Which Enzymes are Responsible for the Biodegradation of Noladin Ether?

Which Enzymes are Responsible for the Biodegradation of Noladin Ether?

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Which enzymes degrade the CB1-specific endogenous cannabinoid 2-arachidonyl glyceryl ether? (Noladin ether)

Noladin is an endocannabinoid. The best-characterised members of this family of molecules are Arachidonoylethanolamine (AEA; anandamide) and 2-arachidonoylglycerol (2-AG). The degradation of these endocannabinoids is reviewed in:

Pamplona, FA et al. (2012) Psychopharmacology of the endocannabinoids: far beyond anandamide. Journal of Psychopharmacology 26:7-22

These endocannabinoids are inactivated via a reuptake system that is present in both neuronal and glial cells, although the protein responsible has apparently not been identified.

AEA is then metabolised to arachidonic acid and ethanolamine by fatty acid amide hydrolase, while 2-AG is converted to arachidonic acid and glycerol by monoacylglycerol lipase.

Because of its ether linkage, noladin cannot be a substrate for either of these enzymes. However the review also describes other possible pathways of endocannabinoid metabolism involving oxidation by cyclooxygenases, lipoxygenases and P450 cytochromes. These are the best candidates for the further metabolism of noladin, but there is apparently no direct evidence for this.

Which Enzymes are Responsible for the Biodegradation of Noladin Ether? - Biology

Pyrethroids and their metabolite 3-phenoxybenzoic acid are of great concern.

Application of microbial food cultures in feed and food during postharvest process.

Cleavage of pyrethroids ester linkage and diphenyl ether bond are discussed.

Key enzyme genes are outlined along with genetic engineering methodologies.

Genetically modified microbes exhibit mineralization potential for pyrethroids.

Study of Biological Degradation of New Poly(Ether-Urethane-Urea)s Containing Cyclopeptide Moiety and PEG by Bacillus amyloliquefaciens Isolated from Soil

The present work for the first time investigates the effect of Bacillus amyloliquefaciens, M3, on a new poly(ether-urethane-urea) (PEUU). PEUU was synthesized via reaction of 4,4′-methylenebis(4-phenylisocyanate) (MDI), l -leucine anhydride cyclopeptide (LACP) as a degradable monomer and polyethylene glycol with molecular weight of 1000 (PEG-1000). Biodegradation of the synthesized PEUU as the only source for carbon and nitrogen for M3 was studied. The co-metabolism biodegradation of the polymer by this organism was also investigated by adding mannitol or nutrient broth to the basic media. Biodegradation of the synthesized polymer was followed by SEM, FT-IR, TGA, and XRD techniques. It was shown that incubation of PEUU with M3 resulted in a 30–44 % reduction in polymer’s weight after 1 month. This study indicates that the chemical structure of PEUU significantly changes after exposure to M3 due to hydrolytic and enzymatic degradation of polymer chains. The results of this work supports the idea that this poly(ether-urethane) is used as a sole carbon source by M3 and this bacterium has a good capability for degradation of poly(ether-urethane)s.

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Microbial degradation of recalcitrant pesticides: a review

Some pesticides such as organochlorines are of critical environmental concern because they are highly persistent due to their stable chemical nature. As a consequence, even after banning, dichlorodiphenyltrichloroethane and endosulfan can be detected at concentrations above permissible limits. Moreover, classical pesticide degradation of these compounds using physiochemical processes is limited. Alternatively, biodegradation using microorganisms isolated in contaminated sites appears promising. For instance, the bacterium Pseudomonas fluorescens degrades aldrin by 94.8%, and the fungus Ganoderma lucidum can bring down the levels of lindane by 75.5%. In addition, the toxicity is reduced by enzymes that perform oxidation, reduction, hydrolysis, dehydrogenation, dehalogenation and decarboxylation. Then, the metabolites are further degraded by mineralisation and cometabolism. The biodegradation process can be manipulated by applying techniques such as bioattenuation, bioaugmentation and biostimulation. This article discusses the latest advances in microbial degradation of recalcitrant pesticides.

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Biotransformation and biodegradation

Anaerobic biodegradation of TCP

Few studies have been done aimed at establishing the possibilities for biodegradation and bioremediation of TCP. Growth-supporting biodegradation of halogenated compounds is generally based on one of the following processes: (1) chemotrophy with an oxidizable electron donor (hydrogen, lactate) and use of the halogenated compounds as a physiological electron acceptor (anaerobic conditions) (2) use of the halogenated compound as a carbon and energy source with an external electron acceptor (oxygen, nitrate) (3) fermentative metabolism, in which the halogenated compound serves both as electron donor and (indirectly) as electron acceptor. Biotransformation processes not linked to growth may also be important. Such cometabolic transformations are due to the broad substrate spectrum of many microbial enzymes, the general reactivity of cofactors, or the formation of reactive intermediates in the catalytic cycle of some enzymes. Examples are reductive dechlorination by cobalamin cofactors of anaerobic bacteria, and oxidative transformation by broad-specificity metal-containing monooxygenases of aerobic bacteria. Obviously, a biodegradation process that stimulates growth of the active organisms is preferable in a bioremediation situation since it allows adaptation at the population level, leading to an increase of the amount of active biomass during the treatment process.

Reductive biotic transformation of TCP has been demonstrated under anaerobic conditions (Löffler et al. 1997 Peijnenburg et al. 1998 Hauck and Hegemann 2000) (Fig. 3) and sequential aerobic–anaerobic conditions (Long et al. 1993). Reductive dechlorination of both TCP and chloroethanes was observed with an enrichment culture that dechlorinated 1,2-dichloropropane, and propene and 1,2-dichloropropane were detected as products (Löffler et al. 1997). Experiments in which various halogenated aliphatic compounds were incubated with anaerobic sediments indicated zero-order conversion kinetics for TCP and dichloromethane, whereas most other organohalogens were transformed according to first order kinetics (Peijnenburg et al. 1998).

Anaerobic biotransformations of TCP. Both reductive dehalogenation (RD) and dihaloelimination reactions (DHE) are observed. Formation of allylchloride may occur by dihaloelimination (Yan et al. 2009) or possibly via 1,3-dichloropropene (broken arrows)

Recently, two strains (BL-DC-8 and BL-DC-9) of an anaerobic Gram-negative bacterium were isolated from contaminated groundwater at a Superfund site located near Baton Rouge, and characterized as belonging to the new species Dehalogenimonas lykanthroporepellens (Moe et al. 2009 Yan et al. 2009). These bacteria utilize TCP as an electron acceptor under anaerobic conditions but not chlorinated alkenes. Hydrogen was the electron donor. For both strains, allyl chloride was detected as the main initial dechlorination product (Fig. 3). However, allyl chloride is unstable and is hydrolyzed abiotically to allyl alcohol, whereas in the presence of cysteine or sulfide, allyl chloride was transformed to allyl mercaptan, S-allyl mercaptocysteine and allyl sulfides. The mechanism of the reductive dechlorination reaction is not completely clear, as the enzymes responsible for TCP dechlorination have not yet been isolated and characterized. In freshwater environments, transformation of TCP into allyl chloride followed by the formation of allyl alcohol could be toxic to fish and aquatic life (Ewell et al. 1986).

Bioremediation of contaminated groundwater through in situ reductive dechlorination can be performed by injecting a compound such as hydrogen, lactic acid or another oxidizable organic substrate that is used by microorganisms to produce hydrogen, which induces reductive dechlorination and serves as electron donor (Tratnyek 2008). At a site in California, 99.9 % reduction of TCP contamination has been found over a period of 1,000 days. However, biotic dechlorination through hydrogen-releasing compounds may be applicable only at low concentrations, such as less than 1 mg TCP/l (CH2M HILL 2005).

Aerobic cometabolic conversion

Various halogenated aliphatic hydrocarbons can be transformed in a cometabolic manner by broad-specificity monooxygenase involved in hydrocarbon degradation, such as methane monooxygenase (Hanson et al. 1990 Oldenhuis et al. 1989). The soluble methane monooxygenase produced by cells of the methanotrophic bacterium Methylosinus trichosporium OB3b can convert TCP, giving rise to dichloropropanols after subsequent reduction (Bosma and Janssen 1998 Fig. 4). However, TCP is a poor substrate for the enzyme as compared to other pollutants such as trichloroethylene. The conversions are analogous to those catalyzed by cytochrome P450 in mammalian systems (Fig. 1). The major drawback of such cometabolic conversions is product toxicity. In case of TCP conversion by methane monooxygenase, the insertion of oxygen preferentially occurs on the terminal carbon atom, which yields chlorinated carbonyl compounds that may undergo elimination to produce 2-chloroacrolein, a very reactive compound.

Conversions of 1,2,3-trichloropropane initiated by methane monooxygenase (MMO) produced by M. trichosporium OB3b cells. Reduction to alcohols is caused by alcohol dehydrogenase activity (DH)

The aerobic conversion of TCP reported by Leahy et al. (2003) using a mixture of hydrocarbon-degrading bacteria is probably based on similar reactions, but the products were not identified. Aromatic hydrocarbon-degrading bacteria produce monooxygenases that are capable of chlorinated hydrocarbon degradation through similar oxidative reactions as the methane monooxygenase of methanotrophs. For example, the toluene monooxygenase of a pseudomonad was described to convert chlorinated hydrocarbons (Newman and Wackett 1997).

Recalcitrant behavior of TCP towards growth-supporting aerobic biotransformation

Microbial transformation of TCP to CO2, H2O and HCl by oxidative metabolism with oxygen as an electron acceptor and by reduction to lesser chlorinated propanes and HCl is thermodynamically possible (Dolfing and Janssen 1994). However, no aerobic organisms, enrichment cultures, or bioreactors have been described that demonstrate the use of TCP as a growth-supporting oxidizable substrate. Various attempts to enrich TCP-degrading microorganisms from environmental samples, including from sites with a long history of TCP or epichlorohydrin pollution, or to obtain TCP degradation in continuous-flow columns inoculated with samples from contaminated sites, have failed (Bosma et al. 2002). This indicated that TCP is indeed a very recalcitrant compound and nature has not yet evolved aerobic organism that are adapted to it. The fact that the thermodynamic calculations indicate that aerobic oxidation of TCP is energetically favorable suggests biochemical hurdles instead of another fundamental reason as the cause of the apparent recalcitrance of TCP.

An example of such a biochemical hurdle is toxicity of intermediates. In case of halogenated aliphatic compounds, several reactive intermediates occur along the metabolic pathways, requiring optimization of fluxes to prevent accumulation of such reactive intermediates to toxic levels (van Hylckama Vlieg and Janssen 2001). It also may be due to formation of dead-end side products that are toxic. Formation of such reactive intermediates will act against evolutionary selection of more efficient initial enzymes for TCP metabolism. The recalcitrant nature of a non-natural compound might also be due to presence of structural elements that cannot be recognized and converted by microbial enzymes, which evolved for the conversion of natural compounds (Rieger et al. 2002).

When inspecting the possible pathways for productive aerobic metabolism of TCP, hydrolysis of a carbon–halogen bond as the first step seems the most attractive reaction, because it does not involve reactive intermediates and leads to 1,3-dichloro-2-propanol 2,3-dichloro-1-propanol. These compounds are known to be biodegradable and pure cultures capable of using dichloropropanols for growth under aerobic conditions are known (Effendi et al. 2000 Higgins et al. 2005 van den Wijngaard et al. 1989 Yonetani et al. 2004).

Hydrolysis of carbon–halogen bonds in chlorinated compounds is carried out by a diversity of microbial enzymes called dehalogenases. These belong to different phylogenetic classes, of which the haloalkane dehalogenases that are members of the α/β-hydrolase fold superfamily of proteins are the best characterized (Janssen 2004 Koudelakova et al. 2011). Another prominent class is the HAD-superfamily of haloacid dehalogenases and phosphatases, with dehalogenases that act on 2-chloroacetate and 2-chloropropionate. Haloalkane dehalogenases are known to convert compounds such as 1,2-dichloroethane, 1,2-dibromoethane, 1,3-dichloropropane, 1,2-dichloropropene, and (slowly) hexachlorocyclohexane (Janssen et al. 2005 Poelarends et al. 1999, 2000 Koudelakova et al. 2011). The conversion of TCP by a haloalkane dehalogenase was first described by Yokota et al. (1987) using an enzyme from Corynebacterium strain m15-3, but the activity was very low (k cat/K m = 36 s −1 M −1 ) (Bosma et al. 1999). Sequence analysis and structural studies identified the protein (which is commonly called DhaA) as a member of the α/β-hydrolase fold family. Another dehalogenase that has a low activity with TCP is LinB, and enzyme originally discovered in bacteria that degrade hexachlorocyclohexane (Monincová et al. 2007).

The first DhaA gene sequence was described by Kulakova et al. (1997) in the 1-chlorobutane degrader Rhodococcus rhodochrous NCIMB13064. Poelarends et al. (2000) found that the same gene is geographically widely distributed by using PCR analysis and dehalogenase gene sequencing of different bacteria enriched with other haloalkanes, including 1,3-dichloropropene. Comparison of the genetic organization in different organisms revealed that the haloalkane dehalogenase gene likely originates from Rhodococcus strains, where it is present in an operon together with an alcohol dehydrogenase and an aldehyde dehydrogenase gene, as well as a regulatory gene that influences gene expression. The latter may act as a repressor in the absence of a halocarbon substrate (like 1-chlorobutane). When the dehalogenase gene regions from a 1,2-dibromoethane degrading Mycobacterium and a 1,3-dichloropropene dehalogenating Pseudomonas were examined, it appeared that the repressor gene was absent or inactivated by mutations to allow production of the enzyme in the presence of these new, non-inducing substrates (Poelarends et al. 1999, 2000). In the absence of a functional regulatory gene, inactivation of the repressor causing constitutive expression of a dehalogenase appears a way to allow genetic adaptation and biodegradation.

Lack of microbial growth on TCP and lack of adaptation in column or enrichment experiments is most likely due to the very rare occurrence of a haloalkane dehalogenase gene with a suitable activity in a host organism that is capable of dichloropropanol conversion. Mutations in the haloalkane dehalogenase that would lead to an enhanced substrate range that includes TCP would be unlikely to propagate in an organism that does not grow on the hydrolysis product and thereby provide a selective growth advantage. When DNA sequence databases, both of completed bacterial genomes and environmental sequences, are searched for genes that encode the DhaA-type haloalkane dehalogenase, or the haloalcohol dehalogenases known to be involved in 2,3-dichloro-1-propanol metabolism (except in organisms isolated on these compounds), no hits are found. These genes seem very rare and can only be recovered by appropriate enrichment culture techniques starting with polluted environmental samples.

The evolution of bacteria that have the capacity to degrade TCP aerobically is thus restricted by the selectivity of haloalkane dehalogenases, and the rare occurrence of bacteria growing on dichloropropanols (Fig. 5). Consequently, attempts were made to obtain organisms capable of TCP detoxification by a combination of protein engineering and heterologous gene expression (Bosma et al. 1999, 2002).

Comparison of catabolic pathways for 1,2-dichloroethane (DCE) and TCP. DCE bioremediation has been established at full scale, using bacterial cultures that use DCE as carbon source for growth according to the pathway that is shown (a). It starts with hydrolytic dehalogenation catalyzed by a haloalkane dehalogenase (DhaA). TCP is much more recalcitrant, but productive catabolic pathways can be envisaged (b). The upper routes could proceed from 2-chloroacrylic acid either via dehydrogenation (DH) (Kurata et al. 2005) or dechlorination (Dhl) (Mowafy et al. 2010). The lower route is thought to proceed in the strain constructed by Bosma et al. (2002) in A. radiobacter AD1 expressing a mutants haloalkane dehalogenase (DhaAM2) and involves dehalogenases (Hhe) and epoxide hydrolase (EH)

Engineering enzymes and organisms for TCP conversion

Different reports on the engineering of haloalkane dehalogenase variants with enhanced activity towards TCP have been published. By using error prone PCR and DNA shuffling, Bosma et al. (2002) generated a DhaA mutant (e.g., a variant called DhaAM2 with the mutations C176Y and Y273F) that had three times higher catalytic efficiency (k cat/K m = 280 s −1 M −1 ) than wild-type enzyme. Similarly, Gray et al. (2001) performed in vitro evolution studies which also yielded a mutant with a substitution at position 176 and a mutation close to the N terminus that showed higher activity with TCP as compared to wild-type, and further mutations enhanced the stability of the enzymes.

The strategy to construct a recombinant TCP-degrading strain was based on the use of an improved haloalkane dehalogenase into an organism that grows on the product of hydrolytic dehalogenation, which is 2,3-dichloro-1-propanol. For this, a host was used that could degrade both 2,3-dichloropropanol and 1,3-dichloropropanol: Agrobacterium radiobacter AD1 (van den Wijngaard et al. 1989). First, the wild-type haloalkane dehalogenase gene for DhaA from Rhodococcus was placed under control of a strong constitutive promoter and cloned on a broad host range plasmid (pLAFR3) that was introduced into strain AD1 (Bosma et al. 1999). Growth of the resulting strain was not significant, but after incubation of 25 days 0.7 mM of TCP was converted and a small increase of biomass was observed. The strain did utilize 1,2,3-tribromopropane and 1,2-dibromo-3-chloropropane as sole carbon source, showing for the first time growth on a trihalopropane.

Growth on TCP could be obtained when a DhaA-type dehalogenase with improved activity for TCP was used. The dhaAM2 gene for the improved dehalogenase was constitutively expressed in strain AD1. The resulting strain, A. radiobacter AD1(pTB3-M2), was able to utilize TCP as carbon and energy source under aerobic conditions. After 10 days, 3.6 mM TCP was converted by a culture initially inoculated to an OD450 of 0.14. Due to production of hydrochloric acid, the pH dropped to 6.0 (Bosma et al. 2002).

The construction of a recombinant strain using an improved haloalkane dehalogenase that was expressed under a strong constitutive promoter in a host that degrades a dichloropropanol, is an important step towards obtaining an organism that is suitable for TCP bioremediation under aerobic conditions. However, the system has still limitations and drawbacks (Bosma et al. 2002): (1) Although the initial dehalogenase is significantly improved for TCP conversion (ca. 5-fold as compared to wild-type), the activity of DhaAM2 is still too low to rapidly transform TCP. Consequently, the estimated doubling time of the constructed strain was 90 h, which, for comparison, is much slower than the ca. 10 h measured for the 1,2-dichloroethane-degrader Xanthobacter autotrophicus strain that is used for full-scale groundwater bioremediation. (2) Degradation of TCP was incomplete due to the enantioselective conversion of only the (R)-2,3-dichloropropanol by the host A. radiobacter AD1. The DhaAM2 dehalogenase produced a racemic mixture of (R)- and (S)-2,3-dichloropropanol from TCP. (3) The modified dehalogenase gene for DhaAM2 was introduced into strain AD1 using the cloning vector pLAFR3, which is a transmissible plasmid. Such a plasmid may be modified or lost under stress conditions, or it may be transferred to other bacteria. (4) Application of specialized bacteria in bioremediation operations will likely make use of open systems, such as an immobilized-cells bioreactor from which organisms may detach and end up in effluent water. This may lead to spread of resistance genes (in this case tetracycline) if the engineered organism contains additional antibiotic resistance markers. To remedy these limitations, further improvements are under investigation.


The catabolic potential of naturally occurring organisms towards organic compounds is the result of long evolution processes, whereas the time in which organisms have been tempted to evolve new enzymes, pathways and regulatory mechanisms that allow conversion of xenobiotic industrial chemicals is quite short. The industrial synthesis of compounds such as trichloropropane only started in the first half of the 20th century. Nevertheless, the presence of these synthetic compounds in the biosphere has already triggered the evolution of new metabolic activities, as illustrated by various examples (Janssen 2004 Janssen et al. 2005 Paul et al. 2005).

An important example of bacteria capable of TCP degradation are the strictly anaerobic strains BL-DC-8 and BL-DC-9 of D. lykanthropropepellens, isolated from contaminated groundwater in the USA (Yan et al. 2009). The net dihaloelimination reaction catalyzed by these organisms implies transfer of electrons to TCP, with chloride release. This suggests the possibility of reductive dehalogenation coupled to electron transfer from hydrogen or another electron donor to TCP (dehalorespiration Smidt and de Vos 2004). Since this process could possibly stimulate growth, as indicated by an increase in cell numbers (Yan et al. 2009), genetic- or population-level adaptation of cultures to TCP under anaerobic conditions can be envisaged. This may yield faster growing cultures than those currently described (maximum specific growth rate 0.15–0.17 day −1 ). It would also be highly interesting to identify the genes, proteins and cofactors involved in anaerobic conversion of TCP to allyl chloride and to establish their possible association with energy metabolism. The biochemical basis of dihaloelimination reactions is currently not well understood, although they may be important for different chlorinated substrates (de Wildeman et al. 2003 Smidt and de Vos 2004). For in situ bioremediation, anaerobic transformation may be more attractive than aerobic processes, due to the difficulty of homogeneous oxygen supply and its preferred use for other oxidative processes if TCP is a low-level contaminant.

Anaerobic degradation of TCP was described to produce next to allyl chloride also small amounts of further conjugation products (diallyl sulfide, allyl mercaptan), probably due to abiotic reactions with sulfide (Yan et al. 2009). The chemically labile carbon–halogen bond in allyl chloride, as well as its sensitivity to cleavage by hydrolytic dehalogenases, suggest that more rapid biodegradation of allyl chloride with reduced formation of sulfur conjugates can be achieved when adapted mixed cultures are used. Thus, further studies on the anaerobic metabolism of TCP and allyl chloride, in combination with appropriate enrichment and adaptation strategies, may well lead to more rapid anaerobic degradation as compared to what is currently possible.

Regarding aerobic degradation of TCP, genetic engineering can contribute to the acquisition of new bioremediation organisms, as illustrated by Bosma et al. (2002). To further enhance the biodegradation of TCP, use of a better haloalkane dehalogenase is desirable. By using rational design and directed evolution, the activity of DhaA against TCP was recently improved by Pavlova et al. (2009). Tunnel residues leading to the active site of DhaA were selected as target spots for mutagenesis, based on the notion that substrate binding and/or product release may limit the rate of catalysis. The best variants that were obtained carried three new mutations as compared to variant DhaAM2, and had 36 times higher activity (kcat) than the natural enzyme towards TCP (Table 1). In the degradation pathway of 1,2-dichloroethane (DCE) by X. autotrophicus GJ10, the first step is catalyzed by DhlA, which is a phylogenetically related haloalkane dehalogenase. Since this organism was successfully used for groundwater cleanup at full scale (Stucki and Thüer 1995), it is interesting to compare the catalytic rates of the initial haloalkane dehalogenases (Table 1). The differences in Table 1 are important since kinetic properties and expression levels of the dehalogenases have a major impact on the kinetic properties of chloroalkane degradation (substrate affinity, growth rate) by the host organism (van den Wijngaard et al. 1993). Even though the activity of DhaA31 is significantly improved by directed evolution, the k cat and k cat/K m values of DhaA31 for TCP are still lower than the corresponding values of DhlA for DCE (Table 1). Thus, an engineered organism expressing the evolved DhaA31 will still have a lower affinity for TCP than strain GJ10 for DCE. It is well possible that further variants of haloalkane dehalogenases that convert TCP even better can be obtained. Strategies for laboratory evolution of new enzyme activities are still improving, and recently we were able to obtain complementary TCP dehalogenating mutants that produce almost enantiopure (R)- or (S)-2,3-dichloro-1-propanol. Although dehalogenase enantioselectivity may be unimportant for groundwater and soil bioremediation, it holds great promise for converting TCP waste to economically valuable chiral building blocks for use in the fine chemicals and pharmaceutical industries (van Leeuwen et al. 2012).

Improved conversion of 2,3-dichloropropanol by a better host is under investigation with new isolates that were obtained from a site contaminated with epichlorohydrin and chloropropanols due to leakage of waste from epichlorohydrin manufacture. This organism, a strain of Pseudomonas putida, uses a pathway for 2,3-dichloropropanol degradation that is different from the route detected in Agrobacterium strains (Higgins et al. 2005 van den Wijngaard et al. 1989) and lacks enantioselectivity. However, none of the current dichloropropanol degraders has been selected on the basis of its potential to form a biofilm on a solid support under groundwater flow conditions, and in competition with other bacteria. Furthermore, substrate supply will likely be low, which also may impose physiological requirements on the host organism.

The use of plasmid-based systems, as in the A. radiobacter AD1(pTB3-M2) recombinant (Bosma et al. 2002) is undesirable for the construction of bioremediation organisms, especially when in situ remediation is targeted (de Lorenzo and Timmis 1994 Timmis and Pieper 1999). A recombinant organism applied in situ should be capable of establishing itself an environment where the conditions cannot be controlled (de Lorenzo 2009). This may cause stress, leading to plasmid loss or lysis, as well as to spread of recombinant DNA. The presence of antibiotic resistance-based selection markers and the use of transmissible plasmids can be avoided by employing chromosomal integration, for which efficient transposon-based systems were developed. For example, a modified Tn5 transposon system can be used to integrate a foreign gene into the chromosome, leading to stable integration (de Lorenzo and Timmis 1994). Such cloning vectors have been used successfully to construct strains for environmental applications (Panke et al. 1998).

If an efficient pathway can be assembled or evolved in the laboratory, in a robust host organism that can maintain itself under practical conditions, the prospects of successful application of such a genetically engineered organism for bioremediation are good. The limited success that has been achieved so far in this area, is mainly due to the fact that few recombinant organism have been engineered to degrade compounds which are really recalcitrant and where the poor degradability is due to biochemical factors instead of low solubility, limiting oxygen supply, poor bioavailability, etc. On the other hand, evolution of dehalogenases also occurs in natural environments (Janssen 2004), and it is well possible that at some day, due to continued evolutionary pressure, TCP becomes a degradable compound and that TCP-degrading organisms can be obtained by classical enrichment.

Degradation of polyethylene by Trichoderma harzianum—SEM, FTIR, and NMR analyses

Trichoderma harzianum was isolated from local dumpsites of Shivamogga District for use in the biodegradation of polyethylene. Soil sample of that dumpsite was used for isolation of T. harzianum. Degradation was carried out using autoclaved, UV-treated, and surface-sterilized polyethylene. Degradation was monitored by observing weight loss and changes in physical structure by scanning electron microscopy, Fourier transform infrared spectroscopy, and nuclear magnetic resonance spectroscopy. T. harzianum was able to degrade treated polyethylene (40 %) more efficiently than autoclaved (23 %) and surface-sterilized polyethylene (13 %). Enzymes responsible for polyethylene degradation were screened from T. harzianum and were identified as laccase and manganese peroxidase. These enzymes were produced in large amount, and their activity was calculated using spectrophotometric method and crude extraction of enzymes was carried out. Molecular weight of laccase was determined as 88 kDa and that of manganese peroxidase was 55 kDa. The capacity of crude enzymes to degrade polyethylene was also determined. By observing these results, we can conclude that this organism may act as solution for the problem caused by polyethylene in nature.

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Biosorption and degradation of decabromodiphenyl ether by Brevibacillus brevis and the influence of decabromodiphenyl ether on cellular metabolic responses

There is global concern about the effects of decabromodiphenyl ether (BDE209) on environmental and public health. The molecular properties, biosorption, degradation, accumulation, and cellular metabolic effects of BDE209 were investigated in this study to identify the mechanisms involved in the aerobic biodegradation of BDE209. BDE209 is initially absorbed by wall teichoic acid and N-acetylglucosamine side chains in peptidoglycan, and then, BDE209 is transported and debrominated through three pathways, giving tri-, hepta-, octa-, and nona-bromodiphenyl ethers. The C–C bond energies decrease as the number of bromine atoms on the diphenyl decreases. Polybrominated diphenyl ethers (PBDEs) inhibit protein expression or accelerate protein degradation and increase membrane permeability and the release of Cl − , Na + , NH4 + , arabinose, proteins, acetic acid, and oxalic acid. However, PBDEs increase the amounts of K + , Mg 2+ , PO4 3− , SO4 2− , and NO3 − assimilated. The biosorption, degradation, accumulation, and removal efficiencies when Brevibacillus brevis (1 g L −1 ) was exposed to BDE209 (0.5 mg L −1 ) for 7 days were 7.4, 69.5, 16.3, and 94.6 %, respectively.

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Rhodococcus sp. strain B1 could degrade 100 mg/L butachlor within 5 days. Butachlor was first hydrolyzed by strain B1 through N-dealkylation, which resulted in the production of butoxymethanol and 2-chloro-N-(2,6-dimethylphenyl)acetamide. Butoxymethanol could be further degraded and utilized as the carbon source for the growth of strain B1, whereas 2-chloro-N-(2,6-dimethylphenyl)acetamide could not be degraded further. The hydrolase designated ChlH, responsible for the N-dealkylation of the side chain of butachlor, was purified 185.1-fold to homogeneity with 16.1% recovery. The optimal pH and temperature of ChlH were observed to be 7.0–7.5 and 30 °C, respectively. This enzyme was also able to catalyze the N-dealkylation of other chloroacetamide herbicides the catalytic efficiency followed the order alachlor > acetochlor >butachlor > pretilachlor, which indicated that the alkyl chain length influenced the N-dealkylation of the chloroacetamide herbicides. This is the first report on the biodegradation of chloroacetamide herbicides at the enzyme level.

Metabolic repertoire of Sphingomonas wittichii strain RW1

Substantial research has been done on the metabolic functionalities of Sphingomonas wittichii strain RW1 with respect to degradation of array of dioxin congeners. This strain is considered to be the best characterized among organisms with dioxin degradation capability (Nojiri and Omori 2002). It was isolated from Elbe River (Germany) through enrichment culture technique and grows within the mesophilic temperature range (30 °C) (Wittich et al. 1992). It belongs to the Alphaproteobacteria class which assimilates l -arabinose, d -mannitol, maltose, and phenyl-acetic acid. The microorganism grew on both DD and DF as the sole source of carbon and energy. Wilkes et al. (1996) showed the ability of RW1 to oxidize some PCDD/F congeners. The bacterium degraded quite a number of mono- and dichlorinated DFs and DDs, but failed to degrade the highly chlorinated analogues. 2-chlorodibenzofuran (2-CDF) and 3-CDF were degraded to their corresponding chlorinated salicylic acids which is an evidence of dioxygenation of the aromatic nuclei at the angular position. However, in the case of 4-CDF and 2,3-dichlorodibenzofuran (2,3-DCDF) used as substrate for resting cells, metabolites 3-chloro- and 4,5-dichlorosalicylic acids, respectively, were identified. The non-detection of salicylic acid during 4-CDF incubation and accumulation of 3-chlorosalicylic acid in the culture fluid suggested the lack of dioxygenation of the substituted aromatic ring. 4,5-dichlorosalicylic acid recovery was also nearly stoichiometric thus indicating that the growth of the organism occurred at the expense of the non-substituted ring. The absence of dioxygenation on the chlorinated ring of 2,3-DCDF may be due to steric hindrance or electronic influences by the chlorine substituents on the same aromatic ring. Alternatively, salicylic acid may have been entirely consumed by the resting cells, consequently preventing the detection of this metabolite. Also, no metabolites were detected during the incubation of RW1 resting cells with 3,7-DCDF and 2,4,8-trichlorodibenzofuran (2,4,8-TrCDF) in the culture supernatants. It remained unclear whether the recalcitrance was due to electronic effects or steric hindrance caused by the heftiness of the chlorine substituents. Similar results were obtained when the resting cells pre-grown on DF were incubated with the PCDD congeners. In a related study by Hong et al. (2002), DF-grown resting cells of RW1 was used to aerobically transform 2,7-DCDD and 1,2,3,4-TCDD. Forty-seven percent of 2,7-DCDD was biotransformed to 4-chlorocatechol after 120 h incubation. During 1,2,3,4-TCDD incubation however, 37% of the congener was transformed. Examination of the entire culture fluid showed two metabolites (3,4,5,6-tetrachlorocatechol and 2-methoxy-3,4,5,6-tetrachlorophenol) accumulated. About 37% of 1,2,3,4-TCDD had been transformed to 3,4,5,6-tetrachlorocatechol, while 6% was converted to 2-methoxy-3,4,5,6-tetrachlorophenol.

In spite of the pronounced recalcitrance of highly chlorinated congeners of dioxin, Nam et al. (2006) unequivocally demonstrated for the first time, transformation of 1,2,3,7,8-PeCDD and 1,2,3,4,7,8-hexachlorodibenzo-p-dioxin (1,2,3,4,7,8-HeCDD) by S. wittichii strain RW1. Interestingly, both compounds are not only ubiquitous pollutants, they are known to have long half-lives when compared to lightly substituted analogues (McLachlan et al. 1996). According to Nam et al. (2006), nearly 100% of the initially supplied 1,2,3,4,7,8-HeCDD was transformed through 2,2′,3-trihydroxy-hexachlorodiphenyl ether (produced in trace amounts) to 3,4,5,6-tetrachlorocatechol and 2-methoxy-3,4,5,6-tetrachlorophenol as major metabolites. Quite surprisingly, 1,2,3,7,8-PeCDD which is less chlorinated was not transformed to any detectable extent. Similarly too, 1,2,3-TrCDD was transformed through 2,2′,3-trihydroxy-1,2,3-trichlorodiphenyl ether to 3,4,5-trichlorocatechol, whereas, 2,3,7-trichlorodibenzo-p-dioxin (2,3,7-TrCDD) was not metabolized. The dynamics of dioxin transformation in strain RW1 readily suggests that chlorine substitution patterns on the dioxin ring influences biodegradability rather than the overall number of chlorine substituents. Identification of the metabolites extracted from the culture fluid further implied a dioxygenase attack on the less substituted aromatic ring, as depicted in Fig. 5.

Pathways for biotransformation of 1,2,3-trichlorodibenzo-p-dioxin, A and 1,2,3,4,7,8-hexachlorodibenzo-p-dioxin, B by Sphingomonas wittichii RW1 (Nam et al. 2006). The compounds in parentheses are very unstable and undergo spontaneous transformation to their respective hydroxylated metabolites. Dibenzo-p-dioxin-1,10a-dioxygenase (1), 1,10a-dihydroxy-1-hydrodibenzo-p-dioxin dehydrogenase (2), 2,2',3-trihydroxydiphenyl ether dioxygenase (3), 2-hydroxy-6-oxo-6-(2-hydroxyphenoxy)-hexa-2,4-dienoic acid hydrolase (4)

In a different study, Bunz and Cook (1993) purified and characterized angular dioxygenase enzyme system from strain RW1 and called it a three-component ring hydroxylating dioxygenase, designated DF 4,4a-dioxygenase system. The enzyme needs the involvement of a flavoprotein, reductase A2, and an iron–sulfur protein to which is involved in the movement of electrons from NADH to the dioxygenase for activation of oxygen which consists of terminal oxidase dxnA1, a reductase (RedA2), and a ferredoxin as revealed in Fig. 4.

Genes coding for each component in the catabolic routes have been cloned, sequenced, and expressed (Armengaud et al. 1998). Gene-encoding enzymes involved in DD and DF degradation (dxnA1, dxnA2, fdx1, and redA2) are dispersed over the entire chromosome. Likewise, genes responsible for the upper pathway have also been elucidated, cloned, and expressed in E. coli. The authors suggested that the genes are likely participants in oxidation of DF and DD.

The complete genome sequencing of RW1 gave in-depth view on improved studies of its peculiar metabolic competence and the precise locations of the gene sequences of interest. The RW1 genome comprises of two mega plasmids and one chromosome (Miller et al. 2010). The RW1 genome has multiple TonB-dependent outer membrane receptor protein genes which reside within the operons and was suspected to code for enzymes taking part in aromatic hydrocarbon catabolism.

Strain RW1 response to changes in water stress conditions through evolving adaptation which is important for successful application in soil bioremediation strategy was a subject of investigation through the genome-wide expression of the organism in response to induced water stress condition. In addition, the bacterium was perturbed with either non-permeating solute or cell-permeating solute. The responses to the different stress conditions were assessed and compared using membrane fatty acid analyses, growth assays, and transcriptome profiling. The results showed that both the non-permeating and permeating solutes triggered divergent adaptive responses, indicating that the solutes affect cells in major disparate ways (Johnson et al. 2011). Cellular responses of RW1 during DF metabolism were monitored during gene transcription (genome-wide level). Several unessential and related aromatic pathways were proposed as well as high-expression level of non-catabolic genes during the initial exposure to DF (stressor), However, with continuous exposure, DF was perceived as a carbon source and metabolized through various pathways in parallel (Coronado et al. 2012).

In a related study, a genome-wide transposon scanning of the bacterium was investigated to have a perception on the survival ability of RW1 in a stressed environment (Roggo et al. 2013). It also identified putative functions necessary for survival in soil during drought condition. To this extent, RW1 transposon libraries were generated and grown in liquid medium-containing salicylic acid as exclusive carbon and energy source either in the presence or absence of salt stress conditions. On the other hand, libraries were grown in moist sand with salicylic acid. On analysis, no transposon reads were recovered in 10% of the annotated genes of RW1 genome amounting to a total of 579 genes in any of the libraries under any of the growth conditions, thus, indicating those to be important for survival under the stressed conditions. Fatty acid catabolism and oxidative stress response are important for long-term survival of cells in sand. The transcriptome data indicated salient roles in flagellar activity, pili synthesis, trehalose, and polysaccharide synthesis, and reputed cell surface antigen proteins when grown under stress condition in soil. This demonstrated the competence of genome-wide transposon scanning methodologies for evaluation of multiplex traits (Roggo et al. 2013).

The first investigation on protein analysis of RW1 during doxin metabolism was reported by Colquhoun et al. (2012). Some proteins linked to DD/DF degradation were upregulated and cellular stress was increased during DF metabolism. Dioxin dioxygenase was also expressed during growth in both acetate and DF. In another study, 502 proteins were detected when RW1 was grown on DD, DF, and 2-CDD during protein profiling (shotgun proteomics) when compared with growth on acetate. Previous roles of DxnA1A2, DbfB, and DxnB were established and confirmed Swit_3046 dioxygenase and DxnB2 hydrolase role in dioxin degradation (Hartmann and Armengaud 2014).

RNA-Seq (whole transcriptome shotgun sequencing) was also used to profile transcriptional responses of RW1 to DD and DF, while growth in succinic acid served as control. Under these conditions, the protein-coding genes that were expressed differentially in DF were more than 240, while over 300 were expressed in the presence of DD. Stress response genes were upregulated in response to both DD and DF with higher effect in the former than the latter, indicating a higher toxicity of DD when compared with DF (Chai et al. 2016).

Factors determining the biodegradative capacity and environmental behaviour of strain RW1, were performed under near natural conditions (non-sterile DF-contaminated soil) and compared with laboratory culture conditions (RW1 liquid culture in DF) (Moreno-Forero and van der Meer 2015). Reactions were deduced under different growth conditions from analyzing recorded genome-wide expression profiles. RW1 showed stationary phase characteristics, which was evidence of stress, nutrient hunting, and primary metabolism adjustment, if they were not pre-grown in liquid medium with pollutant as established in the soil. Cells multiplying and thriving in sand degraded DF, but displayed a disparate transcriptome indication in liquid culture exposed to drought stress, as well as indication of several ‘soil-specific’ expressed genes was also deduced. Suggestions were made on inoculation efficacy by testing behaviour of strains under conditions as close to the intended microbiome conditions.


Ametoctradin Biodegradation in an ex vivo Soil System

Soil samples were collected from four different locations previously utilized for ametoctradin environmental fate testing and regularly used for agricultural studies: California (CA), New Jersey (NJ), and Germany (LUFA 2.2 and LUFA 2.3) (Table 1). Roughly 1 kg of each soil sample was placed in two replicate containers linked to a CO2-scrubbed air source. Each sample was then either treated with 2.4 μg mL 𠄱 of ametoctradin or remained untreated. After 14 days, the soil samples were harvested and analyzed in triplicate using HPLC-MS/MS. In each case, degradation of ametoctradin was confirmed along with the identification of four major metabolites: M650F01, M650F02, M650F03, and M650F04 (Figure 1). These compounds appeared to be the product of degradation by soil microorganisms capable of metabolizing the long aliphatic chain on the ametoctradin molecule while the amine ring structures remain largely unaltered. The NJ soil sample displayed the highest and most complete level of ametoctradin degradation, with only 13.6% of the original amount of agrochemical remaining after the 14-day period. The M650F03 metabolite appeared to be the main degradation product in every soil type.