremely difficult to pro-duce a single method for pesticide disposal that appliesuniversally. Photochemical or chemical treatments are fre-quently inefficient for the removal of synthetic organochlo-rine pesticides from waters and may also lead to hazardousfinal products. On the other hand, advanced water treatmentprocesses, and mainly the adsorption onto activated carbon,have proved to be the most efficient and reliable methodfor the removal of aqueous-dissolved organic pesticides. In

recent years the ability of microorganisms to metabolizesome pesticides has also received much attention due tothe environmental persistence and toxicity of these chemi-cals. Although in some cases, microbial metabolism of con-taminants may produce toxic metabolites, a variety of mi-croorganisms (many aerobic bacteria and fungi) are knownto utilize organic pesticides as the sole carbon or energysource, such asPseudomonas pickettii, Alcalilgenes eutro-phus, Desulfomonile tiedjei, Phanerochaete chrysosporium,etc. However, conventional activated sludge systems oftenfail to achieve high efficiency in removing pesticides fromwastewater due to the low biodegradability and toxicity orinhibition of organic pesticides to microorganisms[56–63].

Adsorption has been shown to be the most promisingoption for all these non-biodegradable organics for the re-moval from aqueous streams, activated carbons being themost common adsorbent for this process due to its ef-fectiveness and versatility. Activated carbons are usuallyobtained from materials with a high carbon content andpossess a great adsorption capacity, which is mainly deter-mined by their porous structure. Although activated carbon,in granular or powdered form has a good capacity for theadsorption of organic molecules, it suffers from a numberof disadvantages. Activated carbon is quite expensive andthe higher the quality the greater the cost. Both chemicaland thermal regeneration of spent carbon is expensive, im-practical on a large scale and produces additional effluentand results in considerable loss of the adsorbent. This hasled many workers to search for the use of cheap and effi-cient alternative materials such as bagasse pith, carbonizedbark, peat, soil, tree, and eucalyptus barks, chitin, rice husk,wood, fly ash, and carbonized sewage sludge. However,these low-cost adsorbents have generally low adsorptioncapacities so large amounts of adsorbents will be needed[5,6,8,9,17–27,38–50,56–62]. Alternatively, the so-calledbiosorption, i.e. the passive uptake of pollutants from aque-ous solutions by the use of non-growing or non-livingmicrobial mass, thus allowing the recovery and/or environ-mentally acceptable disposal of the pollutants, could alsobe considered. The special surface properties of bacteria,yeasts, fungi and algae enable them to adsorb different kindsof pollutants from solutions. “Biosorption” term is usedto indicate a number of metabolism-independent processes(physical and chemical adsorption, electrostatic interaction,ion exchange, complexation, chelation, and microprecip-itation) taking place essentially in the cell wall ratherthan oxidation though anaerobic or aerobic metabolism(biodegradation). The main attractions of biosorption arehigh selectivity and efficiency, cost effectiveness and goodremoval performance; raw materials which are either abun-dant (sea weeds) or wastes from other industrial operations(fermentation wastes, activated sludge process wastes) canbe used as biosorbents presenting performances often com-parable with those of ion exchange resins. Both living anddead (heat killed, dried, acid and/or otherwise chemicallytreated) biomass can be used to remove hazardous organics,


but maintaining a viable biomass during adsorption is diffi-cult, because it requires a continuous supply of nutrients andavoidance of organic toxicity to the microorganisms. Theuse of dead microbial cells in biosorption is more advan-tageous for water treatment in that dead organisms are notaffected by toxic wastes, they do not require a continuoussupply of nutrients and they can be regenerated and reusedfor many cycles. Dead cells may be stored or used for ex-tended periods at room temperature without putrefactionoccurring. Their operation is easy and their regeneration issimple. Moreover, dead cells have been shown to accumu-late pollutants to the same or greater extent than growingor resting cells. The mechanism of binding by inactivatedbiomass may depend on the chemical nature of pollutant(species, size, ionic charge), type of biomass, its prepara-tion and its specific surface properties and environmentalconditions (pH, temperature, ionic strength, existence ofcompeting organic or inorganic ligands in solution). Ashydrophobic organic pollutants show a high tendency toaccumulate onto microbial cells or sludge, the microbialbiomass could be used as an adsorbent of biological originfor the removal of very low concentration hazardous organ-ics from the wastewater[4–7,10–12,14–16,39–42,64–109].Although biosorption is generally used for the treatment ofheavy metal pollutants in wastewaters[64–76], it can alsobe considered a promising technology for the removal oforganics from industrial waste streams and polluted naturalwaters[4–7,10–12,14–16,39–42,77–113].

A great deal of the biosorption studies are performedin batch systems with single species of organics. Theseprocesses are conceptually simple. A suitable microbialbiomass is contacted with aqueous solution containing or-ganic pollutant molecules or ions. The contacting process isallowed to proceed for a sufficient time for the biomass tosequester these molecules and to reach equilibrium. Thenthe biomass is separated from the liquid phase and thepollutant-containing biomass is either regenerated or dis-posed in an environmentally acceptable manner. A majorconsideration with any biosorption scheme is the sepa-ration of liquid and solids after batch or counter currentcontacting. Centrifugation of filtration, as routinely used inthe laboratory, are not generally practical in industrial pro-cesses; thus, continuous systems such as continuous stirredtank reactors, fluidized bed, moving bed and packed bedcolumns must be used. For continuous operation, the mostconvenient configuration is that of a packed column, muchlike that used for ion exchange. This operating method en-sures the highest possible concentration difference drivingforce. Continuous packed bed sorption has a number ofprocess engineering advantages including high yield op-erations and relatively easy scaling up from a laboratoryscale procedure. The stages in the separation protocol canalso be automated and high degrees of purification can of-ten be achieved in a single step process. A large volumeof wastewater can be continuously treated using a definedquantity of biosorbent in the column. Reuse of microorgan-

ism is also possible. After pollutant loading the pollutantmay be concentrated in a small volume of solid materialor desorbed into a small volume of eluant for recovery,disposal or containment. However, the use of dead biomassin powdered form in the column has some problems, suchas difficulty in the separation of biomass after biosorption,mass loss after regeneration, low strength and density andsmall particle size, which make it difficult to use in col-umn applications. To solve these problems, dead biomasscan be immobilized in a supporting material. Researchershave recognized that immobilizing nonliving biomass in abiopolymeric or polymeric matrix may improve biomassperformance, biosorption capacity, increase mechanicalstrength and facilitate separation of biomass from pollu-tant containing solution. Immobilization also allows higherbiomass concentration, resistance to chemical environmentsand column operations and immobilized systems may bewell suited for non-destructive recovery. Indeed, the use ofimmobilized biomass has a number of major disadvantages.In addition to increasing the cost of biomass pre-treatment,immobilization adversily affects the mass transfer kineticsof organics uptake. When biomass is immobilized the num-ber of binding sites easily accessible to organic moleculesor ions in solution is greatly reduced since the majority ofsites will lie within the bead[64,71,108].

2. Modeling of biosorption in batch and continuoussystems

2.1. Equilibrium modeling of biosorption in a batch system

Equilibrium data, commonly known as adsorptionisotherms, are basic requirements for the design of biosorp-tion systems used for the removal of organic pollutants.The Langmuir, Freundlich, Langmuir–Freundlich, Redlich–Peterson, Brunauer–Emmet–Teller (BET), Radke–Prausnitzare the most frequently used two- and thee-parameters mod-els in the literature describing the non-linear equilibriumbetween adsorbed organic pollutant on the cells (qeq) andorganic pollutant in solution (Ceq) at a constant temperature.

The Langmuir equation which is valid for monolayer sorp-tion onto a surface with a finite number of identical sites isgiven byEq. (1).

qeq = Q0bCeq

1 + bCeq(1)

where parametersQ0 andb are Langmuir constants relatedto maximum adsorption capacity (monolayer capacity) andbonding energy of adsorption, respectively, which are func-tions of the characteristics of the system as well as time[114]. Q0 andb can be determined from the linear plot ofCeq/qeq versusCeq.

The Langmuir equation is used for homogeneous sur-faces. The Freundlich isotherm model assumes neither ho-


mogeneous site energies nor limited levels of sorption. TheFreundlich equation has the general form:

qeq = KFC1/neq (2)

where KF and n are the Freundlich constants related toadsorption capacity and adsorption intensity, respectively[115]. Eq. (2)can be linearized in logarithmic form and Fre-undlich constants can be determined.

The Langmuir–Freundlich model is essentially a Fre-undlich isotherm which approaches an adsorption maximumat high concentrations of adsorbate. An equation mathemat-ically equivalent to the Langmuir–Freundlich equation canalso be obtained by assuming that the surface is homoge-neous, but that the adsorption is a cooperative process dueto adsorbate–adsorbate interactions. The following relationrepresent this model:

qeq = Q0bC1/neq

1 + bC1/neq

(3)

For positive interactions (1/n > 1), the Langmuir–Freundlichconverts to Hill equation[39].

A further empirical model has been developed byRedlich–Peterson to improve the fit by the Langmuir orFreundlich equation[116] and is given byEq. (4).

qeq = KRPCeq

1 + aRPCβeq

(4)

whereKRP, aRP, andβ are the Redlich–Peterson parameters.The exponentβ lies between 0 and 1. Forβ = 1 Eq. (4)converts to the Langmuir form.

The Radke–Prausnitz isotherm model[117] is given as:

qeq = arCpeq

a+ rCp−1eq

(5)

wherea, r, andp are related model constants.The theoretical BET model for multilayer sorption[118]

is:

qeq = BQ0Ceq

(Cs − Ceq)[1 + (B − 1)(Ceq/Cs)](6)

where Cs is the saturation concentration of the adsorbedcomponent;B a constant indicating the energy of interactionbetween the solute and the adsorbent surface, andQ0 is aconstant indicating the amount of solute adsorbed forminga complete monolayer.

In some cases the relationship between equilibrium con-centrations of organics in liquid and solid phases could belinear and defined by simple distribution coefficients. Inthis case the adsorption data are fitted to linear adsorptionisotherm as described inEq. (7) [43,96,97].

qeq = KdCeq (7)

whereKd is distribution coefficient.

When several components are present, interference andcompetition phenomena for adsorption sites occur and leadto a more complex mathematical formulation of the equi-librium. Several isotherms have been proposed to describeequilibrium and competitive adsorption for such a system.These isotherms range from simple models related to theindividual isotherm parameters only (such as competitiveRedlich–Peterson isotherm model), to more complex mod-els related to the individual isotherm parameters and to cor-rection factors (such as modified competitive Langmuir andmodified Freundlich isotherm models).

The modified competitive Langmuir isotherm is writtenas:

qeqi = Q0i bi(Ceqi /ηi)

1 +∑Nj=1bj(Ceqj /ηj)

(8)

whereCeqi and qeqi are the unadsorbed concentration ofeach component at equilibrium and the adsorbed quantityof each component per g of dried biomass at equilibrium,respectively.bi andQ0

i are derived from the correspondingindividual Langmuir isotherm equations.ηi is the Langmuircorrection coefficient of thei component where estimatedfrom competitive adsorption data[119].

The empirical extended form of the Freundlich modelrestricted to binary mixtures can be given byEqs. (9) and(10) for each component of binary system:

qeq1 =KF1C

1/n1+x1eq1

Cx1+y1eq1 C

z1eq2

(9)

qeq2 =KF2C

1/n2+x2eq2

Cx2+y2eq2 C

z2eq1

(10)

whereKF1, KF2 andn1 andn2 are derived from the corre-sponding individual Freundlich isotherm equations and thesix other parameters are the competition coefficients for twospecies[119].

The competitive Redlich–Peterson model related to theindividual isotherm parameters only is given as follows:

qeqi = KRPiCeqi

1 +∑Nj=1aRPj (Ceqj )

βj(11)

where KRPi , aRPi , and βj are the Redlich–Petersonparameters derived from the corresponding individualRedlich–Peterson isotherm equations[119].

2.2. Kinetic modeling of biosorption in a batch system

If the movement of organic pollutant molecule from thebulk liquid to the liquid film or boundary layer surroundingthe biosorbent is ignored, the following sequence of stepscan take place in the biosorption process of porous biosor-bent: transport of solute molecules from the boundary film tothe external surface of the biosorbent (film diffusion), trans-fer of molecules from the surface to the intraparticular active


sites and uptake of molecules by the active sites of sorbent.In the removal of organics from wastewater, it is importantfor design purposes to investigate the mechanisms of ad-sorption and potential rate controlling steps which controlthe adsorption rate. In order to find the contribution of ratecontrolling steps such as external mass transfer, intraparticlediffusion, and adsorption process and also, mass transfer andkinetic models have been used to test the experimental data.

In the first step of adsorption, the film diffusion is an im-portant rate-controlling step and external mass transfer orboundary layer diffusion can be characterized by the initialrate of solute sorption. In this case, adsorption rate is ex-pected to be proportional to the first power of concentration;it means that this step is a first-order process and can bedefined as:

dC

dt= k1,adC (12)

where C is the pollutant concentration in the wastewa-ter remaining at each contact time andk is the first-orderreaction-rate constant. After integration and applyingboundary conditions,t = 0 to t = t andC = C0 to C = C;the integrated form ofEq. (12)becomes

logC0

C= 1

2.303k1,adt (13)

whereC0 is the initial pollutant concentration[15].If intraparticular diffusion is involved in the sorption pro-

cess, the model developed by Weber and Morris[120] canbe used to find the region where intraparticle diffusion israte-limited and to determine intraparticular diffusion rate.In this model, the rate of intraparticular diffusion is a func-tion of t0.5 and can be defined as follows:

q = f

(D t

r2p

)0.5

= K t0.5 (14)

whererp is particle radius,D is the effective diffusivity of so-lute within the particle, andK intraparticular diffusion rate.If intraparticle diffusion is rate-limited; then a plot of adsor-bate uptake (q) versus the square root of time (t0.5) wouldresult in a linear relationship andK value can be obtainedfrom this plot. Moreover, the particle diffusion would be therate-controlling step if the line passes through the origin.

In many cases, the kinetics of biosorption based onoverall adsorption rate by biosorbents are described by thefirst-order Lagergren[121] and pseudo second-order[122]kinetic models.

The first-order rate expression of Lagergren based on thesorption capacity of adsorbent is generally expressed as fol-lows:

dq

dt= k1,ad(qeq − q) (15)

whereq is the amount of adsorbed pollutant on the biosor-bent at timet, and k1,ad is the rate constant of Lager-gren first-order biosorption. After integration and applying

boundary conditions,t = 0 to t = t andq = 0 to q = q; theintegrated form ofEq. (15)becomes

log(qeq − q) = logqeq − k1,ad

2.303t (16)

A straight line of log(qeq−q) versust suggests the appli-cability of this kinetic model. In order to fitEq. (16)to ex-perimental data, the equilibrium sorption capacity,qeq, mustbe known. In many casesqeq is unknown and as adsorptiontends to become unmeasurably slow, the amount sorbed isstill significantly smaller than the equilibrium amount. Forthis reason it is necessary to obtain the real equilibrium sorp-tion capacity,qeq, by extrapolating the experimental data tot = ∞ or by using a trial and error method. Furthermore,in most cases the first-order equation of Lagergren does notfit well for the whole range of contact time and is generallyapplicable over the initial 20–30 min of the sorption process.

The pseudo second-order equation is also based on thesorption capacity of the solid phase. Contrary to the othermodel it predicts the behaviour over the whole range ofadsorption. The pseudo second-order kinetic rate equationis expressed as

dq

dt= k2,ad(qeq − q)2 (17)

wherek2,ad is the rate constant of second-order biosorption.For the boundary conditionst = 0− t = t andq = 0− q =q; the integrated and linear form ofEq. (17)becomes

t

q= 1

k2,adq2eq

+ 1

qeqt (18)

If second-order kinetics are applicable, the plot oft/qagainstt of Eq. (18)should give a linear relationship, fromwhich qeq andk2,ad can be determined from the slope andintercept of the plot and there is no need to know any pa-rameter beforehand.

2.3. Kinetic modeling of biosorption in a continuouspacked bed system

When an organic pollutant containing solution passesthrough a packed bed column, at the beginning, most oforganic pollutant get sorbed on biosorbent so organic pollu-tant concentration in the effluent remains either very low orin some cases is not detectable. As biosorption continues,organic pollutant concentration in the effluent rises, slowlyat first, and then abruptly. When this abrupt rise or break-through occurs, the flow is stopped. The performance ofcontinuous packed bed is described through the concept ofthe breakthrough curve. The time for breakthrough appear-ance (breakthrough time) and the shape of the breakthroughcurve are very important characteristics for determiningthe operation and the dynamic response of a biosorptioncolumn. The general position of the breakthrough curvealong the time or volume axis depends on the capacity ofthe column with respect to the feed concentration and flow


rate. The breakthrough curves show the loading behaviourof pollutant to be removed from solution in a fixed bed andis usually expressed in terms of normalized concentrationdefined as the ratio of effluent pollutant concentration toinlet pollutant concentration (C/C0) as a function of time orvolume of effluent (Veff ) for a given bed height.

Successful design of a column adsorption process requiresprediction of the concentration-time profile or breakthroughcurve for the effluent under given specific operating con-ditions. Developing a model to accurately describe the dy-namic behaviour of adsorption in a fixed bed system is in-herently difficult. Since the concentration of the adsorbateas the feed moves through the bed, the process does not op-erate at steady state. The fundamental transport equationsderived to model the fixed bed with theoretical rigor aredifferential in nature and usually require complex numeri-cal methods to solve. Such a numerical solution is not usu-ally difficult, but often does not fit experimental results es-pecially well. Some solutions for very limiting cases havebeen reported, but in general, complete time-dependent an-alytical solutions to differential equation based models ofthe proposed rate mechanisms are not available. Because ofthis, various simple mathematical models have been devel-oped to predict the dynamic behaviour of the column andthe following models used in the literature to characterizethe fixed bed performance for the removal of organics arepresented here[41,64,71,108,123].

The Adams–Bohart model is used for the description ofthe initial part of the breakthrough curve and the linear formof model is given byEq. (19):

lnC

C0= kAB C0 t − kABN0

Z

U0(19)

wherekAB is the Adams–Bohart kinetic constant,N0 is thesaturation concentration andZ is the column height. Fromthis equation, values describing the characteristic operationalparameters of the column can be determined from a plot ofln C/C0 againstt at a given bed height and flow rate[124].

Besides the prediction of the concentration–time profileor breakthrough curve for the effluent, the maximum adsorp-tion capacity of an adsorbent is also needed in design. Tra-ditionally, the Thomas model is used to fulfil the purpose.The model has the following form:

C

C0= 1

1 + exp(kTh/Q(q0X− C0Veff))(20)

wherekTh is the Thomas rate constant,q0 is the maximumsolid-phase concentration of the solute,Q is the flow rateandX is the amount of sorbent in the column. The linearizedform of the Thomas model is as follows:

ln

(C0

C− 1

)= kThq0X

Q− kThC0

QVeff (21)

The kinetic coefficientkTh and the adsorption capacity ofthe bedq0 can be determined from a plot of ln[(C0/C] − 1]againstt at a given flow rate[125].

Clark [126] defined a new simulation of breakthroughcurves. This model combines the Freundlich equation andthe mass transfer concept and has the following form:

C

C0=(

1

1 + Ae−rt

)1/n−1

(22)

with

A =(Cn−1

0

Cn−1break

− 1

)ertbreak (23)

and

R(n− 1) = r andR = kCl

U0ν (24)

Eq. (24)is the generalized logistic function wheren; Cbreak,tbreak, kCl, andν are the Freundlich constant, the outlet con-centration at breakthrough (or limit effluent concentration),the time at breakthrough, the Clark rate constant and migra-tion rate, respectively. For a particular adsorption process ona fixed bed and a chosen treatment objective, values ofA andr can be determined by usingEq. (24)by non-linear regres-sion analysis, enabling the prediction of the breakthroughcurve according to the relationship betweenC/C0 and t inEq. (24) [126].

Yoon and Nelson[127] have developed a relatively simplemodel which not only is less complicated than other models,but also requires no detailed data concerning the character-istics of adsorbate, the type of adsorbent, and the physicalproperties of adsorption bed. The Yoon and Nelson equationregarding to a single-component system is expressed as:

lnC

C0 − C= kYN t − τkYN (25)

wherekYN is the Yoon and Nelson rate constant,t is thetime required for 50% adsorbate breakthrough andτ is thebreakthrough (sampling) time. The calculation of theoreti-cal breakthrough curves for a single-component system re-quires the determination of the parameterskYN and τ forthe adsorbate of interest. These values may be determinedfrom available experimental data. The approach involves aplot of lnC/(C0 −C) versus sampling time (t) according toEq. (25). If the theoretical model accurately characterizesthe experimental data, this plot will result in a straight linewith slope ofkYN and interceptτkYN [127].

3. Biosorption of dyes

A wide variety of microorganisms including bacteria,fungi and yeasts are used for the biosorption of a broadrange of dyes. Textile dyes vary greatly in their chemistries,and therefore their interactions with microorganisms de-pend on the chemical structure of a particular dye, the spe-cific chemistry of the microbial biomass and characteristicsof the dye solution or wastewater. Depending on the dye


Table 1Data on the biosorption of dyes by various microorganisms

Biosorbent Dye Operation conditions Biosorption capacity Ref.

pH T (◦C) C0 (mg l−1) teq qeq (mg g−1)

Activated sludge Basic Red 29 – 20 500 6 h 113.2 [15]Basic Yellow 24 – 20 500 6 h 105.6Basic Blue 54 – 20 500 6 h 86.6Basic Red 18 – 20 500 6 h 133.9Basic Violet 3 – 20 500 6 h 113.6Basic Blue 4 – 20 500 6 h 157.5Basic Blue 3 – 20 500 6 h 36.5

Activated sludge Reactive Blue 2 7 25 200 1 h 102.0 [91]Reactive Yellow 2 5 25 200 1 h 119.4

Activated sludge Maxilon Red BL-N – 20 200 2 h (123.2) [94]

Aeromonassp. Reactive Blue 5 3 28 200 1 h 124.8 [83]Reactive Red 22 3 28 200 1 h 116.5Reactive Violet 2 3 28 200 1 h 114.5Reactive Yellow 2 3 28 200 1 h 124.3

Aspergillus niger Basic Blue 9 6 – 50 48 h 18.5 (1.2) [88]Acid Blue 29 4 – 50 30 h 13.8 (6.6) [89]Congo Red 6 – 50 42 h 14.7 [16]Disperse Red 1 4 – 50 48 h 5.6 [16]

Aspergillus niger Reactive Brilliant Red – – 250 2 weeks 14.2 [85]

Botrytis cinerea Reactive Blue 19 – – – – 42 (13.0) [84]Sulfur Black 1 – – – – 360 (49.7)

Candidasp. Remazol Blue 2 25 400 24 h 169 [93]Candida lipolytica Remazol Blue 2 25 400 24 h 230Candida membranaefaciens Remazol Blue 2 25 300 24 h 149Candida quilliermendii Remazol Blue 2 25 300 24 h 152Candida tropicalis Remazol Blue 2 25 400 24 h 180Candida utilis Remazol Blue 2 25 300 4 h 113

Candida rugosa Reactive Blue 19 – – – – 8 (8) [84]Reactive Black 5 – – – – 31 (31)Sulfur Black 1 – – – – 407 (308)

Cryptococcuss heveanensis Reactive Blue 19 – – – – 23 (22) [84]Reactive Black 5 – – – – 76 (60)Sulfur Black 1 – – – – 407 (360)

Dekkera bruxellensis Reactive Blue 19 – – – – 19 (36) [84]Reactive Black 5 – – – – 36 (38)Sulfur Black 1 – – – – 589 (527)

Endothiella aggregata Reactive Black 5 – – – – 44 [84]Sulfur Black 1 – – – – 307

Escherichia coli Reactive Blue 5 3 28 200 1 h 89.4 [83]Reactive Red 22 3 28 200 1 h 76.6Reactive Violet 2 3 28 200 1 h 65.5Reactive Yellow 2 3 28 200 1 h 52.4

Fomitopsis carnea Orlamar Red BG – RT 100 20 h 503.1 [14]Orlamar Blue G – RT 100 20 h 545.2Orlamar Red GTL – RT 100 20 h 643.9

Geotrichum fici Reactive Blue 19 – – – – 17 (60) [84]Reactive Black 5 – – – – 45 (7)Sulfur Black 1 – – – – 37 (60)

Kluyveromyces marxianus Remazol Black B – RT 100a 12 h 37 [87]Rem. Turquoise Blue – RT 100a 12 h 98Remazol Red – RT 100a 12 h 68Rem. Golden Yellow – RT 100a 12 h 33Cibacron Orange – RT 100a 12 h 8.5


Table 1 (Continued)

Biosorbent Dye Operation conditions Biosorption capacity Ref.


Kluyveromyces marxianus Remazol Blue 2 25 300 4 h 161 [93]Kluyveromyces waltii Reactive Blue 19 – – – – 14 (20) [84]

Reactive Black 5 – – – – 72 (60)Sulfur Black 1 – – – – 549 (445)

Laminaria digitata Reactive Brilliant Red – – 250 4 weeks 20.5 [85]

Myrothecum verrucaria Orange II – – 200 5 h 70% [80]10B (Blue) – – 200 5 h 86%RS (Red) – – 200 5 h 95%

Phanerochaete chrysosporium Congo red – – 500 2 days 90% [86]

Pichia carsonii Reactive Blue 19 – – – – 5 (3) [84]Reactive Black 5 – – – – 32 (25)Sulfur Black 1 – – – – 549 (499)

Pseudomonas luteola Reactive Blue 5 3 28 200 1 h 102.5 [83]Reactive Red 22 3 28 200 1 h 105.3Reactive Violet 2 3 28 200 1 h 96.4Reactive Yellow 2 3 28 200 1 h 102.6

Rhizopus arrhizus Humic acid – – 500 – 91.9 [82]

Rhizopus arrhizus Reactive Orange 16 2 RT 400 20 h 190 [12]Reactive Blue 19 2 RT 250 20 h 90Reactive Red 4 2 RT 350 20 h 150

Rhizopus arrhizus Remazol Black B 2 35 800 1 h 500.7 [11]Rhizopus oryzae(26668) Reactive Brilliant Red – – 250 4 weeks 102.6 [85]Rhizopus oryzae(57412) Reactive Brilliant Red – – 250 4 weeks 37.2

Rhizopus oryzae Reactive Black 5 – – – – 452 (99) [84]Sulfur Black 1 – – – – 3008 (1107)

Saccharomyces cerevisiae Remazol Blue 2 25 300 4 h 162 [93]Saccharomyces cerevisiae Reactive Blue 19 – – – – 69 (52) [84]Saccharomyces pombe Remazol Blue 2 25 300 4 h 152 [93]

StreptomycetesBW130 Anthraquinone Blue – – 280 14 days 27.0% [81]114Azo-copper Red – – 180 14 days 73.0%171

Azo-reactive Red 147 – – 150 14 days 29.0%Formazan Blue 209 – – 80 14 days 70.0%Phytalocyanine Blue 116 – – 200 14 days 39.0%

Tremella fuciformis Reactive Blue 19 – – – – 35 (41) [84]Reactive Black 5 – – – – 79 (92)Sulfur Black 1 – – – – 892 (934)

Xeromyces bisporus Reactive Blue 19 – – – – 60 (0) [84]Reactive Black 5 – – – – 1 (11)Sulfur Black 1 – – – – 60 (63)

Values in parentheses represent biosorption capacity of live biomass; %: percent removal.a Equilibrium concentration.

and the species of microorganism used different bindingcapacities have been observed (Table 1).

3.1. Microorganisms and dyes used in biosorption:biosorption mechanisms

Zhou and Banks[77] firstly reported the adsorption ofhumic acid caused organic colour in raw water by deadRhizopus arrhizusand they concluded that adsorption was

a biphasic process; the first was fast and independent ofmetabolic energy while the second was slow and dependenton metabolic energy. Based on the examinations by infraredspectra, they concluded that no chemical reaction occurredbetween cell wall and humic acid; just a physical adsorption[77].

Hu [7] demonstrated the ability of bacterial cells isolatedfrom activated sludge process of a textile industry and soilto adsorb 11 reactive dyes including Reactive Blue, Reactive


Red, Reactive Violet, Reactive Yellow and Procion Red G.The author suggested that the cell wall portion ofAeromonassp. had a higher specific adsorption capacity than the in-tact cells due to the larger surface area in the cell walls. At100 mg l−1 dye concentration the colour removal efficien-cies ranged from 12.9 to 94.3% and the maximum specificadsorption capacity ofAeromonassp. was 27.4 mg dye g−1

dried cells for Procion Red G at pH 3.0[7].Brahimi-Horn et al.[80] studied with the intact and dis-

rupted (sonicated)Myrothecum verrucariacells for the re-moval of three acid dyes in order to investigate the possiblerole of intracellular compartmentation of dye in decolour-ization. They found that both external and internal uptakewere important for the binding capacity. On the other handthey observed that the divalent dye (Acid Red) was boundto a greater extent than the monovalent dye (Acid OrangeII) [80].

Zhou and Zimmerman[81] used the actinomyceteStrep-tomycetesBW130 as an adsorbent for the decolourizationof effluents containing anthroquinone, phtlocyanine, and azodyes[81].

In another study of Zhou and Banks[82], they reportedthat chitin/chitosan was the major active component ofR.arrhizusfor humic acid adsorption[82].

In another study of Hu[83], three Gram-negative bacteria(Aeromonassp.,Pseudomonas luteolaandEscherichia coli),two Gram-positive bacteria (Bacillus subtilisandS. aureus)and activated sludge (consisting of both Gram-negative andGram-positive bacteria) were used as biosorbents for the re-moval of reactive dyes of Reactive Blue, Reactive Red, Re-active Violet and Reactive Yellow. Dead cells of test generashowed a higher uptake than the living cells due to increasedsurface area and Gram-negative bacteria had a higher ad-sorption capacity than Gram-positive bacteria due to higherlipid contents in the cell wall portion. Among these microor-ganisms, the specific adsorption capacity ofAeromonassp. for these dyes was the maximum in the range of114–146 mg g−1 at 200 mg l−1 initial dye concentration[83].

Polman and Breckenridge[84] tested 30 species of fila-mentous fungi, yeast and bacteria to remove reactive (Reac-tive Black 5, Reactive Blue 19) and sulfur (Sulfur Black 1)dyes from simulated plant waste effluents. They used bothdead and live forms of each of the species and observed thatamong 28 microbial species, 64% of the dead forms hada higher adsorption capacity for the Reactive Black 5 dyewaste; among the 21 species capable of binding ReactiveBlue 19 dye waste, 71% were more efficient dye bindersin the dead form than in the live form. They suggested thatthis might be due to an increase in the surface area for ad-sorption because of cell rupture upon death. But among the26 species capable of binding Sulfur Black 1 dye waste,54% were more efficient in the live state. They proposedthat this might be due to the chemistry of different dyes[84].

Mittal and Gupta[14] examined dead macrofungus,Fomi-topsis carneafor the sorption of three cationic dyes, Orla-

mar Red BG (ORBG), Orlamar Blue G (OBG), and OrlamarRed GTL (ORGTL)[14].

Gallagher et al.[85] investigated three types of fungi,including Laminaria digitata, Rhizopus oryzae, and As-pergillus niger, to remove Reactive Brilliant Red, a reactivedye suggesting that adsorption was occurred by combinedmechanisms onto a heterogeneous surface[85].

Tatarko and Bumpus[86] studied the biosorption ofCongo Red, a cationic azo dye on autoclavedP. chrysospo-rium, a wood rotting basidiomycete in a agitated batchsystem and observed a higher colour removal (90%)[86].

Bustard et al.[87] used the biomass derived from the ther-motolerant ethanol-producing yeast strainKluyveromycesmarxianusIMB3 for the removal of commonly used textiledyes including Remazol Black B, Remazol Turquoise Blue,Remazol Red, Remazol Golden Yellow and Cibacron Or-ange. They proposed that Cu atom in the Remazol TurquoiseBlue plays a role in the interaction between that dye and thebiosorbent regarding the maximum uptake of this dye by theyeast[87].

Aksu and Tezer[11] studied with driedR. arrhizusfor theremoval of Remazol Black B, a reactive anionic dye fromaqueous solution. They proposed that the biosorption is aresult of interaction between the active groups on the cellsurface of the fungus such as chitin, acidic polysaccharides,lipids, amino acids, and other cellular components of themicroorganism and dye anions which are typically azo-basedchromophores combined with vinyl sulfone reactive groups[11].

Fu and Virarahavan[5,16,88–90]investigated the removalof Basic Blue 9 (cationic), Acid Blue 29 (anionic), CongoRed (anionic), and Disperse Red 1 (nonionic) dyes fromaqueous solutions by biosorption on dead and pretreatedAs-pergillus niger fungus. They found thatA. niger is capa-ble of removing dyes from an aqueous solution. They ex-plored that three major functional groups: carboxyl, aminoand phosphate, and the lipid fraction in the biomass ofA.niger played an important role in the biosorption of thesedyes[5,16,88–90].

Aksu [91] investigated the biosorption of two reactivedyes (Reactive Blue 2 and Reactive Yellow 2) onto driedactivated sludge consisting of mainly both bacteria and pro-tozoa. They suggested that activated sludge has an extensiveuptake capacity for organic pollutants due to acidic polysac-charides, lipids, amino acids and other cellular componentsavailable on the cell wall of bacteria[91].

O’Mahony et al. [12] indicated that the ability ofoven-dried R. arrhizus biomass for the biosorption ofthree commonly used reactive dyes, Cibacron Brilliant Red3B-A (Reactive Red), Remazol Brilliant Blue R (Reac-tive Blue 19), Remazol Brilliant Orange3WR (ReactiveOrange 16) from aqueous solutions in a batch system[12].

Chu and Chen[15,92] studied with the oven-driedactivated sludge biomass in the particle size range of105–297�m as an biosorbent for the removal of selected


anionic dyes such as Direct Orange 39 and Direct Red 83,or for selected non-ionic dyes, such as Disperse Violet 8and Disperse Yellow 54 or for cationic (basic) dyes such asBasic Blue 3 (B-3), Basic Violet 3 (V-3), Basic Yellow 24(Y-24), Basic Red 18 (R-18), Basic Red 29 (R-29), BasicBlue 47 (B-47), Basic Blue 54 (B-54) from wastewater.They observed that the biomass had no affinity for anionicand non-ionic dyes. However, the biomass could integratewith the cationic dyes. Maximum adsorption capacity ofthese basic dyes occurred in the order B-47> R-18 >

V-3 > R-29> Y-24 > B-54 � B-3, They concluded thatthe chemical structure (e.g. molecular structure of colouringgroups, such as anthraquinone, monoazo, oxazine, thiazoleazo, or triarylmethane and the type and number of the posi-tion of the substituents in the dye molecule), basicity (e.g.element of chromophore group, insulated or conjugatedtype), and molecular weight of basic dye molecules havean influence on the adsorption capacity of activated sludgebiomass[15,92].

Aksu and Dönmez[93] reported the biosorption capac-ities and rates of nine yeast species (Saccharomyces cere-visiae, Schizosaccharomyces pombe, K. marxianus, Candidasp.,Candida tropicalis, Candida lipolytica, Candida utilis,Candida quilliermendii, and Candida membranaefaciens)for Remazol Blue reactive dye from aqueous solutions. Theyeasts studied were found to be more effective for concen-trating Remazol Blue dye at different capacities accordingto the dye concentration. They explained the differences be-tween yeast species for dye binding capacity in terms ofthe properties of the yeast (e.g., structure, functional groups,surface area and morphological differences depending onthe yeast division, genera and species). They proposed thatcell walls of yeasts contain polysaccharides as basic build-ing blocks which have ion exchange properties, and alsoproteins and lipids and therefore offer a host of functionalgroups capable of binding dye molecules. These functionalgroups such as amino, carboxylic, sulfydryl, phosphate andthiol groups, differ in their affinity and specificity for dyebinding [93].

Basibuyuk and Forster[94] studied the biosorption ofone acid dye (Acid Yellow 17) and one basic dye (Max-ilon Red BL-N) onto live activated sludge. They chose liveactivated sludge in order to understand the adsorption prop-erties of dye-activated sludge systems since sorption is theprimary mechanism leading to biodegradation in activatedsludge processes commonly used for the treatment of textilewastewaters. The results showed that binding of Acid Yel-low 17 where the colouring group is anionic onto activatedsludge was not promising while Maxilon Red BL-N wherethe colouring group is cationic was adsorbed well by acti-vated sludge. They explained the main reasons for the pooradsorption capacity of acid dye as the negative electricalcharge of activated sludge under normal pH conditions sorepulsion between negatively charged sorbate ions and neg-atively charged sorbent surface, and the number of sulphogroups in acid dye reducing dye adsorption[94].

3.2. Pretreatment of microorganism

Researches have shown that some physical or chemicalpretreatment processes can increase the adsorption capacityof biomass. These pretreatment methods mainly includeddrying, autoclaving, contacting with organic chemicals, suchas formaldehyde, or inorganic chemicals, such as NaOH,H2SO4, NaHCO3, and CaCl2.

Zhou and Banks[77] reported thatR. arrhizuspretreatedwith 10% formaldehyde or autoclaving significantly in-creased the humic acid adsorption capacity compared withliving cells, due to the exposure of latent binding sites[77].

In another study of Zhou and Banks[82], they usedR. ar-rhizusbiomass pretreated with 2 M NaOH for 1 h and theyfound an increased biosorptive capacity. They also observedthat the longer duration of treatment caused further enhance-ment of the biosorption capacity. They suggested that NaOHtreatment could remove proteins and glucans from the cellwall thereby increasing the percentage of chitin/chitosan inthe whole cell fraction. Some chitin may also be transformedto chitosan with concentrated alkaline solution over a longperiod. They suggested chitosan could be the most efficientsequester of humic acid molecules[82].

Hu [83] used live and autoclaved Gram-negative bacteriafor the removal of reactive dyes and indicated that the au-toclaved cells had a higher uptake capacity than living cellsdue to increasing in surface area caused by cell rupture dur-ing autoclaving[83].

Gallagher et al.[85] observed that all methods; includingautoclaving, calcium saturation, NaOH and chitin/chitosanenrichment, they used for the pretreatment ofR. oryzaein-creased the biosorption capacity from 7 to 15%. After mea-suring the porosity and surface area of the biomass ofR.oryza, they explained the increase in biosorption capacityby autoclaving process due to the disruption of the particlestructure. The disruption may cause an increase of surfacearea and monolayer volume and an increase in porosity ofthe particles and thus expose latent sites, consequently in-creasing the dye adsorption. An increase of adsorption ca-pacity by Ca2+ saturation was demonstrated as a fact thatR. oryzaehad a low affinity for Ca2+ ions, which made cal-cium a good activating counter ion which was easy to bereplaced by dyes that formed more stable complexes. Ac-cording to authors, pretreatment by NaOH could generateanionic sites without significant modification of the cell wallstructure and also expose the chitin/chitosan complex of thecell by dissolving certain biopolymers from the surface ofbiomass particles, since chitin/chitosan was suggested as thepredominant biosorbent of the dye[85].

Tatarko and Bumpus[86] used both living and autoclavedcultures ofP. chrysosporiumto decolourize Congo Red andobserved that the autoclaved cells had a higher colour re-moval (90%) than the living cells (70%)[85].

Fu and Viraraghavan[88,89] used some pretreatmentmethods such as autoclaving and contacting with nor-mal chemicals, including 0.1 M NaOH, 0.1 M HCl, 0.1 M


H2SO4, 0.1 M CaCl2, 0.1 M NaHCO3, 0.1 M Na2CO3 and0.1 M NaCl to pretreat the living fungal biomassA. niger.They reported that the effective pretreatment was differ-ent with each of the dyes, Basic Blue 9 and Acid Blue29. Autoclaving increased the biosorption capacity from1.2 mg of Basic Blue 9 per gram of living fungal biomassto 18.5 mg g−1 of autoclaved biomass, while 0.1 M H2SO4pretreatment enhanced the biosorption capacity from 6.6 mgof Acid Blue 29 per gram of living biomass to 13.8 mg g−1

of dead biomass. They suggested that autoclaving coulddisrupt the fungal structure and expose the potential bindingsites for Basic Blue 9, a cationic dye, biosorption, whileH2SO4 pretreatment could change the negatively chargedsurface of the fungal biomass to positively charged and thusincrease the attraction between fungal biomass and AcidBlue 29, an anionic dye[88,89].

Fu and Viraraghavan[90] also studied the removalof Congo Red, an anionic dye by pretreated fungusA. niger. They classified the chemicals used for pre-treatment as acids, alkalis and salts. They used au-toclaved (for 30 min at 121◦C and 18 p.s.i.), 0.1 MNaOH, 0.1 M HCl, 0.1 M H2SO4, 0.1 M CaCl2, 0.1 MNaHCO3, 0.1 M Na2CO3, and 0.1 M NaCl treatedA.niger biomass. They observed that while HCl, H2SO4,CaCl2, NaHCO3, and NaCl pretreatments all increasedbiosorption capacity, NaOH and Na2CO3 pretreatmentsdecreased biosorption capacity. They proposed that fungalbiomass is usually charged negatively on its surface. CongoRed ionizes to give a coloured anion in solution. Thereforethe coloured anions of Congo Red will be repulsed by theanionic groups on the surface of fungal biomass. As auto-claving could cause the disruption o the fungal structureresulting in an increase in porosity[85], more Congo Redions could enter the expanded pores in autoclaved fungalbiomass leading to increased biosorption capacity. Theyexplained the reason why CaCl2 pretreatment increasedbiosorption capacity could be that Ca2+ is divalent andthus could neutralize the negative charge on the surface offungal biomass and change part of the negatively chargedsurface to positively charged. NaOH pretreatment decreasedbiosorption capacity of Congo Red. This is because pretreat-ment by NaOH could generate anionic sites on the surfaceof fungal biomass[85] and thus increase repulsion betweenthe negatively charged surface of the fungal biomass and thecoloured anions of Congo Red. Pretreatment with NaHCO3was found to be the most effective with a biosorption ca-pacity of 14.7 mg g−1 compared with 12.1 mg g−1 of livingbiomass for Congo Red. This could be because bicarbonateion, HCO3

− can either provide protons or accept protons inwater. The protons could neutralize negative charges on thesurface of fungal biomass and change the part of the neg-atively charged surface to positively charged. Meanwhile,they explored that changes in charge density could alsoaffect adsorption affinity for particular dyes. For Acid Blue29 they found the extent of increase in biosorption capacityby NaHCO3 pretreatment was lower than that by HCl and

H2SO4 pretreatment. They attributed the difference to thedifferent molecular structures of the two dyes. Therefore,they suggested that the effective pretreatment is related todye molecules, which is specific for each dye[90].

3.3. Effect of pH on dye biosorption

Since pH is the most important parameter affecting notonly the biosorption capacity, but also the colour of the dyesolution and the solubility of some dyes, various researchershave investigated the effect of pH on colour removal.

Hu [7] demonstrated that the optimal pH for biosorptionof 11 reactive dyes byAeromonassp. cells was at acidicrange. He found that the removal decreased as the pH of thedye solution increased from 3.0 to 11.0. He suggested thatat a lower pH, the association of dye anions with positivelycharged bacterial cell surfaces at acidic pH takes place[7].

Zhou and Banks[82] reported that the biosorption of hu-mic acid byR. arrhizusincreased with decreasing pH. Theysuggested that at lower pH, more of humic acid functionalgroups were uncharged and humic acid had a lower solubil-ity and was thus more adsorbable. On the other hand, thelower pH results in high concentrations of protons whichneutralize the negative charge on bothR. arrhizusand hu-mic acid, leading to increased adsorption[82].

On the other hand Mittal and Gupta[14] studied the ef-fect of pH on the biosorption of three cationic dyes, OrlamarRed BG, Orlamar Blue G and Orlamar Red GTL by deadfungus ofF. carneaand their results showed that colour re-moval decreased with decreasing pH due to repulsive forcesbetween coloured dye cations in solution and biosorbent sur-face charged positively at pH values lower than 3.0[14].

Hu [83] investigated the effect of pH on the removal of sixreactive dyes by three Gram-negative bacteria (P. luteola, E.coli, andAeromonassp.) and reported that the biosorption ofall dyes by all cells increased significantly with decreasingpH. They explained this situation as a fact that dye anionsare electrostatically bonded to positively charged bacterialcell surfaces at low pH[83].

Fu and Viraraghavan[88,89] reported that initial pH ofdye solution significantly influenced the chemistry of bothAcid Blue 29 and Basic Blue 9 molecules and deadA. nigerfungus in an aqueous solution. The effective initial pH of dyesolution was 6.0 and 4.0, respectively, for Basic Blue 9 andAcid Blue 29. At pH of 2.0, no biosorption occurred for Ba-sic Blue 9 due to the high concentration of protons, while atpH of 12, no biosorption occurred for Acid Blue 29[88,89].The same authors[16] also investigated the effect of pHon Congo Red biosorption by NaHCO3 pretreatedA. nigerbiomass in an aqueous solution and they determined the ef-fective pH as 6.0. They explained the biosorption mecha-nism due to pH as follows: the surface of NaHCO3 pre-treated fungal biomass could be partially positively charged,so it might have some negatively charged adsorption sites.So the presence of high concentrations of protons at a lowerpH could neutralize the negative charge on the surface of


NaHCO3 pretreated biomass, leading to increased biosorp-tion capacity. Therefore, a pH of 6.0 was the effective pHfor biosorption of Congo Red. At a high pH, a high con-centration of OH− could neutralize the positively chargedsurface of NaHCO3 pretreated biomass and might form anegatively charged surface again. Thus, it would increasethe repulsion between the coloured anions of Congo Redand the negatively charged fungal biomass and cause a de-crease in biosorption capacity. Meanwhile, there would becompetition between OH− (at high pH) and the colouredanions of Congo Red for the positively charged adsorp-tion sites, which could also decrease biosorption capacity[16].

Aksu and Tezer[11] also examined the effect of initialpH on fungal (driedR. arrhizus) binding of reactive dyeRemazol Black B and they found maximum uptake at pH2.0. They explained the higher uptakes at lower pH valuesby electrostatic attractions between negatively charged dyeanions and positively charged cell surface[11].

The results obtained by O’Mahony et al.[12] also showedthat the maximum biosorption of three commonly used re-active dyes, Reactive Red, Reactive Blue 19, and ReactiveOrange 16 from aqueous solutions by oven-driedR. ar-rhizus biomass was performed at pH 2.0. They explainedthe variation in uptake capacity of theRhizopusbiomassacross the pH range in terms of its effective isoelectricpoint. At pH values below the isoelectric point (<4.0), thebiomass will have a net positive charge. It is expected thatnitrogen-containing functional groups such as amines orimadazoles in the biomass will also be protonated at acidicpH values. These charged sites become available for elec-trostatic binding of anionic reactive dyes[12].

In a study performed by Aksu and Dönmez[93], optimumbiosorption pH was also determined as 2.0 for the removal ofRemazol Blue reactive dye by nine yeast species.C. lipoly-tica showed the maximum biosorption capacity at pH 2.0binding 173.1 mg dye g−1 dry biomass. They explained theenhancement of uptake of reactive dyes at acidic pH in termsof electrostatic interactions between the biomass and thedye particles. Upon dissolution, ionic dyes release coloureddye ions into solution. The adsorption of these charged dyegroups onto the adsorbent surface is primarily influenced bythe surface charge that in turn is influenced by the solutionpH. With diminishing pH increasing numbers of weak basegroups in the biomass become protonated and acquire a netpositive charge. These charged sites become available forbinding anionic groups such as the reactive dye used in thisstudy[93].

3.4. Effect of temperature on dye biosorption

As various textile dye effluents are discharged at rela-tively high temperatures (50–60◦C), so temperature willbe an important design parameter affecting the biosorptioncapacity in the real application of biosorption by biomassin future [4,5].

Zhou and Banks[82] investigated the effect of tem-perature on humic acid biosorption byR. arrhizus. Theyobserved that low temperature (from 36 to 16◦C) caused ahigh biosorption. They suggested that biosorption betweenR. arrhizusand humic acid was an exothermic process andthe mechanism was mainly physical adsorption, dominantat lower temperatures[82].

Hu [83] studied the effect of temperature on the removalof six reactive dyes by three Gram-negative bacteria (P.luteola, E. coli, andAeromonassp.) and reported that tem-perature had slight or no effect on the equilibrium uptakesuggesting the feasibility of directly using dead biomass inthe dyeing wastewater to absorb dyes without decreasingthe temperature of wastewater[83].

Gallagher et al.[85] also decided that biosorption of Re-active Brilliant Red byR. oryzaewas a physical adsorptiondue to increase of biosorption capacity with decreasing tem-perature[85].

Aksu and Tezer[11] also investigated the effect of temper-ature on the biosorption of Remazol Black B reactive dye byR. arrhizusand their results indicated that optimum adsorp-tion temperature was 35◦C and adsorption decreased withfurther increasing temperature due to the decreased surfaceactivity [11].

Chu and Chen[15,92] studied the effect of temperatureon biosorption of Basic Violet 3 and Basic Yellow 24 dyesusing dried activated sludge biomass and they observed thatadsorption capacity decreased from 113.6 to 109.8 mg g−1

for Basic Violet 3 and decreased from 57.0 to 51.3 mg g−1

for Basic Yellow 24 with increasing temperature from 20to 40◦C. These results indicated that both biosorption pro-cesses are exothermic in nature. They determined the activa-tion energy as 3.27 kcal mol−1 for Basic Violet 3 (coloranttriarylmethane) biosorption and 1.45 kcal mol−1 for BasicYellow 24 (colorant triarylmethane) biosorption showingfast and intraparticular diffusion limited adsorptions[15,92].

3.5. Effect of initial dye concentration on dye biosorption

Dye concentration also affects the efficiency of colourremoval. Initial concentration provides an important drivingforce to overcome all mass transfer resistances of the dyebetween the aqueous and solid phases. Hence a higher initialconcentration of dye may enhance the adsorption process.

Bustard et al.[87] observed that although the uptake ofRemazol Golden Yellow dye byK. marxianusIMB3 waslower at lower dye concentrations, the biosorptive capac-ity increased significantly at higher concentrations of dye.These results suggested some form of cooperativity with re-spect to interactions between the dye and the biomass. On theother hand for Cibacron Orange dye biosorption by the samebiomass, they found that the biosorptive capacity increasedto a maximum of 8.5 mg g−1 at a residual dye concentrationof 100 mg l−1 and then decreased rapidly as the equilibriumconcentration increased. They suggested one possible rea-son for this observation that, as the concentration of dye


increases above 100 mg l−1, dye–dye interactions becomeprevalent and these interactions may result in decreasedaffinity of the dye binding sites on the biomass[87].

The results obtained by Aksu and Tezer[11] showed thatthe equilibrium sorption capacity of driedR. arrhizusin-creased with increasing initial Remazol Black B concentra-tion up to 800 mg l−1 while the adsorption yield of dye indi-cated the opposite trend. At 35◦C, they found that when theinitial Remazol Black B dye concentration increased from20.5 to 802.4 mg l−1, the loading capacity of biomass in-creased from 19.3 to 500.7 mg g−1 and the adsorption yieldof biomass decreased from 94.0 to 62.4%[11].

Aksu [89] reported that the equilibrium sorption capacityof dried activated sludge increased with increasing initialdye concentration up to 200 mg l−1 for both Reactive Blueand Reactive Yellow dyes[89].

O’Mahony et al.[12] reported that uptake of each of theReactive Red, Rective Blue 19 and Reactive Orange 16 dyesby oven-driedR. arrhizusincreased with increasing dye con-centration[12].

Chu and Chen[92] investigated the effect of dye concen-tration on adsorption of Basic Yellow 24 using dried acti-vated sludge biomass. Uptake of the dye increased from 18to 90 mg g−1with increasing dye concentration from 50 to300 mg l−1 [92].

Dönmez and Aksu[93] also investigated the effect of ini-tial Remazol Blue concentration on the dye sorption capac-ity and yield of each of nine yeast species between 100 and400 mg l−1 at the initial pH value of 2.0. They found differ-ent binding capacities and yields depending on the speciesand initial dye concentration. However, all the yeast specieswere capable of removing more than 90% of the colouringmaterial at 100 mg l−1 of initial dye concentration[93].

3.6. Effect of salts on dye biosorption

Dyeing processes consume large amounts of salts. So saltconcentration in dye wastewater (or the ionic strength ofsolution) is one of the important factors which influence thebiosorption capacity.

Zhou and Banks[77,82] reported that high ionic strength(high concentration of NaCl) led to high biosorption of hu-mic acid byR. arrhizus. They proposed that the effect ofionic strength was similar to that of a colloid. At higherionic strength, the electrical double layers of bothR. ar-rhizusbiomass and humic acid would be compressed thinner.Therefore, biomass and humic acid could approach closerand thus this would increase van der Waals bonding andhence increase biosorption[77,82].

3.7. Effect of heavy metal ions on dye biosorption

Textile wastewaters may include metal ions beside dyesand salts due to metal-containing dyes used in textile in-dustry. Metal ions would be a factor influencing biosorptionrate and capacity. They might compete with dye molecules

for the binding sites or stimulate the biosorption of dye ontobiomass.

Zhou and Banks[77,82]studied the effect of Cd2+, Cu2+,and Al3+ ions on humic acid adsorption byR. arrhizus. Theyobserved that high concentrations of Cd2+, Cu2+, and Al3+resulted in high biosorption. They suggested that metal ioncould be a bridge betweenR. arrhizusand humic acid, whichwere both negatively charged. So the addition of metal ionswould neutralize their surface charge and thus reduce therepulsive forces between them, leading to their closer con-tact and increase bonding. Metal ions with di- and tri-valentcations could interact with humic acid to form precipitatesor aggregates and thus reduce humic acid solubility and in-crease its biosorption potential[77,82].

O’Mahony et al.[12] investigated the effect of Cd2+ ionson the uptake of each of the Reactive Red, Reactive Blue19, and Reactive Orange 16 dyes by oven-driedR. arrhizus.They observed uptake of each of the dye was diminished bythe presence of 100 mg l−1 Cd2+ ions due to competitionbetween Cd2+ and dye molecules. For Cibacron Red, themaximum reduction was of the order of 20 mg g−1 biosor-bent which represents 12.5% of maximum dye adsorptionlevels. They also found similar reductions in uptake of Re-mazol Blue and Remazol Orange dyes. They explored thatthe presence of high levels of Cd2+ did not significantly de-crease the adsorption capacity of the biomass[12].

3.8. Effect of other dyes on dye biosorption(multicomponent dye biosorption)

To study the biosorption of a dye or dyes from a multi-component dye solution which simulate dyehouse or textilemill effluents is important for design. While the knowledgeof general uptake of single species of dyes by microorgan-isms is increasing, relatively little is known about the com-bined effects of two or more dyes and simultaneous removalof dyes from a mixture of dye solution.

O’Mahony et al.[12] studied with multicomponent dyesolutions containing equal concentrations of the ReactiveRed, Reactive Blue 19, and Reactive Orange 16 dyes to amaximum total dye concentration of 450 mg l−1. They ob-served that uptake of each of the dyes from multicomponentsolution at pH 2.0 byRhizopusbiomass increased with in-creasing solution concentration suggesting a direct compe-tition mechanism and no preferentially dye binding[12].

3.9. Effect of surfactants on dye biosorption

In the dyeing process, surfactants are occasionally usedand thus may be present in dye wastewaters.

Brahimi-Horn et al.[80] observed that the presence ofdetergent in wastewaters may reduce the binding efficiencyof the cells and reported that high concentration of Tween, anonionic surfactant, results in a low adsorption and differentdyes show different effects with the same concentration ofTween. The effect of Tween diminished with time[80].


3.10. Regeneration of biosorbent

One of the important characteristics of a biosorbent iswhether it can be regenerated. Research has shown thatbiomass can be eluted and regenerated by some organic sol-vents such as methanol, ethanol, and some surfactants suchas nonionic Tween, as well as NaOH solution.

Brahimi-Horn et al.[80] used methanol to desorb dyesfrom boundM. verrucaria cells and recovered 39, 35, and53% of Orange II, 10B (blue) and RS (red) dye, respec-tively. They observed that methanol-treated cells still pos-sessed to a large extent the capacity to absorb dyes. Theysuggested that methanol might influence the hydropho-bic/hydrophilic interaction between the dyes and biomass[80].

Zhou and Banks[82] used 0.1 M NaOH to desorb the hu-mic acid fromR. arrhizusbiomass because humic acid sol-ubility increased in water at higher pH and thus humic acidwas eluted from fungal biomass. The average desorption ef-ficiency was above 90%. The results showed thatR. arrhizusbiosorbents could be used for several sorption–desorptioncycles with similar efficiency[82].

Polman and Breckenridge[84] studied the recovery ofReactive Black 5 fromR. oryzaeusing ethanol and Tween 80and the recoveries were 38.4 and 6.5% at the concentrationsof ethanol 60% (v/v in water) and Tween 80 two percent,respectively[84].

3.11. Effect of shaking rate on dye biosorption

Providing an adequate stirring rate in a batch biosorptionprocess is important to overcome external mass transfer re-sistances so the effect of stirring rate on biosorption shouldbe investigated.

Chu and Chen[92] investigated the effect of shaking rateon the biosorption of Basic Yellow 24 using dried activatedsludge biomass with 300–600�m selected range of parti-cle size. They observed that uptake capacity of biomass in-creased from 18 to 53 mg g−1 with increasing shaking ratefrom 40 to160 rpm. The results showed that there is a bound-ary layer surrounding the biomass particles and a decreasein its effect with increasing shaking rate[92].

3.12. Effect of particle size on dye biosorption

Biosorption kinetics is related with surface area of biosor-bent directly so particle size is also one of the importantfactors which affect the biosorption capacity.

Chu and Chen[92] studied the effect of biomass parti-cle size on the biosorption of Basic Yellow 24 using driedactivated sludge biomass with 75–150, 150–300, 300–600,600–1180�m selected ranges of particle size. They observedthat biosorption capacity of biomass increased with decreas-ing particle size. This situation was explained by larger to-tal surface area of smaller particles for the same amount ofbiomass[92].

3.13. Equilibrium modeling of biosorption

Hu [7] observed that the adsorption isotherm BG13 bluereactive dye byAeromonassp. followed the Freundlichmodel (Eq. (2)) with the values ofKF and 1/n of 0.3769and 1.255, respectively[7].

Zhou and Banks[82] reported that humic acid adsorptiononR. arrhizusobeyed the Freundlich isotherm model whichsuggested that biosorption occurred on the heterogeneoussurface[82].

Hu [83] reported that the biosorption equilibrium of sixreactive dyes by dead cells ofAeromonassp.,P. luteolaandE. coli fitted to Freundlich model[83].

Gallagher et al.[85] usedR. oryzaebiomass to adsorbReactive Brilliant Red in solution and observed that bothFreundlich and Langmuir (Eq. (1)) isotherm models fittedbiosorption well, which indicated adsorption by combinedmechanisms onto a heterogeneous surface[85].

Bustard et al.[87] reported that the biosorption of Rema-zol Black B, Remazol Turquoise Blue, Remazol Red ontoK. marxianusIMB3 fitted to Langmuir model while Rema-zol Golden Yellow and Cibacron Orange failed to adherethis model[87].

Aksu and Tezer[11] used the Freundlich and Langmuiradsorption models for the mathematical description of thebiosorption equilibrium of Remazol Black B on driedR.arrhizus and evaluated the isotherm constants at differenttemperatures. Equilibrium data fitted very well to the Fre-undlich model in the studied concentration (20–800 mg l−1)and temperature (25–55◦C) ranges[11].

The Freundlich and Langmuir models were also appliedto the biosorption equilibrium data of Reactive Blue 2 andReactive Yellow 2 dyes onto dried activated sludge by Aksu[91] and both the models were found suitable for describ-ing equilibrium data of both the dyes. The Langmuir andFreundlich constants were used to compare the biosorptivecapacity of the dried biomass for both the dyes[91].

Fu and Virarahavan[88] reported that at initial pH 4,biosorption of Basic Blue 9 byA. nigerfitted the Langmuirequation well; at initial pH 10, the Langmuir and Freundlichisotherm models both fitted biosorption well[88].

The isotherm studies for Acid Blue 29 biosorption byA.niger fungus conducted by Fu and Virarahavan[89] showedthat the Langmuir, Freundlich, and BET (Eq. (6)) isothermmodels all fitted well with the experimental data[89].

Isotherm studies performed by Fu and Virarahavan[16,90] indicated that among the Langmuir, Freundlichand BET isotherm models, none was found suitable, butthe Radke-Prausnitz model (Eq. (5)) was able to describethe biosorption equilibrium of Congo Red on NaHCO3pretreatedA niger [16,90].

Aksu and Dönmez[93] reported that both the Freundlichand Langmuir adsorption models were found suitable fordescribing the biosorption of the Remazol Blue reactive dyeby all the Candidayeasts (exceptC. membranaefaciens).According to Langmuir constants,C. lipolytica exhibited


the highest dye uptake capacity (Q0 = 250 mg g−1) amongthese species[93].

Basibuyuk and Forster[94] applied the Langmuir modelto the biosorption data of basic dye Maxilon Red BL-N bylive acivated sludge and they noted that Langmuir equationfitted the data very well[94].

3.14. Biosorption kinetics

Hu [83] reported that the time required to reach equilib-rium for the biosorption of reactive dyes to dead cells ofAeromonassp.,P. luteola, andE. coli appeared to be veryshort and was roughly within 1 h[83].

Mittal and Gupta [14] described the biosorption ofcationic dyes byF. carnea in the batch adsorber byfirst-order reaction kinetics defined inEq. (16) [14].

Aksu and Tezer[11] applied the pseudo first- andsecond-order kinetic models (Eqs. (16) and (18)) to theexperimental kinetic data of Remazol Black B biosorptionon driedR. arrhizusassuming that the external mass trans-fer limitations in the system can be neglected. The resultsindicated that the dye uptake process followed the pseudosecond-order rate expression and adsorption rate constantsincreased with increasing temperature up to 35◦C anddecreased with increasing concentration[11].

The kinetic studies performed by Fu and Virarahavan[16,88–90]for the biosorption of Acid Blue 29, Basic Blue9 and Congo red by treatedA. niger, indicated that equi-librium was reached in 30, 48, and 42 h, respectively, de-pendent on initial pH of the dye solution. These equilib-rium times were shorter than the time observed by Zhouand Banks[82] (3 days) and Gallagher et al.[85] (4 weeks).They also reported that the Lagergren first-order and pseudosecond-order rate equations were able to provide a realisticdescription of biosorption kinetics for the biosorption of alldyes by treatedA. niger [16,88–90].

Aksu [91] reported that the initial sorption of ReactiveYellow 2 occurred more rapidly by dried activated sludgethan that of Reactive Blue 2. Equilibrium was establishedin 15–30 min for Reactive Yellow 2 and in 30–60 min forReactive Blue 2 at all dye concentrations studied. The re-sults showed that the uptake process of each of the dye isnot Lagergren first-order reaction and that the second-ordermodel, based on the assumption that the rate limiting stepmay be chemical biosorption[91].

Work carried out by Chu and Chen[15] indicated thatequilibrium was established in 6 h for the biosorption of ba-sic dyes by dead activated sludge. They observed that thekinetics of initial adsorption stage of basic dyes by biomassis a first-order process and is controlled by film diffusionaccording toEq. (13) [15]. Chu and Chen[15] also stud-ied the rate processes for the adsorption of Basic Yellow 24dye on dried activated sludge as a function of shaking rate,initial dye concentration, biomass particle size, and dye so-lution temperature. The experimental results indicated thatthere is a boundary layer surrounding the biomass parti-

cles, the kinetics of the adsorption process is mainly con-trolled by intraparticle diffusion due to Weber-Morris theory(Eq. (14)).

Aksu and Dönmez[93] studied the biosorption kineticsof nine yeasts for Remazol Blue reactive dye removal at200 mg l−1 initial dye concentration for the first 240 minof biosorption. The results showed that the biosorption be-haviour and biosorption capacities of all the yeasts were dif-ferent from each other. Initial sorption of Remazol Blue dyeby S. cerevisiae, S. pombe, K. marxianus, andC. utilis yeastcells occurred more rapidly than by otherCandidayeasts.A larger amount of dye was removed by these dried cellsin the first 15 min of contact (69, 68, 58 and 57%, respec-tively). An equilibrium was established in 240 min for thesebiosorbents for all the initial dye concentrations studied andequilibrium did not change subsequently up to 72 h. Theysuggested that for these yeast cells the uptake of dye occurspredominantly by surface binding and that available sites onthe biosorbent are the limiting factor for the biosorption. Incontrast with theCandidayeasts, they proposed that the in-tracellular uptake by these cells appears to be insignificant[93].

The data obtained by Basibuyuk and Forster[94] showedthat a contact time of 2 h was sufficient to achieve equi-librium for the Maxilon Red adsorption by live activatedsludge. The results indicated that the initial part of the Max-ilon Red adsorption followed a first-order process defined byEq. (13)controlled by film diffusion. Although intraparticlediffusion played a significant role, it was not the main ratedetermining step throughout the adsorption. A comparisonof the kinetic models on the overall adsorption rate showedthat the Maxilon Red and live activated sludge system wasbest described by the pseudo second order kinetic model[94].

3.15. Column studies

Banks and Parkinson[79] packed the autoclavedR. ar-rhizusmycelia in a sorption column to remove the organiccolour in raw water attributed to humic acid. They reportedthat active sites for humic acid adsorption on fungal biomassin R. arrhizuswere on the fungal cell wall and were mostprobably the chitin/chitosan components[79].

4. Biosorption of phenols and phenolic compounds

In recent years, a number of studies have focused on somemicroorganisms including bacteria and fungi which are ableto biosorb phenols and chloro- and nitro-phenols. Dependingon the phenolic compound and the species of microorganismused also different binding capacities have been determined(Table 2). The researchers have selected the pH, initial pol-lutant and biomass concentrations and pretreatment methodas important parameters affecting the removal efficiency ofphenolics.


Table 2Data on the biosorption of phenol and phenolic compounds by various microorganisms

Biosorbent Phenolic compound Operation conditions Biosorption capacity Ref.


Activated sludge PCP – 20 1a 3 days 3.3 [95]

Activated sludge Phenol 1.0 25 100 50 min 86.1 [101]o-Chlorophenol 1.0 25 100 350 min 102.4p-Chlorophenol 1.0 25 100 350 min 116.3

Activated sludge PCP 7.0 25 0.5 2 h 2.56 [103]Activated sludge Phenol 1.0 25 500 – 166.6 [104]Anaerobic sludge Phenol 1.0 25 500 – 245.0 [105]

Anaerobic granular sludge 2-CP 7.5 35 1 2 h 0.0016 [43]3-CP 7.5 35 1 2 h 0.02034-CP 7.5 35 1 2 h 0.02192,3-DCP 7.5 35 1 2 h 0.02162,4-DCP 7.5 35 1 2 h 0.06382,5-DCP 7.5 35 1 2 h 0.02912,6-DCP 7.5 35 1 2 h 0.00663,4-DCP 7.5 35 1 2 h 0.05953,5-DCP 7.5 35 1 2 h 0.07262,3,4-TCP 7.5 35 1 2 h 0.04892,3,5-TCP 7.5 35 1 2 h 0.02712,3,6-TCP 7.5 35 1 2 h 0.01322,4,5-TCP 7.5 35 1 2 h 0.03242,4,6-TCP 7.5 35 1 2 h 0.0189PCP 7.5 35 1 2 h 0.2704

Anaerobic granular sludge 2-NP 7.5 29 90 4 h 1.43 [107]4-NP 7.5 29 90 4 h 1.512,4-DNP 7.5 29 90 4 h 1.87

Aspergillus niger Phenol 5.1 21 1 24 h 0.5 [41]

Emericella nidulans 2,4-DCP 6 20 163 3 h 9.1 [102]4-CP 6 20 128.6 3 h 3.0

Kluvera cryocrescens 4-CP 6.1 25 – 2 h 72.3 [106]Mycobacterium chlorophenolicumPCP-1 PCP 7 30 50 1.5 min 23.0 [99]Rhizopus arrhizus PCP – 20 1a 3 days 14.9 [95]

Papermill sludge Phenol 6 20 800 260 h 0.4 [40]2-CP 6 20 800 3 h 0.343-CP 6 20 800 3 h 1.04-CP 6 20 800 3 h 1.02-NP 6 20 800 3 h 0.124-NP 6 20 800 3 h 0.312,4-DCP 6 20 800 3 h 2.73,4-DCP 6 20 800 3 h 5.03,5-DCP 6 20 800 3 h 3.02,4,5-TCP 6 20 800 3 h 2.7

a Equilibrium concentration.

4.1. Microorganisms and phenolics used in biosorption:biosorption mechanisms

Bell and Tsezos[95] reported on the biosorption of pen-tachlorophenol (PCP) onto two types of inactive microbialbiomass; a mixed culture of aerobic activated sludge and apure culture ofR. arrhizus. Based on the examination of theuptake by cell walls, they concluded that biosorption pro-cess involves uptake by both the cell walls and other cellularcomponents of the microorganisms[95]. They obtained thatdead cells ofR. arrhizushad a higher biosorption capacity

for PCP than that of dead activated sludge. Tsezos and Bell[96] also studied the biosorption of PCP by live cells of ac-tivated sludge andR. arrhizus. They found the biosorptionof PCP to be nonlinear and correlated with the octane/waterpartition coefficient (Kow, defined as the ratio of concentra-tions in water immisciblen-octanol phase to one in aqueousphase) on both type of biosorbents within the concentra-tion range (about 10 mg l−1) studied. They related the up-take of PCP to the hydrophobicity of the target pollutantaccording to the hydrophobic partitioning model giving thecorrelation between octanol–water partition coefficient and


the solid phase concentration (qeq) within the concentrationrange studied[96].

Kennedy et al.[43] studied live anaerobic granular sludgefor the sorption of a number of chlorophenols (phenol,most isomers of tri-, di-, and monochlorophenols and pen-tachlorophenol), observing that the component and contentof cellular lipids would influence the adsorption capac-ity. They found no obvious relationship between sorptionand numbers or positions of chlorine substituents. The re-sults showed that PCP was more strongly sorbed than thelesser-chlorinated phenols. Although the correlation coef-ficent (<0.6) indicated thatKow only partly described thesorption of chlorophenols to granular sludge, in general,the uptake capacity of live anaerobic biomass tended toincrease with increasing octanol–water partition coefficientwhich is an indicator of hydrophobicity. They found that thePCP uptake capacity of live anaerobic sludge was much less(20- and 10-fold) than that of live or dead aerobic sludgeobtained by Bell and Tsezos[95] and Tsezos and Bell[96].They attributed these differences in biosorption capacity foranaerobic and aerobic biomass to the varying lipid compo-sition and content of different biomasses. Another factorinfluencing the sorptive capacity of anaerobic biomass wasgiven as the test temperature (35◦C) which is higher thanthat of Tsezos and Bell’s study (20◦C) assuming adsorptionis exothermic[43].

Logan et al. [97] studied the biosorption of PCP by12 species of white rot fungi. In general PCP adsorptionto mycelia was very low, ranging from 1 to 5 mg PCP gmycelium (dry weight basis)−1 at 40 mg l−1 PCP. Severalspecies of fungi, includingP. chrysosporium, T. versicolorand all Ganodermasp. removed >50% of the PCP within24 h [97].

Jacobsen et al.[98] also studied the adsorption and des-orption of PCP to activated sludge biomass[98].

Brandt et al.[99] investigated the biosorption of PCPon the cells ofMycobacterium chlorophenoliciumPCP-1,a Gram-positive bacterium. They proposed that adsorp-tion mechanisms were a combination of reversible andirreversible adsorption processes. They suggested that thereversible adsorption was due to a physical adsorption(ion exchange) of the ionic PCP on the cell wall, whilethe irreversible adsorption was due to interaction betweenvery hydrophobic PCP and hydrophobic cell membrane. Incontrast to the results of Tsezos and Bell[96], they foundthat PCP is only adsorbed on the cell wall and sorptionequilibrium was normally reached in less than 1.5 min[99].

Daughney and Fein[100] observed that 2,4,6-trichloro-phenol (2,4,6-TCP) displayed a strong affinity for the cellwalls of B. subtilis. They described 2,4,6-TCP-B. subtilissorption by a surface complexation model in which boththe negative and the neutral forms of TCP form 1:1 surfacecomplexes with the neutral hydroxyl functional groups ofthe bacteria[100].

Aksu and Yener[39,101] investigated the biosorption ofphenol and monochlorinated phenols (o-chlorophenol and

p-chlorophenol) on the dried activated sludge. They sug-gested that the cell wall of bacteria essentially consisting ofvarious organic compounds such as chitin, acidic polysac-charides, lipids, amino acids and other cellular componentsis responsible for the uptake of organic pollutants. Theydetermined the maximum saturation capacity of biosorbentas 236.8, 281.1, and 287.2 mg g−1 for phenol, o-CP andp-CP, respectively. They found a higher adsorption capac-ity for o- and p-CP than that of phenol. They explainedthis behaviour due to the activation of the aromatic ring ofmonochlorophenol by the chlorine. This favored the forma-tion of donor–acceptor interactions between the phenoliccompound and the groups of the biosorbent surface. Theyalso suggested that biosorption capacity is affected by theposition of –Cl group on the ring. Biosorption ofp-CP hav-ing a chlorine in thepara position tended to be higher thano-chlorophenol containing chlorine in onlyortho position.This may be a result of steric hindrance between the –Cland –OH group in the case ofo-chlorophenol[39,101].

Benoit et al.[102] studied the biosorption of 2,4-dichlor-ophenol (2,4-DCP) and 4-chlorophenol (4-CP) on thefreeze-dried mycelium ofEmericella nidulansandPenicil-lium miczynskii, isolated from composted wheat straw anda soil, respectively. Results obtained with inactivated fungalcells showed thatP. miczynskiihad a higher biosorption ca-pacity for 2,4-DCP than that ofE. nidulansdue to differentbiochemical compositions and physicochemical propertiesof both fungi, and the biosorptive uptake of 2,4-DCP washigher than that of 4-CP. Higher sorption of 2,4-DCP wasexplained with increase in the polarity, decrease in the pKaof the phenol and enhance in the hydrophobicity of the sub-stituted benzene ring due to the increase in chlorosubstitu-tion. The rapid adsorption on fungal cell walls surfaces wasthe main sorption phenomenon for the more hydrophobicmolecules[102].

Jianlong et al.[103] characterized the adsorption be-haviour of pentachlorophenol from aqueous solution toactivated sludge biomass collected from a local biologicalwastewater treatment plant[103].

The ability of dried activated sludge and dried anaero-bic sludge to adsorb phenol was investigated by Aksu andAkpinar [104,105] in the absence and in the presence ofheavy metal ions. They found that dried anaerobic sludgehad a higher adsorption capacity than that of dried activatedsludge[104,105].

Calace et al.[40] studied the sorption capacity of pa-per mill sludge produced from primary (sedimentation)and secondary (biological) treatments of a pulp and paperindustry wastewater, for phenol, 2-chlorophenol (2-CP),3-chlorophenol (3-CP), 4-chlorophenol (4-CP), 2-nitro-phenol (2-NP), 4-nitrophenol (4-NP), 2,4-dichlorophenol(2,4-DCP), 3,4-dichlorophenol (3,4-DCP) 3,5-dichlorophe-nol (3,5-DCP) and 2,4,5-trichlorophenol (2,4,5-TCP) underbatch conditions. The phenols selected were phenol andmono-, di- and trichlorophenols and mono-nitrophenolscharacterized by different pKa and solubility values. The


order of sorption capacity on papermill sludge was: 2-NP=4-NP 2-CP< phenol< 4-CP = 3-CP< 2,4-DCP<

3,4-DCP = 2,4,5-TCP< 3,5-DCP. On the basis of theorder of retention selectivity obtained, they hypothesizedthat hydrophobic interactions drive the phenol sorption onthe sludge surface and that the main chemical features ofphenols playing an important role in sorption mechanismsare solubility and pKa [40].

Rao and Viraraghavan[41] examined the phenol removalcapacity of pretreated non-viableA. nigerbiomass from anaqueous solution containing phenol at a concentration of1 mg l−1. They observed that sulphuric acid-treated biomasspowder was the most effective for the removal of phenol[41].

A biofilm represents a stable ecosystem. The main com-positions of biofilms are bacterial cells, extracellular poly-mers produced by the bacteria (exopolymers), lysis andhydrolysis products, attached organic matter and some inor-ganic compounds. Exopolymers consist mainly of polysac-charides, proteins, uronic acids, humic acids, nucleic acids,lipids, and cell fragments. In biofilms, possible sorptionsites are extracelluar polymeric substances, cell wall, cellmembranes, and cytoplasm. These sites contribute to thesorption properties of biofilms to organic and inorganic sub-stances. Wang et al.[106] investigated the p-chlorophenol(4-CP) biosorption to biofilm components under the con-ditions of temperature 25◦C, pH values of 2.7, 5.3, and6.1 in a batch system. They chose biofilm coated kaolin,bacteria and bacterial exopolysaccharide (EPS) as biofilmcomponents for the adsorption of 4-CP. The bacteria usedin this study was identified asKluvera cryocrescens. Theyexplained that the interaction between 4-CP and biomassis poorly understood. There could be the protonation ofcarboxyl, phosphate, and hydroxyl functional groups onK.cryocrescensbacteria. The bacterial surface develops a neg-ative electric potential due to deprotonation of its surfacefunctional groups. This potential affects the hydrophobicityof the surface, and it influences interactions between thesurface sites and charged species in solution[106].

Karim and Gupta [107] investigated the biosorp-tion of 2-nitrophenol (2-NP), 4-nitrophenol (4-NP) and2,4-dinitrophenol (2,4-DNP) on live anaerobic granularsludge in order to understand their fate in an upflow anaer-obic sludge blanket reactor. They demonstrated that theuptake of each nitrophenol was only weakly correlated totheir respective octanol–water partition coefficients[107].


Tsezos and Bell[96] compared the biosorption of pen-tachlorophenol by live activated sludge andR. arrhizuswiththe results obtained by treated (autoclaved) activated sludgeandR. arrhizus. They observed that the PCP uptake capac-ity of deadR. arrhizuswas six times higher than that of liveR. arrhizus. In the case of activated sludge, however, theyfound that live sludge had higher biosorption capacity. They

attributed this situation to the biodegradation of a portion ofthe PCP[96].

Rao and Viraraghavan[41] examined the phenol removalcapacity of pretreated non-viableA. nigerbiomass from anaqueous solution. They pretreated live fungal pellicles in fivedifferent ways: autoclaving, treatment with 0.1 M HNO3,treatment with 0.1 M H2SO4, treatment with 0.1 M NaOH,treatment with a laboratory detergent (5 g l−1). They foundthat the removal trend was as follows: sulfuric acid pre-treated (50%) > laboratory detergent (42%) > sodium hy-droxide (39.9%) > autoclaved biomass (26.8% > nitric acid(20.6%). They proposed that the difference in the percent-age removal between H2SO4 and HNO3 pretreated biomasscould be due to the difference in the numbers of H+ ionscontributed by them at the same molarity during pretreat-ment. Sodium hydroxide pretreated and laboratory detergentpretreated biomasses gave very close removals. Alkali pre-treatment has also been shown to be a probable cause fordestroying autolytic enzymes that cause putrefaction of thebiomass and for removing lipids and proteins that mask re-active sites[41]. Gallagher et al.[85] have also suggestedthat NaOH pretreatment may also increase the percentageof chitin/chitosan in the whole cell wall fraction by dissolv-ing certain biopolymers from the cell wall. Chitin/chitosanunits may then be responsible for the uptake of phenol inthe case of NaOH pretreated biomass[85].

Benoit et al.[102] pretreatedE. nidulansandP. miczynskiiby freeze drying, autoclaving (121◦C, 103.4 kPa for 20 min)and autoclaving in the presence of formaldehyde at 30 g l−1

and they observed that for both fungi, autoclaving increasedthe 2,4-DCP biosorption capacity 10% higher than that offreeze-dried and autoclaved with formaldehyde cells due tothe solubilization of carbohydrates and peptidic compoundsresulted in modifications of chemical compositions on thefungal cell walls during heat sterilization[102].

4.3. Effect of pH on phenolics biosorption

Brandt et al.[99] observed that the adsorption capacityof M. chlorophenoliciumPCP-1 for PCP increased with de-creasing pH. The overall adsorption capacity at pH 5.4 wasabout eight-fold higher than that at pH 7.0. The effect ofpH on the sorption behaviour was found to be related to theionization of PCP. They demonstrated that the irreversiblyadsorbed PCP is a function of undissociated PCP, while thereversibly adsorbed PCP correlates well with the concentra-tion of ionic PCP[99].

Aksu and Yener[39,101]defined the pH of the sorptionmedium as the most critic parameter in the treatment ofphenol and monochlorinated phenols by the dried activatedsludge that affects biosorption capacity. They obtained thatphenol and monochlorinated phenols were more effectivelyadsorbed to the biomass at very low values of pH and theydetermined the optimum initial pH as 1.0. They explainedthe change in biosorption due to pH with the ionization ofphenols and chlorinated phenols, alterations in the sorbent


surface and the interaction of phenols and chlorinated phe-nols with the cells with primarily electrostatic forces orcomplex formation or electron share in nature or membranetransport[39,101].

Jianlong et al.[103] demonstrated that the initial pH valuewas an important parameter affecting the pentachlorophe-nol adsorption capacity of activated sludge biomass thatincreased with decreasing pH (between 6.0 and 8.0).The equilibrium adsorption capacity dropped from 3.0 to2.0 mg g−1 when the initial pH was increased from 6.0 to8.0. They explained the decrease in adsorption capacity ofbiomass with an increase in pH value due to the fact thatas the pH increased, the overall surface charge on the cellsbecame negative and this led to a lower electrostatic attrac-tion between negatively charged PCP and binding sites ofthe biomass surface[103].

Aksu and Akpinar[104,105]studied the uptake of phenolby both the dried activated sludge and dried anaerobic sludgeas a function of initial pH. The results obtained showed thatthe uptake of phenol increased with decreasing initial pH andwas the greatest at pH 1.0 for both biosorbents[104,105].

Rao and Viraraghavan[41] also studied the effect of initialpH on the phenol removal capacity of sulfuric acid pretreatedA. nigerbiomass. They obtained the maximum removal ofphenol at an initial pH of 5.1 and they reported that an in-crease or decrease in the pH from this optimum pH resultedin a reduction in the biosorption of phenol. They explainedthe biosorption mechanism due to pH as a fact that phenolcould be expected to become negatively charged phenoxideion above a pH of 9.0 because of pKa value of phenol (=9.99at 25◦C). The surface charge on fungal biomass is predom-inantly negative over the pH range of 3.0–10.0. Below a pHof 3.0, the overall surface charge on fungal cells becomespositive. Pretreating it with sulfuric acid could generate pos-itively charged sites on its surface due to the sorption ofan excess of H+ ions. Phenol being weakly acidic will bepartially ionized in solution. These ions will be negativelycharged and will be directly attracted due to electrostaticforces by the positively charged fungal biomass surface.Unionized phenol molecules will also be attracted, possi-bly by physical forces. At a high basic pH range, OH− ionswould compete with the phenol molecules for biosorptionsites. Sorption of an excess of OH− ions could convert aninitial positively charged surface into a negatively chargedsurface. This surface would then repel the negatively chargedphenoxide ions. This was probably the reason why there wasa decrease in biosorption at the basic range with ultimatelyno biosorption taking place at an initial pH of 10.0. At a lowacidic pH of 2.0, phenol molecules get protonated and sub-sequently positively charged. This causes repulsion betweenthe positively charged fungal surface and phenol moleculesleading to a decreased uptake of phenol[41].

Calace et al.[40] studied the sorption capacity of papermill sludge for phenol, and mono, di- and trichlorophenolssolutions at pH 2.0 and 11.0. They found that the amountadsorbed increased with decreasing pH due to increase the

number of hydrophobic adsorptive domains on the surfaceand the undissociated fraction of chlorophenols that are ableto interact at low pH value[40].

Wang et al.[106] investigated the impact of pH valueon the p-chlorophenol (4-CP) biosorption to biofilm com-ponents over the pH range 2.0–6.0 for an initial adsorbateconcentration 20 mg l−1. They observed that the amount ad-sorbed slightly increased with increasing pH from 2.7 to6.1 for the adsorption system of bacterial exopolysaccharidebut it slightly decreased in the systems of biofilm coatedkaolin andK. cryocrescensbiomass. They proposed that4-CP was in the dissociated and undissociated forms in so-lutions within the pH range in these experiments. When pHdecreased, H+ ion concentration increased and the undis-sociated forms of 4-CP increased in proportion. The undis-sociated 4-CP was highly hydrophobic, it was easier to beadsorbed than its dissociated form. The decrease of pHwas useful for 4-CP adsorption on bacteria and kaolin withbiofilm coating. When pH increases, bacterial exopolysac-charide adsorbed more 4-CP[106].

4.4. Effect of initial phenol concentration on phenolicsbiosorption

Aksu and Yener[39,101] observed that the equilibriumsorption capacity of the dried activated sludge for phenol,o-chlorophenol andp-chlorophenol increased with the initialpollutant concentration up to 500 mg l−1. They suggestedthat the increase of loading capacity of biosorbent with theincrease of pollutant concentration may be due to higherprobability of collision between pollutants and biosorbentand sufficient active sites[39,101].

Jianlong et al.[103] also observed that the equilibriumsorption capacity of the activated sludge increased with in-creasing initial PCP concentration up to 0.5 mg l−1 [103].

The results obtained by Aksu and Akpinar[104,105]showed that the equilibrium uptake of phenol by dried aer-obic and anaerobic sludges increased with increasing initialphenol concentrations up to 500 mg l−1 for both biosor-bents. The curvilinear relationship between the amountof phenol adsorbed per unit weight of each microorgan-ism and the residual phenol concentration at equilibriumsuggested that saturation of cell-binding sites occurred athigher concentrations of this component[104,105].

Karim and Gupta[107] studied the sorption 2-NP, 4-NPand 2,4-DNP on live anaerobic granular sludge at differentsorbate (nitrophenols) concentrations changing between 10and 90 mg 1−1 and they found that equilibrium sorption ca-pacity of biomass increased with increasing each initial ni-trophenol concentration[107].

4.5. Effect of biosorbent concentration on phenolicsbiosorption

Brandt et al.[99] demonstrated that the adsorption capac-ity of M. chlorophenoliciumPCP-1 for PCP increased sig-


nificantly with decreasing biomass concentration in the lowconcentration range (below 0.5 g l−1). They explained thissituation by a smaller accessible sorbent surface resultingfrom cells that are closer to each other at high cell density[99].

Jianlong et al.[103] investigated the effect of acti-vated sludge concentration on the adsorption of PCP. Thebiomass concentration varied from 0.5 to 5.0 g l−1. Theyobserved that on one hand, as expected, the percentageof the PCP removal increased with increasing activatedsludge concentration. But conversely, the absorbed amountof PCP per biomass quantity decreased with the increasingbiomass concentration. The adsorption capacity droppedfrom 2.6 to 1.1 mg g−1 when the biomass concentrationincreased from 0.5 to 5.0 g l−1. The drop in adsorption ca-pacity was explained due to binding sites of the biomassremaining unsaturated during the adsorption reaction[103].

4.6. Effect of heavy metal ions on phenolics biosorption

Wastewaters of steel, metal plating, dye, textile, andpainting industries frequently encounter organics along withheavy metal ions. The examining the combined effects ofmetal ions and organics in various combinations is morerepresentative, of the actual environmental problems facedby organisms, than are single organic studies. In mixturesof two or more components in a solution, the synergisticor antagonistic interaction occurring between the compo-nents may affect the individual component uptake by themicroorganism.

The phenol uptake by dried activated sludge in the pres-ence nickel(II) ions and the biosorption of phenol by driedanaerobic sludge in the presence chromium(VI) ions wereinvestigated by Aksu and Akpinar[104,105]with respect toinitial pH and initial concentration of each component in theadsorption medium. They observed that the equilibrium up-take of phenol was diminished by the addition of metal ionsin the concentration range of 25–500 mg l−1 due to antago-nistic interaction between the components for both sorbents.The most logical reason for this behaviour was claimed tobe the competition for similar binding sites on the surface ofcells and/or the screening effect by the second component[104,105].


Tsezos and Bell[96] used distilled deionized water forthe desorption of PCP from live cells of activated sludge andR. arrhizus, and demonstrated that the biosorption of PCPby both live cells was partially irreversible[96].

Kennedy et al.[43] used a method based on contact-ing the loaded biomass with the adsorbate solution forthe desorption of 3-chlorophenol, 3,4-dichlorophenol and2,4,6-trichlorophenol from anaerobic granular sludge. Theyfound that the adsorption of 3,4-DCP and 2,4,6-TCP was

almost completely reversible while 3-CP exhibited a greaterdegree of irreversibility[43].

Brandt et al.[99] found that the desorption of PCP fromM.chlorophenoliciumPCP-1 from was strongly affected by pH.At pH 5.4 the adsorption was almost completely irreversible,while a nearly complete desorption was obtained at pH 7.0[99].

Benoit et al.[102] used 0.01 M CaCl2 to desorb 2,4-DCPand 4-CP from bound freeze-driedE. nidulanscells andfound that biosorption was only partially reversible and thedecrease of irreversibility of biosorption was related to an in-crease in chlorosubstitution[102]. This was consistent withresults of Kennedy et al.[43] who also observed that the ir-reversibility of sorption on bacterial biomass was strongerfor 3-chlorophenol than for 3,4-dichlorophenol[43].

Rao and Viraraghavan[41] found that the desorption ofphenol fromA. niger with distilled deionized water gavethe maximum desorption of only about 5% which suggestedstrong biosorption by the biomass[41].

Karim and Gupta[107] used a dilution media (0.01 MNaOH solution adjusted to pH 7.5) to desorb 2-NP, 4-NP,and 2,4-DNP from bound anaerobic granular sludge anddemonstrated that the sorption nitrophenols was partiallyreversible. About 20–89%, 36–90%, and 29–80% desorptionwas observed for 10–90 mg 1−1 sorbate concentrations of2-NP, 4-NP, and 2,4-DNP, respectively[107].

4.8. Equilibrium modeling of biosorption

Tsezos and Bell[96] reported that the biosorption of PCPto living and dead activated sludge andR. arrhizusbiomasseswas non-linear and could be described by the Freundlichequation[96].

Kennedy et al.[43] investigated the fitting of sorption dataof phenol and a number of chlorophenols by anaerobic gran-ular sludge to the Freundlich equation. The authors reportedthat most chlorophenols had linear sorption isotherms (1/nclose to 1), which were defined by simple distribution co-efficients (Eq. (7)). A linear adsorption isotherm commonlyoccurs when relatively pure, porous sorbents are used and isbeing carried out over a relatively small concentration range.The availability of sorption sites remains constant at all con-centrations up to saturation, typical of the partitioning of asolute between two immiscible solvents. They proposed thatthe granules which consist of a consortium of anaerobic bac-teria and are known to be very porous allowing penetrationof sorbate to sorption sites in the inner portion of the gran-ules, behaved mainly as a homogeneous and pure sorbent.However, these distribution coefficients were found onlyweakly correlated to octanol–water partition coefficients[43].

Jacobsen et al.[98] also observed linear sorptionisotherms in the biosorption of PCP by activated sludgeup to 0.08 mg l−1 of dissolved PCP at pH 6, and up to0.16 mg l−1 at pH 8. They concluded that linear sorptioncoefficients were primarily influenced by pH, although ionic


strength (owing to pH-buffering) and the concentration ofdissolved organic matter also had an impact[98].

Brandt et al.[99] observed that the sorption equilibriumdata of PCP onM. chlorophenoliciumPCP-1 was adequatelydescribed by the Freundlich equation, indicating no clearadsorption saturation. They also investigated the effect ofbiomass concentration onKF andn and found thatKF wassignificantly affected by lower biomass concentration whilen was independent of the biomass concentration[99].

The sorption phenomena of phenol,o-chlorophenol andp-chlorophenol to dried activated sludge were expressed bythe Langmuir and Freundlich adsorption models by Aksuand Yener[39,101]. The adsorption isotherms showed thatthe equilibrium data for all the pollutants fitted well to boththe Langmuir and Freundlich models in the studied concen-tration range studied for all the pollutants[39,101].

Benoit et al.[102] applied the Freundlich model to de-scribe the biosorption equilibrium of 2,4-DCP and 4-CPfrom aqueous solution to freeze-dried and autoclaved withformaldehydeE. nidulans biomasses. These authors ob-tained that in both conditions, sorption isotherms were closeto linearity (1/n close to 1) and autoclaved with formalde-hydeE. nidulanshad lowerKF values for 2,4-DCP and 4-CPwhich could be due to a competitive effect of formaldehydefor the sorption sides[102].

Jianlong et al.[103] applied the Freundlich model todescribe the biosorption equilibrium of pentachlorophenolfrom aqueous solution to activated sludge biomass by chang-ing the initial PCP concentrations from 0.025 to 0.5 mg l−1.Both the biomass concentration and pH value only affectedthe capacity constantKF of the Freundlich equation whilethe intensity constantn remained constant. Attempts to usethe Langmuir equation to fit the adsorption isotherm failedto provide a satisfactory correlation[103].

Aksu and Akpinar[104] defined the equilibrium up-take of phenol in the presence of nickel(II) ions by driedactivated sludge by competitive multi-component Lang-muir, Freundlich and Redlich–Peterson adsorption models(Eqs. (8)–(11)) and estimated model parameters by thenon-linear regression method at two optimum biosorptionpH values determined as 4.5 for nickel(II) and as 1.0 forphenol. The results showed that all the multi-component ad-sorption models adequately predicted the multi-componentadsorption equilibrium data at moderate ranges of concen-tration [104].

Aksu and Akpinar[105] also expressed the equilibriumuptake of phenol by dried anaerobic sludge in the presenceof chromium(VI) ions by multi-component Langmuir, Fre-undlich and Redlich–Peterson adsorption models and theyfound that modified Freundlich model adequately predictedthe multi-component adsorption equilibrium data at moder-ate ranges of concentration[105].

Calace et al.[40] attempted to fit the three equations(Langmuir, Freundlich and Langmuir-Freundlich) to theirexperimental data. The results showed that the biosorptionequilibrium of phenol, mono-, di-, and tri-chlorophenols and

mono-nitrophenols fitted the Hill equation very well, whichis mathematically equivalent to the Langmuir–Freundlichmodel (Eq. (3)) obtained by assuming that the surface is ho-mogeneous, and that the adsorption is a cooperative processinfluenced by adsorbate–adsorbate interactions[40].

Rao and Viraraghavan[41] fitted the Langmuir, Fre-undlich and the Brunauer–Emmet–Teller isotherm modelsto the isotherm data of phenol biosorption by H2SO4 treatedA. nigerbiomass. They found that the adsorption of phenolby treated biomass was best described by the BET model[41].

The adsorption data ofp-chlorophenol by biofilm com-ponents (biofilm coated kaolin, bacterial exopolysaccharideandK. cryocrescens) was analyzed with Langmuir and Fre-undlich models to evaluate the parameters in the adsorptionprocess by Wang et al.[106] at 25◦C. All the adsorptions fit-ted to these two equations. The maximum adsorption capac-ity was 72.3 mg 4-CP g−1 dry bacteria and 6.6 mg 4-CP g−1

biofilm coated kaolin[106].Karim and Gupta[107] fitted the Freundlich equation to

the equilibrium data of three nitrophenols (2-NP, 4-NP and2,4-DNP). Since they found 1/n, the Freundlich constant, as1.0 for all nitrophenols, they concluded that there is a linearrelationship between nitrophenol concentration in the liq-uid and solid phases and equilibrium can be shown by lin-ear adsorption isotherms suggesting a constant-partitioningsorption mechanism[107].


The detailed investigation on sorption of 15 chlori-nated phenols onto anaerobic granular sludge conducted byKennedy et al.[43] indicated that a rapid biosorption oc-curred and equilibrium was reached in less than 2 h for thesorption of all chlorophenols by anaerobic granular sludge[43].

Jacobsen et al.[98] observed that equilibrium conditionsfor sorption and desorption of pentachlorophenol to micro-bial biomass were established within 5 min[98].

Daughney and Fein[100] observed that the sorption anddesorption reactions of 2,4,6-trichlorophenol by the cellwalls of B. subtiliswere reversible and rapid. The adsorp-tion equilibration occurred in approximately 30 min[100].

Aksu and Yener[39,101]observed that initial sorption ofphenol occurred very rapidly by the dried activated sludge;sorption reached equilibrium in 45–60 min at 100 mg l−1 ini-tial phenol concentration. Adsorption proceeded at a slowerrate in the case ofo- andp-chlorophenol. They noted thatalthough the initial phenol biosorption rate by dried acti-vated sludge was higher than that ofo-andp-chlorophenol,the equilibrium uptakes of phenol by biomass were lowerthan that ofo- andp-chlorophenol[39,101].

Benoit et al. [102] indicated that the biosorption of2,4-DCP by treatedE. nidulansandP. miczynskiicells wasfast and the equilibrium was reached within the first 3 h ofcontact[102].


Jianlong et al.[103] indicated that the pentachlorophenolbiosorption equilibrium was reached in less than 2 h underexperimental conditions. No changes in PCP concentrationswere observed during prolonged shaking[103].

Rao and Viraraghavan[41] reported that biosorption ofphenol by treatedA. niger biomass reached equilibriumwithin 24 h.[41].

Kinetic experiments performed by Calace et al.[40]showed that substituted chlorophenol sorption on papermillsludge was rapid (equilibrium was reached after 3 h); con-versely, the time taken by the phenol to reach equilibriumconditions was 260 h. Evaluating the experimental databy using Eq. (17) they found that intraparticle diffusionwas involved in the sorption process but was not the onlyrate-limiting mechanism; several other mechanisms wereinvolved [40].

Wang et al.[106] investigated the kinetic characteristicsof 4-CP adsorption by biofilm components. They observeddifferent times to reach equilibrium in each system. The4-CP adsorption by kaolin with biofilm coating and EPSreached near equilibrium at 60 and 180 min at pH 6.1, re-spectively. But the 4-CP adsorption by bacteriaK. cryocre-scensshowed no evident near equilibrium during 3 h inthe experiment, the adsorption reached equilibrium slowly[106].

4.10. Column studies

Rao and Viraraghavan[40] studied phenol biosorption byimmobilized H2SO4 treatedA. niger biomass in polysul-phone in a packed bed column. They obtained phenol re-moval as 66% from breakthrough curves for up to 2 h at1 mg l−1 phenol concentration in the influent with an ad-justed pH of 5.1 at 7.0 ml min−1 flow rate. The breakthroughdata were fitted to the Thomas model (Eq. (21)). Biosorption

Table 3Data on the biosorption of pesticides by various microorganisms

Biosorbent Pesticide Operation conditions Biosorption capacity Ref.


Activated sludge Lindane – 20 1a 3 days 1.6 [95]Diazinon – 20 1a 3 days 0.5Malathion – 20 1a 3 days 16.9

Bacillus subtilis Lindane – 20 4 4 h 0.6 [110]Bacillus megaterium Lindane – 20 4 4 h 0.7Emericella nidulans 2,4-D 6.0 20 221 3 h 2.1 [102]Escherichia coli Lindane – 20 4 4 h 0.5 [102]Mucor racemosus PCNB – 21 250 6 h 5.1 [112]Rhizopus arrhizus PCNB – 21 250 6 h 4.6 [112]

Rhizopus arrhizus Lindane – 20 1a 3 days 2.7 [95]Diazinon – 20 1a 3 days 0.5Malathion – 20 1a 3 days 13.22-Chlorobiphenyl – 20 1a 3 days 11.1

Sporothrix cyanescens PCNB – 21 250 6 h 2.6 [112]Zooglea ramigera Lindane – 20 4 4 h 2.8 [110]

a Equilibrium concentration.

capacity of the beads for phenol was found to be 0.2 mg g−1

with a correlation coefficient of 0.93[40].Aksu and Gonen[108] used Mowital®B30H resin immo-

bilized dried activated sludge for the biosorption of phenolin a continuous fixed bed. They applied four kinetic mod-els; Adams-Bohart (Eq. (19)), Thomas, Clark (Eq. (22))and Yoon-Nelson (Eq. (25)) models to experimental datato predict the breakthrough curves and to determine thecharacteristic parameters of the column useful for processdesign. All models were found suitable for describing thewhole or a definite part of dynamic behaviour of the columnwith respect to flow rate and inlet phenol concentration.They determined the biosorption capacity of immobilizedsludge for phenol as 6.4 mg g−1 at 52.3 mg l−1 initial phe-nol concentration with an adjusted pH of 1.0 at a flow rateof 0.8 ml min−1 from the Thomas model[108].

5. Biosorption of pesticides

Application of biosorption for pesticides is also possibleand several microorganisms including bacteria and fungihave been studied for the removal of some pesticides.Table 3presents a comparison of biosorption capacities ofsome microorganisms for some pesticides at their workingconditions.

5.1. Microorganisms used in pesticide biosorptionand biosorption mechanisms

Bell and Tsezos[95] and Tsezos and Bell[96] stud-ied the biosorption of lindane, diazinon, malathion and2-chlorobiphenyl which are widely used organochlorine andorganophosphorus insecticides, onto live and dead cells ofactivated sludge andR. arrhizus. They proposed that a part


of the observed biosorptive uptake can be attributed to thecell walls of the microbial biomass. Moreover the biosorp-tion process in the case of diazinon and lindane, involvedan exothermic physical, rather than chemical, mechanism.They found that the octanol/water partition coefficient wasdirectly related to the biosorptive uptake of these organicsonto live and dead biomasses[95,96].

Ju et al.[110] investigated the biosorption of lindane, anorganochlorine pesticide onto dried Gram-negative bacte-ria (E. coli, Zooglea ramigera) and Gram-positive bacteria(Bacillus megaterium, B. subtilis). They proposed that hy-drophobic interaction and van der Waals forces are involvedin the biosorption of lindane. They found that among thefour bacteria,Z. ramigerashowed the maximum uptake ca-pacity [110].

Young and Banks[111] used a heat treated non viable cellsuspension of the fungusR. oryzaefor the removal of lowconcentrations of lindane from aqueous solution in a batchsystem. The results indicated that the mechanism of adsorp-tion was by physical bonding of the negatively charged lin-dane molecule to the negatively charged fungal cell wallwith hydrogen ions acting as the bridging ligand[111].

Lievremont et al.[112] reported on the removal of pen-tachloronitrobenzene (PCNB), a fungicide, from aqueoussolution by dead fungal mycelia ofMucor racemosus, R.arrhizus, and Sporothrix cyanescensand compared withsorption on isolated cell walls of these three strains. Theyproposed that biosorption involved both uptake by the cellwalls and by other cellular components. Size of cells, mor-phology and chemical composition as well as the numberof the active adsorption sites and their distribution may playa significant role in determining uptake capacity. Sorptionof the adsorbate is also dependent on its molecular size andreactivity as well as mobility in the solution phase[112].

Benoit et al.[102] investigated the biosorption charac-teristics of herbicide, 2,4-dichlorophenoyacetic acid (2,4-D)on the freeze-dried fungal mycelium ofE. nidulans, iso-lated from composted wheat straw. They observed a loweradsorption of 2,4-D (2.1 mg g−1 at 221 mg l−1 initial 2,4-Dconcentration) due to the negative charge due to dissociationof the carboxylic group and due to electrostatic repulsioneffects at pH 6.0[102].

Hong et al.[113] studied the biosorption of 1,2,3,4-tetra-chlorodibenzo-p-dioxin (1,2,3,4-TCDD) and some poly-chlorinated dibenzofurans (PCDFs) pesticides byBacilluspumilus. The results showed that dead biomass of microor-ganism could remove these molecules from the mediummore effectively than live cells. They suggested that inaddition to the attachment to microorganisms itself, extra-cellular polymeric substances might also be involved in thebiosorption process[113].


Bell and Tsezos[95] and Tsezos and Bell[96] used thelive and dead activated sludge andR. arrhizusbiomasses.

They prepared both types of pretreated biomass by firstlyautoclaving and then, washing and drying. They found thatuptake of lindane by both live biomass types was less thanthe respective uptake by the same dead cells. Diazinon wasbiosorbed by liveR. arrhizusto approximately the same levelas the dead biomass. Uptake of diazinon by live activatedsludge was observed to be similar to the uptake by deadsludge. For 2-chlorobiphenyl, the uptake by both liveR. ar-rhizusand activated sludge was found higher than the uptakeby the same dead cells so they concluded that it was impos-sible to generalize on the relative magnitude of biosorptionuptake between live and nonviable biomass[95,96].

Ju et al.[110] indicated that the treatment of 1 g of wetcell of E. coli or Z. ramigeraor B. megateriumor B. sub-tilis with 15 mM EDTA for 5 min increased the biosorptioncapacity of each cell—the most significant 156% increasein biosorption of EDTA treatedE. coli was observed—as aresult of increased permeability toward lindane[110].

Benoit et al.[102] pretreated the biomass ofE. nidulansby freeze drying and autoclaving at 121◦C, 103.4 kPa for20 min by adding formaldehyde at 30 g l−1 for the biosorp-tion of 2,4-D. They found that chemical sterilization by ad-dition of formaldehyde at 30 g l−1 decreased the biosorptioncapacity of fungus which could be due to a competitive ef-fect of formaldehyde for the sorption sites[102].

5.3. Effect of pH on pesticide biosorption

Ju et al. [110] investigated the effect of pH between2.93 and 6.88 on the biosorption of lindane byE. coli, Z.ramigera, B. megaterium, and B. subtilis. They observedhigher biosorption under lower pH. They found the isoelec-tric points of all bacteria at pH 2.0, except forE. coli whoseisoelectric point is at 3.0 so all cells are negatively chargedabove these pH values. They proposed that the repulsiveelectrostatic force for the adsorption of organic halide onthe cell surface decreases when a lower pH generates lessnegative charge on cell surfaces. As the cell and lindanemolecules move closer to each other, owing to the decreasein electrostatic force, the van der Waals force is intensifiedand biosorption is enhanced consequently[110].

Young and Banks[111] studied at different pH valueschanging from 2.0 to 10.0 in order to investigate the effectof pH on the biosorption of lindane by heat treatedR. oryzaeand they found that biosorption was most effective at lowpH [111].

5.4. Effect of temperature on pesticide biosorption

Bell and Tsezos[95] studied lindane, diazinon andmalathion biosorption by both dead and live cells of acti-vated sludge andR. arrhizusat two different temperatures(5 and 20◦C) and they found that the biosorption of lin-dane on both type of microorganisms was exothermic andthe negative value of heat of biosorption was in the rangewhere physical adsorption rather than chemisorption was


the dominant mechanism. They also found low adsorptionenergies for diazinon biosorption suggesting that a physi-cal adsorption process was dominant. Diazinon biosorptionby activated sludge was endothermic. Based on the largepositive value of&H it was probable that a chemical re-action is involved in the biosorption of malathion by bothbiosorbents[95].

Ju et al.[110] investigated the biosorption of lindane byE. coli, Z. ramigera, B. megaterium, and B. subtilis as afunction of temperature. They found that except forE. coli,increasing temperature tended to decrease the biosorptioncapacity of lindane for the other three bacteria[110].

Young and Banks[111] investigated the effect of temper-ature on the biosorption of lindane by driedR. oryzain therange of 5–45◦C and they found that adsorption was mosteffective at low temperature[111].

5.5. Effect of initial pesticide concentration on pesticidebiosorption

The results obtained by Ju et al.[110] indicated that the in-crease in initial lindane concentration from 1 to 4 mg l−1 at aconstant cell concentration of 8 g l−1, increased the biosorp-tion capacity of lindane by dried bacteria ofZ. ramigera, E.coli, B. subtilis, andB. megateriumfrom 370 to 2800�g g−1,from 98 to 500�g g−1, from 100 to 600�g g−1, and 100 to700�g g−1, respectively[110].

5.6. Effect of ionic strength on pesticide biosorption

Ju et al.[110] investigated the effect of initial cell con-centration on the biosorption of lindane byB. megateriumat an initial lindane concentration of 4 mg l−1. The resultsobtained showed that lindane biosorption capacity increasedfrom 340 to 700 mg g−1 with increasing cell concentrationfrom 2 to 16 g l−1 [110].

Young and Banks[111] observed that the biosorption ca-pacity of R. oryzaefor lindane increased with increasingbiomass density from 1 to 12 g l−1 [111].

5.7. Effect of ionic strength on dye biosorption

Ju et al.[110] investigated the effect of ionic strength onthe biosorption of lindane byB. megateriumby adding 0.1 Mof NaNO3 to the aqueous lindane solution. The results ob-tained showed that lindane biosorption increased by increas-ing ionic strength. Ionic strength influenced the biosorptionby affecting the surface charge and the double-layer prop-erties of the cells[110].


Bell and Tsezos[95] studied the desorption of lindane, di-azinon and malathion from bound live and dead cells of acti-vated sludge andR. arrhizususing distilled deionized water.They demonstrated that the adsorption of lindane on both

live and dead cells of activated sludge andR. arrhizusandthe adsorption of diazinon on both species of dead biomasseswere completely reversible, indicating a low binding en-ergy characteristics of physical adsorption while diazinonappeared to desorb well from liveR. arrhizusbut not so wellfrom live activated sludge. In the biosorption of malathionthey observed reversibility at 5◦C, irreversibility at 20◦C,suggesting additional uptake at higher temperature[95].

Benoit et al.[102] used 0.01 M CaCl2 to desorb 2,4-Dfrom bound freeze-driedE. nidulansand observed an irre-versible biosorption between 2,4-D and biosorbent[102].

5.9. Equilibrium modeling

Bell and Tsezos[95] applied the Freundlich model tothe biosorption data of lindane, diazinon, malathion and2-chlorobiphenyl on both activated sludge andR. arrhizusand they found that the isotherms for lindane were linear(1/n = 1). A large ultimate adsorption capacity and lowenergy of adsorption could account for the linear isotherm.For diazinon biosorption byR. arrhizusequilibrium data fit-ted the Freundlich model while the isotherm for activatedsludge was essentially linear[95].

Ju et al.[110] defined the lindane biosorption equilibriumby E. coli, Z. ramigera, B. megateriumand B. subtilis interms of the Freundlich model[110].

Young and Banks[111] also applied the Freundlichmodel successfully to the biosorption data of lindane onR.oryzae[111].

Lievremont et al.[112] also described the biosorptionof PCNB on to dead fungal mycelia ofM. racemosus, R.arrhizus, andS. cyanescensby the Freundlich model[112].

Benoit et al.[102] described the equilibrium data of 2,4-Dby the pretreatedE. nidulansby Freundlich model[102].


Ju et al.[110] observed that biosorption of lindane byE.coli, Z. ramigera, B. megateriumandB. subtilisat first wasrelatively rapid and slowed down later, reaching equilibriumwithin 4 h [110].

Lievremont et al.[112] found a contact time of 6 h wassufficient enough to attain equilibrium for the biosorptionof PCNB on to dead fungal mycelia ofM. racemosus, R.arrhizus, andS. cyanescens[112].

The results obtained by Benoit et al.[102] indicated that afast equilibrium between 2,4-D in solution and 2,4-D sorbedon pretreatedE. nidulanswas occurred within the first threehours of contact[102].

6. Discussion

The usage of inactive microorganisms for the removalof organics including dyes, phenolics and pesticides fromwastewaters and the parameters affecting the biosorption


rate and capacity has been reviewed in detail in this paper.Use of untreated (live) or treated (chemical or heat treated)microorganisms as biological adsorbents has attracted at-tention during recent years because of fastness, low cost,easy availability, easy operating conditions, high efficiencyin detoxifying very dilute or concentrated effluents and nonutrient requirements and thus have been proposed to clean avariety of industrial effluents containing organic pollutants.Bacterial, fungal, and yeast strains are shown to be the mainmicroorganism types capable of removing organics fromwastewater. But the literature survey indicated that biosorp-tion studies of organics are very limited and only sorption ofselected toxic organics onto a few types of bacterial, fungaland yeast biomass have been investigated. There is a needto study with other organic pollutants and to develop newstrains which can be provided easily as waste and/or abun-dant biomass or can grow in simple, inexpensive mediumand have high production rate and possess high biosorptioncapacity.

The research indicated that there are various factors in-fluencing biosorption rate and capacity related to structureof organic molecule, microorganism type, and wastewatercharacteristics. Dye, phenolic, and pesticide molecules havemany different and complicated structures and biosorptionis affected considerably by the molecular size, charge, sol-ubility, hydrophobicity, and reactivity of organic molecule.Another factor affecting the biosorption is the mobility ofsorbate molecule in the solution phase and the hydrodynam-ics of contact between the solution and particle phase. Fur-ther research is needed to establish the relationships betweenthe structure of organic molecule and microbial biosorption.The type of biomass has also a significant effect on the ob-served biosorptive uptake. Biosorption of organic pollutantsis found strain specific, and the behaviour and mechanismsof biosorption of different sorbate/biomass systems are notwell understood. It was proposed that the size of cells, mor-phology as well as the number of the active adsorption sitesand their distribution related to surface chemistry may playa significant role in determining uptake capacity. The lack ofunderstanding of the mechanism of organic sorption processhinders the satisfactory estimation of process performanceand limitations, and thus the widespread application ofbiosorption. Further investigation is needed to identify themechanism of organic uptake by biomass. The feasibilityand efficiency of a biosorption process depends not only onthe properties of the sorbate and biosorbent, but also on thecomposition of wastewater and environmental conditions.pH, temperature, concentrations of biomass and organicpollutant, ionic strength, other components in wastewatersuch as metal ions, are one or more selected parametersinvestigated in each study presented in the literature as thecharacteristics of wastewater affecting biosorption capacity.Further study is needed to determine the effects of parame-ters which are not investigated before and other parameterssuch as particle size, stirring rate on the biosorption capac-ity. Despite the fact that industrial effluents contain several

pollutants simultaneously, little attention has been given tosorption of organics from mixture. More information onbiosorption is required to determine the best combinationof organics, biomass types and environmental conditions.

Some investigators observed that the uptake of organicpollutants by dead microbial biomass is greater than or equalto the uptake by the same living microorganisms due to theabsence of metabolic protection against transport of pollu-tants into the cell, increased permeability of the dead cellmembrane and the change of the surface adsorptive proper-ties of the microbial cell following its death. However a gen-eralization concerning the relative magnitude of biosorptiveuptake between live and dead biomass could not be made.For molecules which are not readily biodegradable, the over-all uptake by live biomass appears to be less than that of thesame dead biomass. For more readily degradable moleculesor for strongly adsorbing molecules, the reverse appears tobe true. Additional work on this subject is needed.

Research indicated that some pretreatment methods in-cluding heat-killing, contacting with organic chemicals, suchas formaldehyde, detergent, or inorganic chemicals, suchas NaOH, HCl, HNO3, H2SO4, NaHCO3, and CaCl2 forkilling cells before final drying and granulation effectivelyincreased the biosorption capacity of biomass for organicpollutants in many cases. Additional research is required toimprove the biosorptive capacity of many of these sorbents.

To make the biosorption process more economical it isnecessary to regenerate the spent adsorbent for reusing inmultiple sorption cycles. Numerous methods are reportedin the literature for the elution and regeneration of organicpollutant loaded biomass by some organic solvents, sur-factants and NaOH solution. Adsorption–desorption stud-ies indicated that yield was changed due to reversibility ofbiosorption. A more detailed study on regeneration methodsof spent biosorbents, desorption equilibrium and kinetics isneeded.

Biosorption is also a well known equilibrium separa-tion process for wastewater treatment containing organ-ics. Equilibrium data, commonly known as adsorptionisotherms, are basic requirements for the design of ad-sorption systems and provide information on the capacityof the biosorbent or the amount required to remove aunit mass of pollutant under the system conditions. Lang-muir, Freundlich, Langmuir–Freundlich, Redlich–Peterson,Brunauer–Emmet–Teller, Radke–Prausnitz models are com-monly used for describing the biosorption equilibrium oforganics at a constant temperature. Measured values ofmodel constants indicated significant differences in thecurve shapes and sorption capacities regarding microorgan-ism and organic pollutant. Frequently there is more than onepollutant in wastewaters and, in this case, the equilibriummodelling of multi-component biosorption have to be con-sidered for the real design of biosorption processes. Onlya few studies regarding binary biosorption and equilibriummodeling of binary biosorption of organics are available inthe literature. Further work should be carried out to better


characterize the performance of the biosorbents, studyingnot only organics but also removal of other pollutants fromsolution and, moreover, carrying out multi-component ad-sorption tests in order to study adsorption under competition.Further study is also needed to develop equilibrium modelsto predict the behaviour of each component in a mixture.

Kinetics studies give detailed information on sorbate up-take rates and rate controlling steps such as external masstransfer, intraparticle mass transfer, biosorption process. Sothe structure and particle size of biosorbent are the mostimportant parameters affecting biosorption rate. Althoughbiosorption equilibrium has been defined extensively, how-ever, only limited information is available regarding the ad-sorption rate and kinetics of organic removal processes. Theliterature on kinetic modelling is also rather limited and onlya few studies applying the first and second order kineticmodels to their biosorption data were reported. Moreover,no information is provided on the effect that diffusion of or-ganics into sorbent may have on the rate or nature of sorp-tion. Further investigation is needed to develop predictivemodels for the biosorption process kinetics playing an im-portant role in transferring technologies from the laboratoryto a full-scale application.

For the industrial application of biosorption, immobiliza-tion of biosorbent is necessary for solid/liquid separation.Immobilized biomass beads can then be packed in sorptioncolumn, which is perhaps the most effective device forcontinuous operations. For the optimization of a biosorp-tion process in a packed bed column, the effect of differentoperating conditions such as flow rate, initial pollutant con-centration, particle size, etc. on the biosorption capacityshould be studied and cheap and feasible immobilizingagents should be investigated. Mathematical models whichare the functions of sorption equilibrium, mass transfer andfluid-flow parameters are helpful to determine the overallsorption performance and the shape of the column break-through curve. The literature indicated that only limitedinformation is available on packed bed applications andmodeling of biosorption of organic pollutants and only afew studies were reported. Further study is needed to findout parameters for describing and predicting performanceof dynamic sorption systems and reactor design.

For the removal of organics, fungi, yeast and bacteria arerepresented the efficient classes of biosorbents relative toother biomass types. Although the biosorptive capacity andthe time needed to reach the equilibrium have differed fromone to another combination of microorganism-organic pol-lutant, the studies indicated that biomass had a much morehigher biosorption capacity for dyes than that of phenolicsand pesticides. Research indicated that pH of biosorptionmedium is the most important parameter influencing thebiosorption capacity. Biosorption of anionic (reactive andacidic) dyes by bacterial and fungal cells was significantparticularly under acidic conditions while cationic (basic)dye biosorption was accomplished at pH values higher than3.0. For the biosorption of phenolics and pesticides the pH

value for the highest biosorption was varied between acidicpH values. Unfortunately, the effect of pH on the biosorp-tion capacity is not always given in the literature. It shouldnot be forgotten that the neglect of a possible pH effect maylead to serious deviation of experimental results. In generalthe absorbed amount of organics per biomass quantity in-creased with increasing initial sorbate concentration and de-creased with increasing biomass concentration. The inves-tigators found that the uptake capacity is less than that ofactivated carbon, but it is high enough to provide significantremoval of trace quantities of toxic compounds combinedwith domestic and industrial pollutants.

Microbial adsorption is a promising alternative to replaceor supplement present treatment processes for the removalof very high concentrations of dyes and very low concentra-tions of phenolics and pesticides from the wastewater. Theuse of equilibrium and kinetic biosorptive data in associa-tion with plant operating parameters is useful in improvingthe understanding of the concentration profiles of hazardouspollutants in the final effluent and should be included inmodeling attempts. However, using fungal, yeast and bac-terial biomass to remove organic pollutants in a wastewateris still in the research stage. More studies are needed to de-velop a practical application.

References

[1] Clarke EA, Anliker R. Organic dyes and pigments. In: Handbookof environmental chemistry, anthropogenic compounds, vol. 3, partA. New York: Springer-Verlag, 1980. p. 181–215.

[2] Zollinger H. Azo dyes and pigments. Colour chemistry-synthesis,properties and applications of organic dyes and pigments. NewYork: VCH, 1987. p. 92–100.

[3] Mishra G, Tripathy M. A critical review of the treatment for de-colourization of textile effluent. Colourage 1993;40:35–8.

[4] Banat IM, Nigam P, Singh D, Marchant R. Microbial decolouriza-tion of textile-dye containing effluents: a review. Bioresour Technol1996;58:217–27.

[5] Fu Y, Viraraghavan T. Fungal decolourization of wastewaters: areview. Bioresour Technol 2001;79:251–62.

[6] Robinson T, Mcmullan G, Marchant R, Nigam P. Remediationof dyes in textile effluent: a critical review on current treat-ment technologies with a proposed alternative. Bioresour Technol2001;77:247–55.

[7] Hu TL. Sorption of reactive dyes byAeromonasbiomass. WaterSci Technol 1992;26:357–66.

[8] Juang R-S, Tseng R-L, Wu F-C, Lee S-H. Adsorption behaviour ofreactive dyes from aqueous solutions on chitosan. J Chem TechnolBiotechnol 1997;70:391–9.

[9] Karcher S, Kornmuller A, Jekel M. Removal of reactive dyesby sorption/complexation with cucurbituril. Water Sci Technol1999;40:425–33.

[10] Sumathi S, Manju BS. Uptake of reactive textile dyes byAspergillusfoetidus. Enzyme Microbial Technol 2000;27:347–52.

[11] Aksu Z, Tezer S. Equilibrium and kinetic modelling of biosorptionof Remazol Black B byR. arrhizus in a batch system: effect oftemperature. Process Biochem 2000;36:431–9.

[12] O’Mahony T, Guibal E, Tobin JM. Reactive dye biosorption byRhi-zopus arrhizusbiomass. Enzyme Microbial Technol 2002;31:456–63.


[13] Moran C, Hall ME, Howell RC. Effects of sewage treatment ontextile effluent. J Soc Dyers Colour 1997;113:272–4.

[14] Mittal AK, Gupta SK. Biosorption of cationic dyes by deadmacro fungusFomitopsis carnea: batch studies. Water Sci Technol1996;34:157–81.

[15] Chu HC, Chen KM. Reuse of activated sludge biomass: I. Re-moval of basic dyes from wastewater by biomass. Process Biochem2002;37:595–600.

[16] Fu Y, Viraraghavan T. Removal of Congo Red from an aque-ous solution by fungusAspergillus niger. Advances Environ Res2002;7:239–47.

[17] Gupta GS, Prasad G, Singh VH. Removal of chrome dye fromaqueous solutions by mixed adsorbents: fly ash and coal. WaterRes 1990;24:45–50.

[18] Slokar YM, Le Marechal AM. Methods of decolouration of textilewastewaters. Dyes Pigments 1997;37:335–56.

[19] El-Geundi MS. Colour removal from textile effluents by adsorptiontechniques. Water Res 1991;25:271–3.

[20] McKay G, Poots JP. Kinetics and diffusion processes in colourremoval from effluent using wood as an adsorbent. J Chem TechnolBiotechnol 1980;30:279–92.

[21] Lambert SD, Graham NJD, Sollars CJ, Fowler GD. Evaluation ofinorganic adsorbents for the removal of problematic textile dyesand pesticides. Water Sci Technol 1997;36:173–80.

[22] Low KS, Lee CK. Quaternized rice husk as sorbent for reactivedyes. Bioresour Technol 1997;61:121–5.

[23] Ramakrishna KR, Viraraghavan T. Dye removal using low costadsorbents. Water Sci Technol 1997;36:189–96.

[24] Lee CK, Low KS, Gan PY. Removal of some organic dyes byacid-treated spent bleaching earth. Process Biochem 1999;34:451–65.

[25] Morais LC, Freitas OM, Gonçalves EP, Vasconcelos LT, GonzalezBeça CG. Reactive dyes removal from wastewaters by adsorptionon eucalyptus bark: variables that define the process. Water Res1999;33:979–88.

[26] Ho YS, McKay G. Comparative sorption kinetic studies of dyeand aromatic compounds onto fly ash. J Environ Sci Health1999;A34:1179–204.

[27] Otero M, Rozada F, Calvo LF, Garcia AI, Moran A. Kinetic andequilibrium modelling of the methylene blue removal from solutionby adsorbent materials produced from sewage sludge. Biochem EngJ 2003;15:59–68.

[28] Razo-Flores E, Luijten M, Donlon B, Lettinga G, Field J. Biodegre-dation of selected azo dyes under methanogenic conditions. WaterSci Technol 1997;36:65–72.

[29] Manu B, Chaudhari S. Anaerobic decolourization of simu-lated textile wastewater containing azo dyes. Bioresour Technol2001;82:225–31.

[30] Glenn JK, Gold MH. Decolourization of several polymeric dyes bythe lignin degrading basidiomycete,Phanerochaete chrysosporium.Appl Environ Microbiol 1983;45:1741–7.

[31] Knapp JS, Newby PS. The decolourization of a chemical industryeffluent by white rot fungi. Water Res 1999;33:575–7.

[32] Kapdan IK, Kargi F, McMullan G, Marchant R. Effect of en-vironmental conditions on biological decolourization of textiledyestuff byC. versicolor. Enzyme Microbial Technol 2000;26:381–7.

[33] Meehan C, Banat IM, McMullan G, Nigam P, Smyth F, MarchantR. Decolourization of Remazol Black-using a thermotolerant yeast,Kluyveromyces marxianusIMB3. Environ Int 2000;26:75–9.

[34] Ramalho PA, Scholze H, Cardoso MH, Ramalho MT, Oliveira-Campos AM. Improved conditions for the aerobic reductive de-colourisation of azo dyes byCandida zeylanoides. Enzyme Micro-bial Technol 2002;31:848–54.

[35] Pearce CI, Lloyd JR, Guthrie JT. The removal of colour from textilewastewater using whole bacterial cells: a review. Dyes Pigments2003;58:179–96.

[36] Dönmez G. Bioaccumulation of the reactive textile dyes byCandidatropicalis growing in molasses medium. Enzyme Microbial Technol2002;30:363–6.

[37] Aksu Z. Reactive dye bioaccumulation bySaccharomyces cere-visiae. Process Biochem 2003;38:1437–44.

[38] Patterson JW. Wastewater treatment technology. USA: Ann ArborScience Pub. Inc.; 1977.

[39] Aksu Z, Yener J. A comparative adsorption/biosorption study ofmono-chlorinated phenols onto various sorbents. Waste Manage2001;21:695–702.

[40] Calace N, Nardi E, Petronio BM, Pietroletti M. Adsorption ofphenols by papermill sludges. Environ Pollut 2002;118:315–9.

[41] Rao JR, Viraraghavan T. Biosorption of phenol from an aque-ous solution by Aspergillus niger biomass. Bioresour Technol2002;85:165–71.

[42] Perrich JR. Activated carbon adsorption for waste water treatment.USA, Florida: CRS Press; 1981.

[43] Kennedy KJ, Lu J, Mohn WW. Biosorption of chlorophenols toanaerobic granular sludge. Water Res 1992;26:1085–92.

[44] Callega G, Serna J, Thirumaleswara SGB. Kinetics of adsorptionof phenolic compounds from wastewater onto activated carbon.Carbon 1993;31:691–7.

[45] Abdo MSE, Nosier SA, El-Tawil YA, Fadi SM, El-Khairy MI.Removal of phenol from aqueous solutions by mixed adsor-bents: Maghara coal and activated carbon. J Environ Sci Health1997;A32:1159–69.

[46] Streat M, Patrick JW, Camporro-Perez MJ. Sorption of phenol andp-chlorophenol from water using conventional and novel activatedcarbons. Water Sci Res 1995;29:467–72.

[47] Edgehill RU, Lu GQ. Adsorption characteristics of carbonized barkfor phenol and pentachlorophenol. J Chem Technol Biotechnol1998;71:27–34.

[48] Brasquet C, Roussy J, Subrenat E, Le Cloirec P. Adsorption andselectivity of activated carbon fibers application to organics. EnvironTechnol 1996;17:1245–52.

[49] Daifullah AAM, Girgis BS. Removal of some substituted phenolsby activated carbon obtained from agricultural waste. Water Res1998;32:1169–77.

[50] Gupta VK, Sharma S, Yadav IS, Mohan D. Utilization of bagassefly ash generated in the sugar industry for the removal of phenoland p-nitrophenol from wastewater. J Chem Technol Biotechnol1998;71:180–6.

[51] Chitra S, Chandrakasan G. Response of phenol degradingPseu-domonos pictorumto changing loads of phenolic compounds. JEnviron Sci Health 1996;A31:599–619.

[52] Garcia IG, Pena PRJ, Venceslada JLB, Martin AM, Santos MAM,Gomez ER. Removal of phenol from olive mill wastewater usingPhanerochaete chrysosporium, Aspergillus niger, Aspergillus ter-reus and Geotrichum candidum. Process Biochem 2000;35:751–8.

[53] Aksu Z, Bülbül G. Investigation of the combined effects of externalmass transfer and biodegradation rates on phenol removal usingimmobilized P. putida in a packed bed column reactor. EnzymeMicrobial Technol 1998;22:397–403.

[54] Dapaah SY, Hill GA. Biodegradation of chlorophenol mixtures byPseudomonas putida. Biotechnol Bioeng 1992;40:1353–8.

[55] Perez RR, Benito GG, Miranda MP. Chlorophenol degradationby Phanerochaete chrysosporium. Bioresour Technol 1997;60:207–13.

[56] Schwartz HG. Adsorption of selected pesticides on activated carbonand mineral surfaces. Environ Sci Technol 1967;1:332–7.

[57] Thacker NP, Vaidya MV, Sipani M, Kalra A. Removal technologyfor pesticide contaminants in potable water. J Environ Sci Health1997;B32:483–96.

[58] Gonzalez-Pradas E, Villafranca-Sanchez M, Gallege-Campo A,Urena-Amate D, Fernandez-Perez M. Removal of atrazine from


aqueous solution by natural and activated bentonite. J Environ Qual1997;26:1288–91.

[59] Kouras A, Zouboulis A, Samara C, Kouimtzis T. Removal ofpesticides from aqueous solutions by combined physicochemicalprocesses—the behaviour of lindane. Environ Pollut 1998;103:193–202.

[60] Gupta VK, Jain CK, Ali I, Chandra S, Agarwal S. Removal oflindane and malathion from wastewater using bagasse fly ash—asugar industry waste. Water Res 2002;36:2483–90.

[61] Sotelo JL, Ovejero G, Delgado JA, Martı̀nez I. Adsorption oflindane from water onto GAC: effect of carbon loading on kineticbehaviour. Chem Eng J 2002;87:111–20.

[62] Aksu Z, Kabasakal E. Batch adsorption of 2,4-dichlorophenoxy-acetic acid (2,4-D) from aqueous solution by granular activatedcarbon. Sep Purif Technol 2004;35:223–40.

[63] Shelton DR, Khader S, Karns JS, Pogell BM. Metabolism of twelveherbicides byStreptomyces. Biodegradation 1996;7:129–36.

[64] Volesky B. Detoxification of metal-bearing effluents: biosorptionfor the next century. Hydrometallurgy 2001;59:203–16.

[65] Huang C-P, Huang C-P, Morehart AL. The removal of Cu(II) fromdilute aqueous solutions bySaccharomyces cerevisiae. Water Res1990;24:433–9.

[66] Nourbakhsh M, Sag Y, Ozer D, Aksu Z, Kutsal T, Çaglar A. A com-parative study of various biosorbents for removal of chromium(VI)ions from industrial waste waters. Process Biochem 1994;29:1–5.

[67] Brady JM, Tobin JM. Adsorption of metal ions byRhizopus ar-rhizusbiomass: characterization studies. Enzyme Microbial Technol1994;16:671–5.

[68] Kapoor A, Viraraghavan T. Fungal biosorption—an alternative treat-ment option for heavy metal bearing wastewaters: a review. Biore-sour Technol 1995;53:195–206.

[69] Veglio F, Beolchini F. Removal of metals by biosorption: a review.Hydrometallurgy 1997;44:301–16.

[70] Puranik PR, Paknikar KM. Biosorption of lead and zinc fromsolutions usingStreptoverticillium cinnamoneumwaste biomass. JBiotechnol 1997;55:113–24.

[71] Aksu Z. Biosorption of heavy metals by microalgae in batch andcontinuous systems. In: Tam NFY, Wong Y-S (Eds.). Algae forwaste water treatment. Germany: Springer Verlag and Landes Bio-science; 1998. p. 37–53.

[72] Yetis U, Ozcengiz G, Dilek FB, Ergen N, Erbay A, Dilek A.Heavy metal biosorption by white-rot fungi. Water Sci Technol1998;38:323–30.

[73] Dönmez G, Aksu Z, Oztürk A, Kutsal T. A comparative studyon heavy metal biosorption characteristics of some algae. ProcessBiochem 1999;34:885–92.

[74] Yu Q, Matheickal JT, Yin P, Kaewsarn P. Heavy metal uptakecapacities of common marine macro algal biomass. Water Res1999;33:1534–7.

[75] Wong JPK, Wong YS, Tam NFY. Nickel biosorption by twochlorella species,C. vulgaris(a commercial species) andC. miniata(a local isolate). Bioresour Technol 2000;73:133–7.

[76] Say R, Denizli A, Arica MY. Biosorption of cadmium(II),lead(II) and copper(II) with the filamentous fungusPhanerochaetechrysosporium. Bioresour Technol 2001;76:67–70.

[77] Zhou JL, Banks CJ. Removal of humic acid fraction byRhizopusarrhizus: uptake and kinetic studies. Environ Technol 1991;12:859–69.

[78] Mou DG, Lim KK, Shen HP. Microbial agents for decolourizationof dye wastewater. Biotechnol Adv 1991;9:613–22.

[79] Banks CJ, Parkinson ME. The mechanism and application of fungalbiosorption to colour removal from raw water. J Chem TechnolBiotechnol 1992;54:192–6.

[80] Brahimi-Horn MC, Lim KK, Liany SL, Mou DG. Binding of textileazo dyes byMirothecium verrucariaOrange II, 10B (blue) and RS(red) azo dye uptake for textile wastewater decolourization. J IndMicrobiol 1992;10:245–61.

[81] Zhou W, Zimmermann W. Decolourization of industrial effluentscontaining reactive dyes by actinomycetes. FEMS Microbiol Lett1993;107:157–62.

[82] Zhou JL, Banks CJ. Mechanism of humic acid colour removalfrom natural waters by fungal biomass biosorption. Chemosphere1993;27:607–20.

[83] Hu T-L. Removal of reactive dyes from aqueous solution by differentbacterial genera. Water Sci Technol 1996;34:89–95.

[84] Polman JK, Breckenridge CR. Biomass-mediated binding and re-covery of textile dyes from waste effluents. Text Chem Colour1996;28:31–5.

[85] Gallagher KA, Healy MG, Allen SJ. Biosorption of synthetic dyeand metal ions from aqueous effluents using fungal biomass. In:Wise DL. (Ed.). Global Environmental Biotechnology. UK: Elsevier;1997. p. 27–50.

[86] Tatarko M, Bumpus JA. Biodegradation of Congo Red byPhane-rochaete chrysosporium. Water Res 1998;32:1713–7.

[87] Bustard M, McMullan G, McHale AP. Biosorption of textile dyes bybiomass derived fromKluveromyces marxianusIMB3. BioprocessEng 1998;19:427–30.

[88] Fu Y, Virarahavan T. Removal of a dye from an aqueous solution bythe fungusAspergillus niger. Water Quality Res J Can 2000;35:95–111.

[89] Fu Y, Virarahavan T. Removal of C.I. Acid Blue 29 from an aqueoussolution byAspergillus niger. AATCC Mag 2001;1:36–40.

[90] Fu Y, Viraraghavan T. Dye biosorption sites inAspergillus niger.Bioresour Technol 2002;82:139–45.

[91] Aksu Z. Biosorption of reactive dyes by dried activated sludge:equilibrium and kinetic modelling. Biochem Eng J 2001;7:79–84.

[92] Chu HC, Chen KM. Reuse of activated sludge biomass: II. Therate processes for the adsorption of basic dyes on biomass. ProcessBiochem 2002;37:1129–34.

[93] Aksu Z, Dönmez G. A comparative study on the biosorption char-acteristics of some yeasts for Remazol Blue reactive dye. Chemo-sphere 2003;50:1075–83.

[94] Basibuyuk M, Forster CF. An examination of the adsorption char-acteristics of a basic dye (Maxilon Red BL-N) on to live activatedsludge system. Process Biochem 2003;38:1311–6.

[95] Bell JP, Tsezos M. Removal of hazardous organic pollutants byadsorption on microbial biomass. Water Sci Technol 1987;19:409–16.

[96] Tsezos M, Bell JP. Comparison of the biosorption and desorptionof hazardous organic pollutants by live and dead biomass. WaterRes 1989;23:561–8.

[97] Logan BE, Alleman BC, Amy GL, Gilbertson RL. Adsorption andremoval of pentachlorophenol by white rot fungi in batch culture.Water Res 1994;28:1533–8.

[98] Jacobsen BN, Arvin E, Reinders M. Factors affecting sorptionof pentachlorophenol to suspended microbial biomass. Water Res1996;30:13–20.

[99] Brandt S, Zeng A-P, Deckwer W-D. Adsorption and desorptionof pentachlorophenol on cells ofMycobacterium chlorophenolicumPCP-1. Biotechnol Bioeng 1997;55:480–9.

[100] Daughney CJ, Fein JB. Sorption of 2,4,6-trichlorophenol byBacillussubtilis. Environ Sci Technol 1998;32:749–52.

[101] Aksu Z, Yener J. Investigation of the biosorption of phenol andmonochlorinated phenols on the dried activated sludge. ProcessBiochem 1998;33:49–655.

[102] Benoit P, Barriuso E, Calvet R. Biosorption characterization ofherbicides, 2,4-D and atrazine, and two chlorophenols on fungalmycelium. Chemosphere 1998;37:1271–82.

[103] Jianlong W, Yi Q, Horan N, Stentiford E. Bioadsorption of pen-tachlorophenol (PCP) from aqueous solution by activated sludgebiomass. Bioresour Technol 2000;75:157–61.

[104] Aksu Z, Akpinar D. Modelling of simultaneous biosorption ofphenol and nickel(II) onto dried aerobic activated sludge. Sep PurifTechnol 2000;21:87–99.


[105] Aksu Z, Akpinar D. Competitive biosorption of phenol andchromium(VI) from binary mixtures onto dried anaerobic activatedsludge. Biochem Eng J 2001;7:183–93.

[106] Wang W, Zhang X, Wang D. Adsorption ofp-chlorophenol bybiofilm components. Water Res 2002;36:551–60.

[107] Karim K, Gupta SK. Biosorption of nitrophenols on anaerobicgranular sludge. Environ Technol 2002;23:1379–84.

[108] Aksu Z, Gönen F. Biosorption of phenol by immobilized activatedsludge in a continuous packed bed: prediction of breakthroughcurves. Process Biochem 2004;39:599–613.

[109] Voerman S, Tammes PML. Adsorption and desorption of lindaneand dieldrin by yeast. Bull Environ Contam Toxicol 1969;45:271–7.

[110] Ju Y-H, Chen T-C, Liu JC. A study on the biosorption of lindane.Colloids Surf B 1997;9:187–96.

[111] Young E, Banks CJ. The removal of lindane from aqueous solutionusing a fungal biosorbent: the influence of pH, temperature, biomassconcentration and culture age. Environ Technol 1998;19:619–25.

[112] Lièvremont D, Seigle-murandi F, Benoit-guyod J-L. Removal ofPCNB from aqueous solution by a fungal adsorption process. WaterRes 1998;32:3601–6.

[113] Hong H-B, Hwang S-H, Chang Y-S. Biosorption of 1,2,3,4-tetrachlorodibenzo-p-dioxin and polychlorinated dibenzofurans byBacillus pumilus. Water Res 2000;34:349–53.

[114] Langmuir I. The adsorption of gases on plane surfaces of glass,mica, and platinum. J Am Chem Soc 1918;40:1361–8.

[115] Freundlich H. Adsorption in solution. Phys Chem Soc 1906;40:1361–8.

[116] Redlich OJ, Peterson DL. A useful adsorption isotherm. J PhysChem 1959;63:1024.

[117] Radke CJ, Prausnitz JM. Adsorption of organic solutions from diluteaqueous solution on activated carbon. Ind Eng Chem 1972;11:445–51.

[118] Weber Jr WJ. Adsorption. In: Physicochemical processes for waterquality control, New York: Wiley 1972. p. 206–11.

[119] Bellot JC, Condoret JS. Modelling of liquid chromatography equi-libria. Process Biochem 1993;28:365–76.

[120] Weber WJ, Morris JC. Kinetics of adsorption on carbon fromsolution. J Sanit Eng Div Am Soc Civ Eng 1963;89SA2:31–9.

[121] Lagergren S. Zur theorie der sogenannten adsorption gelöster stoffe.Kungliga Svenska Vetenskapsakademiens. Handlingar 1898;24:1–39.

[122] McKay G, Ho YS. Pseudo-second order model for sorption pro-cesses. Process Biochem 1999;34:451–65.

[123] Guibal E, Lorenzelli R, Vincent T, Le Cloirec P. Application ofsilica gel to metal ion sorption: Static and dynamic removal ofuranyl ions. Environ Technol 1995;16:101–14.

[124] Bohart G, Adams EQ. Some aspects of the behaviour of charcoalwith respect to chlorine. J Am Chem Soc 1920;42:523–44.

[125] Thomas HC. Heterogeneous ion exchange in a flowing system. JAm Chem Soc 1944;66:1664–6.

[126] Clark RM. Evaluating the cost and performance of field-scale granu-lar activated carbon systems. Environ Sci Technol 1987;21:573–80.

[127] Yoon YH, Nelson JH. Aplication of gas adsorption kinetics. I. Atheoretical model for respirator cartridge service time. Am Ind HygAssoc J 1984;45:509–16.

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