Biodegradation of Pollutant

LITERATURE REVIEW

2.3 TREATMENT TECHNOLOGIES FOR REMOVALS OF PHENOL, THIOCYANATE, AMMONIA-N, NITRATE-N AND PYRIDINE

2.3.4 Biodegradation of Pollutant

Treatment of phenol, thiocyanate and ammonia is elaborately studied in individual or in common using various bioreactors. Individual biodegradation of these pollutants are discussed in present section.

2.3.4.1 Phenol biodegradation

Many microorganisms are capable of degrading phenol through the action of variety of enzymes. Wide range of bacteria like Pseudomonas sp., Bacillus sp., Acinetubacter sp., Corynbacterium sp., Enterobacter sp., Alcaligenes sp. Streptomyces sp., Serratia sp., are reported to be efficient in phenol degradation in anaerobic, anoxic and aerobic environments (An et al. 2001; Prieto et al. 2002; Neumann et al. 2004; Fang et al. 2006;

Nilotpala and Ingle, 2007; Jiang et al. 2007; Ho et al. 2009). Loh et al. (2000) reported biodegradation of phenol from wastewater is generally more cost effective than the physicochemical treatment process. Efficient phenol removal is reported in temperature of psychrophilic, mesophilic and or thermophilic conditions (Scully et al. 2006; Fang et al.

1996; 2006).

Cleavage of the aromatic ring is typically achieved via the ortho (intradiol) or meta (extradiol) pathways (Yang and Humphrey, 1975). The most well known key intermediates resulting from the biodegradation of aromatic compounds are catechol, protocatechuic acid and gentisic acid. These intermediates further undergo ring fission following the Krebs cycle to yield other metabolites, such as pyruvic acid, acetic acid, succinic acid and acetyl- CoA (Loh and Chua, 2002). Degradation of phenol follows a sequence of (a) hydroxylation to catechol, (b) ring cleavage via catechol-2,3-dioxygenase to 2- hydroxymuconic semialdehyde (HMSA) for meta-pathway, and via catechol-1,2- dioxygenase to cis, cis-muconate for ortho pathway, (c) HMSA is either oxidized to 4- oxalocrotonate or hydrolyzed to 2-oxopent-4-enoate in case of meta and cis, cis-muconate gets converted into muconolactone for ortho-cleavage (Kwon and Yeom, 2009; Cao and Loh, 2008). Phenol degradation in anaerobic, anoxic and aerobic environment is given in equation 2.1a, 2.1.b and 2.1.c with generation of final product bicarbonate, water and carbon dioxide etc (Fang et al. 1996; Chakraborty and Veeramani, 2006).

- +

6 5 2 3

C H OH+ 17H O6HCO + 34H (Eq.2.1a)

- - +

6 5 3 2 3 2

C H OH+ 5.6NO + 0.2H O6HCO + 0.4H + 2.8N (Eq.2.1b)

6 5 2 2 2

C H OH+ 7O 3H O+ 6CO (Eq. 2.1c)

∆G^o(aq)for reaction shown in equation 2.1 (a), 2.1 (b) and 2.1 (c) are (-) 2549 kJ.mol^-1, (-) 2802 kJ.mol^-1 and (-) 2866 kJ.mol^-1, respectively, suggesting phenol degradation in anaerobic, anoxic and aerobic environments are thermodynamically favorable and aerobic treatment of phenol is the most favorable.

2.3.4.2 Thiocyanate degradation

Thiocyanate wastewater is usually treated by an activated sludge process, where microbial activity degrades this substance as source of nitrogen, sulfur carbon and, energy (Kim and Katyama, 2000). Thiocyanate degradation capacity was initially reported to be limited to strains of neutrophilicThiobacilli sp. by various researchers (Katayama & Kuraishi, 1978;

Smith and Kelley, 1988). However, several new thiocyanate-oxidizing bacteria are identified being capable of growth on thiocyanate at high pH and presence of high concentration of salt (Sorokin et al. 2001). Two major pathways for thiocyanate metabolism have been identified. The first reaction pathway involves carbonyl sulfide as an intermediate, whereas cyanate is an intermediate in the second pathway. Hung and Palvosthis, (1997) reported thiocyanate biodegradation in aerobic condition to proceed as follows: first, thiocyanate is hydrolyzed to cyanate (OCN^-) and sulfide (S^2-); second, cyanate is hydrolyzed to ammonium (NH4+) and bicarbonate (HCO3-) ions; finally, sulfide is oxidized to sulfate (SO42-). Thus, the overall degradation reaction is expressed as equation (2.2a).

- - + 2- +

2 2 3 4 4

SCN + 3H O + 2O HCO + NH + SO + H (Eq. 2.2a) (∆G^o(aq)= -824.65 kJ.mol^-1)

In absence of oxygen, mineralization of SCN^- generates sulfide, ammonia and HCO3- as shown in equation 2.2 (b) (Hung and Pavlostathis, 1997). Not much information is available about the possibility of anaerobic growth with thiocyanate.

- + 2-

2 2 4

SCN + 2H O CO + NH + S (Eq. 2.2b)

(∆G^o(aq)= -9.1 kJ.mol^-1)

An early publication of De Kruyff et al. (1957) reported that Thiobacillus denitrificans can grow with thiocyanate, aerobically or anaerobically, in the presence of nitrate as the electron acceptor, reducing the latter completely to N2, while Thiobacillus thioparus only reduced nitrate to nitrite in the presence of thiocyanate. Andreoni et al. (1988) reported thiocyanate-dependent denitrification by a mixed bacterial population in a thiocyanate waste-treatment plant. Sorokin et al. (2007) reported anaerobic thiocyanate oxidation by denitrifying isolates proceeded through intermediate cyanate, similar to aerobic thiocyanate degradation utilizing halo-alkaliphiles genus Thioalkalivibrio with ammonia, sulfate and nitrogen gas as the final products as shown in equation 2.2 (c-e).

- - 2- - +

3 2 4 2

5SCN + 8NO + H O5SO + 4N + CNO + 2H (Eq. 2.2c)

- - +

3 3 2

5CNO + 5HCO + 10H 5NH + 10CO (Eq. 2.2d)

- - + - 2-

3 2 3 4 3 2 2

SCN + 1.6NO + 0.2H O+ 1.6H + HCO SO + NH + 2CO + 0.8N (Eq. 2.2e) (∆G^o(aq)= -824.01 kJ.mol^-1)

There are differences between autotrophic and heterotrophic metabolism of thiocyanate as well as differences within each group of organisms in anoxic environment. The autotrophic isolate studied by Youatt (1954) used the cyanate pathway, whereas the isolate studied by Katayama and Kuraishi, (1978) used the carbonyl sulfide pathway. Mason et al. (1994) isolated two heterotrophs Acinetobacter jeunii and Pseudomonas fluorescens that were able to metabolize thiocyanate. They suggested that the former organism used the carbonyl sulfide pathway and the latter the cyanate pathway. Sorokin et al. (2007) isolated a new species from a hypersaline lake and named Thiohalophilus thiocyanooxidanswhich was an incomplete denitrifier able to use thiocyanate as an electron donor and nitrite as an electron acceptor. Thiohalophilus thiocyanooxidans followed both carbonyl sulfide pathway and cyanate pathway for thiocyanate degradation through generation of carbonyl sulfide or cyanate as intermediate (Sorokin et al. 2007).

2.3.4.3 Nitrogen removal (a) Ammonia

Conventional microbial nitrogen removal is based on autotrophic nitrification and heterotrophic denitrification. The removal involves (i) aerobic nitrification (i.e., the conversion of NH4+ to NO2- and further to NO3- by ammonia oxidizing bacteria (AOB) and nitrite oxidizing bacteria (NOB), respectively) with molecular oxygen as the electron acceptor and (ii) anoxic denitrification i. e., the reduction of nitrate or nitrite to N2, mostly catalyzed by heterotrophic bacteria. AOB and NOB are slow growers and influenced by many environmental factors such as pH, temperature, C/N ratio, unsaturated fatty acids, dissolved oxygen, ammonia, and nitrite concentrations (Stephen et.al. 1998; Svenson et.al.

2000; Avrahami et.al. 2003). The fraction of nitrifying bacteria in the biomass is important in nitrogen removal systems, since nitrifiers are considered poor competitors for oxygen compared to heterotrophs (Rittmann et. al. 1999). The relevant nitrification reactions are as shown in equation 2.3 (a) - (c)

+ - +

4 2 2 2

NH + 1.5O NO + 2H + H O (Eq. 2.3a)

- -

2 2 3

NO + 0.5O NO (Eq. 2.3b)

+ - +

4 2 3 2

NH + 2O NO + 2H + H O (Eq. 2.3c)

(∆G^o(aq)= -204.54 kJ.mol-¹)

(b) Nitrate and nitrite removal

Biological denitrification is a reliable method for nitrogen removal from wastewater. The anoxic denitrification (i.e., the conversion of NO3- and NO2- to gaseous nitrogen) is accomplished with a variety of electron donors in presence of heterotrophic denitrifying bacteria using nitrate/nitrite as electron acceptor (Grabinska-Loniewska, 1991; Tam et al.

1992; Akunna et al. 1993). This process results in simultaneous removal of nitrate/nitrite and organic/ or inorganic matter used as electron donor.

The anoxic denitrification involves the following reactions as shown in equation [2.3(d) and (e)]

- + - -

3 2 2

2NO + 10H + 10 e N + 2OH + 4H O (Eq. 2.3d) (∆G^o(aq)= -1040.04 kJ.mol^-1)

- + - -

2 2 2

2NO + 6H + 6 e N + 2OH + 2H O (Eq. 2.3e) (∆G^o(aq)= -714.36 kJ.mol^-1)

As nitrification and denitrification are carried out under different conditions and by different microorganisms, experience shows that these processes have to be separated in time or space to function effectively. The conventional nitrification/denitrification reactions have been known for a long time. The nitrification reaction consumes a large amount of oxygen, requiring 4.2 g of oxygen for each gram of ammonium nitrogen nitrified (EPA, 1975). During denitrification, the requirement of organic carbon is significant. For example, 2.47 g of methanol is required per gram of nitrate nitrogen for complete denitrification (McCarty et al. 1969). Therefore if the electron donor is present in adequate amount in wastewater there is feasibility of high denitrification efficiency with high nitrogen removal.

2.3.4.4 Pyridine degradation

Chemically, the pyridine ring is susceptible to reduction and these characteristic have been exploited by microorganisms for evolving mechanisms for pyridine ring degradation (Liu et al. 1994). Two general strategies of bacterial pyridine degradation involve (i) hydroxylation reactions, followed by reduction, and (ii) (aerobic) reductive pathway not initiated by hydroxylations (Kaiser et al. 1996).

Pyridine mineralization in anaerobic, anoxic and aerobic conditions is reported to be occurred as shown in Equation 2.4 a-c (Liu et al. 1994)

+ +

5 5 2 2 4 4

C H N + 4.5H O + H  2.25CO +2.75 CH +NH (Eq. 2.4a) (ΔG^o(aq) = -169 kJ/mole)

+ +

5 5 3 2 2 2 4

C H N + 4.4NO + 5.4H ^  5CO +2.2N +3.2H O +NH (Eq. 2.4b) (ΔG^o(aq) = -2277 kJ/mole)

+ +

5 5 2 2 2 4

C H N + 6O + 3H  5CO ++ 2H O + NH (Eq. 2.4c) (ΔG^o(aq) = - 2682.4 kJ/mole)

Adav et al. (2007) observed no inhibitory effect of pyridine on phenol degradation in aerobic environment upto pyridine concentration of 1500 mg/L. Li et al. (2001) reported complete removal of pyridine in anoxic environment within 12-24 h at initial pyridine concentration of 20-100 mg/L in absence of any other pollutant. Li et al. (2009) also reported complete removal of pyridine by an aerobic strain Streptomyces within 8 days from initial concentration of 2000 mg/L at pH of 7. Sun et al. (2011) observed simultaneous degradations of pyridine and phenol by an aerobic strain Rhodococcus, using phenol as carbon source and pyridine as the nitrogen source. Pyridine and its derivatives are reported to be toxic to the anaerobic process (Gijzen et al. 2000). Blum et al. (1986) reported pyridine have inhibitory effect on anaerobic digestion in presence of phenol.