Information

Which Enzymes are Responsible for the Biodegradation of Beta-endorphin?

Which Enzymes are Responsible for the Biodegradation of Beta-endorphin?


We are searching data for your request:

Forums and discussions:
Manuals and reference books:
Data from registers:
Wait the end of the search in all databases.
Upon completion, a link will appear to access the found materials.

Which enzymes are responsible for the biodegradation of the endogenous opioid peptide, beta-endorphin?


As far as I have been able to find out, there isn't a definitive answer to this question.

β endorphin is a peptide, and it would seem that there are numerous brain peptidase enzymes that are implicated in the hydrolysis of β endorphin and other peptide neurotransmitters. These include aminopeptidase N, membrane-dipeptidase A, angiotensin-converting enzyme and neutral endopeptidase (also called enkephalinase).

For an early (1984) review of this topic see here; there doesn't seem to be a more up-to-date review, but there is quite a bit of literature on the effects of inhibitors of these various enzymes.


Usage of Potential Micro-organisms for Degradation of Plastics

Plastics are high molecular weight organic source materials. It is necessary to devise systems to decompose plastic polymers because their disruptive effects are threatening the ecosystem. Biotic and abiotic strategies are being employed to convert plastics into monomers. The objective of both techniques is to reduce polymers to monomers. Microbes act on monomers for their degradation by releasing enzymes on polymers. The rate of microbial degradation is affected by both the environmental conditions as well as by polymer characteristics. Different methods are used to check the rate of biological degradation However, some plastics oppose microbial action. The environment condition and polymer characteristics affect the rate of degradation. Different approaches are used to check the rate of biological degradation. The need of the time is to generate bio based plastics material which can be degraded efficiently. These polymers can be recycled by degradation to monomers and then convert back to petrochemical products. This will contribute to fulfill the increasing demand of organic fuels and may serve as next generation fuel. There is no effective technique that can degrade plastics with efficacy, so scientists are struggling to develop techniques which not only degrade these polymers but also results into beneficial products. This review is an attempt to organize some of the most common strategies for degradation of various types of polymers along with a list of potential microbes capable of feeding on them.

Main article text

Abbreviations

PE: Polyethylene PP: Polypropylene PVC: Polyvinyl Chloride PS: Polystyrene PUR: Poly Urethane PET: Poly Ethylene Tetraphthalate PEA: Polyethylene Adipate PCL: Polycaprolactone PES: Polyethylene Succinate PLA: Poly Lactic Acid PHB: Poly Hydroxyl Butyrate PBS: Poly Butylenes Succinate PBSA: Poly Butylenesuccinate-co-adipate PHA: Polyhydroxy Alkanoate PVA: Polyvinyl Alcohol PEG: Polyethylene Glycol PES: Polyethersulfone

Introduction

Plastics are vital hydrocarbons occurring both in natural as well as in synthetic forms. “Plastic” word comes from “plastikos” having Greek origin and implies to any material which can be molded in any shape. Generally thechemical structure of plastics comprises high molecular weight and long chains of hydrocarbon polymers [1].

Plastics have wide range of applications ranging from industrial, agricultural to domestic market. Examples include common use of polyethylene soil mulching in the agriculture sector [2]. The use and demand of polymers is growing day by day due to its applications in every field of life. Large scale synthetic plastic production initiated in 1950. There is a 20-fold increase in plastic production over five decades since 1964 [3]. Synthetic polymer production has enormously progressed for the last 60 years. This time period allowed polymeric materials to step into every domain of human life [4]. A global estimate suggested plastics production up to three hundred and eleven million tons in 2014 [3,5] and reached up to three hundred and thirty five million tons by 2015 which clearly depicts its enormously rapid production which will become twice in the next two decades whereas will reach nearly four-folds by 2050 [3].

Plastics are actually the polymers derived from carbon source including fossil fuels and renewable resources. The discovery of the chemical process leading to the production of long chain, high molecular weight synthetic polymers from organic sources was the novel advancement in the field of chemical sciences. These long chain polymers are notorious for their distinguished features such as strength, malleability, low-weight, easy and inexpensive production [2]. Moreover, the polymeric materials can be categorized into natural polymers, synthetic polymers and blends of polymers, whereas the petrochemical compounds extracted from coal, oil and natural gas are the prime sources of these polymers [6].These organic and inorganic raw materials are substantially composed of carbon, hydrogen, silicon, nitrogen, oxygen and chloride [7]. About 80% of the global plastic usage is of synthetic origin which includes polyethylene (PE), polypropylene (PP), polyvinyl chloride (PVC), polystyrene (PS), polyurethane (PUR) and polyethylene terephthalate (PET) [8].

The division of plastics is majorly based on two class’s i.e. biodegradable and non-biodegradable plastics. Biodegradable plastics are known for some major advantages (1) they help in enriching the soil by returning to the soil through the process of composting with organic wastes (2) it causes the abatement of injuries to wild animals which occurs mainly by dumping of conventional plastics (3) as they are degraded naturally in the environment so, it reduces the need for labor and ultimately labor cost for removal of plastic waste from the environment (4) this feature also enhances the durability and stability of landfills by decreasing the amount of waste. (5) Effective microbial and enzymatic activity reduces them to monomers and oligomers [9].There are various non-biodegradable polymers which are well documented and mainly includes PE, PP, PS and PVC [6]. However, high accumulation of non-biodegradable plastics in the environment due to inappropriate waste management practices and uncontrolled waste disposals are the cause of destruction on earth [6,10].

Extensive production of synthetic non-biodegradable plastics which initiated in the last century is contributing about 10% of the total waste disposed of either properly or improperly [11]. Non-biodegradable plastic waste is wreaking havoc for terrestrial and marine environments as large scale pollution [11,12]. A study suggested that China is the largest plastic waste producing country in the world with 8.82 million metric tons per year (Figure 1). An estimate indicated that 10 to 20 million tons of plastics find their way into the oceans annually [3]. Also, plastic debris has not spared Greenland and Barant Seas having hundreds of thousand pieces per square kilometers [3,13]. In addition, the Antarctic Ocean has also been affected by plastic litter. These regions are mainly inhabited by microplastics (


Contents

The increasing amount of bacterial genomic data provides new opportunities for understanding the genetic and molecular bases of the degradation of organic pollutants. Aromatic compounds are among the most persistent of these pollutants and lessons can be learned from the recent genomic studies of Burkholderia xenovorans LB400 and Rhodococcus sp. strain RHA1, two of the largest bacterial genomes completely sequenced to date. These studies have helped expand our understanding of bacterial catabolism, non-catabolic physiological adaptation to organic compounds, and the evolution of large bacterial genomes. First, the metabolic pathways from phylogenetically diverse isolates are very similar with respect to overall organization. Thus, as originally noted in pseudomonads, a large number of "peripheral aromatic" pathways funnel a range of natural and xenobiotic compounds into a restricted number of "central aromatic" pathways. Nevertheless, these pathways are genetically organized in genus-specific fashions, as exemplified by the b-ketoadipate and Paa pathways. Comparative genomic studies further reveal that some pathways are more widespread than initially thought. Thus, the Box and Paa pathways illustrate the prevalence of non-oxygenolytic ring-cleavage strategies in aerobic aromatic degradation processes. Functional genomic studies have been useful in establishing that even organisms harboring high numbers of homologous enzymes seem to contain few examples of true redundancy. For example, the multiplicity of ring-cleaving dioxygenases in certain rhodococcal isolates may be attributed to the cryptic aromatic catabolism of different terpenoids and steroids. Finally, analyses have indicated that recent genetic flux appears to have played a more significant role in the evolution of some large genomes, such as LB400's, than others. However, the emerging trend is that the large gene repertoires of potent pollutant degraders such as LB400 and RHA1 have evolved principally through more ancient processes. That this is true in such phylogenetically diverse species is remarkable and further suggests the ancient origin of this catabolic capacity. [3]

Anaerobic microbial mineralization of recalcitrant organic pollutants is of great environmental significance and involves intriguing novel biochemical reactions. [4] In particular, hydrocarbons and halogenated compounds have long been doubted to be degradable in the absence of oxygen, but the isolation of hitherto unknown anaerobic hydrocarbon-degrading and reductively dehalogenating bacteria during the last decades provided ultimate proof for these processes in nature. While such research involved mostly chlorinated compounds initially, recent studies have revealed reductive dehalogenation of bromine and iodine moieties in aromatic pesticides. [5] Other reactions, such as biologically induced abiotic reduction by soil minerals, [6] has been shown to deactivate relatively persistent aniline-based herbicides far more rapidly than observed in aerobic environments. Many novel biochemical reactions were discovered enabling the respective metabolic pathways, but progress in the molecular understanding of these bacteria was rather slow, since genetic systems are not readily applicable for most of them. However, with the increasing application of genomics in the field of environmental microbiology, a new and promising perspective is now at hand to obtain molecular insights into these new metabolic properties. Several complete genome sequences were determined during the last few years from bacteria capable of anaerobic organic pollutant degradation. The

4.7 Mb genome of the facultative denitrifying Aromatoleum aromaticum strain EbN1 was the first to be determined for an anaerobic hydrocarbon degrader (using toluene or ethylbenzene as substrates). The genome sequence revealed about two dozen gene clusters (including several paralogs) coding for a complex catabolic network for anaerobic and aerobic degradation of aromatic compounds. The genome sequence forms the basis for current detailed studies on regulation of pathways and enzyme structures. Further genomes of anaerobic hydrocarbon degrading bacteria were recently completed for the iron-reducing species Geobacter metallireducens (accession nr. NC_007517) and the perchlorate-reducing Dechloromonas aromatica (accession nr. NC_007298), but these are not yet evaluated in formal publications. Complete genomes were also determined for bacteria capable of anaerobic degradation of halogenated hydrocarbons by halorespiration: the

1.4 Mb genomes of Dehalococcoides ethenogenes strain 195 and Dehalococcoides sp. strain CBDB1 and the

5.7 Mb genome of Desulfitobacterium hafniense strain Y51. Characteristic for all these bacteria is the presence of multiple paralogous genes for reductive dehalogenases, implicating a wider dehalogenating spectrum of the organisms than previously known. Moreover, genome sequences provided unprecedented insights into the evolution of reductive dehalogenation and differing strategies for niche adaptation. [7]

Recently, it has become apparent that some organisms, including Desulfitobacterium chlororespirans, originally evaluated for halorespiration on chlorophenols, can also use certain brominated compounds, such as the herbicide bromoxynil and its major metabolite as electron acceptors for growth. Iodinated compounds may be dehalogenated as well, though the process may not satisfy the need for an electron acceptor. [5]

Bioavailability, or the amount of a substance that is physiochemically accessible to microorganisms is a key factor in the efficient biodegradation of pollutants. O'Loughlin et al. (2000) [8] showed that, with the exception of kaolinite clay, most soil clays and cation exchange resins attenuated biodegradation of 2-picoline by Arthrobacter sp. strain R1, as a result of adsorption of the substrate to the clays. Chemotaxis, or the directed movement of motile organisms towards or away from chemicals in the environment is an important physiological response that may contribute to effective catabolism of molecules in the environment. In addition, mechanisms for the intracellular accumulation of aromatic molecules via various transport mechanisms are also important. [9]

Petroleum oil contains aromatic compounds that are toxic to most life forms. Episodic and chronic pollution of the environment by oil causes major disruption to the local ecological environment. Marine environments in particular are especially vulnerable, as oil spills near coastal regions and in the open sea are difficult to contain and make mitigation efforts more complicated. In addition to pollution through human activities, approximately 250 million litres of petroleum enter the marine environment every year from natural seepages. [10] Despite its toxicity, a considerable fraction of petroleum oil entering marine systems is eliminated by the hydrocarbon-degrading activities of microbial communities, in particular by a recently discovered group of specialists, the hydrocarbonoclastic bacteria (HCB). [11] Alcanivorax borkumensis was the first HCB to have its genome sequenced. [12] In addition to hydrocarbons, crude oil often contains various heterocyclic compounds, such as pyridine, which appear to be degraded by similar mechanisms to hydrocarbons. [13]

Many synthetic steroidic compounds like some sexual hormones frequently appear in municipal and industrial wastewaters, acting as environmental pollutants with strong metabolic activities negatively affecting the ecosystems. Since these compounds are common carbon sources for many different microorganisms their aerobic and anaerobic mineralization has been extensively studied. The interest of these studies lies on the biotechnological applications of sterol transforming enzymes for the industrial synthesis of sexual hormones and corticoids. Very recently, the catabolism of cholesterol has acquired a high relevance because it is involved in the infectivity of the pathogen Mycobacterium tuberculosis (Mtb). [1] [14] Mtb causes tuberculosis disease, and it has been demonstrated that novel enzyme architectures have evolved to bind and modify steroid compounds like cholesterol in this organism and other steroid-utilizing bacteria as well. [15] [16] These new enzymes might be of interest for their potential in the chemical modification of steroid substrates.

Sustainable development requires the promotion of environmental management and a constant search for new technologies to treat vast quantities of wastes generated by increasing anthropogenic activities. Biotreatment, the processing of wastes using living organisms, is an environmentally friendly, relatively simple and cost-effective alternative to physico-chemical clean-up options. Confined environments, such as bioreactors, have been engineered to overcome the physical, chemical and biological limiting factors of biotreatment processes in highly controlled systems. The great versatility in the design of confined environments allows the treatment of a wide range of wastes under optimized conditions. To perform a correct assessment, it is necessary to consider various microorganisms having a variety of genomes and expressed transcripts and proteins. A great number of analyses are often required. Using traditional genomic techniques, such assessments are limited and time-consuming. However, several high-throughput techniques originally developed for medical studies can be applied to assess biotreatment in confined environments. [17]

The study of the fate of persistent organic chemicals in the environment has revealed a large reservoir of enzymatic reactions with a large potential in preparative organic synthesis, which has already been exploited for a number of oxygenases on pilot and even on industrial scale. Novel catalysts can be obtained from metagenomic libraries and DNA sequence based approaches. Our increasing capabilities in adapting the catalysts to specific reactions and process requirements by rational and random mutagenesis broadens the scope for application in the fine chemical industry, but also in the field of biodegradation. In many cases, these catalysts need to be exploited in whole cell bioconversions or in fermentations, calling for system-wide approaches to understanding strain physiology and metabolism and rational approaches to the engineering of whole cells as they are increasingly put forward in the area of systems biotechnology and synthetic biology. [18]

In the ecosystem, different substrates are attacked at different rates by consortia of organisms from different kingdoms. Aspergillus and other moulds play an important role in these consortia because they are adept at recycling starches, hemicelluloses, celluloses, pectins and other sugar polymers. Some aspergilli are capable of degrading more refractory compounds such as fats, oils, chitin, and keratin. Maximum decomposition occurs when there is sufficient nitrogen, phosphorus and other essential inorganic nutrients. Fungi also provide food for many soil organisms. [19]

For Aspergillus the process of degradation is the means of obtaining nutrients. When these moulds degrade human-made substrates, the process usually is called biodeterioration. Both paper and textiles (cotton, jute, and linen) are particularly vulnerable to Aspergillus degradation. Our artistic heritage is also subject to Aspergillus assault. To give but one example, after Florence in Italy flooded in 1969, 74% of the isolates from a damaged Ghirlandaio fresco in the Ognissanti church were Aspergillus versicolor. [20]


Petro-Plastics With Ester-Bond Back-Bones and Side-Chains

Polyethylene terephthalate (PET) and polyurethane (PU) plastics have heteroatoms in the main chain. Plastics composed of polymers with carbon and hetero atoms in the main chain have improved thermal stability relative to polymers with only carbon backbone (Venkatachalam et al., 2012). These polymers are susceptible to hydrolytic attack of e.g., ester or amide bonds (Muller et al., 2001). Plastics with heteroatoms in the main chain can be degraded by photo-oxidation, hydrolysis and biodegradation (Muller et al., 2001). This results in the formation of smaller fragments and carboxylic end groups.

Polyethylene Terephthalate

PET Degrading Microorganisms

Polyethylene terephthalate is one of the major synthetic petro-plastics ( Figure 1 ) that is produced in very large amounts globally. Its worldwide production accounted to 56 million tons in 2013 (Neufeld et al., 2016). PET consists of aromatic polyesters with high glass transition temperatures (Tg) of approximately 75�ଌ in air. However, Tg decreases to 60�ଌ in aqueous solution (Kawai et al., 2014). At temperatures above the Tg, the amorphous regions of PET become flexible and more accessible to microbial degradation and/or enzymatic attack. With polymer degradation, a decrease in Tg was observed as a result of reduction in the average chain length, due to the higher motility of shorter chains (Odusanya et al., 2013).

PET is used in a wide-variety of applications, such as in manufacturing bottles, containers, textile fibers, and films. PET polymers differ in crystallinities based on its usage. While, most PET used for manufacturing textiles and bottles has high crystallinity (30�%), PET used for packaging has less crystallinity (approximately 8%) (Kawai et al., 2014). Commercially available low-crystalline PET (PET-GF) has approximately 6𠄷% crystallinity (Ronqvist et al., 2009 Kawai et al., 2014).

Although most biodegradable plastics are polyesters [e.g., polyhydroxyalkanoate, PCL, polybutylene succinate, polybutylene succinate-co-adipate, and poly(butylene adipate-co-terephthalate) (PBAT)], PET, which is also a polyester, but is considered to be recalcitrant to biodegradation (Marques-Calvo et al., 2006). However, several microorganisms capable of metabolizing these polymers have been identified in recent years ( Table 2 ). These include the bacterium Ideonella sakaiensis 201-F6, is able to depolymerize PET polymers and utilize the terephthalate subunits as a carbon and energy source for metabolism and growth (Yoshida et al., 2016). Biotechnological conversion of PET through pyrolysis and conversion to polyhydroxyalkanoate has been demonstrated using different Pseudomonads (Ward et al., 2006 Kenny et al., 2008 Guzik et al., 2014). Pyrolysis of PET resulted in terephthalate, which was used as feedstock for P. putida GO16 (Kenny et al., 2012). The degradation rate of PET films depends on the crystallinity, purity of films and orientation of the polymer chains. The degradation of the commercially available standard PET film (pure and amorphous PET) at 50ଌ was found significantly lower (approximately 5%), however, the degradation increased with high temperatures (55, 60, and 65ଌ) to more than 30% (Oda et al., 2018).

Mechanism of PET Biodegradation

Yoshida et al. (2016) demonstrated that PET-digesting enzyme labeled as PETase, converts PET to mono(2-hydroxyethyl) terephthalic acid (MHET), with minimal amounts of terephthalic acid (TPA) and bis(2-hydroxyethyl)-TPA as secondary products ( Figure 4 ). Another enzyme, MHETase (MHET-digesting enzyme), further hydrolyzes MHET into the two monomers, TPA and EG ( Figure 4 ). Li et al. (2019) has reported the metabolism of EG, the second component of PET besides terephthalate, in P. putida KT2440. The metabolism of EG and its derivatives has resulted in different oxidation products such as glycolaldehyde, glyoxal, glycolate, and glyoxylate ( Figure 5 ), which have a variety of value-added applications. These find use as reactive building blocks in the production of agro-, aroma-, and polymer chemicals, or pharmaceuticals (Sajtos, 1991 Mattioda and Christidis, 2000 Yue et al., 2012). Several microorganisms have been reported to utilize EG such as those from Acetobacter and Gluconobacter (DeLey and Kersters, 1964), Acinetobacter and halophilic bacterium, T-52 (ATCC 27042) (Gonzalez et al., 1972 Caskey and Taber, 1981), Flavobacterium species (Child and Willetts, 1978), Hansenula (Harada and Hirabayashi, 1968), Candida, Pichia naganishii AKU4267, and Rhodotorula sp. 3Pr-126 (Kataoka et al., 2001).

Microbial degradation of Polyethylene Terephthalate (PET) (adapted from Austin et al., 2018). PETase, polyethylene terephthalate (PET) hydrolase or PET-digesting enzyme BHET, bis(2-hydroxyethyl) terephthalic acid MHET, mono(2-hydroxyethyl) terephthalic acid TPA, terephthalic acid EG, Ethylene glycol.

Ethylene glycol metabolism by Pseudomonas putida. PedE and PedH are the PQQ-dependent alcohol dehydrogenases (ADHs) PedI, PP_0545 and PP_2049 are the NADH-dependent aldehyde dehydrogenases (ALDHs). These enzyme encoding genes forms the phenylethanol degradation (Ped) cluster of Pseudomonas putida (adapted from Muckschel et al., 2012).

The genetic mechanism for the pathways enabling EG metabolism has been well demonstrated in P. putida. P. putida strain JM37 was able to utilize EG as a sole source of carbon and energy. However, while P. putida KT2440 was able to use EG a carbon source, it did not grow well. Both strains were able to metabolize EG and produced glycolic acid and glyoxylic acid ( Figure 5 ). Initially, the diol is oxidized into glyoxylate in a series of reactions catalyzed by a set of dehydrogenases genes that encode the PP_0545, PedI, PedE, and PedH enzymes (Muckschel et al., 2012 Wehrmann et al., 2017). The conversion of EG to glyoxylate yields three reducing equivalents, either in the form of PQQH2, NADH, or in a direct coupling to the electron transport chain. Glyoxylate can be further metabolized by the Glyoxylate carboligase B (GclB) enzyme or AceA enzyme involved in the glyoxylate shunt (Blank et al., 2008) yielding two molecules of CO2 and two reducing equivalents.

In P. putida strain JM37, the activity of two additional pathways, namely, Gcl and GlcB, leads to rapid metabolism of EG without accumulation of the intermediates and/or oxalic acid (Muckschel et al., 2012). In case of P. putida U and P. aeruginosa, the Pyrroloquinoline quinone (PQQ)-dependent alcohol dehydrogenases involved are the PedE and PedH, and the ExaA (formerly QedH) enzymes (Arias et al., 2008). These periplasmic enzymes have been found essential for growth utilizing ethanol as a carbon source, since it catalyzes the oxidation of the substrate into acetaldehyde (Gorisch, 2003). Franden et al. (2018) has reported improved growth and EG utilization upon overexpression of Glyoxylate carboligase (gcl) operon in the engineered P. putida KT2440 (ATCC 47054) strain. In addition, the engineered strain enables conversion of EG to medium-chain-length polyhydroxyalkanoates (mcl-PHAs) (Mohanan et al., 2020).


Systems biology, network theory and biodegradation

Although originally genetic modification appeared to be the solution to environmental pollution, it appears that biodegradation processes are framed in a complex web of metabolic and regulatory interactions, difficult to approach with the traditional molecular approaches ( Cases & de Lorenzo, 2005b). The recent emergence of ‘omics’ technologies (genomics, proteomics and metabolomics) and the application of ideas and methods for network analysis have offered new insights into the biodegradation process with a new ‘systems biology’ perspective.

Systems biology has been set up to examine complex biological interactions and processes using a comprehensive approach ( Kitano, 2001). This field began to develop in the 1960s, but it was not institutionalized until this century. It is clear that scientists have always been aware of the need to integrate the information produced by the detailed study of individual proteins and genes and that the reductionist approach was only the first step towards understanding the entire process of life. However, for a long time, the experimental procedures only allowed the analysis of a few proteins at a time. New techniques, and in particular large-scale approaches, have made a more comprehensive view possible, in which the system is not divided into parts to study them individually, but it is possible to look into the interactions between the parts and how they influence the behaviour of the system ( Noble, 2006 Sauer, 2007). This systems biology vision of biology is, at least initially, highly related to application of the network theory to the study of functional properties, and to deciphering of the mechanisms involved in the organization of biological systems, with the ultimate goal of modelling and predicting their response to internal and external, for example environmental, variations ( Feist, 2007 Feist & Palsson, 2008).

In the last few years, we have seen growing activity addressing the issues related to understanding of the structure of connections between the elements of various complex systems described at the level of their interactions, including social (networks of scientific collaboration or the World Wide Web), technological (network connections between routers) and finally biological systems (networks between genes or metabolic regulation). Elements are represented by nodes, and the relations between them by edges. The number of edges connected to a node can vary (node degree) and each one of them can have an associated numerical value (weight). Edges can have a direction, such as a catalytic reaction or a signal transduction step, or can be bidirectional, as in protein–protein interaction networks. The application of this concept to biological networks, and metabolic networks in particular, has allowed an initial interpretation of their structure, behaviour and evolution. Three properties of these networks have attracted considerable attention because they reveal the basic organizational principles of biological systems. Metabolic networks are often ‘scale-free’, ‘small world’ and hierarchical. While some of these properties are still controversial (see for instance Arita, 2005 Khanin & Wit, 2006), they provide a useful conceptual framework for the analysis of global biological properties.

The ‘scale-free’ nature of a metabolic network implies a high heterogeneity in the number of connections of their nodes (defined by the chemical compounds). While most of them present a low connectivity, that is, they participate in very few metabolic reactions, a few nodes, called hubs, have a high connectivity ( Barabasi & Bonabeau, 2003). A particularly important property of scale-free networks is its robustness. Random removal of nodes in these networks is more likely to affect a node connected to the few other nodes than to a very connected one, thus preserving the basic structure and behaviour. That is, mutation of a randomly selected enzyme will most likely affect a peripheral pathway, having little effect on the general metabolism or the physiology of the cell. On the other hand, the selective removal of hubs can lead to abrupt changes in the system ( Albert, 2000), or, in other words, the mutation of a main enzyme of a central pathway, such as the Krebs cycle, can be extremely deleterious for the cell. The existence of hubs in this kind of networks has been explained by a mechanism called ‘preferential coupling’. According to this model, ‘scale-free’ networks are the result of a growth process during which new nodes are added to the system by connecting to nodes that already have many links ( Jeong, 2000). In biological terms, this means that a novel metabolic reaction would be more likely selected by evolution if the product can in turn be used in many other metabolic reactions.

Metabolic networks are also ‘small world’. Such networks are an intermediate state between a completely randomly connected network and very regular ones (those in which all nodes have a similar number of connections). In regular networks, if one node A is connected to two others, B and C, B and C also tend to be connected. The frequency with which this occurs is called the clustering coefficient. ‘Small world’ and regular networks have large clustering coefficients, while random networks present a small one. In turn, what separates regular and ‘small world’ networks is the average distance between nodes. While regular networks tend to have long average distances, random and ‘small world’ networks present short average distances between nodes. In other words, in ‘small world’ networks, nodes tend to form highly interconnected clusters that are well connected among them. In metabolic networks, these properties translate into the easy interconversion of metabolites and also into enhanced metabolic stability, because disturbances in the concentration of a particular metabolite are rapidly buffered by the whole metabolic network ( Jeong, 2000 Wagner & Fell, 2001).

Finally, metabolic networks are hierarchical. This means that enzymes are clustered into interconnected groups working together to perform a relatively discrete function, and that these modules are connected among them by specific nodes that acquire a critical importance to maintain the flow in the full system. These nodes are normally highly connected, that is, they are hubs, and have the additional property of presenting a low clustering coefficient, or in other words, their neighbouring nodes are not normally connected among them ( Ravasz, 2002). In biological systems, hierarchical modularity agreed with the notion that evolution can act at several levels of organization simultaneously: at the level of particular modules and at the level of interconnection of these modules.

Additionally, evolution has operated by copying existing modules, adapting them to relatively different tasks, which increases the complexity of the system/organism. In this way, hierarchical networks arise from duplication of nodes that form clusters, a process that, in principle, could be repeated indefinitely ( Barabasi & Oltvai, 2004). On the other hand, the evolution of the integration between modules remains an open question, as does their impact on the network structure and behaviour ( Parter, 2007 Tamames, 2007).

All these system biology approaches have been possible due to the wealth of knowledge accumulated on cellular metabolism, and its formalization and categorization in rich databases such as KEGG ( Kanehisa, 2006) or Metacyc ( Karp, 2002). In the particular case of biodegradation pathways, although numerous experimental studies have provided information about the biochemical reactions for individual biodegradation pathways ( Warhurst, 1994 Seeger, 1995 Casellas, 1997), it was not until 1995 that the University of Minnesota pioneered the compilation of the University of Minnesota Biocatalysis/Biodegradation Database (UM-BBD), collecting information on many biodegradation reactions and pathways of recalcitrant chemicals ( Ellis, 2006). In the last version, this database contains information on over 900 compounds, over 600 enzymes, nearly 1000 reactions and about 350 microorganism entries. This resource includes data about bacterial species in which reactions have been described, the conditions under which they take place and bibliographic references.

It is interesting to note that typically systems biology approaches consider the ‘complete system’ to be a single cell, a formulation that is clearly insufficient in the case of biodegradation where the ‘system’ is a complex environment including multiple biotic and abiotic components. An initial step in this direction was the work of Pazos (2003), which considered all the biodegradation reactions known as part of a single supraorganism metabolic network, assuming that metabolic activities and substrates and intermediate compounds flow freely in the environment. This extreme assumption is supported by several previous observations of biodegradation scenarios. On the one hand, the coordination of microbial communities to mediate biodegradation, the transference of substrates and products between species and communities, is well documented ( Pelz, 1999 Abraham, 2002). Also, the importance of HGT as a mechanism that incorporates biochemical capabilities into bacterial communities through the direct transfer of catabolic genes ( Dejonghe, 2000) and the overall consideration of the biodegradation communities as a pool of genes needed to carry out the degradation in different parts of the pathway have also been reported. Finally, atmospheric phenomena mobilize and disperse large amounts of polluting compounds to sites distant from their place of origin ( Carrera, 2002). All these facts make credible the model proposed by Pazos and colleagues, which considers the biodegradation process as a single interconnecting network (metabolic cooperation) where the boundaries between bacterial species are blurred (easy to incorporate new capabilities by HGT) and without a precise geographical location (dispersion of pollutants). Another concept, this time from the original UM-BBD, was to add an abstract node to the network representing the central metabolism, a node that is used as a central point towards which the distances of all the other nodes can be calculated.

The first important observation studying this global biodegradation network was that it has a free-scale structure similar to other metabolic networks, allowing a rapid degradation of many compounds: the average distance to the central metabolism is only 3.3 steps. This observation, that the biodegradation network, a supraorganism metabolism, presents properties similar to single-organism ones, poses interesting questions on how they have evolved and what are the selection mechanisms at the ecosystem level. The study of the biodegradation network topology also revealed a funnel-like structure in which several pathways converge to a set of common intermediates, which constitute the hubs of the network, and that are placed closer to the central metabolism ( Fig. 1). This property is specific to the biodegradation network and diverges for the hierarchical modularity of the general metabolism. This observation confirms and generalizes the early postulates of the seminal work of Ramos & Timmis (1987) besides, their importance as basic knowledge of the biodegradation process also has relevance for bioremediation applications, and as shown by Ramos and Timmis, when designing an artificial pathway using recombinant DNA technology, it is more favourable to link it to the general metabolism through one of the hub compounds.

The global biodegradation networks of chemical compounds. Network constructed with information from UM-BBD and Metarouter. The larger circle represents the central metabolism. The linear pathways converging on particular intermediates forming a funnel topology can be easily observed. Overlaid are the trends described in Pazos and colleagues. Reactions in exterior part of the network are more rare, and present in less bacterial species, suggesting a more recent appearance in evolution.


Recent developments in microbial biotransformation and biodegradation of dioxins

Polychlorinated dibenzo-p-dioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs), commonly known as dioxins (PCDD/Fs), are toxic environmental pollutants formed from various sources. Elimination of these pollutants from the environment is a difficult task due to their persistent and ubiquitous nature. Removal of dioxins by biological degradation (biodegradation) is considered a feasible method as an alternative to other expensive physicochemical approaches. Biodegradation of dioxins has been extensively studied in several microorganisms, and details concerning biodiversity, biodegradation, biochemistry and molecular biology of this process have accumulated during the last three decades. There are several microbial mechanisms responsible for biodegradation of dioxins, including oxidative degradation by dioxygenase-containing aerobic bacteria, bacterial and fungal cytochrome P-450, fungal lignolytic enzymes, reductive dechlorination by anaerobic bacteria, and direct ether ring cleavage by fungi containing etherase-like enzymes. Many attempts have been made to bioremediate PCDD/Fs using this basic knowledge of microbial dioxin degradation. This review emphasizes the present knowledge and recent advancements in the microbial biotransformation, biodegradation and bioremediation of dioxins.


Role of Microbial Enzymes in the Bioremediation of Pollutants: A Review

A large number of enzymes from bacteria, fungi, and plants have been reported to be involved in the biodegradation of toxic organic pollutants. Bioremediation is a cost effective and nature friendly biotechnology that is powered by microbial enzymes. The research activity in this area would contribute towards developing advanced bioprocess technology to reduce the toxicity of the pollutants and also to obtain novel useful substances. The information on the mechanisms of bioremediation-related enzymes such as oxido-reductases and hydrolases have been extensively studied. This review attempts to provide descriptive information on the enzymes from various microorganisms involved in the biodegradation of wide range of pollutants, applications, and suggestions required to overcome the limitations of their efficient use.

1. Introduction

The quality of life on the Earth is linked inextricably to the overall quality of the environment. Unfortunately the progress in science, technology, and industry a large amount ranging from raw sewage to nuclear waste is let out or dumped into the ecosystem thereby posing a serious problem for survival of mankind itself on earth.

In the past, wastes were traditionally disposed by digging a hole and filling it with waste material. This mode of waste disposal was difficult to sustain owing to lack of new place every time to dump. New technologies for waste disposal that use high-temperature incineration and chemical decomposition (e.g., base-catalyzed dechlorination, UV oxidation) have evolved. Although they can be very effective at reducing wide a range of contaminants but at the same time have several drawbacks. These methods are complex, uneconomical, and lack public acceptance. The associated deficiencies in these methods have focused efforts towards harnessing modern-day bioremediation process as a suitable alternative.

Bioremediation is a microorganism mediated transformation or degradation of contaminants into nonhazardous or less-hazardous substances. The employability of various organisms like bacteria, fungi, algae, and plants for efficient bioremediation of pollutants has been reported [1, 2]. The involvement of plants in the bioremediation of pollutants is called as phytoremediation. The process of phytoremediation is an emerging green technology that facilitates the removal or degradation of the toxic chemicals in soils, sediments, groundwater, surface water, and air (RTDF). Genetically, engineered plants are also in use. For instance arsenic is phytoremediated by genetically modified plants such as Arabidopsis thaliana which expresses two bacterial genes. One of these genes allows the plant to modify arsenate into arsenite and the second one binds the modified arsenite and stores it in the vacuoles [2].

The process of bioremediation mainly depends on microorganisms which enzymatically attack the pollutants and convert them to innocuous products. As bioremediation can be effective only where environmental conditions permit microbial growth and activity, its application often involves the manipulation of environmental parameters to allow microbial growth and degradation to proceed at a faster rate (Figure 1).


The process of bioremediation is a very slow process. Only certain species of bacteria and fungi have proven their ability as potent pollutant degraders. Many strains are known to be effective as bioremediation agents but only under laboratory conditions. The limitation of bacterial growth is under the influence of pH, temperature, oxygen, soil structure, moisture and appropriate level of nutrients, poor bioavailability of contaminants, and presence of other toxic compounds. Although microorganisms can exist in extreme environment, most of them prefer optimal condition a situation that is difficult to achieve outside the laboratory [1, 3–5]. Most bioremediation systems operate under aerobic conditions, but anaerobic environments may also permit microbial degradation of recalcitrant molecules. Both bacteria and fungi rely on the participation of different intracellular and extracellular enzymes respectively for the remediation of recalcitrant and lignin and organopollutants [1, 6].

2. Enzymes

2.1. Introduction to Enzymes

Enzymes are biological catalysts that facilitate the conversion of substrates into products by providing favorable conditions that lower the activation energy of the reaction. An enzyme may be a protein or a glycoprotein and consists of at least one polypeptide moiety. The regions of the enzyme that are directly involved in the catalytic process are called the active sites. An enzyme may have one or more groups that are essential for catalytic activity associated with the active sites through either covalent or noncovalent bonds the protein or glycoprotein moiety in such an enzyme is called the apoenzyme, while the nonprotein moiety is called the prosthetic group. The combination of the apoenzyme with the prosthetic group yields the holoenzyme.

2.2. Enzyme Nomenclature

Enzyme names apply to a single catalytic entity, rather than to a series of individually catalyzed reactions. Names are related to the function of the enzyme, in particular, to the type of reaction catalyzed [7].

2.3. Enzyme Classification

The ultimate identification of a particular enzyme is possible through its enzyme commission (E.C.) number. The assignment of E.C. numbers is described in guidelines set out by the International Union of Biochemistry. All known enzymes fall into one of these six categories. The six main divisions are (1) the oxidoreductases, (2) the transferases, (3) the hydrolases, (4) the lyases, (5) the isomerases, and (6) the ligases (synthetases). Oxidoreductases catalyze the transfer electrons and protons from a donor to an acceptor. Transferases catalyze the transfer of a functional group from a donor to an acceptor. Hydrolases facilitate the cleavage of C–C, C–O, C–N, and other bonds by water. Lyases catalyze the cleavage of these same bonds by elimination, leaving double bonds (or, in the reverse mode, catalyze the addition of groups across double bonds). Isomerases facilitate geometric or structural rearrangements or isomerizations. Finally, ligases catalyze the joining of two molecules [7].

3. Microbial Enzymes in Bioremediation

3.1. Microbial Oxidoreductases

The detoxification of toxic organic compounds by various bacteria and fungi [9] and higher plants [10] through oxidative coupling is mediated with oxidoreductases. Microbes extract energy via energy-yielding biochemical reactions mediated by these enzymes to cleave chemical bonds and to assist the transfer of electrons from a reduced organic substrate (donor) to another chemical compound (acceptor). During such oxidation-reduction reactions, the contaminants are finally oxidized to harmless compounds (ITRC 2002).

The oxidoreductases participate in the humification of various phenolic substances that are produced from the decomposition of lignin in a soil environment. In the same way, oxidoreductases can also detoxify toxic xenobiotics, such as phenolic or anilinic compounds, through polymerization, copolymerization with other substrates, or binding to humic substances [11]. Microbial enzymes have been exploited in the decolorization and degradation of azo dyes [1, 12, 13].

Many bacteria reduce the radioactive metals from an oxidized soluble form to a reduced insoluble form. During the process of energy production, bacterium takes up electrons from organic compounds and use radioactive metal as the final electron acceptor. Some of bacterial species reduce the radioactive metals indirectly with the help of an intermediate electron donor. Finally precipitant can be seen as the result of redox reactions within the metal-reducing bacteria [2].

Chlorinated phenolic compounds are among the most abundant recalcitrant wastes found in the effluents generated by the paper and pulp industry. These compounds are produced upon the partial degradation of lignin during pulp bleaching process. Many fungal species are considered to be suitable for the removal of chlorinated phenolic compounds from the contaminated environments. The activity of fungi is mainly due to the action of extracellular oxidoreductase enzymes, like laccase, manganese peroxidase, and lignin peroxidase, which are released from fungal mycelium into their nearby environment. Being filamentous, fungi can reach the soil pollutants more effectively than bacteria [14].

Water polluted with phenolic compounds can be de-contaminated by plants with the help of enzymes exuded by their roots. The plant families of Fabaceae, Gramineae, and Solanaceae are found to release oxidoreductases which take part in the oxidative degradation of certain soil constituents. Phytoremediation of organic contaminants has been generally focused on three classes of compounds: chlorinated solvents, explosives, and petroleum hydrocarbons [15, 16].

3.1.1. Microbial Oxygenases

Oxygenases belong to the oxidoreductase group of enzymes. They participate in oxidation of reduced substrates by transferring oxygen from molecular oxygen (O2) utilizing FAD/NADH/NADPH as a cosubstrate. Oxygenases are grouped into two categories the monooxygenases and dioxygenases on the basis of number of oxygen atoms used for oxygenation. They play a key role in the metabolism of organic compounds by increasing their reactivity or water solubility or bringing about cleavage of the aromatic ring. Oxygenases have a broad substrate range and are active against a wide range of compounds, including the chlorinated aliphatics. Generally the introduction of O2 atoms into the organic molecule by oxygenase results in cleavage of the aromatic rings. Historically, the most studied enzymes in bioremediation are bacterial mono- or dioxygenases. A detailed study of the role of oxygenases in biodegradation process is available [18–20].

Halogenated organic compounds comprise the largest groups of environmental pollutants as a result of their widespread use as herbicides, insecticides, fungicides, hydraulic and heat transfer fluids, plasticizers, and intermediates for chemical synthesis. The degradation of these pollutants is achieved by specific oxygenases. Oxygenases also mediate dehalogenation reactions of halogenated methanes, ethanes, and ethylenes in association with multifunctional enzymes [19].

3.1.2. Monooxygenases

Monooxygenases incorporate one atom of the oxygen molecule into the substrate. Monooxygenases are classified into two subclasses based on the presence cofactor: flavin-dependent monooxygenases and P450 monooxygenases. Flavin-dependent monooxygenases contain flavin as prosthetic group and require NADP or NADPH as coenzyme. P450 monooxygenases are heme-containing oxygenases that exist in both eukaryotic and prokaryrotic organisms. The monooxygenases comprise a versatile superfamily of enzymes that catalyzes oxidative reactions of substrates ranging from alkanes to complex endogenous molecules such as steroids and fatty acids. Monooxygenases act as biocatalysts in bioremediation process and synthetic chemistry due to their highly region-selectivity and stereoselectivity on wide range of substrates. Majority of mono-oxygenase studied previously are having cofactor, but there are certain monooxygenases which function independent of a cofactor. These enzymes require only molecular oxygen for their activities and utilize the substrate as reducing agent [8, 22].

The desulfurization, dehalogenation, denitrification, ammonification, hydroxylation, biotransformation, and biodegradation of various aromatic and aliphatic compounds are catalyzed by monooxygenases. These properties have been explored in recent years for important application in biodegradation and biotransformation of aromatic compounds [8]. Methane mono-oxygenase enzyme is the best characterized one, among monooxygenases. This enzyme is involved in the degradation of hydrocarbon such as substituted methanes, alkanes, cycloalkanes, alkenes, haloalkenes, ethers, and aromatic and heterocyclic hydrocarbons [23, 24] (Figure 2). Under oxygen-rich conditions, mono-oxygenase catalyzes oxidative dehalogenation reactions, whereas under low oxygen levels, reductive dechlorination takes place. Oxidation of substrate can lead to de-halogenation as a result of the formation of labile products that undergo subsequent chemical decomposition [19, 25, 26].


3.1.3. Microbial Dioxygenases

Dioxygenases are multicomponent enzyme systems that introduce molecular oxygen into their substrate. Aromatic hydrocarbon dioxygenases, belong to a large family of Rieske nonheme iron oxygenases. These dioxygenases catalyze enantiospecifically the oxygenation of wide range of substrates. Dioxygenases primarily oxidize aromatic compounds and, therefore, have applications in environmental remediation. All members of this family have one or two electron transport proteins preceding their oxygenase components. The crystal structure of naphthalene dioxygenase has confirmed the presence of a Rieske (2Fe–2S) cluster and mononuclear iron in each alpha subunit [4].

The catechol dioxygenases serve as part of nature’s strategy for degrading aromatic molecules in the Environment. They are found in the soil bacteria and involved in the transformation of aromatic precursors into aliphatic products. The intradiol cleaving enzymes utilize Fe(III), while the extradiol cleaving enzymes utilize Fe(II) and Mn(II) in a few cases [17] (Figure 3).


3.2. Microbial Laccases

Laccases (p-diphenol:dioxygen oxidoreductase) constitute a family of multicopper oxidases produced by certain plants, fungi, insects, and bacteria, that catalyze the oxidation of a wide range of reduced phenolic and aromatic substrates with concomitant reduction of molecular oxygen to water [9, 29]. Laccases are known to occur in multiple isoenzyme forms each of which is encoded by a separate gene [30], and in, some cases, the genes have been expressed differently depending upon the nature of the inducer [31].

Many microorganisms produce intra and extracellular laccases capable of catalyzing the oxidation of ortho and paradiphenols, aminophenols, polyphenols, polyamines, lignins, and aryl diamines as well as some inorganic ions [29, 32, 33]. Laccases not only oxidize phenolic and methoxy-phenolic acids (Figure 4), but also decarboxylate them and attack their methoxy groups (demethylation). These enzymes are involved in the depolymerization of lignin, which results in a variety of phenols. In addition, these compounds are utilized as nutrients for microorganisms or repolymerized to humic materials by laccase [34]. Among the biological agents, laccases represent an interesting group of ubiquitous, oxidoreductase enzymes that show promise of offering great potential for biotechnological and bioremediation applications [9].


The substrate specificity and affinity of laccase can vary with changes in pH. Laccase can be inhibited by various reagents such as halides (excluding iodide), azide, cyanide, and hydroxide [35]. Different laccases appear to have differing tolerance toward inhibition by halides, indicating differential halide accessibility. Laccase production is sensitive to the nitrogen concentration in fungi. High nitrogen levels are usually required to obtain greater amounts of laccase. Recombinant laccase can be produced by either homologous or heterologous means [9].

3.3. Microbial Peroxidases

Peroxidases (donor: hydrogen peroxide oxidoreductases) are ubiquitous enzymes that catalyze the oxidation of lignin and other phenolic compounds at the expense of hydrogen peroxide (H2O2) in the presence of a mediator. These peroxidases can be haem and nonhaem proteins. In mammals, they are involved in biological processes such as immune system or hormone regulation. In plants, they are involved in auxin metabolism, lignin and suberin formation, cross-linking of cell wall components, defense against pathogens, or cell elongation [37, 38].

The hemeperoxidases have been classified into two distinct groups as found only in animals and found in plants, fungi, and prokaryotes. The second group peroxidases have been subdivided into three classes on the basis of sequence comparison. Class I is intracellular enzymes including yeast cytochrome c peroxidase, ascorbate peroxidase (APX) from plants, and bacterial gene-duplicated catalase peroxidases. Class II consists of the secretory fungal peroxidases such as lignin peroxidase (LiP) and manganese peroxidase (Mnp) from Phanerochaete chrysosporium, and Coprinus cinereus peroxidase or Arthromyces ramosus peroxidase (ARP). The main role of class II peroxidases appears to be the degradation of lignin in wood. Class III contains the secretory plant peroxidases such as those from horseradish (HRP), barley, or soybean. These peroxidases seem to be biosynthetic enzymes involved in processes such as plant cell wall formation and lignifications [37, 38].

Nonhaem peroxidases are not evolutionarily linked and form five independent families. They are thiol peroxidase, alkylhydroperoxidase, nonhaem haloperoxidase, manganese catalase and NADH peroxidase. Among all these thiol peroxidase is the largest and having two subfamilies such as glutathione peroxidases and peroxy redoxins [38].

3.3.1. Classification of Peroxidase Enzymes

Peroxidases have been classified into many types based on its source and activity (http://peroxibase.toulouse.inra.fr/). Among peroxidases, lignin peroxidase (LiP), manganese-dependant peroxidase (MnP), and versatile peroxidase (VP) have been studied the most due to their high potential to degrade toxic substances in nature.

(1) Microbial Lignin Peroxidases
Lignin peroxidases are heme proteins secreted mainly by the white rot fungus during secondary metabolism. In the presence of cosubstrate H2O2 and mediator like veratryl alcohol LiP degrade lignin and other phenolic compounds. During the reaction, H2O2 gets reduced to H2O with the gaining of electron from LiP, (which itself gets oxidized). The LiP (oxidized) with gaining an electron from veratryl alcohol returns to its native reduced state, and veratryl aldehyde is formed. Veratryl aldehyde then again gets reduced back to veratryl alcohol by gaining an electron from substrate. This result in the oxidation of halogenated phenolic compounds, polycyclic aromatic compounds and other aromatic compounds followed by a series of nonenzymatic reactions (Figure 5) [40, 41]. Lignin peroxidase (LiP) plays a central role in the biodegradation of the plant cell wall constituent lignin. LiP is able to oxidize aromatic compounds with redox potentials higher than 1.4 V (NHE) by single-electron abstraction, but the exact redox mechanism is still poorly understood [42].


(2) Microbial Manganese Peroxidases
MnP is an extracellular heme enzyme from the lignin-degrading basidiomycetes fungus, that oxidizes Mn 2+ to the oxidant Mn 3+ in a multistep reaction. Mn 2+ stimulates the MnP production and functions as a substrate for MnP. The Mn 3+ , generated by MnP, acts as a mediator for the oxidation of various phenolic compounds. The resulting Mn³ + chelate oxalate is small enough to diffuse into areas inaccessible even to the enzyme, as in the case of lignin or analogous structures such as xenobiotic pollutants (Figure 6) buried deep within the soil, which are not necessarily available to the enzymes [41].


(3) Microbial Versatile Peroxidases
VP enzymes are able to directly oxidize Mn 2+ , methoxybenzenes, phenolic aromatic substrates like that of MnP, LiP, and horseradish peroxidase. VP has extraordinary broad substrate specificity and tendency to oxidize the substrates in the absence of manganese when compared to other peroxidases. It has also been demonstrated that VP is able to oxidize both phenolic and nonphenolic lignin model dimers [43]. Therefore, a highly efficient VP overproduction system is desired for biotechnological applications in industrial processes and bioremediation of recalcitrant pollutants [27, 44].

4. Microbial Hydrolytic Enzymes

The pollution of soil and water by industrial chemicals and petroleum hydrocarbons is a serious problem of the modern world. Due to their extensive use, they are found as environmental contaminants in numerous aquatic and terrestrial ecosystems. The use of bioremediation technologies for removing these contaminants provides a safe and economic alternative to commonly used physical-chemical treatment. Bacterial activity is the major process involved in the hydrolysis of organic pollutants (Table 1). Extracellular enzyme activity is a key step in degradation and utilization of organic polymers, since only compounds with molecular mass lower than 600 daltons can pass through cell pores [46].

Hydrolytic enzymes disrupt major chemical bonds in the toxic molecules and results in the reduction of their toxicity. This mechanism is effective for the biodegradation of oil spill and organophosphate and carbamate insecticides. Organochlorine insecticides such as DDT and heptachlor are stable in well-aerated soil but readily degrade in anaerobic environments [12, 46, 47]. Hydrolases also catalyze several related reactions including condensations and alcoholysis. The main advantages of this enzyme class are ready availability, lack of cofactor stereoselectivity, and tolerate the addition of water-miscible solvents. Hydrolases belong to group 3 of enzyme class and may further be classified according to the type of bond hydrolyzed [48].

Extracellular hydrolytic enzymes such as amylases, proteases, lipases, DNases, pullulanases, and xylanases have quite diverse potential usages in different areas such as food industry, feed additive, biomedical sciences, and chemical industries [49]. The hemicellulase, cellulase, and glycosidase are of much importance due to its application in biomass degradation [39].

4.1. Microbial Lipases

Lipase degrades lipids derived from a large variety of microorganisms, animals and plants. Recent works have shown that lipase is closely related with the organic pollutants present in the soil. Lipase activity was responsible for the drastic reduction total hydrocarbon from contaminated soil. Research undertaken in this area is likely to progress the knowledge in the bioremediation of oils spill [50, 51]. Lipases have been extracted from bacteria, plant, actinomycetes, and animal cell. Among these microbial lipases are more versatile because of their potent application in industries. These enzymes can catalyze various reactions such as hydrolysis, interesterification, esterification, alcoholysis and aminolysis [52].

Lipases are ubiquitous enzymes which catalyze the hydrolysis of triacylglycerols to glycerol and free-fatty acids. Lipolytic reactions occur at the lipid-water interface, where lipolytic substrates usually form equilibrium between monomeric, micellar, and emulsified states. Lipases have been classified into two types on the basis of criteria such as (a) enhancement in enzyme activity as soon as the triglycerides form an emulsion and (b) lipases with a loop of protein (lid) covering on the active site [53].

Triglyceride is the main component of natural oil or fat. This can hydrolyze consecutively to diacylglycerol, monoacylglycerol, glycerol, and fatty acids. Glycerol and fatty acids are widely used as raw materials, for instance, monoacylglycerol is used as an emulsifying agent in the food, cosmetic, and pharmaceutical industries. The study made on triolein hydrolysis from Candida rugosa lipase in the biphasic oil-water system as proven to be effective. The lipase adsorbs on to the oil-water interface in the bulk of the water phase. The lipase then breaks the ester bonds of triolein to produce consecutively diolein, monoolein, and glycerol. During the catalysis oleic acid is formed at each consecutive reaction stage. The glycerol formed is hydrophilic and thus dissolves into the water phase [36] (Figure 7).


Proposed mechanism for triolein hydrolysis by Candida rugosa lipase in biphasic oil-water system. CE represents the enzyme concentration in the bulk of the water phase [36].

Lipase activity was found to be the most useful indicator parameter for testing hydrocarbon degradation in soil [50, 51]. Lipase is of much interest in the production of regiospecific compounds which are employed in pharmaceutical industry. Along with its diagnostic usage in bioremediation, lipase has many potential applications in food, chemical, detergent manufacturing, cosmetic, and paper making industries, but its production cost has restricted its industrial use [53, 54].

4.2. Microbial Cellulases

Cellulases now promise the potential of converting waste cellulosic material into foods to meet burgeoning population and have been the subject of intense research [55]. Some organisms produce cell bound, cell envelope associated, and some extra cellular cellulases. Extracellular cellulases, hemicellulases, and pectinases have been shown to be constitutively expressed at very low levels by some bacteria and fungi [56, 57].

Cellulases are usually a mixture of several enzymes. At least: three major groups of cellulases are involved in the hydrolysis process (1) endoglucanase (EG, endo- 1,4-D-glucanohydrolase) which attacks regions of low crystallinity in the cellulose fiber, creating free chain ends (2) exoglucanase or cellobiohydrolase (CBH, 1,4-b-D-glucan cellobiohydrolase) which degrade the cellulose molecule further by removing cellobiose units from the free chain ends (3) β-glucosidase which hydrolyzes cellobiose to glucose units. Along with major enzymes, some ancillary enzymes are also present. During the enzymatic hydrolysis, cellulose is degraded by the cellulases to reducing sugars (Figure 8) that can be fermented by yeasts or bacteria to ethanol [58].


Cellulase enzymes are capable of degrading crystalline cellulose to glucose. Cellulases have been used in the manufacture of detergents since early 1990s. Cellulases remove cellulose microfibrils, which are formed during washing and the use of cotton-based cloths. This is also observed as the colour brightening and material softening in the textile industry. Alkaline cellulases are produced by Bacillus strains and neutral and acidic cellulases by Trichoderma and Humicola fungi. In paper and pulp industry, cellulases have been employed for the removal of ink during recycling of paper. The cellulases are added during brewing to increase the juice liberation from fruit pulp and for the production of ethanol from cellulosic biomass [59].

4.3. Microbial Proteases

Proteases hydrolyze the breakdown of proteinaceous substance which enter atmosphere due to shedding and moulting of appendages, death of animals, and also as byproduct of some industries like poultry, fishery, and leather. Proteases belong to group of enzymes that hydrolyze peptide bonds in aqueous environment and synthesize them in nonaqueous environment. Proteases have wide range of applications in food, leather, detergent, and pharmaceutical industry [60, 61].

Proteases are divided as endopeptidases and exopeptidases based on the catalysis of peptide chain. Endo peptidases further grouped based on the position of active site such as serine endopeptidase, cysteine peptidase, aspartic endopeptidases, and metallopeptidases. The enzymes whose reaction mechanism is completely elucidated are grouped under (Figure 9). The exopeptidases act only near the terminal amino or carboxylic position of chain. The protease that acts on free amino, and carboxyl terminals are called as aminopeptidase and carboxypeptidase, respectively.


The endopeptidase acts on the inner regions of peptide chain. The presence of free amino and carboxyl terminal will have negative impact on enzyme activity [62].

Proteases have been used in the manufacture of cheese and detergent manufacturing industry since many years. The alkaline proteases are used in leather industry for the removal of hairs and parts which are present on the animal skin. Proteases have been employed for the production of dipeptide aspartame, which is a noncalorific artificial sweetener. In the pharmaceutical industry, a varying and specific proteases are used in developing effective therapeutic agents. Clostridial collagenase or subtilisin is used in combination with broad-spectrum antibiotics in the treatment of burns and wounds [63].

References

  1. M. Vidali, “Bioremediation. An overview,” Pure and Applied Chemistry, vol. 73, no. 7, pp. 1163–1172, 2001. View at: Google Scholar
  2. M. Leung, “Bioremediation: techniques for cleaning up a mess,” Journal of Biotechnology, vol. 2, pp. 18–22, 2004. View at: Google Scholar
  3. F. Bernhard-Reversat and D. Schwartz, “Change in lignin content during litter decomposition in tropical forest soils (Congo): comparison of exotic plantations and native stands,” Comptes Rendus de l'Academie de Sciences—Serie IIa, vol. 325, no. 6, pp. 427–432, 1997. View at: Google Scholar
  4. M. Dua, A. Singh, N. Sethunathan, and A. Johri, “Biotechnology and bioremediation: successes and limitations,” Applied Microbiology and Biotechnology, vol. 59, no. 2-3, pp. 143–152, 2002. View at: Publisher Site | Google Scholar
  5. L. D. Dana and J. W. Bauder, A General Essay on Bioremediation of Contaminated Soil, Montana State University, Bozeman, Mont, USA, 2011.
  6. K. E. Hammel, “Fungal degradation of lignin,” in Driven by Nature: Plant Litter Quality and Decomposition, G. Cadisch and K. E. Giller, Eds., pp. 33–45, CAB International, Wallingford, UK, 1997. View at: Google Scholar
  7. A. L. Lehninger, D. L. Nelson, and M. M. Cox, Lehninger’s Principles of Biochemistry, W.H. Freeman, New York, NY, USA, 4th edition, 2004.
  8. P. K. Arora, A. Srivastava, and V. P. Singh, “Application of Monooxygenases in dehalogenation, desulphurization, denitrification and hydroxylation of aromatic compounds,” Journal of Bioremediation & Biodegradation, vol. 1, pp. 1–8, 2010. View at: Google Scholar
  9. L. Gianfreda, F. Xu, and J. M. Bollag, “Laccases: a useful group of oxidoreductive enzymes,” Bioremediation Journal, vol. 3, no. 1, pp. 1–25, 1999. View at: Google Scholar
  10. J.-M. Bollag and J. Dec, “Use of Plant material for the removal of pollutants by polymerization and binding to humic substances,” Tech. Rep. R-82092, Center for Bioremediation and Detoxification Environmental Resources Research Institute The Pennsylvania State University, University Park, Pa, USA, 1998. View at: Google Scholar
  11. J.-W. Park, B.-K. Park, and J.-E. Kim, “Remediation of soil contaminated with 2,4-dichlorophenol by treatment of minced shepherd's purse roots,” Archives of Environmental Contamination and Toxicology, vol. 50, no. 2, pp. 191–195, 2006. View at: Publisher Site | Google Scholar
  12. P. P. Williams, “Metabolism of synthetic organic pesticides by anaerobic microorganisms,” Residue Reviews, vol. 66, pp. 63–135, 1977. View at: Google Scholar
  13. Q. Husain, “Potential applications of the oxidoreductive enzymes in the decolorization and detoxification of textile and other synthetic dyes from polluted water: a review,” Critical Reviews in Biotechnology, vol. 26, no. 4, pp. 201–221, 2006. View at: Publisher Site | Google Scholar
  14. O. Rubilar, M. C. Diez, and L. Gianfreda, “Transformation of chlorinated phenolic compounds by white rot fungi,” Critical Reviews in Environmental Science and Technology, vol. 38, no. 4, pp. 227–268, 2008. View at: Publisher Site | Google Scholar
  15. N. Durán and E. Esposito, “Potential applications of oxidative enzymes and phenoloxidase-like compounds in wastewater and soil treatment: a review,” Applied Catalysis B, vol. 28, no. 2, pp. 83–99, 2000. View at: Publisher Site | Google Scholar
  16. L. A. Newman, S. L. Doty, K. L. Gery et al., “Phytoremediation of organic contaminants: a review of phytoremediation research at the University of Washington,” Soil and Sediment Contamination, vol. 7, no. 4, pp. 531–542, 1998. View at: Google Scholar
  17. L. Que and R. Y. N. Ho, “Dioxygen activation by enzymes with mononuclear non-heme iron active sites,” Chemical Reviews, vol. 96, no. 7, pp. 2607–2624, 1996. View at: Google Scholar
  18. P. K. Arora, M. Kumar, A. Chauhan, G. P. Raghava, and R. K. Jain, “OxDBase: a database of oxygenases involved in biodegradation,” BMC Research Notes, vol. 2, article 67, 2009. View at: Publisher Site | Google Scholar
  19. S. Fetzner and F. Lingens, “Bacterial dehalogenases: biochemistry, genetics, and biotechnological applications,” Microbiological Reviews, vol. 58, no. 4, pp. 641–685, 1994. View at: Google Scholar
  20. S. Fetzner, “Oxygenases without requirement for cofactors or metal ions,” Applied Microbiology and Biotechnology, vol. 60, no. 3, pp. 243–257, 2003. View at: Publisher Site | Google Scholar
  21. B. Dedeyan, A. Klonowska, S. Tagger et al., “Biochemical and molecular characterization of a laccase from Marasmius quercophilus,” Applied and Environmental Microbiology, vol. 66, no. 3, pp. 925–929, 2000. View at: Publisher Site | Google Scholar
  22. P. C. Cirino and F. H. Arnold, “Protein engineering of oxygenases for biocatalysis,” Current Opinion in Chemical Biology, vol. 6, no. 2, pp. 130–135, 2002. View at: Publisher Site | Google Scholar
  23. B. G. Fox, J. G. Borneman, L. P. Wackett, and J. D. Lipscomb, “Haloalkene oxidation by the soluble methane monooxygenase from Methylosinus trichosporium OB3b: mechanistic and environmental implications,” Biochemistry, vol. 29, no. 27, pp. 6419–6427, 1990. View at: Publisher Site | Google Scholar
  24. S. Grosse, L. Laramee, K.-D. Wendlandt, I. R. McDonald, C. B. Miguez, and H.-P. Kleber, “Purification and characterization of the soluble methane monooxygenase of the type II methanotrophic bacterium Methylocystis sp. strain WI 14,” Applied and Environmental Microbiology, vol. 65, no. 9, pp. 3929–3935, 1999. View at: Google Scholar
  25. S. Fetzner, “Bacterial dehalogenation,” Applied Microbiology and Biotechnology, vol. 50, no. 6, pp. 633–657, 1998. View at: Publisher Site | Google Scholar
  26. J. P. Jones, E. J. O'Hare, and L. L. Wong, “Oxidation of polychlorinated benzenes by genetically engineered CYP101 (cytochrome P450cam),” European Journal of Biochemistry, vol. 268, no. 5, pp. 1460–1467, 2001. View at: Publisher Site | Google Scholar
  27. D. W.S. Wong, “Structure and action mechanism of ligninolytic enzymes,” Applied Biochemistry and Biotechnology, vol. 157, no. 2, pp. 174–209, 2009. View at: Publisher Site | Google Scholar
  28. H. Wariishi, K. Valli, and M. H. Gold, “Manganese(II) oxidation by manganese peroxidase from the basidiomycete Phanerochaete chrysosporium. Kinetic mechanism and role of chelators,” The Journal of Biological Chemistry, vol. 267, no. 33, pp. 23688–23695, 1992. View at: Google Scholar
  29. C. Mai, W. Schormann, O. Milstein, and A. Huttermann, “Enhanced stability of laccase in the presence of phenolic compounds,” Applied Microbiology and Biotechnology, vol. 54, no. 4, pp. 510–514, 2000. View at: Google Scholar
  30. P. Giardina, R. Cannio, L. Martirani, L. Marzullo, G. Palmieri, and G. Sannia, “Cloning and sequencing of a laccase gene from the lignin-degrading basidiomycete Pleurotus ostreatus,” Applied and Environmental Microbiology, vol. 61, no. 6, pp. 2408–2413, 1995. View at: Google Scholar
  31. M. I. Rezende, A. M. Barbosa, A.-F. D. Vasconcelos, R. Haddad, and R. F.H. Dekker, “Growth and production of laccases by the ligninolytic fungi, Pleurotus ostreatus and Botryosphaeria rhodina, cultured on basal medium containing the herbicide, Scepter® (imazaquin),” Journal of Basic Microbiology, vol. 45, no. 6, pp. 460–469, 2005. View at: Publisher Site | Google Scholar
  32. M. A. Ullah, C. T. Bedford, and C. S. Evans, “Reactions of pentachlorophenol with laccase from Coriolus versicolor,” Applied Microbiology and Biotechnology, vol. 53, no. 2, pp. 230–234, 2000. View at: Google Scholar
  33. S. Rodríguez Couto and J. L. Toca Herrera, “Industrial and biotechnological applications of laccases: a review,” Biotechnology Advances, vol. 24, no. 5, pp. 500–513, 2006. View at: Publisher Site | Google Scholar
  34. J. S. Kim, J. W. Park, S. E. Lee, and J. E. Kim, “Formation of bound residues of 8-hydroxybentazon by oxidoreductive catalysts in soil,” Journal of Agricultural and Food Chemistry, vol. 50, no. 12, pp. 3507–3511, 2002. View at: Publisher Site | Google Scholar
  35. F. Xu, “Catalysis of novel enzymatic iodide oxidation by fungal laccase,” Applied Biochemistry and Biotechnology, vol. 59, no. 3, pp. 221–230, 1996. View at: Google Scholar
  36. H. Hermansyah, A. Wijanarko, M. Gozan et al., “Consecutive reaction model for triglyceride hydrolysis using lipase,” Jurnal Teknologi, vol. 2, pp. 151–157, 2007. View at: Google Scholar
  37. A. N. P. Hiner, J. H. Ruiz, J. N. Rodri et al., “Reactions of the class II peroxidases, lignin peroxidase and Arthromyces ramosus peroxidase, with hydrogen peroxide: catalase-like activity, compound III formation, and enzyme inactivation,” The Journal of Biological Chemistry, vol. 277, no. 30, pp. 26879–26885, 2002. View at: Publisher Site | Google Scholar
  38. D. Koua, L. Cerutti, L. Falquet et al., “PeroxiBase: a database with new tools for peroxidase family classification,” Nucleic Acids Research, vol. 37, supplement 1, pp. D261–D266, 2009. View at: Publisher Site | Google Scholar
  39. O. Schmidt, Wood and Tree Fungi, Springer, Berlin, Germany, 2006.
  40. S. Yoshida, “Reaction of manganese peroxidase of Bjerkandera adusta with synthetic lignin in acetone solution,” Journal of Wood Science, vol. 44, no. 6, pp. 486–490, 1998. View at: Google Scholar
  41. R. Ten Have and P. J. M. Teunissen, “Oxidative mechanisms involved in lignin degradation by white-rot fungi,” Chemical Reviews, vol. 101, no. 11, pp. 3397–3413, 2001. View at: Publisher Site | Google Scholar
  42. K. Piontek, A. T. Smith, and W. Blodig, “Lignin peroxidase structure and function,” Biochemical Society Transactions, vol. 29, no. 2, pp. 111–116, 2001. View at: Publisher Site | Google Scholar
  43. F. J. Ruiz-Dueñas, M. Morales, M. Pérez-Boada et al., “Manganese oxidation site in Pleurotus eryngii versatile peroxidase: a site-directed mutagenesis, kinetic, and crystallographic study,” Biochemistry, vol. 46, no. 1, pp. 66–77, 2007. View at: Publisher Site | Google Scholar
  44. T. Tsukihara, Y. Honda, R. Sakai, T. Watanabe, and T. Watanabe, “Exclusive overproduction of recombinant versatile peroxidase MnP2 by genetically modified white rot fungus, Pleurotus ostreatus,” Journal of Biotechnology, vol. 126, no. 4, pp. 431–439, 2006. View at: Publisher Site | Google Scholar .
  45. E. Vasileva-Tonkova and D. Galabova, “Hydrolytic enzymes and surfactants of bacterial isolates from lubricant-contaminated wastewater,” Zeitschrift fur Naturforschung, vol. 58, no. 1-2, pp. 87–92, 2003. View at: Google Scholar
  46. R. Lal and D. M. Saxena, “Accumulation, metabolism and effects of organochlorine insecticides on microorganisms,” Microbiological Reviews, vol. 46, no. 1, pp. 95–127, 1982. View at: Google Scholar .
  47. C. Sánchez-Porro, S. Martín, E. Mellado, and A. Ventosa, “Diversity of moderately halophilic bacteria producing extracellular hydrolytic enzymes,” Journal of Applied Microbiology, vol. 94, no. 2, pp. 295–300, 2003. View at: Publisher Site | Google Scholar
  48. R. Margesin, A. Zimmerbauer, and F. Schinner, “Soil lipase activity—A useful indicator of oil biodegradation,” Biotechnology Techniques, vol. 13, no. 12, pp. 859–863, 1999. View at: Publisher Site | Google Scholar
  49. R. Riffaldi, R. Levi-Minzi, R. Cardelli, S. Palumbo, and A. Saviozzi, “Soil biological activities in monitoring the bioremediation of diesel oil-contaminated soil,” Water, Air, and Soil Pollution, vol. 170, no. 1–4, pp. 3–15, 2006. View at: Publisher Site | Google Scholar
  50. M. P. Prasad and K. Manjunath, “Comparative study on biodegradation of lipid-rich wastewater using lipase producing bacterial species,” Indian Journal of Biotechnology, vol. 10, no. 1, pp. 121–124, 2011. View at: Google Scholar
  51. D. Sharma, B. Sharma, and A. K. Shukla, “Biotechnological approach of microbial lipase: a review,” Biotechnology, vol. 10, no. 1, pp. 23–40, 2011. View at: Publisher Site | Google Scholar
  52. B. Joseph, P. W. Ramteke, and P. A. Kumar, “Studies on the enhanced production of extracellular lipase by Staphylococcus epidermidis,” Journal of General and Applied Microbiology, vol. 52, no. 6, pp. 315–320, 2006. View at: Publisher Site | Google Scholar
  53. J. W. Bennet, K. G. Wunch, and B. D. Faison, Use of Fungi Biodegradation, ASM Press, Washington, DC, USA, 2002.
  54. J. E. Rixon, L. M.A. Ferreira, A. J. Durrant, J. I. Laurie, G. P. Hazlewood, and H. J. Gilbert, “Characterization of the gene celD and its encoded product 1,4-β-D-glucan glucohydrolase D from Pseudomonas fluorescens subsp. cellulosa,” Biochemical Journal, vol. 285, no. 3, pp. 947–955, 1992. View at: Google Scholar
  55. M. Adriano-Anaya, M. Salvador-Figueroa, J. A. Ocampo, and I. García-Romera, “Plant cell-wall degrading hydrolytic enzymes of Gluconacetobacter diazotrophicus,” Symbiosis, vol. 40, no. 3, pp. 151–156, 2005. View at: Google Scholar
  56. Y. Sun and J. Cheng, “Hydrolysis of lignocellulosic materials for ethanol production: a review,” Bioresource Technology, vol. 83, no. 1, pp. 1–11, 2002. View at: Publisher Site | Google Scholar
  57. M. Leisola, J. Jokela, O. Pastinen, and O. Turunen, 6.54.2.10 Industrial use of Enzymes—Essay, Laboratory of Bioprocess Engineering, Helsinki University of Technology, Helsinki, Finland, 2006.
  58. C. J. Singh, “Optimization of an extracellular protease of Chrysosporium keratinophilum and its potential in bioremediation of keratinic wastes,” Mycopathologia, vol. 156, no. 3, pp. 151–156, 2003. View at: Publisher Site | Google Scholar
  59. A. K. Beena and P. I. Geevarghese, “A solvent tolerant thermostable protease from a psychrotrophic isolate obtained from pasteurized milk,” Developmental Microbiology and Molecular Biology, vol. 1, pp. 113–119, 2010. View at: Google Scholar .
  60. M. B. Rao, A. M. Tanksale, M. S. Ghatge, and V. V. Deshpande, “Molecular and biotechnological aspects of microbial proteases,” Microbiology and Molecular Biology Reviews, vol. 62, no. 3, pp. 597–635, 1998. View at: Google Scholar

Copyright

Copyright © 2011 Chandrakant S. Karigar and Shwetha S. Rao. This is an open access article distributed under the Creative Commons Attribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.


Petroleum Microbiology

Microbial Desulfurization and Denitrogenation

A variety of microbial strains including Rhodococcus, Nocardia, Agrobacterium, Mycobacterium, Gordona, Klebsiella, Xanthomonas, and Paenibacillus are capable of selective desulfurization of organic sulfur. When certain species such as Rhodococcus erythropolis are cultured aerobically, they exhibit the ability to desulfurize compounds such as DBT without degrading the carbon ring structure. These strains can use the DBT-released sulfur as sole source of sulfur for growth. The sequence of catabolism of DBT by Rhodococcus is mediated by two monooxygenases and a desulfinase, and results in successive production of dibenzothiophene -5-oxide (DBTO), dibenzene-5,5-dioxide (DBTO 2), 2-(2-hydroxybiphenyl)-benzenesulfinate (HPBS), and 2-hydroxybiphenyl (HBP) with associated release of inorganic sulfur.

Desulfurization genes have been manipulated by directed evolution and gene shuffling approaches to broaden substrate specificity and improved biocatalysts have been engineered. Deletions in desulfurizing bacteria of enzymes in the biodesulfurization pathway such as dibenzothiophene sulfone monooxygenase (DszA) or hydroxyphenyl benzene sulfinase (DszB) have created opportunities to use these microbial cell biocatalysts for production of potentially valuable sulfur-containing metabolic intermediates as products. Molecular manipulations, involving use of a RhodococcusE. coli shuttle vector, were used to construct the recombinant strain, Rhodococcus sp. T09, which could desulfurize alkylated DBT and BT and could use both DBT and benzothiophene (BT) as the sole sulfur source. Resting cells of these strains could also desulfurize alkylated DBT in oil–water, two-phase systems.

Biodesulfurizations carried out in two-phase aqueous-alkane solvent systems exhibited increased sulfur removal rates as compared with aqueous systems. In treating crude oil, it is necessary to apply intensive high energy mixing and/or addition of a surfactant to create a two-phase microbial biodesulfurization system with high interfacial areas, and then after desulfurization to implement a de-emulsification step. The key technoeconomic challenge to the viability of biodesulfurization processes is to establish cost-effective means of implementing the two-phase bioreactor system and de-emulsification steps as well as the product recovery step. It has been found that use of multiple-stage, airlift reactors can reduce mixing costs and centrifugation approaches facilitate de-emulsification, desulfurized oil recovery, and recycling of the cells. Other identified goals relate to improving microbial cell reaction kinetics and to achieving continuous growth and regeneration of the biocatalyst in the same system, rather than in a separate reactor. Extent of biodesulfurization varies dramatically with the nature of the oil feedstock, and especially with feedstock physical properties and the extent to which the feedstock has been refined. De-emulsification extents in the ranges of 20–60%, 20–60%, 30–70%, 40–90%, 65–70%, and 75–90% were observed for crude oil, light gas oil, middle distillates, diesel, hydrotreated diesel, and cracked stocks, respectively.

While the 1990 Clean Air Act Amendment set sulfur content of diesel fuel at a maximum of 500 ppm, lower sulfur standards are enforced in some jurisdictions and future values for diesel fuel are expected to be in the region of 30 ppm. HDS technologies cannot achieve the future required 30 ppm levels and, the de-emulsification extent ranges above suggest current microbial desulfurization technology is not cost-effective for heavy or middle distillates of crude oil. However, a combination of biodesulfurization and HDS technology has the potential to achieve the future required 30 ppm level.

Bacteria exhibit some general similarities in the pathways where oxygenases play an important role in the initial attack in the transformation of nitrogen compounds. Some species of Alcaligenes, Bacillus, Beijerinckia, Burkholderia, Comamonas, Mycobacterium, Pseudomonas, Serratia, and Xanthomonas can utilize indole, pyridine, quinoline, and carbazole compounds. Pyrrole and indole are easily degraded, but carbazole is relatively resistant to microbial attack. The genes responsible for carbazole degradation by Pseudomonas sp. have been identified and cloned. Gene manipulations have created recombinant strains able to transform a wide range of aromatic compounds including carbazole, N-methylcarbazole, N-ethylcarbazole, dibenzofuran, DBT, dibenzo-p-dioxin, fluorene, naphthalene, phenanthrene, anthracene, and fluoranthene.

Thus, effective biodesulfurization and biodenitrogenation require removal of sulfur and nitrogen through specific enzymatic attack of the C–S and C–N bonds, respectively, but without C–C bond attack, thereby preserving the fuel value of the residual products. Critical to biorefining process development will be the design of a cost-effective, two-phase bioreactor systems with subsequent oil–water separation and product recovery. From a practical implementation perspective, denitrogenation and desulfurization processes need to be integrated and in practice may have to be combined with physical HDS treatments.


Biotechnology in the Pulp and Paper Industry

A. Ferraz , . R. Mendonça , in Progress in Biotechnology , 2002

3.2. Molecular weight distribution of residual cellulose

Biodegradation of cellulose during biopulping by C. subvermispora is often considered low, because glucan (or cellulose) losses are very low even after long biodegradation periods ( Table 1 ). However, the glucan loss measurement is based on the glucose released after acid hydrolysis of the wood samples ( 11, 23 ). This means that glucan loss represents the polymer degraded to carbon dioxide and water. Depolymerization of cellulose obviously starts before the polymer is mineralized, and a better estimation of the cellulose loss during biopulping needs a direct evaluation of the cellulose DP in biotreated samples. This has been evaluated by determining both, the alpha-cellulose contents in biodegraded wood samples and the molecular weight distribution of the residual cellulose. Figure 3 shows that, at the beginning of P. taeda biodegradation by C. subvermispora, neither glucan loss nor alteration in the yield of alpha-cellulose occurs. However, long biodegradation periods bring about a considerable decrease in the yield of alpha-cellulose, while the glucan loss remains low. In the case of Eucalyptus grandis hardwood biodegradation, the alpha-cellulose loss is already significant after two weeks of biotreatment ( 24 ). This suggests that degradation reactions start at the cellulose backbone, even when no significant glucan loss is detected. This kind of degradation reactions can generate low molecular mass glucans, which are soluble in the alkaline solution used for alpha-cellulose preparation, resulting in a decrease in the alpha-cellulose yield. The assimilation rate of these low-molecular-mass fragments by C. subvermispora seems to be very slow, since no significant increase in the glucan loss is detected even after long biodegradation periods. The degradation of the cellulose backbone can also be evidenced by size exclusion chromatography. Figure 4 shows the HPSEC of alpha-cellulose tricarbanilates prepared from P. taeda samples biodegraded by C. subvermispora. The average DP of cellulose was almost the same during the first 30 days of biodegradation, when wood weight losses were no higher than 3%. After this period, alpha-cellulose DP starts to decrease, which shows that although glucan losses are negligible in periods no longer than 90 days, after 30 days of biotreatment, the polymer loses its integrity. These data are in accordance with a report that shows a poor cellulolytic complex in C. subvermispora. Despite having some endo-glucanases activity (which might be responsible for cellulose depolymerization), C. subvermispora is not able to produce significant amounts of cellobiohydrolases responsible for releasing cellobiose from glucans produced by the endo-cellulases ( 25 ).

Figure 3 . Glucan (-•-) and alpha-cellulose (-▴-) losses during biopulping of P. taeda (filled symbols) and E. grandis (open symbols) by C. subvermispora ( 24 ).

Figure 4 . HPSEC of alpha-cellulose tricarbanilates prepared from P. taeda biotreated by C. subvermispora


Watch the video: 2-Minute Neuroscience: Beta-Endorphin (May 2022).


Comments:

  1. Durn

    I'm sure this is not true.

  2. Hampton

    Author, where can you find such a design? I really liked ...

  3. Gaheris

    I absolutely agree with you. There is something in this and I think this is a great idea. I completely agree with you.

  4. Gerhard

    I'm sorry, this is not exactly what I need. There are other options?

  5. Damon

    It agree, very useful piece



Write a message