**Microbial Degradation of Persistent Organophosphorus Flame Retardants**

Shouji Takahashi, Katsumasa Abe and Yoshio Kera

Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/53749

### **1. Introduction**

### **1.1. Flame retardants**

Flame retardants (FRs) are chemicals used in polymers to protect the public from accidental fires by preventing or retarding the initial phase of a developing fire (EFRA, 2007). These chemicals are now found in numerous consumer products, including construction materials, upholstery, carpets, electronic goods, furniture and also children's products such as car seats, strollers and baby clothing. FRs have become indispensable to modern life, and have saved numerous lives by preventing unexpected fires across the globe.

FRs are divided into two general classes based on their relation to host polymers: addi‐ tive and reactive FRs (WHO, 1997). Additive FRs are simply mixed with host polymers. The lack of chemical bonding between the FRs and host polymers enables the FRs to leach out of or volatilize from host polymers over time into the ambient environment. Reactive FRs are incorporated into host polymers by covalent bonding into the polymer backbone, and are thus less likely to leach into the environment. Additive FRs are main‐ ly used in thermoplastics, textiles and rubbers, whereas reactive FRs are usually used in thermoset plastics and resins (SFT, 2009a).

FRs are sub-divided into six groups characterized by their chemical composition: 1) alumi‐ num hydroxide, 2) brominated, 3) organophosphorus, 4) antimony oxides, 5) chlorinated and 6) other FRs. These groups account for 40%, 23%, 11%, 8%, 7% and 11% of the annual FR global consumption in 2007, respectively (Beard & Reilly, 2009). The total market for FRs in the United States, Europe and Asia in 2007 amounted to about 1.8 million tons.

© 2013 Takahashi et al.; licensee InTech. This is an open access article distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited. © 2013 Takahashi et al.; licensee InTech. This is a paper distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

### **1.2. Organophosphorus flame retardants**

Organophosphorus flame retardants (PFRs) are based primarily on phosphate esters, phosphonate esters and phosphite esters. The total consumption of FRs in Europe was an estimated 465,000 tons in 2006, of which 20% comprised PFRs (KLIF, 2010). Of the PFRs consumed, 55% were chlorinated. Halogenated PFRs are the preferred form of FRs because halogen inhibits flame formation in organic materials, and non-halogenated PFRs are typically used as flame-retardant plasticizers (KLIF, 2010).

brittleness of flame-resistant rigid or semirigid polyurethane foams. More recently, it has been used as a flame-retarding plasticizer and viscosity regulator in unsaturated polyest‐ er resin (accounting for around 80% of current use) (EURAR, 2009). TCEP-containing polymers are commonly used in the furniture, textile and building industries (for exam‐ ple, more than 80% of the TCEP consumption in the EU is invested in roofing insula‐ tion). TCEP is also used in car, railway and aircraft materials, and in professional paints. Since the 1980s, TCEP has been progressively replaced by other flame retardants, pri‐ marily tris(1-chloro-2-propyl) phosphate (TCPP). Consequently, global consumption of TCEP in the EU, which exceeded 9,000 tons in 1989, declined to below 4,000 tons by

Microbial Degradation of Persistent Organophosphorus Flame Retardants

tris(2-chloropropyl) phosphate (EURAR, 2009)

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93

Tris(2-chloroethyl) phosphate Tris(β-chloroethyl) phosphate 2-chloroethanol phosphate Phosphoricacid,tris(2-chloroethyl) ester Tris(2-chloroethyl) orthophosphate Tris(chloroethyl) phosphate

> TCEP TClEP

0.00114 Pa (20°C, extrapolated)

1997. TCEP is no longer produced in the EU (EURAR, 2009).

Synonym:

Abbreviation:

*n*-Octanol/water partition coefficient: tris(1,3-dichloro-2-propyl) phosphate (US EPA, 2005)

Tris(1,3-dichloro-2-propyl) phosphate Tris-(2-chloro-,1-chloromethyl-ethyl) phosphate 1,3-dichloro-2-propanol phosphate Phosphoricacid, tris(1,3-dichloro-2 propylester) Tris(1,3-dichloroisopropyl) phosphate Tris(1-chloromethyl-2-chloroethyl) phosphate Tri(β, β'-dichloroisopropyl) phosphate

> TDCPP TDCP

Molecular weight: 430.91 285.49

Melting point: -58°C <-70°C

Density: 1.52 1.4193 (25°C)

Vapor pressure: 0.01 mmHg (30°C) 43 Pa (136.9°C)

Water solubility: 42 mg/L 7.82 g/L (20°C)

**Table 1.** General aspect of Tris(1,3-dichloro-2-propyl) phosphate (TDCPP) and tris(2-chloroethyl) phosphate (TCEP)

Physical state: Viscous, clear liquid Clear, transparent, Low viscosity liquid

Boiling point: 236-237°C at 5 mm Hg Decomposition at 320°C at 1013 hPa

2.4 1.78

Cas number: 13674-87-8 115-96-8

### **1.3. Tris(1,3-dichloro-2-propyl) phosphate and tris(2-chloroethyl) phosphate**

Tris(1,3-dichloro-2-propyl) phosphate (TDCPP) and tris(2-chloroethyl) phosphate (TCEP) are typical examples of additive chlorinated PFR (Fig. 1 and Table 1).

**Figure 1.** Chemical structure of tris(1,3-dichloro-2-propyl) phosphate (TDCPP) and tris(2-chloroethyl) phosphate (TCEP)

TDCPP is a viscous colorless to light yellow liquid and is produced by the epoxide opening of epichlorohydrin in the presence of phosphorus oxychlorine (ATSDR, 2009). TDCPP is used primarily in flexible polyurethane foams but also in rigid polyurethane foams, resins, plastics, textile coatings and rubbers (California EPA, 2011). TDCPP was a common ingredi‐ ent of sleepwear for children in the 1970s, but was voluntarily withdrawn by manufactures in 1977 because of its proven mutagenicity (California EPA, 2011). However, the PFR can still be found in many baby products (Stapleton et al., 2011). Currently, TDCPP is used mostly in flexible polyurethane foams for upholstered furniture and automotive products. TDCPP consumption has increased following the ban on common FR polybrominated di‐ phenyl ethers (PBDEs). Consequently, total TDCCP production has increased, being an esti‐ mated 4,500-22,700 tons in the United States in 2006 and <10,000 tons in Europe in 2000 (van der Veen & de Boer, 2012).

TCEP is colorless to pale yellow liquid and is highly soluble in water (Fig. 1 and Table 1). The compound is chemically synthesized via condensation of phosphorus oxychlor‐ ide and chloroalkyl alcohol at low temperatures and pressures to avoid formation of al‐ kyl chlorides (ATSDR, 2009). Previously, the main purpose of TCEP was to reduce the brittleness of flame-resistant rigid or semirigid polyurethane foams. More recently, it has been used as a flame-retarding plasticizer and viscosity regulator in unsaturated polyest‐ er resin (accounting for around 80% of current use) (EURAR, 2009). TCEP-containing polymers are commonly used in the furniture, textile and building industries (for exam‐ ple, more than 80% of the TCEP consumption in the EU is invested in roofing insula‐ tion). TCEP is also used in car, railway and aircraft materials, and in professional paints. Since the 1980s, TCEP has been progressively replaced by other flame retardants, pri‐ marily tris(1-chloro-2-propyl) phosphate (TCPP). Consequently, global consumption of TCEP in the EU, which exceeded 9,000 tons in 1989, declined to below 4,000 tons by 1997. TCEP is no longer produced in the EU (EURAR, 2009).

**1.2. Organophosphorus flame retardants**

92 Environmental Biotechnology - New Approaches and Prospective Applications

Organophosphorus flame retardants (PFRs) are based primarily on phosphate esters, phosphonate esters and phosphite esters. The total consumption of FRs in Europe was an estimated 465,000 tons in 2006, of which 20% comprised PFRs (KLIF, 2010). Of the PFRs consumed, 55% were chlorinated. Halogenated PFRs are the preferred form of FRs because halogen inhibits flame formation in organic materials, and non-halogenated

Tris(1,3-dichloro-2-propyl) phosphate (TDCPP) and tris(2-chloroethyl) phosphate (TCEP)

**Figure 1.** Chemical structure of tris(1,3-dichloro-2-propyl) phosphate (TDCPP) and tris(2-chloroethyl) phosphate

TDCPP is a viscous colorless to light yellow liquid and is produced by the epoxide opening of epichlorohydrin in the presence of phosphorus oxychlorine (ATSDR, 2009). TDCPP is used primarily in flexible polyurethane foams but also in rigid polyurethane foams, resins, plastics, textile coatings and rubbers (California EPA, 2011). TDCPP was a common ingredi‐ ent of sleepwear for children in the 1970s, but was voluntarily withdrawn by manufactures in 1977 because of its proven mutagenicity (California EPA, 2011). However, the PFR can still be found in many baby products (Stapleton et al., 2011). Currently, TDCPP is used mostly in flexible polyurethane foams for upholstered furniture and automotive products. TDCPP consumption has increased following the ban on common FR polybrominated di‐ phenyl ethers (PBDEs). Consequently, total TDCCP production has increased, being an esti‐ mated 4,500-22,700 tons in the United States in 2006 and <10,000 tons in Europe in 2000 (van

TCEP is colorless to pale yellow liquid and is highly soluble in water (Fig. 1 and Table 1). The compound is chemically synthesized via condensation of phosphorus oxychlor‐ ide and chloroalkyl alcohol at low temperatures and pressures to avoid formation of al‐ kyl chlorides (ATSDR, 2009). Previously, the main purpose of TCEP was to reduce the

Cl

Cl

O

Cl

P

O

O

O

Tris(2-chloroethyl) phosphate (TCEP)

PFRs are typically used as flame-retardant plasticizers (KLIF, 2010).

are typical examples of additive chlorinated PFR (Fig. 1 and Table 1).

Cl

Cl

O

Cl

P

O

O

der Veen & de Boer, 2012).

Cl

Cl

(TCEP)

O

Tris(1,3-dichloro-2-propyl) phosphate (TDCPP)

**1.3. Tris(1,3-dichloro-2-propyl) phosphate and tris(2-chloroethyl) phosphate**

Cl


**Table 1.** General aspect of Tris(1,3-dichloro-2-propyl) phosphate (TDCPP) and tris(2-chloroethyl) phosphate (TCEP)

### **1.4. Occurrence and behavior of TDCPP and TCEP in the environment**

TCEP and TDCPP have been detected in various environments worldwide, including in‐ door and outdoor air, surface and ground waters, and even drinking water (Tables 2 and 3). It is unlikely that these compounds are produced naturally. Their environmental presence is thus considered to be the result of human activity. Because these PFRs are physicochemical‐ ly and microbiologically stable in the environment and are also reportedly toxic, they are a serious threat to human and ecosystem health.

**Environment Concentration Location Country Reference**

hospital, hotel, prison, library, office shops

Outdoor air: <0.04-0.072 ng m3-1 nearby main road Sweden Marklund et al., 2003

5-8 pg m3-1 sea Northern pacific

49-780 pg m3-13 sea East Indian

680-6,180 ng L-1 raw water of waste disposal site

treatment

<250 ng L-1 water after drinking water treatment

Drinking water: 1.2-2.4 ng L-1 water after drinking water

<0.04-0.14 ng m3-1 remote area from main roadSweden Marklund et al., 2003 n.d.-5 pg m3-1 sea Arctic ocean Moller et al., 2012 16-52 pg m3-1 sea Japan Moller et al., 2012

n.d.-220 pg m3-13 sea Indian ocean Moller et al., 2012 80 pg m3-1 sea Southern ocean Moller et al., 2012

~50 ng L-1 river Germany Andresen et al., 2004 2-24 ng L-1 rain Germany Regnery & Püttmann, 2009 5-40 ng L-1 snow Germany Regnery & Püttmann, 2009 <19 ng L-1 river Austria Martinez-Carballo et al., 2007 <3.0-19 ng L-1 river Austria Martinez-Carballo et al., 2007

<1,335 ng L-1 lake Italy Bacaloni et al., 2008 108-448 ng L-1 rain Italy Bacaloni et al., 2008

Surface water: 10-18 ng L-1 river Germany Andresen & Bester, 2006

ocean

archipelago, Philippine sea

<0.6 ng m3-1, <8.7 ng m3-1

Indoor dust: 0.2-67 μg g-1 home, cinema, university,

1.3 ng m3-1 newly constructed house Japan Saito et al., 2007

<0.08-6.64 μg g-1 house Belgium van den Eede et al., 2011 <0.08-56.2 μg g-1 store Belgium van den Eede et al., 2011 2.2-27 μg g-1 home Belgium Bergh et al., 2011 3.9-150 μg g-1 day care Belgium Bergh et al., 2011 3.3-91 μg g-1 work place Belgium Bergh et al., 2011 <1.1 μg g-1 house Spain Garcia et al., 2007 <0.09-56.1 μg g-1 house United States Stapleton et al., 2009 0.069-18 μg g-1 hotel Japan Takigami et al., 2009 <127 μg kg-1 house Japan Kanazawa et al., 2010

house and office Japan Saito et al., 2007

Sweden Marklund et al., 2003

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Microbial Degradation of Persistent Organophosphorus Flame Retardants

Moller et al., 2012

Moller et al., 2012

Japan Kawagoshi et al., 1999

Germany Andresen & Bester, 2006

United States Stackelberg et al., 2004

### *1.4.1. TDCPP*

Detected air concentrations of TDCPP have attained up to150 ng m3-1 in Sweden houses, and in Belgium office and stores, they have reached 73 ng m3-1 (Table 2). In outdoor air, TDCPP levels near a main road in Sweden ranged from <0.04-0.072 m3-1, and significant amounts have been detected globally in air borne particles over the Pacific, Indian, Arctic and South‐ ern Oceans. TDCPP has been also found in indoor dust at relatively higher concentrations. Levels of TDCPP have tended to be higher in public buildings than in domestic buildings.

With respect to water environments, TDCPP concentrations have been detected at up to ~50 ng L-1 in German rivers and at 1,335 ng L-1 in Italian lakes. In these countries, it also occurs in rain and/or snow, as a result of volatilization from host materials. A much higher TDCPP concentration was detected in raw water at a disposal site in Japan, suggesting that the com‐ pound leaches and migrates to water sources. In the United States and Germany, TDCPP has even been detected in drinking water processed in treatment plants (DWTs). Relatively higher concentrations of TDCPP occur in landfill site sediments. Much higher concentra‐ tions still have been found in sediments near a car demolition site in Norway.

TDCPP has been also detected in the effluents of sewage treatment plants (STPs) and waste water treatment plants (WWTPs) in European countries and Japan, revealing that effluents are a source of aquatic TDCPP contamination. Comparable levels have been observed in the influents, indicating that the compound persists in the treatment plants. Degradation of TDCPP in the environment has been reported as low. Together, these observations suggest that TDCPP is likely to accumulate in the environment.



**1.4. Occurrence and behavior of TDCPP and TCEP in the environment**

94 Environmental Biotechnology - New Approaches and Prospective Applications

serious threat to human and ecosystem health.

*1.4.1. TDCPP*

TCEP and TDCPP have been detected in various environments worldwide, including in‐ door and outdoor air, surface and ground waters, and even drinking water (Tables 2 and 3). It is unlikely that these compounds are produced naturally. Their environmental presence is thus considered to be the result of human activity. Because these PFRs are physicochemical‐ ly and microbiologically stable in the environment and are also reportedly toxic, they are a

Detected air concentrations of TDCPP have attained up to150 ng m3-1 in Sweden houses, and in Belgium office and stores, they have reached 73 ng m3-1 (Table 2). In outdoor air, TDCPP levels near a main road in Sweden ranged from <0.04-0.072 m3-1, and significant amounts have been detected globally in air borne particles over the Pacific, Indian, Arctic and South‐ ern Oceans. TDCPP has been also found in indoor dust at relatively higher concentrations. Levels of TDCPP have tended to be higher in public buildings than in domestic buildings.

With respect to water environments, TDCPP concentrations have been detected at up to ~50 ng L-1 in German rivers and at 1,335 ng L-1 in Italian lakes. In these countries, it also occurs in rain and/or snow, as a result of volatilization from host materials. A much higher TDCPP concentration was detected in raw water at a disposal site in Japan, suggesting that the com‐ pound leaches and migrates to water sources. In the United States and Germany, TDCPP has even been detected in drinking water processed in treatment plants (DWTs). Relatively higher concentrations of TDCPP occur in landfill site sediments. Much higher concentra‐

TDCPP has been also detected in the effluents of sewage treatment plants (STPs) and waste water treatment plants (WWTPs) in European countries and Japan, revealing that effluents are a source of aquatic TDCPP contamination. Comparable levels have been observed in the influents, indicating that the compound persists in the treatment plants. Degradation of TDCPP in the environment has been reported as low. Together, these observations suggest

Sweden Marklund et al., 2005a

Sweden Staaf & Ostman, 2005

tions still have been found in sediments near a car demolition site in Norway.

**Environment Concentration Location Country Reference** Indoor air: <0.04-18 ng m3-1 office and store Norway SFT, 2008

> hospital, hotel, prison, library, office shops

> electronic dismantling facility recycling plant

<73 ng m3-1 work place Belgium Bergh et al., 2011 <61.4 ng m3-1 house Japan Kanazawa et al., 2010

that TDCPP is likely to accumulate in the environment.

<0.2-150 ng m3-1 home, cinema, university,

<0.3-7 ng m3-1 lecture and computer hall,


road traffic as an important source of TCEP emission. TCEP has also been detected globally in air borne particles over the Pacific, Indian, Arctic and Southern Ocean. In Belgium, indoor dust can contain up to 260 μg g-1 TCEP. TCEP concentrations in dusts of public spaces tend

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97

TCEP ranges from <3.0-1,236 ng L-1 in German rivers, lakes and reservoirs. In this country and in Italy, it has also been detected in rain and/or snow, indicating that, like TDCPP, TCEP volatilizes from its host materials. Groundwater TCEP levels up to 754 ng L-1 have been reported in Germany, suggesting that TCEP primarily mobilizes into water rather than attaching to soil. TCEP also occurs in drinking water or finished water from DWTs; record‐ ed concentrations are as high as 99, 25 and 1.7 ng L-1 in the United States, Korea and Germa‐ ny, respectively. Much higher concentrations have been observed in raw water of waste disposal sites in Japan. Relatively higher concentrations of TCEP have been detected in land‐ fill site sediments in Japan and Norway (up to 7,400 and 380 μg kg-1, respectively). Especial‐

TCEP has been also detected in STP or WWTP effluents in many countries. Comparable lev‐ els of TCEP are observed in the influents. These observations demonstrate that, like TDCPP,

Also similarly to TDCPP, TCEP has been detected in biological samples, including fishes, crabs, mussels and birds. In Norway, fishes and mussels respectively contain up to 26 and 23 ng g-1 TCEP. In birds and their eggs, TCEP levels can reach up to 6.1 ng g-1. In fishes resid‐ ing near emission sources in Sweden, they reach up to 69 and 160 ng g-1 respectively. Fur‐ thermore, like TDCPP, TCEP has been detected in the breast milk of Swedish women.

3, 9 ng m3-1 lecture room and kindergarten Sweden Tollback et al., 2006

3-15 ng m3-1 lecture room and office room Sweden Bjorklund et al., 2004 <297 ng m3-1 house Japan Kanazawa et al., 2010

1.2 ng m3-1 newly constructed house Japan Saito et al., 2007 <28 ng m3-1 home Belgium Bergh et al., 2011 7.8-230 ng m3-1 day care center Belgium Bergh et al., 2011

house and office Japan Saito et al., 2007

Sweden Marklund et al., 2005a

Sweden Staaf & Ostman, 2005

Sweden Hartmann et al., 2004

ly high concentrations were found in the sediment nearby a car demolition site.

**Environment Concentration Location Country Reference** Indoor air: <0.2-23 ng m3-1 office and store Norway SFT, 2008

hospital, hotel, prison, library,

electronic dismantling facility

office and electronics store

0.4-730 ng m3-1 home, cinema, university,

<0.3-10 ng m3-1 Lecture and computer hall,

<22 ng m3-1 car, theater, furniture store,

office shops

recycling plant

to exceed those in domestic dusts.

TCEP persists in the treatment plants.

<136 ng m3-1, <42.1 ng m3-1

**Table 2.** Occurrence and behavior of TDCPP

TDCPP has also been detected in biological samples, including fishes, mussels and birds. In Norway, fishes and mussels were observed to contain up to 8.1 and 30 ng g-1 of TDCPP, respectively. In bird blood/plasma and eggs respectively, TDCPP levels range from <0.11-0.16 and from <0.72-1.9 ng g-1. In Sweden, freshwater fishes close to emission sources contained 49-140 ng g-1 TDCPP. Worryingly, TDCPP has also been detected in the breast milk of Swedish women.

### *1.4.2. TCEP*

In Sweden, the highest detected air concentration of TCEP was 730 ng m3-1 inside an office furnished with linoleum floor and a new photocopier (Table 3). In outdoor air, it can reach 6.2 ng m3-1 beside a main road, but remote areas harbor less than 0.2 ng m3-1, implicating road traffic as an important source of TCEP emission. TCEP has also been detected globally in air borne particles over the Pacific, Indian, Arctic and Southern Ocean. In Belgium, indoor dust can contain up to 260 μg g-1 TCEP. TCEP concentrations in dusts of public spaces tend to exceed those in domestic dusts.

**Environment Concentration Location Country Reference**

Sludge: 110-330 μg kg-1 Norway SFT, 2008

Influent: 630-820 ng L-1 WWTP Norway SFT, 2008

Effluent: 86-740 ng L-1 WWTP Norway SFT, 2008

Biota: <6.0 ng g-1 fish liver Norway SFT, 2009b

49-140 ng g-1 freshwater fishes close to sources

1,500-4,100 μg kg-1 landfill site Norway SFT, 2008 <250-8,800 μg kg-1 car demolition site Norway SFT, 2008

<709 μg kg-1 waste disposal site Japan Kawagoshi et al., 1999

3.0-260 μg kg-1 Sweden Stackelberg et al., 2004

240-450 ng L-1 STP Sweden Marklund et al., 2005b 330-1,600 ng L-1 STP Japan Ishikawa et al., 1985

130-340 ng L-1 STP Sweden Marklund et al., 2005b 20-120 ng L-1 STP Germany Andresen et al., 2004 19-1,400 ng L-1 WWTP Austria Martinez-Carballo et al., 2007

280-1,400 ng L-1 STP Japan Ishikawa et al., 1985

cod liver and mussel Norway SFT, 2008

16.-5.3 ng g-1 human milk Sweden Sundkvist et al., 2010

TDCPP has also been detected in biological samples, including fishes, mussels and birds. In Norway, fishes and mussels were observed to contain up to 8.1 and 30 ng g-1 of TDCPP, respectively. In bird blood/plasma and eggs respectively, TDCPP levels range from <0.11-0.16 and from <0.72-1.9 ng g-1. In Sweden, freshwater fishes close to emission sources contained 49-140 ng g-1 TDCPP. Worryingly, TDCPP has also been detected in

In Sweden, the highest detected air concentration of TCEP was 730 ng m3-1 inside an office furnished with linoleum floor and a new photocopier (Table 3). In outdoor air, it can reach 6.2 ng m3-1 beside a main road, but remote areas harbor less than 0.2 ng m3-1, implicating

Norway Sundkvist et al., 2010

<0.3-6.7 ng g-1 fish muscle Norway SFT, 2009b <0.72-1.9 ng g-1 bird egg Norway KLIF, 2010 <0.11-0.16 ng g-1 bird blood and plasma Norway KLIF, 2010 <0.6-8.1 ng g-1 whole fish Norway SFT, 2009b <1.5 ng g-1 seabird liver Norway SFT, 2009b <0.3-1.2 ng g-1 whole fish liver Norway SFT, 2009b

Norway KLIF, 2010

WWFP

Sediment: <0.15-54 μg kg-1 lake and fjord at vicinity of

96 Environmental Biotechnology - New Approaches and Prospective Applications

<5.0 ng g-1, <10-<30 ng g-1

**Table 2.** Occurrence and behavior of TDCPP

the breast milk of Swedish women.

*1.4.2. TCEP*

TCEP ranges from <3.0-1,236 ng L-1 in German rivers, lakes and reservoirs. In this country and in Italy, it has also been detected in rain and/or snow, indicating that, like TDCPP, TCEP volatilizes from its host materials. Groundwater TCEP levels up to 754 ng L-1 have been reported in Germany, suggesting that TCEP primarily mobilizes into water rather than attaching to soil. TCEP also occurs in drinking water or finished water from DWTs; record‐ ed concentrations are as high as 99, 25 and 1.7 ng L-1 in the United States, Korea and Germa‐ ny, respectively. Much higher concentrations have been observed in raw water of waste disposal sites in Japan. Relatively higher concentrations of TCEP have been detected in land‐ fill site sediments in Japan and Norway (up to 7,400 and 380 μg kg-1, respectively). Especial‐ ly high concentrations were found in the sediment nearby a car demolition site.

TCEP has been also detected in STP or WWTP effluents in many countries. Comparable lev‐ els of TCEP are observed in the influents. These observations demonstrate that, like TDCPP, TCEP persists in the treatment plants.

Also similarly to TDCPP, TCEP has been detected in biological samples, including fishes, crabs, mussels and birds. In Norway, fishes and mussels respectively contain up to 26 and 23 ng g-1 TCEP. In birds and their eggs, TCEP levels can reach up to 6.1 ng g-1. In fishes resid‐ ing near emission sources in Sweden, they reach up to 69 and 160 ng g-1 respectively. Fur‐ thermore, like TDCPP, TCEP has been detected in the breast milk of Swedish women.



**Environment Concentration Location Country Reference**

treatment

4-99 ng L-1 water after drinking water treatment

<99 ng L-1 water after drinking water treatment

14, 25 ng L-1 water after drinking water treatment

WWFP

Drinking water:0.74-1.7 ng L-1 water after drinking water

Sediment: <0.16-8.5 μg kg-1 lake and fjord at vicinity of

Biota: 0.5-5.0 ng g-1

13-26 ng g-1

<5 ng g-1, <10-23 ng g-1

<754 ng L-1 Germany Fries & Püttmann , 2001

27-380 μg kg-1 landfill site Norway SFT, 2008 2,300-5,500 μg kg-1 car demolition site Norway SFT, 2008

Sludge: <9-<19 μg kg-1 Norway SFT, 2008

Influent: 2,000-2,500 ng L-1 STP Norway SFT, 2008

Effluent: 1600-2,200 ng L-1 STP Norway SFT, 2008

<160 μg kg-1 river Austria Martinez-Carballo et al., 2007 <7,400 μg kg-1 waste disposal site Japan Kawagoshi et al., 1999

6.6-110 μg kg-1 Sweden Marklund et al., 2005b

90-1,000 ng L-1 STP Sweden Marklund et al., 2005b 290, 180 ng L-1 STP Germany Meyer & Bester, 2004 983-1,123 ng L-1 municipal STWs Germany Fries & Püttmann , 2003 <0.025-0.3 ng L-1 STP Spain Rodriguez et al., 2006 540-1,200 ng L-1 STP Japan Ishikawa et al., 1985

350-890 ng L-1 STP Sweden Marklund et al., 2005b 350, 370 ng L-1 STP Germany Meyer & Bester, 2004 214-557 ng L-1 municipal STWs Germany Fries & Püttmann , 2003 <0.025-0.7 ng L-1 STP Spain Rodriguez et al., 2006 500-1,200 ng L-1 STP Japan Ishikawa et al., 1985

1.8-3.2 ng kg-1 whole fish Norway SFT, 2009b

<0.6-4.7 ng g-1 sea bird liver Norway SFT, 2009b <0.17-19 ng g-1 beach crab Norway KLIF, 2010 <0.06-0.11 ng g-1 blue mussel Norway KLIF, 2010 <1.7-8.6 ng g-1 burbot liver Norway KLIF, 2010 <0.08-0.21 ng g-1 trout Norway KLIF, 2010 <0.33-6.1 ng g-1 bird egg Norway KLIF, 2010 <0.17-6.0 ng g-1 bird blood and plasma Norway KLIF, 2010

fish muscle and liver Norway SFT, 2009b

cod liver and mussel Norway SFT, 2008

Germany Andresen & Bester, 2006

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99

United States Stackelberg et al., 2007

United States Stackelberg et al., 2004

Korea Kim et al., 2007

Norway KLIF, 2010

Microbial Degradation of Persistent Organophosphorus Flame Retardants


**Environment Concentration Location Country Reference**

office shops

hospital, hotel, prison, library,

Outdoor air: 0.51-6.2 ng m3-1 nearby main road Sweden Marklund et al., 2003

159-282 pg m3-1 sea Northern pacific

19-156 pg m3-1 sea East Indian

4,230-87,400 ng L-1 raw water of waste disposal site

<0.2 ng m3-1 remote area from main road Sweden Marklund et al., 2003 126-585 pg m3-1 ocean Arctic ocean Moller et al., 2012 273-1,961 pg m3-1 sea Japan Moller et al., 2012

46-570 pg m3-1 sea Indian ocean Moller et al., 2012 74 pg m3-1 sea Southern ocean Moller et al., 2012 Surface water: <3-184 ng L-1 lake and reservoir Germany Regnery & Püttmann , 2010

> 12-130 ng L-1 river Germany Andresen & Bester, 2006 13-130 ng L-1 river Germany Andresen et al., 2004 <1,236 ng L-1 river Germany Fries & Püttmann , 2003 11-196 ng L-1 rain Germany Regnery & Püttmann, 2009 121 ng L-1 rain Germany Fries & Püttmann , 2003 19-60 ng L-1 snow Germany Regnery & Püttmann, 2009 13-130 ng L-1 river Austria Martinez-Carballo et al., 2007

<33 ng L-1 lakes Italy Bacaloni et al., 2008 7 ng L-1 river Italy Bacaloni et al., 2007 19-161 ng L-1 rain Italy Bacaloni et al., 2008

14-347 ng L-1 river and sea water Japan Ishikawa et al., 1985 14-81 ng L-1 lake and river Korea Kim et al., 2007 Ground water: 3-9 ng L-1 Germany European Commission DG

<312 ng L-1 Germany Fries & Püttmann , 2003

ocean

archipelago, Philippine sea

Indoor dust: 0.19-94 μg g-1 home, cinema, university,

98 Environmental Biotechnology - New Approaches and Prospective Applications

<140 ng m3-1 work place Belgium Bergh et al., 2011

<0.08-2.65 μg g-1 house Belgium van den Eede et al., 2011 <33 μg g-1 house Belgium Bergh et al., 2011 <0.08-5.46 μg g-1 store Belgium van den Eede et al., 2011 2.5-150 μg g-1 day care center Belgium Bergh et al., 2011 1.3-260 μg g-1 work place Belgium Bergh et al., 2011 0.25-1.56 μg g-1 house Spain Garcia et al., 2007 <308 μg g-1 house Japan Kanazawa et al., 2010 0.082-2.3 μg g-1 hotel Japan Takigami et al., 2009

Sweden Marklund et al., 2003

Moller et al., 2012

Moller et al., 2012

Japan Kawagoshi et al., 1999

ENV, 2011


alga *Pseudokirchneriella*, ErC10 (10% growth-rate inhibition) was recorded as 2.3 mg L-1*.* Thus, TDCPP is classified as N; R51/53, denoting "Toxic to aquatic organisms, may cause longterm adverse effects in the aquatic environment". In addition, an LC50 of 23 mg kg-1 has been

LD50=3,160 mg kg-1 male rat EURAR, 2008

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101

**Toxicity Organism Reference** Acute toxicity LD50=6,800 mg kg-1 male rabbit US EPA, 2005

> LD50=2,670 mg kg-1 male mice LD50=2,250 mg kg-1 female mice LD50=2,236 mg kg-1 male rat LD50=2,489 mg kg-1 female rat

convoluted tubule epithelium

Cytotoxicity hepatocytes and neuronal cells Crump et al., 2012 Neurotoxicity *in vitro* PC12 cells Dishaw et al., 2011 Carcinogenicity rat California EPA, 2011 Genotoxicity *in vivo Salmonella typhimurium* California EPA, 2011 *in vitro* mouse, Chinese hamster and rat cells

Toxic to aquatic organisms fishes, invertebrates and algae EURAR, 2008 LC50=1.1 mg L-1 rainbow trout (96 h) EC50=3.8 mg L-1 *Daphnia magana* (48 h)

LC50=23 mg kg-1 earthworm *Eisenia*

NOEC=17 mg kg-1 plant *Mustard* Alter hormone levels human and zebra fish cells Liu et al., 2012

ErC10=2.3 mg L-1 algae

**Table 4.** Toxicological information of TDCPP

*1.5.2. TCEP*

LOEC=1.0 mg L-1 *Daphnia* for reproduction (21 days) NOEC=0.5 mg L-1 *Daphnia* for reproduction (21 days)

NOEC=2.9 mg kg-1 earthworm *Eisenia* for reproduction

Decreased sperm quality human Meeker & Stapleton, 2010

Rats given oral doses of TCEP absorb over 90% of the compound within 24 h, with marked accumulations in liver, kidney, fat and the gastrointestinal tract (EURAR, 2009). In animals,

reported for a terrestrial organism, the earthworm *Eisenia*.

Chronic toxicity LOAEL=5 mg kg-1 day-1 rat for hyperplasia and

**Table 3.** Occurrence and behavior of TCEP

### **1.5. Toxicological information of TDCPP and TCEP**

Since the toxic effects of TCEP and TDCPP have been regarded as marginal compared to those of PBDEs, they have been extensively used. However, their non-negligible toxic‐ ities have been revealed in a number of studies (Tables 4 and 5). Together with their persistence in the environment, the environmental contamination of both compounds has become of serious concern.

### *1.5.1. TDCPP*

Rats given oral doses of TDCPP absorb more than 90% of the compound within 24 h, with the highest concentrations being observed in kidney, liver and lung (EURAR, 2008). The acute toxicity of oral TDCPP has been reported as low, with LD50 values ranging from 2,250 mg kg-1 for female mice to 6,800 mg kg-1 for male rabbits (Table 4). In a 2-year chronic toxici‐ ty study in rats, the lowest observable adverse effect level (LOAEL) was 5 mg kg-1 day-1. In that study, statistically significant relationships between TDCPP dose and tumor incidences were observed in both male and female rats. Consequently, TDCPP is today classified as Carc. Cat. 3; R40 and Cat. 2; H351, denoting "limited evidence of a carcinogenic effect" and "suspected of causing cancer", respectively.

A number of TDCPP genotoxicity studies have been conducted in whole mammals that have resulted in negative conclusions regarding genotoxicity (Albemarle Corp. & ICL North America Inc., 2011). However, *in vitro* studies using bacteria and mammalian cells have sug‐ gested that TDCPP exerts genotoxic effects, and an *in vivo* study showed its covalent bind‐ ing to DNA (US EPA, 2005; Morales & Matthews, 1980).

Similarly, neurotoxicity studies of TDCPP involving hens and rats reveal no clear evidence that TDCPP is neurotoxic. However, a study based on undifferentiated and differentiating PC12 cells showed its potential neurotoxicity (Dishaw et al., 2011).

Whether, and to what extent, TDCPP is toxic to humans remains unknown. However, TDCPP has been shown to alter sex hormone balance in human cell lines, via alteration of steroidogenesis or estrogen metabolism (Liu et al., 2012). In addition, TDCPP concentrations in house dusts have been linked to altered hormone levels and decreased semen quality in men (Meeker & Stapleton, 2010).

TDCPP is regarded as toxic to aquatic organisms (EURAR, 2008). An acute toxicity study on fish trout yielded an LC50 value of 1.1 mg L-1. Acute and chronic toxicity studies conducted on the invertebrate *Daphnia* produced an EC50 value of 3.8 mg L-1. In a chronic study on the alga *Pseudokirchneriella*, ErC10 (10% growth-rate inhibition) was recorded as 2.3 mg L-1*.* Thus, TDCPP is classified as N; R51/53, denoting "Toxic to aquatic organisms, may cause longterm adverse effects in the aquatic environment". In addition, an LC50 of 23 mg kg-1 has been reported for a terrestrial organism, the earthworm *Eisenia*.


**Table 4.** Toxicological information of TDCPP

### *1.5.2. TCEP*

**Environment Concentration Location Country Reference**

<160 ng g-1 freshwater fishes close to sources

100 Environmental Biotechnology - New Approaches and Prospective Applications

**1.5. Toxicological information of TDCPP and TCEP**

**Table 3.** Occurrence and behavior of TCEP

has become of serious concern.

"suspected of causing cancer", respectively.

men (Meeker & Stapleton, 2010).

ing to DNA (US EPA, 2005; Morales & Matthews, 1980).

PC12 cells showed its potential neurotoxicity (Dishaw et al., 2011).

*1.5.1. TDCPP*

1.5-69 ng g-1 marine fishes Sweden Sundkvist et al., 2010

201-8.2 ng g-1 human milk Sweden Sundkvist et al., 2010

Since the toxic effects of TCEP and TDCPP have been regarded as marginal compared to those of PBDEs, they have been extensively used. However, their non-negligible toxic‐ ities have been revealed in a number of studies (Tables 4 and 5). Together with their persistence in the environment, the environmental contamination of both compounds

Rats given oral doses of TDCPP absorb more than 90% of the compound within 24 h, with the highest concentrations being observed in kidney, liver and lung (EURAR, 2008). The acute toxicity of oral TDCPP has been reported as low, with LD50 values ranging from 2,250 mg kg-1 for female mice to 6,800 mg kg-1 for male rabbits (Table 4). In a 2-year chronic toxici‐ ty study in rats, the lowest observable adverse effect level (LOAEL) was 5 mg kg-1 day-1. In that study, statistically significant relationships between TDCPP dose and tumor incidences were observed in both male and female rats. Consequently, TDCPP is today classified as Carc. Cat. 3; R40 and Cat. 2; H351, denoting "limited evidence of a carcinogenic effect" and

A number of TDCPP genotoxicity studies have been conducted in whole mammals that have resulted in negative conclusions regarding genotoxicity (Albemarle Corp. & ICL North America Inc., 2011). However, *in vitro* studies using bacteria and mammalian cells have sug‐ gested that TDCPP exerts genotoxic effects, and an *in vivo* study showed its covalent bind‐

Similarly, neurotoxicity studies of TDCPP involving hens and rats reveal no clear evidence that TDCPP is neurotoxic. However, a study based on undifferentiated and differentiating

Whether, and to what extent, TDCPP is toxic to humans remains unknown. However, TDCPP has been shown to alter sex hormone balance in human cell lines, via alteration of steroidogenesis or estrogen metabolism (Liu et al., 2012). In addition, TDCPP concentrations in house dusts have been linked to altered hormone levels and decreased semen quality in

TDCPP is regarded as toxic to aquatic organisms (EURAR, 2008). An acute toxicity study on fish trout yielded an LC50 value of 1.1 mg L-1. Acute and chronic toxicity studies conducted on the invertebrate *Daphnia* produced an EC50 value of 3.8 mg L-1. In a chronic study on the

Sweden Sundkvist et al., 2010

Rats given oral doses of TCEP absorb over 90% of the compound within 24 h, with marked accumulations in liver, kidney, fat and the gastrointestinal tract (EURAR, 2009). In animals, TCEP appears to be mainly toxic to brain, kidney and liver. Toxicity studies have implicated TCEP as moderately toxic; in rats, oral administration yields an LD50 of 430-1,230 mg kg-1 and skin contact reveals a low acute dermal toxicity (LD50 >2,150 mg kg-1) (Table 5). A 2-year chronic toxicity study of TCEP yielded LOAELs of 44 mg kg-1 day-1 in rats and 175 mg kg-1 day-1 in mice. The same study indicated that TCEP is potentially neurotoxic, with no ob‐ served adverse effect levels (NOAELs) in rats and mice being 88 mg kg-1 day-1 and 175 mg kg-1 day-1, respectively.

TCEP is toxic to aquatic organisms, being classified as N; R51/53 (EURAR, 2009). Short term exposure to TCEP is mildly-moderately adverse to the aquatic invertebrate organisms *Daph‐ nia* and *Planaria,* and TCEP presents low acute toxicity to killifish, trout and goldfishes.

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The toxic effects of TCEP in humans are largely unknown. However, neurotoxic signs have been reported in a 5-year old child who slept in a room with wood paneling containing 3% TCEP (Ingerowski & Ingerowski, 1997). In addition, an epidemiological study of children in school environments found a potential association between the TCEP content in air-bone dusts and impaired cognitive ability (UBA, 2008). TCEP has been further reported to alter

The persistence of chlorinated FRs TCEP and TDCPP in current waste water and drinking water treatment processes has accelerated the investigation of alternative water treatment

Echigo *et al.* showed that TDCPP in distilled water and an effluent from a solid waste land‐ fill site is effectively degraded by O3/vacuum UV or O3/H2O2 process, although degradation products were not determined in this study (Echigo et al., 1996). Westerhoff *et al.* reported that >20% of approximately 30 ng L-1 of TCEP in surface water samples can be removed with powdered activated carbon, but that other adsorptive processes, metal salt coagulation and lime softening, and oxidative processes (chlorination and ozonation) are ineffective (West‐ erhoff et al., 2005). Lee *et al.* showed that > 90% removal efficiency of 100 μg L-1 of TCEP in river and sea waters is possible using tight nanofiltration membranes with a low molecular weight cutoff of approximately 200 (Lee et al., 2008). Watts *et al.* demonstrated that the high‐ er removing efficacy (> 95%) of 5 mg L-1 of TCEP in a water is achieved by a UV/H2O2 ad‐ vanced oxidation process with the highest UV fluence at 6,000 mJ cm-2 (Watts & Linden, 2008). In this study, the generation of stoichiometric amount of chloride ion was observed. In addition, Benotti *et al.* reported that UV/TiO2 supplemented with H2O2 can decrease the concentration of TCEP in a river water, although the degradation was not so effective and

**2. Microbial degradation and detoxification of TDCPP and TCEP**

FRs have been widely distributed commercially and are necessary to prevent or reduce mor‐ tality from accidental fires. However, the leaching of additive FRs has led to global contami‐ nation of the environment. The chlorinated PFRs TCEP and TDCPP persist in the environment and exhibit varying toxic effects, raising concerns about their effects on human and ecological health. Although several physicochemical methods for removing TCEP and TDCPP have been reported (as described above), biotechnological techniques offer an attrac‐ tive alternative, being potentially cost-effective, eco-friendly and enabling *in situ* remedia‐ tion of contaminants. However, prior to recent isolation of TCEP- and TDCPP-degrading bacteria by our group, no biological degrading agent for such compounds was known.

the sex hormone balance in human cells, as well as in fish cells.

**1.6. Removal technique for TDCPP and TCEP**

techniques that will dispel these compounds.

not completed (Benotti et al., 2009).


**Table 5.** Toxicological information of TCEP

In the 2-year study, increased incidences of adenomas and carcinomas were linked to TCEP exposure, revealing TCEP as a potential carcinogen (EURAR, 2009). TCEP is thus classified as Carc. Cat. 3; R40. Because TCEP additionally exhibits reproductive toxicity in rats and mice, it is also classified as Repr. Cat. 2; R60, denoting "may impair fertility". TCEP at envi‐ ronmental concentrations has been reported to affect the expression of cell cycle regulatory genes in primary cultured rabbit renal proximal tubule cells (Ren et al., 2008).

TCEP is toxic to aquatic organisms, being classified as N; R51/53 (EURAR, 2009). Short term exposure to TCEP is mildly-moderately adverse to the aquatic invertebrate organisms *Daph‐ nia* and *Planaria,* and TCEP presents low acute toxicity to killifish, trout and goldfishes.

The toxic effects of TCEP in humans are largely unknown. However, neurotoxic signs have been reported in a 5-year old child who slept in a room with wood paneling containing 3% TCEP (Ingerowski & Ingerowski, 1997). In addition, an epidemiological study of children in school environments found a potential association between the TCEP content in air-bone dusts and impaired cognitive ability (UBA, 2008). TCEP has been further reported to alter the sex hormone balance in human cells, as well as in fish cells.

### **1.6. Removal technique for TDCPP and TCEP**

TCEP appears to be mainly toxic to brain, kidney and liver. Toxicity studies have implicated TCEP as moderately toxic; in rats, oral administration yields an LD50 of 430-1,230 mg kg-1 and skin contact reveals a low acute dermal toxicity (LD50 >2,150 mg kg-1) (Table 5). A 2-year chronic toxicity study of TCEP yielded LOAELs of 44 mg kg-1 day-1 in rats and 175 mg kg-1 day-1 in mice. The same study indicated that TCEP is potentially neurotoxic, with no ob‐ served adverse effect levels (NOAELs) in rats and mice being 88 mg kg-1 day-1 and 175 mg

LD50>2,150 mg kg-1 rat for dermal EURAR, 2009

rabbit renal proximal tubule cells Ren et al., 2008

EURAR, 2009

LOAEL=175 mg kg-1 mouse for kidney morphology (2

NOAEL=88 mg kg-1 day-1 rats (16 weeks by gavage)

NOAEL=175 mg kg-1 day-1 mouse (16 weeks by

NOAEL=175 mg kg-1 day-1 mouse for fertility

In the 2-year study, increased incidences of adenomas and carcinomas were linked to TCEP exposure, revealing TCEP as a potential carcinogen (EURAR, 2009). TCEP is thus classified as Carc. Cat. 3; R40. Because TCEP additionally exhibits reproductive toxicity in rats and mice, it is also classified as Repr. Cat. 2; R60, denoting "may impair fertility". TCEP at envi‐ ronmental concentrations has been reported to affect the expression of cell cycle regulatory

**Toxicity Organism Reference** Acute toxicity LD50=430-1,230 mg kg-1 rat EURAR, 2009

Chronic toxicity LOAEL=44 mg kg-1 day-1 rat for kidney lesions (2 years) EURAR, 2009

Neurotoxicity rat and mouse EURAR, 2009

Reproductive toxicity rat and mouse EURAR, 2009

Carcinogenicity rat and mouse SCHER, 2012

Toxic to aquatic organisms killifish, trout and goldfish EURAR, 2009

Alter sex hormone balance human cells and Zebra fish Liu et al., 2012

genes in primary cultured rabbit renal proximal tubule cells (Ren et al., 2008).

years)

102 Environmental Biotechnology - New Approaches and Prospective Applications

gavage)

kg-1 day-1, respectively.

Alter cell cycle regulatory protein

**Table 5.** Toxicological information of TCEP

expression

The persistence of chlorinated FRs TCEP and TDCPP in current waste water and drinking water treatment processes has accelerated the investigation of alternative water treatment techniques that will dispel these compounds.

Echigo *et al.* showed that TDCPP in distilled water and an effluent from a solid waste land‐ fill site is effectively degraded by O3/vacuum UV or O3/H2O2 process, although degradation products were not determined in this study (Echigo et al., 1996). Westerhoff *et al.* reported that >20% of approximately 30 ng L-1 of TCEP in surface water samples can be removed with powdered activated carbon, but that other adsorptive processes, metal salt coagulation and lime softening, and oxidative processes (chlorination and ozonation) are ineffective (West‐ erhoff et al., 2005). Lee *et al.* showed that > 90% removal efficiency of 100 μg L-1 of TCEP in river and sea waters is possible using tight nanofiltration membranes with a low molecular weight cutoff of approximately 200 (Lee et al., 2008). Watts *et al.* demonstrated that the high‐ er removing efficacy (> 95%) of 5 mg L-1 of TCEP in a water is achieved by a UV/H2O2 ad‐ vanced oxidation process with the highest UV fluence at 6,000 mJ cm-2 (Watts & Linden, 2008). In this study, the generation of stoichiometric amount of chloride ion was observed. In addition, Benotti *et al.* reported that UV/TiO2 supplemented with H2O2 can decrease the concentration of TCEP in a river water, although the degradation was not so effective and not completed (Benotti et al., 2009).

### **2. Microbial degradation and detoxification of TDCPP and TCEP**

FRs have been widely distributed commercially and are necessary to prevent or reduce mor‐ tality from accidental fires. However, the leaching of additive FRs has led to global contami‐ nation of the environment. The chlorinated PFRs TCEP and TDCPP persist in the environment and exhibit varying toxic effects, raising concerns about their effects on human and ecological health. Although several physicochemical methods for removing TCEP and TDCPP have been reported (as described above), biotechnological techniques offer an attrac‐ tive alternative, being potentially cost-effective, eco-friendly and enabling *in situ* remedia‐ tion of contaminants. However, prior to recent isolation of TCEP- and TDCPP-degrading bacteria by our group, no biological degrading agent for such compounds was known.

### **2.1. Isolation and characterization of TDCPP- and TCEP-degrading bacteria**

### *2.1.1. Enrichment of TCEP and TDCPP-degrading bacteria*

### *2.1.1.1. Enrichment cultivation of TCEP and TDCPP-degrading bacteria*

To obtain microorganisms that can degrade TDCPP and TCEP, we used an enrichment cul‐ ture technique in which one of TDCPP or TCEP served as the sole phosphorus source (Taka‐ hashi et al., 2008). Forty six environmental samples (soils and sediments) in Japan were cultivated at 30°C in minimal medium containing approximately 20 μM of each compound. Significant degradation of TCEP and TDCPP was seen in ten and three of the samples, re‐ spectively. In the first cultivation round, each compound had disappeared within 2 to 5 days; successive sub-cultivations reduced the degradation time to within one day. The en‐ richment cultures displaying the highest degradation efficacy against TCEP and TDCPP were designated 67E and 45D, respectively. Culture 67E completely degraded 20 μM of TDCPP in 3 h and TCEP in 6 h (Fig. 2A and B), while culture 45D completely degraded the same concentration of TDCPP in 3 h and TCEP in 24 h. During the degradations, 2-CE was liberated from TCEP and 1,3-DCP from TDCPP, indicating that the degradation pathway in‐ volved hydrolysis of phosphoester bonds.

ated chloride ions from 2-CE and 1,3-DCP, respectively. After 120 h reaction, the propor‐ tion of chloride ion was approximately 100% and 68.5% of the total chlorine contained in the supplied 2-CE and 1,3-DCP, respectively. This shows that both cultures can deha‐ logenate their respective chloroalcohols and can therefore potentially detoxify chlorinated

**Time (h) Time (h)**

**Figure 3.** Effect of exogenous phosphate on the degradation of TCEP (A) and TDCPP (B) and the chloride ion forma‐ tion from TCEP (C) and TDCPP (D). The enrichment cultures, 67E (A and C) and 45D (B and D), were cultivated on 20 µM of TCEP or TDCPP as the sole phosphorus source, respectively, with various concentrations of inorganic phosphate (NaH2PO4): 0 mM (closed circles), 0.02 mM (closed triangle), 0.2 mM (closed squares) and 2 mM (closed diamonds). Control culture without cell inoculation is indicated by open circles. Each data point represents the mean of at least

Phosphate-sufficient conditions are well known to repress the expression of genes involved in phosphorus utilization. We thus examined the effect of exogenous inorganic phosphate

*2.1.1.3. Effect of exogenous phosphate on the degradation ability of enrichment cultures*

**Cl**

**-(μM)**

**Time (h) Time (h)**

**0**

**0 2 4 6**

**0 30 60 90 120**

**5**

**10**

**15**

**20**

**25**

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105

**TDCPP (μM)**

**B**

PFRs in the environment.

**TCEP (μM)**

**0**

two independent determinations.

**Cl**

**-(μM)**

**0 2 4 6**

**C D**

**0 30 60 90 120 150**

**5**

**10**

**15**

**20**

**25**

**A**

**Figure 2.** Degradation of TCEP (A) and TDCPP (B) by enrichment cultures. The enrichment cultures, 67E (circles) and 45D (triangles), were cultivated on 20 µM of TCEP or TDCPP as the sole phosphorus source.

### *2.1.1.2. 2-CE and 1,3-DCP degradation ability of enrichment cultures*

The metabolites 2-CE and 1,3-DCP are also persistent and toxic: 1,3-DCP is a known gen‐ otoxin and carcinogen (NTP & NIEHS, 2005), while 2-CE exhibits genotoxicity, fetotoxici‐ ty and cardiotoxicity (National Toxicology, 1985). We analyzed whether the cultures can degrade the metabolites by measuring chloride ion formation. Cultures 67E and 45D liber‐ ated chloride ions from 2-CE and 1,3-DCP, respectively. After 120 h reaction, the propor‐ tion of chloride ion was approximately 100% and 68.5% of the total chlorine contained in the supplied 2-CE and 1,3-DCP, respectively. This shows that both cultures can deha‐ logenate their respective chloroalcohols and can therefore potentially detoxify chlorinated PFRs in the environment.

**2.1. Isolation and characterization of TDCPP- and TCEP-degrading bacteria**

To obtain microorganisms that can degrade TDCPP and TCEP, we used an enrichment cul‐ ture technique in which one of TDCPP or TCEP served as the sole phosphorus source (Taka‐ hashi et al., 2008). Forty six environmental samples (soils and sediments) in Japan were cultivated at 30°C in minimal medium containing approximately 20 μM of each compound. Significant degradation of TCEP and TDCPP was seen in ten and three of the samples, re‐ spectively. In the first cultivation round, each compound had disappeared within 2 to 5 days; successive sub-cultivations reduced the degradation time to within one day. The en‐ richment cultures displaying the highest degradation efficacy against TCEP and TDCPP were designated 67E and 45D, respectively. Culture 67E completely degraded 20 μM of TDCPP in 3 h and TCEP in 6 h (Fig. 2A and B), while culture 45D completely degraded the same concentration of TDCPP in 3 h and TCEP in 24 h. During the degradations, 2-CE was liberated from TCEP and 1,3-DCP from TDCPP, indicating that the degradation pathway in‐

**0**

**Time (h) Time (h)**

**Figure 2.** Degradation of TCEP (A) and TDCPP (B) by enrichment cultures. The enrichment cultures, 67E (circles) and

The metabolites 2-CE and 1,3-DCP are also persistent and toxic: 1,3-DCP is a known gen‐ otoxin and carcinogen (NTP & NIEHS, 2005), while 2-CE exhibits genotoxicity, fetotoxici‐ ty and cardiotoxicity (National Toxicology, 1985). We analyzed whether the cultures can degrade the metabolites by measuring chloride ion formation. Cultures 67E and 45D liber‐

**TDCPP (μM)**

**0 3 6 9 12**

**5**

**10**

**15**

**20**

**25**

*2.1.1. Enrichment of TCEP and TDCPP-degrading bacteria*

104 Environmental Biotechnology - New Approaches and Prospective Applications

volved hydrolysis of phosphoester bonds.

**0 6 12 18 24**

45D (triangles), were cultivated on 20 µM of TCEP or TDCPP as the sole phosphorus source.

*2.1.1.2. 2-CE and 1,3-DCP degradation ability of enrichment cultures*

**A B**

**0**

**5**

**10**

**15**

**TCEP (μM)**

**20**

**25**

*2.1.1.1. Enrichment cultivation of TCEP and TDCPP-degrading bacteria*

**Figure 3.** Effect of exogenous phosphate on the degradation of TCEP (A) and TDCPP (B) and the chloride ion forma‐ tion from TCEP (C) and TDCPP (D). The enrichment cultures, 67E (A and C) and 45D (B and D), were cultivated on 20 µM of TCEP or TDCPP as the sole phosphorus source, respectively, with various concentrations of inorganic phosphate (NaH2PO4): 0 mM (closed circles), 0.02 mM (closed triangle), 0.2 mM (closed squares) and 2 mM (closed diamonds). Control culture without cell inoculation is indicated by open circles. Each data point represents the mean of at least two independent determinations.

### *2.1.1.3. Effect of exogenous phosphate on the degradation ability of enrichment cultures*

Phosphate-sufficient conditions are well known to repress the expression of genes involved in phosphorus utilization. We thus examined the effect of exogenous inorganic phosphate on TDCPP and TCEP degradations and chloride ion formation (Fig. 3). At concentrations of 0.02, 0.2 and 2 mM, exogenous inorganic phosphate did not significantly inhibit TCEP and TDCPP degradation by the respective cultures (Fig. 3A and B), but chloride ion formation was enhanced at concentrations up to 0.2 mM (Fig. 3C and D). From these results, we con‐ cluded that efficient PFR detoxification could be achieved by optimizing the inorganic phos‐ phate concentration.

**46.5%**

**39.4%**

**Denaturant (%)**

**60%**

**0 3 48 144 0 3 48 144**

**Time (h)**

*2.1.2.1. Isolation of TDCPP- and TCEP-degrading bacteria*

inorganic phosphate. The arrowheads indicated the DNA fragments sequenced.

*2.1.2. Isolation and characterization of TDCPP- and TCEP-degrading bacteria*

peated three times, and a single colony was named strain TCM1 (Fig. 5B).

**TCEP 20 μM TECP 20 μM <sup>+</sup> NaH2PO4 200 μM**

**P source**

**A B**

**Denaturant (%)**

**C1**

**C2**

**55%**

**Figure 4.** DGGE profile of the enrichment cultures 67E (A) and 45D (B) during cultivation in the presence of absence of

We attempted to isolate the bacteria responsible for degrading TDCPP and TCEP in the cul‐ tures 67E and 45D. (Takahashi et al., 2010). In the case of 45D, isolation was achieved by lim‐ iting dilution method. The culture was repeatedly serially diluted in a minimal medium containing 20 μM of TDCPP and cultivated at 30°C. Finally, the culture was spread onto a minimal agar plate containing 232 μM of TDCPP as the sole phosphorus source. A single colony grown on the plate was named strain TDK1 (Fig. 5A). In the case of 67E, the culture was spread onto a minimal agar plate containing 232 μM of TCEP as the sole phosphorus source and incubated at 30°C. Single colonies were then cultivated in a minimal medium containing 20 μM of TCEP as the sole phosphorus source. This isolation procedure was re‐

**0 6 36 96192 0 6 36 96 192 Time (h)**

**TDCPP 20 μM TDCPP 20 μM <sup>+</sup> NaH2PO4 200 μM**

**D1**

**D2 D3**

**P source**

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Microbial Degradation of Persistent Organophosphorus Flame Retardants

### *2.1.1.4. Bacterial communities of enrichment cultures*

To profile the bacterial communities in the cultures, we performed denaturing gradient gel electrophoresis (DGGE) analysis (Fig. 4). In the absence of inorganic phosphate, two bands (C1 and C2) were observed in the fingerprint of TCEP-supplemented 67E, which persisted throughout cultivation (Fig. 4A). With inorganic phosphate added, the intensity of C2 mark‐ edly decreased at later incubation stages (Fig.4A). In 45D supplemented with TDCPP, a sin‐ gle band (D3) was observed at the beginning of cultivation, but at later times two additional bands (D1 and D2) appeared, regardless of the presence or absence of inorganic phosphate (Fig. 4B). However, with inorganic phosphate added, the intensity of D2 and D3 decreased while that of D1 increased at the late stage of cultivation (Fig. 4B). The nucleotide sequence of C1 and D1 was affiliated with the genus *Acidovorax*, that of D2 with the genus *Aquabacteri‐ um*, and C2 and D3 were assigned to the genus *Sphingomonas* (Table 6). Together with the effect of exogenous inorganic phosphate on chlorinated PFRs degradation with liberation of chloride ions, these results imply that the *Sphingomonas*-related bacteria hydrolyze the PFRs, and that the *Acidovorax*-related bacteria dehalogenate the chloroalcohols. Among these bac‐ terial genera, a strain of S*phingomonas* sp. has been reported to hydrolyze some organophos‐ phate pesticides, such as chlorpyrifos (Li et al., 2007). However, bacteria that are known to dehalogenate the chloroalcohols were not identified in the enrichment cultures, suggesting that a new member, possibly *Acidovorax* sp., is responsible for dehalogenating the chloroal‐ cohols in the cultures.


**Table 6.** Phylogenetic affiliation of microorganisms represented by bands in DGGE profiles of the enrichment cultures 67E and 45D.

**Figure 4.** DGGE profile of the enrichment cultures 67E (A) and 45D (B) during cultivation in the presence of absence of inorganic phosphate. The arrowheads indicated the DNA fragments sequenced.

### *2.1.2. Isolation and characterization of TDCPP- and TCEP-degrading bacteria*

### *2.1.2.1. Isolation of TDCPP- and TCEP-degrading bacteria*

on TDCPP and TCEP degradations and chloride ion formation (Fig. 3). At concentrations of 0.02, 0.2 and 2 mM, exogenous inorganic phosphate did not significantly inhibit TCEP and TDCPP degradation by the respective cultures (Fig. 3A and B), but chloride ion formation was enhanced at concentrations up to 0.2 mM (Fig. 3C and D). From these results, we con‐ cluded that efficient PFR detoxification could be achieved by optimizing the inorganic phos‐

To profile the bacterial communities in the cultures, we performed denaturing gradient gel electrophoresis (DGGE) analysis (Fig. 4). In the absence of inorganic phosphate, two bands (C1 and C2) were observed in the fingerprint of TCEP-supplemented 67E, which persisted throughout cultivation (Fig. 4A). With inorganic phosphate added, the intensity of C2 mark‐ edly decreased at later incubation stages (Fig.4A). In 45D supplemented with TDCPP, a sin‐ gle band (D3) was observed at the beginning of cultivation, but at later times two additional bands (D1 and D2) appeared, regardless of the presence or absence of inorganic phosphate (Fig. 4B). However, with inorganic phosphate added, the intensity of D2 and D3 decreased while that of D1 increased at the late stage of cultivation (Fig. 4B). The nucleotide sequence of C1 and D1 was affiliated with the genus *Acidovorax*, that of D2 with the genus *Aquabacteri‐ um*, and C2 and D3 were assigned to the genus *Sphingomonas* (Table 6). Together with the effect of exogenous inorganic phosphate on chlorinated PFRs degradation with liberation of chloride ions, these results imply that the *Sphingomonas*-related bacteria hydrolyze the PFRs, and that the *Acidovorax*-related bacteria dehalogenate the chloroalcohols. Among these bac‐ terial genera, a strain of S*phingomonas* sp. has been reported to hydrolyze some organophos‐ phate pesticides, such as chlorpyrifos (Li et al., 2007). However, bacteria that are known to dehalogenate the chloroalcohols were not identified in the enrichment cultures, suggesting that a new member, possibly *Acidovorax* sp., is responsible for dehalogenating the chloroal‐

67E C1 *Acidovorax* sp.

45D D1 *Acidovorax* sp*.*

C2 *Sphingomonas* sp.

D2 *Aquabacterium* sp.

D3 *Sphingomonas* sp.

**Table 6.** Phylogenetic affiliation of microorganisms represented by bands in DGGE profiles of the enrichment cultures

**Phylogenetic affiliation Species**

phate concentration.

cohols in the cultures.

67E and 45D.

**Culture Band**

*2.1.1.4. Bacterial communities of enrichment cultures*

106 Environmental Biotechnology - New Approaches and Prospective Applications

We attempted to isolate the bacteria responsible for degrading TDCPP and TCEP in the cul‐ tures 67E and 45D. (Takahashi et al., 2010). In the case of 45D, isolation was achieved by lim‐ iting dilution method. The culture was repeatedly serially diluted in a minimal medium containing 20 μM of TDCPP and cultivated at 30°C. Finally, the culture was spread onto a minimal agar plate containing 232 μM of TDCPP as the sole phosphorus source. A single colony grown on the plate was named strain TDK1 (Fig. 5A). In the case of 67E, the culture was spread onto a minimal agar plate containing 232 μM of TCEP as the sole phosphorus source and incubated at 30°C. Single colonies were then cultivated in a minimal medium containing 20 μM of TCEP as the sole phosphorus source. This isolation procedure was re‐ peated three times, and a single colony was named strain TCM1 (Fig. 5B).

### *2.1.2.2. Identification of TDCPP- and TCEP-degrading bacteria*

Both strains were short-rod-shaped bacteria (0.8-1.0 × 1.0-2.5 μm) and produced yellow, cir‐ cular, convex colonies with smooth, glistening surfaces on a nutrient agar plate. As carbon sources, both strains assimilated glucose, maltose and L-arabinose; in addition, strain TCM1 assimilated potassium gluconate, while strain TDK1 assimilated D-mannose, *N*-acetyl-Dglucosamine, and D, L-malate. Both strains tested negative for indole, urease, arginine dihy‐ drolase, nitrate reduction, gelatine hydrolysis, and glucose fermentation, and were positive for esculin hydrolysis. TCM1 and TDK1 tested negative and positive for cytochrome oxi‐ dase, respectively. The morphological and physiological characteristics of the strains were similar to those of *Sphingomonas* spp. Furthermore, the 16S rRNA gene sequence of the strains is closely related to those of sphingomonads, comprising the genera *Sphingomonas*, *Sphingobium*, *Novosphingobium* and *Sphingopyxis* (Takeuchi et al., 2001). The phylogenetic tree constructed from the sequences of these genera showed that strains TCM1 and TDK1 belong to *Sphingobium* and *Sphingomonas*, respectively

strain TCM1 grew moderately on tributyl phosphate and slightly on tris(2-butoxyethyl) phosphate, triethyl phosphate and trimethyl phosphate. These results demonstrate that the strains can degrade not only TDCPP and TCEP but also other PFRs, and that the strains

TCEP・TDCPP (μM)

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**Figure 6.** Degradation of TDCPP and TCEP by strains TCM1 (A) and TDK1 (B) and generation of 2-CE and 1,3-DCP (C and D). The cultivations were performed aerobically at 30°C in a minimal medium containing 20 μM of TCEP or TDCPP as the sole phosphorus source. (A and B) Open circles and triangles represent the concentrations of TCEP and TDCPP, respectively, and their filled forms represent concentrations for autoclaved control cells. (C and D) Open circles and triangles represent the concentrations of 2-CE and 1,3-DCP, respectively. Each data point represents the mean of at

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have different substrate specificity for trialkyl phosphates.

A B

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least two independent determinations.

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TCM1

Figure 1. SEM micrographs of TCEP- and TDCPP-degrading bacteria *Sphingobium* sp. strain TCM1 (A) and *Sphingomonas* sp. stain TDK1 (B). **Figure 5.** SEM micrographs of TCEP- and TDCPP-degrading bacteria *Sphingobium* sp. strain TCM1 (A) and Sphingomo‐ *nas* sp. stain TDK1 (B).

### *2.1.2.3. Degradation ability of TCEP and TDCPP-degrading bacteria*

Both strains completely degraded 20 μM of TDCPP within 6 h (Fig. 6A and B). Strain TDK1, however, was 48 times less effective in degrading TCEP than TCM1 (TCEP degradation time was 144 h for TDK1, versus 3 h for TCM1) (Fig. 6A and B). During the degradations, 1,3- DCP and 2-CE were detected in the cultures of both strains and were not further degraded (Fig. 6C and D). These results showed that the strains degrade the compounds by hydrolyz‐ ing their phosphotriester bonds. To date, TCM1 and TDK1 are the only isolated microorgan‐ isms reported to degrade the persistent PFRs.

We then analyzed whether the strains can degrade other PFRs by utilizing them as sole phosphorus source. Both strains grew on tris(2,3-dibromopropyl) phosphate, tricresyl and triphenyl phosphates. Stain TDK1 did not grow on all trialkyl phosphates tested, whereas strain TCM1 grew moderately on tributyl phosphate and slightly on tris(2-butoxyethyl) phosphate, triethyl phosphate and trimethyl phosphate. These results demonstrate that the strains can degrade not only TDCPP and TCEP but also other PFRs, and that the strains have different substrate specificity for trialkyl phosphates.

*2.1.2.2. Identification of TDCPP- and TCEP-degrading bacteria*

108 Environmental Biotechnology - New Approaches and Prospective Applications

to *Sphingobium* and *Sphingomonas*, respectively

*nas* sp. stain TDK1 (B).

**A B**

*2.1.2.3. Degradation ability of TCEP and TDCPP-degrading bacteria*

isms reported to degrade the persistent PFRs.

Both strains were short-rod-shaped bacteria (0.8-1.0 × 1.0-2.5 μm) and produced yellow, cir‐ cular, convex colonies with smooth, glistening surfaces on a nutrient agar plate. As carbon sources, both strains assimilated glucose, maltose and L-arabinose; in addition, strain TCM1 assimilated potassium gluconate, while strain TDK1 assimilated D-mannose, *N*-acetyl-Dglucosamine, and D, L-malate. Both strains tested negative for indole, urease, arginine dihy‐ drolase, nitrate reduction, gelatine hydrolysis, and glucose fermentation, and were positive for esculin hydrolysis. TCM1 and TDK1 tested negative and positive for cytochrome oxi‐ dase, respectively. The morphological and physiological characteristics of the strains were similar to those of *Sphingomonas* spp. Furthermore, the 16S rRNA gene sequence of the strains is closely related to those of sphingomonads, comprising the genera *Sphingomonas*, *Sphingobium*, *Novosphingobium* and *Sphingopyxis* (Takeuchi et al., 2001). The phylogenetic tree constructed from the sequences of these genera showed that strains TCM1 and TDK1 belong

Figure 1. SEM micrographs of TCEP- and TDCPP-degrading bacteria *Sphingobium* sp. strain TCM1 (A) and *Sphingomonas* sp. stain TDK1 (B).

**Figure 5.** SEM micrographs of TCEP- and TDCPP-degrading bacteria *Sphingobium* sp. strain TCM1 (A) and Sphingomo‐

Both strains completely degraded 20 μM of TDCPP within 6 h (Fig. 6A and B). Strain TDK1, however, was 48 times less effective in degrading TCEP than TCM1 (TCEP degradation time was 144 h for TDK1, versus 3 h for TCM1) (Fig. 6A and B). During the degradations, 1,3- DCP and 2-CE were detected in the cultures of both strains and were not further degraded (Fig. 6C and D). These results showed that the strains degrade the compounds by hydrolyz‐ ing their phosphotriester bonds. To date, TCM1 and TDK1 are the only isolated microorgan‐

We then analyzed whether the strains can degrade other PFRs by utilizing them as sole phosphorus source. Both strains grew on tris(2,3-dibromopropyl) phosphate, tricresyl and triphenyl phosphates. Stain TDK1 did not grow on all trialkyl phosphates tested, whereas

**Figure 6.** Degradation of TDCPP and TCEP by strains TCM1 (A) and TDK1 (B) and generation of 2-CE and 1,3-DCP (C and D). The cultivations were performed aerobically at 30°C in a minimal medium containing 20 μM of TCEP or TDCPP as the sole phosphorus source. (A and B) Open circles and triangles represent the concentrations of TCEP and TDCPP, respectively, and their filled forms represent concentrations for autoclaved control cells. (C and D) Open circles and triangles represent the concentrations of 2-CE and 1,3-DCP, respectively. Each data point represents the mean of at least two independent determinations.

### **2.2. Microbial detoxification of TDCPP and TCEP by two bacterial strains**

We have successfully isolated TCEP- and TDCPP-degrading bacteria. However, neither strain can degrade the resulting toxic and persistent metabolites 2-CE and 1,3-DCP. Elimina‐ tion of the metabolites is required before the strains can be used to degrade TDCPP and TCEP in practice. Fortunately, bacteria with chloroalcohol-degrading ability have been welldocumented. We thus attempted to completely detoxify the PFRs by combining strain TCM1 with bacteria capable of degrading the chloroalcohols (Takahashi et al., 2012a; Taka‐ hashi, et al., 2012b).

*2.2.1.2. Optimum TDCPP and 1,3-DCP degradation conditions of strains TCM1 and PY1*

and 9.5 for strains TCM1 and PY1, respectively.

A B

100

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60

Relative activity (%)

40

20

0

two independent determinations.

0 10 20 30 40 50 60 Temperature (°C)

*2.2.1.3. Complete detoxification of TDCPP by mixed bacteria cells*

We then determined the optimum temperature and pH for lyophilized cell activity (Fig. 8). At pH 9.0 for strain TCM1 and pH 8.5 for strain PY1, the highest activity of TCM1 and PY1 cells occurred at 30°C (2.53 μmol h-1 OD660-1) and 35°C (1.31 μmol h-1 OD660-1), respectively (Fig. 8A). At 30°C, the highest activity of TCM1 and PY1 cells occurred at pH 8.5 (2.48 μmol h-1 OD660-1) and pH 9.5 with 50 mM Tris-H2SO4 (0.95 μmol h-1 OD660-1), respectively (Fig. 8B). We thus established the optimum temperature as 30°C and 35°C and the optimum pH as 8.5

Relative activity (%)

**Figure 8.** Effect of temperature and pH on the degradation activity of strains TCM1 and PY1. (A) effect of tempera‐ ture: TDCPP hydrolyzation activity of strain TCM1 cells (closed circle) and 1,3-DCP dehalogenation activity of strain PY1 cells (open circle) were, respectively, assayed in 50 mM Tris-H2SO4 buffer (pH 9.0) and 50 mM Tris-H2SO4 buffer (pH 8.5). (B) effect of pH: TDCPP hydrolyzation activity of strain TCM1 cells (closed symbols) and 1,3-DCP dehalogenation activity of strain PY1 cells (open symbols) was assayed at 30°C in 50 mM MOPS-NaOH buffer (circle, pH 6.0-7.5), Tris-H2SO4 buffer (triangle, pH 7.5-9.5), and glycine-NaOH buffer (square, pH 9.0-12.0). Each datum represents means of

Based on the optimum conditions, we set the reaction temperature to 30°C and pH to 9.0 (50 mM Tris-H2SO4) for TDCPP detoxification by mixed bacteria (Fig. 9). Under these condi‐ tions, the respective activities of strains TCM1 and PY1 were 2.21 and 0.92 μmol h-1 OD660-1. In the detoxification reaction using a mixture of TCM1 and PY1 cells (OD660 0.05 and 0.2, re‐ spectively), approximately 50 μM of TDCPP disappeared within 1 h, and 1,3-DCP and chlor‐ ide ions were formed to levels of approximately 100 and 120 μM, respectively, after 2 h (Fig. 9A). This result suggests incomplete detoxification of TDCPP due to low 1,3-DCP dehaloge‐ nation activity. Increasing the strain PY1 population to an OD660 of 4.0 decreased the TDCPP hydrolyzation rate of TCM1 cells, but completely eliminated the resulting 1,3-DCP after 10 h (Fig. 9B). At the same time, chloride ion concentration had reached its theoretical value ex‐

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pH 56789 10 11 12 13

### *2.2.1. Microbial detoxification of TDCPP using Sphingobium sp. strain TCM1 and Arthrobacter sp. strain PY1*

Several 1,3-DCP-degrading bacteria have been reported, including *Arthrobacter* sp. strains PY1 (Yonetani et al., 2004) and AD2 (van den Wijngaard et al., 1991), *A. erithii* H10a (Assis et al., 1998), *Agrobacterium radiobacter* strain AD1 (van den Wijngaard et al., 1989), and *Coryne‐ bacterium* sp. strain N-1074 (Nakamura et al., 1991). Of these, *Arthrobacter* sp. strain PY1 ex‐ hibits high 1,3-DCP degradation ability. Therefore, we attempted to detoxify TDCPP by cohabitation of strain TCM1 and *Arthrobacter* sp. PY1 in a resting cell reaction (Fig. 7) (Takahashi et al., 2012a).

**Figure 7.** Complete detoxification of TDCPP by *Sphingobium* sp. strain TCM1 and 1,3-DCP-degrading bacterium Ar‐ *throbacter* sp. strain PY1.

### *2.2.1.1. Freezing and lyophilization of strains TCM1 and PY1 cells*

For resting cell preparation, we first examined the effect of freezing and lyophilization on the activity of strains TCM1 and PY1. The TDCPP-hydrolyzing activity of strain TCM1 in‐ tact cells was 1.07 μmol h-1 OD660-1, whereas respective activities of frozen and lyophilized cells were 0.90 and 0.84 μmol h-1 OD660-1. On the other hand, the 1,3-DCP-dehalogenating ac‐ tivity of strain PY1 intact cells was 0.22 μmol h-1 OD660-1, with respective frozen and lyophi‐ lized cell activities of 0.23 and 0.26 μmol h-1 OD660-1. These results reveal that freezing and lyophilization treatments cause no significant decline in degradation activities of the strains.

### *2.2.1.2. Optimum TDCPP and 1,3-DCP degradation conditions of strains TCM1 and PY1*

**2.2. Microbial detoxification of TDCPP and TCEP by two bacterial strains**

110 Environmental Biotechnology - New Approaches and Prospective Applications

hashi, et al., 2012b).

(Takahashi et al., 2012a).

Cl

Cl

O

*throbacter* sp. strain PY1.

O

Cl

Cl Cl

O P O

Tris(1,3-dichloro-2-propyl) phosphate (TDCPP)

Cl

*2.2.1.1. Freezing and lyophilization of strains TCM1 and PY1 cells*

O H3 PO4

*strain PY1*

We have successfully isolated TCEP- and TDCPP-degrading bacteria. However, neither strain can degrade the resulting toxic and persistent metabolites 2-CE and 1,3-DCP. Elimina‐ tion of the metabolites is required before the strains can be used to degrade TDCPP and TCEP in practice. Fortunately, bacteria with chloroalcohol-degrading ability have been welldocumented. We thus attempted to completely detoxify the PFRs by combining strain TCM1 with bacteria capable of degrading the chloroalcohols (Takahashi et al., 2012a; Taka‐

*2.2.1. Microbial detoxification of TDCPP using Sphingobium sp. strain TCM1 and Arthrobacter sp.*

Several 1,3-DCP-degrading bacteria have been reported, including *Arthrobacter* sp. strains PY1 (Yonetani et al., 2004) and AD2 (van den Wijngaard et al., 1991), *A. erithii* H10a (Assis et al., 1998), *Agrobacterium radiobacter* strain AD1 (van den Wijngaard et al., 1989), and *Coryne‐ bacterium* sp. strain N-1074 (Nakamura et al., 1991). Of these, *Arthrobacter* sp. strain PY1 ex‐ hibits high 1,3-DCP degradation ability. Therefore, we attempted to detoxify TDCPP by cohabitation of strain TCM1 and *Arthrobacter* sp. PY1 in a resting cell reaction (Fig. 7)

> OH Cl Cl

1,3-Dichloro-2-propanol (1,3-DCP)

3H 6H2 O 6HCl <sup>2</sup>

**Figure 7.** Complete detoxification of TDCPP by *Sphingobium* sp. strain TCM1 and 1,3-DCP-degrading bacterium Ar‐

For resting cell preparation, we first examined the effect of freezing and lyophilization on the activity of strains TCM1 and PY1. The TDCPP-hydrolyzing activity of strain TCM1 in‐ tact cells was 1.07 μmol h-1 OD660-1, whereas respective activities of frozen and lyophilized cells were 0.90 and 0.84 μmol h-1 OD660-1. On the other hand, the 1,3-DCP-dehalogenating ac‐ tivity of strain PY1 intact cells was 0.22 μmol h-1 OD660-1, with respective frozen and lyophi‐ lized cell activities of 0.23 and 0.26 μmol h-1 OD660-1. These results reveal that freezing and lyophilization treatments cause no significant decline in degradation activities of the strains.

3

*Sphingobium* sp. strain TCM1 *Arthrobacter* sp. strain PY1

3

OH OH OH

Glycerol

We then determined the optimum temperature and pH for lyophilized cell activity (Fig. 8). At pH 9.0 for strain TCM1 and pH 8.5 for strain PY1, the highest activity of TCM1 and PY1 cells occurred at 30°C (2.53 μmol h-1 OD660-1) and 35°C (1.31 μmol h-1 OD660-1), respectively (Fig. 8A). At 30°C, the highest activity of TCM1 and PY1 cells occurred at pH 8.5 (2.48 μmol h-1 OD660-1) and pH 9.5 with 50 mM Tris-H2SO4 (0.95 μmol h-1 OD660-1), respectively (Fig. 8B). We thus established the optimum temperature as 30°C and 35°C and the optimum pH as 8.5 and 9.5 for strains TCM1 and PY1, respectively.

**Figure 8.** Effect of temperature and pH on the degradation activity of strains TCM1 and PY1. (A) effect of tempera‐ ture: TDCPP hydrolyzation activity of strain TCM1 cells (closed circle) and 1,3-DCP dehalogenation activity of strain PY1 cells (open circle) were, respectively, assayed in 50 mM Tris-H2SO4 buffer (pH 9.0) and 50 mM Tris-H2SO4 buffer (pH 8.5). (B) effect of pH: TDCPP hydrolyzation activity of strain TCM1 cells (closed symbols) and 1,3-DCP dehalogenation activity of strain PY1 cells (open symbols) was assayed at 30°C in 50 mM MOPS-NaOH buffer (circle, pH 6.0-7.5), Tris-H2SO4 buffer (triangle, pH 7.5-9.5), and glycine-NaOH buffer (square, pH 9.0-12.0). Each datum represents means of two independent determinations.

### *2.2.1.3. Complete detoxification of TDCPP by mixed bacteria cells*

Based on the optimum conditions, we set the reaction temperature to 30°C and pH to 9.0 (50 mM Tris-H2SO4) for TDCPP detoxification by mixed bacteria (Fig. 9). Under these condi‐ tions, the respective activities of strains TCM1 and PY1 were 2.21 and 0.92 μmol h-1 OD660-1. In the detoxification reaction using a mixture of TCM1 and PY1 cells (OD660 0.05 and 0.2, re‐ spectively), approximately 50 μM of TDCPP disappeared within 1 h, and 1,3-DCP and chlor‐ ide ions were formed to levels of approximately 100 and 120 μM, respectively, after 2 h (Fig. 9A). This result suggests incomplete detoxification of TDCPP due to low 1,3-DCP dehaloge‐ nation activity. Increasing the strain PY1 population to an OD660 of 4.0 decreased the TDCPP hydrolyzation rate of TCM1 cells, but completely eliminated the resulting 1,3-DCP after 10 h (Fig. 9B). At the same time, chloride ion concentration had reached its theoretical value ex‐ pected from the initial TDCPP concentration, demonstrating that complete detoxification of TDCPP is achievable using strains TCM1 and PY1.

*2.2.2.1. Optimum TCEP degradation condition of strain TCM1*

A B <sup>40</sup>

volved in the degradation of both compounds.

OD660 of 0.8 with 50 μM of Co2+.

0 0

2

to stop the reaction.

4

6

8

10

12

TCEP (µM)

We first determined the optimum temperature and pH for TCEP degradation by strain TCM1 in a resting reaction using lyophilized cells. At pH 7.4, the highest activity was ob‐ tained at 30°C (14.1 nmol min-1 OD660-1). Maintaining this temperature and varying the pH, the highest activity was recorded at pH 8.5 (14.6 nmol min-1 OD660-1). These optimum condi‐ tions were identical to those for TDCPP, suggesting that the same enzyme(s) might be in‐

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Under the optimum conditions, TCM1 cells completely eliminated 10, 20 and 50 μM of TCEP within 3 h, but the generated 2-CE was approximately 50% of its theoretical value based on the initial TCEP concentrations (Fig. 11). Phosphotriesterase that can hydrolyze or‐ ganophosphorus pesticides structurally similar to TCEP, such as chlorpyrifos, require two zinc ions for catalysis, and enzyme activity can be maximized by replacing Zn2+ with Co2+ (Omburo et al., 1992). A bacterial phosphodiesterase that can hydrolyze alkyl phosphodiest‐ ers similarly requires divalent metals (Gerlt & Wan, 1979). We therefore examined the effect of Co2+ as well as cell amount on TCEP hydrolysis (Fig. 11). In the reaction using approxi‐ mately 10 μM of TCEP without Co2+, 2-CE reached 21.2 μM (OD660 of 0.8) after 3 h. Addition of 50 μM Co2+ resulted in an increase of 2-CE to 32.3 μM, equivalent to the theoretical value of 30 μM (Fig. 11B). These results showed that complete hydrolysis can be achieved at an

Time (h) Time (h)

**Figure 11.** Effect of Co2+ and cell amount on TCEP hydrolysis by strain TCM1-resting cells. The reactions were per‐ formed at 30°C using the resting cells at OD660 of 0.4 (circles) or 0.8 (triangles) with (open symbols) or without (closed symbols) 50 μM Co2+ in 50 mM Tris-H2SO4 buffer (pH 8.5) containing 10 μM TCEP, and TCEP (A) and 2-CE (B) were determined. Each datum represents the mean of two independent determinations. The inconsistency of the initial concentrations of TCEP at zero time with the set-up ones was mainly attributed to reaction progress in several minutes

0 0 1 2 123

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20

30

2-CE (µM)

**Figure 9.** Complete detoxification of TDCPP by the mixed resting cells of strains TCM1 and PY1. The reactions were performed at 30°C with 50 μM TDCPP in 50 mM Tris - H2SO4 buffer (pH 9.0), and TDCPP (circles), 1,3-DCP (triangles) and chloride ion (squares) were determined. Cell concentrations of strains TCM1 and PY1 for each reaction were, re‐ spectively, OD660 of 0.05 and 0.2 (A) and 0.04 and 4.0 (B). Each datum represents means of two independent determi‐ nations.

### *2.2.2. Microbial detoxification of TCEP using Sphingobium sp. strain TCM1 and Xanthobacter autotrophicus strain GJ10*

Several 2-CE-degrading bacteria have been reported, including *Xanthobacter autotrophicus* strain GJ10 (Janssen et al., 1985), *Pseudomonas putida* strain US2 (Strotmann et al., 1990) and *P. atutzeri* strain JJ (Dijk et al., 2003). Among these, the degradation of 2-CE by *X. autotrophicus* strain GJ10 has been well characterized. Therefore, we attempted to detoxi‐ fy TCEP by co-habitation of strain TCM1 and *X. autotrophicus* strain GJ10 (Fig. 10) (Taka‐ hashi et al., 2012b).

**Figure 10.** Complete detoxification of TCEP by *Sphingobium* sp. strain TCM1 and 2-CE-degrading bacterium Xantho‐ *bacter autotrophicus* strain GJ10.

### *2.2.2.1. Optimum TCEP degradation condition of strain TCM1*

pected from the initial TDCPP concentration, demonstrating that complete detoxification of

TDCPP・1,3-DCP (µM)

Cl ー (µM)

**Figure 9.** Complete detoxification of TDCPP by the mixed resting cells of strains TCM1 and PY1. The reactions were performed at 30°C with 50 μM TDCPP in 50 mM Tris - H2SO4 buffer (pH 9.0), and TDCPP (circles), 1,3-DCP (triangles) and chloride ion (squares) were determined. Cell concentrations of strains TCM1 and PY1 for each reaction were, re‐ spectively, OD660 of 0.05 and 0.2 (A) and 0.04 and 4.0 (B). Each datum represents means of two independent determi‐

*2.2.2. Microbial detoxification of TCEP using Sphingobium sp. strain TCM1 and Xanthobacter*

Several 2-CE-degrading bacteria have been reported, including *Xanthobacter autotrophicus* strain GJ10 (Janssen et al., 1985), *Pseudomonas putida* strain US2 (Strotmann et al., 1990) and *P. atutzeri* strain JJ (Dijk et al., 2003). Among these, the degradation of 2-CE by *X. autotrophicus* strain GJ10 has been well characterized. Therefore, we attempted to detoxi‐ fy TCEP by co-habitation of strain TCM1 and *X. autotrophicus* strain GJ10 (Fig. 10) (Taka‐

Cl 3

3H 6H2 O 6HCl <sup>2</sup>

**Figure 10.** Complete detoxification of TCEP by *Sphingobium* sp. strain TCM1 and 2-CE-degrading bacterium Xantho‐

3

Cl

O H3 PO4

> 2-Chloroethanol (2-CE)

*Sphingobium* sp. strain TCM1 *Xanthobacter autotrophicus* strain GJ10

OH

100

200

300

400

0

60 40 20

0

0 2 4 8 6 10 12 Time (h)

Cl ー (µM)

100

200

300

400

0

Glycolic acid

OH

O

OH

80 100 120

TDCPP is achievable using strains TCM1 and PY1.

012 Time (h) 0.5 1.5

60 40 20

TDCPP・1,3-DCP (µM)

nations.

0

*autotrophicus strain GJ10*

hashi et al., 2012b).

Cl

O P O

O

*bacter autotrophicus* strain GJ10.

O

Cl

Tris(2-chloroethyl) phosphate (TCEP)

80 100 120

A B

112 Environmental Biotechnology - New Approaches and Prospective Applications

We first determined the optimum temperature and pH for TCEP degradation by strain TCM1 in a resting reaction using lyophilized cells. At pH 7.4, the highest activity was ob‐ tained at 30°C (14.1 nmol min-1 OD660-1). Maintaining this temperature and varying the pH, the highest activity was recorded at pH 8.5 (14.6 nmol min-1 OD660-1). These optimum condi‐ tions were identical to those for TDCPP, suggesting that the same enzyme(s) might be in‐ volved in the degradation of both compounds.

Under the optimum conditions, TCM1 cells completely eliminated 10, 20 and 50 μM of TCEP within 3 h, but the generated 2-CE was approximately 50% of its theoretical value based on the initial TCEP concentrations (Fig. 11). Phosphotriesterase that can hydrolyze or‐ ganophosphorus pesticides structurally similar to TCEP, such as chlorpyrifos, require two zinc ions for catalysis, and enzyme activity can be maximized by replacing Zn2+ with Co2+ (Omburo et al., 1992). A bacterial phosphodiesterase that can hydrolyze alkyl phosphodiest‐ ers similarly requires divalent metals (Gerlt & Wan, 1979). We therefore examined the effect of Co2+ as well as cell amount on TCEP hydrolysis (Fig. 11). In the reaction using approxi‐ mately 10 μM of TCEP without Co2+, 2-CE reached 21.2 μM (OD660 of 0.8) after 3 h. Addition of 50 μM Co2+ resulted in an increase of 2-CE to 32.3 μM, equivalent to the theoretical value of 30 μM (Fig. 11B). These results showed that complete hydrolysis can be achieved at an OD660 of 0.8 with 50 μM of Co2+.

**Figure 11.** Effect of Co2+ and cell amount on TCEP hydrolysis by strain TCM1-resting cells. The reactions were per‐ formed at 30°C using the resting cells at OD660 of 0.4 (circles) or 0.8 (triangles) with (open symbols) or without (closed symbols) 50 μM Co2+ in 50 mM Tris-H2SO4 buffer (pH 8.5) containing 10 μM TCEP, and TCEP (A) and 2-CE (B) were determined. Each datum represents the mean of two independent determinations. The inconsistency of the initial concentrations of TCEP at zero time with the set-up ones was mainly attributed to reaction progress in several minutes to stop the reaction.

### *2.2.2.2. Optimum 2-CE degradation condition of strain GJ10*

We prepared resting cells of intact, frozen and lyophilized cells of *X. autotrophicus* strain GJ10 and examined their 2-CE degradation activity. Activity was detected only in frozen cells at 4.93 pmol min-1 OD450-1, four orders lower than the TCEP degradation activity of strain TCM1. This low 2-CE degradation activity might be attributable to the lack of coenzyme regeneration of enzymes involved in the degradation process. We next exam‐ ined 2-CE degradation in a growing cell reaction. The growing cells completely degrad‐ ed approximately 180 μM of 2-CE within 24 h. The degradation ability was estimated to be a minimum of 7.5 μM h-1, comparable to the TCEP degradation ability of strain TCM1-resting cells (approximately 10 μM h-1). This result shows that growing cells of strain GJ10 can degrade 2-CE effectively.

### *2.2.2.3. Complete detoxification of TCEP by two bacterial strains*

Based on the results described above, we examined whether combining TCEP hydrolysis by TCM1 resting cells and 2-CE degradation by GJ10 growing cells would completely detoxify TCEP (Fig. 12). TCM1 resting cells abolished 9.6 μM of TCEP within 4 h, releas‐ ing 2-CE at 29.0 μM, equivalent to that estimated from the initial TCEP concentration, and consistent with complete TCEP hydrolysis (Fig. 12A and B). The generated 2-CE was abolished by GJ10 growing cells within 48 h, and chloride ion concentration reached 30.2 μM after 144 h, equivalent to that estimated from the generated 2-CE (Fig. 12C and D). Taken together, these results demonstrate that complete detoxification of TDCPP can be achieved using strains TCM1 and GJ10.

2-CE (µM)

30

10

8

6 4

TCEP (µM)

2

0

12

20

10

0

of two independent determinations.

**Acknowledgements**

0 48 96 144

0 h 4 h

C D

A B

TCM1 - - + +

ND

Cl ー (µM)

Time (h) Time (h)

**Figure 12.** Complete detoxification of TCEP by *Sphingobium* sp. strain TCM1-resting cell reaction (A and B) and the following *X. autotrophicus* GJ10-growing cell reaction (C and D). The resting cell reaction was performed at 30°C with (+) or without (-) strain TCM1 cells at OD660 of 0.8 in 50 mM Tris-H2SO4 buffer (pH 8.5) containing 10 μM TCEP and 50 μM Co2+, and TCEP (A) and 2-CE (B) were determined. The growing cell reaction was performed at 30°C with (closed symbols) or without (open symbols) strain GJ10 cells in a medium containing the generated 2-CE as the sole carbon source, and 2-CE (C) and chloride ion (D) was determined. ND means not detected. Each datum represents the mean

This research was supported in part by a Grant-in-Aid for Scientific Research (B) (to Y. K) from the Ministry of Education, Science, Sports, and Culture of Japan, by the River Environ‐ ment Fund (REF) in charge of the Foundation of River and Watershed Environment Man‐

2-CE (µM)

0

30

20

10

0

10

20

30

Microbial Degradation of Persistent Organophosphorus Flame Retardants

http://dx.doi.org/10.5772/53749

115

0 48 96 144

ND ND

TCM1 - - + +

0 h 4 h

### **3. Concluding remarks**

We have successfully isolated two novel bacterial strains capable of degrading the persistent and potential toxic PFRs, TCEP and TDCPP, which have become worldwide environmental contaminants. The two strains TCM1 and TDK1 belong to *Sphingobium* sp. and *Sphingomonas* sp. respectively. The strains are the first microorganisms reported to degrade the persistent PFRs. They degrade the compounds by hydrolyzing their phosphotriester bonds to produce metabolites 1,3-DCP from TDCPP and 2-CE from TCEP, which are themselves toxic and non-self-biodegradable. In a successful attempt to completely detoxify the FPRs, we com‐ bined TCM1 with the 1.3-DCP-degrading bacterium *Arthrobacter* sp. strain PY1 (for TDCPP degradation), and with the 2-CE-degrading bacterium *X. autotrophicus* strain GJ10 (for TCEP degradation). This is the first description of microbial FPR detoxification. The bacteria and the microbial detoxification techniques may prove useful for the bioremediation of sites con‐ taminated with intractable compounds. Further studies on the PFRs-degrading bacteria as well as the chloroalcohols-degrading bacteria, and on the detoxification techniques, could help to establish more efficient detoxifications, and could also provide novel insights into microbial degradation of organophosphorus compounds. We are now working towards elu‐ cidating the enzymes and the genes involved in the degradation processes.

**Figure 12.** Complete detoxification of TCEP by *Sphingobium* sp. strain TCM1-resting cell reaction (A and B) and the following *X. autotrophicus* GJ10-growing cell reaction (C and D). The resting cell reaction was performed at 30°C with (+) or without (-) strain TCM1 cells at OD660 of 0.8 in 50 mM Tris-H2SO4 buffer (pH 8.5) containing 10 μM TCEP and 50 μM Co2+, and TCEP (A) and 2-CE (B) were determined. The growing cell reaction was performed at 30°C with (closed symbols) or without (open symbols) strain GJ10 cells in a medium containing the generated 2-CE as the sole carbon source, and 2-CE (C) and chloride ion (D) was determined. ND means not detected. Each datum represents the mean of two independent determinations.

### **Acknowledgements**

*2.2.2.2. Optimum 2-CE degradation condition of strain GJ10*

114 Environmental Biotechnology - New Approaches and Prospective Applications

*2.2.2.3. Complete detoxification of TCEP by two bacterial strains*

TDCPP can be achieved using strains TCM1 and GJ10.

strain GJ10 can degrade 2-CE effectively.

**3. Concluding remarks**

We prepared resting cells of intact, frozen and lyophilized cells of *X. autotrophicus* strain GJ10 and examined their 2-CE degradation activity. Activity was detected only in frozen cells at 4.93 pmol min-1 OD450-1, four orders lower than the TCEP degradation activity of strain TCM1. This low 2-CE degradation activity might be attributable to the lack of coenzyme regeneration of enzymes involved in the degradation process. We next exam‐ ined 2-CE degradation in a growing cell reaction. The growing cells completely degrad‐ ed approximately 180 μM of 2-CE within 24 h. The degradation ability was estimated to be a minimum of 7.5 μM h-1, comparable to the TCEP degradation ability of strain TCM1-resting cells (approximately 10 μM h-1). This result shows that growing cells of

Based on the results described above, we examined whether combining TCEP hydrolysis by TCM1 resting cells and 2-CE degradation by GJ10 growing cells would completely detoxify TCEP (Fig. 12). TCM1 resting cells abolished 9.6 μM of TCEP within 4 h, releas‐ ing 2-CE at 29.0 μM, equivalent to that estimated from the initial TCEP concentration, and consistent with complete TCEP hydrolysis (Fig. 12A and B). The generated 2-CE was abolished by GJ10 growing cells within 48 h, and chloride ion concentration reached 30.2 μM after 144 h, equivalent to that estimated from the generated 2-CE (Fig. 12C and D). Taken together, these results demonstrate that complete detoxification of

We have successfully isolated two novel bacterial strains capable of degrading the persistent and potential toxic PFRs, TCEP and TDCPP, which have become worldwide environmental contaminants. The two strains TCM1 and TDK1 belong to *Sphingobium* sp. and *Sphingomonas* sp. respectively. The strains are the first microorganisms reported to degrade the persistent PFRs. They degrade the compounds by hydrolyzing their phosphotriester bonds to produce metabolites 1,3-DCP from TDCPP and 2-CE from TCEP, which are themselves toxic and non-self-biodegradable. In a successful attempt to completely detoxify the FPRs, we com‐ bined TCM1 with the 1.3-DCP-degrading bacterium *Arthrobacter* sp. strain PY1 (for TDCPP degradation), and with the 2-CE-degrading bacterium *X. autotrophicus* strain GJ10 (for TCEP degradation). This is the first description of microbial FPR detoxification. The bacteria and the microbial detoxification techniques may prove useful for the bioremediation of sites con‐ taminated with intractable compounds. Further studies on the PFRs-degrading bacteria as well as the chloroalcohols-degrading bacteria, and on the detoxification techniques, could help to establish more efficient detoxifications, and could also provide novel insights into microbial degradation of organophosphorus compounds. We are now working towards elu‐

cidating the enzymes and the genes involved in the degradation processes.

This research was supported in part by a Grant-in-Aid for Scientific Research (B) (to Y. K) from the Ministry of Education, Science, Sports, and Culture of Japan, by the River Environ‐ ment Fund (REF) in charge of the Foundation of River and Watershed Environment Man‐ agement (FOREM) (to Y. K.), by a grant from the Uchida Energy Science Promotion Foundation (to S. T.) and by the Kurita Water and Environment Foundation (to S. T).

[9] Benotti, M. J., Stanford, B. D., Wert, E. C., & Snyder, S. A. (2009). Evaluation of a pho‐ tocatalytic reactor membrane pilot system for the removal of pharmaceuticals and

Microbial Degradation of Persistent Organophosphorus Flame Retardants

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### **Author details**

Shouji Takahashi\* , Katsumasa Abe and Yoshio Kera

Department of Environmental Systems Engineering, Nagaoka University of Technology, Ka‐ mitomioka, Nagaoka, Niigata, Japan

### **References**


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**Chapter 6**

**Continuous Biotechnological Treatment of Cyanide**

**Contaminated Waters by Using a Cyanide Resistant**

Various industries release a combination of free cyanide and cyanide complexes into the en‐ vironment via a variety of disposal methods, particularly as wastewater. These industries utilise cyanide based compounds in various operations, including: the beneficiation of met‐ als, electroplating, case hardening, automotive manufacturing, circuitry board manufactur‐ ing, and in chemical industries [23]. Cyanide is often found in organic, hydrocarbon chains or as inorganic, transition, alkali and alkali earth metal complexes [20]. Many cyanide com‐ plexes are highly unstable, thus temperature, pH and light can degrade the components to

There is an overwhelming popularity in industry for the use of chemical treatment methods for the treatment of free cyanide and cyanide complexes compared to biochemical treatment methods. Chemical remediation methods like alkaline chlorine oxidation are commonly used to treat cyanide contaminated wastewater [23, 24]. Chemical oxidation is particularly ineffective in the treatment of cyanide-metal complexes containing heavy metals, such as copper, nickel and silver, due to the slow reaction rate [23]. The excess quantity of chlorine used in the treatment process increases the chemical oxygen demand (COD) of the wastewa‐ ter thereby rendering the water undesirable for reuse, toxic to aquatic life and may produce organic substances. In order to reduce operational costs, some manufacturers partially treat the wastewater, resulting in untreated and/or partially decomposed cyanide being dis‐ charged. Other methods of treatment include copper catalysed hydrogen peroxide oxida‐

> © 2013 Santos et al.; licensee InTech. This is an open access article distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use,

© 2013 Santos et al.; licensee InTech. This is a paper distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

distribution, and reproduction in any medium, provided the original work is properly cited.

**Species of** *Aspergillus awamori*

Additional information is available at the end of the chapter

form free cyanide which is the most toxic form of cyanide [20, 26].

Bruno Alexandre Quistorp Santos, Seteno Karabo Obed Ntwampe and

James Hamuel Doughari

http://dx.doi.org/10.5772/53349

**1. Introduction**

