Biotechnology-based microbial degradation of plastic additives.

Plastic additives are agents responsible to the flame resistance, durability, microbial resistance, and flexibility of plastic products. High demand for production and use of plastic additives is associated with environmental accumulation and various health hazards. One of the suitable methods of depleting plastic additive in the environment is bioremediation as it offers cost-efficiency, convenience, and sustainability. Microbial activity is one of the effective ways of detoxifying various compounds as microorganisms can adapt in an environment with high prevalence of pollutants. The present review discusses the use and abundance of these plastic additives, their health-related risks, the microorganisms capable of degrading them, the proposed mechanism of biodegradation, and current innovations capable of improving the efficiency of bioremediation.

Plastic additive biodegradation, Microbial remediation, Health risk of plastic additives

In modern-day society, the environment has been the “sink” of man-made wastes mostly comprised of plastics. The total virgin plastics that have been produced up to date is 8300 million metric tons for which only 9% have been recycled or reprocessed into secondary material, 12% have been incinerated and 79% have accumulated in landfills and the natural environment (Nelsen, et al. 2020; Geyer et al. 2017). Landfilling is the simplest, cheapest, and most cost-efficient way of waste disposal. In the U.S., 292.4 million metric tons of municipal solid waste (MSW) were generated in 2018, wherein 35.4% were used for recycling and energy recovery and 50% were left discarded for which 12.2% were plastics (EPA, 2021). Plastic refuse, in the long term, releases its decomposition by-products that infiltrates ground water resources. Groundwater near landfill sites exhibit higher concentration of organic pollutants, exposing those who depend on these water resources to biotoxicity (Hossain et al. 2014). In addition, insoluble pollutants accumulate in the soil, harming soil-dwelling organisms. Mismanaged plastic waste is also a global environmental problem. The estimated amount of mismanaged plastic wastes that entered the ocean is 4.8–12.7 million metric tons per year (Law et al., 2020; Jambeck et al. 2015). Aside from microplastics that could influence biogeochemical cycles (Guzzetti et al. 2018), plastic fragments impose a great danger to marine life (Koelmans et al. 2013). Intoxication of microplastics by aquatic microorganisms may impose a risk to higher organisms that tend to feed on them. Plastic fragments include plasticizers, monomers and oligomers that are gradually released to the environment (hereafter referred as ‘additives’). These additives displayed high toxicity even at nanomolar concentration (Koch and Calafat 2009; Oehlmann et al. 2009; Teuten et al. 2009). Several remediation technologies were offered to deplete plastic additive pollution in the environment. Physical remediation performs treatment by separating the pollutant from the contaminated zone and concentrating it in a collection reservoir (Fox 1996). Physical treatment exhibits high percentage collection of pollutants from the medium, however, additional treatment is required to completely dispose and destroy the collected contaminants. Chemical treatments, as another approach in remediation, comprise techniques that chemically transform organic pollutants into less harmful substances. Examples are reduction with ultraviolet or alkaline reagents, oxidation, and solidification/stabilization (Fox 1996). Chemical treatments impose great potential for remediation, however, several techniques exhibit limitations. Oxidation promotes generation of carbon dioxide, a ubiquitous air pollutant, and solidification/stabilization may display issues about its long-term effectiveness (Fox 1996). Overall, physical and chemical methods are both high-cost technologies, tedious and capable of altering the natural ecosystem (Ghoreishi et al. 2017). Among the remediation techniques, bioremediation, specifically microbial remediation, gained the upper hand in terms of sustainability, cost efficiency, and convenience. Microbial remediation offers easy applicability in situ (Das and Dash 2014). It explores the utilization of the ability of microorganisms to adapt to a polluted environment, where toxic concentration of an element or compound exists, and use it for their own advantage such as but not limited to cell growth and energy production (Boopathy 2000). Global use of physico-chemical remediation technologies dwarfs that of bioremediation, highlighting the underutilization of the sustainable bioremediation process. Several studies already raised concern on the leakage of additives from plastic products. It is known that additives that constitute a higher percentage by weight (w/w) in plastic products exhibit more toxicity even at low concentration than those at lower percentage. Despite the risk brought by plastic additive pollution in the environment, there is a lack of collective review regarding its bioremediation. This review aims to describe the potential of microbial bioremediation in degrading plastic additives. Furthermore, we anticipate future researches to focus on microorganisms in the remediation of plastic additives. We limit this review to plastic additives that are more than or equal to 1% w/w of the plastic product. In addition, one major rule of thumb for the migration of additives is that compounds with molecular weight of more than 600 g/mol will have a low tendency to migrate (Hansen et al. 2013). Thus, we cover only additives that migrate easily from the plastic product whose molecular weight do not exceed 600 g/mol. We provide here additional insights to complement a review paper about microbial remediation of plastic additives (Sanchez 2021). We also emphasize in this review the extent of use, environmental contamination, as well as the health impact of these plastic additives. We also evaluated the feasibility of several microorganisms in field applications. The structures of the representative plastic additives covered in this review is summarized in Fig. 1. Table 1 provides a summary of the microorganism identified in recent years that are capable of degrading plastic additives. Modification in bioremediation technique, efficiency, and notable capabilities of the organism and discoveries of the study are listed on the remarks section of Table 1. Table 1. | Pollutant | Species | Efficiency | Remarks | References | |---|---|---|---|---| | DEHP | Bacillus subtilis No. 66 | Degraded 80% of 5 mM DEHP Appropriate for soil reclamation | Hydrolysis of alkyl side chains of the PAEs forming phthalic acid Phthalic acid was then reduced Microorganisms that are capable to use the following enzymes have the potential to degrade PAEs into safer forms Hydrolase Carboxylic acid reductase Dehydrogenase Carboxylic acid reductase played a vital role in terms of degradation of PAEs as it reduces phthalic acid. The optimal pH for this enzyme is 7.5 (Finnigan et al. 2017) Degradation pathway of Gordonia alkanivorans YC-RL2 showed direct decarboxylation of phthalic acid. A possible decarboxylase enzyme was used by YC-RL2, converting phthalic acid to benzoic acid | Quan et al. (2005) | | Rhodococcus sp. strain WJ4 | Tolerated 200 mg/L of DEHP in liquid culture (96.4% degradation) Tolerated 1000 mg/L of DEHP in soil culture (55% degradation) Appropriate for harsh environmental sinks | Wang et al. (2015) | || | Agromyces sp. MT-O | Can tolerate DEHP at concentration range of 200 mg/L to 1000 mg/L Degraded 82.1% of 100 mg/kg of DEHP in non-sterilized soil for 12 days Degraded 82.1% of 100 mg/kg of DEHP in non-sterilized soil for 12 days Applicable in wide range of environmental condition | Zhao et al. (2016) | || | Gordonia alkanivorans YC-RL2 | Can tolerate DEHP at concentration range of 100 mg/L to 800 mg/L > 94% degradation efficiency Applicable in a wide range of environmental condition (temperature, pH, salinity) | Nahurira et al. (2017) | || | Burkholderia pyrrocinia B1213 | Degraded 98% of 500 mg/L DEHP Provided a clearer picture of degradation pathway of DEHP | Li et al. (2019) | || | Microbial consortium (CM9) Rhodococcus, Niabella, Sphingopyxis, Achromobacter, Tahibacter, and Xenophilus | Degraded 94.85% and 100.00% of DEHP (1000 mg/L) within 24 h and 72 h, respectively degradation through de-esterification and β-oxidation | Bai et al. (2020) | || | Achromobacter sp. RX | Used 99.3% of DEHP as sole carbon source after 96 h Masson pine seed powder (co-metabolic substrate) and Tween-80 (solubilizing agent) enhanced the biodegradation | Wang et al. (2021) | || | Ochrobactrum anthropi L1-W | Degraded 98.7% of DEHP (200 mg/L) in 72 h Also capable of degrading five other phthalates | Nshimiyimana et al. (2020) | || | Rhodococcus jostii PEVJ9 | Complete degradation of DEHP at minimal level of 1–100 μg/L in the presence of self-assembled monolayer-silver nanoparticles on bacterial cells | Annamalai et al. (2020) | || | Enterobacter spp. YC-IL1 | Degraded DEHP and hard to degrade butylbenzyl phthalate (BBP) and dicyclohexyl phthalate (DCHP) at 100%, 81.15%, and 50.69% efficiency, respectively, for 7 days DEHP degradation rate in artificially contaminated soil with 86% removed in 6 days | Lamraoui et al. (2020) | || | Cupriavidus oxalaticus E3 | Degraded 87.4%–94.4% of DBP and 82.5%–85.6% of DEHP at an initial amount of each phthalate of 200 mg/L after 60 h | Docking study showed that the conserved catalytic triplet structure (Ser140, His284, and Asp254) in the active sites are essential in catabolism of phthalate monoesters (PMEs) | Chen et al. (2021) | | | Paracoccus kondratievae | Simultaneous degradation of di-methyl phthalate (DMP), di-ethyl phthalate (DEP), di-butyl phthalate (DBP), di-isobutyl phthalate (DIBP) and DEHP, with DMP and DEP as the preferred substrates Slow DEHP degradation (200 mg/L) with t1/2 if 84.5 h | Two esterases belonging to families IV and VI were obtained | Xu et al. (2020) | | | Fusarium culmorum | NR | Esterase activity is induced in high concentration of DEHP (1500 mg/L) Identified five esterase isoforms (26.4, 31.7, 43, 73.6, and 125 kDa) | González-Márquez et al. (2020) | | | Burkholderia pyrrocinia B1213 | Degraded 98% of 500 mg/L DEHP Provided a clearer picture of degradation pathway of DEHP | Li et al. (2019) | || | DIDP, DINP and Other Phthalates | Ruegeria lacuscaerulensis, Thiomicrospira crunogena, Pseuodalteromonas tetraodonis, Alteromonas litorea, Alteromonas addita, Ectothiorhodospira haloalkaliphile, Desulfuromonas acetoxidans, Thalassobacillus devorans, Methylonatrum kenyense, and Salegentibacter salinarum | Most abundant species in a DINP and DIDP polluted sediments and potential phthalate degrading species Individual, synergistic, and competitive effects on degradation should be deeply observed | Chen et al. (2017) | | | Gordonia sp. | Degraded 80% of 500 mg/L phthalic esters for 30 h | Takao et al. (2002) | || | Consortium of saline soil bacterial Serratia sp., Methylobacillus sp., Achromobacter sp., Pseudomonas sp., Stenotrophomonas sp., Methyloversatilis sp., Delftia sp. and Brevundimonas sp. | Degraded 99% of DINP (500 mg/L) in 168 h, t1/2 of 12.76 h | de-esterification and β-oxidation pathway | Pereyra-Camacho et al. (2021) | | | Gordonia sp. | Degraded 91.25% of DOP with different initial concentrations (100–2000 mg/L) | Wang et al. (2020) | || | Rhodococcus ruber YC-YT1 | Entirely degraded 100 mg/L of phthalate mixture in 5 days Can degrade long-chained and branched phthalic esters Can be carried out in an electrolytic environment | Yanchun et al. (2017) | || | PCB | Aquamicrobium defluvii DSM11603 Y15403 | Depleted 100% of 0.5 mM 4-chlorobiphenyl after 4 days Depleted 53% of 0.5 mM 2,2′,4,4′,5,5′-hexachlorobiphenyl No depletion of 2,2′,3,4′,5,5′,6-heptachlorobiphenyl Degradation efficiency depends on the number of chlorine substituents | Growth is evident in single to six chlorine substituents The strain could also degrade other aromatic substrates, suggesting its capability is not only limited to dechlorination but also hydrolysis of the substrates | Chang et al. (2013) | | Rhodococcus erythropolis LG12 EU852376 | Can tolerate high salt concentration Degradation products are chlorine substituted, indicating its insufficiency for remediation | Metabolites should be nonchlorinated to conclude that the degradation is cost-effective | Chang et al. (2013) | | | Dehalococcoides mccartyi CG-1, CG-4 & CG-5 | Enriched by serial transfer using tetrachloroethene PCBs’ chlorine removal of 23.7, 27.7, and 83.6 µM, respectively | Tetrachloroethene is a good preliminary substrate before introducing the strain in the PCB mixture | Wang et al. (2014) | | | Paraburkholderia xenovarans LB400 | Capable of biotransforming 76% of the PCB mixture | Biodegradation was limited to lightly chlorinated PCBs | Bako et al. (2021) | | | Pleurotus pulmonarius LBM 105 | Degraded 65.50 ± 8.09% of PCBs in 21 days Toxicity was reduced by 46.47% | Proteomic analysis of the the Pleurotus pulmonarius LBM 105 revealed that oxidative metabolism was highly involved in the degradation of the PCBs | Chelaliche et al. (2021) | | | Pseudomonas extremaustralis ADA-5 | Strain was able to use 9.75% of available decachlorobiphenyl (DCB) as its sole carbon source Was able to bioaccumulate 19.98% of the DCB biomass | Bacterial membrane lipids were found to be altered in the presence of DCB | Lopez et al. (2021) | | | Pleurotus pulmonarius LBM 105 Trametes sanguinea LBM 023 | Biodegradation efficiency of 95.4% for single-culture experiments using Pleurotus pulmonarius LBM 105 Biodegradation efficiency of 55.4% for experiments using a co-culture of Pleurotus pulmonarius LBM 105 and Trametes sanguinea LBM023 | Comparative analysis of the degradation efficiency of single and co-culture experiments revealed that using a single culture of Pleurotus pulmonarius LBM 105 has has much a higher degradation efficiency | Benitez et al (2021) | | | Anabaena PD-1 | Aroclor 1254 biodegradation rate of 85% for 25 days Could degrade dioxin-like PCBs Further understanding of degradation mechanism is required | Highly efficient against meta-chlorine substituted biphenyls since the halide is an electron withdrawing group | Zhang et al. (2015) | | | TBBP A | Dehalococcoides mccartyi CBDB1 | Full debromination of TBBPA Generated BPA as metabolite Cell growth was inhibited due to TBBPA toxicity Additional remediation is required for the depletion of BPA | Yang et al. (2015) | | | Comamonas sp. JXS-2–02 | Newly isolated strain was able to degrade 86% of TBBPA after 10 days | Strain was the first to be isolated under anaerobic conditions | Peng et al. (2013) | | | Arthrobacter sp. YC-RL1 | Substrate analysis confirmed successful biotransformation of TBBPA concentrations ranging 0.2–300 mg/L | Strain was able to efficiently remediate other bisphenols such as BPA (0.2–600 mg/L) and BPF (0.2–600 mg/L) | Ren et al. (2016) | | | Bacillus brevis Bacillus pumilus | Biougmentation of an augmented sludge reactor and membrane bioreactor using both bacterial strains showed an 83% removal rate of TBBPA in wastewater | High removal efficiency can be traced to both biosorption and biodegradation activity | Islam et al. (2018) | | | Pycnoporus sanguineus | Growth was inhibited due to presence of Cr (VI) in the medium Higher concentration of TBBPA enhanced the removal of both TBBPA and Cr (VI) | Cr (VI) normally exists in TBBPA-polluted environment as both were typically used together in electronic devices | Feng et al. (2017) | | | PBDE | Pseudomonas spp. | Biostimulation was observed in depletion of less-brominated congeners when electron donors were added Could deplete BDE-209 for up to 11% in 90 days | Uses biostimulation to attract microorganism that has the potential to degrade PBDE | Qiu et al. (2012) | | Enterococcus casseliflavus | Degraded 10% to 50% BDE-209 (1 mg/L) after 7 days incubation | The high cell surface hydrophobicity of the bacterial cell wall helps in absorbing BDE-209 2D gel electrophoresis identified 50 differentially expressed proteins in treated vs untreated samples | Tang et al. (2016) | | | Phanerochaete chrysosporium | Degraded 55% to 66% BDE-47 (0.5 mg/L) with Cd2+ concentration of 0 to 10 mg/L | Extracellular enzyme may be responsible for the process BDE-47 is degraded by hydroxylation | Cao et al. (2017) | | | Phlebia lindtneri JN45 | Degraded 77.3% BDE-209 (20 mg/L) in 30 days Adding glucose showed marked increase in degradation | Proposed mechanism involves debromination, hydroxylation, and ring opening reactions | Xu and Wang (2014) | | | Pseudomonas putida | Degradation rate constants were 2–3 times larger when glucose was added to the PBDE system Glucose does not alter the degradation pathway of PBDE Degradation rates were 1–2 folds higher when biphenyl was added to the PBDE system Lateral deoxygenation took place when biphenyl was added to the PBDE system, resulting to the formation of toxic para-OH PBDE | The addition of glucose/biphenyl provides additional carbon source for bacteria, helping them not to fully rely on BDE as it may be difficult for them to survive in such environment if deficient in nutrient | Lv et al. (2016) | | | BPA | Microbial consortia | Degradation half-life is 13.1, 10.8, and 10.2 h for concentrations of 10, 20 and 50 mg/L (Dark) Metabolites were not hazardous | Eio et al. (2014) | | | Pseudomonas aeruginosa PAb1 | Promote cell growth rate of 0.841 h-1, Metabolites were phenol, acetophenone, and hydroquinone and p-hydroxybenzoic acid | Vijayalakshmi et al. (2018) | || | Pseudomonas putida YC-AE1 | Resistant to BPA for up to 1000 mg/L Completely degrade 500 mg/L BPA within 72 h | Eltoukhy et al. (2020) | || | Microbial community Sphingonomas spp. Sph-1 Sphingonomas spp. Sph-2, | Sphingonomas spp. was able to degrade 50 mg/L BPA Pseudomonas sp. and Pusillimonas sp. residing on the similar community were observed as BPA non-degraders, however, degraded the by-products of BPA by Sphingonomas spp. | The BPA-degradation pathway involved the cross-feeding of BPA by the Sphingonomas spp with BPA intermediates by Pseudomonas sp. and Pusillimonas sp. | Yu et al. (2019) | | | Shewanella haliotis MH137742 | Degraded 75 mg/L of BPA | de Santana et al. (2019) | || | Triclosan | Alpha-proteobacteria | Bacterium was known to exist predominantly in RL TCS was observed to deplete from 30 µg/g to 1.84 µg/g after 56 days using di-RL homologues | Guo et al. (2016) | | | Sphingobium | Glutathione transferase-dependent biodegradation Sphingobium was detected in the wastewater sample as a major TCS-degrading bacterium by using DNA-SIP (stable isotope probing) The consortia shows that it contributes to the 18.54% metabolism of TCS; the rest is accounted to the nitrification process of the wastewater treatment system | High accuracy and convenience are observed in DNA-SIP as thert do not require redundant purifications. Potential degrading-bacteria could be easily detected by this technology | Dai et al. (2021) | | | Amaricoccus as the most abundant micorbe | With the aid of the following surfactants: rhamnolipid (RL), sophorolipid (SL) and sodium dodecyl benzene sulfonate (SDBS), the biodegradation of TCS, with the influence of nitrification, was enhanced by 1.25, 1.23, and 1.14 times | Jia et al. (2020) | || | Georgenia, Soehngenia, Comamonas, Pseudomonas, Desulfovibrio, and Sulfurospirillum | Biodegradation accounts for 49–68% of TCS removal. Synergistic with sorption which contributed 32–51% of TCS removal | Novel technology showing sulfur-driven TCS removal | Zhang et al. (2021) | | | Dehalococcoides mccartyi CG1 | Slow dichlorination of TCS of 0.062 µM/day Addition of electron acceptors increased the rate to 0.245 µM/day | Zhao et al. (2020) | || | Citrobacter freundii KS2003 | Able to degrade 99.57 ± 0.6% of 250 mg/L of TCS | After o-cleavage, subsequent biodegradation by KS2003 wherein catechol 1,2-dioxygenase acted as an important enzyme | Kumari et al. (2021) | NR Not reported Plasticizers Plasticizers are chemical compounds added to plastics to improve their flexibility which typically comprise 10% to 70% w/w of the plastic product (Hansen et al. 2013). In 2005, phthalates accounted for 88% of global consumption of plasticizers; however, due to its health-related risks, it was reduced to 65% in 2017 (IHS Markit 2018). The most common phthalate plasticizer is di(2-ethylhexyl) phthalate (DEHP) which contributes to 51% of the global consumption of phthalates, followed by diisodecyl phthalate (DIDP) and diisononyl phthalate (DINP) (Fig. 1) that constitutes 21% and 11% of phthalate demand, respectively (Peijnenburg 2008). a. Di(2-ethylhexyl) phthalate (DEHP) The annual production of DEHP globally is around 45–113 thousand tons and usually used as a plasticizer for polyvinyl chloride (PVC) plastics (EPA, 2020). DEHP and its phthalic acid esters intermediates are ubiquitous in the environment and is mostly found in wastewater (Ren et al. 2017). DEHP exposure causes pregnancy disorders and birth defects. Findings suggest that prenatal exposure to DEHP could lead to shorter pregnancy duration and might cause low birth weight on infants (Latini et al. 2003). High plasma concentration of DEHP was observed to be correlated with endometriosis, a condition in which the lining of the womb (endometrium) grows outside the uterus (Cobellis et al. 2003). DEHP disturbs estradiol production and suppresses ovulation due to the lack of stimulation of luteinizing hormone (LH) (Davis et al. 1994). Phthalates exposure is also associated with autism spectrum disorder (ASD) (Testa et al. 2012). Several bacterial strains isolated mainly from soil and sediment samples have reported to degrade DEHP up to a concentration as high as 1000 mg/L and at varying rates ranging from a few hours to days (Table 1). The complete degradation of DEHP proceeds via a series of esterase, β-oxidation, decarboxylation and dioxygenase action (Fig. 2). Survival of a bacteria in tough environments is strongly accounted to endospore formation, a capability exhibited by most of those in genus Bacillus (Sonenshein et al. 2002). Biodegradation of DEHP together with other alkyl phthalates was observed using Bacillus subtilis No. 66 (Quan et al. 2005). The addition of 8% of culture medium of strain No. 66 in the soil successfully degraded 5 mM DEHP for up to 80% even in the presence of other organisms. B. subtilis are well-known promoters of biofilm formation (Bais et al. 2004). Adsorption of phthalates on biofilm biomass is found to be the main driving force for the removal of phthalates in the aqueous phase which further resulted in increased biodegradation efficiency (Wen et al. 2016). B. subtilis also promotes plant growth in polluted soil by acting as biocontrol agent and stimulating nutrient availability in soil (Nagórska et al. 2007). Rhodococcus is a diverse genus of bacteria and at par with Pseudomonas when it comes to degradation of a wide variety of organic compounds (Larkin et al. 2005). Rhodococcus sp. strain WJ4 can withstand and degrade 200 mg/L of DEHP in liquid culture for up to 96.4% and 1000 mg/L of DEHP in soil culture for up to 55% (Wang et al. 2015). Having a low lysis efficiency, Rhodococcus is a perfect candidate for degradation of DEHP residing in harsh environmental sinks such as industrial effluents and landfills (Kauffmann et al. 2004). Agromyces sp. MT-O strain is also a promising agent for complete degradation of DEHP, producing mono(2-ethylhexyl) phthalate (MEHP) and phthalic acid (PA) which can subsequently be mineralized (Fig. 2) (Zhao et al. 2016). In a microbially induced mineralization, organic pollutants are transformed to carbonate precipitate which is less toxic and a more stable compound. The DEHP concentration in the bioaugmentation set-up in this study was limited to 100 mg DEHP per kg of soil to avoid the inhibition of microbial biomass, basal respiration, and catalase activity (Zhao et al. 2016). In 12 days, 82.1% and 73.9% of DEHP was removed in non-sterilized soil and sterilized soil, respectively. This implies that MT-O strain may interact with other microorganisms that helps enhance DEHP degradation. When inoculated in a mineral salt media with DEHP as sole carbon source, the strain almost completely degraded 200 mg/L of DEHP and can withstand up to 1000 mg/L of DEHP although the degradation rate was lower at high DEHP concentration. This means that bioremediation would be less efficient in extremely DEHP-polluted sites. Gordonia alkanivorans YC-RL2, a bacterial strain isolated from petroleum-contaminated soil and can tolerate high concentrations of DEHP ranging from 100 mg/L to 800 mg/L (Nahurira et al. 2017). Degradation rate of this strain exceeds 94% and metabolites can be further utilized for subsequent cell growth (Fig. 2). Factors affecting bioremediation are temperature, electron acceptors/donors, nutrients, pH, metabolites, energy sources, and temperature (Boopathy 2000). Interestingly, YC-RL2 displays versatility in carrying out the degradation of DEHP in soil, having a temperature range of 10 °C–50 °C, pH condition of 6.0–11.0, and salinity level of 0–12% w/v of NaCl (Nahurira et al. 2017). Burkholderia pyrrocinia B1213, isolated from fertile soil near the vegetable oil factory in China (Li et al. 2018), can degrade 98% of 500 mg/L of DEHP to produce MEHP, mono-butyl phthalate (MBP), PA and 4-oxo-hexanoic acid as metabolites (Li et al. 2019). Detection of these metabolites provides a clearer picture of the biodegradation pathway. β-oxidation of MEHP was responsible for the formation of MBP where the alkyl chain length was reduced from 8 to 4. PCA was not detected as a metabolite, suggesting that PCA was rapidly converted to 3-oxohexanedoic acid to 4-oxo-hexanoic acid before the complete mineralization in the cell (Fig. 2). b. Diisodecyl phthalate (DIDP), Diisononyl phthalate (DINP) and other phthalates DIDP is a mixture of isomers of 10-carbon branched dialkyl chain phthalates (CAS registry numbers 68515–49-1 (C-10 rich) and 26,761–40-0) and one of the leading phthalate plasticizers which is commonly used as an additive in PVC. Several studies observed that DIDP are less toxic than the other phthalate counterparts. Evaluation of DIDP showed that it does not exhibit mutagenicity (Mckee et al. 2000), cytotoxicity (Ghisari and Bonefeld-Jorgensen 2009), and teratogenicity (Waterman et al. 1999). However, significant health threats were observed when DIDP is mixed with other phthalate compounds. Phthalate mixtures, despite the weak individual phthalate compound toxicity is, therefore, a serious environmental problem. Phthalate mixture containing DIDP interferes with the thyroid hormone system (Ghisari and Bonefeld-Jorgensen 2009). In vivo evaluation of DIDP in rats showed that it enhanced the activity of sodium/iodide symporter–a protein responsible for the active transport if iodide in the thyroid (Breous et al. 2005). Mixtures of phthalates that include DIDP also exhibited estrogenic endocrine disrupting activities (Chen et al. 2014; Ghisari and Bonefeld-Jorgensen 2009). Similar to DIDP, DINP is a complex mixture of DINP isomers (CAS registry number of 68,515–48-0 and 28,553–12-0) commonly used as plasticizer in PVC (EPA 2019). Compared to DIDP, several literature studies have reported its adverse health effect as an individual phthalate. Although it does not exhibit mutagenicity (Mckee et al. 2000) and developmental toxicity (Waterman et al. 1999), DINP can cause rodent liver tumor through peroxisome proliferation (Kaufmann et al. 2002; Reisenbichler and Eckl 1993). DINP could aggravate atopic dermatitis in vivo (Koike et al. 2010) and its phthalate mixture exhibit enhanced estrogenic activity (Chen et al. 2014). Furthermore, phthalate mixture containing DINP interferes with the thyroid hormone system (Ghisari and Bonefeld-Jorgensen 2009). DIDP and DINP are less prevalent than DEHP; thus, only a few studies have explored the remediation of these phthalates. Most studies have focused on the remediation of phthalate mixture rather than individual phthalate substance. Bioremediation in a study using indigenous soil microorganisms done in a slurry phase reactor exhibited more than 70% efficiency of phthalate removal (Ferreira and Morita 2012). The condition was monitored at pH and temperature range of 7.8 to 8.4 and 17 °C to 25 °C, respectively. This was also applied ex situ using soil contaminated with wastes from the plasticizer industry. The biodegradation followed first-order kinetics with efficiency of 61% pollutant removal wherein DIDP was also reduced from 1800 (± 193) mg/L to 141 (± 3) mg/L after 120 days. Bacteria strains from an ocean dredge material disposal site were found to potentially degrade phthalates (Chen et al. 2017). The middle part of the site was polluted with 2849 (± 2588) ng/g of total phthalic esters, wherein the concentration of DINP is 381 (± 293) ng/g and for DIDP is 91.2 (± 73.6) ng/g. The most abundant bacterial species in this site are Ruegeria lacuscaerulensis (7.87%), Thiomicrospira crunogena (5.45%), Pseudoalteromonas tetraodonis (3.07%), Alteromonas litorea (2.69%), Alteromonas addita (2.31%), Ectothiorhodospira haloalkaliphile (1.62%), Desulfuromonas acetoxidans (1.62%), Thalassobacillus devorans (1.56%), Methylonatrum kenyense (1.15%), and Salegentibacter salinarum (1.09%). This study provided a good preliminary screening of the potential species that are capable of efficiently degrading phthalate esters, DIDP and DINP. However, their individual microbial activity, the synergy within the consortia, and the competitive effects in a biofilm must be thoroughly observed and analyzed. A species belonging to the genus Gordonia was capable of scavenging phthalates at more than 80% of 500 mg/L phthalic ester mixture for 30 h (Takao et al. 2002). The strain Gordonia sp. Lff was able to degrade more than 91.25% of di-n-octyl phthalate (DOP) with different initial concentrations (100–2000 mg/L) under optimal condition (Wang et al. 2020). This Lff strain showed little influence on the soil bacterial community making it a strong candidate for DOP remediation in various environments. Rhodococcus ruber YC-YT1 (Yanchun et al. 2017) was observed to tolerate and degrade 85% of 100 mg/L DEHP in an inorganic salt medium. In addition, it was able to entirely sequester 100 mg/L DEHP, dipropyl phthalate and butyl benzyl phthalate in an organic salt for 5 days. Hence, the capability of strain YC-YT1 to strongly degrade long-chained and branched phthalic esters suggests its potential to degrade various phthalates such as DIDP and DINP. Moreover, the degradation was carried out in an electrolytic environment implying its applicability to remediate phthalate-polluted wastewater and saltwater resources. Similar to DEHP, DIDP and DINP undergo degradation thru the de-esterification and β-oxidation pathway (Pereyra-Camacho et al. 2021). Flame retardants Flame retardants are common additives constituting 12%–18% w/w of the plastic product (Hansen et al. 2013). Flame retardants extinguish the flame propagation of the plastic product and are categorized into four groups: halogenated flame retardants, phosphorus-based flame retardants, melamine flame retardants, and inorganic hydroxide flame retardants (Levchik 2007). Halogenated flame retardants dominate the global consumption of flame retardants (Murphy 2001) and can be subcategorized either as chlorine-containing and bromine-containing flame retardants. The main commercial brominated fire retardants are tetrabromobisphenol A (TBBPA) and polybrominated diphenyl ethers (PBDE) (Segev et al. 2009). Polychlorinated biphenyls (PCB) have structural and behavioral similarity with PBDEs (Birnbaum and Bergman 2010). a. Polychlorinated biphenyls (PCB) PCBs were used as flame retardants in plastic products, in power supplies such as capacitors and transformers, in pesticides and in adhesives. Known to be a persistent organic pollutant, it is no longer produced and used in most countries. However, pollution occurs due to the redistribution of PCB that is already present in the soil and water (EPA 2000). Aside from being a persistent substance in the environment, it also confers various health risks. Dopamine levels tend to decrease after exposure to PCB (Seegal et al. 1986). Prenatal exposure could also delay embryonic development (Lindenau et al. 1994). Furthermore, PCBs are genotoxic and have a potential to promote hepatocellular carcinoma (Ludewig and Robertson 2013). Aquamicrobium defluvii DSM11603 Y15403 isolated from a sewage treatment plant was able to degrade PCB having one to six chlorine substituents (Chang et al. 2013). The culture grew for about 30–70 µg/ml and depleted 100% of 0.5 mM 4-chlorobiphenyl after 4 days. However, the increase in the number of chlorine substituents in PCB like in 0.5 mM 2,2′,4,4′,5,5′-hexachlorobiphenyl resulted in a decrease to 53% depletion efficiency. In addition, no depletion of 2,2′,3,4′,5,5′,6-heptachlorobiphenyl was observed in this isolate. In the same study, Rhodococcus erythropolis LG12 EU852376 was isolated in the same sewage treatment plant and displayed biphenyl degradation and salt tolerance. Hence, the isolate could work even in soils with high NaCl concentration and nitrate concentration. However, analysis of the metabolites from the degradation of 4-chlorbipenyl showed the presence of chlorinated compounds (Fig. 3). Since chlorinated organic compounds are still harmful in the environment (Henschler 1994), further dechlorination is needed. To produce environmentally safe by-products, introduction of a dehalogenating or genetically engineered bacteria that have the enzyme capable of halogen cleavage from the dead-end metabolites and by-product is thereby warranted. Dehalococcoides is a promising genus of bacteria towards dechlorination of PCBs; however, the challenge in PCB remediation is the acquisition of enriched dechlorinators strains. Dehalococcoides sp. CBDB-1 is one of the few bacterial strains that shows dechlorination activity in PCB; however, the enrichment of the dechlorinating bacteria in commercial PCBs (such as Aroclor 1260) was unsuccessful (Adrian et al. 2009). To resolve this, the less stable tetrachloroethene was used as an alternative electron acceptor instead of PCB in order to enrich the Dehalococcoides in three isolated PCB-degrading microbial communities, CG-1, CG-4, and CG-5. The Dehalococcoides in CG-1, CG-4, and CG-5 was further enriched to 85.2%, 44.2%, and 86.3%, respectively (Wang et al. 2014). Tetrachloroethene was degraded and chlorine substituents were progressively cleaved to trichloroethene then to dichloroethane. 10 serial transfers of the cultures acquired > 90% of the Dehalococcoides. This was confirmed after transcriptomic analyses and quantitative polymerase chain reaction which showed that the gene involved in tetrachloroethene and PCB dechlorination was similar. PCB dechlorination capability was retained after confirming that the three cultures were able to support growth in Arochlor 1260 containing media. After 150 days, growth of CG-1, CG-4, and CG-5 cultures were promoted by 8.2-, 12.0-, and 16.2-fold, respectively, with chlorine removal of 23.7, 27.7, and 83.6 µM. Anabaena PD-1 isolated from PCB-contaminated paddy soils in Taizhou, Zhejiang, China was capable of degrading Aroclor 1254 (mixture of PCB having 54% by weight of chlorine) with biodegradation rate of 85% for 25 days (Zhang et al. 2015). In addition, it also showed a high degradation percentage of dioxin-like PCBs. However, analysis of the degradation products was not done in this study. Moreover, no gene and enzyme identification were done to fully understand the mechanism of remediation. Acidovorax sp. KKS102 can degrade PCB; however, the by-products are still toxic since they are still chlorine-containing. This problem has been resolved using a recombinant glutathione S-transferase (GST) from the homolog of the gene BphK (biphenyl upper pathway K) in Acidovorax sp. KKS102 (Shehu and Alias 2019). The C10F and A180P mutated recombinant proteins showed an increase in GST dichlorination activity relative to the wild-type protein. b. Tetrabromobisphenol A (TBBPA) TBBPA is a widely used flame retardant mainly in electronic devices and electronic circuit boards. In addition to the effectivity of brominated compounds in terms of fire extinguishing, TBBPA is not subjected to any regulatory restriction and accounts for the largest production of flame retardants, amounting to around 170,000 tons of the global demand (BSEF 2012; Makinen et al. 2009). Accumulation of TBBPA in the environment displays a great threat to a wide variety of organisms. Accumulation of TBBPA in plants induces oxidative stress, in which the toxic dose level depends on the antioxidative capacity of the plant (Sun et al. 2008). It can suppress thyroid hormone triiodothyronine (T3) which was observed in Rana rugosa tadpole’s tail shortening (Kitamura et al. 2005; Sun et al. 2009). Thus, TBBPA could affect the metamorphosis of amphibians. Moreover, TBBPA in Wistar rats increased T3 levels and decreased the T4 levels, suggesting its adverse effect on reproduction and development (Van der Ven et al. 2008). Egg production of zebrafish (Danio rerio) tends to decrease when exposed to at least 0.047 µM of TBBPA (Kuiper et al. 2007). Cytotoxic evaluation of TBBPA shows its capability to induce cell death in TM4 Sertoli cells which is important for sperm development (Ogunbayo et al. 2008). TBBPA is an immunotoxin and significantly inhibits the expression of interleukin-2 receptor a chain (CD25), a protein essential for proliferation of activated T cells (Pullen et al. 2003). TBBPA exhibits nephrotoxicity at high doses in newly born rats (Fukuda et al. 2004). In vivo study of TBBPA in rats show significant changes of glutathione, malondialdehyde, and 5-aminolevulinate dehydratase, suggesting that TBBPA is hepatotoxic (Szymańska et al. 1999). As mentioned in PCB degradation, Dehalococcoides mccartyi bacteria strains are known for its capability to degrade halogenated compounds. Strain CBDB1 was able to fully debrominate TBBPA to form BPA via reduction/dehalogenation (Fig. 4); however, no cell growth in the spent media was observed during the transformation (Yang et al. 2015). This growth inhibition may be due to the high lipophilicity of TBBPA that penetrates the cell membrane of the bacteria (Kefeni et al. 2011). Furthermore, the by-product BPA is an endocrine disrupting chemical; thus, it is necessary that this metabolite be further degraded. TBBPA and Cr (VI) coexist with each other in electronic wastes. Chromium (VI) is a toxic metallic substance added to electronic devices to prevent rust (Salnikow and Zhitkovich 2008). Their coexistence is an important factor to the bioremediation of TBBPA. Cr (VI) affects the TBBPA degradation capability of Pycnoporus sanguineus as Cr (VI) inhibits the fungal growth, intracellular proteins synthesis, cell viability, and ATP enzyme activity (Feng et al. 2017). However, when TBBPA exists at higher concentration, it enhances the removal of both TBBPA and Cr (VI) due to improved intracellular proteins synthesis and ATP enzyme activity. c. Polybrominated diphenyl ethers (PBDEs) PBDEs are structurally similar to PCB and a common additive used to retard fire in a wide variety of consumer and electronic products (Siddiqi and Clinic 2003). Landfill, being the most significant reservoir of electronic equipment waste, collects 1747 million metric tons of PBDE, whereas the incineration releases 44.9 kg/year of decaBDE (U.S. Environmental Protection Agency, 2010). The Stockholm convention bans the use of tetraBDE, heptaBDE, and decaBDE, leading several companies to resort to the use of alternative flame retardant such as TBBPA (Fromme et al. 2016). Despite the restriction, PBDEs are still prevalent in the environment and is one of the common organic pollutants present in soil. PBDEs can disrupt the thyroid hormone. Exposure to BDE-28, -47, -99, -100, and -153 during prenatal stage may cause a 27.6% decrease of thyroid-stimulating hormones to children under 3 years old (Vuong et al. 2018). Prenatal exposure to BDE-99, -47, -100, and -153 was also associated with an increase of total thyroxine level in cord blood (Ding et al. 2017). Electronic waste recycling workers who are directly exposed to PBDE bioaccumulation exhibit altered thyroid hormone-regulated gene expression due to the interference of PBDE thyroid hormone signaling effects (Zheng et al. 2017). BDE -17, -100, -47, and -49 was also correlated with domestic feline hyperthyroidism (Walter et al. 2017). Biostimulation is a conventional technique in bioremediation wherein an electron acceptor is added to stimulate microbial activity. Enrichment of microbial consortia by five different electron donors (methanol, ethanol, pyruvate, lactate, and acetate) showed no significant increase to dehalogenation of deca-BDE (Qiu et al. 2012). This indicates that generation of bromide ion was not obviously affected by the addition of exogenous electron donors. Pseudomonas spp. identified as the predominant species in this consortium was able to reduce BDE-209 by 11% after 90-day incubation (Qiu et al. 2012). However, degradation of less brominated congeners was stimulated when electron acceptors were added although lactate and pyruvate inhibited the degradation of BDE-209. Thus, the microbial consortium is recommended to the remediation of PBDE containing nine bromine substituent or less, with an addition of electron acceptors for better results. Pseudomonas putida was employed for the bioremediation of PBDE and the effect of glucose and biphenyl as an additional carbon source on the efficiency of aerobic cometabolism was investigated (Lv et al. 2016). Degradation rate constants of four PBDE congeners (BDE47, BDE15, BDE3, and DE) were improved by 2–3 times when glucose was added. On the other hand, cometabolism of the four PBDE congeners with biphenyl improved its degradation rate by 1–2 folds. The cometabolism with glucose showed that PBDE degradation will only start when a huge amount of glucose is consumed. This implies that the P. putida was enriched, and the increase of the cell population results in the faster degradation of PBDE. Moreover, the glucose-PBDE system and single system share a similar pathway. On the other hand, the addition of biphenyl provided a larger amount of carbon source, this helped to improve cell population. It is important to identify the new metabolic pathway when cometabolism is performed to recognize potentially toxic metabolites that are being generated in the process. As illustrated on Fig. 5, PDBE cometabolism with biphenyl undergoes dioxygenation processes, resulting in two dihydroxylated by-products. The para-OH PBDE by-product is one of the most toxic metabolites as it strongly disrupts nuclear hormone receptor activity (Kojima et al. 2009; Qiu et al. 2009). Bisphenol-A as stabilizers Stabilizers are additives that are added to the plastic products to hinder their degradation. There are two types of stabilizers: UV and heat stabilizers. UV stabilizers inhibit oxidative degradation of plastic products upon exposure to ultraviolet light while heat stabilizers prevent oxidation of plastic when exposed to heat, mechanical stress and loading (Hahladakis et al. 2018; Murphy 2001). Bisphenol-A (BPA) is an organic pollutant that is commonly used as stabilizer which comprises 0.05%–3% w/w of the plastic product (Brevik and Burgess 2012; Hansen et al. 2013). Aside from being a stabilizer, BPA is a highly produced chemical commonly used as a monomer for polycarbonate plastic and also used for epoxy resins (American Chemistry Council 2012; Møller et al. 2012). As a result, BPA has become ubiquitous in nature which mostly pollutes water resources (Michałowicz 2014; Staples et al. 1998). BPA, due to structural similarity with estrogen, imposes the risk for endocrine disruption. Several studies showcase the capability of BPA to contribute to some endocrine-related diseases. BPA is associated with diabetes mellitus (Shankar and Teppala 2011), polycystic ovarian syndrome (Takeuchi et al. 2004), and obesity (Carwile and Michels 2011; Takeuchi et al. 2004). BPA may also encourage developmental reprogramming which may increase cancer susceptibility (Keri et al. 2007; Walker and Ho 2012). In utero exposure to BPA increases the enhancer of Zeste Homolog 2, a protein known as a promoter of viral carcinogenesis in the breast (Doherty et al. 2010; Sanna et al. 2018). Low-dose exposure to BPA during prenatal stage can alter gene expression, specifically the sex-specific effects of hypothalamic estrogen receptors α and β expression and hippocampal and hypothalamic oxytocin expression (Arambula et al. 2016). BPA can also alter steroidogenic enzymes, disturbing ovarian steroidogenesis resulting in complications during pregnancy (Zhou et al. 2008). BPA conversion into bacterial biomass and metabolites diplaying low to no estrogenic activity can be accomplished by a bacterial consortia through catabolic pathways (Eio et al. 2014). The consortia which were isolated from an activated sludge in a wastewater treatment plant were able to deplete BPA for up to 50 mg/L. Light is also observed as a major factor for the degradation as it inhibits the process. In the absence of light, the degradation half-life is 13.1, 10.8, and 10.2 h for concentrations of 10, 20, and 50 mg/L, respectively. Degradation is not a single step process; thus, using a microbial consortium as a bioremediation agent offers advantage than a single microbe. BPA depleted by one species offers no guarantee that the products will be non-hazardous. The remaining species of the consortia will possibly work on degrading the hazardous by-products into less harmful compounds. Degradation of BPA using Pseudomonas aeruginosa PAb1 promoted cell growth rate and released the metabolites phenol, acetophenone, hydroquinone and p-hydroxybenzoic acid (Vijayalakshmi et al. 2018). The bacterium was isolated from an effluent of a thermal paper industry which widely uses BPA as antioxidant of the product. Moreover, the bacterium was identified to exist in a condition where pollution parameters were significantly higher than the typical limit. Hence, PAb1 might exhibit durability and survive when applied to other environmental sinks that possess worse physicochemical conditions. Triclosan (TCS) as biocide Biocides were used to preserve materials and prevent fungal, microbial, and insect activity from damaging the material (Horn et al. 2003). In 2007, global demand for biocidal products reached $6 billion, with 5–6% projected growth per year (Karsa 2007). Biocides were typically added (0.001–1% w/w) in soft PVC and foamed polyurethane (Hansen et al. 2013). An antimicrobial agent, TCS was used for formulation of hygienic products, pesticide, fungistat, and as material preservative for adhesives, fabrics, vinyl, plastic products, textiles, and rubber (EPA 2008; FDA 2017). TCS were also used as a preservative for one-time use medical devices such as surgical scrubs, surgical sutures, catheters, and ureteral stents (Yueh and Tukey 2016). TCS was used for over 50 years; however, it is now limited to medical care settings and recently banned for the formulation of antibacterial soaps due to its insignificant effect for sanitation (FDA 2016; Schweizer 2001). TCS kills the bacteria by inhibiting enoyl-(acyl carrier protein) reductase, an enzyme essential for the lipid synthesis of the cell (Macri 2017). The most common environmental sink of TCS are surface water bodies such as wastewater effluent, river, and lake (Singer et al. 2002). TCS is an endocrine disruptor that commonly affects the thyroid hormone (Crofton et al. 2007; Veldhoen et al. 2006) and is strongly associated with decrease immunity function (Rees Clayton et al. 2011). TCS could also disrupt LH and follicle-stimulating hormone (FSH) synthesis which will also affect the production of androgen resulting in a lower sperm production (Kumar et al. 2009). Carcinogenicity potential of TCS was observed in vivo as findings suggest that TCS promotes liver fibrogenesis and tumorigenesis (Yueha et al. 2014). Aside from its persistence in the environment and low bioavailability, TCS are strongly uncooperative to microbial bioremediation. Hence, more innovative bioremediation techniques are necessary. With the use of biosurfactant (e.g. rhamnolipid) in bioremediation, interaction between TCS and microbial cells could be strongly improved. Alpha-proteobacteria (e.g., Sphingomonadaceae and Caulobacteraceae) were observed to be dominant in resisting TCS in a rhamnolipid-enhanced aerobic biodegradation system (Guo et al. 2016). Hence, it is a promising candidate for the degradation of TCS. It was observed that the concentrations of rhamnolipid (RL) is directly related to the aerobic degradation. TCS was observed to be depleted from 30 µg/g to 1.84 µg/g after 56 days using di-RL homologues (Guo et al. 2016). However, efficiency of biosurfactant-enhanced bioremediation is dependent on the environmental condition. Application of the RL is selective on a specific environment. It was observed that the ideal conditions were pH 8–9, 20–35 °C, 0.001–0.1 mol/L NaCl, and high concentration of dissolved oxygen. Moreover, surfactants such as RL, sophorolipids, and sodium dodecyl benzene sulfonate which do not affect nitrification were found to enhance the biodegradation of TCS in nitrification systems where Amaricoccus was identified as the major TCS degrader (Jia et al. 2020).

and future direction The continuous increase in demand for plastics results in the accumulation of plastic waste in the environment. These wastes gradually release the additives that were once used to produce and enhance the quality of these plastic products. Even though plastics contain little to trace amounts of additives, the enormous volume of plastic waste makes plastic additives a serious environmental hazard. The ubiquitous nature of plastic additives in the environment prompted researches that focuses on its bioremediation. Microorganisms are well known for its capability to evolve and survive in harsh environments. Thus, there is a huge potential for microorganisms to endure a plastic additive-polluted environment. Biodegradation of the plastic additives by microorganism was made complicated as these pollutants tend to appear stable and/or highly toxic. Since bioremediation is not a fully utilized technology, several limitations can be observed in some pollutants, including, but not limited to the following: (1) difficulty in enrichment of degrading bacteria in microbial community as observed in PCB remediation by Dehalococcoides sp. (Adrian et al. 2009); (2) cell growth inhibition due to high toxicity of pollutant as in the case of TBBPA degradation in Dehalococcoides mccartyi CBDB-1 (Yang et al. 2015); (3) dead-end by-products that exhibit toxicity; and (4) low degradation efficiency. To enhance the degradation of organic pollutants, several techniques are offered: (1) use of biosurfactants to increase the rate of solubilization; (2) enrichment of strains using a less toxic carbon source as discussed in acquisition of PCB-dechlorinating strains enriched by tetrachloroethene; (3) biostimulation (Qiu et al. 2012); (4) cometabolism (Lv et al. 2016); and (5) gene modification in order to produce a degrading enzyme essential to the decomposition of organic pollutant (Ohtsubo et al. 2004). The emergence of systems biology during the past decades potentially resolved several complications encountered in past trials of bioremediation. Such complications include, but are not limited to the following: (1) inefficient natural biodegradation due to slow evolution of some microbes to develop catabolic traits; (2) non-feasibility of catabolic networks and enzymatic reaction kinetically and thermodynamically; and (3) stress response of genetically modified cell to the changes in metabolic module and exposure to pollutant (Dvořák et al. 2017). Combination of the approaches of metabolic engineering and synthetic biology is possible through systems biology. Improving and modifying cellular properties such as product yield by the organism, production of new enzymes essential for biodegradation, substance uptake, and cell robustness against toxic compounds are warranted in order to realize the full potential of microbial bioremediation (Nielsen et al. 2014). PET plastics account for the 7.7% global demand distribution by resin type in 2018 (Plastic Europe 2019). Ideonella sakaiensis 201-F6 is a promising candidate in degradation of polyethylene terephthalate (PET) plastic. PETase produced by Ideonella sakaiensis 201-F6 exhibits structural similarity with cutinase and lipase (Austin et al. 2018). Mutation of two active-site residues of PETase in order to narrow the binding cleft improved the degradation (Austin et al. 2018). Leaf-branch compost cutinase (LCC) also received much attention due to its capability to depolymerize PET (Sulaiman et al. 2012). In a recent study, LCC was enhanced via (1) mutagenesis of the enzyme to enhance the PET-depolymerization specific activity and (2) addition of disulfide bridge to provide higher thermostability (Tournier et al. 2020). The depolymerization of PET was further improved to 90% conversion of 200 g PET in just a span of 2 h. Moreover, the group was able to recreate new PET plastics out of products of depolymerization, a potential solution to high global plastic volume. This study has been successful in harnessing the potential of microbial enzymes in the biodegradation of plastics. This has also demonstrated the feasibility of circular economy by reducing the global production of virgin plastics as they were able to recreate plastics out of the enzyme degradation products.

The authors would like to express their gratitude to the De La Salle University - University Research Coordination Office (DLSU - URCO) for their support. Declarations Conflict of Interest The authors declare no conflict of interest. Ethics approval and consent to participate The article does not contain any studies involving human participants or animals.

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