REVIEW

Insights into the interactions of cyanobacteria with uranium

Celin Acharya • Shree Kumar Apte

Received: 5 April 2013 / Accepted: 23 September 2013 / Published online: 8 October 2013

� Springer Science+Business Media Dordrecht 2013

Abstract Due to various activities associated with nuclear

industry, uranium is migrated to aquatic environments like

groundwater, ponds or oceans. Uranium forms stable car-

bonate complexes in the oxic waters of pH 7–10 which

results in a high degree of uranium mobility. Microorgan-

isms employ various mechanisms which significantly influ-

ence the mobility and the speciation of uranium in aquatic

environments. Uranyl bioremediation studies, this far, have

generally focussed on low pH conditions and related to

adsorption of positively charged UO22? onto negatively

charged microbial surfaces. Sequestration of anionic ura-

nium species, i.e. [UO2(CO3)22-] and [UO2(CO3)3

4-] onto

microbial surfaces has received only scant attention. Marine

cyanobacteria are effective metal adsorbents and represent

an important sink for metals in aquatic environment. This

article addresses the cyanobacterial interactions with toxic

metals in general while stressing on uranium. It focusses on

the possible mechanisms employed by cyanobacteria to

sequester uranium from aqueous solutions above circum-

neutral pH where negatively charged uranyl carbonate

complexes dominate aqueous uranium speciation. The

mechanisms demonstrated by cyanobacteria are important

components of biogeochemical cycle of uranium and are

useful for the development of appropriate strategies, either

to recover or remediate uranium from the aquatic

environments.

Keywords Cyanobacteria � Uranium � Interaction

mechanisms � Bioremediation � Biorecovery

Introduction

Increasing contamination of the environment by uranium

on account of its mining and disposal of tailings, nuclear

power/weapons production, nuclear testing or nuclear

accidents is a worldwide problem. Microbial interactions

with metals form an important part of the natural biogeo-

chemical processes and have important consequences for

human society. It is therefore vital to advance our under-

standing of the metal–microbe interactions to develop

suitable bioremediation strategies for metal-contaminated

sites. The versatility of microbial systems to remove heavy

metals and radionuclides from their immediate environ-

ment is well recognized. Microbes influence the environ-

mental fate of metals by employing diverse physico-

chemical and biological mechanisms, effecting changes in

the mobility and speciation of metals. As a consequence,

microbial bioremediation of radionuclide pollutants is

being actively explored currently.

Cyanobacteria represent a morphologically diverse group

of oxygenic, gram-negative photosynthetic prokaryotes,

which are widely distributed in freshwater, marine and ter-

restrial environments (Fogg et al. 1973). These organisms

respond and adapt to most stress conditions and are often

abundant in metal-contaminated environments (Kanamaru

et al. 1994). They can tolerate, accumulate and detoxify

metal contaminants in aquatic environments, where they

abound, thereby affecting the mobility and bioavailability of

metals (Li et al. 2004; Zhou et al. 2004; Gardea Torresdey

et al. 1998). Their mass cultivation is economic and feasible

which qualifies them as suitable bioremediation agents for

C. Acharya � S. K. Apte (&)

Molecular Biology Division, Bhabha Atomic Research Centre,

Trombay, Mumbai 400 085, India

e-mail: aptesk@barc.gov.in; sksmbd@barc.gov.in

C. Acharya

e-mail: celin@barc.gov.in

123

Photosynth Res (2013) 118:83–94

DOI 10.1007/s11120-013-9928-9

the recovery and recycling of target metals (Garcia-Meza

et al. 2005; Wang et al. 2005).

This article reviews cyanobacterial interactions with the

toxic metals, focussing primarily on uranium. The importance

of bacteria, microalgae, yeast and fungi has received most

attention in uranium bioremediation, while cyanobacterial

interactions with uranium are rather poorly understood and

need to be addressed more vigorously. The aquatic environ-

ments directly or indirectly receive uranium contamination.

Understanding the basic behaviour of cyanobacteria, repre-

senting an important sink for metals in aquatic environment,

is expected to aid the development of bioremediation strate-

gies for U-contaminated aquatic environments.

Uranium speciation, toxicity and microbial interactions

in aquatic systems

Uranium exists primarily as U(VI) in oxic aqueous systems

in the form of free divalent oxocomplex, UO22? at pH B 5.

The ability of any metal to bind to or traverse across the

cell surface of organisms depends on its speciation

(Markich 2002). Aqueous speciation of uranium undergoes

major changes within a pH range of 5–7 because of com-

plexation with the carbonates and hydroxides (Choppin

2007). In waters at pH 7–10, the soluble carbonate com-

plexes of UO22?, i.e. [UO2(CO3)2

2-] and [UO2(CO3)34-],

are the predominant anionic species, with the latter being

more dominant at higher pH values (Aleissa et al. 2004).

The aforesaid carbonate species are the more probable

uranium species in water because phosphates are generally

not detected in the pond or seawater.

Uranium contamination in surface, ground or natural

waters (ponds, lakes, sea water), resulting from activities

such as mining, storage of radioactive waste, nuclear energy

production, is a subject of intense public concern. Uranium

has no biological function and is known to be toxic to humans

(Cothern and Lappenbusch 1983) and microbes (Plummer

and Macaskie 1990; Suzuki and Banfield 2004). The chem-

ical toxicity studies for uranium have reported inhibition of

growth of aquatic microflora including algae, cyanobacteria

and other aquatic microorganisms at 1 mg U L-1 in fresh-

water systems, and bactericidal activity of this radionuclide

at a concentration of 100 mg L-1 (Driver 1994). Uranium

concentrations in aquatic systems range from 30 lg L-1 in

the fresh surface or groundwater, 0.01–6.6 lg L-1 in river

water, to 3 lg L-1 in sea water (Markich 2002). These

aforesaid aquatic environments receive direct or indirect U

contamination due to activities associated with the nuclear

industry. The safe level of uranium in groundwater is esti-

mated to be 30 lg L-1 (US EPA 2003). Toxic concentra-

tions, as high as 11.7 g L-1, of depleted uranium at some

highly contaminated sites have also been reported (Riley and

Zachara 1992). Certain contaminated surface ponds have

revealed highly elevated concentrations of uranium, i.e.

1,140 lg L-1 (Aleissa et al. 2004).

The bioavailability of uranium to aquatic microorgan-

isms poses a critical issue for aquatic habitats exposed to

uranium release from U ore deposits. Microorganisms can

potentially affect the form and distribution of uranium in the

environment by oxidizing, reducing, binding, immobilizing,

complexing or by precipitating the metal. Earlier investi-

gations have revealed the inability/lesser efficiency of sev-

eral marine microalgae/cyanobacteria/fungi to sequester

uranium from aqueous solutions/sea water above pH 6.0

(Sakaguchi et al. 1978; Nakajima et al. 1982; Tsezos and

Noh 1984). This is due to the formation of stable carbonato

complexes of uranyl ions in sea water which suppress uranyl

adsorption by these organisms. However, a non-specific

adsorption of anionic uranyl carbonate onto negatively

charged surface of Bacillus subtilis in the pH range of 7–9

was demonstrated through thermodynamic modelling

(Gorman-Lewis et al. 2005). A filamentous green fresh

water algae, Spirogyra, accumulated uranium, from pond

water (pH ranging from 7.63 to 8.41) and marine water (pH

8.35), at elevated levels (140–1.140 lg L-1) under natural

conditions. Uranium was suggested to be absorbed as

anionic uranyl carbonate complexes resulting from ion

exchange between the anionic (uranyl) complexes and

hydroxyl group on the algal cell surface (Aleissa et al.

2004). A green alga, Chlamydomonas reinhardtii showed

maximum uranium uptake at pH 7 as compared to pH 5. This

phenomenon was attributed to reduced concentrations of

protons with increased pH, thereby reducing the competition

for uranium for physiologically active binding sites on the

cell surface (Fortin et al. 2007). Powdered form of lake

harvested biomass, consisting predominantly of Microcystis

aeruginosa from cyanobacterial water bloom, demonstrated

an optimal pH for uranyl binding between 4 and 8 (Li et al.

2004).

Microbial response to uranium toxicity

Uranium is known for its chemical toxicity rather than

radiotoxicity and its contamination in surface or ground

water poses health hazards. In microbial systems, no spe-

cific mechanism has been attributed to uranium toxicity.

Although, numerous studies have been done on uranium

toxicity, very limited knowledge exists regarding molecu-

lar mechanisms underlying microbial exposure to uranium.

A whole genome transcriptomic analysis of uranium-

stressed Caulobacter crescentus has shown that uranium

imposes less direct oxidative damage to cells as compared

to cadmium and chromium. Uranium exposure resulted in a

2.9-fold induction of superoxide dismutase (Mn sod A)

84 Photosynth Res (2013) 118:83–94

123

gene as compared to cadmium and chromium stress which

showed 18.9- and 14.1-fold induction of the same gene (Hu

et al. 2005). In a recent study, it has been demonstrated that

uranium exerts acute toxicity by binding to pyrroloquino-

line quinone (PQQ) in Pseudomonas aeruginosa. Uranium

binding to PQQ (a redox cofactor of a number of bacterial

dehydrogenases which protects the cells from oxidative

stress in vivo) resulted in the inhibition of growth and

metabolism of Pseudomonas (Vanengelen et al. 2011). A

thermoacidophilic strain, Metallosphaera prunae, isolated

from a smouldering heap on a uranium mine in Thuringen,

Germany, on exposure to high concentrations of soluble

uranium (1,238 mg L-1) for 15 min, was shown to abort

its transcriptional and translational processes to resist U

toxicity. Transcriptomic analysis of uranium-exposed

Metallosphaera prunae indicated the possible role of sid-

erophore in U sequestration, as the genes encoding an

operon homologous to iron complex transport system in

other microbes, were found to be highly induced (Muk-

herjee et al. 2012).

Uranium has no biological role and its transportation into

the microbial cells occurs due to an increased membrane

permeability resulting from uranium toxicity (Suzuki and

Banfield 1999). There is no evidence for uranium transporters

in microbes and intracellular accumulation of uranium occurs

mostly through passive diffusion (Suzuki and Banfield 2004).

Under low pH conditions, UO22? is the predominant aqueous

uranium species which is considered very toxic due to its high

positive charge (Suzuki and Banfield 2004). Some of the

important microbial interactions with uranium at low to

neutral pH include binding to the various ligands on cell

surface (Fowle and Fein 2000; Fein et al. 1997), chelation by

extracellular polysaccharides (EPS) (Mohamed 2001), intra-

cellular polyphosphates (Merroun et al. 2006; Swift and

Forciniti 1997; Suzuki and Banfield 2004), binding to S-layer

proteins (Merroun et al. 2005) and siderophores (Vijayar-

aghavan et al. 2013), precipitation as inorganic mineral phase

(Macaskie et al. 2000; Martinez et al. 2007; Appukuttan et al.

2006) or reduction to insoluble U(IV) (Lovley et al. 1991).

Hydroxamate-type siderophores have been shown to be

produced in the presence of uranium and their complexation

with uranium has been demonstrated in a marine cyanobac-

terium Synechococcus elongatus BDU130911 (Vijayaragh-

avan et al. 2013). TEM analysis of Arthrobacter,

Desulfovibrio and Sphingomonas cells which had accumu-

lated uranium, showed needle-like fibrils in the cytoplasm

(Merroun et al. 2006). Extracellular bioprecipitation of ura-

nium as uranyl phosphate has been found to diminish uranyl

sensitivity of bacterial cells (Misra et al. 2012). A recent

investigation showed that the conductive pili (anchored to

cell envelope) of Geobacter sulfurreducens catalysed the

extracellular reduction of U(VI) to U(IV) and that the

expression of pili enhanced the rate of uranium

immobilization, preventing permeation of U inside the peri-

plasm and preserving the vital function of cell envelope and

cell’s viability (Cologgi et al. 2011). Electron microscopic

observations on uranium-exposed Acidithiobacillus ferroox-

idans cells have revealed intracellular accumulation of ura-

nium in polyphosphate bodies, which appeared as dense dark

granules in the cytoplasm (Merroun et al. 2003).

Cyanobacteria’s mettle with toxic metals

Cyanobacteria are endowed with attributes to respond to

toxic metals to lessen or prevent the metal toxicity. These

mechanisms modify the metal speciation, leading to

decreased or increased mobility of metals. Such mecha-

nisms include extracellular sequestration, intracellular

compartmentalization, organic or inorganic precipitation,

active transport and synthesis of the metal-binding pro-

teins such as metallothioneins. Some of these mechanisms

harboured by cyanobacteria for detoxification of metal

contaminants are discussed below.

Sequestration by cell surface and associated

components

The cyanobacterial cell surface harbours functional groups,

such as carboxyl, phosphoryl, hydroxyl and amine ligands

which bind to the metal ions to form metal–ligand surface

complexes (Yee et al. 2004; Phoenix et al. 2002). This

potential is available even when the cells are dead. These

organisms bind the heavy metals onto the cell wall or EPS

found outside its cell wall restricting the metal transport to

the cell interior (Philippis and Vincenzini 1998). Adsorp-

tion studies of copper, cadmium and lead with Calothrix

sp. KC97 have shown that the reactive sites on the cell

surface are heterogeneously distributed between EPS and

the cell wall, the carboxyl group being the most prominent

site for metal binding (Phoenix et al. 2002). It is suggested

that the carboxyl groups represent the most important sink

for metal ions at near neutral pH (Yee et al. 2004).

Polikarpov (1966) has proposed that the radionuclides

present in aquatic environment are accumulated by the

marine microorganisms through direct adsorption from the

water and this property is independent of cellular metabo-

lism. The cell surfaces of cyanobacteria have several com-

ponents apart from EPS which contribute to metal

adsorption. Cell surface-associated mucilaginous sheaths of

Gloeothece magna have been shown to bind cadmium and

magnesium effectively (Mohamed 2001). Cells of Ana-

baena cylindrica produce extracellular polypeptides which

complex with copper, zinc and iron (Fogg and Westlake

1955). Another potential metal complexation mechanism is

through siderophores. Siderophores are low molecular

Photosynth Res (2013) 118:83–94 85

123

weight Fe(III) coordination compounds which are secreted

by microorganisms to enable accumulation of iron.

Although specific for iron, the siderophores have been

shown to sequester gallium, chromium, nickel, uranium and

thorium (Macaskie and Dean 1990). Metal exclusion by cell

wall, membrane or cell envelope results in structural alter-

ation of the cells. Copper-stressed cells of Synechocystis sp.

PCC 6803 developed thickened calyx around the cell wall

which was found to be responsible for binding copper ions

on cell surface (Gardea Torresdey et al. 1996).

Intracellular accumulation

The cyanobacterial cells have been shown to concentrate

several metals within the cells. The metal ions cross the

cyanobacterial membranes with the help of channels,

termed as porins, by active or passive mechanisms (Bev-

eridge 1981; Swift and Forciniti 1997). Intracellularly, the

metal ion sequestration is facilitated either by polyphos-

phate bodies or small, cysteine-rich proteins called metal-

lothioneins (MTs) (Pettersson et al. 1988; Daniels et al.

1998). Live cyanobacterial cells have been shown to con-

centrate metal ions such as Pb, Sr, Mn, Al, Zn, Cu, Cd, Hg

in the intracellular polyphosphates (Jensen et al. 1982;

Rachlin et al. 1984; Swift and Forciniti 1997).

Another aspect of intracellular metal accumulation by

cyanobacteria is the synthesis of metal-binding proteins

known as MTs. These are low molecular weight, cysteine-

rich proteins which bind metal ions in metal thiolate

clusters. Their synthesis has been shown to increase in

response to increased concentrations of metals such as

cadmium, copper and zinc (Blindauer et al. 2002). MTs

form complexes with metals and prevent accumulation of

potentially toxic-free metal ions within the cytosol, thereby

offering protection to the cells (Blindauer et al. 2002).

SmtA protein from Synechococcus PCC7942 is the only

fully characterized prokaryotic metallothionein (Turner

and Robinson 1995). A Zn metallothionein-like sequence

has also been reported in Anabaena PCC7120 and Syn-

echocystis PCC6803 (Blindauer et al. 2002). However, the

cyanobacterial MTs have not been reported to bind or

detoxify uranium so far.

Active metal transport across the cytoplasmic membrane

occurs through import systems, using ABC-type or P-type

ATPases, involving ATP hydrolysis as the energy source

(Nies 2003). While the ABC transporters are known to

translocate biomolecules like peptides, amino acids, sugars

or inorganic ions (Mikkat and Hagemann 2000), P-type

ATPases are membrane transporters which carry metal ions

Table 1 Comparison of uranium-binding capacities of various

biomass

Biomass Uconc.

(mg L-1)

pH Loading

(mg g-1)

References

Rhizopus

arrhizus

50–1,000 2–5 180 Tsezos and

Volesky (1981)

Deinococcus

radiodurans

10 4 57.04 Suzuki and

Banfield (2004)

Cystoseira

indica

500 4 198 Khani et al. (2008)

Sargassum

fluitans

200 4 560 Yang and Volesky

(1999)

Bacillus subtilis n.a 4 600 Sakaguchi (1996)

Lentinus sajor-

caju

200 4.5 128 Bayramoglu et al.

(2006)

Trichoderma

harzianum

100 4.5 196 Akhtar et al.

(2007)

Catenella

repens

100 4.5 303 Bhat et al. (2008)

Saccharomyces

cerevisiae

10 4.6 12 Sakaguchi and

Nakajima (1991)

Aspergillus

niger

10 4.6 29 Sakaguchi and

Nakajima (1991)

Arthrobacter

simplex

10 4.6 58 Sakaguchi and

Nakajima (1991)

Pseudomonas

sp.

n.a 5 410 Sar and D’souza

(2001)

Talaromyces

emersonii

30–300 5 323 Bengtsson et al.

(1995)

Scenedesmus

obliquus

5 5 75 Zhang et al. (1997)

Peltigera sp. 100 4–5 42 Haas et al. (1998)

Synechococcus

elongatus

23.8 6 66.93 This work

Anabaena

torulosa

23.8 6 30.94 This work

Chlorella

regularis

1 8 Nil Sakaguchi et al.

(1978)

Microcystis

aeruginosa

100 8 44 Li et al. (2004)

Rhizopus

arrhizus

0.003 8 2.2 Tsezos and Noh

(1984)

Synechococcus

elongatus

1 8 1.76 Sakaguchi et al.

(1978)

Synechococcus

elongatus

0.003 7.8 2.96 Acharya et al.

(2013)

Synechococcus

elongatus

5 7.8 13.3 Acharya et al.

(2013)

Synechococcus

elongatus

11.9–238 7.8 124 Acharya et al.

(2009)

Anabaena

torulosa

11.9–238 7.8 220 This work

n.a., not available

86 Photosynth Res (2013) 118:83–94

123

and maintain homeostasis of the cytoplasmic metals

(Arnesano et al. 2002). ABC-type Mo and Zn transporters

have been shown to exist in Synechocystis sp. and Syn-

echocystis PCC6803, respectively (Self et al. 2001; Cavet

et al. 2003), while P-type ATPase having affinity for

copper has been reported both in Synechococcus PCC 7942

and Synechocystis PCC6803 (Axelsen and Palmgren 1998).

No such transporter for uranium is known in bacteria.

Uranium sequestration by marine cyanobacteria

Most of the studies on uranyl bioremediation have focussed

on low pH conditions, where UO22? is the predominant

aqueous species. However, very little work has been done

on uranyl adsorption onto bacteria at or above circum-

neutral pH (C7.5) where the aqueous uranium speciation is

so complex. In groundwater, aqueous calcium uranyl car-

bonate complex has been identified to dominate uranium

speciation in mid to high pH solutions. Such aqueous

complexation results in high degree of uranium mobility

(Gorman-Lewis et al. 2005). Although there are reports on

adsorption of uranyl carbonate complexes onto mineral

surfaces (Gorman-Lewis et al. 2005), the evidence for such

adsorption onto anionic microbial surfaces is scanty

(Table 1).

Uranium sequestration from aqueous solutions

above circumneutral pH

We investigated the uranium-binding abilities and the

underlying mechanisms in two selected marine cyanobac-

teria—a unicellular strain, Synechococcus elongatus

BDU75042, and a filamentous, nitrogen-fixing strain,

Anabaena torulosa, from micromolar concentrations of

uranyl carbonate at pH 7.8. S. elongatus BDU75042 was

procured from National Facility for Marine Cyanobacteria

(NFMC) Tiruchirapalli, India, whereas A. torulosa was

isolated earlier in our laboratory (Apte and Thomas 1980)

from saline paddy fields of Trombay, Mumbai, India. The

presence of uranyl carbonate species [UO2(CO3)22-] in the

experimental solutions (BG 11 media without phosphate)

was confirmed by UV-Vis absorption spectrophotometry

which showed absorbance peaks at 434, 448 and 464 nm

(Acharya et al. 2009).

Synechococcus elongatus BDU75042 cells, exposed to

23.8 mg L-1 U (or 100 lM) at pH 7.8 for 5 h, bound 72 % U

resulting in a loading of 53.5 mg U g-1 dry weight (Acharya

et al. 2009). Energy dispersive X-ray fluorescence (EDXRF)

spectroscopy of uranium-loaded biomass revealed all com-

ponents of UL X-rays (ULl, ULa, ULb1 and ULb2) confirming

the association of uranium with the cells (Acharya et al.

2009). Such U-loaded cells exhibited black deposits around

the cell margins as compared to control untreated cells

(Fig. 1a, b). Treatment of U-loaded cells with 0.1 N HCl

showed loss of black deposits from the cell surface (Fig. 1c)

along with *80 % U desorption. Most of the bound uranium

was found to be associated with the EPS, suggesting its

interaction with the surface active ligands. Fourier-transform

infrared (FT-IR) spectroscopy suggested the amide groups

and the deprotonated carboxyl groups on the cyanobacterial

cell surface as likely to be involved in uranyl adsorption. The

X-ray diffraction (XRD) analyses revealed the identity of the

uranium deposits associated with the cell biomass as uranyl

carbonate hydrate. The uranyl-binding efficiency of the heat-

killed or the non-viable Synechococcus cells was similar to

that of live cells, corroborating their extracellular localization

(Acharya et al. 2009).

The filamentous, heterocystous cyanobacterium, A. to-

rulosa, was also found to bind uranium efficiently from

aqueous solutions containing 23.8 mg L-1 U (or 100 lM)

uranyl carbonate at pH 7.8. The uranyl sequestration

kinetics exhibited (a) an initial rapid phase, binding 48 %

uranium within 30 min resulting in a loading of

56 mg U g-1 of dry weight, followed by (b) a slower

phase, binding 65 % uranium with resultant loading of

77.35 mg U g-1 in 24 h (Acharya et al. 2012). However,

unlike S. elongatus, the heat-killed A. torulosa cells or the

EPS derived from such cells exhibited limited uranyl

binding (*26 %) as compared to live cells (65 %),

highlighting the importance of cell viability for optimum

uranyl-binding capacity of A. torulosa (Acharya et al.

2012). Cells challenged with 23.8 mg L-1 for 24 h U

showed dense dark granular structures resembling poly-

phosphate bodies as compared to the unchallenged cells

(Fig. 2a–c). Treatment of uranium-loaded cells with 1 N

HCl at 100 �C for 15 min resulted in complete extraction

of total cell-bound uranium and inorganic phosphate

demonstrating co-localization of uranium with acid solu-

ble polyphosphates (Acharya et al. 2012). Further exam-

ination using light, fluorescence and scanning electron

microscopy-based imaging coupled with energy dispersive

X-ray (EDX) spectroscopy in this filamentous marine

cyanobacterium, A. torulosa, revealed the presence of

acid soluble, novel surface-associated polyphosphate

bodies (SAPBs) and their interaction with uranium

(Acharya and Apte 2013).

Adsorption isothermal data interpreted by Langmuir

model over a concentration range of 11.9–238 mg L-1 of

uranyl carbonate at pH 7.8 demonstrated a maximum

loading of 124 and 220 mg U g-1 dry weight biomass in S.

elongatus and A. torulosa, respectively (Table 1). A

remarkable loading (220 mg U g-1) was observed in A.

torulosa, which surpassed the reported economically fea-

sible adsorption threshold limit of 15 % of biomass dry

weight (Volesky 1990).

Photosynth Res (2013) 118:83–94 87

123

Uranium sequestration from saline solutions

above circumneutral pH

Sea water is an inexhaustible and green source of uranium

with an estimated uranium content of 4.5 billion tonnes

which is 1,000 times more than the terrestrial deposits of

uranium (Heitkamp and Wagener 1982). However, the

uranium concentration in sea is very low, i.e. 13 nM or

3 lg L-1. Uranium sequestration from simulated sea

water containing 3 lg L-1 U by the marine cyanobacte-

rium, S. elongatus BDU75042, was assessed over short

(24 h–5 days) and long (38 days) exposure time periods

(Acharya et al. 2013). The organism could remove

90–98 % uranium resulting in a loading of 42 lg U g-1

in 5 days. Under continuous replenishment conditions

over a prolonged duration, S. elongatus BDU75042

(contained in dialysis bags and suspended in simulated sea

water) demonstrated superiority over the other tested

chemical and biological alternatives in terms of high

uranium loading values (2960 lg g-1) in 4 weeks, toler-

ating well the high salinity of sea water (0.5 M NaCl).

Nearly 85–90 % of cell-bound uranium could be desorbed

using 0.1 N HCl. The organism could also rapidly

sequester uranium (13,306 lg U g-1 in 24 h) from

aqueous solutions supplemented with 0.6 M NaCl and

*5 mg L-1 [UO2(CO3)2]2- at pH 7.8 (simulated brine

reject solutions) (Acharya et al. 2013). Table 1 compares

the uranyl-binding capacities of S. elongatus BDU75042

and A. torulosa with various other microorganisms

reported in the literature.

Recovery of uranium from sea water is being explored for

over six decades in efforts to secure uranium resources for

energy production in future. Various inorganic materials,

chelating polymers, nanomaterials, biopolymers, etc. have

(c)

5µm

(b)

5µm

(a)

5µm

Fig. 1 Light microscopy of uranium-exposed S. elongatus cells. The

mid-exponential phase cells were incubated a under control condi-

tions or b were exposed to 23.8 mg L-1 uranyl carbonate

[UO2(CO3)2]2- at pH 7.8 for 5 h resulting in loading of

53.5 mg U g-1 dry weight and c subsequently washed with 0.1 N

HCl. Cells were observed using bright field microscopy in a Carl

Zeiss Axioscop 40 microscope, with oil immersion objectives

(magnification 91,500)

(b)(a) (c)

Fig. 2 Light microscopy of uranium-exposed A. torulosa cells. The

mid-exponential phase cells were incubated a under control condi-

tions or b were exposed to 23.8 mg L-1 uranyl carbonate

[UO2(CO3)2]2- at pH 7.8 for 24 h resulting in loading of

77.35 mg U g-1 dry weight and c the polyphosphate bodies in

U-loaded A. torulosa were stained using standard staining procedures

with 0.05 % toluidine blue (adjusted to pH 1.0) (Ashford et al. 1975).

Cells were observed using bright field microscopy in a Carl Zeiss

Axioscop 40 microscope, with oil immersion objectives. The arrows

indicate the dense dark granular structures (b) or the distinct, dark red

spheres, characteristic of polyphosphate bodies after staining with

toluidine blue (c) in U-loaded A. torulosa cells (magnification

91,500)

88 Photosynth Res (2013) 118:83–94

123

been demonstrated to sequester uranium from sea water with

highest reported adsorbent loading of only 3.2 mg U g-1

with polyacrylamidoxime (PAO) after 180 days of exposure

to sea water (Kim et al. 2013). But prolonged usage of PAO

resins encounters problems of biofouling involving adhesion

and growth of marine microorganisms and algae. Our

studies on uranium sequestration from simulated sea water

with S. elongatus BDU75042 have shown a loading of

2.96 mg U g-1 in 27 days, which is remarkably higher than

that of most of the inorganic/organic/bioadsorbents reported

so far (Acharya et al. 2013). Bioadsorbents such as these

marine cyanobacteria have a strong advantage of being

renewable and environmental friendly for marine systems.

However, further research is necessary for (a) optimization

and development of a commercially feasible biotechnology

for recovery of uranium from sea water, (b) increasing the

uranium adsorption capacity/rate by optimizing the contact

of bioadsorbent with sea water, (c) establishing the adsor-

bent regeneration abilities for long-term cost effectiveness

and (d) possible genetic manipulation to enhance the

inherent U-sequestration capacity.

Uranium sequestration from acidic solutions

Uranium-binding abilities of both the marine cyanobacteria

from micromolar solutions of uranyl nitrate at pH 6 were

also evaluated. On exposure to the test solutions (BG 11

medium devoid of phosphate, as described in Acharya et al.

2009) containing 23.8 mg L-1 [UO2(NO3)2] at pH 6, S.

elongatus bound *90 % of input U within 30 min (Fig. 3a)

0 10 20 30 40 50 150 200 250 3000

20

40

60

80

100

% U

bou

nd to

cel

ls

Time (mins)

S. elongatusA. torulosa

(a)

0 100 200 300 4000

300

600

900

1200

1500

ULβ2

UL β1

ULα

Znkα

Fekα

CakαKkα

Cou

nts

X-ray energy (Channel No.)

MnkαULl

Fekβ

(b)

0 100 200 300 4000

300

600

900

1200

1500

Fekα

Cou

nts

X-ray energy (Channel No.)

CakαKkα Znkα

ULα

ULβ1

ULβ2

(c)

Fig. 3 Uranyl binding by S. elongatus and A. torulosa cells at pH 6.

Mid-log phase cells (equivalent of 0.320 and 0.2 mg dry wt. mL-1 for

S. elongatus and A. torulosa, respectively) were exposed to

23.8 mg L-1 uranyl nitrate in 10 mL experimental medium (BG 11

medium devoid of phosphate) at pH 6 and were assayed for cell-

bound uranium after specified time intervals. The proportion (%) of

uranium bound to the cells is shown in a. b, c The EDXRF spectra of

uranium-loaded Synechococcus (66.93 mg U g-1) and Anabaena

(30.94 mg U g-1) cells, respectively. The components of U L X-rays:

ULl, ULa, ULb1 and ULb2 are seen in the spectra (1 keV = 27

channels)

Photosynth Res (2013) 118:83–94 89

123

and showed a loading of 66.93 mg U g-1 (Table 1). This is

particularly important in the context of contaminated surface

or groundwater where pH ranges from 5 to 9. Groundwater

contaminated with 1 mg L-1 uranyl nitrate has been

reported (Abdelouas et al. 1998). However, A. torulosa

showed only limited uranyl binding (*26 %) (Fig. 3a) and

loading (30.94 mg U g-1) (Table 1) capacity from test

solutions, under identical conditions. Figure 3b, c shows the

EDXRF spectra of U-loaded biomass of S. elongatus and A.

torulosa, respectively at pH 6, wherein components of U L

X-rays, i.e. ULl, ULa, ULb1 and ULb2, were detected con-

firming the association of uranium with the target

cyanobacteria.

Natural marine cyanobacterial strains, such as S. elong-

atus BDU75042 or A. torulosa, which can efficiently

sequester uranium from environmentally relevant concen-

trations (nM to lM) of uranyl carbonate seem to be

promising, both for recovery of uranium from lean

resources like sea water and for remediation of uranium-

contaminated aquatic environments. The marine cyano-

bacteria investigated here exhibit features like rapid

kinetics, high uranyl loading capacity, tolerance to high

salinity, utility for removal of uranium over multiple cycles

of adsorption and desorption and possibilities of using live

cells in free or immobilized form, dead cells or even EPS

therefrom (Acharya et al. 2009, 2012, 2013). While both S.

elongatus BDU75042 and A. torulosa could sequester

uranium from aqueous solutions above pH 7 prevalent in

aquatic environments, the strain BDU75042 appears to

hold a greater promise for uranium immobilization in low

pH wastewater environments common to nuclear waste.

Immobilization of cyanobacterial biomass for metal/

U sequestration

Immobilization of the cyanobacterial biomass in solid and

inert supports allows easier metal recovery along with the

regeneration of biomass, without compromising the natural

binding capacity of biomass for the metal. The free cells or

the cell suspensions have generally low mechanical

strength and smaller particle size. High pressures required

to generate suitable flow rates for metal binding lead to

disintegration of the free cells. These problems can be

appropriately addressed using immobilized biomass.

Among the several methods of immobilization of cyano-

bacteria reported in the literature, entrapment of the

cyanobacterial cells in natural or synthetic polymers is the

most popular method (Mallick and Rai 1994; Garbisu et al.

1993; Prakasham and Ramakrishna 1998). Cyanobacteria

possess high metal adsorption capacity which has encour-

aged their application for detoxification of the effluents in

preference to the conventional wastewater treatment

facilities (Darnall et al. 1986). Immobilized cyanobacteria

show better potential for metal removal than their free

living counterpart. Immobilized Anabaena doliolum

showed an increased uptake of Cu (45 %) and Fe (23 %),

compared to the free living cells (Rai and Mallick 1992).

Synechococcus PCC7942 biomass immobilized in silica

successfully bound copper, lead, nickel and cadmium

under flow-through conditions at pH 5. More than 98 % of

the adsorbed metals could be recovered when treated with

0.1 N HCl, providing a recyclable system for adsorption of

these metal ions (Gardea Torresdey et al. 1998).

Aqueous biopolymer solutions containing polysaccha-

rides of Nostoc muscorum coated/immobilized onto a

uranium-contaminated steel coupons showed ability for

removal of[80 % of the uranium (VI) from such coupons.

The biopolymer–radionuclide complex was then removed

or peeled off the steel coupons as a viscous film, as a dry

powder, or by washing. This ‘‘apply, wait, and remove’’

procedure was proposed to reduce the amount of time spent

in uranium decontamination activities. The metal sorptive

capacity of such biopolymer of Nostoc was found to be up

to 0.2 g U/g biomass (Davison et al. 2001).

There is no report on the usage of immobilized cyano-

bacterial biomass for uranium recovery from the aqueous

solutions above pH [ 7 under flow-through conditions.

The rapid kinetics, high U-loading capacity by the free

(live or dead) cells from micromolar concentrations of

uranyl carbonate and reversibility of uranyl binding in

batch mode (Acharya et al. 2009) prompted us to develop a

workable system using immobilized cyanobacterium, S.

elongatus BDU75042, for uranium recovery under flow-

through conditions. It was shown previously that the

adsorption of uranium by immobilized cells of Strepto-

myces and Chlorella is unaffected by pH values between 4

and 9 as compared to the free cells (Nakajima et al. 1982).

The whole cells of S. elongatus were immobilized in

polyacrylamide gels (15 %), passed through a sieve, with

size range between mesh numbers 5 and 6 (Fig. 4a1), and

packed into a cylindrical fixed bed column (Fig. 4a2).

Uranyl carbonate solutions (238 mg L-1) (Fig. 4a3) were

pumped into the column at constant upward flow rate

(350 ll min-1) using peristaltic pump (Miclins, India)

(Fig. 4a4) and the effluent fractions were collected in a

fraction collector (Bio-Rad) (Fig. 4a5) to measure their

uranium content (Fig. 4a6). A column packed with poly-

acrylamide without S. elongatus cells served as the nega-

tive control (Fig. 4a7) and revealed loss of \10 % of total

input uranium (238 mg L-1) due to its adsorption to

polyacrylamide. All the data for uranium binding presented

here were corrected for this loss. S. elongatus BDU75042

cells were regenerated three times by desorbing U-loaded

biomass using 0.1 N HCl. In the first adsorption cycle,

almost 91.4 % (217.5 mg L-1) of the input, 1 mM

90 Photosynth Res (2013) 118:83–94

123

(238 mg L-1) uranyl carbonate at pH 7.8 could be adsor-

bed within 5 h, loading up to 72.5 mg U g-1 on the

immobilized biomass (Fig. 4b). Nearly 90 % of the bound

uranium could be recovered by 0.1 N HCl (Fig. 4c). In the

consecutive second and third cycles (of 5 h each), 76.4 %

(*182 mg L-1) and *67 % (*160 mg L-1) of input,

238 mg L-1 uranyl carbonate at pH 7.8 was adsorbed to

the immobilized biomass (Fig. 4c) and 90–92 % of

adsorbed uranium (in second and third cycles of regener-

ation) could be desorbed by 0.1 N HCl (Fig. 4c). Although

no clogging was observed, the consecutive regeneration

cycles led to a *26 % decrease in the overall adsorption

performance of S. elongatus BDU75042 cells after three

cycles (Fig. 4c). The decrease might be due to the damage

of the cell surface by acid usage for biomass regeneration.

In our previous studies, uranium binding in S. elongatus

was found to be predominantly a surface phenomenon

(Acharya et al. 2009). Use of immobilized S. elongatus

cells demonstrates their potential for uranium recovery/

remediation from aqueous solutions allowing regeneration

of the biomass for multiple sorption–desorption cycles

above pH 7.

Conclusions

It is well known that cyanobacteria are efficient metal

adsorbents. Based on critical analysis, one of the cyano-

bacteria-based bioadsorbents, immobilized in polysulfone

and named Bio-Fix, obtained from the biomass of a variety

of sources including the cyanobacterium, Spirulina, has

been commercialized and has been demonstrated to treat

wastewater. It could be reused for more than 120 extrac-

tion-elution cycles (Brierley 1990).

Accumulation of the uranium complexes in aquatic

environment poses a threat to humans and the natural

microflora. Tendency of these U complexes (uranyl car-

bonates) to remain in waters of pH 7–10 causes high degree

0

20

40

60

80

100

62.3 mg g-1

72.5 mg g-1

38.6 mg g-1

% U

rem

ova

l by

the

bio

mas

s

30 min 1h 3h 5hTime

28.5 mg g-1

(b)

0

50

100

150

200

250

Adsorption Desorption

Cycle 3Cycle 2

Ad

sorb

ed/D

eso

rbed

Ura

niu

m (

mg

)

Cycle 1

217.5 196182 163

160 148

(c)

(a)

Fig. 4 Experimental set-up for uranyl binding by immobilized S.

elongatus cells. a Cells (30 mg) were immobilized in 15 %

polyacrylamide gel and passed through a sieve, and were filled into

a column (diameter of column 2 cm, length 20 cm). One litre of feed

solution containing 238 mg L-1 uranyl carbonate at pH 7.8 was

pumped into the column at a constant upward flow rate of

350 ll min-1 over a period of 5 h. The various components used

for fixed bed column experiment were: cells immobilized in

polyacrylamide gel (1), column packed with immobilized S.

elongatus cells (2), feed solution (3), peristaltic pump, (4), fraction

collector (5), uranium effluent (6) and a column packed with

polyacrylamide without S. elongatus cells serving as negative control

(7). Uranium was estimated in the effluent at regular intervals. The

proportion (%) of input uranium bound to the biomass in the column

is shown in b and c depicts U binding/recovery over three

regeneration cycles using 0.1 N HCl from U-loaded S. elongatus

biomass

Photosynth Res (2013) 118:83–94 91

123

of U mobility. Ability to remove uranium from aqueous

solutions above circumneutral pH is rare among microbes.

However, selected cyanobacteria, which represent an

important component of aquatic environment such as ponds,

sea or oceans that receive direct or indirect metal contami-

nation, appear to be endowed with such desirable capacity.

Our recent studies have identified two marine cyanobacteria

with high potential for uranyl sequestration from aqueous

solutions, sea water or reverse osmosis (RO) waste. The

U-sequestration capacity of these microbes is quite high, but

is limited by U concentration in sea water. Further feasibility

studies are required to optimize process parameters for the

development of suitable pilot-scale technologies. This is

necessary since oceans appear to be an evergreen resource of

uranium as compared to the terrestrial resources of uranium

which are expected to be depleted/exhausted in the next few

decades (Heitkamp and Wagener 1982).

Fundamental understanding of mechanisms employed

by cyanobacterial cells to resist/alleviate uranium toxicity

will prove useful for the development of strategies for

either uranium recovery from lean sources such as sea

water or remediation from contaminated aquatic environ-

ments. Exploration of possible occurrence of such abilities

in cyanobacterial strains, which are amenable to genetic

manipulation such as Synechococcus PCC7942 or Ana-

baena PCC 7120, will be useful to understand the genetic

basis of such phenomena and open new vistas for

enhancing such capabilities through genetic engineering.

Acknowledgments The authors thank Prof. L. Uma and Dr.

N. Thajuddin, NFMC, Tiruchirapalli, India for providing S. elongatus

strain BDU/75042. The authors thank Dr. Daisy Joseph, Nuclear

Physics Division, BARC, for extending technical help in EDXRF

analyses of uranium in cyanobacterial biomass samples.

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