ARTICLE IN PRESS

Available at www.sciencedirect.com

WAT E R R E S E A R C H 4 1 ( 2 0 0 7 ) 1 3 6 6 – 1 3 7 8

0043-1354/$ - see frodoi:10.1016/j.watres

�Corresponding autE-mail address: a

journal homepage: www.elsevier.com/locate/watres

Removal and recovery of uranium from aqueous solutionsby Trichoderma harzianum

Kalsoom Akhtara, M. Waheed Akhtarb, Ahmad M. Khalidc,�

aBioprocess Technology Division, National Institute for Biotechnology and Genetic Engineering (NIBGE), P.O. Box 577,

Jhang Road, Faisalabad, PakistanbSchool of Biological Sciences, University of the Punjab, Lahore, PakistancDepartment of Chemistry, GC University Faisalabad, Allama Iqbal Road, Faisalabad, Pakistan

a r t i c l e i n f o

Article history:

Received 18 August 2006

Received in revised form

6 December 2006

Accepted 7 December 2006

Available online 8 February 2007

Keywords:

Uranium

Bioaccumulation

Biosorption

T. harzianum

Algae

Kinetic models

Pseudo-second-order kinetics

Multilayer

nt matter & 2007 Elsevie.2006.12.009

hor. Tel.: +92 41 9201371,mkhalid45@gmail.com (

a b s t r a c t

Removal and recovery of uranium from dilute aqueous solutions by indigenously isolated

viable and non-viable fungus (Trichoderma harzianum) and algae (RD256, RD257) was studied

by performing biosorption—desorption tests. Fungal strain was found comparatively better

candidate for uranium biosorption than algae. The process was highly pH dependent. At

optimized experimental parameters, the maximum uranium biosorption capacity of

T. harzianum was 612 mg U g�1 whereas maximum values of uranium biosorption capacity

exhibited by algal strains (RD256 and RD257) were 354 and 408 mg U g�1 and much higher in

comparison with commercially available resins (Dowex-SBR-P and IRA-400). Uranium

biosorption by algae followed Langmuir model while fungus exhibited a more complex

multilayer phenomenon of biosorption and followed pseudo-second-order kinetics. Mass

balance studies revealed that uranium recovery was 99.9%, for T. harzianum, and 97.1 and

95.3% for RD256 and RD257, respectively, by 0.1 M Hydrochloric acid which regenerated the

uranium-free cell biomass facilitating the sorption–desorption cycles for better economic

feasibility.

& 2007 Elsevier Ltd. All rights reserved.

1. Introduction

Biosorption started to gain importance in 1980s and since then

it has demonstrated a potential as an alternative technology for

removal and recovery of both toxic as well as precious metals

from waste water streams (Beolchini et al., 2006; Schiewer and

Volesky, 1995). Biosorption is the property of biomaterials to

bind and concentrate heavy metals/radio-nuclides from very

dilute aqueous solutions. It includes metal uptake by both,

active and passive modes (Kapoor et al., 1999). Passive mode of

metal uptake is termed as biosorption and is a ‘‘non-directional

physico-chemical interaction that may occur between metal/

radionuclide species and cellular compounds of biological

species’’ (Strandberg et al., 1981; Shumate and Strandberg,

r Ltd. All rights reserved.

+92 03008651900; fax: +92A.M. Khalid).

1985). This process requires no energy and is independent of

biological metabolism. This phenomenon has been reported for

a range of biomass types or viable or killed cells (Valdman et al.,

2001). During biosorption, microorganisms immobilize metals

by binding them to their cell walls. However, bioaccumulation

is carried out by the viable cells (Holan and Volesky, 1994) and

the metal uptake can also involve its passage into the cells

across the cell membranes. In living cells, both active and

passive modes may be involved during heavy metal uptake

(Kapoor et al., 1999). Among the major advantages of bioaccu-

mulation and biosorption are their economical nature, eco-

friendly behavior, regeneration of biosorbents for multiple uses,

possibility of selective metal recovery for recycling and there-

fore high efficiency of heavy metal removal from dilute metal

41 9200671.

ARTICLE IN PRESS

Nomenclature

b Langmuir constant

Ci initial metal ion concentration, mg/l

Ce equilibrium metal ion concentration, mg/l

Cf final metal ion concentration, mg/l

Cdes concentration of metal desorbed into the eluent,

mg/l

Es sorption energy, J/mole

k1 pseudo-first-order rate constant

k2 pseudo-second-order rate constant

KF freundlich adsorption capacity

K adsorption distribution coefficient,

mg metal g�1 biosorbent mg�1 ml�1

L ligands on the biosorbents

M metal ions

ML metal–ligand complex

n adsorption intensity

V volume of metal ion solution, L

W dry weight of biosorbent, g

q biosorption capacity, mg/g

qe amount of biosorption at equilibrium, mg/g

qdes eluted metal contents per gram of the biosorbent,

mg/g

qmax Maximum amount of metal ions per unit mass,

mg/g

qt amount of biosorption at time ‘‘t’’, mg/g

WAT E R R E S E A R C H 41 (2007) 1366– 1378 1367

solutions which are hazardous to the environment (Khalid

et al., 1993; Veglio and Beolchini, 1997).

Various types of microbial biomasses including both

heterotrophs (bacteria and fungi) and photoautotrophs (algae

and cyanobacteria) are reported to possess metal-binding

properties and have been studied as potential biosorbents for

selected metals (Tsezos et al., 1997). Most biosorption process

uses biomaterials, which are abundant in nature such as

marine algae, wastes coming from industrial and biological

processes such as fermentation and activated sludge and

activated charcoal (Sag et al., 2003; Nadeem et al., 2006). High

affinity, rapid rate of metal uptake and maximum loading

capacity are some of the important factors for the selecting of

a biosorbent. Therefore, there is an increased interest in the

identification of some new and better biosorbents that show

promising uptake of metallic ions.

Due to widespread applications of nuclear technologies,

there are many potential sources of uranium pollution.

Activities associated with the nuclear industry such as

uranium mining, reactor operations and fuel reprocessing

have brought excessive amounts of uranium into the

environment and this contamination has posed a serious

threat to the quality of surface and ground waters. Therefore,

removal of uranium from aqueous solution by biosorption/

bioaccumulation not only solves the contamination problem

but also makes economical recovery possible.

Current study was aimed to investigate the removal of

uranium by non-viable and viable biomass of fungus,

Trichoderma harzianum, and algae, RD256 and RD257 strains.

These microbial biomasses have been selected after screening

of a wide range of microbes. Although the removal of

radionuclide, uranium has been reported earlier, but to our

knowledge it is the first comprehensive report on its

accumulation by T. harzianum.

2. Materials and methods

2.1. Microorganisms and growth conditions

All microbial cultures used in screening studies were obtained

from NIBGE Culture Collection. Larger surface area of

suspended biomass was obtained by adding 10–15 glass beads

to each flask before autoclaving. Cultures were harvested

during the stationary growth phase by filtration through a

sterilized nylon cloth and washed thrice with distilled water.

The algal strains, RD256 and RD257, were isolated from

uranium mine drainage water samples obtained from Kabul

Kheel, Dera Ghazi Khan, Pakistan, by enrichment and were

separately cultured at 2572 1C under illumination at 1100 Lux

light intensity with a light/dark cycle of 16/8 h for 10 days. The

algal biomass was harvested by centrifugation at 7000 rpm for

20 min (4 1C), washed thrice with distilled water. Yeast strains

were maintained at YPD spread plates, YPD or LB medium at

pH 5.5 (Kuroda et al., 2001).

Wet biomass was determined after blotting the freshly

harvested biomass with commercial grade paper towels to

remove excess water. This wet biomass was stored in a screw-

capped bottle at 4 1C. A known weight from this wet biomass

was then dried at 80 1C in an oven for 24 h or to a constant

weight and factor to calculate dry weight from wet weight

was determined. This freshly harvested biomass stored at 4 1C

was used throughout the experiments for biosorption studies

unless otherwise mentioned.

2.2. Analytical determinations

The concentration of free uranium in the aqueous solution

before and after biosorption, was determined spectrophoto-

metrically (Bhatti et al., 1991). Stock solutions (1000 mg/l)

were prepared from analytical grade uranyl acetate dihydrate

(UO2(CH3COO)2. 2H2O) (BDH). All other chemicals used were

also of analytical grades.

2.3. Biosorption studies

Known amounts of biomass was added to metal solution in a

shaking flask and, final volume was made 100 ml yielding

metal concentrations as specified in each experiment. The

reaction mixtures was then agitated at 100–200 rpm at

2872 1C in an orbital rotary shaker. Periodically, 1.0 ml sample

was withdrawn, centrifuged at 9000g for 15 min and uranium

was determined in the supernatant. An average of triplicate

ARTICLE IN PRESS

WAT E R R E S E A R C H 4 1 ( 2 0 0 7 ) 1 3 6 6 – 1 3 7 81368

observations was reported. Metal ion (mg) biosorbed per gram

(dry weight) of biomass q (mg/g dry weight) was calculated

following Volesky and Holan, 1995:

q ¼ ðCi � Cf ÞV=W,

Ci, corresponded to the control sample, and Cf to the super-

natant solution when W (g) of biomass was suspended in V (l)

volume of the metal solution. Analysis of uranium solution in

control flasks (Ci) revealed that adsorption losses to the flask

walls were negligible.

2.4. Metal elution and regeneration of biomass

Different regenerating solutions like distilled water, HCl,

NaOH, Na2CO3, NaHCO3, (NH4)2SO4, (NH4)2CO3 and ethylene-

diaminetetra-acetic acid disodium salt (Na2EDTA) were tried

to release the accumulated metal. Biomass samples, (0.05 g)

were washed by distilled water and mixed with 30 ml of 0.1 M

regenerating solution and shaken for 3 h (or as mentioned

otherwise) at 100 rpm at 2872 1C. Pulp densities (solid to

liquid ratio) were optimized for maximum elution. All

experiments were carried out in triplicate.

The regenerated biomass was again washed with deionized

water for 3–4 times and the procedure was repeated for the

next de-sorption cycles which were carried out for five times.

The values of qdes were calculated from Cdes as follows:

qdes ¼ Cdes V=W.

The percentage of desorbed metal was evaluated as

Percentage desorption ¼ qdes=q� �

� 100.

Mass balance studies were performed by digesting 0.05 g

biomass with 10 ml of concentrated HCl and HNO3 (3:1);

biosorbed uranium was released and determined as described

earlier.

3. Results and discussion

3.1. Screening of microorganisms for biosorption studies

Twenty-five microbial species belonging to different types of

algae, fungi, and bacteria were screened for uranium biosorp-

tion potential at initial pH values of 3, 4 and 5 by incubating

freshly harvested wet biomass. The biomass harvested after

2, 3 and 10 days (for bacteria, fungi and algae, respectively)

were used and significant differences varying from

674–19075 mg g�1 were observed in biosorption capacity (q)

for uranium ions. These differences in q may be due to

morphological differences or stereo chemical changes that

could have occurred in the polysaccharide structures of cell

walls of these strains (Lo et al., 1999). Uranium uptake

capacities of the magnitudes 17173, 19075 and 17774 mg g�1

obtained for T. harzianum, RD256 and RD257, respectively,

were considerably higher than reported for other fungi, yeast

or bacteria. Although highest value for uranium accumula-

tion reported is with a Citrobacters sp. (8000 mg/g; Macaskie et

al., 1992) but it was due to enzymatic accumulation of

uranium. Therefore the values reported for uranium uptake

by T. harzianum, RD256 and RD257 in the present work have

suggested these biosorbents as the potential candidates for

uranium biosorption and recovery from the bacterial lea-

chates and industrial effluents.

Biosorption capacity values were higher with wet/viable

biomass as compared to dried/non-viable biomass of these

strains. The percentage decrease in biosorption capacities for

dried/non-viable biomass was found to be 18.872%,

19.771.6% and 23.273% for RD256, RD257 and T. harzianum,

respectively. These results are in good agreement with

observations made by Kapoor et al. (1999) who reported

higher values of nickel biosorption capacity by live Aspergillus

niger cells than the dead biomass. Similarly viable S. cerevisiae

sequestered three times more Cu2+ than dead cells (Volesky

and May-Phillips, 1995). The higher values of biosorption

capacity with wet/viable biomass may be due to uptake of

metals in addition to the physical adsorption. Although there

are reports attributing the decrease in biosorption capacity by

non-living biomass to cell surface modifications, which might

have occurred during drying process at high temperature

(Veglio and Beolchini, 1997) but Gonzalez-Munoz et al. (1997)

found that dead/dried cells of Myxococcus xanthus were able to

accumulated up to 2.4 mM of uranium/g and hence were more

efficient biosorbent than the wet/viable biomass.

Uranium biosorption capacities were also found to be

highly dependent on age of the cells used. Thus T. harzianum

gave maximum biosorption (19074 mg/g dry weight) when

harvested after 48–72 h inoculation. On the other hand,

uranium biosorption by algae RD256 and RD257 was found

to augment continuously with their culture age giving the

maximum values of biosorption capacities 19075 and 17774

for RD256 and RD257, respectively. For algal biosorbents, both

biosorption capacities and biomass yields increased with

culture age up to 240 h while S. cerevisiae used after 12 h

growth, remove 4.6 times more uranium from solution than

24 h old cells (Volesky and May-Phillips, 1995). The decrease in

sorption capacity with culture age may be due to change in

the chemical composition of cell wall as well as intracellular

components, since some biopolymers (i.e. chitosan) are

reported to be superior in metal uptake than others (i.e.

mannan and glucan) (Gonzalez-Munoz et al., 1997).

Storage of wet biomass for 4–5 days at 4 1C had no adverse

effects upon the biosorption capacity. Nevertheless, it was

found to decrease up to 25–37% when storage period

prolonged more than 60 days, after this time, though

biosorption capacity was not drastically affected, but the

physical appearance of biomass was changed, which was

more pronounced in case of T. harzianum as compared to

RD256 and RD257.

Uranium uptake by T. harzianum and RD256 increased

concomitantly with the shaking rate from stationary to

100 rpm, a further rise in shaking rate could not produce

any significant change. In case of RD257, uranium biosorption

maxima were attained at 200 rpm. The agitation seems to

improve the diffusion of metal ions on to the surface of

biosorbent.

3.2. Effect of pH on uranium biosorption

The biosorption capacity of microbial biomass for uranium

was strongly affected by the initial pH of aqueous metal

solution; it was found to increase with the pH, exhibiting a

ARTICLE IN PRESS

WAT E R R E S E A R C H 41 (2007) 1366– 1378 1369

maxima of 19674 mg/g at 4.5 for T. harzianum. A further

increase in pH adversely affected the biosorption capacity

(Fig. 1). Uranium sorption capacity by commercial resins

Dowex-SBR-P and IRA-400 was also investigated as a function

of pH. Both resins showed maximum sorption capacities of

16472 and 15775 mg/g dry weight at pH 2.0, above which

there was a sharp decline. It was interesting to observe that

these three biosorbents gave higher values of biosorption

capacity than commercial resins, indicating their superiority

over them. Uranium biosorption pH maxima were found to be

5 for algae, (RD256 and RD257) and 4.0 for T. harzianum which

are in accordance with earlier reports (Bhainsa and D’Souza,

1999; Gonzalez-Munoz et al., 1997; Khalid et al., 1993; Yang

and Volesky, 1999a). Studies could not be undertaken at pH

values higher than 5 as uranium precipitated at higher pH

values, confirming the earlier findings by Tsezos et al. (1997).

The pH of the medium affects both the solubility of metal

ions and ionization state of the functional groups (�COO�,

�OPO32� and �NH�) of the cell wall, which are acidic and carry

negative charges rendering cell walls to be potent scavenger

of cations. At low pH they will be protonated and are not so

freely available for the binding of metals. However, with an

increase in pH, the negative charge density on the cell surface

increases due to deprotonation of the metal-binding sites

thus increasing the attraction of metallic ions with negative

charge and allowing the biosorption on the cell surface.

Several researchers have investigated the effect of pH on

biosorption of heavy metals by using different kinds of

microbial biomass and have reported the similar behavior

(Holan and Volesky, 1994).

At low pH, uranyl ions mainly exist in the form of the

simple uranyl cations UO22+. At the same time, solution having

low pH (o2.5), increases the H3O+ concentration and inten-

sifies the competition between H3O+ and UO22+ for the binding

sites on the adsorbate surface. As the pH of the solution

increases, solubility of uranium decreases due to extensive

hydrolysis (the percentage of hydrolyzed ions UO2OH+,

(UO2)3(OH)5+ and (UO2)2(OH)2

2+ increases), hence adsorption

capacity increases with increasing pH, to a certain limit. The

210

170

130

90

50Sorp

tion

Cap

acity

q(m

g/g)

31 2

Initial p

Fig. 1 – Effect of initial solution pH on uranium ions uptake by (m

IRA-400 (W ¼ 0.05 g, Ci ¼ 100 mg/l, V ¼ 100 ml, temperature ¼ 28

competition between protons and metal ions for the same

carboxyl and sulfate binding sites has been described by a pH

sensitive biosorption isotherm model which allows the

prediction of pH effects on metal binding and the amount

of protons bound (Schiewer and Volesky, 1995). Solution pH

also has an effect on the hydrolysis of the metal-binding site

complex e.g. uranium–chitin complex on fungal walls (Tsezos

et al., 1997).

3.3. Effect of temperature on uranium biosorption

Uranium biosorption was enhanced with increments in

temperature from 10 to 30 1C (3972–5070.4%, Fig. 2): a further

rise in temperature from 30 to 70 1C, decreased the biosorp-

tion capacity (33.0–48.6%) for all the three biosorbents.

Comparison of biosorption capacities of commercial resins

Dowex-SBR-P and IRA-400 with biosorbents revealed that

sorptive potential of resins increased continuously with

temperature. Though the net percentage increase in sorption

capacity was more for commercial resins as compared to that

of microbial biosorbents, but still microbial biosorbents

showed higher values of biosorption potential. Many sug-

gested reasons for this temperature-dependent variation

in metal biosorption exist; such as, increase in uptake

at increased temperature may be due to either a higher

affinity of sites for the metal or an increase in binding sites on

the relevant biomass (Marques et al., 1991); and decrease

in uptake capacity with a decrease in temperature has

been suggested due to decrease mobility of potential

binding groups/moieties on the biosorbent surface (Bengtsson

et al., 1995).

3.4. Effect of contact time on biosorption

Viable biomasses of T. harzianum, RD256 and RD257 adsorbed

6073, 4274 and 3271 percentage of uranium in the initial

15 min from 100 mg/l uranium solution (Fig. 3A). However,

saturation levels were gradually attained after 24 h. Thus,

uranium biosorption is found to be a two stage process,

7654

H of Solution

) RD256 (K) RD257 (+) T. harzianum (.) Dowex-SBR-P and (’)

72 1C, agitation rate ¼ 200 rpm, contact time ¼ 24 h).

ARTICLE IN PRESS

200

175

150

125

100

75

Sorp

tion

Cap

acity

q(m

g/g)

Temperature (°C)

8060 7050403020100

RD256

RD257

T. harzianum

Dower-SBR-P

IRA-400

Fig. 2 – Effect of temperature on uranium biosorption (W ¼ 0.05 g, Ci ¼ 100 mg/l, V ¼ 100 ml, pH ¼ 5.0 for RD256 and RD257; 4.5

for T. harzianum and 2.0 for resins, agitation rate ¼ 200 rpm, contact time ¼ 24 h).

WAT E R R E S E A R C H 4 1 ( 2 0 0 7 ) 1 3 6 6 – 1 3 7 81370

consisting of an initial rapid passive binding of metals to

negatively charged sites on the cell walls, followed by a slower

active uptake of metal ions in to the cells. These results are in

good agreement with biosorption studies of various metals

studied with groups of microorganisms where higher rates of

metal binding have been reported (Gonzalez-Munoz et al.,

1997). The fast biosorption kinetics observed initially is

typical for biosorption process involving no energy-mediated

reactions and metal removal from solution is due to purely

physico-chemical interactions between biomass and metal

solution (Aksu, 2001). Microbial biosorbents significantly out

performed ion exchange resins IRA-400 and Dowex-SBR-P for

their uranium biosorption potential (Fig. 3A) confirming the

earlier reports for activated carbon, coal and ion exchange

resins (Nadeem et al., 2006).

During biosorption of uranium, pH profile was also

monitored and results have been presented in Fig. 3B. In case

of algae, no significant change in pH was noticeable. However,

when T. harzianum was used as biosorbent pH of solution

increased from 4.5 to 6.7. Different reasons could have been

forwarded for this phenomena viz., chitin removed hydrogen

ions from solution, resulting in an increase in the pH of bulk

solution and possibly in an even greater temporary increase

of micro spaces within the cell wall. Therefore, uranium

precipitated quickly within chitin micro spaces of the cell wall

filling substantial void space. An increase of pH could also be

the result of dissolution of some cytoplasmic components or

ions, such as carbonates, released into the solution. The

hypothesis of dissolution of cell components seems also to be

valid during present study, because of some difficulties of

filtration for solutions, which showed an increase in pH.

However, non-significant decrease in biosorption capacity

during sorption–desorption cyclic studies showed that dis-

solved components are not the major contributors towards

biosorption capacity.

The slight decrease in pH during biosorption of uranium by

RD256 and RD257 was due to ion exchange with protons. The

monovalent ions can have even higher affinity to the biomass

in ion exchange with protons because they could replace

single protons on separate binding sites in the biomass.

During ion exchange with protons, the ion exchange stoi-

chiometry would be U/H+¼ 1:2, 1:1, 3:1, 2:2 for UO2

2+, UO2OH+,

(UO2)3(OH)5+ and (UO2)2(OH)2

2+, respectively (Yang and Volesky,

1999b). Uranium biosorption by Rhizopus arrhizus resulted in

exchange of hydrogen ions from biomass for uranyl ions,

showing ion exchange as the principle of metal biosorption

(Treen-Sears et al.,1984).

3.5. Effect of biomass concentration on biosorption

The percentage of uranium removal from aqueous solution

was found to increase concomitantly with increments in

biomass concentration and percentage removal values of

95.1%, 91.3% and 97.5% for RD256, RD257 and T. harzianum,

respectively were attained at 0.5 g dry weight basis of

biosorbent/l of uranium solution. Further increase in biomass

concentration from 0.6 to 1.0 g/l was unable to produce

significant removal due to low metal concentration in

solution (Fourest and Roux, 1992). Biosorption capacity,

however, was increased by increasing biomass concentration

from 0.05 to 0.1 g/l, being maximum at this concentration.

The maximum values of biosorption capacities were 36273,

30873 and 54675 mg/g for RD256, RD257 and T. harzianum

(Fig. 4). Further increase in biomass concentrations lowered

these values to 9675, 9373 and 8376 mg/g for RD256, RD257

and T. harzianum, respectively. This decrease in specific

uptake values with increase in biomass concentration has

been explained by various researchers hypothesizing that

high biomass concentration leads to formation of cell

aggregates, thereby reducing the effective biosorption

area (Aksu, 2001), or an increase in biomass concentration

leads to interference between binding sites (Gadd and White,

1989). However, in our case, the decrease in specific uptake

value with an increase in biomass concentration due to

ARTICLE IN PRESS

210

170

130

90

50

10

Sorp

tion

Cap

acity

q(m

g/g)

pH d

urin

g B

ioso

rptio

n

50403020100

Contact Time (hours)

50403020100

Contact Time (hours)

7

6

5

4

3

2

1

A

B

Fig. 3 – Time course for uranium biosorption capacities (A) and pH changes (B) by (m) RD256, (K) RD257, (+) T. harzianum, (.)

Dowex-SBR-P and (’) IRA-400 (W ¼ 0.05 g, Ci ¼ 100 mg/l, V ¼ 100 ml, temperature ¼ 2872 1C, agitation rate ¼ 200 rpm).

WAT E R R E S E A R C H 41 (2007) 1366– 1378 1371

decrease in metal concentration in solution as evident

from corresponding percentage removal values (Fourest and

Roux, 1992). In fact, in the presence of a high biomass

concentration there is a very fast superficial adsorption on to

the microbial cells that produce a lower metal concentration

in solution than when the cell concentration is lower. Similar

results on the effect of biomass concentration on the

biosorption of metals have been reported for various micro-

organisms.

3.6. Effect of initial uranium concentration on biosorption

Effect of initial uranium ions concentration on its biosorption

by RD256, RD257 and T. harzianum was investigated by

incubating 0.05 g of biosorbents with 100 ml of uranium

solutions of concentrations ranging from 50 to 1100 mg/l. It

was found that the amount of uranium taken up by the cells

increased with an augmentation of uranium concentration

from 50–200 mg/l rapidly for all the three biosorbents used

(Fig. 5A). This increase was more rapid for algal biosorbents as

compared to fungal one. Relatively slow increase was

observed at concentrations greater than 200 mg/l for both

RD256 and RD257. For T. harzianum, the increase in biosorp-

tion capacities was rapid for whole range of concentrations

studied. The highest concentrations of uranium taken up by

RD256, RD257 and T. harzianum were 35476, 40975 and

61276 mg/g dry weight at equilibrium uranium concentration

of 82375, 79673 and 80877 mg/l, respectively. This increase

may be due to higher probability of collision between the

metal ions and biosorbent particles. At high equilibrium

concentrations, uptake of uranium owing to surface binding

was negligible due to saturation of biosorbent-binding sites.

This slow increase in biosorption capacity at higher equili-

brium concentrations could be related to the different

concentration gradient between the solution and the inside

of the microbial cells and an increase in uptake was due to

penetration of metal ions inside the cells rather than surface

adsorption. At very high solute level, solid–liquid equilibrium

becomes limited by diffusion of metal ions into the cell

(Fourest and Roux, 1992).

The adsorption distribution coefficient (K, ratio of the

equilibrium concentration in solid and aqueous phase)

ARTICLE IN PRESS

600

490

380

270

160

500.00 0.20 0.40 0.60 0.80 1.00 1.20

Bio

sorp

tion

Cap

acity

q(m

g/g)

Biomass Concentration (g/l)

% R

emov

al

100

80

60

40

20

0

Fig. 4 – Uranium uptake capacities (__________) and % removal (- - - -) by (m) RD256, (K) RD257 and (+) T. harzianum as a function

of biomass concentration (Ci ¼ 100 mg/l, V ¼ 100 ml, pH ¼ 4.5, temperature ¼ 2872 1C, agitation rate ¼ 100 rpm, contact

time ¼ 24 h).

WAT E R R E S E A R C H 4 1 ( 2 0 0 7 ) 1 3 6 6 – 1 3 7 81372

having unit of (mg metal g�1 biosorbent)/(mg metal ml�1 solu-

tion) or ml g�1 biosorbent was shown in Fig. 5B. A high value

of distribution coefficient is the characteristic of a good

biosorbent. T. harzianum, RD256 and RD257 exhibited max-

imum K values of 80,744, 38,020 and 20,738 ml g�1 dry weight

at Ce of 1.5, 5 and 8.8 mg U l�1 which decreased to 758, 430 and

514 ml g�1 at Ce of 808, 823 and 796 mg U l�1, respectively when

biomass concentration was 0.5 g l�1. The K value for uranium

bioaccumulation by T. harzianum (80,000 ml g�1) was about

eight times greater when compared to uranium distribution

coefficient value of Aspergillus fumigatus (10,000 ml g�1, Bhain-

sa and D’Souza, 1999). Since many industrial separation

processes utilized adsorbents with distribution coefficient as

small as 10 ml g�1 adsorbent, therefore T. harzianum, RD256

and RD257 appeared to be the potential microbial biosorbents

for uranium biosorption.

3.7. Equilibrium and kinetic models

3.7.1. Equilibrium modelsTo examine the relationship between sorbed (qe) and aqueous

concentrations (Ce) at equilibrium, the Langmuir and the

Freundlich adsorption isotherms are widely employed for

fitting the data.

For the fitting of experimental data, the linearized form of

Langmuir model is as follows:

1qe

¼1b

qmax

ðCe þ 1Þqmax

,

where qmax is the maximum amount of metal ions per unit

mass of biosorbent to form a complete monolayer on the

surface. The qe represents the practical limiting adsorption

capacity when the surface is fully covered with metal ions

and allows the comparison of adsorption performance,

particularly in the cases where the sorbent did not reach its

full saturation in experiments (Aksu, 2001). The plot of 1/qe vs.

1/Ce was employed to generate the intercept of 1/b. qmax and

the slope of 1/qmax. The value of qe for the construction of

sorption isotherms is determined as follows:

qe ¼ Ci � Ceð Þ=M.

The empirical Freundlich equation is

qe ¼ KF Ceð Þ1=n.

The above equation can be linearized by taking natural

logarithm as follows:

ln qe ¼ ln KF þ1n

ln Ce.

The Freundlich constants ‘‘KF’’ and ‘‘n’’ can be calculated

from intercept and slope of the straight line (obtained by

plotting ln qe vs. ln Ce), respectively. Langmuir and Freundlich

isotherms have been frequently used to fit experimental data

(Aksu, 2001; Holan and Volesky, 1994) during biosorption

studies of different metals by biomass. In the present studies,

Langmuir (Fig. 6A), Freundlich (Fig. 6B), and Dubinin–Radus-

kevich transformations (Fig. 6C), were found to be linear

(Fig. 6) and values of qmax as calculated from Langmuir model

were in good agreement with that of experimental values in

case of RD256 and RD257, while for T. harzianum theoretical

qmax was smaller than experimental value (Table 1). The

values of KF obtained from Freundlich isotherm were very low

as compared to experimental values for each of the three

biosorbents. While the values of 1/n are o1 suggesting

that biosorbents possess heterogeneous surface with iden-

tical adsorption energy in all sites and the biosorption

of uranium was limited to monolayer and the adsorbed metal

ion interacts only with the active site but not with

other. However this interpretation should be reviewed with

ARTICLE IN PRESS

650

550

450

350

250

150

50

90

80

70

60

50

40

30

20

10

0

Sorp

tion

Cap

acity

(m

g/g)

Dis

trib

utio

n C

oeff

icie

nt (

ml/g

)

(Tho

usan

ds)

9008007006005004003002001000

Equilibrium Concentration (mg/L)

9008007006005004003002001000

Equilibrium Concentration (mg/L)

A

B

Fig. 5 – Effect of initial uranium ion concentrations on (A) uranium biosorption and (B) distribution coefficient by (m) RD256,

(K) RD257 and (+) T. harzianum (0.05 g dry weight was incubated for 24 h with 100 ml of uranium solutions having different

concentrations at 200 rpm and 2872 1C).

WAT E R R E S E A R C H 41 (2007) 1366– 1378 1373

caution, as the biosorption and isotherm exhibit an irregular

pattern due to:

a)

Complex nature of biosorbent.

b)

Presence of varied multiple active sites on the biosorbent

surface.

c)

Change of metallic compounds chemistry in a solution.

Similarly the values of qmax obtained from Dubinin–Radus-

kevich isotherm were very high as compared to experimental

values for each of the three biosorbents. Therefore uranium

biosorption by RD256 and RD257 followed Langmuir model.

Data pertaining to uranium biosorption by A. fumigatus was

found to be best fit to Langmuir model of isotherm giving

maximum loading capacity of 423 mg U/g dry weight (Bhainsa

and D’Souza, 1999). The non-compatibility of Langmuir,

Freundlich and Dubinin–Raduskevich models in case of

T. harzianum substantiates a more complex multilayer

adsorption, indicating that the cell surface is a complex

heterogenous matrix containing an array of possible different

metal binding sites.

3.7.2. Kinetic modelsTo understand the controlling mechanism of biosorption,

kinetic models, pseudo-first-order and pseudo-second-order

were used to interpret the experimental data assuming that

measured concentrations are equal to cell surface concentra-

tions. Lagergren first-order rate equation is represented as

follows (Lagergren, 1898):

dqt=dt ¼ k1 qe � qt

� �.

Integration of above equation applying boundary condi-

tions, qt ¼ 0 at t ¼ 0 and qt ¼ qt at t ¼ t, resulted in

log qe=qe � qt

� �¼ k1tð Þ=2:303

or

log qe � qt

� �¼ log qe � k1tð Þ=2:303.

Plot of log (qe�qt) vs. t should give a straight line to confirm

the applicability of the kinetic model. Pseudo-second-order

ARTICLE IN PRESS

WAT E R R E S E A R C H 4 1 ( 2 0 0 7 ) 1 3 6 6 – 1 3 7 81374

equation can be expressed as

dqt=dt ¼ k2 qe � qt

� �2.

0.12

0.09

0.08

0.05

0.04

0.02

6.50

6.00

5.50

5.00

4.50

4.00

-5.00

-5.50

-6.00

-6.50

-7.00

-7.50

-8.00

0.680.510.340.170.00

I/Ce

I/qe

(E

-1)

In q

eI/

qe

InCe76543210

900750600450300150(Millions)

E2

C

B

A

Fig. 6 – Langmuir (A), Freundlich (B) and Dubinin-

Radushkevich (C) adsorption isotherms plots of uranium

ions biosorption to (m) RD256 (K) RD257 and (+) T.

harzianum, respectively.

Table 1 – Comparison of qmax obtained from Langmuir, Freunduranium biosorption

Adsobents(conc. g/l)

Experimentalqe (mg/g)

Langmuir

qmax b

RD256 35476 363.1 0.0121

RD257 40875 426.6 0.0053

T. harzianum 612 496.0 0.0173

pH ¼ 4.5 for T. harzianum and 5.0 for RD256 and RD257, biomass concent

Integrating and applying the boundry conditions leads to

1= qe � qt

� �� �¼ 1=qe

� �þ k2t

or

t=qt

� �¼ 1=k2q2

e

� �þ 1 þ qe

� �t.

Plot of t/qt vs. t should give a linear relationship. The k2 and

qe can be obtained from the intercept and slope, respectively.

Applications of pseudo-first-order and pseudo-second-or-

der equations to uranium biosorption by T. harzianum, RD256,

D257, Dowex SBR-P and IRA-400 suggested that uranium

biosorption by T. harzianum, RD256 and RD257 followed the

pseudo-second-order rather than pseudo-first-order kinetics

(Fig. 7). The values of qe obtained from pseudo-second-order

kinetic model were in close agreement with that of experi-

mental values (Table 2). On the other hand the value of qe

obtained from pseudo-first-order kinetic were too small as

compared to the experimental values although correlation

coefficient were in the range of 0.92–0.98 and these low values

of qe suggested the insufficiency of pseudo-first-order to fit

the kinetic data. Since even, if the values of equilibrium

uptake calculated from kinetic plots does not equal to the

experimental equilibrium uptake then the reaction is not

likely to follow the particular kinetic model even though this

plot may has high correlation coefficient with the experi-

mental data. Ho (2006) has proposed a pseudo-second-order

kinetic model for the sorption of cadmium onto tree fern.

Application of Weber–Morris equation to kinetic data

revealed that uranium biosorption did not follow this

equation (Fig. 7C), although straight lines were obtained but

these straight lines were not passing through origin, which is

the prerequisite of this model.

3.8. Uranium elution and regeneration of biosorbents

Regeneration of uranium-loaded biomasses was attempted

using various potential eluting agents. The elution efficiency

of hydrochloric acid was 93.9%, 87.5% and 85.6% from T.

harzianum, RD256 and RD257, respectively whereas distilled

water, ammonium sulfate and NaEDTA gave minimum

elution efficiency ranging from 4.2% to 6.5%. The percentage

elution by sodium carbonate after 3 h of incubation was 95.6,

76.4 and 80.8 while with sodium bicarbonate these values

were in the order of 88.5, 80.5 and 95.4% for T. harzianum,

RD256 and RD257, respectively (Table 3). Yang and Volesky

lich and Dubinin–Radushkevich adsorption isotherms for

Freundlich Dubinin Radushkevich

KF n qmax Es (104)

109.0 2.26 557.9 1.51

086.1 1.73 739.5 1.63

120.0 4.22 863.5 1.40

ration ¼ 0.5 g/l, agitation rate ¼ 200 rpm, temperature ¼ 2872 1C

ARTICLE IN PRESS

2.50

2.10

1.70

1.30

0.90

0.50

0.16

0.13

0.10

0.06

0.03

0.00

200

170

140

110

80

50

20

252010 1550

Contact Time (hours)

252010 1550

Contact Time (hours)

542 310

t1/2

log(

qe-q

t)t/

qtqt

C

B

A

Fig. 7 – Pseudo-first-order (A), pseudo-second-order (B) and

Weber–Morris (C) plots of uranium ions biosorption to (m)

RD256, (K) RD257, (+) T. harzianum, (m) Dowex-SBR-P and

(’) IRA-400.

Table 2 – Comparison between adsorption rate constants andpseudo-second-order kinetic models for uranium biosorption

Biosorbents First-order kinetic

k1,ads (min�1) qe (mg/g)

RD256 0.114 77.8

RD257 0.141 100.7

T. harzianum 0.123 49.3

Dowex-SBR-P 0.221 98.3

IRA-400 0.187 113.4

Ci ¼ 111 mg/l, pH ¼ 2.0 for resins; 4.5 for T. harzianum and 5.0 for RD256 a

temperature ¼ 2872 1C.

WAT E R R E S E A R C H 41 (2007) 1366– 1378 1375

(1999a) used 0.1 N hydrochloric acid for regeneration of

Sargassum seaweed biomass column and obtained a very

narrow peak of elution curve reflecting a high efficiency in

uranium recovery. Gonzalez-Munoz et al. (1997) found 0.1 M

sodium carbonate as efficient desorbent of uranium (82%)

from Myxococuss xanthus biomass.

The percentage decrease in dry weight of biosorbents after

elution with various desorbents is given in Table 3. The

maximum decrease in dry weight was resulted from sodium

carbonate and sodium bicarbonate elution. The maximum

decrease was 52.5%. Hydrochloric acid caused 10.9%, 14.7%

and 12.9% decrease in dry weight of T. harzianum, RD256 and

RD257, respectively.

3.8.1. Effect of pulp density on maximum elution efficiencyAn important parameter for metal biosorption is the solid-to-

liquid ratio (S/L) defined as the mass of metal-laden biosor-

bent to the volume of the elutant. It is desirable to use the

smallest possible eluting volume to contain the highest

concentration of the metal but at the same time, the volume

of the solution should be enough to provide maximum

solubility for the metal desorbed. Studies were conducted to

optimize pulp density to obtain maximum uranium elution

efficiency of 0.1 M HCl for T. harzianum, RD256 and RD257,

(Fig. 8). The 0.02, 0.04, 0.05, 0.06 and 0.08 g of uranium-loaded

biomass was incubated with 50 ml of eluting solution. It was

found that percentage elution for all the three biosorbents

remained almost same with an increase in pulp density up to

1.0 g/l of eluent. Maximum elution efficiency obtained after

4 h of incubation at solid/liquid of 1.0 g of uranium-loaded

biomass/l of eluent was 99.8, 97.1 and 95.3% for T. harzianum,

RD256 and RD257, respectively. Although above this pulp

density, values of elution efficiency decreased but still

remained at 95.5%, 89.9% and 86.1% for T. harzianum, RD256

and RD257, respectively even at 1.6 g/l pulp density value.

After optimizing S/L conditions, almost 96–99% recovery

was achieved with 0.1 M hydrochloric acid (Fig. 8). High

elution efficiency and low biomass damage favors hydro-

chloric acid to be employed as a suitable uranium-desorbing

qe estimated to the Lagergren pseudo-first-order and

Second-order kinetic Experimental

k2,ads (g/mg min)

qe (mg/g) qe (mg/g)

0.0062 191.9 19173

0.0042 181.8 17975

0.0051 195.3 19572

0.0053 171.5 16578

0.0032 169.8 15773

nd RD257, biomass concentration ¼ 0.5 g/l, agitation rate ¼ 200 rpm,

ARTICLE IN PRESS

Table 3 – Desorption of uranium from loaded biomass by various desorbents

Desorbents pH % elution after 3 h % decrease in dry weight after elution

T. h RD256 RD257 T. h RD256 RD257

Distilled water 05.04 06.1 04.0 04.2 NC NC NC

HCl 02.12 93.9 87.5 85.6 10.9 14.5 12.9

NaOH 12.40 40.5 22.8 25.6 19.7 13.5 11.2

Na2CO3 10.98 95.6 76.4 80.8 24.9 47.0 44.5

NaHCO3 08.65 88.0 80.5 95.4 23.5 52.5 51.0

(NH4)2SO4 05.62 06.5 04.4 05.4 NC 01.5 NC

(NH4)2CO3 08.31 51.0 78.8 94.2 04.5 01.5 03.0

NaEDTA 04.68 04.5 03.7 05.1 NC NC NC

Results are average of 3 readings with RSDo5; 0.05 g uranium loaded biomass was suspended in 30 ml of desorbent; Concentration of

desorbents: 0.1 M; NC: No change observed; T. h: T. harzianum.

100

95

90

85

% E

lutio

n

80

75

700.4 0.8 1.0 1.2 1.6

RD256

RD257

T. harzianum

Fig. 8 – Effect of pulp density on elution of uranium by 0.1 M

hydrochloric acid after 4 h of incubation at 100 rpm and

2872 1C.

WAT E R R E S E A R C H 4 1 ( 2 0 0 7 ) 1 3 6 6 – 1 3 7 81376

agent for T. harzianum, RD256 and RD257. Another justifica-

tion for preferring hydrochloric acid as eluent is the fact that

it causes no harm to Ca-alginate beads during immobilization

studies whereas these beads get completely dissolved in

sodium carbonate solution.

3.8.2. Sorption/desorption cyclic studiesFor cyclic studies, biomass at concentration of 0.5 g/l of

uranium solution (concentration 100 mg/l) was used for

sorption studies. Elution studies were carried out using

1.0 g/l pulp density. 0.1 M hydrochloric acid was used as

eluent for this sorption/desorption cyclic studies. Both

sorption capacity and elution efficiency were measured after

24 h of incubation at 2872 1C and 100 rpm (Fig. 9). In first

cycle, values of sorption capacity were 190.5, 175.9 and

193.9 mg/g that decreased to 119.7, 149.5 and 171.5 for

RD256, RD257 and T. harzianum, respectively. The order for

decrease in sorption capacity in subsequent five cycles was

RD2564RD2574T. harzianum. The values for elution efficiency

were 92.2%, 97.1% and 99.6% in first cycle that decreased to

85.9%, 90.5% and 98.6% for RD256, RD257 and T. harzianum,

respectively at fifth cycle. T. harzianum showed better

performance in sorption–desorption cyclic studies as

compared to RD256 and RD257 (Table 4). The decrease in

sorption capacity values was found to be 37.3%, 18.3% and

12.1% for RD256, RD257 and T. harzianum, respectively.

Therefore, T. harzianum could be exploited commercially for

the treatment of effluent streams containing uranium in

repeated cycles.

3.9. Mass balance studies

To prove that decrease in concentrations of uranium ions in

solutions during biosorption experiments were really due to

biosorption, mass balance studies were conducted. Table 5

summarizes the comparison of uranium biosorbed and

released after digestion. In case of uranium biosorption, upto

500 mg/l initial concentration of solution, there was no

significant difference between both values for all three

biosorbents (o10 mg/g biomass). At initial concentrations of

800 and 1000 mg/l, the amount of metal ions released after

digestion were less than the amount of ions biosorbed and

values of this difference were in the range of 20–30 and

30–48 mg/l for 800 and 1000 mg/l of initial uranium concen-

trations, respectively. At both these concentrations, the

values of relative standard deviation were also high relative

to the values at low initial concentrations (50–500 mg/l). One

explanation for this large difference may be that at high

initial concentration, some metal ions got precipitated in

solution and apparently this seems to be contributing

towards biosorption but actually it is not doing so. During

washing of biomass before acid digestion, these precipitates

washed away or metal ions not adsorbed strongly detached

from biomass during washing and metal concentration

obtained after digestion was only due to biosorption or

bioaccumulation.

ARTICLE IN PRESS

200

180

160

140

120

100

100

80

60

40

20

0

Sorp

tion

Cap

acity

(m

g/g)

% E

lutio

n

Number of Cycles

4 5321

RD256 (q mg/g)

RD257 (q mg/g)

RD256 (%Elu)

RD257 (%Elu)

T. h (q mg/g)

T. h (%Elu)

Fig. 9 – Batch sorption–desorption cyclic studies for uranium biosorption. Pulp density 0.5 g/l, concentration 100 mg/l, pH 4.5

for T. harzianum and 5.5 for RD256 and RD257 Elution studies: pulp density 1.0 g/l, incubated for 4 h at 100 rpm and 2872 1C.

Table 4 – Percentage decrease in uranium sorptioncapacities after five subsequent sorption-desorptioncycle

Biosorbents Sorption capacity(mg/g)

% decreasein sorption

capacity1st 5th

RD256 190.5 119.5 37.5

RD257 175.9 149.5 15.0

T. harzianum 193.9 171.5 11.6

Table 5 – Mass Balance of Uranium absorbed anddesorbed

Biosorbent Initialsolution

conc. (mg/l)

Uraniumbiosorbed

(mg/g)

Uraniumreleased afteracid digestion

(mg/g)

RD256 50 9373 9273

100 19075 19174

200 30877 30176

500 33875 33274

800 346714 316714

1000 353716 318715

RD257 50 8976 9174

100 18373 18075

200 31573 31374.

500 36875 37072

800 397715 379718

1000 409714 376721

T. harzianum 50 9373 9173

100 19575 19474

WAT E R R E S E A R C H 41 (2007) 1366– 1378 1377

4. Conclusions

This work describes suitable microorganisms which can

efficiently remove uranium from aqueous solutions under

optimized environmental conditions of pH, temperature,

biomass size, contact time.

200 34675 34374

400 42775 42476

� 800 549719 526715

1000 616713 568718

Pulp density 0.5 g/l, pH 4.5 (T. harzianum) and 5.0 (RD256, RD257).

Among algae and fungi, fungus Trichoderma harzianum was

found better uranium biosorbent giving biosorption capa-

city of 612 mg U g�1 which when compared with commer-

cially used resins, was significantly higher.

�

When equilibrium data in batch mode were fitted,

uranium biosorption by Trichoderma harzianum occurred

via complex multi-layer phenomenon following pseudo-

second-order kinetics. Algal biomass, however, exhibited

simple monolayer biosorption pattern as determined by

classical Langmuir model.

�

Recoveries of sorbed uranium could be obtained upto

99.9% by desorbing by HCl, which was capable of

regenerating the biomass, which could be used for five

cycles.

No doubt that earlier studies exist where removal of

uranium/radionuclide have been described, but this is the

first comprehensive report on removal and recovery of this

metal by Trichoderma harzianum.

Acknowledgments

A.M.K. gratefully acknowledges the financial support for this

project from IAEA research contract no.8617/RB. The work

presented here is a part of Ph.D. dissertation of KA. Technical

ARTICLE IN PRESS

WAT E R R E S E A R C H 4 1 ( 2 0 0 7 ) 1 3 6 6 – 1 3 7 81378

assistance of Messers Faqir Muhammad and Habib Shah are

also acknowledged.

R E F E R E N C E S

Aksu, Z., 2001. Equilibrium and kinetic modeling of cadmium (II)biosorption by C. vulgaris in a batch system: effect oftemperature. Separat. Purificat.Technol. 21, 85–294.

Bengtsson, L., Johansson, B., Hacket, T.J., McHale, L., McHale, A.P.,1995. Studies on the biosorption of uranium by Talaromycesemersonii CBS 814.70 biomass. Appl. Microbiol. Biotechnol. 42,807–811.

Beolchini, F., Pagnanelli, F., Toro, L., Veglio, F., 2006. Ionic strengtheffect on copper biosorption by Sphaelotilus natans: equilibriumstudy and dynamic modeling in membrane reactor. Water Res.40, 144–152.

Bhainsa, K.C., D’Souza, S.F., 1999. Biosorption of uranium (VI) byAspergillus fumigatus. Biotechnol. Tech. 13, 695–699.

Bhatti, T.M., Mateen, A., Amin, M., Malik, K.A., Khalid, A.M., 1991.Spectrophotometric determination of uranium (VI) in bacterialleach liquors using arsenazo-III. J. Chem. Technol. Biotechnol.55, 331–341.

Fourest, E., Roux, J.C., 1992. Heavy metal biosorption by fungalmycelial by-products: mechanism and influence of pH. Appl.Microbiol. Biotechnol. 37, 399–403.

Gadd, G.M., White, C., 1989. Removal of thorium from simulatedacid process stream by fungal biomass. Biotechnol. Bioeng. 33,592–597.

Gonzalez-Munoz, M.T., Merroum, M.L., Omar, N.B., Arias, J.M.,1997. Biosorption of uranium by Myxococcus xanthus. Int.Biodet. Biodeg. 40, 107–114.

Ho, S.H., 2006. Second-order kinetic model for the sorption ofcadmium onto tree fern: a comparison of linear and non-linear methods. Water Res. 40, 119–125.

Holan, Z.R., Volesky, B., 1994. Biosorption of lead and nickel bybiomass of marine algae. Biotechnol. Bioeng. 43, 1001–1009.

Kapoor, A., Viraraghavan, T., Cullimore, D.R., 1999. Removal ofheavy metals using fungus Aspergillus niger. Biores. Technol.70, 95–104.

Khalid, A.M., Shemsi, A.M., Akhtar, K., Anwar, M.A., 1993. Uraniumbiosorption by Tricchoderma harzianum entrapped in polyesterfoam beads. In: Torma, A.E., Wey, J.E., Lakkshmanan, V.I. (Eds.),Biohydrometalurgical Technologies, Fossil, Energy, Materials,Bioremediation, Microbial Physiology, vol. II. The Minerals,Metals and Materials Society, Pennysylvania, USA, pp. 309–317.

Kuroda, K., Shibasaki, S., Ueda, M., Tanaka, A., 2001. Cell surface-engineered yeast displaying a histidine oligopeptide (hex-His)has enhanced adsorption of and tolerance to heavy metalions. Appl. Microbiol. Biotechnol. 57 (5), 697–701.

Lagergren, S., 1898. About the theory of so-called adsorption ofsoluble substances, Kungliga svenska Vetenskapsakademiens.Handlingar 24 (4), 1–39.

Lo, W., Chua, H., Lam, K.H., Bi, S.P., 1999. A comparativeinvestigations on the biosorption of lead by filamentousfungal biomass. Chemosphere 39 (15), 2723–2736.

Macaskie, L.E., Empson, R.M., Cheetham, A.K., Grey, C.P., Skarnu-lis, A.J., 1992. Uranium bioaccumulation by a Citrobacter sp. asa result of enzymically mediated growth of polycrystallineHUO2PO4. Science 252, 782–784.

Marques, A.M., Roca, X., Simon-Pujol, M.D., Fuste, M.C., Con-gregado, F., 1991. Uranium accumulation by Pseudomonas sp.EPS-5028. Appl. Microbiol. Biotechnol. 35, 406–410.

Nadeem, M., Mahmood, A., Shahid, S.A., Shah, S.H., Khalid, A.M.,Mckay, G., 2006. Sorption of lead from aqueous solutions bychemically modified carbon adsorbents. J. Hazard. Mater. 138(3), 604–613.

Sag, Y., Tatar, B., Kutsal, T., 2003. Biosorption of Pb (II) and Cu (II)by activated sludge in batch and continuous-flow stirredreactors. Biores. Technol. 87, 27–33.

Schiewer, S., Volesky, B., 1995. Modeling of the proton–metal ionexchange in biosorption. Environ. Sci. Technol. 29, 3049–3058.

Shumate II, S.E., Strandberg, G.W., 1985. Accumulation of metalsby microbial cells. In: Moo, Y.M., Robinson, C.N., Howell, J.A.(Eds.), Comprehensive Biotechnology, vol. 4. Pergamon Press,New York, pp. 235–247.

Strandberg, G., Shumate, S.E., Parrott, J.R., 1981. Microbial cells asbiosorbents for heavy metals: accumulation of uranium bySaccharomyces cerevisiae and Pseudomonas aeruginosa. Appl.Environ. Microbiol. 41, 237–245.

Treen-Sears, M.E., Volesky, B., Neufeld, R.J., 1984. Ion exchange/complexation of the uranyl ions by Rhizopus biosorbents.Biotechnol. Bioeng. 26, 1323–1329.

Tsezos, M., Georgousis, Z., Remoudaki, E., 1997. Mechanism ofaluminium interference on uranium biosorption by Rhizopusarrhizus. Biotechnol. Bioeng. 55 (1), 16–27.

Valdman, E., Erijman, L., Pessoa, F.L.P., Leite, S.G.F., 2001.Continuous biosorption of Cu and Zn by immobilized wastebiomass Sargassum sp. Process Biochem 36, 869–873.

Veglio, F., Beolchini, F., 1997. Removal of metals by biosorption: areview. Hydrometalurgy 44, 301–319.

Volesky, B., Holan, Z.R., 1995. Biosorption of immobilized Sacchar-omyces cerevisiae with successive copper adsorption–desorp-tion cycles. Biotechnol. Lett. 18, 531–536.

Volesky, B., May-Phillips, H.A., 1995. Biosorption of heavy metalsby Saccharomyces cerevisiae. Appl. Microbiol. Biotechnol. 42,797–806.

Yang, J., Volesky, B., 1999a. Biosorption of uranium on Sargassumbiomass. Water Res. 33 (15), 3357–3363.

Yang, J., Volesky, B., 1999b. Modelling uranium–proton ionexchange in biosorption. Environ. Sci. Technol. 33, 4079–4085.