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Review of Literature 20

2.1 Textile Industry in Rajasthan The industrial development of Rajasthan began between 1950 and 1960.

Large and small scale industries sprung up in the Kota, Jaipur, Udaipur, Bhilwara and

other industrial estates of Rajasthan. The key industries of Rajasthan include textile,

rugs, woolen goods, vegetable oil and dyes. Heavy industries include copper and zinc

smelting and the manufacture of railway rolling stock. The private sector industries

include steel, cement, ceramics and glass wares, electronic, leather and footwear,

stone and chemical industries. Textile sector holds for about 20% accountability of

investment made in the state. Rajasthan contributes over 7.5 percent of India's

production of cotton and blended yarn (235,000 tonnes in 2002-03) and over 5

percent of fabrics (60 million sq meters). The state holds a leading position in

spinning of polyester viscose yarn & synthetic suiting (at Bhilwara) and processing,

printing & dyeing of low cost, low weight fabric (at Pali, Balotra, Sanganer and

Bagru). In total the production of textiles in Rajasthan accounts for 21.96 % of the

total national production scenario. Large quantities of spun yarn and hence is the

fourth largest producer in India (http://www.mapsofindia.com/maps/rajasthan/

rajasthanindustry.htm). Besides, Jaipur is also a well-known center for manufacturing

garments primarily for exports. Some knitting units are in process of setting up their

ventures at Neemrana. The economic growth, the state of Rajasthan has witnessed

with the textile sector is depicted in Figure 9.

Figure 9: Industrial Export from Rajasthan (Rs. In crore) with special emphasis to textile industry (Source: Commissioner, Department of Industry; Rajasthan) (State of Environment report for Rajasthan: 2007)

0

200

400

600

800

1000

1200

1400

1600

1999-20002000-20012001-20022002-20032003-2004

Rs.

(in

Cro

res)

Annual industrial Export of diffrent industries

Textile

Gems and Jewellery

Engineering

Marble and Granite

Electrical and Electronics

Wool and Woolens

Chemical and Allied

Drugs and Pharmaceuticals

Plastics and Linoleum

Handicrafts

Handloom


The figure above clearly indicates the significance of textile industry in

Rajasthan. Keeping the fact in mind, that a coin always has two facets; likewise,

textile industry is has proven to be a boon and a bane for societal development. On a

positive front, it is a growing sector with respect to economic growth of Rajasthan

and on the negative side; the industrial growth is always been associated with

augmentation in environmental pollution particularly related to inland water, air,

noise and soil; if not stringently monitored. The industries in the state have been

categorized into red and orange depending upon the environmental pollution:-

•••• Highly polluting industries have been classified as red category which

includes man made fiber manufacturing (Rayon & Others). Yarn and textile

processing industries which involve scouring, bleaching, dyeing, printing or

any effluent / and process that relate to generation of emission.

•••• The moderately polluting industries are classified as orange category and

include yarn and textile manufacturing /processing not involving scouring,

bleaching, dyeing, printing or any effluent /emission generating process

including spinning/weaving unit.

As per the categorization of Rajasthan State Pollution Control Board

(RSPCB), textile units have been included both in red and orange categories, so as to

cover these entire units under the pollution control system (RSPCB office order No.

F.14 (57) Policy / RSPCB/ Plg/ 9219 – 9259 dated 21.12.2010).

Worth mentioning is the fact that textile industry occupies the top notched

position in contributing to the hazardous wastes amongst all industries (State of

Environment Report for Rajasthan, 2007). Table 5 represents the sector specific

generation of hazardous waste by industrial units.


Table 5: Sector – Specific Number of Hazardous Waste Generating Industries

S.No Product Number of Units SS/Med* LS** Total

1 Asbestos and it’s products - 16 16

2 Benefiaciation of ore 1 - 1

3 Cement 14 - 14

4 Ceramic 2 - 2

5 Chemical 23 36 59

6 CETP’s 3 1 4

7 Drug and Pharmaceuticals 8 23 31

8 Dye and it’s intermediates 3 33 36

9 Electronics 2 1 3

10 Engineering 41 41 82

11 Fertilizer 14 1 15

12 Mining 6 2 8

13 Paints, Varnish and ink 2 4 6

14 Pesticides 2 36 38

15 Petroleum Drilling/storage 19 - 19

16 Power plants 3 - 3

17 Primary production of copper/zinc /lead 3 - 3

18 Secondary production of copper/zinc 1 13 14

19 Secondary production of zinc /lead - 32 32

20 Steel rolling mills 2 21 23

21 Tanneries - 9 9

22 Textile 70 19 89

23 Thermal power 2 - 2

24 Waste oil refineries 3 30 33

25 Workshop 2 - 2

26 Miscellaneous 1 1 2 *Small/ Medium Scale **Large Scale (Source - RSPCB Hazardous Waste Report, 2007) Evidently, textile industries are major sources of environmental pollution. As

the textile industries consume large quantities of water and generate waste water in

proportionate order. Pali has always been into limelight as one of the major textile

hubs of Rajasthan which proportionately generates huge volumes of waste water.


It houses around 800 textile units, of these, 200 fall under orange category

and the remaining 600 under the red category, of which 75-80 are non- operational

(Centre for Science and Excellence (CSE), September, 2014). In Rajasthan, the semi

arid and arid belts rely on groundwater resources due to paucity of surface waters

(Bhadra et al. 2013). The deterioration of surface and groundwaters due to

industrial and urban waste has long been recognized (Olayinka, 2004). The

rivers and stream are common sites for disposal of industrial effluents. The textile

dyeing and printing units situated at Pali have been discharging effluents in the

tributary Bandi which owes it’s origin to Luni basin (Khandelwal and Chauhan,

2005). The untreated industrial effluent flows regularly in the dry bed of river about

45 km downstream Pali city (Rathore, 2011). Particularly, the release of effluents

influxed with dye into the water environment is unacceptable because of their colour,

direct release and end products are toxic and may be carcinogenic or mutagenic to life

forms because of benzidine, naphthalene and other aromatic compounds (Suteu et al.

2009; Zaharia et al. 2011).

The groundwater pollution caused by the discharge of untreated and partially

treated textile effluents into river Bandi has been an issue of utmost concern which

awaits immediate and justified addressal. Recent study conducted by The Centre for

Science and Environment (CSE) on pollution caused by by textile mills in Pali was

of the opinion that none of the samples like surface waters, effluents from different

mills assessed for physico-chemical properties were in compliance with the standards

prescribed by Central Pollution Control Board (CPCB) (Times of India, dated

19.09.2014) The depletion in quality of water has an undue effect on human beings

as well as aquatic ecosystem directly or indirectly (Chinda et al. 2004; Ugochukwu,

2004; Emongor et al. 2005).

National Green Tribunal (NGT) had strictly ordered 550 industrial units to

shut down on account of water pollution caused by textile industries in Pali in March,

2014 (Rajasthan Patrika, dated 03/ 06/ 2014).

A study conducted by Geological Survey India (GSI) in January, 2014 was

aimed to assess the nature of discharging effluents from textile industries and it’s


effect on natural resources in their vicinity. This study remarkably revealed that the

effluents had caused irreparable damage to agriculture and groundwater significantly

(Times of India, dated 2.1.2014).

Nevertheless, an initiative of Times of India (TOI) and RSPCB led to the

closure of all textile units in Pali town on account of excessive effluent generation

and it being drained directly into river Bandi (November, 2012). Previously, a similar

initiative taken by Central Groundwater Board (CGB) which was taglined as “In

Pali, a river being killed” was of opinion that due to excessive release of effluent; 50

Million litres per Day (MLD) against the permissible which is 36 MLD, resulted into

alkalinity of groundwater and hence the wells were rendered unfit for drinking

purposes. (September, 2012). A brief account of textile units in Pali city is given in

Table 6.

Table 6: Distribution of Textile Industries in diff erent localities of Pali city

(Rathore, 2012)

S.No Location Number of Units

1 Mandia Road Industrial Area Phase III 473

2 Industrial Area I and II 121

3 Main Mandia Road, Ramdev Nagar, Residential Area, Gandhi Nagar

173

4 Sumerpur Road, Bajrang Bari, Ramleela Maidan, Sojat Road, Village Mandia

95

5 Composite Unit- Maharaja Shri Umaid Mills Ltd., Pali, Marwar

1

6 Total 860

The tools like Geographic Information System (GIS) and Remote Sensing had

been used to assess groundwater pollution due to industrial effluents from textile

units around Pali city. Indiscriminate discharge of these effluents in the past two

decades caused severe damage to the agricultural land and groundwater resources

within 3 km buffer zone of the river. Remote Sensing and Geographic Information

System (GIS) techniques are found to be better tools in assessing the damage in


agricultural crops during 1997 and 2006 with the help of high resolution satellite

images (Bhadra et al. 2013).

2.2 Characteristics of textile effluents and CETP an approach to combat

water pollution

Effluent discharged from the textile industries has variable characteristics in

terms of pH, dissolved oxygen, organic and inorganic chemical content. Together

with industrialization, awareness towards the environmental problems arising due to

effluent discharge is of critical importance. Pollution caused by dye effluent is

mainly due to durability of the dyes in wastewater (Jadhav et al. 2007).

Textile industries are large industrial consumers of waters as well as

producers of wastewaters and contribute to a total of 14% of industrial pollution in

India (business.mapsofindia.com).

This industry involves processing or conversion of raw material/fabric into

finished cloth involving various processes/ operations and consumes large quantities

of water and produces extremely polluted waste effluents (Asia et al. 2006; Andre et

al. 2007). The effluent is primarily composed of extremely diversified inorganic

compounds, polymers and organic substances (Mishra and Tripathy, 1993) in

dissolved, colloidal or suspended forms and is typically coloured due to the presence

of residual dyestuffs (Patel and Pandey, 2008). The effluent is characterized by a

high BOD (Wang et al. 2007; Annuar et al. 2009) which depletes the dissolved

oxygen; both of which are aesthetically and environmentally unacceptable. The

release of these highly coloured effluents can damage directly the receiving waters

(Chen et al. 2003). Because of the stable structure of synthetic dyes used in textile

industries, the retention of colour remains a permanent feature and is not removed by

municipal waste water systems (Shaul et al. 1991; Robinson et al. 2002; Forgacs et

al. 2004; Couto, 2009). It is estimated that globally 280 000 tons of textile dyes are

discharged in textile industrial effluent every year (Jin et al. 2007).

The risk factors for pollution are primarily associated with the different

production steps in which wet processes include scouring, desizing, mercerizing,


bleaching, dyeing and finishing. Desizing, scouring and bleaching processes produce

large quantities of wastewater (Yusuff et al. 2004) and is characterized by high pH,

temperature, detergents, oil, suspended and dissolved solids, dispersants, leveling

agents, toxic and non-biodegradable matter, colour and alkalinity (Nese et al. 2007;

Govindwar et al. 2012; Sharma et al. 2013). Important pollutants in textile effluent

are mainly recalcitrant organics, colour, toxicants and surfactants, chlorinated

compounds. Extreme fluctuations in COD, BOD, pH, colour and salinity of textile

effluents are reported which may be attributed to 8000 different types of chemicals

like salts, metals, surfactants, organic processing assistants, sulphide and

formaldehyde (Zollinger, 1987; Talarposhti et al. 2001; Kanu and Achi, 2011;

Jamaluddin and Nizamuddin, 2012; Sharma et al. 2014). In cases most of the

BOD/COD ratios are found to be around 1:4, indicating the presence of non-

biodegradable substances (Arya and Kohli, 2009). The process of adding colour to

the fibres is known as dyeing which normally requires large volumes of water not

only in the dye bath, but also during the rinsing step. The water let out after the

production of textiles contains a large amount of dyes and other chemicals which are

harmful to the environment. The level of toxicity or harmfulness of the textile

effluents varies among industries (Kiriakidou, 1999). The effluent generated during

the process of rinsing and in dye bath (Dos Santos, 2005) always contains a portion

of unfixed dye (1mg/l) (Gonclaves, 2000) which imparts strong colour to the effluent

(Soares et al. 2006). Rathore et al. 2011 was of opinion that, the colour of textile

effluents is usually subject to daily and seasonal variations. The routine of production

and the other characteristics vary according to the nature of textile manufactured.

The concentration of these unfixed dyes and trace metals like Cr, As, Cu and Zn was

found to be high in textile wastewater (Eswaramoorthi et al. 2008). The presence of

these heavy metals affects the aesthetic value, transparency of water and solubility of

gases in lakes, rivers and other water bodies (Agarval, 2001). Dyes in water give out

a bad colour and can cause diseases like haemorrhage, ulceration of skin, nausea,

severe of skin irritation and dermatitis (Nese et al. 2007). The penetration of sunlight

can be blocked from water surface preventing photosynthesis and increased BOD

which reduces the process of re oxygenation and growth of photoautotrophic


organisms is hampered by and large (Laxman, 2009). Figure 10 represents schematic

representation of unit processes in a textile industry.

Figure 10: Schematic representation of unit processes in a textile industry

Sharma et al. in 2014 explored one of the famous textile hotspots of Jaipur

region, Bagru. This area is internationally famous for bagru prints. The study was

conducted in months of July, August and September 2013 to assess the quality of

water resources affected by textile effluents in terms of pH, electrical conductivity,

total dissolved solids, total suspended solids, dissolved oxygen, chemical oxygen

demand and chloride. The study was found to have non adherence of effluent

standards for discharge of textile effluent prescribed by RSPCB. Thoker et al. in

2012; Jamaluddin and Nizamuddin in 2012, had conducted similar studies to assess

the quality of textile effluents in Chittagong, Bangladesh.

Gray cloth

Settling

Bleaching and washing

Dyeing and printing

Steaming

Washing

Finishing

Folding


Rathore in 2011 and 2012 conducted a study to assess the quality of river

Bandi with special emphasis to parameters like COD, suspended solids, chloride,

sulphate, and chromium and compared with standards laid down by Bureau of Indian

Standards (BIS). A significant reduction in pollution load of river Bandi during

monsoon season was observed which may be attributed to flow of freshwater

upstream of Pali city.

Joshi and Santani in 2013 examined the physicochemical properties like

temperature, pH, colour, odour, COD, BOD, EC, TSS, TDS and inorganic elements

like sodium, potassium and phosphorus levels of effluent discharged by Sumukh

textile mill, Vapi, Gujarat, India.

Textile industry is not only confined to northern and western parts of India,

Tamil Nadu has also emerged as a potential state for the same. The textile mills are

concentrated in Coimbatore, Tirupur, Salem, Palladam, Karur and Erode. Tamil

Nadu has around 3, 50,000 power looms manufacturing cotton fabrics and accounts

for about 30 per cent of India's export of textile products. The Erode district in Tamil

Nadu is well known for marketing of textile products of handloom, powerloom and

readymade garments (Chellasamy and Karuppiah, 2005).

Textile effluents generated by textile dyeing industries of Tirupur and

Coimbatore, Tamil Nadu were thoroughly studied for their physicochemical

parameters. The values of each parameter were found to be well above the laid

standards (Rao and Prasad, 2011; Rajeswari et al. 2013a).

Continuous discharge of untreated textile effluent into Noyyal river basin has

deteriorated the quality of water and has rendered it unfit for irrigational purposes

(Mahalakshmi and Saranathan, 2014). Sundar and Saseedharan in 2006 and 2008

investigated the groundwater quality in Noyyal river basin. The result demonstrated

that due to industrialization, high level of dyeing effluents in and around the river is

responsible for deterioration of groundwater quality (Palanysamy et al. 2008). The

hydro-geochemistry analysis for groundwater of the river basin has been examined to

characterize the groundwater (Sunder and Saseedharan, 2008). Usharani et al. 2010


has also analyzed the physico-chemical and bacteriological characteristics of Noyyal

river and it’s groundwater quality. Santhosh and Devi in 2010 have integrated land

use and land cover detections using remote sensing and GIS for Coimbatore district.

Zahiruddin and Kandasamy, 2013 studied the germination profile of Cajanus cajan

(Red gram), Phaseolus aureus, (Green gram) and Vigna mungo (Black gram) from

the family Fabaceae. When irrigated by textile effluent taken in varying

concentration from three different textile mills of Kanchipuram, Tamil Nadu, India,

significant differences in germination profile were observed.

Studies pertaining to toxicity of textile effluents on soil microflora have been

reported (Kaur et al. 2010). The parameters like germination, delay index,

physiological growth and plant pigments are affected because utilization of textile

effluent for irrigation purposes is rendered unfit (Garg and Kaushik, 2008). In India

the untreated sewage water and wastewater from textile industries containing variety

of chemicals such as aniline, caustic soda, acids, bleaching powder including heavy

metals, are used in irrigating agricultural fields, for growing vegetables and other

crop plants owing to paucity of water resources (Chandra et al. 2009). The effluent

water takes the dissolved toxicants to crop plants and it’s consumers (Kala and Khan,

2009). Vegetables grown in the agricultural fields using untreated textile wastewater

for irrigation are adversely affected. Use of wastewater alters the nutritional value of

the vegetables grown there and in long run consumption of such vegetables may

impose health hazards in human beings, which is a matter of concern. Irrigating

crops with wastewater containing high nutrient load often leads to prolific weed

growth and undesirable growth of microorganisms (Modasiya et al. 2013). Debnath

et al. 2014 studied the effect of untreated textile effluent generated by textile

industries of Sanganer, Jaipur on irrigation of crops. A similar study conducted on

untreated textile effluent in Sanganer was aimed to assess the bioaccumulation of

different heavy metals by Lycopersicon esculentum (wild tomato) when a varying

concentration of effluent was used for irrigation purposes (Khan and Marwari, 2010).

On the contrary, it has also been reported that waste water reuse for

irrigational purposes is a strategic approach for water pollution abatement ( Bhabindra,

2003) and certain components like nitrogen and phosphorus (Zahiruddin et al. 2013)


in textile effluent when present in lower concentration are beneficial for the plant

growth (Kannan and Upreti, 2008) .

Considering, water pollution caused by textile industries as a global concern,

the physico-chemical properties of textile effluents have been envisaged in different

parts of the world.

Imtiazuddin et al. in 2012 conducted an extensive study for a period of four

months i.e. November, December, January and February in 2009-2010. The samples

were collected from different textile mills in Pakistan and analyzed for various

parameters such as TDS, COD, BOD, pH, EC and heavy metals like cadmium,

chromium, copper, iron, manganese, nickel, potassium, phosphorous, sodium,

sulphur, zinc.

The wastewater treatment plants of 11 textile mills (Turkey) for woven fabric

and knit fabric finishing industry were investigated (Nese et al.2007) and treatment

plants were evaluated for their performance by in situ inspections and analyses of

influent and effluent samples. The analysis of effluent indicated non-adherance with

the IWSA (Istanbul Water and Sewerage Administration) discharge standards.

Keeping the problem of water pollution caused by textile industries as a

major target, CETP had been proved to be a remedial alternative to minimize the

excessive problem caused.

Forty percent of the industrial wastewater generated in India comes from

small size industries. With the adoption of the Water Act (1990), small size industries

should treat their effluent in order to limit pollution concentration at the minimum

acceptable standards laid down by the State Pollution Control Boards (SPCB).

Nevertheless, the size of these facilities makes the installation of a standard Effluent

Treatment Plant (ETP) unaffordable (Singh et al. 2011). The onset of CETP’s was

the pioneering effort of Ministry of Environment and Forests (MoEF) aimed at

treating effluents generated by different small scale industries; to devise remedial

measures to which may result into generation of an eco friendly effluent. The first

CETP came into existence in 1985 in Jedtimetla near Hyderabad, Andhra Pradesh to


treat waste waters generated from pharmaceutical and chemical industries

(Maheswari and Dubey, 2000).

The concept of CETP has four underlying objectives given in Figure 11.

Figure 11: Objectives determining the viability of a CETP

It has been clearly shown, that compared to individual ETP’S, CETP are

more cost effective in reaching the effluent concentration standards (Pandey and

Deb, 1998). Taking this into consideration, the massive problem of water pollution

caused by textile industries in Pali, CETP’s have been installed with the aim to

pollution abatement (Dhariwal and Ahmed, 2009). A brief overview of different

CETP’s in Pali has been reported earlier (RSPCB, 2007) (Table 7).

Table 7: Different CETP’s installed for treating textile effluents in Pali city

(RSPCB, 2007)

S.No Year Capacity Purpose Present

Status Cost Performance Problems Units

PWPCTRF

Unit 1 1983 5MLD Textile Operational 38

Lacs Non-compliance

with standards

Needs up gradation

435 in all

Unit II 1997 8.4MLD Textile Operational

Unit III 1999 9MLD Textile Operational 7.5

Crores

To achieve economics of scale in waste to minimise pollution cost

To circumvent the problem of space as CETP is a centralised facility

To reduce the problems of monitoring for the pollution control boards.

To organize the disposal of treated wastes and sludge and to improve the recycling and reuse possibilities.


Assessment of CETP at regular intervals for inlet and outlet parameters has

been an important aspect to be explicitly considered. This justifies the functioning of

CETP but also correlates the significant reduction in pollution load after treatment

(Kore and Kore, 2012). The CETP located in Kagal Five Star MIDC, Kagal Nagpur

was studied for both inlet and outlet parameters like pH, TSS, TDS, oil and grease,

COD and BOD.

Desai and Kore in 2011 conducted a study to evaluate the performance

efficiency of an ETP of a textile industry located in Kagal-Hatkanangale

(Maharashtra Industrial Development Corporation) MIDC area, Kolhapur

(Maharashtra). Both inlet and outlet parameters like pH, BOD, COD and TDS were

taken into account.

Patel et al in 2013 undertook a study to evaluate efficiency of an ETP of a

textile industry located in Narol, Ahmedabad (Gujarat). The parameters such as pH,

BOD, COD, TDS, TSS and ammonical nitrogen (NH3-N) were considered for both

inlet and outlet samples.

Rajeswari et al. in 2013b analysed and compared the efficiency of three

textile focused CETP’s located at Telugupalayam, Tamil Nadu for it’s outlet

parameters like temperature, pH, colour, odour , EC, TDS , TDS, COD, BOD, total

hardness, total alkalinity, chlorides and sulphates.

Singh et al. in 2011 attempted to evaluate the performance of CETP’s in Pali

based on criteria like design, operation, maintenance and administration and

contrasted the outcomes with the standards prescribed by Rajasthan State Pollution

Control Board (RSPCB).

Performance of CETP, Mandia Road, Pali was evaluated not only for it’s

inlet and outlet parameters but all stages of treatment like primary clariflocculator,

aeration tank and secondary clarifier were analysed for physicochemical parameters

like pH, BOD, COD, TSS for a period of 4 months on a regular basis (Dhariwal and

Ahmad, 2009).


Along with the liquid effluent, sludge, a byproduct of effluent treatment can

create a menace for environment, because it contains many toxic chemicals released

by textile industries. Therefore, it requires proper and safe disposal (Patel and

Pandey, 2008).

Patel and Pandey in 2008, identified 4 CETP’s catering to textile industries

Mandia Road, Pali; Balotra, Barmer; Manikapuram and Mannarai in Tirupur, Tamil

Nadu and thoroughly investigated physico-chemical properties and heavy metals

present in sludge generated by textile industries.

A continuous growth in the demand of water in agricultural, industrial,

municipal and recreational sectors had been witnessed throughout the country. Due

to paucity of surface water, in semi arid and arid regions, there is a dependency on

groundwater resources to a great extent. In many areas in Rajasthan, alarming trends

of groundwater contamination have been reported due to discharge of untreated

industrial effluents (Rathore, 2012; Khan et al. 2014). Since 1980s, Pali city area is

facing severe problem of pollution due to discharge from wide spread dyeing

industry into the Bandi river and industries here discharge a variety of chemicals,

dyes, acids and alkalis besides heavy metals and other toxic compounds. Due to the

non-biodegradable nature and longtime persistence in the environment, the toxic

waste often accumulates through trophic levels causing a deleterious biological effect

(Bhadra et al. 2013). It has been mentioned in literature that synthetic dyes are the

major culprit’s which not only pollute the environment but are difficult to treat owing

to their recalcitrant nature (Pereira and Alves, 2012).

2.3 Synthetic Dyes: The main culprits

Colourants (dyes and pigments) are important industrial chemicals.

According to the technological nomenclature, pigments fall into the category of

colourants which are insoluble in the medium to which they are added, whereas dyes

are soluble in the medium. The world’s first commercially successful synthetic dye,

named mauveine, was discovered accidently in 1856 by William Henry Perkin. Such

substances known as colourants with considerable colouring capacity are widely

employed in the textile, pharmaceutical, food, cosmetics, plastics, photographic and


paper industries (Carneiro et al. 2007). Dyes are classified according to their

application and chemical structure and are composed of a group of atoms known as

chromophores, responsible for imparting colour to the dye. This fact is attributed to

(1) absorption of light in the visible spectrum (400–700 nm), (2) presence of atleast

one chromophore (colour-bearing group) or the colour enhancer (Pereira and Alves,

2012), (3) have a conjugated structure of alternating double and single bonds, and (4)

electrons, which possess resonance and stabilize organic compounds (Abrahart,

1977). Exclusion of any of these factors leads to loss of colour. These chromophore-

containing centers are based on diverse functional groups, such as azo,

anthraquinone, methine, nitro, arylmethane, carbonyl and others. In addition,

electrons withdrawing or donating substituents so as to generate or intensify the

colour of the chromophores are denominated as auxochromes. The most common

auxochromes are amine, carboxyl, sulfonate and hydroxyl (Christie, 2001; Dos

Santos et al. 2007; Prasad and Rao, 2010).

The colour imparted to fibers is permanent, such that the colour is not lost

(wash fast) when exposed to sweat and light (light fast), water and many chemical

substances including oxidizing agents and also to microbial attack (Rai et al. 2005;

Saratele et al. 2011). Textile effluents primarily consist of synthetic dyes as

colourants (Pereira and Alves, 2012) that are amongst the major priority pollutants

that accumulate in environment as recalcitrant xenobiotic compounds due to

continuous industrial inputs that have created a serious impact on environment.

Synthetic dyes in particular are stable, toxic organic compounds (Hao et al.

2001; Hu et al. 2009) which are mostly carcinogenic like benzidine and other

aromatic compounds (Przytas et al. 2012) mutagenic (Weisburger, 2002; Mathur et

al. 2007) and pose serious health hazards which may remain persistent for longer

durations (Forgacs et al. 2004; Shedbalkar et al. 2008; Sinha et al. 2009; Kumar and

Praveen, 2011). The end users are the textile industries and account for about two-

thirds of the total dye market (Elisangela et al. 2009) and to reciprocate, generate

enormous volume of aqueous wastes (Sen and Demirer, 2003; Dos Santos, 2005;

Ben Mansour et al. 2012). The dye effluents are discharged from the dyeing process

as some of dye remains unfixed to the fibres (Wong and Yu, 1999; Fang et al. 2004;


Lavanya et al. 2014) and is therefore visually identifiable (Kilic et al. 2007) and in

some cases concentration as low as 1 mg/ml is also detectable (Pandey et al. 2007;

Wijetunga et al. 2010).

A need to establish an economic and effective way of dealing with the textile

dyeing waste in response to the production costs of different unit operations is

necessitated (Park et al. 2007). The Ecological and Toxicological Association of the

Dyestuffs Manufacturing Industry (ETAD) was inaugurated in 1974 (Chequer et al.

2011) with the goals of minimizing environmental damage, protecting users and

consumers and cooperating with government and public concerns in relation to the

toxicological impact of their products (Anliker, 1979). Government legislation is

becoming more and more stringent, especially in more developed countries,

regarding the removal of dyes from industrial effluents (Robinson et al. 2001; Kuhad

et al. 2004; Ogubue and Sawidis, 2011) and hence the removal of bioaccumulated

xenobiotic dyes from textile effluents is a matter of international concern (Gienfrada

and Rao, 2008).

Dyes are usually classified by their Colour Index (CI), developed by the

Society of Dyes and Colourist (1984). Accordingly, dyes are firstly listed by a

generic name based on it’s application to fibre and it’s colour, followed by assigning

a 5-digit CI number based on it’s chemical structure, if known (O'Neill et al.1999).

Examples include Acid Blue 120 (26400), Reactive Red 4 (18105), and Mordant

Yellow 10 (1401). They can be grouped in different classes: acid, basic, direct,

disperse, metallic, mordant, pigment, reactive, solvent, sulphur and vat dyes, which

reflects their macroscopic behaviour and also their prevailing functionalities. They

are used in accordance to their compatibility with the type of textile substrate being

processed Acid, direct and reactive dyes are water-soluble anionic dyes; basic dyes

are cationic, whereas disperse, pigment and solvent dyes are non-ionic (Hao et al.

2000). Disperse dyes are sparingly soluble in water for application in hydrophobic

fibres from an aqueous dispersion. They are often of anthraquinone and sulfide

structure, with many -C =O, -NH- and aromatic groups (Fu and Viraraghavan, 2001).

Based upon the chemical structure or chromophore, they may be catergorized like

azo (monoazo, disazo, triazo, polyazo), anthraquinone, phthalocyanine, triarylmethane,


diarylmethane, indigoid, azine, oxazine, thiazine, xanthene, nitro, nitroso, methine,

thiazole, indamine, indophenol, lactone, aminoketone, hydroxyketone stilbene and

sulphur dyes. Most of the mordant dyes are anionic in nature with few exceptional

cationic forms. In aqueous solution, a net charge on anionic dyes is attributed to the

presence of sulphonate (S03-) groups, while the presence of net positive charge on

cationic dyes is due to protonated amine or sulfur containing groups. Solubility in a

dye molecule is a trait of sulphonic group. Disperse vat dyes (indigo) are insoluble in

water; however in presence of reducing conditions, they are convertible into a 'leuco'

form (soluble in alkaline aqueous solutions) which are penetrated into the fibres

during dyeing process. The dyes with metal-complex exhibit higher light and wash

fastness because of presence of transition metals, such as chromium, copper, nickel

or cobalt that modify the surface chemistry between the dye molecule and the fabric

(Hao et al. 2000).

Based on the application to fibre, different classes of dyes are used Reactive

dyes are most commonly used as they can be applied to both in natural (wool, cotton,

silk) and synthetic (modified acrylics) fibres (O'Neill et al. 1999). The characteristic

feature of Reactive dyes which makes them different from other class of dyes is that

their molecules contain one or more reactive groups capable of forming a covalent

bond with a compatible fibre group. They possess the properties of high wet-fastness,

brilliance and range of hues (Hao et al. 2000). Their usage as synthetic fibres has

enhanced. Acid and basic dyes are used for dyeing all natural fibres (wool, cotton,

silk) and some synthetics (polyesters, acrylic and rayon). Direct dyes are applied

directly to cellulose fibres. Furthermore, they are used for colouring rayon, paper,

and leather and to certain extent nylon. The application of mordant dyes is limited to

the colouring of wool, leather, furs and anodised aluminium. Solvent dyes are used

for colouring inks, plastics, and wax, fat and mineral oil products. The direct dyes are

the second largest dye class in the colour index with respect to the amount of the

dyes. Direct dyes are relatively large molecules with high affinity for cellulose fibres

which is attributed to Van der Waals forces that make them bind to the fibre. About

1600 direct dyes are listed but only 30% of them are in current production. Society of


Dyers and Colourists classifies direct dyes by their dyeing characteristics (Preston,

1986).

Azo dye are the largest group of dyes (Saranraj et al. 2013) with -N=N- as a

chromophore in an aromatic system. There are monazo, disazo, trisazo, and polyazo

dyes depending upon the number of azo-groups present. Diazotisation of a primary

amine, in presence of HCl + NaNO2 at freezing temperature, produces a diazonium

salt which in turn coupled with aromatic compounds, produces an azo-dye (Lavanya

et al, 2014) .The phenomena of diazotization was discovered by a german scientist

Gries in 1858 (Saranraj et al.2013). Of all known dyestuffs in the world, azo dyes

make up about a half, making them the largest group of synthetic colourants and the

most common synthetic dyes released into the environment (Zhao and Hardin, 2007).

Almost one million tons of dyes are annually produced in the world, of which azo

dyes, characterized by an azo-bond (R1–N=N–R2), represent about 70% by weight

(Hao et al.2000).

Azo dyes are the most common synthetic colourants mostly used for yellow,

orange and red hues. They are released to the environment via textile, pharmaceutical

and chemical industries. Structurally, they are characterised by reactive groups that

form covalent bonds with HO-, HN-, or HS- groups in fibres (cotton, wool, silk,

nylon). Approximately, 10–15 % of the dyes are released into the environment

during dyeing of different substrates, such as synthetic and natural textile fibres,

plastics, leather, paper, mineral oils, waxes, and even foodstuffs and cosmetics

(Sponza and Isik, 2005).

Azo dyes are second to polymers in terms of the number of new compounds

submitted for registration in the United States (US) under Toxic Substance Control

Act (TSCA) (Brown and DeVito, 1993). Chemically, the textile azo dyes are

characterized by relatively high polarity and high recalcitrance. Recalcitrance is

difficult to evaluate because of the dependence of degradation on highly variable

boundary conditions (e.g., redox milieu or pH). The azo dyes can accept protons

because of the free electron pair of the nitrogen, and the free electron pair of nitrogen

interacts with the delocalized π-orbital system. Acceptor substituents at the aromatic

ring such as –Cl or –NO2 cause an additional decrease in the basic character of


aminic groups. Donor groups such as –CH3 or –OR (in meta and para position) lead

to an increase in the basicity of aromatic aminic groups. However, donor

substituients in the ortho position can sterically impede the protonation and

consequently decrease the basicity of aminic groups. The azo dyes are characterized

by amphoteric properties when molecules contain additional acidic groups such as

hydroxyl, carboxyl or sulfoxyl substituents. Depending on pH value, the azo dyes

can be anionic (deprotonation at the acidic group), cationic (protonated at the amino

group) or non-ionic. Environmental partitioning is influenced by substituents as well

as the number of carbon atoms and aromatic structure of the carbon skeleton

influence environmental partitioning. Amino group causes a higher boiling point, a

higher water solubility, a lower Henry’s law constant, and a higher mobility in

comparison with hydrocarbons (the amino group can also reduce the mobility by

specific interactions with solids via covalent bonding to carbonyl moieties or cation

exchange) (Börnick and Schmidt, 2006).

Azo dyes are structurally diverse in nature thus they are not uniformly

susceptible to microbial attack (O’ Neill et al. 1999; Couto, 2009) especially in case

of cotton fibres. Azo dyes cannot be degraded under aerobic conditions; under

anaerobic conditions, the azo linkage can be reduced to form aromatic amines which

are colourless but can be toxic, carcinogenic and mutagenic (Cripps et al. 1990;

Pinheiro et al. 2004). Several amino substituted azo dyes including 4-

phenylazoaniline and N-methyl- and N,N-dimethyl-4-phenylazo anilines are

mutagenic and carcinogenic (Yaneva and Georgieva, 2012). Direct dyes constitute

the major part of azo dyes which are primarily used for dyeing cotton fabrics. One

such dye which founds a common use in textile sector with special emphasis to Pali’s

scenario is Direct Red 28 commonly known as CR.

CR is a benzidine-based (Shinde and Thorat, 2013) direct, anionic diazo dye

(Jaladoni-Buan et al. 2010) chemically prepared by coupling tetrazotised benzidine

with two molecules of napthionic acid. Chemically, it is a sodium salt of

benzidinediazo-bis-1- naphtylamine-4 sulfonic acid (Jaladoni-Buan et al. 2010;

Perumal et al. 2012; Shinde and Thorat, 2013). Primarily, it has been reported to be a

carcinogenic direct diazo dye used for colouration of paper products (Cripps et. al.


1990; Jaladoni-Buan et al. 2010). Besides having found it’s immense usage in textile,

it is also used in medicine (as a biological stain) and as an indicator (Perumal et al.

2012). The effluents containing CR are generated from a number of industrial

activities: textiles, printing and dyeing, paper, rubber, plastics industries (Purkait et

al. 2007; Vimonses et al. 2009).

This dye has been known to cause allergic reactions and it’s cytotoxicity for

bacteria and algae are well documented. Besides this, these dyes also exhibit

cytotoxicity; genotoxicity; hematotoxicity; neurotoxicity, as well as carcinogenicity

(Rajamohan and Kartikeyan, 2006; Kumar and Sawhney, 2011). The study

conducted by Puvaneswari in 2006 suggests dye to be carcinogenic for urinary

bladder of humans and tumorigenic in animals. The probable cause for

carcinogenicity underlies the fact that it forms carcinogenic amines such as benzidine

through cleavage of one or more azo groups is expected to metabolize to benzidine

(Mathur et al. 2005). This reason validates it to be under category of banned dyes

(Pielesz, 1999). It has also found to induce mutagenicity (Sharma et al. 2009;

Vimonses et al. 2009; Han et al. 2009; Sabnis, 2010; Zvezdelina et al. 2012). Kumar

and Sawhney in 2011 studied phytotoxicity of CR and the impact of the dye on

parameters of plant growth. The recalcitrance of CR has been attributed to the

presence of aminobiphenyl group and azo bonds, two features generally considered

as xenobiotic (Pinheiro, 2004; Sponza and Isik, 2005). Thus, the treatment of CR

contaminated wastewater is a complex process due to it’s aromatic structure,

imparting the dye it’s physicochemical, thermal and optical stability and resistance to

biodegradation and photodegradation (Purkait et al. 2007; Vimonses et al. 2009;

Smaranda et al. 2011). Consequently, due to the harmful effects of this organic

compound, the wastewaters containing CR must be treated properly before being

discharged to receiving water bodies.

2.4 Microbial Diversity: Exploring the unexplored

Microorganisms comprise of a huge and almost untapped reservoir of myriad

forms of molecular and chemical diversity in nature, as they constitute the most

diverse forms of life and are resourceful in terms of extrapolating some innovative

applications for the mankind. They have been evolving for nearly 4 billion years and

occupy almost every ecological niche on this planet and exhibit a tremendous


potential in exploiting a vast range of energy sources. The literature suggests that it is

the microbial protoplasm which constitutes for about 50% of the living protoplasm

on this planet. Being highly adaptive in nature, micro-organisms have shown their

presence even in extremely diverse environments pertaining to which they have

developed an extensive range of metabolic pathways (Jain et al. 2005). This

exemplifies an excellent example of sustained presence in an environment. In natural

environments, microorganisms are present in mixed populations (Flemming and

Wingender, 2010). Different species with different characteristics live together, often

in very complex communities. Microbial community can be described by separating

the individual components. This segregation forms the basis for most techniques to

study microbial communities in natural environments. The presence of

microorganisms in extreme conditions of stress or contaminated environments

facilitates their use in different biotechnological interventions as their enzymatic

systems are coded by genes which could be exploited. Additionally, the use of such

highly tolerant autochthonous strains in wastewater bioremediation is a focal point of

research (Gomaa and Momtaz, 2007). Only a fraction of the culturable bacteria

present in an environmental sample can be successfully isolated and cultivated

(Hugenholtz, 2002). To circumvent the problem of isolating the different species by

growth condition preferences, DNA-based molecular techniques over the last few

decades have been developed and had revealed an enormous reservoir of unexplored

or unculturable but viable microbes. Un-culturability is a condition that

encompasses: (i) Lack of specific growth requirements (nutritional, temperature and

aeration); (ii) slow-growing strains, out-competed in the presence of fast- growers

and (iii) injured organisms, which are incapable of overcoming the stressful

conditions imposed by cultivation. These categories may not represent specific

taxonomic positions but account for about 99% of the environmental bacterial

diversity. This large genetic diversity can potentially be used as a bioresource,

leading to development of novel biotransformation, bioremediation processes and

bioenergy generation (Kalia et al. 2003 a, b; Lee et al. 2006). Culturing these

unexplored microbes in controlled laboratory environment requires extensive

knowledge of their fastidious growth requirements. This has in fact been the driving

force for development of new methods to access this vast microbial wealth (Wanger

and Loy, 2002; Kalia et al. 2003 a, b; Handelsman, 2004; Green and Keller, 2006).

However, the development of culture-independent methods and the


commercialization of next-generation sequencing technology (Mardis, 2008) have

yielded powerful new tools in terms of time savings, cost effectiveness, and data

production capability. Methods such as 16S rDNA gene clone libraries, Fluorescence

In Situ Hybridization (FISH) or Denaturing Gradient Gel Electrophoresis (DGGE)

are being used to explore the bacterial diversity in waters have been reported

judiciously (Dewettinck et al. 2001; Zwart et al. 2002; 2005; Hoefel et al. 2005; Loy

et al. 2005; Bottari et al,. 2006; Wu et al. 2006; de Figueiredo et al. 2007; Von

Mering et al. 2007; Revetta et al.2010; Gabriel, 2010). More recently, the potential

of the high-throughput 454 pyrosequencing to explore the environmental diversity

has been emphasized (Roh et al. 2010).

16S rDNA gene diversity specifies the idea of species richness (number of

16S rDNA gene fragments from a sample) and relative abundance (structure or

evenness), which are reflective of relative pressures that construct diversity within

biological communities (Manefield et al. 2005; Paul et al. 2006). 16S rDNA based

molecular identification aims at identification, by virtue of it’s universal distribution

among bacteria and the presence of species-specific variable regions. Figure 12

represents characterization of microbial diversity by 16S rDNA sequencing strategy.

Figure 12: Flowchart depicting characterization of microbial diversity by 16S

rDNA analysis (Hobel, 2004)


The selective pressures that shape diversity within communities are reflected

by richness and evenness of bacterial communities. These parameters need to be

measured for assessment of different treatment effects (e.g., physical disturbances,

pollution, nutrient addition, predation, climate change, etc.) on community diversity.

This molecular tool has been extensively used assessing the bacterial phylogeny and

to assign the right taxonomic position of an environmental isolate (Yumoto et al.

2001) it’s applied aspect aimed at bacterial identification of unculturable

microorganism, unique or novel isolates and collections of phenotypically identified

isolates (Drancourt et al. 2000). The polyphasic approaches for microbial diversity

are represented in Figure 13.

Molecular methods have been used to analyze the microbial diversity of a

wide range of environments, which has generated many beneficial findings.

Examples include solid waste composters (Nakasaki et al. 2009), wastewater

treatment plants (Wagner et al. 2002) agricultural soils (Ranjard et al. 2000) and

natural rivers (Brummer et al. 2000). Furthermore, culture-independent approaches

have been used to identify many novel bacterial and archaeal lineages from different

environments (Oren, 2004).

Figure 13: Assessment of microbial diversity through culture dependent and

culture independent approach (Jain et al. 2005)


As a consequence, molecular approaches have lead to the interpretation, that

microbial world that is genetically and functionally more complex and diverse than

previously predicted from culture-dependent approach. However, the main

bottleneck of culture-independent approach lies in the fact that the strains in a

particular sample are indistinguishable and the unique properties of a particular

strain cannot be identified which requires intensive culture dependent approach to

assess microbial activites (Rapp & Giovannoni, 2003). Therefore, it would be very

difficult to use these culture-independent approaches to conduct a detailed study of

an individual strain for the development of an applied technology. In other words,

despite the widespread use of culture-independent approaches, cultural isolation will

continue to be an important but necessary method to generate new technologies.

Figure 14 represent the multidisciplinary approach to identify microbial diversity.

Figure 14: Multidisciplinary approach to analyze and characterize the microbial diversity through culture independent approach. (Hugenholtz, 2002)


This approach of 16 S rDNA sequencing which utilize the molecular biology

tools have resulted in the discovery of entirely new lineages, some of which are

major constituents of environmental communities that were not detected by

traditional conventional systems which have lead to emergence of metagenomics to

access the gene pool of vast diversity of unculturable bacteria (Sharma et al. 2005).

The molecular phylogenetic analysis of bacterial communities by culture-

independent studies has resulted in increase of identifiable bacterial divisions to 40,

13 of which are characterized by environmental DNA sequences from unculturable

bacteria and are henceforth known as candidate divisions (Hugenholtz et al. 1998).

The microbial communities in ETP exist in dynamic consortial forms, the

understanding of which could be made through the knowledge of different co-

existing microbial populations (Manefield et al. 2005) which is non uniform with

changing operational conditions of the reactor (Kapley et al. 2007). Their

contribution to overall degradation of pollutants is likely to provide un paralleled

control over the bioremediation of the effluents (Manefield et al. 2005; Paul et al.

2006).

In wastewater treatment, microbial molecular ecology techniques have been

applied mainly to the study of flocs (activated sludge) and biofilms that grow in

aerobic treatment systems like trickling filters (Sanz and Kochling, 2007). Prokaryotes

are among the most important contributors to the transformation of complex organic

compounds in WWTP. Forster et al. in 2003 has reported the importance of bacterial

assemblages to the proper functioning and maintenance of treatment plants.

Molecular approaches such as 16S rDNA clone libraries (Blackall et al.1998 ;

McGarvey et al. 2004; Otawa et al. 2006), Ribosomal Intergenic Spacer Analyses

(RISA) (Yu and Mohn, 2001; Baker et al. 2003), 16S-restriction fragment length

polymorphism (16S-RFLP) (Baker et al. 2003) Repetitive Extragenic Palindrome

PCR (REP-PCR) (Baker et al. 2003) and fluorescent in situ hybridization (FISH)

( Bjornsson et al. 2002) have already been applied to the study of wastewater-

associated microbial communities.


PCR–DGGE has been successfully implemented in many fields of microbial

ecology to assess the diversity and to determine the community dynamics in response

to environmental variations. DGGE-based approach has been used to study bacterial

diversity in wastewaters using reactors systems (Liu et al. 2002; Casserly and

Erijman, 2003; Kaksonen et al. 2003; Rowan et al. 2003) and activated sludge (Boon

et al. 2002; Ibekwe et al. 2003; Gilbride et al. 2006) revealing the presence of highly

complex bacterial communities. Bacterial diversity of aerated lagoons from WWTP

so far have not been substantially assessed where the degradation of organic matter

takes place. Cloning had been employed to establish with precision the phylogenetic

position of filamentous bacteria in granular sludge (Sekiguchi et al. 2001) or to

determine the prevalent sulfate reducing bacteria in a biofilm (Ito et al. 2002). The

study was extended by Yamada et al. 2005 who found the most prevalent sulfate

reducer belonging to Chloroflexi subphylum. This prokaryote has been associated

with bulking of sludge in treatment plants. The microbial communities residing in

reactors for treating several types of industrial wastewater have also been determined

by means of 16S rDNA cloning and sequencing. Egli et al. 2004 examined the

microbial composition and structure of a rotating biological contactor biofilm for the

treatment of ammonium-contaminated wastewaters. The study revealed the

sequences of several previously undetected and uncommon microorganisms.

Phylogenetic analysis of the sequences obtained showed a narrow range of diversity,

with most of the screened microorganisms belonging to the Methanosarcina sp.

Studies conducted by Zhang et al. in 2005 elucidated the efficacy of cloning

approach in conjunction with in situ hybridization analysis in methanogenic reactor

adapted to phenol degradation.

This essentially requires designing of new specific primers and gene probes

for detection and/or quantification of microorganisms. On the similar lines, Crocetti

et al. 2000 extracted genomic microbial DNA from a sequencing batch reactor

cloned the bacterial 16S rDNA and identified Rhodocyclus sp. and Propionibacter

pelophilus as the microorganisms responsible for the polyphosphate accumulation

taking place in the reactor. Futhermore, the same research group, designed probes for

these species that could correlate phosphorous removal and the number of hybridized

cells in different sludges. Alpha-proteobacteria, whose role in anaerobic/aerobic-


activated sludge phosphorous removal plants has been analysed were characterized

by new probes for in situ hybridization with information provided by 16S rDNA

gene library sequences and DGGE analysis (Beer et al. 2004). A similar approach

had been considered three years earlier by the same authors, but Kong et al. 2001

utilised the combination of cloning, DGGE, and FISH for identification of

predominant microorganisms in an anaerobic sequencing batch reactor (SBR)

without probe design and phosphorus removal. A comparative analysis of further

insights into molecular biology tools for detection of microbial diversity are

represented in Table 8.

Table 8: Comparative analysis of further insights into molecular biology tools

for detection of microbial diversity (Sanz and Kochling, 2007)

S.No Method Outline Advantages Disadvantages

1 tRFLP 16 S based, relies on differences in polymorphism

Relatively simple procedures

Heterogenous size of fragments makes

phylogenetic analysis less confident

2 RISA Phylogeny with intergenic spacer region between 23S and 16S

rDNA sequences

High sensitivity, down to sub-species level

Database for comparative analysis small

in comparison to 16S sequences

3 PCR with genes

Microorganisms containing enzymes

involved in the biodegradation process are detected

Direct detection of the presence of degradative microorganisms.

Subtyping on strain level is possible

Global profiling of microbial

community missing

4 PLFA Profiling of microorganisms by

characteristic fatty acid content

Molecular characterization of microorganisms not relying on genes.

Complementary information to 16S based assays

Not a good choice as a standard

alone method

5 DNA microarray

Multi sample hybridization method

High sample throughput Parallel analysis of

different parameters

Expensive equipment, difficult handling


Studies on bacterial diversity of waste water systems have disseminated

knowledge with respect to potential of indigenous microbes in attenuation of

pollutants from contaminated sites. Studies based on assessing microbial

communities of textile effluents by 16 S rDNA sequencing have been reported. To

comprehend the biochemical attributes of autochthonous strains, it becomes

quintessential to identify them and place them in appropriate divisons.

Karthikeyan and Anbusaravanan in 2013 conducted a study pertaining to

characterization of bacteria isolated from textile effluents influxed with sewage

released into the river Amaravathy, Karur, Tamilnadu , identified bacteria by 16 S

rDNA sequencing as Bacillus cereus AK1968 and Pseudomonas sp. AKDYE14.

Rajeswari et al. in 2013 screened textile effluent adapted bacteria for

degradation of different reactive dyes, from CETP, Tirupur and various textile units

and identified the potential isolate through sequencing of 16S rDNA gene as

Stenotrophomomonas maltophila RSV-2.

Ankita and Saharan in 2013 isolated a potential azo dye degrading bacterium

from grit tank of an ETP of tannery industries at Jajmau, Kanpur, Uttar Pradesh,

India and identified the strain by 16 S r DNA sequencing as Pseudomonas otitidis.

Khadijah et al. 2009 conducted a broad spectrum study for screening reactive

azo dye decolourising bacterial strains from effluent from local textile mill, batik

making site, textile laboratory, lake, streams, market, and domestic wastewater of

Shah Alam city. Of a remarkable 1540 bacterial isolates obtained, two potential dye

decolourisers were molecularly characterized as Chryseobacterium sp. and

Flavobacterium sp.

The isolation of microorganisms from highly contaminated environments

(textile effluents) indicates the natural adaptation of microorganisms to survive in the

presence of toxic dyes. The difference in their rate of decolourisation may be due to

the loss of ecological interaction, which they might be sharing with each other under

natural conditions (Bhimani, 2011) and offers novel bacteria of unique functionality

and potential applications in different biotechnological processes. It also entails a


variety of genes responsible for microbial tolerance or defense against extreme

conditions or xenobiotics present in the media (Gomaa and Momtaz, 2007). Bhimani,

2011 carried out an elaborate study to assess the bacterial diversity of a CETP of

Jetpur, located in heart of Saurashtra, Gujarat which is largely infiltrated by textile

effluents. Of total 37 isolates screened, three exhibited tremendous potential to

degrade textile based azo dye and was characterized by 16 S rDNA sequencing and

identified them as Alcaligenes faecalis JTP-07; Pseudomonas aeruginosa JTP-37;

Lysinibacillus fusiformis JTP-23.

Halophilic bacteria have also shown to be a good source of dye degradation

(Kowsalya et al. 2013; Asad et al. 2014). Prabhakar in 2012 isolated a total of 84

bacterial strains from Kelambakkam Solar Salt Crystallizer ponds (or salterns) and

screened for their ability to produce extracellular tannase and laccase enzymes and

eventually to decolourize three widely used textile dyes- Reactive Blue 81, Reactive

Red 111 and Reactive Yellow 44. Of these 84 strains, 18 strains exhibited tannase

activity and 36 strains showed positive laccase enzyme activity. The 11 bacterial

strains that displayed both tannase and laccase enzyme activity were screened for

their ability to decolourize the three textile azo dyes. The best isolate AMETH72 was

sequenced for it’s 16S rDNA gene and was identified as Halomonas elongate.

Bacillus sp is supposedly unique with respect to it’s adaptability to dye based

environments; both soil contaminated with dyes and effluents at large (Devassy,

2010) and this genus is ubiquitous in nature with ability to degrade a wide range of

substrates. Plausibly, comprehensive reports pertaining to textile dye decolourizing

potential of this inherent species are widely available and are indicative of it’s role in

bioremediation of textile effluents. A study was conducted to isolate a potential

bacterial strain from textile wastewater generated by textile industries in Cairo,

Egypt and establish it’s tolerance to pH, low temperature and hydrogen peroxide.

The partial 16S rDNA gene amplification and sequencing revealed the strain to be a

gram positive rod and were identified as Bacillus macroccanus with (Gomaa and

Momtaz, 2007). In a similar study conducted on soil contaminated with dye and

untreated textile mill effluent, Bacillus sp identified through culture dependent

approach has found to possess versatility in terms of azo dye decolourisation


(Dubey et al. 2010). In an another study, bacteria from soil contaminated with dye

was screened for it’s potential to degrade reactive azo dye, the 16s r DNA gene

sequencing revealed the isolated organism as Enterococcus faecalis strain YZ66

(Sahasrabudhe and Pathade, 2011). Aerobic strains of Streptococcus faecalis and

Bacillus subtilis had been isolated and identified by 16S r DNA sequencing from

textile effluent and had been explored for metabolizing the textile effluent. This

study is represents the remarkable feature of strains adaptability to textile effluents

(Sivaraj et al.2011). In an another study, Bacillus subtilis RA-29 isolated from

garden soil in the vicinity of an industrial town in Himachal Pradesh had been

phylogenetically characterized (Kumar and Sawhney , 2011).

Kowsalya et al. 2013 explored bacterial diversity of effluent samples

collected from textile dye units from various identified sites in Kanchipuram,

Tamilnadu to establish their bioremediative properties with respect to dye

degradation. The isolates were characterized as Halomonas KB1; Bacillus KB2;

Bacillus subtilis KB 3; Bacillus cereus KB4; Staphylococcus sp. A bacterium

isolated and identified as Bacillus cereus from tannery effluent collected at Central

Leather Research Institute (CLRI), Chennai was employed for decolourisation of an

azo dye (Kanagaraj et al. 2011).

Shah et al. 2013 isolated dye decolourizing bacterial isolate from textile dye

effluent from textile industries of Ankleshwar, Gujarat, India and identified it as

ETL-79(Bacillus sp.) through culture dependent approach. An earlier study

conducted by Shah et al. 2013 also revealed the presence of ETL-1 (Klebsiella

oxytoca) and ETL-2 (Bacillus subtillis) out of 84 strains isolated from Common

Effluent Treatment Plant (CETP) of Ankleshwar, Gujarat, India for dye degradation

studies. Pradhan et al. in 2012 conducted a study on waste water samples generated

from sugar industry to screen for dye degrading bacteria and the best dye degrader

was characterized at molecular level was Bacillus sp. SCWS5.

Soil and effluent samples obtained from CETP, Ankleshwar, Gujarat had

been the source of an environmental isolate Pseudomonas steutzeri-ETL-4 as

characterized by 16 S rDNA sequencing. This strain had been judiciously utilsed for


degradation of CR which is a sulphonated diazo direct dye (Shah, 2014). In a similar

study carried out earlier, activated sludge obtained from CETP, Naroda, Gujarat had

been a source of an indigenous bacterial strain Bacillus boroniphilus as identified by

16 S rDNA sequencing to degrade Reactive Yellow 145 dye (Derle et al. 2012).

Subhatra et al. 2013 investigated the potential of bacteria (Bacillus sp.)

isolated from effluent produced by textile industries in Madurai , Tamil Nadu to

establish a synergistic effect on heavy metal tolerance and dye degrading potential.

Milikli and Rao, 2012 investigated soil samples from contaminated sites of

Mangalagiri textile industry, Guntur, Andhra Pradesh, India to screen out the most

potential Bromophenol blue dye degrading bacteria. Through selective screening the

most potential isolate out of the 3 isolates screened, was identified by 16 S rDNA

sequencing as Bacillus subtilis.

Manivannan et al. 2011 explored the dye house effluent collected from a dye

processing industry, Arulpuram, Tirupur region, Tamil Nadu to assess it’s bacterial

diversity to explore the biodegradability of Orange 3G dye. The efficient strains

identifies were Bacillus sp., Escherichia coli and Pseudomonas fluorescens

Balakrishnan and Sowparnika, 2013 studied bacterial diversity of fresh textile

effluent released by textile industries in Erode district, Tamil Nadu, India in terms of

dye degradation. The most promising bacterial isolates identified by biochemical

tests were found out to be Pseudomonas sp., Bacillus sp and Alcaligenes sp.

Balgurunathan and Sudha, 2013 isolated and identified a bacterial strain as

Bacillus licheniformis from effluent sample of Infra Tex textile industry, Perundurai

in Erode district Tamil Nadu, India for it’s use as excellent biomass for removal of

reactive dyes. A similar study was conducted to explore the bacterial diversity at the

same site; Perundurai in Erode district of Tamil Nadu, India in which soil

contaminated with dye was collected at effluent discharge site and the isolates were

screened for a repertoire of textile dyes and the most efficient dye degraders were

Pseudomonas aeruginosa, Alcaligenes faecalis, Proteus mirabilis, Serratia

marcescens and Bacillus licheniformis as identified by a series of biochemical tests

(Neelambari et al. 2013).


A combined study was carried out in which both the textile effluent and

sludge samples from Amanishah nullah, Jaipur were assessed for bacterial diversity

aimed at exploring the dye decolourising potential of the adapted bacterial strains.

Most potential light red dye degrading bacterial strains identified through culture

dependent approach were Pseudomonas sp, Klebsiella sp and Proteus sp from the

effluent and Shigella sp, Morganella sp and Klebsiella sp from the sludge sample

(Sethi et al. 2012). A similar study was carried out to explore potential dye

decolourizing bacterial strains from the textile industry waste located in Erode and

Tripur districts, Tamil Nadu, India. There was 96 morphologically distinct bacterial

isolates were isolated from 12 different sludge, textile effluent and dye contaminated

soil samples. Generic composition of the 96 isolates comprised of Bacillus sp.,

Enterobacteriaceae, Pseudomonas sp., Micrococcus sp.(Sahasrabhude et al. 2014)

Alcaligenes sp., Aeromonas sp., Staphylococcus sp., and Lactobacillus sp (Palani

Velan et al. 2012).

Not only does textile effluents and soil contaminated with dye have been the

source of Bacillus sp., it’s ubiquitous nature is well established from a report which

is suggestive of it’s source being the waste water generated from a carpet industry in

Khairabad, Uttar Pradesh, India (Dubey et al. 2010).

Marine bacteria are vital in recycling nutrients and could be useful for

innovative applications, which are helpful to human beings (Barakat, 2012). A study

was carried out to isolate the degrading bacterial strains from mangrove associated

sediment collected from nearby Roche Park, coastal area of Tuticorin and further to

assess the efficiency of selected bacterial strain in degradation of dyes and identified

the strain as Pseudomonas sp (Sponza et al. 2003) This report is unique in respect

that marine microorganisms have developed unique metabolic and physiological

capabilities that not only ensure survival in extreme habitats but also offer the

potential for the production of metabolites, which would not be observed, from

terrestrial organisms (Raja et al. 2013). Durve et al. 2012 isolated two potential dye

degrading strains from effluent samples of textile processing units in Maharashtra,

India and biochemically identified them as Pseudomonas aeruginosa and

Brevibacillus choshinensis (Surwase et al. 2013).


Kannan et al. 2013 screened the dye-contaminated soil samples collected

from Kallidaikurichi, of Tirunelveli district, Southern Tamil Nadu, which is well

known for handloom production with flourishing dyeing units; to isolate and identify

the bacterial isolates which belonged to genera Pseudomonas sp. (32.5 %) followed

by Bacillus sp. (27.5 %), Aeromonas sp. (15.0 %), Micrococcus sp. (12.5 %) and

Achromobacter sp (12.5 %).

Globally, studies on bacterial diversity for dye degradation studies are also

reproducible. Bacterial strains were isolated from activated sludge (Corona

Wastewater Treatment Plant, CA, USA), turf grass soil (Chino loamy sand, pH 7.4,

Chino Park, CA, USA) and from a natural asphalt soil mixture obtained from the

Rancho La Brea tar pit’s in Los Angeles, CA, USA and characterized the strains

based upon their 16S rDNA sequence similarities. From activated Sludge AS7

Bacillus cereus CCM 2010; AS77 Pseudomonas nitroreducens 0802; AS81

Aeromonas punctata MPT4; AS96 Shewanella putrefaciens LMG 2369; Soil S46

Bacillus thuringiensis Al Hakam; S81 Massilia timonae were isolated (Khalid et al.

2008). A bacterial strain, CK3, with remarkable ability to 45 decolourize the reactive

textile dye Reactive Red 180, was isolated from the activated sludge collected from a

textile mill. Phenotypic characterization and phylogenetic analysis of the 16S rDNA

sequence indicated that the bacterial strain belonged to the genus Citrobacter CK 3

(Wang et al. 2009). Pahlaviani et al. 2011 isolated native bacterial strains from

activated sludge, turf grass soil in Iran for examining their azo dye degrading

potential. The strains were identified by 16 S rDNA sequencing as Bacillus cereus

CCM 2010, Pseudomonas sp. B13T, Shewanella algae ATCC 51192, Shewanella

algae ATCC 51192.

Olukanni et al. 2006 investigated the potential of effluent adapted and non-

adapted bacteria isolated from textile industries wastewater and outlet in Nigeria,

Africa. The effluents from textile industries and drains served as a source of effluent

adapted bacteria and isolates from a municipal landfill were effluent non-adapted

bacteria. Effluent adapted strains of Acinetobacter, Bacillus and Legionella with

potentials for colour removal and COD removal activities were characterized.

Likewise, a similar study conducted by Ajao et al. 2013 was based on dye effluent


from International Textile Industry (Nig) Ltd, Odogunyan Industrial Estate Ikorodu,

Lagos State, Nigeria to explore dye decolourising potential of adapted strains under

immobilized conditions on agar-agar in a bioreactor. The strains were characterized

as Pseudomonas aeruginosa and Bacillus subtilis.

Barakat, 2012 isolated a novel azo dye-degrading bacterium T312D9 strain

from Abou Quir Gulf, Amya industrial pumping station considering enriched

location contaminated with various dyes in Alexandria, Egypt. The identification of

the isolate by 16S rDNA gene sequencing revealed to be Lysobacter sp T312D9.

This marine ecofriendly isolate was exploited for it’s ability to degrade two synthetic

azo dyes (CR and methyl red) considered as detrimental pollutants from industrial

effluents.

Mahmood et al. 2012 screened wastewater (Textile effluents), sludge and

affected soil samples from Hudiara drain near Nishat Mills Limited from Ferozepur

Road Lahore, Pakistan to isolate and identify potential dye degrading bacterial

strains. The isolates were identified as Bacillus subtilus (Isolate 20), Bacillus cereus

(Isolate 3), Bacillus mycoides (Isolate 1), Bacillus sp. (Isolate 5), Pseudomonas sp.

(Isolate 9) and Micrococcus sp. (Isolate 7) by standard physiological, morphological

and biochemical tests. The isolation of good dye-decolourizing species requires

screening, and these isolated strains should have ability to degrade and detoxify

textile dyes (Silveira et al. 2009). Yang et al. 2011 screened heavily dye

contaminated soil and wastewater samples from Changzhou dye manufacturing

industries and Qingtan wastewater treatment factory in Changzhou, China for

isolation of a potential dye degrading bacterial isolate The isolate was identified

through culture independent approach by 16 S rDNA sequencing and designated as

Psedomonas putida WLY. Pseudomonas aeruginosa , an effleuent adapted microbe

isolated from textile waste water selected from a group of dye degrading bacteria

obtained, was chosen for achieving the best decolourisation and for it’s broad

spectrum azo dye biodegradability potential (Hafshejani et al. 2013). Pseudomonas

sp. is widely distributed in a variety of habitats and is able to metabolize organic

contaminants in the environment (Ashiuchi et al. 1999; Poornima et al. 2010).


Sun-Young An et al. 2006 isolated and identified a bacterial strain as

Citrobacter sp. from textile waste water capable of degrading both azo and

triphenylmethane group of dyes. An azo-dye-reducing, endospore-forming bacterium

DSM 13822T isolated from textile industry wastewater has been taxonomically

studied and identified as Paenibacillus azoreducens (McMullan et al. 2001).

Recently, a study focused at isolation of bacteria from activated sludge of a

textile plant wastewater treatment facility was reported (Franciscon et al. 2012)

which was aimed at sequential decolourisation and detoxification of the azo dyes

Reactive Yellow 107 (RY107), Reactive Black 5 (RB5), Reactive Red 198 (RR198)

and Direct Blue 71 (DB71). 16 S r DNA sequence of the isolate VN-15 revealed it to

be Brevibacterium sp.

Activated sludge obtained from obtained from denitrifying reactor of

wastewater treatment plant (WWTP) in textile factory “Giorgetti Bulgaria” AD –

Elin Pelin, was used as a source of high performance bacterial decolourizer AZO29

to reduce Amaranth dye. The isolate was characterized as Pseudomonas sp (Vasileva

et al. 2009). Previous studies had focused on the isolation of bacterial strains from

activated sludge. In the early years, (Horan et al. 1988) had attempted to isolate both

floc-forming and filamentous bacteria from activated sludge flocs. Until recent years,

a few bacterial strains especially filamentous bacteria were cultivated by the method

of dilution plate (Ramothokang et al. 2003) or micromanipulation (Blackall et al.

1996). However, it should be noted that a majority of microorganisms present in any

environment have not been cultivated through general media (Ellis et al. 2003) thus,

obtaining more unexplored microbes is an important prerequisite for better

understanding what functions these strains have in this complex environment (Lu et

al.2006) which showed that β-Proteobacteria was the most dominant group in sludge

samples. In addition, at the genus level, strains of the genera Klebsiella,

Pseudomonas, Bacillus, Aeromonas, Flavobacterium and Acinetobacter (12.7%,

12.7%, 12.7%, 11.1%, 7.9% and 7.9%, respectively) were the most predominant

microorganisms in the collection. REP-PCR has attracted a lot of attention in recent

years, for it’s highly discriminatory, accurate, fast and low-cost advantages. To date,

it has been applied for analyses of biodiversity among bacterial strains in water


(Mohapatra and Mazumder, 2008), soil (Rejili et al. 2009), plants (Sikora and

Redzepovic, 2003). However, there were no reports on the analysis of isolated

bacteria diversity in activated sludge using this method. Meanwhile, it should be

noted that different strains belonging to the same genus may play different roles in

activated sludge. For this reason, it is urgent to obtain more detailed evaluation on

biodiversity among the same genus; REP-PCR can be a suitable candidate reference

method for this purpose (Jin et al. 2011).

In the past years, many researchers focused on the enumerating the bacterial

community presents in activated sludge by using molecular approaches, including

DGGE and 16S rDNA gene clone library analysis (Blackall et al. 1998; Boon et al.

2002; Eschenhagen et al. 2003; Choi et al. 2007). These culture-independent

technologies can reveal more abundance of microbial diversity than culture-

dependent approaches in the terms of bacterial composition. The bacterial

community of a bulking sludge from a municipal WWTP with anoxic-anaerobic-oxic

process was investigated by combination of cultivation and 16S rDNA gene clone

library analysis for understanding the causes of bulking. A total of 28 species were

obtained from 63 isolates collected from six culture media. The most cultivable

species belonged to Proteobacteria including Klebsiella sp., Pseudomonas sp.,

Aeromonas sp. and Acinetobacter sp. At the level of class, β -Proteobacteria (43.6%)

was the dominant group in the bulking sludge (Zhuibai et al. 2011) which was in

accordance with previous studies (Rani et al. 2008). Both the marker and functional

genes had been used in identification (Suizhou et al. 2006) of a dye-decolourizing

bacterial strain DN322 from activated-sludge of a textile-printing WWTP,

Guangzhou, China and was identified as Aeromonas hydrophila based on sequence

analysis of 16S rDNA gene and gyrase beta subunit (gyr). The diversity of bacterial

groups of activated sludge samples that received wastewater from four different

types of industries (Flanders, Belgium) i.e., domestic wastewater, carbohydrate rich

wastewater, wastewater from paper and starch related industries, protein and fat rich

wastewater from food and meat related industries and wastewater from textile

industry was investigated by a nested PCR- DGGE (Denaturing Gradient Gel

Electrophoresis) approach. Specific 16S rDNA primers were chosen for large

bacterial groups (Bacteria and α-Proteobacteria in particular), which dominate


activated sludge communities, as well as for actinomycetes, ammonium oxidisers

and methanotrophs (types I and II) In addition primers for the new Acidobacterium

kingdom were used to observe their community structure in activated sludge. After

this first PCR amplification, a second PCR with bacterial primers yielded 16S rDNA

gene fragments that were subsequently separated by DGGE, thus generating `group-

specific DGGE patterns (Boon et al. 2002).

2.5 Dye removal techniques

The addressal of environmental issues pertaining to presence of colour in

wastewater needs some stringent laws to be reinforced so as to develop treatment

technologies so that the effluents prior to their release may be treated properly and

threat to the environment may be minimized upto certain extent. The important

methods useful in management of environmental contaminants in the environment

include the following (Rajendran and Gunasekran, 2006) (Figure 15).

Figure 15: Methods aimed at management of environmental contaminants

(Rajendran and Gunasekran, 2006)

Management of environmental contaminants

Ongoing strategies

In process treatment

End-of-pipe treatment

long term strategies

Remediation of polluted sites

Modification of existing process

Introduction of new processes and

products


Though many technologies have been available for clean up purpose, only a

few of them have been proved to be of routine application value. The choice of

technology is influenced by many factors. The available technologies can be

categorized as potential and time proven trusted technologies. The available

technologies are assessed based on certain criteria (Rajendran and Gunasekran,

2006). Figure 16 represents criteria for available technologies for “clean up”.

Figure 16: Criteria for available technologies for “clean up” (Rajendran and

Gunasekran, 2006).

Several physical, chemical and biological methods for the dyes removal from

the wastewater are available (Pagga and Laboureur, 1983; Cooper, 1993;

Vandevivere, et al. 1998; Hao, et al. 2000; Robinson et al. 2001; Blumel, et al. 2002;

Ali, 2010; Archana et al. 2012). The physicochemical treatment include membrane

filteration, coagulation/ flocculation, precipitation, flotation, adsorption, ion exchange,

ion pair extraction, ultrasonic mineralization, electrolysis, advanced oxidation

(Chlorination, bleaching, ozonization, fenton oxidation and photocatalytic oxidation)

and chemical reduction.

Temporary or permanent solution

More conversion or total conversion

Cost effectiveness

Time taken for treatment

Regulatory standards


The textile organic dyes must be removed from industrial wastewaters by

effective and viable treatments at treatment plant or on site following two different

treatment concepts as: (1) separation of organic pollutants from water environment, or

(2) the partial or complete mineralization or decomposition of organic pollutants.

Separation processes are based on fluid mechanics (sedimentation, centrifugation,

filtration and flotation) or on synthetic membranes (micro- ultra- and nanofiltration,

reverse osmosis). Additionally, physico-chemical processes (i.e. adsorption,

chemical precipitation, coagulation-flocculation, and ionic exchange) can be used to

separate dissolved, emulsified and solid-separating compounds from water

environment (Robinson et al. 2001; Anjaneyulu et al. 2005; Zaharia, 2005; Babu et

al. 2007; Suteu et al. 2009a; Suteu et al. 2011a; Zaharia et al. 2009; Zaharia et al.

2011). The partial and complete mineralization or decomposition of pollutants can be

achieved by biological and chemical processes (biological processes in connection

with the activated sludge processes and membrane bioreactors, advanced oxidation

with O3, H2O2, UV (Dos Santos et al. 2004 ; Oztekin et al. 2010 ; Wiesmann et al.

2007 ; Zaharia et al. 2009). A strategic approach pertaining to different separation

processes (sedimentation, filtration, membrane separation), and some physico-

chemical treatment steps (i.e. adsorption; coagulation-flocculation with inorganic

coagulants and organic polymers; chemical oxidation; ozonation; electrochemical

process) are integrated into a specific order in the technological process of

wastewater treatment for decolourisation or large-scale colour and dye removal

processes of textile effluents. To introduce a logical order in the description of

treatment methods for textile dye and colour removal, the relationship between

pollutant and respective typical treatment technology is taken as a reference. The first

treatment step for textile wastewater is the separation of suspended solids and

immiscible liquids from the main textile effluents by gravity separation (e.g., grit

separation, sedimentation including coagulation/flocculation), filtration, membrane

filtration, air flotation, and/or other oil/water separation operations.

The following treatment steps are applied to soluble pollutants, when these

are transferred into solids (e.g., chemical precipitation, coagulation/flocculation, etc.)

or gaseous and soluble compounds with low or high dangerous/toxic effect (e.g.,


chemical oxidation, ozonation, wet air oxidation, adsorption, ion exchange, stripping,

nanofiltration/reverse osmosis). Solid-free wastewater can either be segregated into a

biodegradable and a nonbiodegradable part, or the contaminants responsible for the

non-biodegradable wastewater part that can be decomposed based on physical and/or

chemical processes. After an adequate treatment, the treated wastewater can either be

discharged into a receiving water body. An outline of different treatment processes

for textile effluents are outlined in Table 9.

2.6 Biodegradation of dyes: Applied aspect of bacterial diversity

Traditional wastewater treatment technologies are markedly ineffective for

handling wastewater of synthetic textile dyes because of the chemical stability of the

pollutants (Forgacs et al. 2004). Additionally, the water recycling issue remains

unaddressed (Soares et al. 2004). Physico-chemical methods have long been used to

treat textile wastewater influxed with synthetic dyes. The major disadvantage of

physicochemical methods is primarily the high cost, low efficiency, limited

versatility, need for specialized equipment, interference by other wastewater

constituents, and the handling of the generated waste (Van der Zee and Villaverde

2005; Kaushik and Malik, 2009). Physical methods can effectively remove colour,

but the dye molecules are not degraded, becoming concentrated and requiring proper

disposal. With the chemical techniques, although the dyes are removed,

accumulation of concentrated sludge creates a menace. A possibility of secondary

pollution problem always arises because of the excessive amounts of chemicals

involved (Amoozegar, 2011). Recently, other emerging techniques like advanced

oxidation processes, which are based on the generation of very powerful oxidizing

agents such as hydroxyl radicals- have been applied with success in pollutant

degradation (Arslan and Balcioglu, 1999). Although these methods are efficient for

the treatment of waters contaminated with pollutants, they are very costly and

commercially unattractive compounds (Alexander, 1994; Bennet et al. 2002). These

compounds may be biodegradable, persistent or recalcitrant (Nikaido and Glazer,

2007) which may be cleaved into smaller compounds by viable microbes (Marinescu

et al. 2009).


Table 9: Various current and emerging dye separation and elimination treatments applied for textile effluents with their principal advantages and limitations (adapted from Robinson et al. 2001; Anjaneyulu et al. 2005; Babu et al. 2007)

S.No Treatment methodology Treatment stage Advantages Limitations

PHYSICO-CHEMICAL METHODS

1 Precipitation,

Coagulationflocculation

Pre/main treatment Short detention time and low capital costs. Relatively good removal efficiencies.

Agglomerates separation and treatment. Selected operating condition.

2 Electrokinetic Coagulation Pre/main treatment Economically feasible. High sludge production.

3 Fenton process Pre/main treatment Effective for both soluble and insoluble coloured contaminants. No alternation in volume

Sludge generation; problem with sludge disposal. Prohibitively expensive.

4 Ozonation Main treatment Effective for azo dye removal. Applied in gaseous state: no alteration of volume.

Not suitable for dispersed dyes. Releases aromatic dyes. Short half-life of ozone (20 min).

5 Oxidation with

NaOCl

Post treatment Low temperature requirement. Initiates and accelerates azo bond cleavage.

Cost intensive process. Release of aromatic amines.

ADSORPTION WITH SOLID ADSORBENTS

6 Activated carbon

Economically attractive.

Pre/post

Treatment

Good removal efficiency of wide variety of dyes.

Very expensive; cost intensive regeneration

Process.

7 Peat Pre treatment Effective adsorbent due to cellular structure. No activation required.

Surface area is lower than activated carbon.

8 Coal ashes Pre treatment Economically attractive. Good removal efficiency. Larger contact times and huge quantities are required.

Specific surface area of adsorption is lower than activated carbon.



9 Wood chips/

Wood sawdust

Pre treatment Effective adsorbent due to cellular structure. Economically attractive. Good adsorption

capacity for acid dyes.

Long retention times and huge quantities are

required.

10 Silica gels Pre treatment Effective for basic dyes Side reactions prevent commercial application

11 Irradiation Post treatment Effective oxidation at lab scale Requires a lot of dissolved oxygen

12 Photochemical process Post treatment No sludge production Formation of byproducts

13 Electrochemical oxidation Pre treatment No additional chemicals required and the end products are non-dangerous/hazardous.

Cost intensive process; mainly high cost of

Electricity

14 Ion exchange Main treatment Regeneration with low loss of adsorbents Specific application; not effective for all dyes

BIOLOGICAL TREATMENT

15 Aerobic process Post treatment Partial or complete decolourisation for all classes of dyes

Expensive treatment

16 Anaerobic process Main treatment Resistant to wide variety of complex coloured compounds. Bio gas produced is used for stream generation.

Longer acclimatization Phase

17 Single cell (Fungal, Algal & Bacterial)

Post treatment Good removal efficiency for low volumes and concentrations. Very effective for specific colour removal.

Culture maintenance is cost intensive. Cannot cope up with large volumes of wastewater.



EMERGING TECHNOLOGIES

18 Other advanced

oxidation process

Main treatment Complete mineralization ensured. Growing number of commercial applications Effective pre-treatment methodology in integrated systems and enhances biodegradability.

Cost intensive process

19 Membrane

filtration

Main

treatment

Removes all dye types; recovery and reuse of chemicals and water.

High running cost. Concentrated sludge production. Dissolved solids are not separated in this process

20 Photocatalysis

Post treatment Process carried out at ambient conditions. Inputs are no toxic and inexpensive. Complete mineralization with shorter detention times. Effective for small amount of coloured compounds.

Expensive Process

21 Sonication

Pre treatment Simplicity in use. Very effective in integrated systems.

Relatively new method and awaiting full scale application.

22 Enzymatic

Treatment

Post treatment Effective for specifically selected compounds. Enzyme isolation and purification is tedious.

23 Redox mediators Pre/ supportive

Treatment

Easily available and enhances the process by increasing electron transfer efficiency

Concentration of redox mediator may give antagonistic effect. Also depends on biological activity of the system..

24 Engineered

wetland systems

Pre/post treatment Cost effective technology and can be operated with huge volumes of wastewater

High initial installation cost. Requires expertise and managing during monsoon becomes difficult


The high electrical energy demand and the consumption of chemical reagents

are common problems. The development of efficient, economic and eco friendly

technologies to decrease dye content in wastewater to acceptable levels at affordable

cost is of utmost importance (Couto, 2009). Biological methods are generally

considered environmentally benign because they lead to complete mineralization of

organic pollutants at effectively low cost (Pandey et al. 2007). They also dissipate

BOD, COD and suspended solids. The main limitation relates, in some cases to the

toxicity of some dyes and/or their degradation products to the organisms used in the

process.

Natural attenuation or bioattenuation is the dilution of contaminants in the

environment through biological processes (aerobic and anaerobic biodegradation,

plant and animal uptake), physical processes (advection, dispersion, dilution,

diffusion, volatilization, sorption/desorption), and chemical responses (ion exchange,

complexation, abiotic transformation) (Joutey et al. 2013). Figure 17 represents natural

attenuation of pollutants.

The resultant product may not be essentially desirable as the simpler

compounds may be more toxic than the parent compound (Nikaido and Glazer,

2007). Biodegradation in current scenario involves aerobic microorganisms, which

utilize molecular oxygen as reducing equivalent acceptor during respiration

(Anjanyelu, 2005; Zaharia, 2012). Under anaerobic environment (anoxic and

hypoxic environments) biodegradation also occurs, and survival of microorganisms

(Singh, 2008) is attained by using sulphates, nitrates and carbon dioxide as electron

acceptors (Birch et al. 1989). Mineralization is complete biodegradation process

(Joutey et al. 2013) in which the end products are CO2, water and other inorganic

compounds (Nikaido and Glazer, 2007) and other simpler forms like cleaved ring

structures (Vallero, 2010). An involvement of enzyme in biodegradation describes

the situation as gratuitous biodegradation in which an enzyme transforms a

compound other than it’s natural substrate, provided the unnatural substrate is able to

occupy the enzyme’s active site and the enzyme may exhibit it’s catalytic activity

(Nikaido and Glazer, 2007).


Figure 17: Natural attenuation for reduction of environmental contaminants

Natural attenuation

Physical

Advection

Dispersion

Dilution

Diffusion

volatalization

Sorption/ desorption

Chemical

Ion exchange

Complexation

Abiotic tranformation

Biological

Aerobic Anaerobic

Plant/ animal uptake


Biodegradation is one of the energy dependent (Alves and Pereira, 2012)

natural attenuation processes that helps to eliminate xenobiotic compounds from the

environment by microbially mediated catalyzed reduction in complexity of chemical.

Extracting a microbe from the environment and exposing it to a target

contaminant under controlled conditions is a strategic approach to break down target

component into simpler non –toxic forms. This can be achieved by bioremediation

(Vallero, 2010). Bioremediation process can be divided into three phases or levels.

First, through natural attenuation, contaminants are reduced by indigenous

microorganisms without any human augmentation. Second, biostimulation is

application of nutrients and oxygen to the systems to enhance their effectiveness

and to accelerate biodegradation. Finally, during bioaugmentation, microorganisms

are added to the systems. These supplemental organisms should be more efficient

than native flora to degrade the target contaminant (Diez, 2010).

A feasible remedial technology requires microorganisms being capable of

quick adaptation and efficient uses of pollutants of interest in a reasonable period of

time (Seo et al. 2009). Many factors influence microorganisms to use pollutants as

substrates or co metabolize (Nikaido and Glazer, 2007; Vallero, 2010) them, like, the

genetic potential and certain environmental factors such as temperature, pH, and

available nitrogen and phosphorus. The rate and extent of degradation is dependent

upon environmental factors (Frit’sche and Hofrichter, 2008). Therefore, applications

of Genetically Engineered Microorganisms (GEM) with high degradative capability in

bioremediation have received a great deal of attention. However, ecological,

environmental concerns and regulatory constraints are major obstacles for testing

GEM in the field (Menn et al. 2008). Figure 18 represents bioremediation of polluants

utilizing biodegradation approach.


Figure 18: Bioremediation of pollutants through natural attenuation utilizing the combinatorial effect of bioaugmentation, biostimulation and GEM.

One means of reducing the environmental and public health risks is to change

the chemical structure of compounds by living organisms or enzymes (Rajendran and

Gunsekaran, 2006) so that do not bind or block receptor sites on cells.

Biotransformation of pollutants can be achieved by three specific mechanisms

(Vallero, 2010).

• Use of compound as an electron acceptor

• Use of compound as an electron donor

• Co-metabolism

Co-metabolism is the ability of an organism to transform a non growth

susbtrate as long as growth substrate or other transformable compound is also present

(Nikaido and Glazer, 2007) Wackett in 1996 defined co metabolism as an interaction

“between enzyme specificity and metabolic regulation, the metabolic interdependence

of microorganisms and co-substrate requirements in the catabolism of xenobiotic

compounds”. These processes can occur together and simultaneously (Batelle, 2009).

Mechanism of biotransformation is represented in Figure 19.


Figure 19: Mechanisms underlying biotransformation of pollutants (Vallero,

2010)

Among the current pollution control technologies for textile effluents, an

advantage of biological treatment over certain physicochemical treatment methods is

biodegradation of synthetic dyes by different microbes. It is emerging as an effective

and promising approach (Charumathi and Das, 2010) as that > 70% of the organic

material measured by the COD test can be converted to biosolids (Forgacs et al.

2004; Anjaneyulu et al. 2005) and is found to be an economic, effective, biofriendly,

and environmentally benign process (Verma and Madamwar, 2003; Pearce et al.

2003; Chen et al. 2003; Jirasripongpun et al. 2007; Kalyani et al. 2008; Shedbalkar

et al. 2008; Ozdemir et al. 2008 ; Gopinath et al. 2009 ; Ramchandran et al. 2013)

amount of sludge produced is also very less (Carvalho et al. 2008; Saratale et al.

2009) as contrasted with physicochemical methods (Zhang et al. 2004;

Amoozegar, 2011) which produce enormous amount of sludge and the emit toxic

substances (Johnson et al. 1978) efficiency for colour removal particularly for

sulphonated azo dyes is less (Banat et al. 1996). Therefore, bioremediation is a

publicly acceptable treatment technology (Hao et al. 2000; Supaka et al. 2004) which

may prove to be a green solution to the problem of environmental soil and water

pollution in future (Ali, 2010; Archana et al. 2012) and at the same time biological

treatment process for decolourisation of industrial effluents is ambiguous, different

and divergent (Anjaneyulu et al. 2005). Several excellent reviews have been

published on the biodegradation or bioremediation, both generally (Prescott et al.

2008; Chatterjee et al. 2008) or specifically, of xenobiotic compounds (Austin et al.


1977; Chaudhry and Chapalamadugu, 1991; Chauhan et al. 2008; Chowdhury et al.

2008). Various microorganisms such as bacteria both gram positive and gram

negative (Sani and Banerjee, 1999; Kodam et al. 2005; Moosvi et al. 2005; Kalyani

et al. 2009; Wang et al. 2009) as well as fungi (Balan and Monteneiro, 2001; Verma

and Madamwar, 2005; Taskin and Erdal, 2010) , yeasts, algae have been reported to

remove dyes (Stolz et al. 2001; Saranraj et al.2013; Ramchandran et al. 2013).

Suitable strains of microorganisms must be adapted to textile effluents prior

decolourisation activity (Saratale et al. 2009; Phugare et al. 2011) further, identified

and characterized to optimize the best conditions for effective biological treatment of

azo dyes (Grossmann et al. 2014). Generally, removal of dyes by microorganisms

takes place by biosorption which is accumulation of chemicals and dyes by microbial

mass (Hu, 1992; Bras et al. 2001) primarily by cell membranes and/or cell walls

through physical adsorption, electrostatic interaction, ion exchange, chelation and

chemical precipitation and the structure remains intact (Ali, 2010) but differs by

biomass type (Barr and Aust, 1992; Ali et al. 2007) and dye biosorption or

biodegradation by microorganism is judged by the colour of cell mat (Chen et al.

2003). This option is viable in cases when conditions are not favourable for growth

and maintainance of microbial population (Miao. Y, 2010) Biosorption of dyes does

not eradicate the problem because the pollutant is not destroyed but instead

entrapped into the matrix of the adsorbent (the microbial biomass). The disposal of

the microbial biomass containing adsorbed dyes itself is a big hurdle in their

proposed role in biocleaning of coloured waters (Chander and Arora, 2007).

Biosorption by living fungi has found be an efficient mechanism of dye

removal (Fu and Viraraghavan, 2001) and the biosorption of dyes may be of interest

in biorecovery of these synthetic chemicals from spent dye baths.

The bioremediation process is a pollution control technology (Shah et al.

2013) mediated by bioengineering the capabilities of intrinsic microorganisms

(bacteria, fungi and yeasts) which completely decolourise and mineralize the

pollutants on site from the dye laden effluent (Nachiyar et al. 2012). The advantages

of such a decolourisation and degradation (of azo dyes) process arise from the fact

that it is an aerobic treatment system where toxic intermediates like aromatic amines


produced in the effluent by abiotic and biotic means are completely mineralized by

the bacteria themselves (Barsing et al. 2011). The ubiquitous nature of bacteria

makes them invaluable tools in effluent biotreatment (Olukanni et al, 2006).

Microorganisms having high growth rate in polluted environment and minimal

nutritional requirements can be used to achieve good result in decolourisation as well

as dye degradation experiments (Chang et al. 2001). Bacterial strains selected by

adaptation or under selective pressure of environmental pollution (Chen et al. 2003;

Madamwar and Keharia, 2003) from textile effluents have been shown to

decolourize textile dyes (Saratale et al. 2009; Hemapriya et al. 2010; Phugare et al.

2011; Shah et al. 2013) under favourable conditions to optimize it’s degradation

potential (Novotny et al. 2004; Lucas et al. 2008).

Biodegradation of dyes is a biologically mediated energy-dependent process

which leads to breakdown the breakdown of dye into various byproducts through the

action of various enzymes (Kaushik and Malik 2009). Biodegradation of synthetic

dyes not only results in decolourisation of the dyes but also in fragmentation of the

dye molecules into smaller and simpler parts (breakdown products). Decolourisation

of the dye occurs when the chromophoric center of the dye is cleaved (Kaushik and

Malik, 2009). Bacterial decolourisation is normally faster (Kalyani et al. 2009). It is

well known that bacteria degrade azo dyes reductively under anaerobic conditions to

colourless aromatic amines can be toxic, mutagenic or carcinogenic (Isik and

Sponza, 2007). The carcinogenicity of an azo dye may be due to the dye it’self or

aryl amine derivatives produced during the reductive biotransformation of an azo

linkage (Dawkar et al. 2009). These colourless aromatic amines should be degraded

further because these may be toxic, mutagenic, and carcinogenic to humans and

animals (Chen, 2006). Aromatic amines formed during anaerobic cleavage of the azo

dyes could be further degraded in an aerobic treatment system (Kuhad et al. 2004;

Van der Zee and Villaverde, 2005).

According to the concept of combined anaerobic–aerobic treatment (Carvalho et

al. 2008; Lin et al. 2010) azo dyes should be removed from the water phase by

(anaerobic) reduction in which a reduction of the bond in the molecules takes place

by hydroxylation followed by (aerobic) oxidation of the dyes constituent aromatic


amines and complete mineralization of the reactive dye molecule takes place

(Lavanya et al. 2014). This treatment holds promise as a method to completely

remove azo dyes from wastewater (Van der Zee and Villaverde, 2005). Conversely,

aerobic treatment followed by anaerobic treatment of biological declourisation has

been reported as a viable alternative (Ong et al. 2005) as certain dyes are susceptible

to anoxic/anaerobic decolourisation. Under aerobic conditions, most of the azo dye

metabolites are quickly degraded by oxidation of the substituents or of the side

branches (Zaharia et al. 2012). Anaerobic reduction of the azo bond by bacteria

seems to be better suited for the decolourisation of azo dyes in treatment systems

(Amoozegar et al. 2011).

The putative advantages of this method are:

(1) The depletion of oxygen is easily accomplished in static cultures, which

enables anaerobic, facultative anaerobic and aerobic bacteria to reduce the

azo dyes.

(2) The reactions take place at neutral pH values and are expected to be

extremely unspecific when low molecular redox mediators are involved.

(3) The reduction rates generally increase in the presence of other carbon

sources.

The reduction equivalents that are formed during anaerobic oxidation of these

carbon sources are used finally for the reduction of the azo bond. The main

microorganisms contributing to biodegradation of organic compounds are bacteria

(e.g. Bacillus subtilis, Aeromonas Hydrophilia, Bacillus cereus, Klebsiella

pneunomoniae, Acetobacter liquefaciens, Pseudomonas species, Pagmentiphaga kullae,

Sphingomonas), fungi (e.g., white-rot fungi: Phanerochaete chrysosporium,

Hirschioporus larincinus, Inonotus hispidus, Phlebia tremellosa, Coriolus versicolour),

algae (e.g. Chlorella and oscillotoria species). However, in the presence of specific

oxygen-catalysed enzymes called azo reductases, some aerobic bacteria are able to

reduce azo compounds and produce aromatic amines (Stolz, 2001). The factors

which govern the process of biodegradation are summarized in Table No 10.


Table 10: Factors responsible for dye degradation by microbial process (Khan

et al. 2012)

Factors Description

pH The pH has a major effect on the effciency of dye decolourisation the optimal pH for colour removal in bacteria is often between 6.0 and 10.0. The tolerance to high pH is important in particular for industrial processes using reactive azo dyes, which are usually performed under alkaline conditions. The pH has a major effect on the effciency of dye decolourisation, the optimal pH for colour removal in bacteria is often between 6.0 and 10.0 (Chen et al. 2003; Guo et al. 2007; Kilic et al. 2007).

Temperature Temperature is also again a very important factor for all processes associated with microbial vitality, including the remediation of water and soil. It was also observed that the decolourisation rate of azo dyes increases uptothe optimal temperature, and afterwards there is a marginal reduction in the decolourisation activity.

Dye concentration

Earlier reports show that increasing the dye concentration gradually decreases the decolourisation rate, probably due to the toxic effect of dyes with regard to the individual bacteria and/or inadequate biomass concentration, as well as blockage of active sites of azo reductase by dye molecules with different structures.

Carbon and Nitrogen sources

Dyes are deficient in carbon and nitrogen sources, and the biodegradation of dyes without any supplement of these sources is very difficult. Microbial cultures generally require complex organic sources, such as yeast extract, peptone, or a combination of complex organic sources and carbohydrates for dye decolourisation and degradation.

Oxygen and agitation

Environmental conditions can affect the azo dyes degradation and decolourisation process directly, depending on the reductive or oxidative status of the environment, and indirectly, influencing then microbial metabolism. It is assumed that under anaerobic conditions reductive enzyme activities are higher; however a small amount of oxygen is also required for the oxidative enzymes which are involved in the degradation of azo dyes.

Electron donor Dyes with simpler structures and low molecular weights exhibit higher rates of colour removal, whereas the removal rate is lower in the case of dyes with substitution of electron withdrawing groups such as SO3H, SO2NH2 in the para position of phenyl ring, relative to the azo bond and high molecular weight dyes.

Redox mediator

Redox mediators (RM) can enhance many reductive processes under

anaerobic conditions, including azo dye reduction


Utilization of microorganism consortia offers considerable advantages over

the use of pure cultures in the degradation of synthetic dyes (Banat et al. 1996;

Pearce et al. 2003; Alves and Pereira, 2012). The individual strains may attack the

dye molecule at different positions or may use decomposition products produced by

another strain for further decomposition (Forgacs et al. 2004). Furthermore, mixed

culture studies may be more comparable to practical situations. With the increasing

complexity of a xenobiotic, one cannot expect to find complete catabolic pathways in

a single organism; a higher degree of biodegradation and even mineralization n1ight

be accomplished when co-metabolic activities within a microbial community

complement one another (Nigam et al. 1996; Khadijah et al. 2009).

Using mixed cultures instead of monocultures, higher degrees of

biodegradation and mineralization can be achieved due to synergistic metabolic

activities of the microbial community (Ramalho et al. 2004; Kehehra et al. 2005; Ali,

2010). The individual strains can attack dye molecules at different positions, yielding

metabolic end products that may be toxic; these can be further metabolised as

nutrient sources to carbon dioxide, ammonia and water by another strain. Other

species present may not be involved in bioremediation at all, but can stabilise the

overall ecosystem (Kandelbauer and Gilibitz, 2005). This type of mineralization is

the safest way to assure that no potentially harmful and unrecognized intermediate

degradation products are released into the environment. Mixed consortia usually do

not require sterile conditions and have greater stability towards changes in the

prevailing conditions (pH, temperature and feed composition) compared with pure

cultures (Ramalho et al. 2004). Therefore, the use of mixed cultures is a good

strategy for bioreactors.

A synergistic action (enzymatic) of fungal-bacterial consortium leads to the

enhanced degradation and detoxification of azo dyes and, thus provides an alternate

way for efficient removal of contaminants (Khelifi et al.2009b; Su et al.2009; Qu et

al.2010). Moreover the high rates of dye decolourisation by fungal-bacterial

synergism suggest it to be an appropriate powerful tool for the efficient degradation

and detoxification of azo dyes as well as textile effluent (Khelifi et al. 2009; Su et al.

2009; Qu et al. 2010; Kadam et al. 2011). Eco-friendly, efficient and short


degradation times are some of the highlights of fungal-bacterial synergism over

individual cultures. Such synergisms are more effective due to the concerted

metabolic activities, which might attack dye molecules at various positions or utilize

intermediate degradation metabolites for further mineralization into non-toxic forms

(Keck et al. 2002; Chen and Chang, 2007). It is known that, addition of intermediate

metabolites of dye decolourizing culture into another culture could enhance the azo

dye decolourisation rates (Chang et al. 2004). Different consortial approaches have

been studied due to their enhancing degradation abilities. Table 11 represents dye

decolourisation and dye degradation studies by adapted microbes under optimized

process parameters.

2.7 Enzymes involved in biodegradation

Microorganisms can decolourize the dyes with different enzyme systems.

Fungal enzymes are non-specific towards different structures of dyes and thus

oxidize a wide range of them (Aust, 1990). Fungi have been extensively studied to

degrade textile dyes due to their extracellular oxidoreductive, nonspecific and non

stereoselective enzyme system, including lignin peroxidase, laccase, manganese

peroxidase and tyrosinase (Stolz, 2001; Hofrichter, 2002; Kaushik and Malik, 2009).

The bacterial biodegradation is associated with it’s intracellular and extracellular

oxidoreductive enzyme system such as azo reductase, DCIP-reductase and laccase

(Chen et al. 2003; Kalyani et al. 2008; Telke et al. 2009a).

Biochemically, Azoreductases [NAD(P)H: 1-(4`-sulfophenylazo)-2-naphthol

oxidoreductase] are the non-specific cytoplasmic enzymes that catalyzes or

biotransform (Ramlan et al. 2011) the reductive cleavage of the azo bridge (-N=N-)

in azo dyes to produce colourless amine products (Stolz, 2001; Ramlan et al. 2011)

more toxic than the parent compound (Kumar et al. 2006; Stingley et al. 2010;

Mendes et al. 2011a ;Ramchandran et al. 2013; Kolekar et al. 2013). These enzymes

are found to be more efficient under anoxic and static conditions.


Table 11: Dye decolourisation and dye degradation studies by adapted microbes under optimized process parameters

S.

No

Source Organism Dye used OPTIMIZATION PARAMETERS Mechanism % Colour

Removal

Reference

pH Temperature

(◦C)

Duration Carbon Nitrogen Dye concentration

(mg/l)

1 Dye contaminated sludge

Pseudomonas aeruginosa Remazol Orange 3R (RO3R)

7 30 15 minutes

Glucose and starch

Yeast extract

50 Enzymatic 98 Surwase et al.2013

2 Textile effluent Bacillus licheniformis Reactive Red 2

9 37 7 days Glucose NH4NO3 50 Enzymatic 80 Balagurunathan and Sudha et

al.2013

3 Environmental treatment Plant

Pseudomonas otitidis Reactive Blue

250

7 37 8 hr. Glucose Yeast extract

100 Enzymatic 96.98 Bhatt et al.2012

4 Textile effluent Bacillus boroniphilus Reactive Yellow

145

7 30 24 hr. Glucose Yeast Extract

50-150 Enzymatic 100 Derle et al.,2012

5 Dye contaminated site

Bacterial-fungal consortium(Pseudomonas sp .SUK 1 and Aspergillus

Ochraceus NCIM-1168)

Rubine GFL 8 37 30 hr. Lactose Yeast extract

100 Enzymatic 95 Govindwar et al. 2012

6 Effluent disposal sites

Pseudomonas aeruginosa Mordant Black 17

7 37 48 hr. Glucose Ammonium nitrate

100 Enzymatic 86 Karunya et al.2013

7 Waste water treatment plant

Bacterial consortium

SpNb 1

Reactive Red M8B

7.5 37 48 hr. Lactose Yeast Extract

300 Enzymatic 96.75

Bhatt et al. ,2012

8 Tannery effluent Shigella sp Acid Red 113 7 37 72 hr. Dextrose L-methionine

200 - 99 Sivranjani et al.2013


S.

No

Source Organism Dye used OPTIMIZATION PARAMETERS Mechanism % Colour

Removal

Reference

pH Temperature

(◦C)

Duration Carbon Nitrogen Dye concentration

(mg/l)

9 Textile effluent Bacillus sp –ETL 79 Crystal Violet 8 35 24 hr. Dextrose Peptone 100 Enzymatic 95 Shah et al.2013

10 National chemical laboratory, Pune

Micrococcus glutamicus NCIM 2168

Reactive Red 195

5 37 18 hr. Sucrose Yeast extract

100 Enzymatic 94.25 Sahasrabudhe et al.,2014

11 Grit tank of Tannery effluent treatment plant

Pseudomonas otitidis SA 1 CR 7 37 24 hr. - - 100 Biosorption 94 Saharan et al.2013

12 Sugarcane waste water

Bacillus sp SCWS 5 CR

7 35 16 hr. Sucrose Beef extract 150 Enzymatic 100 Pradhan et al.2012

14 Soil contaminated by textile effluent

Staphylococcus hominis Acid orange 7 35 60 hr. Glucose Yeast extract

100 Enzymatic 90 Singh et al.2014

15 Tannery division, Central Leather Research Institute, Chennai

Bacillus subtilis Acid blue 113 7 37 50 hr. Starch peptone 200 Enzymatic 90 Gurulakshmi et al.2008


On the contrary, a report suggests, azoreductases obtained from different

microorganisms are known to have broad specificities in their enzymatic reactions

(Nakanishi et al. 2001) and variations can exist among the same organism (Ghosh et

al. 1992). Identification, purification and characterization of azoreductase constitute

a straightforward approach for the development of azo dye biodegradation systems

(Syed et al. 2009; Milikli and Rao, 2012; Bhatt et al. 2013).

Classifiying an azoreductase based on it’s primary amino acid level is

difficult due to low homology. However, a classification scheme based on the

secondary and tertiary amino acid analysis has been developed (Abraham, 2007).

Based on function, another classification scheme is used in which azoreductases are

categorized as either flavin-dependant azoreductases (Nakanishi et al. 2001; Chen et

al. 2004; Chen et al. 2005) or flavin-independent azoreductases (Blumel and Stolz,

2003). The flavin-dependent azoreductases are further organized into three groups;

(1) NADH only (Nakanishi et al. 2001; Chen et al. 2004), (2) NAD(P)H only (Chen

et al. 2005) or (3) both (Ghosh et al. 1992) (Figure20).

Figure 20: Classification of azoreductases based upon function

The reaction catalyzed by azo reductases occurs only in presence of reducing

equivalents like FADH and NADH. Most of the azo dyes have sulphonate substituent

groups possessing high molecular weight which does not pass through cell

membranes. Therefore, the reducing activity of the dye does not depend on the

intracellular uptake of the dye (Robinson et al. 2001). Russ et al. 2000 suggested that


bacterial membranes are almost impermeable to flavin containing cofactors and,

therefore, the transfer of reduction equivalents by flavins from the cytoplasm to the

sulphonated azo dyes is restricted. Thus, a mechanism other than reduction by

reduced flavins formed by cytoplasmic flavin-dependent azoreductases must be

responsible for sulphonated azo dye reduction in bacterial cells with intact cell

membranes. One such mechanism involves the electron transport-linked reduction of

azo dyes in the extra-cellular environment. To achieve this, the bacteria must

establish a link between their intracellular electron transport systems and the high

molecular weight, azo dye molecules.

For such a link to be established, the electron transport components must be

localized in the outer membrane of the bacterial cells (in the case of gram-negative

bacteria), where they can make direct contact with either the azo dye substrate or a

redox mediator at the cell surface (Myers and Myers, 1992). In addition, low

molecular weight redox mediator compounds can act as electron shuttles between the

azo dye and an NADH dependent azo reductase that is situated in the outer

membrane. These mediator compounds owe their origin to the metabolism of certain

substrates by the bacteria or they may be added externally from some other source

(Russ et al.2000). Many proteins and enzymes are not naturally present in an

aqueous either outside or inside of cells. In gram negative bacteria, periplasmic

proteins are present which are partially immobilized between outer and inner

membranes of the cell envelope (Scopes, 2004). Kudlich et al. 1997 support the

suggestion that the membrane-bound azo reductase activity, mediated by redox

compounds, is different from the soluble cytoplasmic azo reductase that is

responsible for the reduction of non-sulphonated dyes that permeate through the cell

membrane. Therefore, the membrane-bound and the cytoplasmic azo reductases are

two different enzyme systems (Kudlich et al. 1997). Although the final reduction of

the azo dyes in the cell supernatants is a dominantly chemical redox reaction, the

redox mediators depend on cytoplasmic reducing enzymes to supply electrons (Yoo

et al. 2001). It is also possible that this chemical redox reaction works in conjunction

with a direct enzymatic reaction involving an azo reductase, which may be a


dehydrogenase enzyme that is synthesized throughout the cytoplasm and secreted

without accumulation inside the cell (Bragger et al. 1997).

Depending upon oxygen utilization, there are two broad classes of

azoreductases; - the true azoreductases produced under aerobic conditions and those

that are produced under anaerobic conditions. Aerobic azoreductases catalyze

reductive metabolism of azo dyes in the presence of molecular oxygen. Previous

studies conducted by Zimmermann and co-workers in 1982 and 1984 on

Pseudomonas strains K22 and KF46 revealed two azoreductases capable of

decolourizing Orange I and II respectively. The enzymes were purified, characterized

and compared. It was found that both azoreductases were monomeric flavin-free

enzymes that preferentially used NAD(P)H and to some extent NADH as co-factors.

They also demonstrated the ability to reductively cleave sulfonated substrates in

addition to the carboxylated growth substrates. Azoreductases transfer the reducing

equivalents originating from the oxidation of organic substrates to the azo dyes.

Studies with some aerobic and facultative aerobic bacteria demonstrated that azo

compounds can be utilized as the sole source of carbon, with azo bond cleavage

being facilitated by enzymes that are thought to be either intracellular or membrane

bound (Van der Zee et al. 2002). These specifically adapted strains synthesize true

azoreductases which reductively cleave the azo group in the presence of molecular

oxygen.

Azoreductase activity has been identified in several bacteria, such as

Xenophilus azovorarts KF46 (Blumel et al. 2001), Pseudomonas luteola (Hu, 1998),

Rhodococcus (Heiss et al. 1992), Shigella dysenteriae Type I (Ghosh et al. 1992),

Klebsiella pnumoniae RS-13 (Wong and Yuen, 1996) and Clostridium perfringens

(Rafii et al. 1997). Cloning of genes encoding azoreductase were also carried out

from various bacteria, such as Geobacillus stearothermophilus 0Y1-2 (Ohnishi et al.

2011), Xenophilus azovorans KF46F (Blumel et al. 2002) and Escherichia coli

(Nakanishi et al. 2001). Recently, such genes were found in Bacillus anthracis (Read

et al.2003) and Bacillus cereus ATCC 10987 (Rasko et al. 2004). Because of their

high biodegradation capacity, they are of considerable biotechnological interest and

their application in the decolourisation process of wastewaters has been extensively

investigated (Young and Yu, 1997).


Enzymes have a number of features that make them more viable in relation to

conventional catalysts: they are biodegradable catalysts, allow the operation at low

and high substrate concentrations, allow the operation over a wide range of pH,

temperature and salinity, have no delays associated with the acclimatization of

biomass, have a reduced sludge formation, are simple and easy to control (Nicell et

al. 1993). Another advantage of using pure enzymes instead of the microorganism is

that the expression of enzymes involved in dye degradation is not constant with time,

but dependent on the growth phase of the organisms, possess greater specificity,

better standardization, easy handle and storage (Roges-Milgarejo et al. 2006).

Conversely, Roges-Milgarejo et al .in 2006 was of an opinion that the expression of

an azoreductase was not dependent on bacterial growth rates and is influenced by

inhibitors that may be present in the effluent. Synthetic or natural redox mediators

have, to be added many times to the enzymatic bath in order to achieve the total

capacity of the enzyme(s) or even to make their work possible (Alves and Pereira,

2012). In order to increase the potential use of enzymes in a wastewater

bioremediation process, their immobilisation is recommended for biochemical

stability and reuse, thereby reducing the cost (Duran and Esposito, 2000;

Kandelbauer et al. 2004). The major drawback of using enzyme preparations is that

once the enzymes become inactivated, it is of no use. Because enzymes can be

inactivated by the presence of the other chemicals, it is likely that enzymatic

treatment will be most effective in streams that have the highest concentrations of

target contaminants but the lowest concentration of other compounds that could

interfere with their action (Alves and Pereira, 2012). The recent biotechnological

advances have allowed the production of cheaper and more available enzymes

through improved purification and isolation processes. All these advantages add to

their high specificities and catalytic activities with the possibility of designing

enzymes with the exactly desired properties through genetic engineering and

computational design, suggest the potential application of this process in the

treatment of effluents (Call and Mucke, 1997).

2.8 Genes behind the Azoreductase

Azoreductase activity in azo dyes decolourisation has been extensively

examined to elucidate azo dye reduction mechanism (Chen et al. 2005, Deller et al.


2006, Wang et al. 2007, Ryan et al. 2010a, Ryan et al. 2010b and Feng et al. 2012).

Only few reports have reported the regulation of azoreductase gene expression

(Töwe et al. 2007, Liu et al. 2009a; Ryan et al. 2010a). Ryan et al. 2010a reported an

increase in mRNA levels for azoreductase genes (ppazoR1, ppazoR2 and ppazoR3)

from Pseudomonas aeruginosa in the presence of azo dyes. A significant increase in

azoreductase mRNA levels including azoR1 and azoR2 have been observed in B.

subtilis in the presence of quinines (Töwe et al. 2007). It was reported that azoR1 and

azoR2 are negatively regulated by redox-sensing transcription factors YodB and

YkvE, respectively (Töwe et al. 2007; Leelakriangsak et al. 2008). Azoreductases

AzoR1 and AzoR2 not only have azoreductase activity but also have quinone

reductase activity that plays a role in bacterial protection thiol-specific stress

(Nishiya and Yamamoto, 2007; Töwe et al. 2007, Leelakriangsak et al. 2008,

Leelakriangsak and Borisut, 2012). More recently, evidence was presented that

azoreductase possess quinone reductase and nitroreductase activity (Rafii and

Cerniglia, 1993, Liu et al. 2008a, Liu et al. 2009a). The flavin-dependent

azoreductases AZR, AzoR from Rhodobacter sphaeroides and E. coli, respectively,

overexpressed in E. coli have quinone reductase activity by reducing quinone

compounds as substrate. Moreover, the quinone compounds were better substrates

for AzoR than the model azo dye substrate Methyl Red (Liu et al. 2009a).

Interestingly, the presence of quinone compound accelerated the azo dye

decolourisation of overexpressed azoreductase AZR (Liu et al. 2009b; Parshetti et al.

2010). A significant increase in the enzyme activities of azoreductase and NADH-

DCIP reductase over a period of methyl orange decolourisation by K. rosea MTCC

1532 was observed. A similar result of an increase in azoreductase and DCIP

reductase activity was also observed when Alishewamella sp. KMK6 exposed to dyes

(Kolekar et al. 2013). A putative azoreductase gene (so3585) of Shewanella

oneidensis was found to up-regulate in response to heavy metal (Murgerfeld et al.

2009). However, the results were suggestive reduction of an azo dye is not the

primary function of the SO3585 protein in vivo. Figure 21 represents a proposed

catalytic cycle for an azoreductase enzyme.


Figure 21: A proposed catalytic reaction of azoreductase. Azoreductase reduces the azo compound via Ping Pong Bi Bi mechanism, with two cycles consuming NAD(P)H, reducing the azo substrate to a hydrazine (partially reduced intermediate) in the first cycle and to two amines in the second cycle (Bin et al. 2004, Ryan et al. 2010b, Wang et al. 2010).

2.9 Genetic Engineering of dye degrading organisms

There are at least four principal approaches to GEM development for

bioremediation application (Menn et al. 2008) (Figure 22)

Figure 22: Approach to GEM development for dye degradation

These include: (1) Modification of enzyme specificity and affinity; (2)

Pathway construction and regulation; (3) Bioprocess development, monitoring and


control; (4) Bioaffinity, bioreporter sensor applications for chemical sensing, toxicity

reduction and end point analysis.

Identification, isolation, and transfer of genes encoding degradative enzymes

can greatly help in designing microbes with enhanced degradation capabilities

known as super degraders. Of the two approaches, acclimatization is natural, since

in this case the built-in genetic setup of the microorganism is not disturbed; only

some components are enabled. On the other hand, in genetic engineering, the natural

genetic set-up of a microorganism is changed by incorporating a new gene or genes.

By cloning and transferring genes encoding for dye degrading enzymes, organisms

could be designed that combine the abilities of mixed cultures within a single

species. Molecular cloning of the gene encoding azoreductase enzyme followed by

protein purification is likely to be crucial for further characterization and application

of this enzyme (Suzuki et al. 2001; Wang et al. 2007; Ryan et al. 2010a; Wang et al.

2010; Mendes et al. 2011a). A number of genes conferring the ability of dye

decolourising have been identified. Successful decolourisation of an azo dye using

Escherichia coli carrying the azoreductase gene from a wild-type Pseudomonas

luteola has been reported (Chang et al. 2000; Chang and Kuo, 2000). This approach

could become a useful alternative for shortening the extended time-periods otherwise

needed to adapt appropriate cultures and isolated strains, respectively. CotA-laccase,

a bacterial enzyme from Bacillus subtilis cloned and over-expressed in E. coli, has

proved to be a thermoactive and intrinsically thermostable enzyme with a high

capacity for the decolourisation of azo and anthraquinonic dyes (Pereira et al. 2009a,

b). The expression level of CotA-laccases in different E. coli host strains, growing

under different culture conditions, was compared and a high-throughput screenings

for the oxidation of dyes with high potential redox developed by (Brissos et al.

2009). Chen et al. 2010 described the cloning of azoreductase gene azoB from

Pigmentiphaga kullae K24. The recombinant azoreductase expressed in E. coli

exhibited optimal for activity of Orange I at pH 6.0 at temperatures between 37 and

45°C. Both NADH and NAD(P)H can be used as an electron donor but NAD(P)H is

preferred. The gene was characterized (Ooi et al. 2007). The recombinant

azoreductase expressed in E. coli exhibitted broad pH stability between 6 and 10


with an optimal temperature of 60-80°C. AzrA effectively decolourized Methyl Red,

Orange I, Orange II and Red 88. No enzyme activity was detected for Orange G and

New Coccin. In addition, the enzyme activity of AzrA was oxygen insensitive and

required NADH as electron donor for dye reduction. Similar results have also been

described for azoreductase enzyme activity extracted from B. velezensis and P.

aeruginosa (Nachiyar and Rajakumar, 2005, Bafana et al. 2008). Furthermore, a

gene encoding NAD(P)H-flavin azoreductase (Azo1) from the skin bacterium

Staphylococcus aureus ATCC 25923 overexpressed in E. coli demonstrated that this

azoreductase is able to decolourize a wide range of structurally complex azo dyes

(Chen et al. 2005). The Azo1 cleaved the model azo dye Methyl Red and sulfonated

azo dyes Orange II, Amaranth and Ponceau. However, no enzyme activity was

observed when Orange G was used as substrate. Recently, the gene encoding an

FMN dependent NADH azoreductase AzrG from thermophilic Geobacillus

stearothermophilus was cloned and expressed in recombinant E. coli (Matsumoto et

al. 2010). The optimal temperature of AzrG was 85°C for Methyl Red degradation

and enzyme also showed a wide range of degrading activity towards several

tenacious azo dyes such as Acid Red 88 Orange I and CR.

Physiochemical properties, enzyme characterization and kinetic studies can

be investigated by obtaining purified azoreductase from whole cell extract from the

source organism or recombinant cell extract (Nachiyar and Rajakumar, 2005; Wang

et al. 2007; Gopinath et al. 2009; Punj and John, 2009; Mendes et al. 2011a;

Morrison et al. 2012). Purification from whole cell extracts from the source organism

employs classical purification procedures which require many steps such as

ammonium sulfate precipitation followed by ion exchange and affinity

chromatography methods (Maier et al. 2004; Nachiyar and Rajakumar, 2005; Punj

and John, 2009; Kolekar et al. 2013). However, in most cases recombinant DNA

techniques permit the construction of fusion proteins in which specific affinity tags

are added to the protein sequence of interest (Bin et al. 2004; Wang et al. 2007).

Therefore, the purification of the recombinant fusion proteins is simplified by

employing affinity chromatography methods. In addition, the expression and

purification of recombinant proteins facilitate the production and detailed


characterization of virtually any protein. Native molecular weight of a protein can be

determined by native gel electrophoresis and/or size exclusion chromatography

(Moutaouakkil et al. 2003; Deller et al. 2006; Ooi et al. 2007; Wang et al. 2007).

The purification and characterization experiments of enzymes were conducted and

the results indicated that the enzyme activity differs in substrate specificity and

preferential coenzymes serving as electron donors. Concludingly, characterization of

recombinant azoreductases provide information for understanding these azoreductases

properties such as enzyme stability and activity, kinetic constants, cofactor

requirement, substrate profile, structure and mechanism (Wang et al. 2007, Ooi et al.

2009, Macwana et al. 2010, Ryan et al. 2010b, Mendes et al. 2011a). A broad range

of substrate specificity and thermostability are important factors in determining the

range of biologically degradable of azo dyes. Table 12 represents summarized

information of cloning and expression of recombinant azoreductase in Escherichia

coli.

Table 12: Cloning and overexpression of azoreductases (Leelakriangsak, 2013)

Organism Gene Expression vector

Molecular weight

Co-factor Reference

Pseudomonas putida MET94

ppAzoR pET-21a Homodimer 40 kDa

FMN, NAD(P)H

Mendes et al. 2011a

Pseudomonas aeruginosa

paazor1 paazor2 paazor3

pET-28b pET-28b pET-28b

Tetramer 110 kDa 23 kDa* 26 kDa*

FMN, NAD(P) H NADH NADH

Wang et al. 2007 Ryan et al. 2010b Ryan et al. 2010b

Pigmentiphaga kullae K24

azoB pET-11a Monomer 22 kDa

NAD(P)H Chen et al. 2010

Geobacillus stearothermophilus

azrG pET-3a Homodimer 23 kDa

FMN, NADH

Matsumoto et al. 2010

Bacillus sp. B29 azrA azrB azrC

pET-3a pET-3a pET-3a

Homodimer 48 kDa Homodimer 48 kDa Homodimer 48 kDa

FMN, NADH FMN, NADH FMN, NADH

Ooi et al. 2007 Ooi et al. 2009 Ooi et al. 2009

Bacillus subtilis yvaB, (azoR2)

p-Bluescript Homodimer 45 kDa

NADH Nishiya and Yamamoto, 2007


Organism Gene Expression vector

Molecular weight

Co-factor Reference

Escherichia coli acpD (azo R)

pET-22b Homodimer 46 kDa

FMN, NADH

Nakanishi et al. 2001

Enterococcus faecium

acpD pET-15b 23 kDa* NAD(P)H Macwana et al. 2010

Enterococcus faecalis

azoA PET-11a Homodimer 43 kDa

FMN, NADH

Chen et al. 2004

Xenophilus azovorans KF46F

azoB pET-11a Monomer 30 kDa

NAD(P)H Blumel et al. 2002

Clostridium perfringens

azoC pET-15b Tetramer 90.4 kDa

FAD, NADH

Morrison et al. 2012

*Molecular weight determined by SDS-PAGE

2.10 Confirmatory studies on dye degradation

Dyes, as colourants, absorb in the visible region of the spectra and each one

has a maximal wavelength, depending on it’s visible colour therefore, the easiest way

to monitor dye degradation is by means of spectrophotometry, following the decrease

in it’s absorbance. By this technique, all the molecules present are quantified, and

intermediates and degradation products will contribute to the spectra absorbance.

Various basic and advanced instrumental techniques of chromatography such as Gas

Chromatography (GC), High Performance Liquid Chromatography (HPLC), Nuclear

Magnetic Resonance Spectroscopy (NMR), Mass Spectrometry (Ion-trap, MALDI)

and Capillary Electrophoresis (CE) are available to assist in the isolation and

characterization of the intermediates and products of dye degradation, thereby giving

new insight into the mechanism of biodegradation. Prior procedures of extraction of

the aqueous sample with an organic solvent or filtration are adopted when a

heterogeneous catalyst or solid reactant is employed, or when a pre-separation is

needed (Pereira and Alves, 2012).

Shah, 2014 conducted the study to investigate the decolourisation and

degradation of azo dyes using bacteria isolated from textile dye effluent. Three

different bacterial species were isolated and the isolates were identified as Bacillus

subtilis, Pseudomonas aeruginosa, and Psuedomonas putida. Decolourisation was

attained in the range from 65%-95% with 500 mg/l of dye amended with trace


amounts of yeast extract, glucose and sucrose and then The degradation product after

decolourisation was examined by TLC and FTIR.

Sahasrabhude et al. 2014 studied the effect of Micrococcus glutamicus NCIM

2168 on degradation of Reactive red 195 and exhibited a tremendous decolourisation

potential of 94.25% in 18 hr. under static conditions. Decolourisation of dye was

attainable over a wide range of pH from 5-8 and in a narrow temperature from 37-

40°C. Effect of various carbon and nitrogen sources on decolourisation revealed

enhanced decolourisation in presence of sucrose followed by glucose, yeast extract

and peptone. Degradation of the dye was confirmed by UV-Vis spectroscopy, TLC,

HPLC and GC-MS.

Surwase et al. 2013 studied the biotransformation Remazol Orange 3R

(RO3R) by Pseudomonas aeruginosa strain BCH. 98 % decolourisation was attained

within 15 min. The RO3R was transformed to the N-(7 amino 8 hydroxynapthalen-

2yl) actamide (m/z, 198), Acetamide (m/z, 59) and Napthalen-1-ol (m/z, 144).

Pseudomonas sp. has been a focal point in many research investigations

pertaining to dye degradative studies. An effluent adapted Pseudomonas sp was

studied for decolourisation and degradation of azo dye and decolourisation attainable

was 98% at 50mg/L within five hr. in static anoxic condition. The optimum pH and

temperature for the decolourisation was 8.0 & 37°C respectively. The

biodegradation was monitored by FTIR analysis (Shah et al. 2013). Another study

was carried out to explore the decolourisation and degradation of Direct Blue 71 by

Pseudomonas aeruginosa. The bacterium was able to decolourize the dye medium to

70.43 % within 48 h under microaerophilic conditions. The degradation metabolites

formed were studied using UV–Vis techniques, HPLC, Fourier Transform Infra Red

(FTIR) spectroscopy and nuclear magnetic resonance spectroscopy analysis. Data

obtained provide evidence for the formation of aromatic amines and their subsequent

oxidative biodegradation (Hafshejani et al. 2013).

Bhatt and co-workers in 2012 screened bacteria for their dye decolourising

activity and the most promising bacterial isolate was used for further dye degradation

studies. The strain showed complete decolourisation of the selected dye (RB 250-100


mg/L) within 8 hr in static condition. The optimum pH, temperature, inoculum size

and carbon and nitrogen sources for the decolourisation was studied at pH 7.0, 37°C,

glucose (0.2 %) and nitrogen (0.5 %) respectively. The biodegradation was

monitored by UV-Vis, HPTLC and FTIR analysis.

A bacterial isolate Bacillus licheniformis was able to degrade Reactive Red 2

dye, optimally at pH 9, temperature at 37°C, dye concentration of 50 mg/l at 20%

inoculum size. Glucose, NH4NO3 were found to be the best additional carbon and

nitrogen sources. The extracellular enzyme from Bacillus licheniformis was studied

for dye decolourisation potential. Biodegradation was confirmed by analyzing the

product using TLC an GC-MS analysis indicated the formation of 2, 4-dichloro-6-

[(1H-indazol-5-ylimino)-methyl]-phenol, benzene sulfonamide, 1H indole and urea

as final metabolites formed by Bacillus licheniformis (Sudha and Balgurunathan,

2013).

Derle et al.in 2012 isolated strain of Bacillus boroniphilus which showed

appreciable ability of decolourisation of Reactive Yellow 145 dyes and exhibited

maximum decolourisation 98% in static condition and supplemented with urea and

yeast extract. Further, biodegradation of azo dye was analyzed by TLC, UV-Vis

spectrophotometry and FTIR, results showed that -N=N- (azo bond) get converted

into –NH2 (amino group), which proves accomplishment of biodegradation of

Reactive Yellow 145.

The marine ecofriendly bacteria Lysobacter sp. T312D9 isolated from Abou

Quir Gulf, Alexandria, Egypt represent an inexpensive and promising marine

bacterium for removal of both methyl and CR (Barakat, 2012). The biodegradation

was analyzed by (GC-MS) analysis and FTIR. In another study, biodegradation of

CR by Fusarium sp TSF-01 was monitored by FTIR and generation of new peaks

other than the parent compound confirmed biodegradtion of the dye by fungus

(Shinde and Thorat, 2013).

GC-MS has been widely used to identify products of dyes degraded with

bacterial monoculture and consortia both in immobilized and non-immobilized forms

in textile effluents (Soundararajan et al. 2012).


Not only does textile effluents or soil contaminated by dyes had proved to the

richest source dye degraders, tannery effluent does house some efficient dye

degrading bacteria belonging to Bacillus cereus, the decolourisation potential of

which has been well established for the Acid Black azo dye. The optimum activity of

Bacillus cereus was found at pH 7.3, temperature 37ºC, and duration of 4 days. The

decolourisation rate was found out to be 80% and 96% for the agitated and static

conditions. UV–Vis spectra analysis showed that the peaks which appeared in the

visible region of the treated dye disappeared, indicating complete decolourisation of

dye. The FTIR analysis for the Bacillus cereus treated sample showed the

transformation of azo linkage into N2 or NH3 or incorporated into complete biomass

(Kanagraj, 2011).

Bhimani, 2011 carried out a study on biodegradation of azo dyes (Reactive

Black 5 and Green B) by bacterial strains isolated from textile effluent obtained from

Jetpur, Gujarat. The biodegradation of azo dyes was preliminarily investigated by

TLC and further confirmed by (HPLC).

In another study aimed at monitoring biodegradtion of reactive dyes, bacteria

were screened for their biodegradative potential of commonly used textile dyes like

Reactive Lanasol Black B (RLB), Eriochrome Red B (RN) and 1, 2 metal complexes

I. Yellow (SGL) and an appreciable decolourisation ranging from 57%-100% was

attained in 15 hr. The biodegradation products of RLB formed during anaerobic and

sequential anaerobic/aerobic treatments were analyzed by HPLC. Peaks at different

retention times were observed in the anaerobic stage, and these peaks completely

disappeared at the end of anaerobic/aerobic incubation. This result clearly indicates

that the dye had been catabolized and utilized by isolates (Pahlaviani et al.2011).

Actinomycetes have also exhibited a tremendous potential to decolourise dye

stuffs. Streptomyces krainskii, SUK -5 was found to decolourize and degrade textile

dye Reactive blue–59.This azo dye was decolourized and degraded completely by

Streptomyces krainskii SUK–5 at 24 hour in shaking condition in the nutrient

medium at pH 8. Induction in the activity of Lignin Peroxidase (LiP) and NADH-

DCIP Reductase and MR reductase represents their role in degradation.


2.11 Colour removal in Real Dye Waste Water (RDWW): An approach to

combat dye influxed wastewaters

RDWW not only comprises of dyes, but is also influxed with surfactants,

salts, chelators, precursors. Despite of immense decolourisation efficacy of microbial

strains, colour removal in RDWW remains a challenging task (Wesenberg et al.

2003). Bioremediation systems were commonly applied to treat industrial effluents

contaminated with different types of synthetic dyes by bacteria, fungi, yeast and

actinomycetes respectively by bioaccumulation (Ramchandran et al. 2013).

Bioaccumulation of chemicals/ pollutants process takes place by respiration or food

intake wherein the concentration of target pollutant is more in vivo than the external

environment (Mebrat, 2006). Fungal strains are preferred over bacterial counterparts

for remediating RDWW as they are capable of growing over a wide range of pH and

can resist dye toxicity even at high concentrations (Kaushik and Malik, 2009).

Treatment of RDWW collected from Tirupur and Erode districts of Tamil Nadu were

treated by dye degrading bacteria in monocultures and consortial forms (Mohan et al.

2013). A similar approach had been adopted to treat textile effluent collected from

CETP of Jetpur, Gujrat by Lysinibacillus fusiformis JTP-23 utilising a sequential static

and agitating incubation (Bhimani, 2011). Exiguobacterium sp. RD3 isolated from dye

contaminated site had be known to degrade textile effluent (Dhanve et al. 2014).

Exposing the target organism prior it’s degradation mechanism has been exemplified

has been a theme of research interest (Jalandoni –Buan et al. 2010). Likewise, a

similar approach to study biodegradation of RDWW, a comparison between effluent

adapted and effluent non-adapted strains has been established (Leena and Selva Raj,

2008).

Studies based on treating the effluents by industrial scale bioreactors have

also been stated in literature (Ali, 2010; Devassy, 2010). To monitor the exclusive

effect of fungal-bacterial consortium on biodegradation of RDWW, the approach

based at sterilizing the effluent. This strategy overcomes the inclusive effect of

residential microflora on biodegradation of real dye wastewater (Lade et al. 2012).

review of literature 20 - inflibnetshodhganga.inflibnet.ac.in/bitstream/10603/89379/8/08_chapter...

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