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Review

Algal–bacterial processes for the treatment of hazardouscontaminants: A review

Raul Munoza,b, Benoit Guieyssea,�

aDepartment of Biotechnology, Center for Chemistry and Chemical Engineering, Lund University, P.O. Box 124, S-22100 Lund, SwedenbDepartamento de Ingenierıa Quımica y Tecnologıa del Medio Ambiente, Universidad de Valladolid, Paseo del Prado de la Magdalena,

s/n, Valladolid, Spain

a r t i c l e i n f o

Article history:

Received 9 March 2006

Received in revised form

14 June 2006

Accepted 15 June 2006

Keywords:

Heavy metals

Industrial wastewater

Microalgae

Organic pollutants

Photobioreactors

Photosynthesis

nt matter & 2006 Elsevie.2006.06.011

thor. Tel.: +46 46 2224228;Benoit.guieysse@biotek.lu

A B S T R A C T

Microalgae enhance the removal of nutrients, organic contaminants, heavy metals, and

pathogens from domestic wastewater and furnish an interesting raw material for the

production of high-value chemicals (algae metabolites) or biogas. Photosynthetic oxygen

production also reduces the need for external aeration, which is especially advantageous for

the treatment of hazardous pollutants that must be biodegraded aerobically but might

volatilize during mechanical aeration. Recent studies have therefore shown that when proper

methods for algal selection and cultivation are used, it is possible to use microalgae to

produce the O2 required by acclimatized bacteria to biodegrade hazardous pollutants such as

polycyclic aromatic hydrocarbons, phenolics, and organic solvents. Well-mixed photobior-

eactors with algal biomass recirculation are recommended to protect the microalgae from

effluent toxicity and optimize light utilization efficiency. The optimum biomass concentration

to maintain in the system depends mainly on the light intensity and the reactor

configuration: At low light intensity, the biomass concentration should be optimized to avoid

mutual shading and dark respiration whereas at high light intensity, a high biomass

concentration can be useful to protect microalgae from light inhibition and optimize the light/

dark cycle frequency. Photobioreactors can be designed as open (stabilization ponds or high

rate algal ponds) or enclosed (tubular, flat plate) systems. The latter are generally costly to

construct and operate but more efficient than open systems. The best configuration to select

will depend on factors such as process safety, land cost, and biomass use. Biomass harvest

remains a limitation but recent progresses have been made in the selection of flocculating

strains, the application of bioflocculants, or the use of immobilized biomass systems.

& 2006 Elsevier Ltd. All rights reserved.

Contents

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2800

2. The potential of microalgae for treating hazardous contaminants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2800

2.1. Direct use of algae . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2800

2.2. Photosynthetic aeration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2801

r Ltd. All rights reserved.

fax: +46 46 2224713..se (B. Guieysse).

ARTICLE IN PRESS

WAT E R R E S E A R C H 4 0 ( 2 0 0 6 ) 2 7 9 9 – 2 8 1 52800

3. Microbial selection . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2802

3.1. Microalgae tolerance . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2802

3.2. Microbial interactions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2803

3.3. Microbial growth rate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2803

3.4. Microalgae predominance . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2804

3.5. Inoculation and selection . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2804

4. Photobioreactor design . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2804

4.1. Open bioreactors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2804

4.2. Closed photobioreactors. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2805

4.3. Mixing. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2805

4.4. Biomass harvesting and biomass retention . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2805

4.5. Biomass concentration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2807

4.6. Surface/volume ratio . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2807

4.7. Hydraulic retention time (HRT) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2808

5. Influence of environmental parameters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2808

5.1. pH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2808

5.2. Temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2809

5.3. Light supply . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2809

5.4. Dissolved oxygen concentration (DOC) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2809

5.5. Predators. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2809

6. Future prospects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2809

6.1. Potential uses of the algal–bacterial biomass . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2809

6.2. Combining wastewater treatment with CO2 mitigation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2810

7. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2810

Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2810

References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2811

1. Introduction

Microalgae play an important role during the tertiary treat-

ment of domestic wastewater in maturation ponds or

the treatment of small–middle-scale municipal wastewater

in facultative or aerobic ponds (Aziz and Ng, 1993; Abeliovich,

1986; Mara and Pearson, 1986; Oswald, 1988, 1995).

They enhance the removal of nutrients, heavy metals and

pathogens (Table 1) and furnish O2 to heterotrophic aerobic

bacteria to mineralize organic pollutants, using in turn

the CO2 released from bacterial respiration (Fig. 1). Photo-

synthetic aeration is therefore especially interesting to

reduce operation costs and limit the risks for pollutant

volatilization under mechanical aeration and recent

studies have shown that microalgae can indeed support the

aerobic degradation of various hazardous contaminants

(Munoz et al., 2004; Safonova et al., 2004). Unfortunately,

microalgae are usually quite sensitive towards the hazardous

compounds (Aksmann and Tukaj, 2004; Borde et al., 2003)

and special care must be taken to improve microbial

activity. Hazardous pollutants include a wide range of toxic

and/or persistent substances that can be found in all

environmental compartments. This review will however

focus on the application of algal-based processes for the

detoxification of industrial effluents which biological treat-

ment requires aerobic conditions (biodegradation of recalci-

trant and toxic contaminants) and external oxygen supply

(i.e. highly loaded wastewater). Guidelines for the design,

start-up, and operation of algal–bacterial processes are

provided and discussed, and the areas for further research

are identified.

2. The potential of microalgae for treatinghazardous contaminants

2.1. Direct use of algae

The mechanisms involved in microalgal nutrient removal from

industrial wastewater are similar than that from domestic

wastewater treatment (Table 1). Nutrients are also not con-

sidered as hazardous pollutants and this will not be discussed

further. However, algal-based treatment is especially interest-

ing in the case of N-containing contaminants whose biode-

gradation normally leads to NH4+ or NO3

� release. For instance,

the net amount of NH4+ produced per mole of acetonitrile

biodegraded decreased from 0.74 mol mol�1 in mechanically

aerated batch processes to 0.46 mol mol�1 in photosynthetically

oxygenated batch processes due to algal assimilation (Munoz et

al., 2005a,b). This ratio was further decreased to 0.17 mol mol�1

when the algal–bacterial process was operated in continuous

mode at a HRT of 3.5 d (Munoz et al., 2005a, b).

Heavy metals represent an important group of hazardous

contaminants often found in industrial wastewater (Kratoch-

vil and Volesky, 1998; Volesky, 2001). Microalgae can be

efficiently use to remove these pollutants (Tables 1 and 2)

and a specific metal uptake of 15 mg gBiomass�1 at 99% removal

efficiency has been reported, showing that the process is

competitive compared to other treatment methods (Cani-

zares-Villanueva, 2000). The removal of heavy metals by algae

is therefore well described in the literature and will not be

discussed further in this review (for general reviews, see

Wilde and Benemann, 1993; Perales-Vela et al., 2006).

Microalgae can finally biodegrade hazardous organic pollu-

tants and Chlorella, Ankistrodesmus or Scenedesmus species

ARTICLE IN PRESS

Table 1 – Main applications of microalgae during WWT

Application Comment References

BOD removal Microalgae release 1.5–1.92 kg O2 kg�1 of microalgae produced

during photoautotrophic growth and oxygenation rates of

0.48–1.85 kg O2 m�3 d�1 have been reported in pilot-scale ponds

or lab-scale tank photobioreactors treating municipal or

artificially contaminated wastewater

Grobbelaar et al., 1988; Martinez Sancho

et al., 1993; McGriff and McKinney, 1972;

Munoz et al., 2004; Oswald, 1988

Nutrient removal Microalgae assimilate a significant amount of nutrients because

they require high amounts of nitrogen and phosphorous for

proteins (45–60% of microalgae dry weight), nucleic acids and

phospholipids synthesis. Nutrient removal can also be further

increased by NH3 stripping or P precipitation due to the raise in

the pH associated with photosynthesis

Laliberte et al., 1994; Oswald, 2003;

McGriff and McKinney, 1972; Nurdogan

and Oswald, 1995; Vollenweider, 1985

Heavy metal

removal

Photosynthetic microorganisms can accumulate heavy metals

by physical adsorption, ion exchange and chemisorption,

covalent bonding, surface precipitation, redox reactions or

crystallization on the cell surface. Active uptake that often

involves the transport of the metals into the cell interior is often

a defensive tool to avoid poisoning or it serves to accumulate

essential trace elements. Microalgae can also release

extracellular metabolites, which are capable of chelating metal

ions. Finally, the increase in pH associated with microalgae

growth can enhance heavy metal precipitation

Chojnacka et al., 2005 Kaplan et al., 1995;

Kaplan et al., 1987; Rose et al., 1998;

Travieso et al., 1996; Van Hille et al., 1999.

Wilde and Benemann, 1993; Yu and

Wang, 2004

Pathogen removal Microalgae enhance the deactivation of pathogens by raising

the pH value, the temperature and the dissolved oxygen

concentration of the treated effluent

Aiba, 1982; Mallick, 2002; Mezrioui et al.,

1994 ; Robinson, 1998; Schumacher et al.,

2003

Heterotrophic

pollutant removal

Certain green microalgae and cyanobacteria are able to use

toxic recalcitrant compounds as carbon, nitrogen, sulphur or

phosphorous source

Semple et al., 1999; Subaramaniana and

Uma, 1997

Biogas production CH4 production from the anaerobic digestion of algal–bacterial

biomass allows substantial economical savings

Eisenberg et al., 1981; Oswald, 1976

Toxicity

monitoring

Microalgae are used in toxicity tests or in studies of microbial

ecology as they are sensitive indicators of ecological changes

Day et al., 1999

CO2

Microalgalphotosynthesis

O2

Bacterialoxidation

Light

BiomassOrganicmatter

Fig. 1 – Principle of photosynthetic oxygenation in BOD

removal processes.

WAT E R R E S E A R C H 40 (2006) 2799– 2815 2801

have been successfully used for the treatment of olive oil mill

wastewater and paper industry wastewater (Abeliovich and

Weisman, 1978; Narro, 1987; Pinto et al., 2002, 2003; Tarlan

et al., 2002). Lima et al. (2003) reported p-nitrophenol removal

of 50 mg l�1 d�1 by a consortium of Chlorella vulgaris and

Chlorella pyrenoidosa under non-optimized conditions, which

was close to the 100 mg l�1 d�1 achieved with Pseudomonas sp.

by Kulkarni and Chaudhari (2006). However, heterotrophic

microalgae can be out-competed by heterotrophic bacteria in

continuous open systems because microalgae often exhibit

lower specific growth rates than bacteria (Semple et al., 1999;

Lee, 2001). The applicability of pollutant biodegradation by

algae therefore remains uncertain and should be further

investigated.

2.2. Photosynthetic aeration

Mechanical aeration accounts for more than 50% of the total

energy consumption of typical aerobic wastewater treatments

(Tchobanoglous et al., 2003): Hence, microalgae can improve

the energy-efficiency of BOD removal from domestic waste-

water by providing O2 to the heterotrophic aerobic bacteria

(Fig. 1). This synergistic relationship can also be used for the

economical treatment of hazardous contaminants, which is

also safer as there is less risk of pollutant or aerosol release

than during intensive mechanical aeration (Brandi et al., 2000;

Hamoda, 2006). This is especially advantageous knowing that

many recalcitrant and toxic compounds are much easier to

degrade aerobically than anaerobically. For instance, a micro-

algae–bacteria consortium was successfully used for the

degradation of black oil and the detoxification of industrial

wastewater in Russia (Safonova et al., 1999, 2004). Likewise,

Chlorella sorokiniana was able to support the aerobic degrada-

tion of phenanthrene, acetonitrile, phenol, and salicylate by

pollutant-specific bacteria without any external O2 supply

(Borde et al., 2003, Guieysse et al., 2002; Munoz et al., 2005a, b).

Salicylate was thus totally converted under photosynthetic

oxygenation into biomass by the symbiotic consortium

Ralstonia basilensis–C. sorokiniana with an excess of O2 produc-

tion according to the following reactions (Borde et al., 2003):

ARTICLE IN PRESS

Table 2 – Reported studies on heavy metal accumulation by microalgae

Metal Biomass Accumulationcapacity

(mg gBiomass�1 )

Adsorptionremoval rate(mg l�1 d�1)

Experimentalset-up

Reference

Zn Chlorella vulgaris — 114.2 1-l column reactor

with microalgae

immobilized in

k-carrageenan

Travieso et al., 1999

Cr Scenedesmus acutus — 3.5

Cd Chlorella vulgaris — 2.5

Co Scenedesmus

obliquus

0.82 Rotary biofilm

reactor

Travieso et al., 2002

Zn Euglena gracilis 7.5 — 500-ml E-flasks,

free

microorganisms

Fukami et al., 1988

Cd Chlorella

Homosphaera

8.4 1.44 500-ml E-flasks,

free

microorganisms

Zn Chlorella

Homosphaera

15.6 2.67 Costa and Leite

(1990)

Cd Chlorella vulgaris 2.6 — 1-l E-flasks, free

microorganisms

Khoshmanesh

et al. (1996)

Chlorella

pyrenoidosa

2.8 —

Chlamydomonas

reinhardtii

2.3 —

Al Scenedesmus

subspicatus

6.8 — 50-ml

polyethylene-

flasks, free

microorganisms

Schmitt et al., 2001

Cd 7.3 —

Cu 13.2 —

Hg 9.2 —

Cd Chlorella

sorokiniana

192 — Column reactor

with algae

immobilized on a

vegetable sponge

Akhtar et al., 2003

WAT E R R E S E A R C H 4 0 ( 2 0 0 6 ) 2 7 9 9 – 2 8 1 52802

Salicylate mineralisation by R. basilensis (Borde et al., 2003)

C7H6O3 þ 0:396 NO�3 þ 0:396 Hþ þ 4:0795 O2

) 5:02 CO2 þ 1:515 H2Oþ 1:98 CH1:7O0:4N0:2.

C. sorokiniana Photosynthesis

5:02 ðCO2 þ 0:7609 H2Oþ 0:15 NO�3 þ 0:1782 Hþ þ 0:0094 PO3�4 ) CH1:7O0:4N0:15P0:0094 þ 1:4243 O2Þ

C7H6O3 þ 1:149 NO�3 þ 0:047 PO3�4 þ 2:3047 H2Oþ 1:291 Hþ ) 3:070 O2 þ 1:98 CH1:7O0:4N0:2 þ 5:02 CH1:7O0:4N0:15P0:0094

.

CH1.7O0.4N0.2 (Atkinson and Mavituna, 1983) and

CH1.7O0.4N0.15P0.0094 (Oswald, 1988) represent the biomass

compositions of bacteria and algae, respectively.

The same consortium was able to remove sodium salicylate

at a maximum rate of 87mg l�1 h�1 in a continuous enclosed

photobioreactor (Munoz et al., 2004). This corresponded to an

oxygenation capacity of 77mg O2 l�1 h�1 close to that of large-

scale mechanical surface aerators (125 mg O2 l�1 h�1; Boon,

1983). Likewise, 2.3 g acetonitrile l�1 d�1 was removed in a

continuous column photobioreactor inoculated with C. soro-

kiniana and a bacterial consortium, which was comparable

to the 0.91g l�1 d�1 achieved by Dhillon and Shivaraman (1999)

in a continuous 19-l trickling filter bioreactor or the 1.04g l�1 d�1

reported by Manolov et al. (2004) in a 20-l aerobic packed-

bed reactor. These results clearly illustrate the potential

advantages of photobioreactors for the treatment of in-

dustrial wastes. However, new limitations might arise from

the fact that microalgae are generally more sensitive to

hazardous pollutants and grow at slower rates than their

pollutant-degrading bacterial partner. Special care must there-

fore be given in selecting the consortia and supporting

microalgal activity.

3. Microbial selection

3.1. Microalgae tolerance

Microalgae are generally sensitive to toxic pollutants and

are even recommended as test microorganisms for the

ARTICLE IN PRESS

Microalgae Bacteria

- Temperature increase- pH increase - DOC increase - Bactericides

+ Growth promoters + DOC decrease

- Algaecide

+ CO2 consumption+ Extracellular matter

Fig. 2 – Positive (dashed line) and negative (plain line)

interactions between microalgae and bacteria.

WAT E R R E S E A R C H 40 (2006) 2799– 2815 2803

measurement of acute toxicity (OECD201, 1984; Chen and Lin,

2006). Heavy metals are particularly strong inhibitors of

microbial photosynthesis (Clijsters and Vanassche, 1985) that

can also cause morphological changes in the shape and size

of microalgae cells (Pena-Castro et al., 2004; Travieso et al.,

1999). Salicylate removal under photosynthetic oxygenation

by C. sorokiniana was therefore totally inhibited in the

presence of 2 mg Cu2+ l�1 (Munoz et al., 2006a). However, the

system was efficiently protected by pre-treating the effluent

with the algal–bacterial biomass generated during salicylate

degradation (Munoz et al., 2006a). Microalgae are also

sensitive to organic pollutants as Chen and Lin (2006) showed

that in an air-tight environment (i.e. simulating closed

photobioreactors), PCP inhibited Pseudokircheneriella subcapita-

ta (EC50-48 h of 0.004–0.013 mg l�1) more than Daphnia magna

(EC50-48 h of 0.55 mg l�1; Kuhn et al., 1989). Chlorella are more

tolerant with PCP EC50-96 h values ranging from 0.05 (Mostafa

and Helling, 2002) to 3.77 mg l�1 (Iannacone et al., 2001) but

remain more sensitive than activated sludge microflora (IC50

value of 31.2 mg l�1; Chan et al., 1999). Hence, microalgae are

more likely to be inhibited during the treatment of hazardous

compounds than their associated degrading bacteria (which

are normally better equipped to resist their substrate). For

instance, 10 mg phenanthrene l�1 totally inhibited the growth

of C. sorokiniana whereas a phenanthrene-degrading Pseudo-

monas strain used to form the consortium easily biodegraded

this compound at 25 mg l�1 (Borde et al., 2003).

Microalgae are also sensitive to the combined effect of high

NH3 concentrations and high pH values because NH3 un-

couples the electron transport in photosystem II and com-

petes with H2O in the oxidation reactions leading to O2

generation (Azov and Goldman, 1982). For instance, Abelio-

vich and Azov (1976) observed a decline in the efficiency of a

high rate algal pond (HRAP) when NH3 concentrations and pH

were simultaneously above 2 mM and 8, respectively. Like-

wise, Munoz et al. (2005b) reported the complete inhibition of

C. sorokiniana at a total NH3/NH4+ concentration of 15 mM and

pH 8.7 during the photosynthetically oxygenated treatment of

2 g l�1 of acetonitrile in a 50-l column photobioreactor.

The use of NH3-tolerant microalgae can improve the process

stability as Ogbonna et al. (2000a) reported no significant

effect on the growth of C. sorokiniana at 22 mM NH3

whereas Spirulina platensis was nearly completely inhibited

by 11 mM NH3.

Resistant strains can be obtained by genetic manipulation,

cell acclimation to progressively higher pollutant concentra-

tions, or isolation from contaminated sites where indigenous

microorganisms have already been exposed to the target

contaminants (Malik, 2004). For instance, Essam et al. (2006)

isolated a C. vulgaris–Alcaligenes consortium from the treat-

ment plant of a coking factory effluent containing phenolics

that was able to treat simulated wastewater. However, the

algae were still inhibited by un-characterized organic com-

pounds present in the real wastewater. This problem was

solved by pre-treating the effluent (UV irradiation or activated

carbon adsorption). A different approach was used by Munoz

et al. (2003a) to prevent algal inhibition during the treatment

of phenanthrene by a C. sorokiniana–Pseudomonas sp. con-

sortium: the culture was mixed with an immiscible, biocom-

patible organic phase (silicone oil) that was used to lower the

aqueous concentration of phenanthrene (thereby lowering its

toxicity). Thus, the consortium was able to biodegrade

phenanthrene initially supplied at 200–500 mg l�1 at a max-

imum rate of 24.2 mg l�1 h�1 without problems of solvent

emulsification (common under intensive mechanical aera-

tion). For more information about two-liquid-phase systems,

see the reviews of Daugulis (2001) and Deziel et al. (1999).

3.2. Microbial interactions

The symbiotic microalgal–bacterial relationship is clear when

microalgae provided the O2 necessary for aerobic bacteria to

biodegrade organic pollutants, consuming in turn the CO2

released from bacterial respiration (Fig. 1). However, micro-

algae and bacteria do not limit their interactions to a simple

CO2/O2 exchange (Fig. 2). Microalgae can have a detrimental

effect on bacterial activity by increasing the pH, the dissolved

oxygen concentration (DOC) or the temperature of the

cultivation broth, or by excreting inhibitory metabolites

(Oswald, 2003; Schumacher et al., 2003). They can however

enhance bacterial activity by releasing extracellular com-

pounds as shown by Wolfaardt et al. (1994) who observed that

diclofop methyl removal by a bacterial consortium increased

up to 36% when actively growing algae or their metabolites

were added to the culture. Similarly, bacterial growth can

enhance microalgal metabolism by releasing growth-promot-

ing factors (Fukami et al., 1997; Gonzalez and Bashan, 2000) or

by reducing O2 concentration in the medium (Mouget et al.,

1995; Paerl and Kellar, 1978). De-Bashan et al. (2002),

for instance, reported that the presence of Azospirillum

brasilense enhanced ammonium and phosphorous removal

by C. vulgaris. Bacteria can also inhibit microalgae by

producing algicidal extracellular metabolites (Fukami et al.,

1997).

3.3. Microbial growth rate

Due to their larger size, microalgae generally grow at slower

rates than heterotrophic bacteria (Fenchel, 1974). In particu-

lar, toluene-degrading Pseudomonas sp. can grow at specific

growth rates of 0.4–0.8 h�1 (Reardon et al., 2000) whereas even

the fast growing Chlorella can hardly grow at rates higher than

0.2 h�1 (Lee, 2001). Hence, pollutant removal is often limited

by O2 production in algal–bacterial systems, which is directly

linked to microalgal activity (Guieysse et al., 2002; Munoz

ARTICLE IN PRESS

Fig. 3 – Aeral view of the Cyanotech Corporation’s

microalgae production facility in Kona, Hawaii. Courtesy of

Cyanotech (USA).

WAT E R R E S E A R C H 4 0 ( 2 0 0 6 ) 2 7 9 9 – 2 8 1 52804

et al., 2004, 2005a, b). Rapidly growing microalgae, which also

exhibit high O2 production rates, should therefore be pre-

ferred as O2 suppliers. Munoz et al. (2003b), for instance,

compared the ability of C. sorokiniana, C. vulgaris, Scenedesmus

obliquus, and Selenastrum capricornutum to support the biode-

gradation of salicylate by a Ralstonia basilensis strain and

showed that C. sorokiniana exhibited the highest specific

growth rates (0.045 h�1) and supported the fastest pollutant

removal rates (18 mg salicylate l�1 h�1). In comparison, Sc.

obliquus exhibited a specific growth rate of 0.013 h�1 and

supported the degradation of 5 mg salicylate l�1 h�1.

3.4. Microalgae predominance

The predominance of slow growing microalgae can be

difficult to maintain in continuous systems due to contam-

ination by small and rapidly growing microalgae (Hoffman,

1998). Closed photobioreactors allow for a better species

control and should therefore be preferred (Tredici, 1999) in

situations where slow growing algae are required (i.e. self-

aggregating microalgae). Attempts to sustain specific micro-

algal populations by manipulating the operational variables

have not always been successful as for instance, Benemann et

al. (1980) failed to maintain Oscillatoria sp. and Micractinium sp.

by microscreening and recirculation of the biomass into the

photobioreactor. However, Wood (1987) successfully estab-

lished the predominance of a Stigeoclonium strain by combin-

ing a short hydraulic retention time (HRT), to wash the freely-

suspended microalgae, with crossflow microscreening of the

target strain. The effluent composition and seasonal environ-

mental conditions also strongly influence microbial predomi-

nance (Fukami et al., 1997; Mara and Pearson, 1986): Euglena

and Chlamydomonas dominate at high organic loads in sewage

treatment while Chlorella and Scenedesmus are the most

abundant species at medium loads (Martinez Sancho et al.,

1993). Euglena and Scenedesmus species also predominate over

Chlorella below 15 1C due to their higher tolerance to low

temperatures (Mara and Pearson, 1986).

3.5. Inoculation and selection

Because microalgae activity and sensitivity usually limit the

removal rate of hazardous pollutants in algal–bacterial

systems, it is important to select fast growing and highly

resistant microalgae. Fortunately, rapidly growing Chlorella

and Scenedesmus sp. naturally dominate most continuous

microalgal-based treatment systems (Garcia et al., 2000a;

Martinez Sancho et al., 1993) and Chlorella species are also

considered as highly resistant microalgae (Palmer 1969;

Munoz et al., 2003b). To start-up facultative (with algae on

the top layer) or maturation (aerobic) ponds for domestic

wastewater treatment, it is therefore sufficient to fill up the

systems with freshwater in order to allow for the develop-

ment of algae and heterotrophic bacteria (UNEP, 2005; Mara

and Pearson, 1998). Raw sewage or activated sludge can be

used when fresh water is not available. A similar strategy is

used for HRAP (water from other ponds can also serve as

inoculum, Tryg Lundquist, Lawrence Berkeley Laboratory,

USA, personal communication) and could be applied for the

treatment of hazardous contaminants. This would permit the

co-selection of the bacteria and algae and ensure that the

microorganisms are compatible with each other. However,

isolation with pre-selected specific strains might be necessary

when, for instance, the target contaminants are too recalci-

trant or too toxic or where there is a need for specific

microalgae (for pigment production, easier harvesting, etc.).

Microbial interaction effects and microbial stability should

then be carefully investigated (Munoz et al., 2003b).

4. Photobioreactor design

As seem above, pollutant removal by algal–bacterial consortia

is often limited by oxygen supply. Hence, without taking into

account any economical consideration, photobioreactors for

the treatment of pollutant-laden effluents and photobioreac-

tors for microalgal mass cultivation (Fig. 3) share the same

basic design criteria: high light utilization efficiency, good

scalability, efficient mixing, control over the reaction condi-

tions, and low hydrodynamic stress on the photosynthetic

cells (Borowitzka, 1999;Lee and Lee, 2003; Pulz, 2001; Tredici,

1999).

4.1. Open bioreactors

Photosynthetic microorganisms can be cultivated in open or

closed reactors (Chaumont, 1993; Molina-Grima, 1999). Typi-

cal aerobic ponds used for WWT are large and shallow open

ditches without internal mixing (Mara and Pearson, 1986,

Racault and Boutin, 2005). They are generally designed upon a

surface-loading criterion such as for instance, 11 m2 per

population equivalent (p.e.) (European Commission, 2001).

These systems, which are not specifically designed to

optimize microalgal activity, were early challenged by Oswald

(1988) who designed HRAPs in order to match algal growth

and O2 production with the BOD of the receiving wastewater.

These are 2–3 m wide and 0.1-–0.3 m depth shallow open

ponds built in a raceway configuration (Fig. 3), lined with PVC,

clay or asphalt to avoid infiltration and range from 1000 to

5000 m2 in large-scale applications (Abeliovich, 1986; Molina-

Grima, 1999). Under optimal conditions, HRAPs can treat up to

35 g �BOD �m�2 d�1 (175 g BOD �m�3 d�1 in a 0.2 m deep pond)

ARTICLE IN PRESS

WAT E R R E S E A R C H 40 (2006) 2799– 2815 2805

compared to 5–10 g �BOD �m�2�d�1 (5-10 g BOD �m�3 d�1 in a

1 m deep pond) in waste stabilization ponds (Racault and

Boutin, 2005). This superior design also allows for continuous

operation at 2–6 d HRT (Mara and Pearson, 1986) compared to

10–40 d in traditional ponds (Crites and Tchobanoglous, 1998).

However, given the merits of HRAPs, there are nowadays only

a few full-scale systems in operation (De la Noue et al., 1992;

Mara and Pearson, 1986).

4.2. Closed photobioreactors

Enclosed photobioreactors offer higher photosynthetic effi-

ciencies and better control than open systems (less risks of

pollutant volatilization and predation). They can also be built

vertically in order to minimize space requirement (Pulz, 2001;

Tredici, 1999) and minimize water looses by evaporation

which can be very significant in open systems (Pulz, 2001).

Unfortunately, closed systems are also more expensive to

construct (need for transparent materials such as Plexiglas,

glass, PVC, etc.) and difficult to operate and scale up. Enclosed

photobioreactors are often designed as tubular or flat plate

photobioreactors arranged in a horizontal, inclined, vertical

or spiral manner (Fig. 4) (Tredici, 1999). Tubular photobior-

eactors (Fig. 5) are the easiest to scale up by increasing the

length and number of tubes and by the connection of several

units via manifolds (Borowitzka, 1999). They also exhibit

higher light utilization efficiencies than flat plate photobior-

eactors because of the larger reactor surface area per unit of

occupied land (Tredici and Zittelli, 1998). Thus, oxygenation

rates of up to 4.3 kg O2 m�3 d�1 have been achieved in tubular

reactors (Torzillo et al., 2003). This is significantly higher than

the oxygenation rates in ponds and HRAP reported above and

is comparable to the maximum oxygenation capacity of

mechanical surface aerators (3 kg O2 �m�3 d�1, Boon 1983).

Few studies are available on the application of algae for the

treatment of hazardous pollutants and industrial wastes

(Table 3). To the best of our knowledge, only two commercial

Exhaust Air

CO2 enriched air CO2 enriched air

Exh

(A) (B)

Fig. 4 – Schematic representation of (A) a vertical spiral (Biocoil

photobioreactor for mass algal cultivation and fed with air enri

enclosed photobioreactors have so far been tested for waste-

water treatment: the Bio-Fence manifold tubular reactor

(Applied Photosynthetic Limited, Manchester, United King-

dom) with a total volume ranging from 0.050 to 1 m3, and a

helical tubular reactor called Biocoil (Biotechna-Graesser A.P.

Ltd, Australia) with a maximum working volume of 10 m3

(Tredici, 1999). Unfortunately, no data is available on the

removal and oxygenation rates achieved in these systems.

Hence, only few guidelines, mainly based on the oxygenation

capacity (BOD removal capacity) achievable in each config-

uration, can be given to design photobioreactors for large-

scale treatment (Table 4).

4.3. Mixing

Algal ponds are typically operated as plug-flow systems.

However, homogenous conditions in the reactor are prefer-

able during the treatment of toxic effluent as pollutant

dilution lower the risk of microalgae inhibition. Mixing also

limits the formation of anaerobic zones and more generally

reduce any mass transfer limitations (Grobbelaar, 2000). The

device used for mixing should be selected to reduce shear

stress imposed to the microalgal cells (Barbosa et al., 2004;

Mitsuhashi et al., 1995). Gudin and Chaumont (1991) reported

an increase of up to 75% in microalgal productivity when

pumps were replaced by an airlift system to suspend the cells.

Paddle wheels are therefore often used for algal mass

cultivation in open ponds and in HRAP as they provide a cost

efficient gentle mixing.

4.4. Biomass harvesting and biomass retention

Biomass harvesting is necessary to ensure a good effluent

quality (low suspended solids concentration) and prevent cell

washout during continuous operation (Evans and Furlong,

2003; Munoz et al., 2004, 2005; Richmond, 1983). Unfortu-

nately, none of the common industrial approaches (filtration,

CO2 enriched Air

Exhaust Airaust Air

(C)

), (B) an inclined tubular column, and (C) a vertical flat-plate

ched with CO2.

ARTICLE IN PRESS

WAT E R R E S E A R C H 4 0 ( 2 0 0 6 ) 2 7 9 9 – 2 8 1 52806

centrifugation, microstraining, etc.) have been proven to be

economical and suitable for large-scale microalgae removal

(Hoffman, 1998). Wastewater pond effluents are therefore

often characterized by high TSS (Total Suspended Solids)

values, which is especially problematic in the case of

industrial effluents since the biomass might contain heavy

metals or hydrophobic organic compounds.

Microalgal flocculation followed by gravity sedimentation is

the most common harvesting technique during wastewater

treatment because of the large volumes treated and the low

value of the biomass generated (Nurdogan and Oswald, 1996;

Molina-Grima et al., 2003). Unfortunately, this approach is not

always efficient, especially in the case of the small, rapidly

growing Chlorella or Scenedesmus sp. (Garcıa et al., 2000b).

Instead, multicellular cyanobacteria of the genus Spirulina or

Fig. 5 – Outdoor tubular photobioreactor from the Easy

Algaes production facility, Cadiz, Spain. Courtesy of Easy

Algae (Spain).

Table 3 – Organic pollutant removal by algal–bacterial or micro

Compound Experimentalsystem

Microorga

Acetonitrile 600 ml Stirred Tank

Reactor (STR)

C. sorokiniana

consorti

Acetonitrile 50-l column

photobioreactor

C. sorokiniana

consorti

Black oil 5-ml tubes Chorella/Scen

alcanotrophic

Black oil 100 l tank Chorella/Scen

Rhodococcu/Ph

Phenanthrene 2-l STR with silicone oil

at 10%

C. sorokin

Pseudomonas

Phenanthrene 50 ml tubes with

silicone oil at 20%

C. sorokin

Pseudomonas

Phenol 600 ml STR with

NaHCO3 at 8 g l�1

C. vulgaris/A

sp.

Phenol 100 ml E-flasks Anabaena v

Salicylate 600 ml STR C. sorokiniana/

basilen

p-Nitrophenol — C. vulgar

pyrenoid

the self-aggregating Phormidium bohneri have been success-

fully applied in wastewater treatment of farm effluents

(Olguin, 2003). Gutzeit et al. (2005) also recently described a

self-aggregating algal-bacterial process for domestic WWT

where algal-bacterial flocks ranging from 400 to 800mm were

easily removed by gravity. In our laboratory, biomass auto-

flocculation was observed during the continuous degradation

of salicylate supported by a C. sorokiniana when the photo-

bioreactor was operated at high HRT (unpublished data).

Microalgal autoflocculation can be caused by electrostatic

interactions among the cell walls as a result of Ca/Mg

carbonate or ortophosphate precipitation at high pH (Oswald,

1988). Bioflocculation can occur due to the microalgal release

of long-chain polymers (Garcia et al., 1998). However these

mechanisms are still poorly understood and hard to induce.

The addition of chemical flocculants such as lime, alum or

polyferric sulfate is efficient and reliable but chemical

flocculants remain expensive and increase the effluent

salinity. Instead, chitosan is an edible, economical

(2 US$ �kg�1, 2002) and non-toxic flocculant that is efficient

for the removal of freshwater microalgae (Divakaran and

Sivasankara, 2002). Biomass removal efficiencies of 90% were

thus obtained using 15 mg chitosan � l�1 in our laboratory

during the batch degradation of acetonitrile by an algal-

bacterial consortium (unpublished data). The use of biofloc-

culants from bacteria present within the microcosms is

another very interesting alternative that should be further

investigated (Oh et al., 2001). Finally, recent developments

made in the construction of membrane bioreactors have

made this technology most affordable and increasingly

popular for wastewater treatment (Yang et al., 2006). Such

algal consortia

nisms Removal rate(mg l�1 d�1)

Reference

/bacterial

um

2300 Munoz et al., 2005a

/bacterial

um

432 Munoz et al., 2005b

edesmus/

bacteria

— Safonova et al., 1999

edesmus/

ormidium

5.5 Safonova et al., 2004

iana/

migulae

192 Munoz et al., 2005c

iana/

migulae

576 Munoz et al., 2003a

lcalıgenes 90 Essam et al., 2006

ariabilis 4.4 Hirooka et al., 2003

Ralstonia

sis

2088 Munoz et al., 2004

is/C.

osa

50 Lima et al., 2003

ARTICLE IN PRESS

Table 4 – Comparison of large-scale photobioreactors

Reactor Max.oxygenationcapacity (kgO2 m�3 d�1)a

Lightutilizationefficiency

Scalability Example ofdesign criteria &

featuresb

Reference

WSP 0.01c Very low Easy 11 m2 per

equivalent person,

1 m depth

Racault and

Boutin, 2005

HRAP 0.3–0.38 Low Easy Raceways of 2–3 m

wide and 0.1–0.3 m

deep ponds

Molina-Grima et

al., 1999

Tubular 5.4–6.9 Very high Easy Tubes of 10–100 m

length and 3–6 cm

+

Lee and Low, 1991

Flat plate 6.5–8.3c Very high Difficult Light path 1–5 cm Hu et al., 1996

Tubular (coil) 1.8–2.3c Very high Easy Tube diameter

2–3 cm, cylindrical

structure 8 m

height, 2 m +

Borowitzka, 1999

Vertical

column

3.1–2.4c High Difficult 0.3–0.5 m + and

2–4 m high

columns

Miron et al., 1999

a Except for WSP and the tubular photobioreactor, the oxygenation rates were calculated from reported biomass productivities and conversion

factors of 1.5–1.92 kg O2 kg�1 microalgae.b According to Tredici (1999), Borowitzka (1999), and Janssen et al. (2003).c Based on a pond depth of 1 m, Racault and Boutin (2005).

WAT E R R E S E A R C H 40 (2006) 2799– 2815 2807

bioreactors have been used for the production of algal

pigments (Rossignol et al., 2000) but their potential for algal-

based wastewater treatment must still be proven.

Biomass immobilization is an efficient mean of retaining

biomass during WWT (Nicolella et al., 2000) and microalgae

immobilization in polymeric material such as carrageenan,

chitosan, or alginate has been reported by various authors

(Chevalier and De la Noue, 1985; Lau et al., 1995; Robinson et

al., 1998). However, these matrices are weak and costly, which

has limited their large-scale application (Hoffman, 1998).

Another approach consists on using enclosed photobioreac-

tors where the algal-bacterial microcosm is attached onto the

reactor walls (Munoz et al., 2006b). For instance, Craggs et al.

(1996) successfully operated a shallow open photobioreactor

with the algal-bacterial biomass attached onto the reactor

base for the treatment of agricultural run-off and domestic

wastewater (algal-turf scrubber). Such systems could be

advantageously designed to reduce effluent toxicity (Fig. 6).

4.5. Biomass concentration

Microalgae concentration determines light utilization effi-

ciency (the energy stored as new biomass per unit of light

absorbed, Janseen et al., 2003) in photobioreactors. It there-

fore also controls the oxygenation and pollutant removal

rates achieved in the system. Munoz et al. (2004), for instance,

reported an increase of 44% in salicylate removal when the

biomass concentration was increased from 0.4 to 0.6 g l�1 in a

closed photobioreactor. However, a decrease of 15% on

salicylate removal efficiency was observed when the algal-

bacterial biomass increased from 0.6 g l�1 to 1.3 g l�1. Indeed,

when the biomass concentration reaches a critical value, all

the light provided to the system is used for photosynthesis

and the oxygenation rate reaches a maximum. Increasing the

biomass concentration further only causes mutual shading

and algal dark respiration to occur (Grobbelaar and Soeder,

1985), which reduces the amount of oxygen available to the

bacteria. However, at high light intensities, mutual shading

can be used to increase the frequency of light/dark cycles at

which the cells are exposed in order to optimize the

photosynthetic activity (Hu et al., 1996; Richmond, 2004).

The higher the light intensity, the higher should be the

biomass concentration (Hu et al. 1996). However, the use of

high algal densities to maintain high light/dark frequencies

requires efficient mixing without damaging the cells (Hu et al.

1996). Predicting the optimum biomass concentration under

natural illumination is also very difficult because the light

intensity onsite greatly varies in time.

4.6. Surface/volume ratio

Since the economic cost of artificial lighting is prohibitive,

sunlight must power oxygenation in algal-bacterial photo-

bioreactors. Hence, a crucial design parameter of these

systems is the illuminated surface to volume ratio (Table 4)

that determines the volumetric microalgae growth rate and

therefore the volumetric O2 production and pollutant removal

rates. Oxygenation capacities estimated from outdoors

ARTICLE IN PRESS

Pollutant

O2

Biofilm Bulkliquid

Bulkliquid

Biofilm

Light

Light

(A)

(B)

Fig. 6 – Theoretical dissolved oxygen (dashed line) and

pollutant (plain line) concentration profiles through the

biofilm of (A) a vertical flat photobioreactor with the biofilm

attached on the reactor wall and illuminated from the sides

and (B) a horizontal algal turf reactor with microalgae

attached on the reactor base and illuminated from above. In

the flat reactor, the most active microalgae which are

directly exposed to light are not directly exposed to the

reactor bulk liquid. The concentration of dissolved oxygen is

therefore expected to decrease through the biofilm as a

result of bacterial consumption. At the same time, toxic

pollutants are consumed and their concentration decreases

through the biofilm, which protects the active microalgae

towards pollutant toxicity. In an algal turf reactor, the most

active microalgae which are directly exposed to light are

also directly exposed to the bulk liquid and therefore, to the

highest possible pollutant concentration.

WAT E R R E S E A R C H 4 0 ( 2 0 0 6 ) 2 7 9 9 – 2 8 1 52808

photobioreactors suggest that horizontal or inclined tubular

and flat plate photobioreactors are the most efficient config-

urations for wastewater bioremediation due to their high

illuminated surface to volume ratio (Lee, 2001). However, the

optimum surface/volume ratio is also dependant on factors

such as the biomass concentration established (or main-

tained) in the system, the hydrodynamic regime, and the

impinging light intensity (Molina-Grima et al., 1999).

High illuminated surface/volume ratio generally means

high land requirement, especially when open reactors are

used. For this reasons, waste stabilization ponds are generally

recommended for small-scale decentralized applications

(Crites and Tchobanoglous, 1998) where the expenses for

land are balanced out by lower costs of the initial investment,

operation and management, and water collection. There is

however no clear definition on the size range of decentralized

or small wastewater treatment plant as Crites and Tchoba-

noglous (1998) define them as facilities handling less than

approx. 3800 m3 of wastewater d�1 (approx. 32 000 p.e. using a

equivalent of 120 l p.e.�1 d�1) whereas the European Commis-

sion (2001) classify small and medium sized communities as

500–5000 p.e. Based on the same criteria, algal–bacterial

processes should be suitable for treating up 60–4000 m3

wastewater d�1 at loads of 30–1800 kg d�1, depending on the

local land value. Essam et al. (2006), for instance, reported

complete phenol removal from a synthetic coking wastewater

at 6 d HRT in an algal–bacterial photobioreactor. Thus, the

full-scale treatment of 300 m3 wastewater d�1 should then be

achieved in a 6000 m2�0.3 m HRAP, the size of a 550 p.e. pond

treatment. Unfortunately, not enough current data are avail-

able on the specific use of algal–bacterial processes for

industrial waste treatment to better predict their applicability.

Small-scale decentralized wastewater treatment could also

allow water reuse onsite and reduce the need for transporta-

tion of hazardous wastes.

4.7. Hydraulic retention time (HRT)

HRAP are traditionally operated at 2–6 d HRT (Mara and

Pearson, 1986) and similar values have been reported in

enclosed photobioreactors (Essam et al., 2006; Munoz et al.,

2005a). Munoz et al. (2004), for instance, reported complete

salicylate removal and high DOCs at 2.7 HRT. When the HRT

was decreased from 2.7 to 1.7 d, both the removal efficiency

and the DOC decreased from 18 to approx. 0.5 mg l�1,

indicating that the process became limited by O2 supply and

therefore by the algal activity. Complete pollutant removal

was however achieved at 0.9 d HRT by sedimentation and

recirculation of a portion of the biomass produced (Munoz

et al., 2004).

5. Influence of environmental parameters

5.1. pH

Microalgal CO2 uptake can cause the pH to rise to 10–11 in

HRAPs and high pH values (up to 9) were also recorded during

salicylate biodegradation by an algal–bacterial consortium in

an enclosed photobioreactor (Munoz et al., 2003b). This

increase, which is beneficial for the disinfection of pathogens,

can also cause a decrease in the pollutant removal efficiency

(Oswald, 1988; Schumacher et al., 2003) as complete bacterial

inhibition at pH above 10 is commonly observed in stabiliza-

tion ponds (Mara and Pearson 1986; Oswald 1988). It is

however difficult to dissociate the direct effects of pH on

microbial growth from collateral effects such as modifications

in the CO2/HCO3�/CO3

�2 and NH3/NH4+ equilibria or in phos-

phorus and heavy metal availability (Laliberte et al., 1994).

The pH also influences N and P removal via NH3 volatilization

and orthophosphate precipitation at a high pH (9–11) (Craggs

et al., 1996; Garcia et al., 2000b; Nurdogan and Oswald, 1995).

Fortunately, it is relatively easy to control the pH in biological

systems.

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WAT E R R E S E A R C H 40 (2006) 2799– 2815 2809

5.2. Temperature

The efficiency of microalgae-based treatments normally

decreases at low temperatures (Abeliovich, 1986). Munoz

et al. (2004) observed that the removal efficiency doubled

when the temperature increased from 25 to 30 1C using a

symbiotic microcosm formed by a C. sorokiniana and a R.

basilensis strain (the activities of both microorganisms in-

creased with the temperature in the tested range). However,

Chevalier et al. (2002) demonstrated that a cold-adapted

cyanobacteria strain was suitable for nutrient removal at

average temperature of 15 1C. Likewise, Gronlund (2004)

described a pilot-scale HRAP capable to support 90% BOD

removal at 2.5 d HRT at temperature below 10 1C and light

intensity below 200mE m�2s�1 (Swedish subartic region,

latitude 631N). These studies therefore show the wastewater

treatment with cold-adapted photosynthetic strains in opti-

mized bioreactors is possible despite the decrease in biologi-

cal activity with temperature inherent to any biological

methods.

Excessive temperature at high light intensities and high

biomass concentrations can also arise from the fact algae

convert a large fraction of the sunlight into heat (Abeliovich,

1986). Temperature control by external heat exchanger or

water spray have been proposed to ensure a stable microalgal

population but their costs remain often prohibitive, even for

high-quality algal mass cultivation (Tredici, 1999). An alter-

native to temperature control is the combination of micro-

algal strains with similar characteristics (in terms of O2

supply, inhibition and harvesting) but with different optimum

growth temperatures (Morita et al., 2001).

5.3. Light supply

Sunlight intensity greatly varies during the day and during

the year. Algal activity increases with light intensity up to

200–400mE m�2 s�1, where the photosynthetic apparatus be-

comes saturated, to decrease at higher light intensities

(Ogbonna and Tanaka, 2000b; Sorokin and Krauss, 1958).

Photoinhibition has therefore been observed during the

central hours of a sunny day when irradiance can reach up

to 4000mE m�2 s�1 (Rebolloso Fuentes et al., 1999). It is more

likely to occur at low microalgal concentration, such as during

start-up (Goksan et al., 2003), because the light intensity to

which microalgae are actually exposed is not reduced by

mutual shading (Evers, 1991; Contreras-Flores et al., 2003;

Richmond, 2000). Careful photobioreactor designing can also

avoid excessive damage of the photosynthetic apparatus

by distributing the light irradiating a certain land area

onto a larger surface (Torzillo et al., 2003). Reducing the

size of the antenna of photosynthetic cells using molecular

tools reduces light adsorption and usually allows higher

photosynthesis rates under high light intensities (Melis et al.,

1999).

Periodical absence of light (or periods of low light intensity)

causes a halt (or sever reduction) of photosynthesis, which

generally leads to the occurrence of anaerobic conditions in

the reactor. However, photosynthesis and pollutant removal

normally resume once light is available again. Waste stabili-

zation ponds are therefore designed to cope with natural

diurnal or seasonal light intensity fluctuations by, for

instance, increasing the HRT in the system (Tadesse et al.,

2004). High HRT, or the use of storage tanks during period

of low light intensities, are also important to avoid increases

of toxic pollutant concentrations and inhibition. In a pilot-

scale closed photobioreactor inoculated with a C. sorokinia-

na–Comamonas sp. consortium, oxygen production and acet-

onitrile removal dropped when illumination was stopped for

10 h but it quickly recovered each time illumination was

resumed (Munoz et al., 2005b). Wastewater storage during

nighttime should therefore no affect the overall process

efficiency.

5.4. Dissolved oxygen concentration (DOC)

High DOC levels can generate photo-oxidative damage on

microalgal cells and therefore decrease treatment efficiency

(Oswald, 1988; Suh and Lee, 2003). For instance, Matsumoto

et al. (1996) reported a 98% decrease in the photosynthetic O2

production rate when the DOC increased from 0 to 29 mg l�1

(E350%). O2 supersaturation in enclosed photobioreactors

designed for mass algal cultivation can reach up to 400%,

which severely inhibits microalgal growth (Lee and Lee, 2003).

Fortunately, O2 supersaturation does not constitute a

severe problem in biodegradation processes due to the

continuous O2 consumption by heterotrophic bacteria. For

instance, the DOC was always very low (E0 mg l�1) during the

biodegradation of acetonitrile and salicylate in the batch

mode when the pollutants were present and being degraded.

However, it also always rapidly increased after complete

pollutant depletion (Guieysse et al., 2002; Munoz et al., 2005a).

High O2 concentrations are therefore a good indication of

complete pollutant depletion in continuous processes (Munoz

et al., 2004). Further research should be conducted to

investigate if the DOC can be used for process control to

optimize, for instance, the biomass concentration in the

system.

5.5. Predators

Infections by parasitic fungi like Chytridium sp. or the

development of food chains in the photobioreactor can

cause unexpected process failure (Abeliovich and Dikbuck,

1977). Fortunately, these potential problems can easily

be avoided by daily operating the process at low O2

levels for a short period of time (1 h) in order to

suppress the growth of higher aerobic organisms (Abeliovich,

1986).

6. Future prospects

6.1. Potential uses of the algal–bacterial biomass

Algal–bacterial biomass can be used for various purposes

(Table 5). However, biomass produced from wastewater will

seldom be suitable for the production of food or even high

value chemicals due to high-quality requirements and public

acceptance. Likewise, fertilization should only be conducted if

the biomass does not contains heavy metals or recalcitrant

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Table 5 – Potential uses of algal–bacterial biomass

Algae application Examples/comments References

Human food source Drinks, noodles, health products Liang et al., 2004

Animal feed Tetraselmis sp., Spirulina sp. and Chaetoceros sp. are currently employed as

food source for shrimps or salmonids production

Borowitzka, 1997; Day et al.,

1999

High-value

biomolecules

Astaxanthin, ascorbic acid, b-carotene, glycerol, or poly-b-

hydroxybutyrate

Ghirardi et al., 2000; Pulz

and Gross, 2004; Tsygankov,

2001

Fertilizer Because algae contain large amounts of nitrogen and phosphorus,

algal–bacterial biomass from wastewater treatment represents an

interesting inexpensive fertilizer. Two million hectares for rice

cultivation were thus fertilized in India in 1977

Oswald and Benemann,

1977

Biogas production Anaerobic digestion of biomass into CH4 and CO2 Munoz et al., 2005a

Biofuels Liquid fuels can be produced from the thermochemical liquefaction or

pyrolysis of microalgae. Certain microalgae also have the capability to

accumulate oils in their cells.

Sawayama et al., 1999

WAT E R R E S E A R C H 4 0 ( 2 0 0 6 ) 2 7 9 9 – 2 8 1 52810

compounds (which are often found in industrial effluents).

Hence, the best option remains to use the algal–bacterial

biomass for energy production by anaerobic digestion into

biogas. Due to CO2 fixation by the algae, all the organic matter

biodegraded is converted into biomass under photosyntheti-

cally oxygenated treatment. This represents a considerable

gain in the carbon available for CH4 production compared to

classical aerated processes where approx. 50% of the original

carbon is lost as CO2 released in the atmosphere. For instance,

Munoz et al. (2005a) reported more than a 100% gain in CH4

production when algal–bacterial biomass produced from

acetonitrile treatment was used instead of bacterial biomass

alone. Furthermore, the biogas produced from the digester

can also be sparged and treated in the algal pond to convert

the CO2 anaerobically produced (biogas usually contains

approx, 60% CH4 and 40% CO2) into algal biomass (Eisenberg

et al., 1981; Mandeno et al., 2005). This further improves the

overall C-mitigation and energy-recovery efficiency of the

system.

6.2. Combining wastewater treatment with CO2

mitigation

As algal cultures require large amounts of nutrients, it can be

very advantageous to combine CO2 fixation from gaseous

streams (i.e. combustion systems) with wastewater treatment

(Nakamura, 2003). This could also help removing hazardous

combustion products such as NOx and SOx (Nagase et al.,

1998, 2001). Thus, 26.0 g CO2 m�3 h�1 was fixed and 0.92 g

NH3 �m�3 h�1 was removed when flue gas and wastewater

from a steel making plant were simultaneously treated (Yun

et al., 1997). Benemann et al. (2003) concluded that productiv-

ities near the theoretical maximum, high-energy prices,

and greenhouse gas abatement credits would however be

required to make this process economically realistic. With

current oil prices and the increase pressure to reduce CO2

emissions and dependence to fossil fuels, this might already

be happening.

7. Conclusions

Algal–bacterial systems are efficient for the treatment of

hazardous pollutants but remains limited by the difficulty of

harvesting the biomass formed, the high land requirement of

open systems, or the high construction costs of enclosed

photobioreactors. Hence, suitable applications will be found

when the effluents to be treated contain hazardous volatile

pollutants, where combined removal capacities (organic

pollutants/nutrients/heavy metals) are desired, or when the

biomass produced can be commercialized. In such cases, the

additional costs brought about by land use, reactor construc-

tion and biomass harvesting will be justified by the gains in

safety and energy savings achieved.

Before algal–bacterial processes can widely be implemented

for the treatment of industrial wastes, more research is still

needed to (1) select ‘‘extreme’’ algal strains capable to grow

under wider and more extreme conditions of light, pH,

pollutant concentrations, etc.; (2) understand and control

the mechanisms of autoflocculation and bioflocculation to

improve harvesting and biomass control; (3) scale-up and

model photobioreactors to provide better design guidelines;

and (4) develop new treatment methods such as membrane

photobioreactors or combined physical–biological processes

to improve biomass control and protect algae against

inhibitory effects.

Acknowledgements

This work is dedicated to Professor William J. Oswald

(1919–2005) who was a pioneer in the development of algae-

based wastewater treatment. The financial support from SIDA

(The Swedish International Development Cooperation

Agency, projects SWE-2002-205 and SWE-2005-439) and the

Spanish Ministry for Science and Education (Juan De La Cierva

Program, JCI-2005-1881-5 Contract) are gratefully acknowl-

edged.

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