Before the advent of petrochemistry, great amount of organic materials and chemicals were synthesized from biogenic feedstock. In recent times, almost all these materials and chemicals are derived from fossil feedstock (Palm et al., 2016). One of such materials that are derived from petroleum is synthetic plastics. The exploration of petroleum in response to the current increase in the demand for plastic materials stands as a matter of grave environmental concern, ranging from global warming, human health risks to even ecosystem toxicity (Rodriguez-Perez et al., 2018; Yogesh et al.,2012).
The use of petrochemical based plastics has been associated with a plethora of adverse environmental impacts owing to its characteristic non-degradable nature. These synthetic plastic polymers persist in the environment and have been described to be eco-hazardous. According to Ong et al. (2017), synthetic plastic polymers contribute significantly to the amount of solid waste in the environment. Accumulation of these solid wastes have been found to lead to emission of greenhouse gases (Piemonte, 2011).
Taking into cognizance the recalcitrant nature of petrochemical plastics in the environment, Kourmentza et al. (2017) opined that replacing synthetic plastics with PHAs would present enormous societal and environmental benefits. The search for new materials as suitable replacements of fossil fuel-based plastics has been focused on biopolymers such as Polyhydroxyalkanoates (PHAs) with comparable physicochemical properties to fossil fuel-based plastics (Rodriguez-Perez et al., 2018). In the long-term, to meet the vision of a fossil-free circular economy, fossil fuels and feedstock will have to be phased out (Palm et al., 2016).
Due to their potential applications in variety of industries, PHAs have attracted widespread interest (Zhu et al., 2013). This is as a result of their structural diversity and close analogy to plastics, high molecular weight, chiral polymeric structure, and biodegradable aliphatic esters (Hwan et al., 2008). Ong et al. (2017) described PHAs as an attractive material owing to its associated mechanical properties and biodegradability. As an upshot of its natural degradability in the environment, utilization of PHA is a step closer in the direction of a greener environment with the aim of plummeting the dependency on the non-degradable synthetic plastic polymers.
Despite the comparative advantages of PHAs over petrochemical based plastics, extensive commercialization and industrial development of PHAs is still besieged with high production cost, thus resulting in higher prices when equated to petrochemical plastics (Kourmentza et al., 2017). Fernandez-Dacosta (2015) added that the high costs of PHA production are projected at 20–80% higher than for their petrochemical counterparts. According to Rodriguez-Perez et al. (2018), present industrial production of PHA employs expensive raw materials and chemicals as sources of organic matter, which entails high costs at industrial scale. This has been described by Rodriguez et al. (2018) as a major factor responsible for the high cost of production on an industrial scale. In addition, Kourmentza et al. (2018) stated that cost of downstream processing and low manufacturing yield are also threats to the industrial scale production of PHAs.
According to Kaur and Roy (2015), the process economics can be improved by design and implementation of efficient bioprocess strategies for improving the overall process kinetics, and thus ensuing higher PHA concentrations and productivities
Global dependence on petroleum derived plastics has increased considerably over the years. The world production of plastics is over one hundred million tons per year (Yogeshet al., 2012). The issue of environmental pollution in relation to the disposal of non-biodegradable plastic has serious environmental impact, and this is a growing concern (Zakariaet al., 2008). Plastics are not readily biodegradable in the natural environment, thus they will remain almost indefinitely in the environment, causing pollution (Ojomuet al., 2004). There is therefore pressing need to replace this environmentally damaging product with more friendly alternatives.

The aim of this study is to isolate microorganisms that can utilize cheap, agro-renewable substrate to accumulate large amount of polyhydroxyalkanoate (PHA) as a basis for gradual replacement of petroleum derived plastics.
The objectives of the study are to:
1. Isolate polyhydroxyalkanoate (PHA) producing bacteria from palm oil mill effluent (POME) and environment.
2. Screen the isolates and select the best producer(s).
3. Produce PHA from POME.
4. Evaluate the optimum conditions for PHA production from POME by the selected strain(s).
5. Characterize and identify the selected strain(s).
6. Characterize the recovered PHA.

Petrochemical plastics are polymers of prime importance to human beings. These plastics have grown into a major industry and significant to human’s existence. According to Singh et al. (2015), petrochemical plastics are designed in such a way as to provide the continual performance and trustable qualities that ensure their long life-span, hence making them to be resistant to both natural and chemical breakdown. The persistence of synthetic plastics contributes to the adverse eco-hazardous impacts of these materials when they are disposed into the waste streams.
Over the years, the use of petrochemical-based plastics is affiliated with some issues such as the length of time required for decomposition in nature (de Paula et al., 2016), threat to wildlife existence (Yates and Barlow, 2013) and the toxins produced during the degradation process (Bharti and Swetha, 2016); these features have given petrochemical plastics a negative eco-image.
In our environment today, petrochemical plastics stand out as the major non-recyclable, non-biodegradable material on the planet (Bharti and Swetha, 2016). Plastic residues accumulate extensively in landfills and in terrestrial and aquatic ecosystems, which affect wildlife and human health (Aires da Silva, 2014). The large quantities and bright colors of plastic waste make them greatly detectable in waste streams and as litter (Yates and Barlow, 2013). In places where plastic wastes are discarded indiscriminately, they accumulate in drainages where they obstruct the flow of water; these plastic wastes also end up in landfills where they create a breeding ground for pests and pathogens.
The non-degradable nature of petrochemical plastics ensure that the problems associated with their disposal persists as long as the plastic wastes are present in the environment. Harshvardhan and Jha (2013) even described plastic debris as one of the prime contaminants of the marine environment. Current methods that have been adopted to manage solid plastic wastes (SPW) include recycling, disposal into landfills, energy recovery, gasification, pyrolysis and incineration; these methods have their drawbacks. In addition, some efforts have been made to degrade these synthetic plastics using suitable microorganisms (Harshvardhan and Jha, 2013).
Increasing cost and decreasing space of landfills makes disposal of solid plastic wastes into landfills a non-viable choice (Al-Salem et al., 2009). However, this practice is what is obtainable in most developing nations. Recycling, which is a preferred method of managing plastic wastes in developed countries owing to its benefits of reducing carbon dioxide (CO2) emissions, fossil fuel use and minimizing landfill deposition is faced with challenges such as difficulty in sorting of recyclable plastics thus involving extensive labor costs and production of plastics with altered quality (Chidambarampadmavathy et al., 2016).
Other challenges facing the adoption of other methods of managing SPWs include potential release of hazardous substances such as dioxins, polychlorinated biphenyls and furans into the atmosphere as in the case of incineration, and also huge cost of large scale operation associated with gasification/pyrolysis (Chidambarampadmavathy et al., 2016).
The enormous negative consequences associated with the use of synthetic plastics have kindled the search for a more ecofriendly polymer that can replace these synthetic plastics especially in terms of variety of applications. The continued clamor for biodegradable plastics with a green agenda have been on the increase since the increased awareness to fulfil the goals and objectives of green chemistry and environmental sustainability. The depletion of petrochemical reserves has also contributed to the special interest in the use of biopolymers such as bioplastics (de Paula et al., 2016).
As the natural environment is continuously polluted by these hazardous plastics, the development and production of eco-friendly biodegradable plastics is rapidly expanding in order to reduce our dependency on petroleum derived synthetic plastics.
According to Palm et al. (2016), the term ‘bioplastics’ often leads to confusion because it includes both plastics that are bio- and fossil-based. A plastic can be a bioplastic in three different ways; (1) bio-based and non-biodegradable, (2) bio-based and biodegradable or (3) fossil-based and biodegradable.
Biodegradable polymeric materials such as polynucleotides, polyamides, polysaccharides, polyoxoesters, polythioesters, polyanhydrides, polyisoprenoids and polyphenols have been described as potential replacements for synthetic plastics (Singh et al., 2015). However, amongst all the biodegradable polymers, PHAs have received greater attention owing to their thermoplastic and biodegradable properties.
PHAs are biodegradable polymers that accumulate naturally in bacteria via fermentation with materials such as vegetable oils, sugar or industrial waste in a culture medium (Hassan et al., 2013). They are also defined as biopolymers that are synthesized by microorganisms as lipid inclusions for energy storage in granular forms within the cellular structure (Raza et al., 2018; Rodriguez-Perez et al., 2018). PHAs serve as water insoluble storage compounds which are synthesized by microorganisms as granules during times of environmental stress conditions or imbalanced growth conditions such as limited essential nutrients and excess carbon sources (Hassan et al., 2013; Mo?ejko-Ciesielska and Kiewisz, 2016). These polyesters are stored within cells as energy storage materials by numerous microbes. PHA plays a role similar to that of fats in humans (Hiroe et al., 2016). Cruz et al. (2015) described PHAs as biodegradable and biocompatible polyesters; they can be extracted and then formulated and processed for plastic production (Rodriguez-Perez et al., 2018).
According to Pappalardo et al. (2014), over 150 different known monomers can be formulated with PHAs to give rise to material blends with tremendously different properties and characteristics depending on the side chain-length and on the substrate the synthesizing bacteria are fed with. The synthesis of PHAs in bacteria allows the potential stereospecificity essential to ensure the biodegradability and biocompatibility of the polymer (Guerra-Blanco et al., 2018).
Raza et al. (2018), described PHAs as “green plastics” that have positive social and environmental impact when put side by side with conventional plastics in terms of production and recycling. PHAs can be degraded by bacteria under aerobic condition into CO2 and H2O when disposed into waste streams (Singh et al., 2015).
PHAs have been used to replace petrochemical plastics in diverse areas. When used for in vivo application, PHAs do not possess acute and chronic health effects. Among PHAs, poly(3-hydroxybutyrate) (P3HB) and poly(3-hydroxyvalerate) (P3HV) have been described to be the most relevant for real applications (Guerra-Blanco et al., 2018).
Polyhydroxyalkanoates (PHAs) are a class of linear polyesters consisting of hydroxy acid monomers (HA) connected by an ester bond. This bond is produced by connecting the carboxylic group of a monomer with the hydroxyl group of a neighboring one (Philip et al., 2007).

Figure 1: General structure of Polyhydroxyalkanoates
Where n = 1 to 4, x =100 to 300,000 while R is the alkyl side chain, C1 – C13
(Mo?ejko-Ciesielska and Kiewisz 2016).

R group Carbon number PHA polymer
methyl C4 Poly(3-hydroxybutyrate)
ethyl C5 Poly(3-hydroxyvalerate)
propyl C6 Poly(3-hydroxyhexanoate)
butyl C7 Poly(3-hydroxyheptanoate)
pentyl C8 Poly(3-hydroxyoctanoate)
hexyl C9 Poly(3-hydroxynonanoate)
heptyl C10 Poly(3-hydroxydecanoate)
octyl C11 Poly(3-hydroxyundecanoate)
nonyl C12 Poly(3-hydroxydeodecanoate)
decyl C13 Poly(3-hydroxytridecanoate)
undecyl C14 Poly(3-hydroxytetradecanoate)
dodecyl C15 Poly(3-hydroxypentadecanoate)
tridecyl C16 Poly(3-hydroxyhexadecanoate)

Depending on the number of carbon atoms in the monomers, PHAs are classified mainly into two distinct groups: short chain length PHAs, (PHASCL) and medium chain length PHAs, (PHAMCL). Short-chain-length PHAs (PHASCL) contain 3-hydroxy acids ranging from 3 to 5 carbons, while medium-chain-length PHAs (PHAMCL) contain 3-hydroxy acids ranging from 6 to 16 carbons. Majority of PHASCL are very rigid and brittle biothermoplastics, owing to their high crystallinity (50–70%), while PHAMCL are more elastic and/or viscous materials, characterized by low crystallinity degrees, glass transition temperatures and melting temperatures (Cruz et al., 2015; Yogesh et al., 2012). Owing to these favorable features, PHAMCL and their copolymers are currently appealing the attention of industrial and, particularly, biomedical applications where flexible biocompatible biomaterials are vital (Pappalardo et al., 2014).
According to Raza et al. (2018), the disparity in the properties and chemical compositions of PHAs can be attributed to the variations in the structure of the monomers contained in PHAs. Based on the branched length of the monomers or the distance between the ester linkages in the polymer backbones, Guerra-Blanco et al. (2018) stated that the mechanical properties of PHAs range from brittle to flexible and elastic.
Chen (2009) stated that the molecular weights of PHA are dependent on PHA production strains and the N-terminus of PHA synthase. The physical properties of PHB varies with the molecular weight of the PHB, which is influenced by the microorganism used in the production, growth conditions, and the purity of the sample obtained. Some of these physical properties include crystallization, and tensile strength (Singh et al., 2015).
Some properties of PHAs that enhances its anaerobic biodegradation in sediments include hydrophobicity, good resilience to hydrolytic attack, and resistance to ultraviolet (Raza et al., 2018).
PHAs amass as intracellular inclusions or granules (50–500nm) conceivably shielded by phospholipid monolayers, the surface of which contains protein gears of the PHA metabolic machinery. The considerable number of proteins on the surface of PHA granules suggests that they represent supramolecular complexes with specific functions, rather than being simple carbon and energy stockpiling systems produced during periods of nutrient imbalance (Prieto et al., 2015). PHAs have commonly been described as storage materials for carbon and energy in bacteria (Mo?ejko-Ciesielska and Kiewisz, 2016). However, they also serve as sinks for reducing equivalents for some microorganisms. Due to their insoluble nature inside the bacterial cytoplasm, PHAs serve as ultimate storage compounds, which exert insignificant increase in osmotic pressure (Singh et al., 2015). Using P. putida, Prieto et al. (2015) described the concept of the multifunctionality of PHAs using the term ‘carbonosomes’ in figure 2.
PHA synthases (PhaC), depolymerase (PhaZKT), phasins (PhaF, PhaI) and acyl-CoA synthase – together referred to as granule-associated proteins (GAPs) – have been identified as carbonosome components in P. putida and are all involved in PHA metabolism and granule formation.
Studies have shown that bacteria capable of synthesizing PHAs as storage materials can subsist periods of famine compared to those incapable of using PHAs as their energy-reserve material. This is because PHAs decelerates the rate of cell autolysis and hence ensures the mortality of the bacteria containing it (Singh et al., 2015).
In their review, Prieto et al. (2015) described PHA (granules) as supramolecular complexes of biopolyesters and proteins which are vital for granule segregation during cell division, and for the functioning of the PHA metabolic route as a continuous cycle. They further stated that PHA cycle also determines the number and size of bacterial cells.

Figure 2: PHA granules location and its segregation during cell division in P. putida KT2440.

A. The PHA granule or carbonosome is a multifunctional complex formed by all enzymes and proteins involved on the PHA machinery.

B. The PHA granules are located lengthwise the cell, forming the characteristic needle array structure in the wild type P. putida KT2440.

C. The lack of PhaF phasin results in a different phenotype; the PHA granules in P. putida KT40F mutant strain remained accumulated in one of the cell poles.

D. Transmission electron microscopy (TEM) image of wild type strain population. The segregation is homogeneous in the wild type strain and the previously formed PHA granules are distributed to the daughter cells keeping the needle array organization.

E. TEM image of PhaF mutant cell population. The granules in the mutant strain remain agglomerated in one of the daughter cells resulting in a heterogeneous population.

F. The impact of cells containing PHA granules in the microbial community due to extracellular depolymerase activities. The extracellular PHA consumption implies an advantage for other microorganisms in the niche providing them an extra source of carbon and energy.

PHA was first discovered in 1926 by Lemoigne while studying Bacillus megaterium. The metabolism of PHA is not a unidirectional metabolic process in which the PHAs are either polymerized or depolymerized, but a bidirectional dynamic process in which there is a continuous cycle of synthesis and degradation. According to Prieto et al. (2015), this mechanism helps to adapt the carbon flow to the transient demand for metabolic intermediates, thus balancing carbon resources, ensuring optimal growth under changing environmental conditions.
According to Chen and Jiang (2018), the PHA biosynthetic pathways and their related enzymes have been extensively studied, and expressly linked to the various pathways leading to the formation of numerous PHA monomers. Depending on the carbon source being fed to the bacteria, there are three possible pathways that lead to the synthesis of PHAs (Figure 3.) (Chem et al. 2015). Prieto et al. (2015) stated that there are two major enzymes whose activities are key to the PHA cycle; these enzymes are PHA synthase and PHA depolymerase. Stating further, they buttressed the essence of PHA synthase and depolymerase adding that the synchronized actions of PHA synthase and depolymerase ensures the carbon flow to the transient demand for metabolic intermediates to balance the storage and use of carbon and energy.
The pathways for the biosynthesis of PHA differs because of two reasons which include carbon source(s) and the PHA product(s) of choice (Singh et al., 2015). PHA biosynthetic pathways are intricately linked with the bacterium’s central metabolic pathways. Acetyl-CoA generated from these central metabolic pathways is used to generate 3-hydroxybutyryl-CoA which is the substrate for the enzyme PHA synthase involved in PHA synthesis (Singh et al., 2015). The biosynthesis of PHA in most bacteria is initiated by the condensation of two molecules of acetyl-CoA by 3-ketothiolase to form aceto-acetyl-CoA, which is reduced in an enantiomerically selective reaction by aceto-acetyl-CoA reductase to R-(-)-3 hydroxy butyryl-CoA and is incorporated into PHA by PHA synthase. The acetoacetyl-CoA reductase involved in the synthesis is generally considered to be specific for NADPH, although in some bacteria the enzyme has been reported to have activity with NADH as well as NADPH (Kalia et al.,2000). Three metabolic phases of the biosynthesis of PHA have been described by Kalia et al., (2000):
Phase 1: A carbon source suitable for biosynthesis of the PHA enters the cell from the environment.
Phase 2: Anabolic or catabolic reactions – or both, in which the compound is converted into a hydroxyacyl coenzyme A thioester which is a substrate of the PHA synthase.
Phase 3: The PHA synthase catalyzes the formation of the ester bond with the concomitant release of coenzyme A.
PHA synthase and PHA depolymerase function in the synthesis and depolymerization of PHA polymers respectively. With regards to PHA depolymerization, physical and chemical parameters in the immediate environment influence the degree of PHA depolymerization. The depolymerization of PHAs produces simple monomers, oligomers and free fatty acids, wherein the bacteria consequently assimilate the intermediates and end-products of the depolymerization for their subsequent growth (Sathiyanarayanan et al., 2016). Many of the known PHA synthases are either PHASCL synthases or PHAMCL synthases. According to Chen et al. (2015), only a few synthases have been reported to be capable of polymerizing both short-chain and medium-chain length PHA monomers; such enzymes are termed low-specificity PHA synthases. Two representatives of the low-specificity PHA synthases are PhaC61-3, first reported by Doi and colleagues, and PhaC21317 described by Chen et al. (2015), and both were found in Pseudomonas spp. The specificity of the PHA synthases can also be manipulated by molecular evolution or by chimera formation.
According to Prieto et al. (2015), PHA accumulation involves:
(i) indirect PHA precursor routes that link the catabolism of other carbon sources (non-PHA related) to fatty acid and PHA metabolism;
(ii) central pathways such as fatty acid ß-oxidation and de novo fatty acid synthesis; these provide direct PHA precursor routes that convert fatty acid or acetyl-CoA from non-PHA-related intermediates, respectively, into different (R)-3-hydroxyalkanoyl-CoAs, the substrates of PHA synthases associated with the carbonosome; and
(iii) specific PHA metabolism (the PHA cycle) encoded by the PHA cluster.

Figure 3: PHA biosynthesis pathway (Mo?ejko-Ciesielska and Kiewisz 2016).

Both prokaryotic and eukaryotic microorganisms can produce different types of PHAs. Different bacteria produce different types of PHAs. Fluorescent Pseudomonas strains, for example, are well known to accumulate mcl-PHAs as they have mcl-PHA synthases for the synthesis of PHAs with 6–14 carbon atoms (Raza et al., 2018). PHA-synthesizing microorganisms can be found in almost all imaginable habitats and ecological niches such as marine microbial mats, estuarine sediments, the rhizosphere, groundwater sediments and ecosystems of anthropogenic origin. These habitats which can be described as hotspots for PHA producers are characterized by fluctuating nutrient contents, e.g. waste water treatment plants (Koller et al., 2015). In addition to these, another niche of PHA-synthesizing organisms is engineered ecosystems with fluctuating nutrient contents to meet the metabolic demands of PHA-synthesizing organisms during starvation periods (Saharan et al., 2014).
Several bacteria, isolated from natural habitats, have been tried likewise, many have been re-constructed to produce PHAs. These include various species of Pseudomonads belonging to rRNA-DNA homology group I, which are known as microbial producers of PHAs built from hydroxyl-acyl-CoA derivatives via different metabolic pathways of the bacterial fermentation of sugar or lipids (Pappalardo et al., 2014).
The capacity of synthesis or catabolism of PHAs has been described in species belonging to three major kingdoms viz Archaea, Bacteria and Eukaryotes. These include organisms present in a wide variety of habitats, including free-living species, parasites, symbionts and predators, and they might be aerobics or anaerobics, and chemotrophs or phototrophs (Prieto et al., 2015). The following genera of microorganisms have been identified as PHA producers Cupriavidus, Methylobacterium, Rhodopseudomonas, Pseudomonas, Alcaligenes, Rhodobacter, Enterobacter, Burkholderia, Chelatococcus, Azotobacter, Aeromonas, Sinorhizobium, Comomonas, Thermus, Rhodopseudomonas, Corynebacterium (Sathiyanarayanan et al., 2016).
The ecological niches of these PHA synthesizing microorganisms are characterized by natural or accidental exposed to elevated levels of organic matter. These niches can also be characterized by growth-limited conditions such as can be seen in dairy wastes, hydrocarbon contaminated sites, pulp and paper mill wastes, agricultural wastes, activated sludges of treatment plants, rhizosphere, and industrial effluents (Saharan et al., 2014).
Hitherto, more than 90 genera of both Gram-positive and Gram-negative bacteria have been discovered to be capable of PHAs production under both aerobic and anaerobic conditions (Raza et al., 2018). Many microbes can store intracellular inorganic and/or organic inclusions which are surrounded by phospholipids. When microorganisms are faced with stress conditions such as carbon: nitrogen: phosphorus unbalance, microorganisms store energy via the synthesis of PHAs (Guerra-Blanco et al., 2018). The prevalent ability to metabolize PHAs, plus the fact that ‘pha genes’ have been horizontally transferred between different phylogenetic groups is an indication that PHAs afford some advantage to the microorganisms that synthesize them (Prieto et al., 2015). Singh et al. (2015) reported that some bacteria are capable of producing PHAs up to 90% (w/w) of dry cells weight during the reduction of essential nutrients like nitrogen, phosphorus or magnesium.
Many bacterial species can synthesize polyhydroxylbutyrate (PHB), but medium-chain-length PHAs are chiefly, though not solely, synthesized by fluorescent pseudomonads such as P. putida (Prieto et al., 2015). According to Chen (2009), the most commonly used wild strains for industrial production of PHAs are Ralstonia eutropha (formerly called Alcaligenes eutrophus, Wautersia eutropha, or Cupriavidus necator). These microorganisms have been used for the production of poly-(R)-3-hydroxybutyrate (PHB), poly((R)-3-hydroxybutyrate-co-4-hydroxybutyrate) (P3HB4HB) and poly((R)-3-hydroxybutyrate-co-(R)-3-hydroxyvalerate) (PHBV). The disparity in the ability of different bacterial species to synthesize varieties of PHAs as opined by Prieto et al. (2015) is based on differences in the in vivo substrate specificity of PHA polymerase or synthase the enzyme responsible for the assembly of PHA monomeric precursors (R)-3-hydroxyacyl-CoAs and the specialization of metabolic and regulatory networks in each species.
According to Prieto et al. (2015), the most widely studied reference medium-chain-length PHA producers are P. putida KT2440 (and its rifampicin-resistant mutant KT2442) and Pseudomonas oleovorans GPo1 (ATCC 29347). Koller et al. (2015) also stated that rapid and reliable tracing of PHA in natural habitats is essential to discover novel, powerful PHA-producing organisms which are adapted to extreme environments or to unusual substrates.

Current industrial production of PHAs make use of pure cultures of selected strains (e.g. Cupriavidous necator) and of ad-hoc designed unbalanced growth media (e.g. glucose and propionic acid in a nitrogen-poor mineral medium) thus making PHA production expensive, mainly because of the associated costs of culture maintenance, substrate formulation and both substrate and reactor sterilization (Villano et al., 2013).
In comparison to MMCs, the use of monoseptic microbial cultures is associated with high expense in terms of cost due to major sterility demands, and sophisticated requirements for equipment and control devices (Koller et al., 2015).
The use of MMCs for PHA production is based on the principle of natural selection and competition (‘evolutionary engineering’) instead of genetic or metabolic engineering. According to Koller et al. (2015), this is accomplished by exercising selective pressure for a desired metabolism on a microbial consortium by choosing appropriate feeding and cultivation conditions in the bioreactor. This way, the ecosystem and not the microbes are engineered.
Koller et al. (2015) in their work outlined the comparative benefits of MMCs over monoseptic cultures in PHA production. These benefits include enhanced capacity to amass PHA at low cost, ease of adaptation to inexpensive, complex substrates such as agro-industrial and domestic waste streams. Furthermore, they stated that auspicious results in terms of yield and specific productivity can be obtained by supplying MMCs either with pure substrates like acetate, or with complex waste streams as low-cost substrates.
Though not at the industrial scale yet, the production of PHA using mixed microbial cultures (MMCs) appears promising because it does not require the maintenance of sterile conditions and it makes easier the use of low-cost feedstocks, such as agro-industrial waste effluents. The MMC-based PHA production process brings the advantage of simultaneously reducing the polluting load of the waste stream and requires different stages which are strictly interconnected (Villano et al., 2013).
Despite the comparative advantages of MMCs approach over the use of monoseptic cultures, Koller et al. (2015) remarked that the using MMCs will not be a better option if the production of highly uniform PHA is desired. Elucidating their position, they added that MMCs are composed of varied PHA-producing species each of them accumulating PHAs of varying molar masses, molar mass distribution, crystallinity, and monomeric composition.

The use of genetic engineering was described by Koller et al. (2015) as a veritable tool for the optimization of the biosynthetic performance of PHA producing strains. The objective of the use of genetic engineering for strain improvement is the transfer of genetic information for enhanced PHA production from eminent PHA accumulating strains to non-or less efficient-PHA producing strains. However, Kaur and Roy (2015) stated that an essential factor to consider when dealing with recombinant organisms is the issue of the stability of the plasmids of the organisms. This is important because the capability of the engineered strain to be used in industrial production over several runs of fermentation (production cycles) without losing its function has to be established.
According to Chen et al. (2015), in silico metabolic modeling can been applied in improving the PHA production capacity of an organism. Citing an example, Chen et al. (2015) explained that PHA titer of a strain of P. putida was improved by 100% using metabolic modelling.

Singh et al. (2015) outlined some important factors that influence the commercial production of PHAs. These factors include (i) the selection of proper bacterial strains (ii) cheap carbon sources (iii) proficient fermentation process and (iv) recovery techniques used in downstream processing.
According to Guerra-Blanco et al. (2018), the monomeric composition and the yield of PHAs synthesized by microbes can be controlled by taking advantage of two important factors that influence the culture conditions. These factors are substrate composition and pH.
Several process parameters are important for the efficient PHA production. These factors influence the production of PHA individually and interactively, either positively or negatively. These factors include pH, microenvironment, substrate concentration, nitrogen, phosphorous, iron concentration and VFA (Mohan et al., 2013). Microenvironment
There are three microenvironments to be considered which are aerobic, micro-aerophilic and anaerobic environments. According to Mohan et al. (2013), microenvironment has the greatest influence on PHA biosynthesis. During PHA production and many other strictly aerobic bioprocesses, gas-to-liquid mass transfer of oxygen is often rate limiting and has a substantial influence on the overall productivity (Blunt et al., 2017). Micro-aerophilic condition was considered by Mohan et al. (2013) to be the optimal microenvironment condition that favors PHA synthesis. There is dissolved oxygen (DO) concentration which reduces ATP availability thus preventing biomass accumulation. The low DO suppresses the assimilative activities needed for the growth and induces cell accumulation of PHA. PHB is favored by high ADP and NADH level which is abundant at low DO concentration. Reddy and Mohan.(2012) documented increased PHA accumulation in fruit waste substrate under microareophilic condition in comparison to aerobic condition.
Aerobic microenvironment is characterized by high DO concentration which has been noted to drive microorganisms towards biomass accumulation, protein and cellular component production, high level of ATP and NAD+ favours higher growth rate and a subsequent decrease in PHB in aerobic condition (Mohan et al., 2013). pH
This is another factor that influences PHA biosynthesis strongly. Mohan et al. (2013) suggested neutral pH to be the optimal PH for efficient synthesis as most enzyme needed for the biosynthesis are active at neutral pH than at acidic and alkaline pH. Substrate concentration
PHA biosynthesis is always directly proportional to carbon/substrate concentration up to certain concentration at which it becomes inhibitory. This relation/behaviour is caused by substrate inhibition as high carbon/substrate lead shock (Mohan et al., 2013). Also, type of substrate used influences the quality and quantity of PHA produced. Nutrient concentration
The major nutrient investigated and varied in PHA biosynthesis are Carbon, Nitrogen and Phosphorous. PHA accumulation varies inversely as Nitrogen and Phosphorous concentration. High level of nitrogen and phosphorous encourages protein synthesis thus accumulation but low Nitrogen and Phosphorous triggers PHA synthesis as cells are stressed. Volatile Fatty Acid (VFA) concentration and composition
VFA are volatile fatty acid which are short chain fatty acid of 2 to 6 carbon atoms produced by the breakdown (fermentation) of carbohydrate. There are 3 most common VFA which are acetate, propionate and butyrate. VFA concentration is directly proportional to the PHA production, also, composition of VFA influences the type of PHA produced (Mohan et al., 2013). The type of organic acids obtained during the acidogenic fermentation influences the monomeric composition of the final PHAs and subsequently determines the physical and mechanical properties of the final PHA (Villano et al., 2013).

An essential aspect of any fermentation process is the appropriate medium composition that would encourage biomass growth and metabolite production (Kaur and Roy, 2015). PHAs are commercially produced by fermentation using PHA synthesizing microorganisms. According to Daly et al. (2018), the most common microorganisms employed in the commercial production of PHAs are Ralstonia eutropha or genetically modified Escherichia coli. The production of PHB as an energy storage compound can be triggered within these microorganisms by supplying carbon substrate in excess (such as sugar, oil, or agricultural waste) concurrent with a shortage of nitrogen or phosphorus (Daly et al., 2018).
The cost involved in PHA production is very much higher when compared to the production cost of petrochemical plastics. As earlier mentioned, this factor still presents as a major drawback in the commercialization and industrialization of PHAs (Mo?ejko-Ciesielska and Kiewisz, 2016). The common substrates used in the production of PHAs are simple and pure carbon sources, such as sugars (e.g. glucose or sucrose). However, the high price of those feedstocks accounts for up to 50% of the overall PHA production costs (Cruz et al., 2015). Hence utilizing renewable feedstocks such as complex wastewaters and agro-industrial wastes, has been explored as a means to ameliorate the high cost associated with PHA production.
Conventional carbon sources used for PHA production are of pure raw materials, comprising of pure carbohydrates (glucose, sucrose, maltose, starch), fatty acids and its derivatives, methanol and alkanes (Aslan et al., 2016). Jiang et al. (2016) explained that the substrates for PHA production are primarily derived from food-based carbon sources, raising concerns over the sustainability of their production in terms of their impact on food prices. Thus, to make the PHA production economically feasible and sustainable, Aslan et al. (2016); Ivanov et al. (2015); Hassan et al. (2013) suggested the use of cheap and renewable biomasses such as horticultural agro-wastes as alternative carbon sources. Kesharvarz and Roy (2010) however included that whatever cheap media to be used for PHA production, should contain the necessary requirements to encourage high productivity of PHA.
Considering the abundance, diversity and lack of proper treatment approach, Amulya et al. (2015) described the use of waste/wastewater for PHA production as a propitious approach, which provides dual benefits of waste remediation with simultaneous value addition. Since PHAs can be produced from renewable carbon resources, Pappalardo et al. (2014) stated that PHAs can play a positive role in agriculture, the environment and the economy, by contributing to the preservation of finite fossil resources, such as mineral oil and coal, and are neutral with regards to CO2 emissions.
Chen (2009); Koller et al. (2015) further opined that for improved large-scale production of PHA, variety of biowastes can be used to lower the production cost. Moreover, Chee et al. (2010); Koller et al. (2015) also agreed that utilizing waste materials for PHA production saves cost of waste disposal.
Despite the benefits of utilizing various waste streams to improve the economics of PHA production, Yang et al. (2011) outlined the questionability of the waste streams being optimal for a sustainable biological polymer production. Difficulty in control of production processes and presence of harmful components (which could inhibit bacterial growth and in turn limit PHA production) as drawbacks in the use of biowaste streams for PHA production. Stating further, Yang et al. (2011) added that difficulty of control of production processes and lack of reproducibility are also factors associated with using waste materials for the production of PHA.
According to Hermann-Krauss et al. (2013), in addition to adopting a suitable inexpensive raw material, proficient, stable and rapid-growing microbial PHA production strains are also needed to ensure an optimum PHA production and thus improve the economics of production. Chee et al. (2010) also added that microorganisms proficient at producing PHAs are capable of utilizing waste biomass as inexpensive carbon sources.
Plethora of researches have reported the use of low cost substrates such as cardboard paper (Bhuwal et al., 2013), oil palm biomass (Hassan et al., 2013), food waste (Amulya et al., 2015; Reddy and Mohan, 2012), organic fraction of municipal solid waste (Ivanov et al., 2015), spent wash effluents (Amulya et al., 2014), paper mill wastewater (Jiang, 2012), waste vegetable oil (Hwan et al., 2008), Zhu et al. (2013) for PHA production. Excluding conventional carbon sources, cheaper carbon sources such as fruit pomace, waste frying oil, and animal-derived waste have been employed in the production of PHAMCL. Heinrich et al. (2016) even stated that exhaust gases and gases arising from gasification and pyrolysis of organic waste referred to as synthesis gas (syngas) have been explored as potential substrates for PHA production. He further stated that using these exhaust gases for industrial production of PHAs promotes environmentally friendly and cost-efficient production processes for these so called second generation biopolymers. However, so far only PHASCL have been synthesized from these gaseous feedstocks, as there is no known microorganism capable of synthesizing PHAMCL from syngas. Yang et al. (2011) opined that there is need to treat the waste materials to be used as substrate in order to generate more desirable feedstock for production. PHA Production from Plant Oils
Plant oils can serve as sustainable carbon source for PHA production (Walsh et al., 2015). According to Younas et al. (2015), plant oils such as soybean, sunflower, palm, canola, cartamo, soja, maize, and olive oil have been shown to be excellent inexpensive carbon sources for PHA production. Walsh et al. (2015) also added that the product of hydrolysis of plant oils can also be used as carbon source for PHA production. When compared to processed sugars and fatty acids, plant oils are cheaper carbon sources for PHA production (Younas et al., 2015; Chee et al., 2010). Besides being inexpensive, plant oils have been shown to produce higher PHA yields per gram of oil in comparison with sugars (Sudesh et al., 2011). Chee et al. (2010) also added that besides providing higher PHA yield of 0.6 to 0.8g of PHA per gram of oil, plant oils also yield higher cell biomass when compared to that obtained from sugars. This is because plant oils contain a complex mix of triglycerides (Sudesh et al., 2011) and as well plant oils have higher carbon content per weight compared to sugars (Chee et al., 2010).
Vegetable oils, waste streams from oil mills or used oils, which are cheaper than purified oils can serve as substrates for PHA production by microbial fermentation owing to their complete esters of glycerin and higher monocarboxylic acids content (Ciesielski et al., 2015; Verliden et al., 2011). Corn oil as sole source of carbon was used by Hwan et al. (2008) for PHB production, they recorded a 37.34% (w/w) of intracellular PHA. According to Panadare and Rathod (2015); Aslan et al. (2016), cooking oils, specifically waste cooking oils have been regarded as potential substitute carbon feedstock for PHA production. Although cost-effective, use of virgin oils gives higher PHA yield compared to waste oils. They attributed this to the presence of impurities in the waste oils. According to Panadare and Rathod (2015), thermal reactions occurring during the frying accounts for the difference in composition of waste cooking oil and vegetable oil. Because of the chemical reactions that occur during frying at elevated temperature and prolonged periods, waste cooking oil generally consists of 70% triacylglycerol TAG (fresh cooking oil consists of 95% TAG) while the remaining fraction consists of oil degradation products (Sudesh et al., 2011). Aslan et al. (2016) further identified triacylglycerides (TAGS) from the fatty acids of the waste oil as valuable alternative feedstock for the production of medium chain length-PHAs long chain length-PHAs.
Verliden et al. (2011) compared the use of waste vegetable frying oil and pure vegetable oil for the production of PHB. They reported from their observation that more biopolymer production was achieved using waste frying oil than pure vegetable oil. The findings of Verliden et al. (2011) differed greatly from that of Aslan et al. (2016). In another study done by Morais et al. (2014) using fat-containing waste produced from the margarine manufacturing process for PHA synthesis, margarine waste was shown to be a promising substrate for P(3HB) synthesis. In a similar study involving utilizing fatty wastes as substrate for PHA production, pure fatty acids were shown to give very high biomass and PHA yields when used as single substrates for PHA production (Walsh et al., 2015). Likewise, Povolo et al. (2012) acknowledged that fatty wastes from animal and vegetable origin could serve as viable substrates for PHA production. Younas et al. (2015); Lopez-Cuellar et al. (2011) studied PHA production using canola oil as carbon source. Their investigation showed that canola served as a good carbon source for PHA synthesis. Lopez-Cuellar et al. (2011) obtained 92% PHA on dry cell weight bases.
Plant oils like palm oil, soybean oil, sunflower oil etc. are preferred to sugars as sole carbon source for PHAs production as they are cheaper (Abid et al., 2016). Although plant oils produce higher quantity of PHAs when used as carbon source as stated by Abid et al. (2016), Younas et al. (2015) in their work observed that glucose as carbon source gave greater yield to canola oil. However, Younas et al. (2015) attributed this to the simple and easy metabolism associated with glucose as compared to oils. Panandare and Rathod (2015) stated that the use of waste cooking oil for PHB production gave the same yield of 1.2g/L which is the same when using glucose as carbon source. PHA Production from Lignocellulosic Biomass
Depending mainly on the global region of their occurrence, a vast number of highly interesting sources of lignocellulosic materials can be found, e.g. rice straw, corn straw and bagasse. These materials have favourable compositions with rather low amounts of lignin in the range of 10% (w/w) and a high percentage of carbohydrates. To implement the concepts of ‘refineries’ starting from lignocellulosic waste, the utilization of the entire plant has to be aspired to. Rice husks, e.g., contain high amounts of silicon dioxide that can be utilized for production of silicon. Since starch, sucrose and lactose are a major food source for humans and animals, derivation of glucose and other simple sugars from lignocelluloses has received great attention (Jiang et al., 2016). Lignocellulose contains about 40%–50% cellulose, 20%–50% hemicelluloses and 20%–30% lignin (Jiang et al., 2016). However, lignin is not fermentable due to its complex aromatic nature. Since lignocellulosic materials have been recommended as potential substrates for low-cost PHB production, only a few studies have utilized this biomass for PHB production.
Prior to the use of lignocellulosic biomass, delignification is of essence to improve the bioavailability of the usable sugars present in the lignocellulosic biomass. The pretreatment methods commonly employed include dilute acid hydrolysis, steam-explosion, hydrothermal treatment, lime treatment and ammonia treatment (Jiang et al., 2016). Davis et al. (2013) in their study subjected grass biomass (a lignocellulosic substrate) to 2% NaOH at 120oC and hot water at 120oC as a way of pretreating the biomass prior to PHA production. This resulted in improved digestibility and glucose concentration. Zhang et al. (2013) used oil palm empty fruit bunch (OPEFB), a lignocellulose biomass, as carbon source and Bacillus megaterium R11 in the production of PHB. They reported a very high PHB production of 9.31 g/L and 12.48 g/L for OPEFB hydrolysate containing 45 g/L and 60 g/L sugar, respectively, both higher than what has been obtained from previous studies involving lignocellulosic biomass. They credited their high PHB yield to the high biomass loading hydrolysis they adopted in their work.
Paper mill wastewater, a lignocellulosic material has been used as a substrate for PHA production. The common composition of paper mill wastewater is cellulose and hemicellulose (Jiang, 2012). Munir et al. (2015) reported the use of paper mill wastewater as a cheap carbon source for PHA production. They however acknowledged that glucose encouraged PHA production and rapid growth of PHA producers than paper mill wastewater.

According to Koller et al. (2015), when selecting media for PHA production, such media should fulfill the following criteria;
? It should be inexpensive
? It should be available in sufficient amounts
? It should have constant quality during the whole year
? It should not be endangered during off-season of harvest (“off-season availability”).
? It should be as invariable as possible in its composition,
? It should be stable, and resistant against rapid microbial spoilage.
In choosing a media for effective PHA production, Kesharvarz and Roy (2010) suggested that the type of microorganism to be used for production either wild type or recombinant and whether it needs nutrient limiting condition should be put into consideration. Chanprateep (2010) in addition stated that the selection of media for PHA production should not focus only on market prices but also on availability and global price consistency. For waste products to be used massively, the consistency and reliability of the raw material, storage issues and the correct balance of the ingredients will need to be critically considered (Kesharvarz and Roy, 2010). Anaerobic fermentation of wastes generates effluents that are rich in VFAs viz., acetic acid, butyric acid, propionic acid and valeric acid that can serve as source of carbon for the growth of aerobic bacteria and their survival during carbon depleted conditions by storing these VFA as PHAs (Amulya et al., 2015). The feasibility of PHA production from the effluent of sugar factories, oil mills, wood mills, paper mills or municipal wastes has been studied (Jiang, 2012). Although these studies have suggested the use of agro-wastes and industrial by-products as substrates, Chanprateep (2010) opined that the use of these waste materials may incur additional production cost. Chanprateep (2010) buttressed this point by attributing the additional costs to pretreatment steps, extended cultivation times, and purification. Jiang (2012) also stated that the PHA storage capacity obtained by utilizing waste as raw material was still significantly lower than the microbial enrichments selected on synthetic substrates. Consistency and reliability of the raw material, storage issues and the correct balance of the ingredients are also some more factors hampering the extensive use of waste biomass for bioplastic production (Kesharvarz and Roy, 2010). There is also the issue of insufficient data about the market potential of biowastes (Morone et al., 2015). The availability of cheap carbon source at the biorefinery location is one of the important factors influencing the selection of biomass as a carbon source (Zahari et al., 2015).

Figure 4. Schematic illustration of factors impacting sustainability of PHA production (Koller et al., 2017).

Following biosynthesis, cells are disrupted and PHB is recovered from the fermentation media. Several studies have projected several recovery techniques that improve the yield and purity of extraction and as well lessen cost of production. Many researchers have remarked DSP as playing a critical role in the overall production of PHAs. The DSP of PHAs is very much associated with high cost because of the materials employed in the recovery and purification of PHAs from the fermentation broth (Daly et al., 2018).
According to Aramvash et al. (2015), the most common DSP techniques are classified as chemical, physical, and enzymatic methods, as stand-alone extraction techniques or a combination of such techniques. Among these procedures, solvent extraction is a common industrial method of recovering PHB that features high efficiency, low degradation of the biopolymer, and elimination of endotoxins from the recovered biopolymer. The easiest process of extracting PHAs is employing solvents, such as chloroform, dichloromethane, dichloroethane, ethylene carbonate or acetone. These solvents first make the cell membrane more permeable and then extracts the PHAs. The polymers are then concentrated and precipitated with methanol or ethanol
Fernandez-Dacosta et al. (2015) stated that employing organic solvents for the DSP of PHAs leads to enhanced purity and recovery yield; the quality of the PHB obtained from solvent-based DSP is comparable to that of a commercial polymer.
Judging from an environmental and economic point of view, Fernandez-Dacosta et al. (2015) stated that the employing chemical digestion in DSP appears to be more attractive than the solvent-based approach. However, organic solvents can pose adverse environmental impacts if the solvent is not completely recovered due to the excess chemicals being released into the environment.
Most organic solvents used in the DSP of PHAs are halogenated solvents (e.g chloroform). Aramvash et al. (2015) supported the position of Fernandez-Dacosta et al. (2015) on the deleterious effects of these solvents on the environment. As an alternate method, Aramvash et al. (2015) opined that a simple, practical, efficient, and cost-effective DSP technique employing non-halogenated solvents should be developed; this will be a step further in improving the applicability of PHAs especially PHB in the biomedical sector (Daly et al., 2018). However, halogen-free solvent-based techniques require improvement of the parameters that influence the overall process to make it suitable on an industrial scale (Aramvash et al., 2018). Examples of these non-halogenated solvents that have been remarked as suitable for DSP are ethyl acetate and butyl acetate are non-toxic solvents.

Currently, there has been heightened attention of both the public and private sector towards the production of PHA (Palm et al., 2016). Efforts have been made and are still being investigated to improve the competitiveness of bioplastics over petrochemical plastics. Many researches hitherto have acknowledged the high cost of PHA production. Examining the techno-economic analysis of PHA production, Fernandez-Dacosta et al. (2015) stated that there are three main factors that the high cost of PHA production can be attributed to. These factors are; (i) the energy input for the sterilization of the fermentation equipment (ii) the PHA yield on the substrate, and (iii) the efficiency of the downstream processing (DSP).
To lessen the cost of PHA production, Chen (2009) proposed the use of genetic engineering technology, pathway modification or even synthetic biology approaches. These will help to develop super PHA production strains that can grow to high cell density within a short period of time utilizing cheap substrates under less demanding fermentation conditions, such as micro-aerobic conditions.
Besides developing versatile strains as a means of lessening the manufacturing cost of PHAs, Fernandez-Dacosta et al. (2015), identified the downstream processing of PHAs as another area that could be developed to reduce the cost of production. In this light, Fernandez-Dacosta et al. (2015), stated that the development of a competitive DSP facility that could release the intracellular PHAs synthesized by the microorganisms is needed to further reduce manufacturing costs; this will also go a long way in reducing the adverse environmental impacts of the organic solvents used for the downstream processing of PHAs.

PHAs are exciting materials with interesting applications in diverse fields. According to Wang et al. (2017), PHAs have been considered as sustainable materials for applications in versatile areas of packaging plastics, medical materials, drug matrices, biofuels, and food additives. Currently, PHAs are used as substitutes to commercial petrochemical polymers in various industries to prepare foils, films, molded goods and paper coatings (Sathiyanarayanan et al., 2016). Stating further, they added that PHAs can also be utilized for making day-to-day objects such as cardboards, milk cartons, moisture blocks in diapers, hygienic cloths, cages, adhesives, non-woven fabrics, shampoo bottles and cosmetics. Figure 5 shows the various potential applications of PHAs in everyday life.

Figure 5. Applications of PHAs in various sectors. (Masood et al., 2014).

Several studies have shown that PHAs are promising implant materials owing to their assorted and dominant mechanical, biodegradable and tissue compatible properties. Importantly, common PHA biodegradation products including oligomers and monomers are also not toxic to the cells and tissues. Therapeutic applications of PHA includes medical implant such as heart valve tissue engineering, vascular tissue engineering, bone tissue engineering, cartilage tissue engineering, nerve conduit tissue engineering as well as drug delivery carrier matrix (Junyu et al., 2017).
PHAs have also been employed in drug delivery systems. PHAs have been used as glaucoma drainage implants for orthopaedic post-operative infections. Using these biodegradable polymers in glaucoma drainage implants has helped to overcome the main limitation in the use of glaucoma drainage implants which is the generation of fibrotic tissue response (Nigmatullin et al., 2015).
The importance of good packaging materials in the food industry cannot be overemphasized. Packaging provides protection from environmental, chemical and physical damage (during transport). Packaging materials also offer a barrier to prevent the penetration of light, oxygen and moisture from deteriorating foodstuffs (Lambert et al., 2015). PHAs have been designated as packaging materials of preference owing to their unique features that gives them edge over conventional plastics. These features include good tensile strength, printability, flavour and odour barriers, resistance to grease and oil and high stability to temperatures (Tabone et al. 2010).
Agriculture is another highly potential market for applications of nanocomposite bioplastic from bacterial biomass containing PHAs. PHAs find application as suitable replacement of dark plastic mulch, which is currently used to suppress weeds, reduce water evaporation from soil, and warm soil for earlier planting (Ivanov et al., 2015). Stating the main benefit that can be derived from using film from nanocomposite bioplastic from bacterial biomass containing PHAs, Ivanov et al. (2015) stated that it can provide all the benefits of traditional plastic mulch except the natural ability to degrade. Another agricultural application of nanocomposite bioplastic from bacterial biomass with PHAs is manufacturing of slow-release fertilizers using bioplastic coating or embedding of fertilizers in bioplastic granules, bars, or films (Ivanov et al., 2015).

Biodegradability is defined as the capacity of a substance to be broken down, especially into innocuous products, by the action of microorganisms. Bacteria and fungi are the main participants in the process of biodegradation in the natural world. The breakdown of materials provides them with precursors for cell components and energy for energy-requiring processes. PHAs can degrade under both aerobic and anaerobic conditions. They can also be degraded by thermal means or enzymatic conditions (Yogesh et al., 2012). Biodegradability and biocompatibility are two key features of PHAs as these features are represent the main comparative advantages of PHAs over petrochemical plastics.
PHAs are stable in air, inert, resistant to moisture and are water insoluble. However, they can be fully degraded to water and carbon dioxide under aerobic conditions and to methane and carbon dioxide under anaerobic conditions by microorganisms in soil, sea, lake water and sewage (Anjum et al., 2016).
PHA degradation has been studied in various in vitro studies. PHAs degrade in a very slow process, which takes several months to reach a detectable mass loss (Nigmatullin et al., 2015). PHA is completely degraded by many species of soil bacteria, which use it as an energy source (Luzier, 1992). The polymer is first degraded by extracellular enzymes to monomeric and dimeric hydroxybutyrate, which are then taken up by the cells and metabolized. The rate of PHA degradation depends upon surface area, microbial activity, pH temperature, moisture and the presence of other nutrients (Luzier, 1992). PHA (e.g. P(3HB) and P(3HB-co-3HV)s) are degraded in both aerobic and anaerobic environments by the action of extracellular enzymes from microbial populations (Luzier, 1992).

The mechanism of PHA degradation involves the production of PHA depolymerases by the degrading microorganisms into water-soluble oligomers and monomers that can be recycled as carbon sources for the degrading microorganism (Reddy et al., 2003). Likewise, most PHA-producing bacteria can carry out intracellular degradation of PHAs. During intracellular degradation, the PHA depolymerase in the cell breaks down P(3HB) to yield 3-hydroxybutyric acid. A dehydrogenase subsequently acts on 3-hydrobutyric acid and oxidises it to acetylacetate and a ?-ketothiolase acts on acetylacetate to break it down to acetyl-CoA. The ?-ketothiolase enzyme plays an important role in both the biosynthetic and the biodegradation pathways (Anjum et al., 2016).
PHA degradation is influenced by the porosity and surface area of the materials, and molecular structure of monomer unit, which manifests in variations in crystallinity and hydrophobicity (Nigmatullin et al., 2015). In addition, Anjum et al. (2016) stated that additional factors that influence PHA degradation include microbial activity of the disposal environment, pH, temperature, moisture and presence of other nutrient materials.

Palm oil mill effluent is waste water generated during the processing of oil palm. The major wastes generated during palm oil production in a mill are generally classified as solid and liquid wastes. Solid wastes typically consist of palm kernel shells (PKS), mesocarp fruit fibres (MF) and empty fruit bunches (EFB). The process of extraction of palm oil requires a lot of water to steam sterilize the palm fruit bunches and clarify the extracted oil. the separated waste water sludge commonly referred to as POME is a brown slurry, which is composed of 4 – 5% solids, 0.5 – 1% residual oil and about 95% water (Nwoko and Ogunyemi, 2010). It is estimated that for each tonne of crude palm oil (CPO) that is produced, 5.0 tonnes-7.5 tonnes of water are required and more than 50% of this water ends up as waste water sludge commonly referred to as palm oil milling effluents POME (Okogbenin et al., 2014).
Due to the regular release of huge volumes of the POME generated, annual disposal remains an environmental challenge (Okogbenin et al., 2014). The common approach used in managing this waste is via treatment-oriented approach and almost all palm oil mills adopt the open ponding system to treat POME. According to Lee et al. (2015), this management system requires a large footprint due to extended retention time of 20–200 days. Also, the open emission of methane gas generated from the anaerobic pond contributes substantially to global warming. These drawbacks have prompted the development of better POME management system (Lee et al., 2015).
Sustainable exploitation of POME wastewater involves its bioconversion into organic compounds which could be used as substitute sources of renewable energy and/or valuable chemicals which will generate additional revenues for the industry (Mumtaz et al., 2010). Several studies have reported the use of POME in the production of biofuels (such as biohydrogen, bioethanol), organic acids (e.g citric acid) oil palm-based activated carbon and compost (after mixing with empty fruit bunch) (Mumtaz et al., 2010).
POME contains substantial quantities of valuable plant nutrient that vary according to the degree of treatment to which it is subjected. The potential use of recovery of water and organic matters from POME has been applied for various applications (Hassan et al., 2002). The ideal life cycle of eco-friendly exposure for PHA production from renewable resources like POME is a closed-loop process (as depicted in Figure 6).

Figure 6. Proposed cycle loop of regenerating waste from POME to biodegradable plastics (modified from Luzier, W. D. 1992).

The production of PHA will subsequently serve as the feed to a microbial fermentation process (at the end of cycle), to promote the environmental friendly effect. Ideally, this process occurs aerobically (in natural and tropical conditions), yielding water (H2O) and CO2 in the same proportions that were originally used in photosynthesis. The harmless production of end products could also be generated from microbial fermentation to produce biofuel energy.

POME is an acidic brownish colloidal suspension (Lee et al., 2013) containing large amounts of organic substances (Din et al., 2014). Din et al. (2014) described POME as having a chemical oxygen demand (COD) 50,000 mg L-1, oil and grease (4,000 mg L-1), COD (50,000 mg L-1), and BOD (25,000 mg L-1). Lee et al. (2013) studied the use of POME as substrate for bioplastic production. They showed that showed that fermented POME, which had high VFA content and high molar ratio of VFA-C:N:P, was a suitable feedstock for PHA production.
Owing to its high organic acids content, POME has been exploited for PHA production. However, POME is usually present in a complex form, which cannot be directly utilized by PHA-producing bacterial species. Therefore, anaerobic treatment has been proposed by many researchers for efficient hydrolysis and acidogenesis of wastes to short-chain volatile fatty acids (VFAs), which are fatty acids with a carbon chain of six carbons or fewer like acetic, butyric, and propionic acids and followed by PHA production (Mumtaz et al., 2010).
During the first stage of anaerobic digestion, acidogenic bacteria degrade the residual oil and lignocellulosic materials in POME into organic acids. Organic acids have been proven to be potential carbon sources for PHA production (Hassan et al., 2013). As proposed by Hassan et al. (2002), the combination of POME treatment and PHA production can provide a zero discharge system for palm oil mills. Several studies on the utilization of clarified organic acids for PHA production have indicated that polymers from POME-derived organic acids are comparable to commercially available organic acids (Hassan et al., 2013).

Figure 7. Bioconversion strategies for energy and polyhydroxyalkanoate production from oil palm biomass (Hassan et al., 2013).