Virgin polymers via pyrolysis – A review of heteroatom removal options

Fuel Processing Technology 254 (2024) 108031 Available online 13 January 2024 0378-3820/© 2024 The Authors. Published by Elsevier B.V. This is an open access article under the CC BY license (http://creativecommons.org/licenses/by/4.0/). Virgin polymers via pyrolysis – A review of heteroatom removal options Jan Snow a , b , * , Pavel Kur´ aˇ n b , Aleˇ s Kaˇ sp´ arek a , Pavel Leˇ stinský c , Robert Suchopa d a ORLEN UniCRE a.s., Revoluˇ cní 1521/84, 400 01 Ústí nad Labem, Czech Republic b Faculty of Environment, Jan Evangelista Purkynˇ e University in Ústí nad Labem, Pasteurova 3632/15, Ústí nad Labem 400 96, Czech Republic c Institute of Environmental Technology, CEET, VSB-TUO, 17. listopadu 2172/15, 708 00 Ostrava, Czech Republic d ORLEN Unipetrol RPA s.r.o., Z´ aluˇ zí 1, 436 70 Litvínov, Czech Republic ARTICLE INFO Keywords: Plastic waste Chemical recycling Pyrolysis Contaminants Product upgrading ABSTRACT This review is dedicated to the removal of heteroatoms from plastic pyrolysis liquid products for use in the petrochemical industry. The chapters are devoted to removing individual groups of heteroatoms and summarizing the current scientific knowledge on the subject. Attention is given to the possibilities of heteroatom removal at all stages of the recycling process except sorting. Most of the findings in this area relates to the halogen removal, where high efficiencies can be achieved already in the pyrolysis process. In contrast, the removal of other heteroatoms has mainly been studied in the liquid product, usually in a hydrogen atmosphere and in the presence of a catalyst. It seems economically feasible to remove the heteroatoms as early as possible in the recycling process. This can be achieved in part by washing the waste plastic in water, which can remove a large proportion of the heteroatoms present as impurities. The work highlights the need for comprehensive mapping of heteroatoms in products and, in many cases, the need for more data regarding their removal. Finally, conclusions are drawn for further research in this area. 1. Introduction Many things come to mind when considering plastics. On the undesirable side, there are mostly environmental pollution and health risks, while on the other side, there are all the benefits that come with the great properties of these materials. To deal with the negative side, there are several things that society needs to address if we are to continue to produce and benefit from plastics. First, the attitude of consumers and governments needs to change to recognize the value of these materials even at the end of their life cycle. Data from PlasticsEurope suggests that public awareness is moving in the right direction, at least in Europe [1]. Nevertheless, of the 29.5 million tonnes of plastics recollected in 2022, approximately 2/3 were either incinerated or landfilled in the EU, which is considered a progressive political body in terms of recycling policy. Considering further legislative measures such as limiting municipal waste landfilling [2], the EU is also fulfilling its Paris Agreement commitments under the framework of the Green Deal [3]. In addition, plastics have been identified as a key value chain in the Circular Economy Action Plan of the European Commission, and according to the Packaging and Packaging Waste Directive, 50% of plastic packaging should be recycled by 2025 [4]. Many plastics producers also call for a mandatory recycled content of 30% in plastic packaging by 2030 [5]. It follows that the emphasis on recycling will continue to grow in the coming years, and the question is whether the technology is ready for this the challenge. At present, recycling techniques can be broadly classified as mechanical and chemical, each with its advantages and pitfalls. The benefit of mechanical recycling is that it is the fastest and easiest way, mainly for PET, PE, and PP, to be reintroduced as new products. Basically, it involves their melting and re-molding [6]. However, this process cannot be repeated countless times and the product quality deteriorates mainly due to additives, impurities, and other polymers in the feedstock [7]. Some initiatives even produce products from blends of different polymers and other materials, such as wood [8]. Whether or not this is considered, this is the end of simple recycling for these alternative materials. There is clearly a need for chemical recycling technology that can produce new virgin polymers even from more diverse feedstocks. Chemical recycling can also address the challenge of a limited plastic waste pool, where market competition for more valuable fractions has already begun, and players able to utilize more diverse and complex feedstocks may have a significant advantage. There are several approaches to chemical recycling that should work * Corresponding author at: ORLEN UniCRE a.s., Revoluˇ cní 1521/84, 400 01 Ústí nad Labem, Czech Republic. E-mail addresses: [email protected] (J. Snow), [email protected] (P. Kur´ aˇ n), [email protected] (A. Kaˇ sp´ arek), [email protected] (P. Leˇ stinský), [email protected] (R. Suchopa). Contents lists available at ScienceDirect Fuel Processing Technology journal homepage: www.elsevier.com/locate/fuproc https://doi.org/10.1016/j.fuproc.2024.108031 Received 19 October 2023; Received in revised form 18 December 2023; Accepted 1 January 2024 Fuel Processing Technology 254 (2024) 108031 2 in sync with mechanical recycling to harness the best of both methods. These can be broadly divided into solvolysis and thermochemical methods, including pyrolysis, which is particularly suitable for PE, PP, and PS [9], but potentially also for PVC and other polymers [10]. In contrast, PET is a favorable feedstock for solvolysis processes [11] and is undesirable for pyrolysis due to decreased dehalogenation efficiency [12] and formation of solid terephthalic and benzoic acids that cause system clogging [13]. Pyrolysis as a process of heating organic material, is far from new. Our ancestors used it to produce charcoal [14], which is a solid product of pyrolysis. More recently, once we were able to prevent air access to the feedstock and reduce the residence time of the products in the reaction zone, we were able to produce gaseous and liquid products consisting of various hydrocarbons depending on the feedstock. Given the planned transition away from combustion engines, the responsible use of these products can be narrowed down to the petrochemical industry. However, unlike crude oil, the pyrolysis products of plastic waste contain a large part of the periodic table, from sodium to biogenic elements, metals, and halogens [15]. This prevents the use of pyrolysis products within petrochemical processes, as corrosion [16], catalyst poisoning, and process fouling may occur [17]. In addition, the use of sorted waste for mechanical recycling leaves only complex waste streams, such as sorting line rejects, e-waste, etc., as potential feedstocks for pyrolysis. These streams are expected to contain higher levels of contaminants and therefore the issue of heteroatom removal is critical. Regarding heteroatoms in liquid pyrolysis products, Kusenberg et al. [18] compiled a comprehensive review mapping the heteroatoms likely to be found in the pyrolysis liquid from municipal plastic waste (MPW) and the implications of their presence for the steam cracker, which according to Palos et al. [19] has the highest potential to utilize the waste plastic pyrolysis liquid, along with FCC, hydroprocessing, and coker units. In their follow-up study, Kusenberg et al. [20] provided a general overview of possible approaches to improve the properties of the pyrolysis liquid product. However, despite the many reviews on the pyrolysis of plastics [20–28], we have not found a study that focuses in more detail on the removal of heteroatoms, with the exception of works that deal exclusively with dehalogenation [29–31]. Moreover, the authors often focus on the hydrocarbon composition and do not consider the heteroatoms in the products [32], so there is a significant body of work describing processes using different catalysts that are certainly affected by these impurities as well as the quality of the final product [33]. Therefore, this review aims to summarize the current scientific knowledge and state of the art to give the reader a comprehensive source of information on possible ways to mitigate problematic elements in pyrolysis products prior to their introduction into the petrochemical industry, and to identify knowledge gaps for further pursuit of the outlined recycling goals. 2. Method The work is organized into chapters according to groups of elements rather than individual feedstocks, as there is usually an overlap in the occurrence of individual elements in common pyrolysis feedstocks. The overall emphasis of the article and discussion is on the use of the liquid pyrolysis product, which is considered the most perspective in terms of subsequent petrochemical utilization [34]. Sorting is not addressed on the assumption that sorted feedstock can be used for mechanical recycling or solvent methods, while pyrolysis has to deal with the remaining, more complicated waste streams. For this reason, preference is given to studies dealing with real wastes or more complex model mixtures. In each section, the efficiency of heteroatom removal, e.g. dechlorination, debromination, desulfurization, demetallization, etc., is referred to as DE-X. D stands for specific de-reaction, E for efficiency, and X =Cl, Br, S, etc. Tables with studies listed by DE have been included to compare efficiencies. Although different analytical methods and approaches to the experimental part can influence efficiencies, general trends can still be seen from the tables. Where possible, DE values were taken as reported by the authors or extracted using the open-source software WebPlotDigitizer [35]. If not stated, the DE was calculated as the difference between the heteroatom content before and after preor post-treatment. For on-line methods, the difference in heteroatom content after pyrolysis with and without purification treatment was used for DE calculation. Studies that did not address overall DE are usually not mentioned. The reader should also be aware that authors report different DE values depending on whether they refer to char, liquid, or gas. For example, a study by Chen et al. [36] shows that 79% Br fixation in the solid residue does not translate into Br removal in the liquid. Therefore, the DE values in the tables refer to the pyrolysis feedstock in the case of pre-treatment or directly to the liquid, which is the main subject of this work. Each chapter is divided according to the location of the purification technique in the processing chain, i.e. pre-treatment, on-line treatment, and post-treatment. Where applicable, a distinction is made between the location of the purification agents within the reactor (insitu) or in a separate vessel (ex-situ). There are some discrepancies in the use of these terms. While some authors refer to in-situ when the agent is mixed with the feedstock and to ex-situ when the agent and feedstock are not mixed [37], in this paper we have sided with those authors who further distinguish between in-situ liquid phase (agent mixed with the feedstock) and in-situ vapor phase (agent above the feedstock) [38]. Thus, ex-situ refers to cases where the pyrolysis products have left the reactor and react with agents in a separate vessel, often at a different temperature. 3. Halogens Halogens appear in products after pyrolysis of plastics containing various organohalogen additives or plastics with halogen directly incorporated into the polymer structure, such as PVC, PTFE, etc. A major group comprises brominated flame retardants (BFRs) such as polybrominated diphenyl ethers (PBDEs), hexabromocyclododecanes (HBCDs) or tetrabromobisphenol A (TBBPA), the latter of which can also be incorporated into the polymer structure of epoxy resins. BFRs are often stable compounds that can accumulate in the environment. Due to their toxicity, many of them have been banned, but there are new BFRs that have different structures and are therefore legally produced [39]. Another widespread group of halogenated additives are plasticizers with flame retardant properties, such as short and medium-chain chlorinated paraffins (SCCP-MCCP) [40]. One of the most challenging sources of the above additives and polymers is waste electrical and electronic equipment (WEEE), which contains high levels of both Br and Cl. Table 1 summarizes the common compounds formed during the pyrolysis of these substances and materials. The main concern with halogens is the corrosion that occurs directly during the pyrolysis of plastics and in the subsequent condensation stage. Fig. 1 shows a picture of the condenser inlet from our laboratory unit where severe pitting was observed after several trials with larger quantities of PVC. The rest of the unit was not damaged in this way, and the problem was solved by a new Teflon inlet that prevented corrosion Table 1 Halogenated products from pyrolysis of common halogenated plastics. PVC [41,42] WEEE [43] PTFE [44–46] HCl HBr tetrafluoroethylene chloroalkanes Bromomethane hexafluoropropene chlorocycloalkanes Bromophenols octafluorocyclobutane chloroalkenes Bromobisphenols 1and 2octafluorobutylene chlorocycloalkenes 1-bromo-3-phenoxy-2propanol chlorobenzenes 1,3-dibromo-2-propanol chloroalkylbenzenes Dibromodimethylbenzene J. Snow et al. Fuel Processing Technology 254 (2024) 108031 3 from condensing HCl. Although proper design of the condensation stage is important, corrosion caused by organochlorides in downstream petrochemical processes is of even greater concern. As reported by Alanazi et al. [16] on the naphtha hydrotreater unit, even 10 ppm of organochlorides in the feed can lead to severe corrosion and technology shutdown. In comparison, the industrial threshold is considered to be 2–3 ppm. Among the halogens, much work has been done on the removal of Br and Cl, which is summarized below. In the case of iodine and fluorine, the latter is expected, for example, in feedstocks containing polytetrafluoroethylene (PTFE) [47], while iodine is not a structural component nor is it commonly added as an additive to plastics and is therefore not reported in pyrolysis products. 3.1. Fluorine In the case of fluorine, there appears to be no method that addresses the issue of dehalogenation to this date. Due to the high energies required to break the C-F bond, the C-C bonds are broken instead to form the primary product tetrafluoroethylene (TFE) and difluorocarbon radicals. While TFE yields can reach 97% under certain conditions [44], subsequent secondary reactions lead to other products such as hexafluoropropene, octafluorocyclobutane, and other fluorocarbons [47,48]. The mechanisms of fluorocarbon formation during pyrolysis are well documented in the literature [45,49]. Therefore, rigorous sorting of these materials is necessary to avoid higher fluorine content in pyrolysis oils. Subsequent pyrolysis of sorted PTFE can provide high yields of the above monomers for further use. 3.2. Chlorine Chlorine is a common heteroatom in MPW, entering the liquid pyrolysis product mainly from chlorinated plastics such as PVC [50] or contamination from inorganic salts [51]. In terms of its removal, chlorine has received the most attention among the listed elements. 3.2.1. Pre-treatment Based on the literature on dechlorination, four possible approaches have been identified: washing, solvothermal, mechanochemical, and microwave pre-treatment. An overview is given in Table 2. Although most of the listed studies were performed with model mixtures of often finely ground virgin plastics, functional mechanisms can be identified and extended to more complex feedstocks. Starting with solvent-based methods, Table 2 shows that up to 74% of the initial Cl content was removed from the pyrolysis feedstock by washing the feedstock in water [51,52] or organic solvent [53], which involves removal of surface contaminants by mechanical forces combined with dissolution of soluble impurities. However, when Genuino et al. [51] pyrolyzed the washed feedstock at 500 ◦C, the DE-Cl in the final liquid product was only 33% because washing mainly removes inorganically bound Cl, which also leaves the system as HCl during pyrolysis. To further increase the DE-Cl, the authors increased the temperature of the solvent or added additives to the solution. It should be noted that once the conditions exceed the threshold for polymer decomposition, the liquid product is produced. However, these approaches, which would otherwise be listed in the on-line section, are listed in this section for comparison. The results of studies using solvents in the pre-treatment of PVC/ PVDC-containing feedstocks in Table 2 suggest that several factors affect dechlorination, such as temperature, pressure, and reaction time. There is also the effect of the used solvent, additives, and feedstock composition. The reaction mechanism proceeds via two main coinciding dechlorination pathways - substitution of a nucleophile, usually from -OH to -Cl, and HCl elimination followed by dissolution of HCl in the solvent [69,88,89]. The substitution pathway leads to the formation of alcohols, which, along with the rest of the feedstock, can be further transformed via hydrolysis, cyclization, oxidation, and further thermal decomposition to yield other O-containing compounds [57,90,91]. This was recently confirmed by Xiu et al. [54] who performed FT-IR analysis of the solid residue after hydrothermal treatment of PVC and reported the substitution of -Cl by -OH groups. After adding pieces of LCD panels, which acted as a catalyst, the same team reported the formation of a more stable alkene‑carbonyl structure. Unfortunately, we have not found a study on the subsequent pyrolysis that addresses the interfering oxygen from the pre-treatment [59]. On the plus side, the authors report that there was little to no formation of chlorinated hydrocarbons in both supercritical [92] and subcritical [57,90,93] conditions. The predominant mechanism differs between studies due to differences in experimental conditions. In general, the rate of substitution and formation of oxygenates increases with reaction time [69], temperature [90], water density [94], and pressure [72]. Therefore, the substitution of -OH to -Cl may decrease with shorter reaction time and lower hydrothermal treatment pressure, as suggested by Zhao et al. [72]. However, this may also be related to the lower overall DE-Cl reported by the authors. Dechlorination typically occurs above 200 ◦C and increases with temperature [57,66,90,95]. Authors have also experimented with additives, mainly alkaline agents [64], acid catalysts [64], and various forms of metals [61,64,74,96]. Table 2 shows that especially alkaline agents such as NaOH in small amounts significantly improved DE-Cl. For example, Wang et al. [59] report DE-Cl improvement from 81% to 99.9% using 0.3 M NaOH. However, the effect of additives decreases with increasing temperature [59,74], suggesting that the trade-off between higher temperature and additive addition should be considered. The solvent effect can be demonstrated by comparing studies in which the authors used DMSO [60] and water [55] in combination with NaOH to dechlorinate PVC. In the latter study, the authors achieved 100% DE-Cl at 250 ◦C, whereas in the former study, the authors reported 99% DE-Cl at 80 ◦C after a similar time (3–5 h). According to the authors, DMSO could be recovered and reused, which may justify the use of this more expensive solvent. In a later study, a similar group of authors [56] reported the use of NaOH/EG solution to achieve 100% De-Cl at 190 ◦C, but only for a particle size of 120 μ m. In the case of pellets (4 mm), the DE-Cl was about 14%. In addition to the effect of particle size, which may limit some of the described approaches due to shredding requirements, it is appropriate to discuss the influence of feedstock composition. Table 2 shows that most of the studies were carried out on PVC without any interfering Fig. 1. Image of the condenser inlet after several trials with high PVC levels. J. Snow et al. Fuel Processing Technology 254 (2024) 108031 4 components in the waste plastics. The results of the listed studies indicate that the DE-Cl decreases with the complexity of the feedstock and that high DE-Cl is achieved with more demanding conditions [63]. The hydrothermal treatment of probably the most complex feedstock, municipal solid waste (MSW), was investigated by Prawisudha et al. [97]. The authors report that the organic chlorine content (insoluble in water) was reduced by about 80% after treatment at 225 ◦C. However, chlorine was retained in the solid product and would require subsequent water washing to remove inorganically bound Cl. This is consistent with another study in a similar setting using a model mixture of PE, PP, PS, PVC, and cellulose, where the authors achieved a 58% reduction in total Cl. When Ca(OH) 2 was added, the Cl removal from the hydrotreated product decreased to 17% due to the formation of CaCl 2 [61]. This suggests that elements that can form chlorides in the feedstock may inhibit dechlorination. The authors mention the possibility of Cl removal by subsequent washing, which they address in another article [52], reporting 74% removal from hydrotreated MSW and 59% from PVC. Another examined group of dechlorination methods is mechanochemical pre-treatment, which relies on the transfer of mechanical energy to the feedstock during the grinding process instead of thermal energy from heating, and for which different types of mills can be used [98]. From the individual studies summarized in Table 2, it can be seen that most studies used pure, finely ground PVC powder [75,76,78,80], with only a few studies using post-consumer PVC [77,79,82], and one using more complex wastes such as automotive shredder residue (ASR) [81]. Although efficiency generally decreases with feedstock complexity, the authors achieved relatively high efficiencies even with waste PVC (99%) [77] and ASR (92%) [81] using a ball mill reactor with NaOH/EG solution at 190 ◦C, which is one of the lowest temperatures found for more complex feedstocks. Furthermore, based on the life cycle assessment and environmental impact of the above system, grinding appears to be less energy intensive and therefore likely to be more Table 2 Overview of pre-treatment dechlorination methods listed by dechlorination efficiency (DE-Cl). Method Solvent Agents Feedstock Particle size T;p t (h) DE-Cl Ref. Washing H 2 O – hydrotreated MSW* n.r.* 60 ◦C; atm. 0.5 74% [52] H 2 O – MPW 12 mm 85 ◦C n.r. 66% [51] H 2 O – hydrotreated PVC n.r.* 60 ◦C; atm. 0.5 59% [52] light naphtha – motor oil flasks 3 mm n.r.* 0.5 31% 1,4) [53] Solvothermal H 2 O LCD PVC pellets 3–5 mm 250 ◦C; 5 MPa 1.5 100% [54] H 2 O NaOH PVC pellets 0.2 g 250 ◦C 3 100% [55] ethylene glycol NaOH flexible PVC 120 μ m 190 ◦C 4 100% [56] H 2 O – PVC powder <0.25 mm 235 ◦C 15 100% [57] H 2 O H 2 O 2 PVC powder 500 μ m 350 ◦C 1 100% [58] H 2 O NaOH 3P 2) ,PVC 0.1–5 mm 240 ◦C 2 99.9% [59] DMSO NaOH PVC powder <44 μ m 80 ◦C; atm. 3 99% [60] H 2 O NaOH, H 2 O 2 , Pd/AC waste PVC n.r.* 180 ◦C 16 99% [61] H 2 O K 2 CO 3 PVC from wires <3 mm 250 ◦C; 3.8 MPa 4 99% [62] CO 2 – display housing 841 μ m 398 ◦C; 0.6 MPa 1.3 99% [63] H 2 O – PVC powder n.r.* 300 ◦C 1 97% [64] H 2 O Ammonia PVC pipe 50 mg 250 ◦C; 10 MPa 2 96% 1) [65] H 2 O – PVC from wires <1 mm 300 ◦C; 8,6 MPa 3 96% [66] H 2 O CaO PVC/agent pellet 1 ×8.5 ×0.5 cm 200 ◦C 1 95% [67] MeOH – medical waste 2 ×2 cm 300 ◦C 1 95% [68] ethylene glycol NaOH PVDC powder 150–180 μ m 190 ◦C 3 93% [69] H 2 O – PVC pipe 50 mg 250 ◦C; 10 MPa 2 85% 1) [65] glycerol – PVC film several mm 240 ◦C 2 82% [70] H 2 O – model medical waste n.r.* 300 ◦C 0.5 82% [71] H 2 O ZnO/CoO PVC powder n.r.* 200 ◦C 1 70% [64] H 2 O Na 2 CO 3, Ni(NO 3 ) 2 ⋅6H 2 O PVC powder ~0.1 mm 220 ◦C 0.5 65% [72] H 2 O – 3P 2) ,PVC,cellulose n.r.* 215 ◦C; 2.4 MPa 1.5 58% [73] H 2 O Cu(NO 3 ) 2 PVC powder 0.1–0.2 mm 220 ◦C 0.5 58% [74] H 2 O NaOH PVC powder n.r.* 200 ◦C 1 21% [64] H 2 O Ca(OH) 2 3P 2) ,PVC,cellulose n.r.* 225 ◦C; 2.4 MPa 1.5 17% [73] ethylene glycol NaOH flexible PVC 4 mm 190 ◦C 4 14% 1) [56] H 2 O – PVC powder n.r.* 200 ◦C 1 13% [64] ethylene glycol – PVDC powder 150–180 μ m 190 ◦C 3 6% [69] Mechanochemical – CaO,SiO 2 ,Al 2 O 3 PVC powder 133 μ m ambient 4 ~100% [75] – oyster-shells PVC powder 40–300 μ m ambient 8 ~100% [76] ethylene glycol NaOH waste PVC (100 g) 1 cm 190 ◦C 5 99% [77] ethylene glycol NaOH flexible PVC pellets 150–250 μ m 190 ◦C 2 97% [78] – eggshells waste PVC <160 μ m n.r.* 4 95% [79] – KOH PVC powder 133 μ m ambient 6 94% 1) [80] ethylene glycol NaOH ASR 3) n.r.* 190 ◦C; 0.1 MPa 5 92% [81] ethylene glycol NaOH rigid PVC pellets 150–250 μ m 190 ◦C 6 84% [78] – eggshells waste PVC (100 g) <1 mm ambient 12 50% [82] Microwave H 2 O NaOH flexible PVC <1 mm 235 ◦C (1 kW) 0.5 100% [83] ethylene glycol NaOH flexible PVC <1 mm 160 ◦C 0.17 99.7% [84] – AC, Zn-Mn ferrite powder PVC powder n.r.* n.r.*(22 W/h) 1 98% [85] – AC, Zn-Mn ferrite powder PVC flooring material 5x3x2 mm n.r.*(60 W/h) 1 97% [85] – SiC PVC, PVDC, CPE 5 ×5 mm 300 ◦C (800 W) n.r.* 87% [86] – – PVC powder 141 μ m 220 ◦C (1.1 kW) 1.5 68% [87] – – PVC, PVDC, CPE 5 ×5 mm 300 ◦C (800 W) n.r.* 66% [86] 1) based on non-numerical data extracted using the open-source software WebPlotDigitizer [35] 2) mixture of PE, PP, and PS 3) automotive shredder residue 4) DE of the final pyrolysis liquid *MSW - municipal solid waste; n.r. - not reported. J. Snow et al. Fuel Processing Technology 254 (2024) 108031 5 environmentally friendly than heating [81]. Methods using solid Cl sorbents such as eggshells [79], NaOH or KOH [80] achieved lower efficiencies (<95%). Efficiency decreases further during scale-up with increasing amounts of feed and particle size [72]. In addition, further washing is required to remove chlorides generated from the products after dechlorination. For more detailed information, the reader is referred to reviews on the mechanochemical treatment of halogenated organic pollutants [99] and various wastes, aimed at both mechanical [100] and chemical recycling [30,101]. Finally, the last group of pre-treatment methods found in the literature can selectively remove Cl from the feed by microwave (MW) irradiation due to the polarity of the C-Cl bond. Authors report several advantages of this approach, such as lower required temperature, faster dechlorination, and the need for less concentrated reagent solutions [83,84,87]. However, as the complexity of the feedstock increases, the DE-Cl decreases, as observed by Moriwaki et al. [85], who compared pure and various additivated PVC plastics and reported that more energy is needed in the case of waste material. For example, 98% DE-Cl was achieved with pure PVC at ~ 22 Wh, compared to 26% DE-Cl with pipe material, which required 36 Wh to achieve similar efficiency. For flooring, the energy requirement was >60 Wh. It should also be noted that an additional microwave absorbent was added to the feedstock to improve the DE-Cl. However, an additional MW absorbent is not always necessary, as shown in a study combining NaOH/EG solution with MW irradiation [84]. Based on the listed literature and relevant efficiencies in Table 2, the highest DE-Cl was achieved by combining MW heating with NaOH solutions [83,84] or additional MW absorbents [85,86]. However, DE-Cl was not generally increased compared to conventional heating sources used in hydrothermal or on-line methods, which are discussed below. Microwave heating can also be used during pyrolysis as discussed in the review by Hu et al. [102]. The reported efficiencies can be compared to the stepwise pyrolysis studies described below, where the authors often used similar model mixtures and reported similar efficiencies by simply adjusting the temperature program [42,103]. Therefore, the overall economics of the pretreatment and other benefits or pitfalls, such as oxygen introduction and wastewater production, should be considered before further scaling the technology. 3.2.2. On-line treatment PVC is known to degrade in two main steps upon heating, with much of the Cl being released as HCl in the 1st step, which occurs at temperatures up to 360 ◦C [104]. The degradation of SCCP-MCCP also follows a similar dehydrochlorination mechanism [40,41]. However, it has a lower thermal stability, as reported by Xin et al. [105], who observed 92% weight loss at 210–330 ◦C. The dehydrochlorination step is well documented and used in stepwise pyrolysis, which is usually carried out in two steps, the first one up to 360 ◦C, and after sufficient time, the temperature is further increased in the second step to ensure maximum yield of liquid product. The other approach is to use dechlorination agents, usually various solid sorbents and catalysts, to remove Cl that passes into the liquid product during the second decomposition step. It is also feasible to combine stepwise pyrolysis with dechlorination agents. Table 3 summarizes the studies that have addressed the dechlorination of liquid products during pyrolysis of plastic waste or model mixtures that simulate the waste composition. The studies listed are categorized according to the achieved DE-Cl, with only the significant or optimal results included in each section (agent, stepwise, and agent +stepwise). In particular, we excluded studies on the pyrolysis of PVC alone [106] and pure binary mixtures [107], with the exception of one [108], in order to focus on the actual waste. Studies were also excluded if there was no analysis of the chlorine content in the liquid [109] or if the results could not be compared due to insufficient data on the chlorine content in the liquid product [110]. In addition, the feedstock-to-agent ratio (F/A) is included to allow a better evaluation of individual agents, as agents with a lower F/A may give better results that may not reflect the actual DE-Cl of the agent. This is relevant for scenarios where large amounts of agent result in complete dehalogenation [111], making it difficult to interpret the calculated DE-Cl without factoring in the amount of agent used. 3.2.2.1. Stepwise pyrolysis. From Table 3, it can be concluded that the DE-Cl decreases with increasing feedstock complexity towards MPW. The exception was found in the study by Park et al. [119], where the lower DE-Cl of waste LDPE/PVC stepwise pyrolysis may have been caused by the combination of lower temperature (298 ◦C) and shorter residence time (10 min) of the feedstock in the 1st step. In order to elucidate the cause of the negative influence of MPW, the DE-Cl of stepwise pyrolysis with simpler virgin polymer mixtures should be investigated, since the degradation and potential for dechlorination via stepwise pyrolysis are well described in the literature mainly for PVC and model, often binary, mixtures [41,42,131]. There are even some contradictory results reported, which should be pointed out. According to the TGA results of Miranda et al. [104], virgin PE, PS, and especially PP can stabilize the HCl evolution via H-abstraction by Clradicals, which leads to a decrease of the autocatalytic decomposition by HCl. Yuan et al. [132] came to analogous conclusions in a TGA analysis of polymer mixtures and reported a significant reduction of DE-Cl for LLDPE and PP. However, this is not fully consistent with the subsequent study by Miranda et al. [42], where the authors obtained comparable HCl release during the 1st step at 360 ◦C in the pyrolysis of mixed plastics, although the composition of chlorinated hydrocarbons differed from that observed in the pyrolysis of pure PVC. Some studies have also reported no change in HCl evolution based on Py-GC–MS [133] and TGA [134] of polymer mixtures with PVC. Cz´ eg´ eny et al. [135] even report improved DE-Cl in the presence of ABS and PET, which seems to contradict earlier studies by a similar group of authors [136] and others [137]. The above findings are not fully consistent, and the linking factor may be the difference in scale, as it appears that the micro-scale results from TGA and Py-GC–MS do not correlate with the larger scale results from pyrolysis, especially in the case of PET, where authors often report a significant negative influence [120]. This probably indicates the influence of parameters such as the residence time of the products in the reaction zone, which increases the likelihood of interactions between HCl and other hydrocarbons. Negative interactions between feedstock components play a key role in dechlorination, yet there is insufficient data in this regard. PET/PVC interactions are the most widely described. The authors have reported the formation of chlorinated derivatives of terephthalic and benzoic acids, which are known degradation products of PET [138], together with a plausible mechanism [137]. On an industrial scale, Fukushima et al. [13] reported that the largest pyrolysis plant to date experienced clogging and corrosion caused by benzoic acid. This problem was solved by adding Ca(OH) 2 to the feedstock for neutralization. Until recently, there was limited information available to quantitatively assess the influence of PET on the stepwise pyrolysis DE-Cl. In a previous study, we found that the addition of PET (12 wt%) to the model plastic mixture during stepwise pyrolysis resulted in the formation of 2chloroethyl benzoate and a significant decrease (30%) in HCl production associated with a large increase in liquid Cl content from 86 ±12 ppm to 3591 ±944 ppm [12]. In a study by L´ opez et al. [118], PET probably also caused lower DE-Cl, although the authors did not specifically address this issue. Furthermore, in the same study, the authors indirectly addressed additives and their possible negative effect on dechlorination, as they observed that the addition of CaCO 3 to the feedstock during conventional pyrolysis increased the Cl content in the liquid. However, the same effect was not observed after stepwise pyrolysis. Similarly, in a previous study, we observed a significant increase in Cl content under similar conditions with Ca(OH) 2 [114], but no effect after stepwise pyrolysis with CaCO 3 [12]. The results indicate that some of the Cl is fixed in the reactor as CaCl 2 , and L´ opez et al. concluded that during stepwise J. Snow et al. Fuel Processing Technology 254 (2024) 108031 6 Table 3 Overview of on-line dechlorination methods sorted by calculated dechlorination efficiency. Method Feedstock Temperature 1) Yield difference 2) Cl content (ppm) DE-Cl 5) F/A 6) ref. Without 3) With 4) Stepwise pyrolysis 3P*, PVC 330 ◦C; − − 4% 390 n.d.* 100% – [112] 3P*, PVDC 300 ◦C; −m.d.* n.r.* 15 m.d.* – PE, PS, PVC 330 +380 ◦C; −m.d.* n.r.* 44 99.6% – [113] 4P*, PVC 350 ◦C; − +4% 7789 86 99% – [114] waste 3P*, PVC, PET 300 ◦C; −m.d.* n.r.* 6 98% 7) – [115] waste LDPE, PP, PVC 300 ◦C; − − 6% 3685 464 87% – [116] WEEE 300 ◦C; −0% 11,532 3283 72% – [117] 4P*, PVC, cellulose 350 ◦C; −m.d.* n.r.* 2974 71% 8) – [12] 4P * *, PVC, PET 350 ◦C; m.d.* n.r.* 3591 67% 8) – [12] 3P*, PVC, waste PET 300 ◦C; − − 7% 5000 2000 60% – [118] MPW 350 ◦C; −0% 269 167 38% – [12] waste LDPE/PVC 298 ◦C; − − 20% 425 367 9) 14% 9) – [119] Agents Ca-C (CaCO 3 ) - in-situ 3P*, PVC -;430 ◦C +4% 360 n.d.* 100% 2.5 [111] Fe-C (Fe 3 O 4 ) - in-situ HIPS-Br, PVC -;430 ◦C −10% 4300 n.d.* 100% 2.5 [111] Ca-C (CaCO 3 ) - in-situ 3P*, PVC, HIPS-Br -;430 ◦C m.d.* 1120 n.d.* 100% 2.5 [120] Ca-C (CaCO 3 ) - in-situ 3P*, ABS-Br, PVC -; 450 ◦C −5% 4972 113 98% 5 [121] CaO - ex-situ waste LDPE, PP, PVC -; 400 ◦C −1% 3685 62 98% 5 [116] Ca-C (CaCO 3 ) - in-situ 3P*, PVC, HIPS-Br, PET -;430 ◦C +3% 2850 100 96% 2.5 [120] Ca(OH) 2 , red mud, NiSAPO-11 - in-situ 4P*, PVC, ethylene propylene dimer -; 520 ◦C −7% 8) 4364 228 95% 20 [122] CaCO 3 - in-situ 3P*, ABS-Br, PVC -; 450 ◦C +5% 4972 355 93% 5 [121] Ca(OH) 2 – in-situ MPW -; 719 ◦C 0% 502 50 90% 28 [123] oyster shell – in-situ MPW -; 722 ◦C −1% 502 57 89% 10.5 [123] CaO – in-situ MPW -; 711 ◦C −2% 502 58 88% 18.7 [123] Fe-C (Fe 3 O 4 ) - in-situ 3P*, ABS-Br, PVC -; 450 ◦C +8% 4972 1014 80% 5 [121] Al-Mg - in-situ 3P*, PVC -; 420 ◦C +9% 10,300 3000 71% 5 [108] MgO - in-situ PP, PVC -; 380 ◦C −23% 16,500 7600 54% 5 [108] Al(OH) 3 - in-situ MPW -; 500 ◦C +19% 618 373 40% 10 [124] β-zeolite - in-situ MPW -; 500 ◦C +28% 618 399 35% 10 [124] Ni-Mo-catalyst - in-situ MPW -; 500 ◦C +21% 618 416 33% 10 [124] FCC - in-situ MPW -; 500 ◦C +29% 618 422 32% 10 [124] FeOOH - in-situ 3P*, ABS-Br, PVC -; 450 ◦C +4% 4972 3370 32% 5 [121] MoO 3 - in-situ MPW -; 500 ◦C +20% 618 451 27% 10 [124] Y-zeolite - in-situ MPW -; 500 ◦C +29% 618 457 26% 10 [124] HZSM-5 - in-situ MPW -; 500 ◦C +25% 618 487 21% 10 [124] γ-Al 2 O 3 PP,PVC -; 380 ◦C −9% 16,500 13,000 21% 5 [108] Ni-Mo-catalyst - in-situ MPW -; 500 ◦C +8% 8) 1285 1135 12% 20 [125] β-zeolite - in-situ MPW -; 500 ◦C +13% 1285 1273 1% 20 [125] Y-zeolite - in-situ MPW -; 500 ◦C +11% 8) 1285 1322 −3% 20 [125] CaCO 3 - in-situ 3P*, PVC, waste PET -; 500 ◦C −18% 5000 6000 −20% 21.5 [118] red mud – in-situ 3P*, waste PVC, PET -; 440 ◦C +5% 2000 5000 −150% 10 [126] ZSM-5 - in-situ 3P * *, waste PVC, PET -; 440 ◦C −22% 2000 12,000 −500% 10 [127] Agents þstepwise pyrolysis Ca(OH) 2 - ex-situ 4P*, PVC 350 ◦C; 300 ◦C +4% 7789 10 99.9% 10 [114] CaO - ex-situ waste LDPE, PP, PVC 300 ◦C; 400 ◦C −10% 3685 11 99.7% 5 [116] Ca(OH) 2 - ex-situ waste LDPE, PVC 304 ◦C; 699 ◦C −20% 425 9 97.9% 0.5 [119] hydrotalcite +β-zeolite (1/1) - ex-situ MPW 350 ◦C; 300 ◦C +3% 269 12 95.5% 10 [12] Na 2 CO 3 – ex-situ MPW 300 ◦C; 541 ◦C −3% 704 65 91% m.d.* [128] red mud - in-situ MPW 300 ◦C; 430 ◦C +7% 7435 1185 84% 5 [129] Ca-C (CaCO 3 ) - ex-situ MPW 350 ◦C; 350 ◦C −3% 1170 590 75% 2.5 [111] Ca(OH) 2 - ex-situ MPW 350 ◦C; 300 ◦C +3% 269 100 63% 10 [12] red mud - in-situ waste PP, PE, PVC 330 ◦C; 440 ◦C +9% 39,900 15,900 61% 5 [130] CaCO 3 - in-situ 3P*, PVC, waste PET 300 ◦C; 500 ◦C −3% 5000 2000 60% 21.5 [118] CaO - ex-situ MPW 300 ◦C; 546 ◦C +2% 704 287 59% m.d.* [128] red mud - in-situ PE, PS, PVC 360 ◦C; 440 ◦C −7% 17,500 15,600 11% 5 [130] Fe-C(Fe 3 O 4 ) - in-situ MPW, HVGO* 300 ◦C; 430 ◦C −10% 1127 1283 −14% 10 [129] Ca(OH) 2 - in-situ waste LDPE, PVC 302 ◦C; 706 ◦C −18% 425 578 −36% 0.5 [119] ZSM-5 - in-situ 3P*, waste PVC, PET 300 ◦C; 440 ◦C −22% 2000 3000 −50% 10 [127] 1) temperature of 1st pyrolysis step; temperature during reaction with agents 2) yield difference calculated as yield from 4) - yield from 3) 3) one-step pyrolysis without further modifications 4) pyrolysis with dechlorination treatment 5) calculated by comparing the Cl content in the final liquid product from 3) and 4) 6) feedstock to agent ratio 7) based on Cl in gases trapped in NaOH 8) calculated from Cl content after stepwise pyrolysis [114] without addition of PET and cellulose 9) calculation based on non-numerical data extracted using the open-source software WebPlotDigitizer [35]. *n.r. - not reported; n.d. - not detected; m.d. - missing data for calculation; 3P - PE, PP, and PS mixture; 4P - LDPE, HDPE, PP, and PS mixture; HVGO - heavy vacuum gas oil. J. Snow et al. Fuel Processing Technology 254 (2024) 108031 7 pyrolysis there is more time for the HCl to be fixed as chloride. However, this does not explain the difference in Cl content after conventional and conventional+additive runs. Therefore, it is likely that the increased amount of Cl in the liquid is due to reactions of CaCl 2 with other pyrolysis products, such as H 2 O or CO 2 , as suggested by Sophonrat et al. [110]. Water and CO 2 are mostly removed during the 1st step [139], and therefore CaCl 2 is more likely to withstand the entire pyrolysis process and remain in the char. This finding serves as another argument for stepwise pyrolysis, as CaCO 3 is a widely used filler additive [140]. The effect of other commonly used additives on dehalogenation remains unclear. However, according to Zheng et al. [141], the plasticizer dioctyl phthalate and the stabilizer calcium stearate are associated with a significant reduction in HCl evolution. It is worth noting that both additives have either carboxylic groups that may be susceptible to cleavage and subsequent C-Cl bond formation, as we have proposed in the case of PET [12], or in the second case, the CaCl 2 may also be formed. Interactions of other specific additives during dechlorination have not been found in the literature. While some studies have reported interactions resulting in increased Cl liquid content when comparing virgin and waste PVC [142], Yuan et al. [143], reported enhanced HCl release from waste PVC and no negative effect on dechlorination. However, neither study identified a specific additive to explain the cause. There may be other additives such as Fe 2 O 3 , ZnO, Sb 2 O 3 , and various organic additives that may interfere with the stepwise pyrolysis by a mechanism similar to that of CaCO 3 , i.e. formation of unstable chlorides, or as reported for dioctyl phthalate, formation of organochloride compounds. Regarding MPW, relatively few studies dealing with stepwise pyrolysis were found. According to Table 3, the level of DE-Cl during stepwise pyrolysis was significantly lower for MPW and WEEE compared to model plastic mixtures. In mixtures with pure polymers such as HDPE, LDPE, PP, and PS, PVC degradation seems to follow a similar trajectory as when pyrolyzed alone, and DE-Cl exceeds 99% [104]. Similar DE-Cl has been reported in the case of the vented screw conveyor, even when a waste plastic mixture containing PET was used, indicating the importance of rapid HCl removal and proper feedstock mixing [115]. Although conventional and stepwise pyrolysis cannot be compared in this case, the Japanese study [13], which includes information from a large-scale commercial operation, should be considered. Dealing with MPW with relatively high PET and PVC content, stepwise pyrolysis in a screw extruder at 350 ◦C removed about 96% Cl, and the authors also describe different DE-Cl with changes in screw design. The optimum temperature for dechlorination appears to be in the range of 300–360 ◦C, which is consistent with the findings for virgin PVC. However, while temperature and residence time are relatively easy to adjust, feedstock composition is often hard to define and interactions with heteroatoms in MPW need to be considered. In addition, cellulose, which has recently been identified as a material that negatively affects dechlorination, may be present in significant amounts in MPW [139,144]. Apart from studies reporting inhibition of HCl production during pyrolysis of cellulose/PVC mixture [145] and more chlorinated compounds in the products [146], a significant decrease in DE-Cl of stepwise pyrolysis is evident from Table 3, where it decreased from 99 ±0.2% to 71 ±13% upon addition of 12 wt% cellulose to the feedstock [12]. It can be concluded that PET, cellulose, and possibly other heteroatom-containing substances interact with the HCl formed during the 1st step of stepwise pyrolysis, leading to a significant decrease in DECl. HCl can shift polymer degradation to lower temperatures, resulting in a decrease in DE-Cl. This is associated with the formation of organochlorine compounds that lead to the dissipation of HCl autocatalysis and, more importantly, the retention of Cl until the temperature is further raised in the next step. Therefore, if the feedstock contains PET and cellulose, it is more likely that chlorinated compounds will be present in the liquid from the 2nd step. Thus, the use of dechlorination agents prior to the condensation stage may be justified. 3.2.2.2. Dechlorination agents. The DE-Cl of the agents is related to the interactions in the feedstock, and their efficiency is based on their ability to remove both HX (X =Cl, Br) and organically bound halogens. This is illustrated by Park et al. [116] who conducted experiments on washing the products of a two-stage pyrolysis of a waste mixture of LDPE, PP, and PVC in water. They separated HCl by washing and reported that most of the Cl remained in the organic fraction. When evaluating the DE-Cl of the agents in Table 3, it is clear that the agents listed have different characteristics that allow them to be classified into the following groups: Alkaline agents - CaCO 3 [118,121], Ca(OH) 2 [12,114,119,123], CaO [116,123,128], MgO [108], Na 2 CO 2 [128], hydrotalcite [12], oyster shells [123], Al(OH) 3 [124]. Acidic agents - β-zeolite [124,125], Y-zeolite [124,125], ZSM−5 [124,127], FCC [124], γ-Al 2 O 3 [108]. Metals and metal oxides - FeOOH [121], MoO 3 [124], red mud [126,130]. Mixtures and composites - Fe-C (Fe 3 O 4 ) [111,121], Ca-C (CaCO 3 ) [111,120,121], Ca(OH) 2 /red mud/NiSAPO-11 [122], Al-Mg [108], hydrotalcite/β-zeolite [12], Ni-Mo catalyst [124,125]. The most widely employed group was alkaline agents, which, according to Table 3, together with composites and mixtures of agents, gave the best results. In contrast, metals, metal oxides, and acidic agents showed lower levels of DE-Cl, except when combined with alkaline agents [12,122]. Also, as in the case of stepwise pyrolysis, the DE-Cl of many of the listed sorbents falls sharply when used during MPW pyrolysis compared to model mixtures, as seen in the examples of Fe-C (Fe 3 O 4 ) composites [111,121,129], Ca-C (CaCO 3 ) [111,120], and others. Among the studies listed, the highest DE-Cl and the lowest Cl content in the liquid from pyrolysis of MPW or similar complex feedstocks containing PET were found in our previous study [12]. There, we combined stepwise pyrolysis with a mixture of hydrotalcite and β-zeolite in a 1:1 ratio, placed ex-situ, to obtain a liquid with 12.2 ±1.2 ppm, reaching 95% DE-Cl. We connected the results with the findings of Jiang et al. [147], who reported a more complicated dechlorination in the case of aromatic hydrocarbons compared to aliphatic ones. Similar to our results, the authors report a synergistic effect towards chlorobenzene dechlorination after combining acidic (alumina) and alkali agents (Na 2 CO 3 , CaCO 3 ). This may explain the reduced DE-Cl of otherwise active alkali agents when pyrolyzing more complex mixtures, as more organohalogens are formed, making Cl inaccessible to the agents. Another factor may be related to the finding of Sophonrat et al. who observed lower DE-Cl of CaO at a temperature of 400 ◦C in the presence of cellulose. In their work, the authors discuss the likely cause as interactions of cellulose pyrolysis products such as H 2 O with CaCl 2 [110]. With respect to the location of the agent, Table 3 shows that good results have generally been achieved with the ex-situ setups [116], also in combination with stepwise pyrolysis [12,114,116,119,128], which can further reduce the exposure of the agent and thus increase its time on stream. In the case of in-situ setups, the results are less conclusive. On the one hand, there are results of complete or high levels of dehalogenation with model mixtures in both vapor [111,121] and liquid phase contact [122]. However, the same amount of agent showed a 25% decrease in DE-Cl during MPW pyrolysis [111]. A similar trend was also reported for the addition of PET (5 wt%) to the model mixture [120], confirming that MPW pyrolysis places higher demands on the agents due to the discussed interactions in the feedstock. As for the mechanisms, we have found three mechanisms that have been described and proposed in the literature so far. First, for HCl, there is sorption of inorganic Cl via a neutralization reaction forming chlorides [148]. For organohalogens, the reaction can follow either direct catalytic elimination or dissociative adsorption followed by β-H elimination [36]. HCl fixation and dechlorination of organic chlorine can also occur sequentially or simultaneously, as suggested by Yanik et al. [129], who observed these phenomena with red mud and other Fe-based agents. The last mechanism found involves a combination of alkali and acidic J. Snow et al. Fuel Processing Technology 254 (2024) 108031 8 agents, likely proceeding by formation of the carbocation R + via zeolitemediated hydrocarbon cracking followed by C-Cl bond cleavage in the presence of an alkaline sorbent [12,147]. The latter mechanism is illustrated in Fig. 2. Considering the large-scale use of the agents, long-term studies are still needed to assess their deactivation, the possibility of regeneration, and the economic feasibility, considering other options such as subsequent hydrogenation treatment. According to L´ opez et al. [149], regeneration is possible, at least for ZSM-5 and hopefully for other agents. 3.2.3. Post-treatment The subsequent (off-line) dechlorination of the pyrolysis liquid is less investigated in the literature than the on-line dechlorination, which is reflected in Table 4. The available literature can be divided into processes in inert and hydrogen atmospheres. Although hydrotreatment is theoretically feasible for removing heteroatoms and improving liquid stability [150,151], studies dealing with the removal of heteroatoms from actual pyrolysis liquid from plastics are scarce. There has been some research on noble metal catalysts, of which Pd appears to be the most efficient [152,153]. However, these studies have been carried out on model compounds and require H 2 as a sweeping gas to clean the surface from formed metal chlorides, which cause poisoning of the active sites [154]. Although various metal supports can partially help in this regard [155], there is also the problem of coke formation [156] and reduced DE-Cl in more complex feedstocks. The latter was observed by Murena and Gioia [157] in the N-containing feedstock. The authors attribute this fact to the preferential adsorption of basic nitrogen compounds, which limits hydrodechlorination or the possible formation of salt deposits due to the reaction of HCl with basic nitrogen compounds [158]. In their study, Karayıldırım et al. [159] removed part of the Cl from a PVC/LDPE/HVGO mixture by stepwise pyrolysis at 350 ◦C and the partially dechlorinated product with 700 ppm was subjected to hydrocracking over commercial sulfided DHC −8. They report no chlorine compounds in the liquid based on GC-AED analysis. However, they did not use the same analysis as after the dechlorination step, so the total Cl content in this case is unknown. Regarding the actual hydrogenation of the pyrolysis liquid, Miller et al. [160] reported high DE-Cl (84–97%) at 450 ◦C (200 psig H 2 ) in the presence of 1 wt% HZSM-5 and 2–3 wt% Na 2 CO 3 , which is another example of acid-alkali dechlorination synergy. In addition, Akimoto et al. [161] applied supercritical water treatment to the liquid product from the pyrolysis of MPW and reported 100% dechlorination after 15 min at 375 ◦C in 0.1 M NaOH aqueous solution. However, the authors did not report the oxygen content after the treatment, which is likely to increase as discussed above. Several authors report an alternative approach that does not require hydrogen or solvents. It is a re-pyrolysis or reheating of the liquid product in an inert atmosphere using similar agents and dechlorination mechanisms as the on-line treatment described above. Unfortunately, the authors often did not give specific values for Cl content or yield after the reaction, which makes the results and Table 4 unnecessarily difficult to interpret. Nevertheless, the rate of dechlorination is evident, and the authors report a high level of DE-Cl (99%) in an inert atmosphere with iron oxide‑carbon composite (Fe 3 O 4 ) [111,163] and FeCl 2 /SiO 2 [162]. The authors also compared the DE-Cl of ZnO, MgO, red mud, and prepared catalysts, showing better DE-Cl and stability of the latter [164]. The above studies were carried out in a continuous flow reactor, which also allowed analysis of the agent performance over time, where the iron-based catalyst maintained high DE-Cl throughout the 16–25 h period [162–164]. High DE-Cl has also been reported after treatment of MPW/HVGO liquid mixture in a two-step system with SA-1 at 430 ◦C and Fe-C(Fe 3 O 4 ) or red mud at 350 ◦C [129]. However, the results do not show the benefits of HVGO addition. Slightly lower DE-Cl (69%) was found in a study by LopezUrionabarrenechea et al. [166], using red mud in a closed batch reactor. The authors mention the physical and chemical fixation of chlorine. However, the decrease in DE-Cl may be due to reactions of fixed chlorine, as Lingaiah et al. [164] reported the continuous formation of HCl by interactions of chlorinated hydrocarbons with FeCl 2 . A recent report by Ye et al. [168] suggests that the stability of FeCl 2 may be reduced due to the higher pressure (50–100 bar) used by the Spanish team. Re-entry of already reacted Cl can probably be avoided in flow reactors where reaction conditions can be more precisely adjusted. The batch setting, but with a separate agent bed, was used by Wang et al. for the post-treatment of MPW-derived pyrolysis oil [167]. However, zeolite 4A, MDC-7, and Ni-based catalysts did not show higher efficiencies for Cl removal. To conclude, post-treatment methods can further reduce the Cl content in the liquid. Nevertheless, the potential loss of liquid product and the overall economics of such a process due to reheating or the use of hydrogen must be considered. 3.3. Bromine Used for its flame retardant properties, Br is found in many waste streams depending on the applications where this property is required, such as WEEE, ASR, textile and household waste, construction plastic Fig. 2. Proposed mechanism of dechlorination in β-zeolite/hydrotalcite agent mixture [12]. J. Snow et al. Fuel Processing Technology 254 (2024) 108031 9 Table 4 Overview of post-treatment dechlorination methods sorted by dechlorination efficiency. Agent Liquid origin Conditions 1) Yield difference 2) Cl content (ppm) DE-Cl 5) ref. Before 3) After 4) SA-1 +Fe-C(Fe 3 O 4 ) MPW,HVGO* 350 ◦C, N 2 -5% 728 0 100% [129] Water, NaOH (0.1 M) MPW 375 ◦C, 22 MPa 0% 62 0 100% [161] FeCl 2 /SiO 2 3P*,PVC 350 ◦C, He m.d.* 1600 6) 21 6) 99% [162] Fe-C (Fe 3 O 4 ) 3P*,PVC 350 ◦C, He m.d.* n.r.* n.r.* 99% 6) [163] Fe-C (Fe 3 O 4 ) 3P*,PVC 350 ◦C, He m.d.* n.r.* n.r.* 99% 6) [111] red mud MPW,HVGO* 350 ◦C, N 2 -8% 728 14 98% [129] α -Fe 2 O 3 3P*,PVC 350 ◦C, He m.d.* n.r.* n.r.* 97% 6) [163] ZSM-5 +Na 2 CO 3 MPW 450 ◦C, 200psig H 2 m.d.* 50–70 2–8 84–97% [160] α -Fe 2 O 3 3P*,PVC 350 ◦C, He m.d.* n.r.* n.r.* 96% 6) [111] γ-Fe 2 O 3 MPW 350 ◦C, He m.d.* 570 6) 31 6) 95% [164] Fe/HY zeolite MPW 450 ◦C, N 2 +4% 900 60 93% [165] ZnO MPW 350 ◦C, He m.d.* 570 6) 48 6) 92% [164] Fe-C (Fe 3 O 4 ) MPW 350 ◦C, He m.d.* 570 6) 70 6) 88% [164] red mud MPW 350 ◦C, He m.d.* 570 6) 84 6) 85% [164] MgO MPW 350 ◦C, He m.d.* 570 6) 139 6) 76% [164] red mud 3P*,PET, PVC 325 ◦C, 100 bar −25% 178,000 56,000 69% [166] HY zeolite MPW 450 ◦C, N 2 +2% 900 320 64% [165] zeolite 4A MPW 380 ◦C, N 2 −5% 1600 1000 38% [167] MDC-7 MPW 380 ◦C, N 2 −14% 1600 1200 25% [167] Ni-based catalyst MPW 380 ◦C, N 2 −5% 1600 1200 25% [167] DHC-8 PVC,LDPE, HVGO* 450 ◦C, 6.5 MPa H 2 +3%(120 min) 700 n.r.* m.d.* [159] 1) atmospheric pressure unless stated otherwise 2) yield difference calculated as the difference between the input liquid and the liquid obtained after treatment 3) Cl content before treatment 4) Cl content after treatment 5) taken from reference or calculated by comparing the Cl content in the liquid product from 3) and 4) 6) based on non-numerical data extracted using the open-source software WebPlotDigitizer [35] *n.r. – not reported; m.d. – missing data for calculation; 3P - PE, PP, and PS mixture, HVGO - heavy vacuum gas oil Table 5 Overview of pre-treatment methods sorted by debromination efficiency. Method Solvent Agents Feedstock Particle size T;p t (h) DE-Br ref. Washing H 2 O Na 2 CO 3 PCB smelting flue dust powder 30 ◦C 2 97% 1) [178] H 2 O – PCB smelting flue dust powder 30 ◦C 2 64% [178] Solvothermal toluene – ABS <1 mm 60 ◦C 2 100% [196] MeOH NaOH PCB <0.5 mm 250 ◦C; 6.1 MPa 1 100% [192] H 2 O Ammonia PCB 3 ×1 cm 300 ◦C 1 100% [193] H 2 O, methanol – PCB 2 ×1 cm 350 ◦C; 13 MPa 1 100% [183] H 2 O, CH 3 COOH – PCB 3 ×1 cm 220 ◦C; 3 MPa 1 99.9% [195] CO 2 – display housing 841 μ m 398 ◦C; 0.6 MPa 1.3 99.5% [63] H 2 O – PP, HIPS 3 ×2 mm 280 ◦C; 7 MPa 1 98.8% [186] H 2 O – PCB 10 ×1.5 cm 400 ◦C; 30 MPa 2 98% [181] H 2 O KOH HIPS 3 mm 280 ◦C; 7 MPa 1 97% 2) [185] H 2 O NaOH HIPS n.r.* 450 ◦C; 31 MPa n.r.* 97% [184] H 2 O NaOH PCB <0.3 mm 250 ◦C 3 95% [180] H 2 O NaOH TV housing 6 ×3 ×3 mm 300 ◦C 3 94% [204] H 2 O – computer housing <1 mm 350 ◦C; 12 MPa 1 93% [182] H 2 O NaOH, Na 2 S HIPS n.r.* 220 ◦C; 2 91% [194] H 2 O – PCB fine dust 275 ◦C; 6 MPa 3 64% [205] H 2 O NaOH PCB <0.07 mm n.r.* 2 54% 3) [179] ethylene glycol NaOH HIPS <250 μ m 190 ◦C 24 42% [188] MeOH – ABS <841 μ m 90 ◦C 2 37% [206] IPA – WEEE <0.9 mm 132 ◦C 6 36% [197] Mechanochemical ethylene glycol NaOH ASR <2.8 mm 190 ◦C 1 100% [187] ethylene glycol NaOH HIPS <250 μ m 190 ◦C 24 98% [188] – Fe PP, DecaBDE few mm – 8 90% [190] – CaO PP, DecaBDE few mm – 8 80% [190] – CaO PCB <0.6 mm – 12 65% [191] – Si-Al PP,PE 75 μ m – 8 55% [189] – Si-Al ABS 75 μ m – 8 46% 2) [189] Microwave IPA/hexane – ABS,HIPS,butadiene,PC,PP n.r.* 100 ◦C 0.67 88% [202] H 2 O, TEG KOH HIPS n.r.* 250 ◦C 0.5 85% [201] 1) 2nd leaching after leaching in water with 64% DE-Br 2) based on non-numerical data extracted using the open-source software WebPlotDigitizer [35] 3) DE after subsequent pyrolysis *n.r. - not reported. J. Snow et al. Fuel Processing Technology 254 (2024) 108031 16 removal (0.6 ppm) [323]. The Ni-Mo catalyst also showed better DE-S than Co-Mo [324] and Mo [325] catalysts. Using H 2 and tetralin as hydrogen donors, Djandja et al. [326] reported the lowest N and S content with commercial Pt/C catalyst compared to activated carbon, Ir/C, Ru/C, and Pd/C. The highest efficiencies were 86% and 99.6%, respectively. Other works utilizing feedstocks other than liquid from plastics have been found in relation to desulfurization [327–329], where the authors emphasize the different susceptibility of sulfur towards its removal. For more stable sulfur compounds, oxidative desulfurization has been proposed, consisting of oxygenation in the first step, followed by removal of the oxidized compounds with a selective extraction solvent [330]. Information on deoxygenation was found in the study by Vasile et al. [276], who hydrotreated pyrolysis liquid from WEEE over DHC-8 and MoNi-AC catalysts, showing 68% DE-O in the latter case. However, MoNi-AC did not show any improvement in DE-O compared to the run without catalyst. It is also worth noting that the oxygen content in the treated liquid was only 5.5 ppm. Since the process of removing heteroatoms coincides with the hydrogenation of olefinic and aromatic hydrocarbons, there is constant competition for catalytic sites. Of the heteroatoms mentioned, nitrogen is reported to be the least reactive [331]. Therefore, its removal may not be sufficient when treating more complex feeds. Several review articles have been devoted to this topic [332–334]. 4.3.2. Extraction and adsorption Works utilizing either extraction of heteroatoms into liquid solvents or sorption onto solid agents have been found for common petroleum feedstocks. For more information, the reader is referred to the reviews by Haruna et al. [316] for sulfur and Prado et al. [332] for nitrogen. As an example, Feng [335] reports DE-N of polar nitrogen compounds between 95 and 99% in gasoline using HCl or alumina at room temperature. However, in this case, certain less polar or non-polar substances are likely to remain in the oil. Another option is to use ionic liquids, as shown in the case of the removal of various thiophenes [336]. It is noteworthy that ionic liquids tend to be highly selective, so not all substances are removed equally. Nevertheless, ongoing advances in ionic liquid technology may provide additional opportunities. Compared to the other methods mentioned, these can be seen as separation processes and thus do not deal specifically with heteroatoms but rather with whole hydrocarbons, which can lead to significant yield losses if heteroatoms are abundant. Their advantage of low temperature requirements may be outweighed by the cost of the solvent/sorbent and its regeneration or lack thereof [332]. 4.4. Summary Biogenic elements can be removed to some extent in all processing steps prior to petrochemical utilization. The least demanding is washing, which can remove inorganics, paper or food residues in particular. In addition, less stable heteroatomic compounds can be partially removed by stepwise pyrolysis or pyrolysis with agents, although reported efficiencies are usually low. The highest reported efficiencies for O, N, S, and P with agents were 25%, 43%, 75%, and 93%, respectively. Studies using H 2 O have not focused on heteroatoms or their removal, but rather on composition from the point of hydrocarbons. The process requires high temperature and pressure to reach the properties of supercritical water. In terms of heteroatoms, there is a possibility of oxygen introduction into the product, which increases with the condition severity. Heteroatom removal can also be improved in an alkaline solution. The use of hydrogen is an option both during and after pyrolysis, but most of the literature is devoted to the latter, using know-how and catalysts from conventional hydrotreatment in refineries. Among the catalysts used, the highest removals have generally been achieved with Ni-Mo-based catalysts, with DE-N and DE-S reaching close to 100%. 5. Metals In the pyrolysis feedstock, metals are usually present as plastic additives, mostly as pigments or fillers [40]. According to Eriksen et al. [337], the impurities adhering to the plastics represent a smaller fraction of the metals that can be removed by washing. However, this may differ among feedstocks, and for some metals, the removal efficiency may be significant [51]. During pyrolysis, metals can leave the reactor as volatile organometallic compounds or solid particles, increasing the metal content in the liquid, as shown by Kusenberg et al. [15] who analyzed products from three different pyrolysis feedstocks. Although most of the metals remained in the solid residue, the total metal content was still 563–790 ppm, with Al, Ca, Cu, Fe, K, Mg, Na, Si, and Ti being the most abundant. After calculating the percentage of metal recovery in the liquid, the value can be almost negligible compared to the solid residue [123,338]. However, even trace metals can be detrimental to downstream processes. The possible concentrations of metals in the pyrolysis liquid products have been reported elsewhere [18,339]. For plastic feedstocks, the following metal removal options have been investigated in the literature: washing [51,340], acidic/alkaline leaching [341], solvothermal processes [184,194] followed by separation [342], on-line use of sorbents [124,240,343], separation/filtration of fine particles [123,338], and distillation/fractionation [292]. The concept of demetallization by hydrotreatment was found only for more conventional feedstocks [344,345]. 5.1. Pre-treatment For metals, washing also seems to be an essential part of the pretreatment. The results show that the content of all monitored elements (Ca, Al, Na, K, Mg, Fe, Si), except Ti, was significantly reduced after the washing sequence [51]. As shown in Fig. 5, washing decreased the total content of analyzed metalloids and alkali metals by up to 78% and 59%, respectively. High DE-M of washing has also been reported for residues from the plastics sorting line, where the highest DE-M among the analyzed metals (77% and 30%) was observed for Zn and Ni, respectively [340]. Although the authors did not analyze the metal content in the resulting liquid pyrolysis products, it can be assumed that the removal would also be reflected there to some extent. Plastics from WEEE are a special case for demetallization. During their treatment, it is worth paying attention, for example, to the Sb content, as Sb 2 O 3 is a common synergistic additive in flame-retardant plastics. Alternative approaches to its removal include dissolution in a process such as CreaSolv® and subsequent separation by decanter centrifuge, where the authors report 79–97% DE-Sb [342]. High efficiency (98%) has also been achieved with SCW treatment [184], and Zhan et al. [194] reported similar efficiency at lower temperature (220 ◦C) using Na 2 S/NaOH solution. Metals such as Cu can be removed by HCl leaching [341]. However, Si was not affected by either alkaline or acid leaching. 5.2. On-line treatment Regarding the demetallization possibilities during pyrolysis, the notion of using several agents, including red mud, CaO, Ni/MCM-41, and natural zeolite, has been found for Sb removal during pyrolysis of WEEE [240,343], reaching 63–81% DE-Sb. Among the possible removal mechanisms, the authors mention physical adsorption of SbBr 3 and chemisorption, forming new compounds such as NiSb. However, from the work of Miskolczi et al. [124], who experimented with different types of zeolites, MoO 3 , Ni-Mo catalyst, and Al(OH) 3 during the pyrolysis of MSW and MPW, it can be concluded that the removal efficiency of all the metals analyzed (Ca, Zn, Fe, Cr, Sb, Pb) was insignificant with all the agents tested. As noted by several authors [123] [338], the DE-M can also be improved by filtration of fine particles before a condensation stage. J. Snow et al. Fuel Processing Technology 254 (2024) 108031 17 5.3. Post-treatment Considering the maximum boiling point of standard steam crackers, which is around 200 ◦C, a significant amount of metals can remain in the distillation residue [292,346]. However, removal may still not be sufficient and subsequent treatment is likely to be required. In this regard, works have been found using solid purification agents [347,348], supercritical water [317,349], solvent extraction [350], and H 2 treatment [344] to treat liquids of different origins. Nevertheless, it can be assumed that the results will also apply to some extent to the liquid from the pyrolysis of plastics. A more detailed review of the treatment of crude oil fractions has been given elsewhere [351,352]. In the case of using purification agents in an inert atmosphere, L´ azaro et al. [348] report relatively high removal of V, Ni, Pb, Cd, and Cu over different agents. The highest DE-M for each element is given in Table 8. The DE-M for individual metals also varied between agents, and while activated carbon was most efficient for Pb removal (89%), it was least efficient for Ni removal (75%). The Cr content was not reduced compared to the pure thermal treatment without agents. The DE-M variation can be attributed to differences in specific surface area and acid-base properties of the agents. The SCW treatment of gasoil spiked with nickel and vanadium compounds proved unsuccessful without the addition of H 2 and a sulfided CoMo/γ-Al 2 O 3 catalyst, which probably acted as a sorbent via the formation of corresponding sulfides. The authors further calculated that hydrolysis of metal-containing compounds is unlikely to occur, and hydrogen is therefore required [317]. In contrast, Tang et al. [349] achieved DE-M >90% for the same metals at 390 ◦C during SCW treatment of heavy oils under N 2 without a catalyst. The authors also compared SCW treatment with pyrolysis under N 2 , demonstrating the advantages of SCW treatment. An alternative requiring lower temperatures may be solvent extraction, as shown for bis(2-ethylhexyl) phosphoric acid, which achieved 90% and 79% DE-M for V and Ni, respectively [353]. The H 2 treatment commonly used in the industry today is the last resort for metal removal. Since the presence of metals causes catalyst deactivation, it may be more cost effective to remove as many metals as possible prior to hydrogenation, as catalysts are irreversibly deactivated and would need to be replaced frequently. Therefore, guard reactors with appropriate catalysts are employed to protect the hydrogenation efficiency of downstream catalysts [354]. In an example given by Kressmann et al. [355], the Hyvahl swing reactor system would need to be changed once a year utilizing the common crude oil containing 250–400 ppm of metals. Commonly used hydrodemetallization catalysts Fig. 5. Metal content in plastic feedstock before and after washing sequences. Data taken from [51]. Table 8 Demetallization efficiency of agents during post-treatment. Agent Liquid origin Conditions Metal Content DE-M 3) ref. Before 1) After 2) CoMo/γ-Al 2 O 3 gasoil 400 ◦C; 25 MPa; H 2 +H 2 O V 12 ppm 0.5 ppm 96% [317] Ni 14 ppm 0.5 ppm 96% limestone mineral waste oil 600 ◦C; 0.1 MPa; N 2 Cu 355 μ g 26 μ g 93% [348] Ni 1752 μ g 197 μ g 89% activated carbon Cd 12 μ g 1 μ g 92% Pb 663 μ g 72 μ g 89% activated pyr. char V 43 μ g 7 μ g 84% activated natural aluminosilicate heavy vacuum fraction 407 ◦C; 8 MPa; H 2 V 128 ppm 15 ppm 88% [344] Ni 40 ppm 9 ppm 78% 1) metal content before treatment 2) metal content after treatment 3) calculated by comparing the metal content from 1) and 2) J. Snow et al. Fuel Processing Technology 254 (2024) 108031 18 are based on alumina or silica promoted by Mo, Co, or Ni oxides [344]. The presence of halogens on the catalyst surface can also improve DE-M [345]. In contrast, Gryglewicz et al. [344] reported higher Ni a V removal with natural aluminosilicate compared to CoMo/γ-Al 2 O 3 , which, on the other hand, showed much higher DE-S. The authors correlated the higher DE-M with the capillary pore structure of natural aluminosilicate, which lacks the hydrogenation activity needed for sulfur removal. 5.4. Summary In general, there is a lack of research on metal removal from plastic pyrolysis products, where other metals not commonly found in crude oil may be present in high concentrations. Furthermore, in most cases a complex screening of the metal content is missing. Based on the most relevant literature, some metals can be removed by washing prior to pyrolysis, while the use of purification agents during pyrolysis has not shown significant DE-M. Since metals tend to agglomerate in the char and solid particles, filtration of gases and liquid products may also prove effective. In terms of post-treatment, some metals can be removed by fractionation/distillation and agglomeration of metals in heavier fractions. Results with common petroleum-derived feedstocks indicate that significant DE-M can be achieved using agents in an inert or H 2 atmosphere. However, a complete screening of relevant metals was not performed and should be pursued in further work dealing with the purification of pyrolysis products. 6. A reflection on economics All of the reviewed options for heteroatom removal have additional capital (CaPex) and operating (OpEx) costs that have to be considered when choosing between treatments and technologies. The key technology costs associated with the reviewed processes are shown in Fig. 6. Note that in the on-line section, the figure only considers conventional pyrolysis and not solvent-based techniques at elevated temperatures, as these have proven effective mostly for single type plastic wastes or model mixtures as described in Table 2, Table 5, and Section 4.2.3. There is also a commercial HYDROPRS™ process by Mura Technology for MPW feedstocks, which uses supercritical steam during pyrolysis that runs at elevated pressure. However, the robustness of the process with respect to heteroatoms and their removal has not been officially disclosed [356]. Therefore, water-utilizing processes and the associated water treatment costs are only considered for preand post-treatment. There are many possible scenarios leading to a liquid product of the quality required for petrochemical processing. The deployment of individual processes is determined by a number of factors such as feedstock composition, the efficiency of individual purification processes, or their CapEx and OpEx costs. Without knowledge of the economics of individual processes, it is difficult to argue for an optimal solution. However, it can be argued to some extent that the fewer steps needed to achieve the required purification, the lower the cost. This argument is supported by the fact that in all the cases studied, perhaps with the exception of washing, there is a need to maintain (on-line) or raise the temperature during the preand post-treatment in order to facilitate the heteroatom reaction with chemical agents or to dilute/decompose polymers in solvents. However, fewer processing steps are not always more effective, as illustrated by the use of a purification agent during pyrolysis. If the presence of heteroatoms in the stream is too high, the agent could be rapidly inactivated, which can significantly increase the cost of this method, depending on the price of the agent or the possibility of its regeneration. In this very possible scenario, other methods may be more cost effective or even necessary to achieve the required product quality. In terms of the cost of each option, the on-line processes have the lowest additional capex and opex costs, as the bed with the purification agent can theoretically be implemented in existing conventional pyrolysis technologies. However, considering the amount of impurities in the feedstock such as MPW, it seems feasible to subject the plastics to the washing sequence and only then proceed with the pyrolysis. Considering that, according to the reviewed literature, a part of the heteroatoms can be removed during the stepwise pyrolysis, this technique seems to be economically feasible, as it only requires to withdraw the products to a separate condensation stage and to prolong the time at lower temperature (300–360 ◦C). The products of the 2nd step can then be subjected to the bed with a purifying agent, effectively utilizing the temperature of the product stream, which is necessary to achieve higher purification efficiencies, as shown for example in Table 3. To remove the most resistant residual heteroatoms, the partially purified product can be Fig. 6. Types of heteroatom removal and corresponding price increasing technology elements. J. Snow et al. Fuel Processing Technology 254 (2024) 108031 19 subjected to hydrotreatment. Here, the removal of most of the heteroatoms in the previous steps ensures minimal hydrogen consumption. In the end, the main driver for the implementation of any of the discussed purification processes will be the quantity and price of usable feedstock on the market, together with legislative measures such as mandatory recycled content, which is already implemented in several US states [357–359] and discussed at the EU level [5,360]. As the availability of less heterogeneous waste plastics decreases, there will be an increased incentive to use more complex feedstocks and to employ heteroatom removal techniques. In fact, this situation has already occurred with the increasing share of chemical recycling in the plastics waste market that is currently underway. 7. Conclusion Several important conclusions can be drawn from this review. Also, regarding the removal of individual heteroatom groups Fig. 7, summarizes the possibility of their removal in each processing step. General remarks •The removal of heteroatoms is crucial for the further utilization of the pyrolysis liquid within current technologies. •There is a lack of studies on comprehensive removal and monitoring of heteroatoms in the products. Although the monitored parameter may improve after certain measures have been taken, there is a possibility of introducing other heteroatoms from additional reactions with catalysts, sorbents, or other treatments. •Compared to the number of studies dealing with purification, many studies deal with the possibility of altering the hydrocarbon composition using different catalysts without considering heteroatoms and catalyst inactivation. •The cost of heteroatom removal potentially increases with additional processes as it requires further heating, use of catalysts, introduction of H 2 , etc. •There is currently no method other than sorting for the removal of fluorine due to the high energy required to break the C-F bond. •There is generally a lack of studies dealing with oxygen removal. •Solvothermal treatment at various stages in the process chain can lead to the formation of oxygenated hydrocarbons and subsequent elevated oxygen levels in the liquid product. •Studies usually do not address the possibility of agent reuse or irreversible inactivation. Another direction of research may be the identification of a possible competition between heteroatoms over the agents. Pre-treatment •Feedstock washing can remove a significant fraction of heteroatoms, including chlorine, biogenic elements, and metals. It can be combined with separation by flotation, which can also remove some of the PVC and PET. On-line treatment •Br and Cl can be effectively removed by combining stepwise pyrolysis with chemical agents placed ex-situ. The removal of other heteroatoms has been scarcely investigated in the literature. •Research into the use of supercritical water has focused mainly on single-type plastics. Post-treatment •High dehalogenation efficiencies have been reported using agents under an inert atmosphere. •Removal of biogenic elements and metals has been pursued mostly by hydrotreatment with conventional catalysts or their modified versions. •During hydrotreatment, the overall purification efficiency is likely to decrease if more heteroatoms can occupy active catalyst sites, typically metals and nitrogen. Ultimately, the product must also meet the hydrocarbon composition requirements for use in the petrochemical industry, which was beyond the scope of this work. The level of heteroatom removal was sufficient in many of the cases studied, given the likely dilution of the pyrolysis liquid product in conventional feedstocks. However, future efforts to utilize liquid products for petrochemicals should report on all relevant heteroatoms whose levels may be altered by the treatment. CRediT authorship contribution statement Jan Snow: Writing – original draft, Writing – review & editing. Pavel Kur´ aˇ n: Supervision, Validation. Aleˇ s Kaˇ sp´ arek: Data curation, Writing – review & editing. Pavel Leˇ stinský: Writing – original draft, Writing – review & editing. Robert Suchopa: Conceptualization, Supervision. Declaration of competing interest The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. Data availability The authors are unable or have chosen not to specify which data has been used. Acknowledgment This publication was created with the state support of the Technology Agency of the Czech Republic under the National Centres of Competence Progamme (registration no. TN02000051/006). Fig. 7. Removal of individual heteroatoms as reported in the literature. * supercritical water treatment. J. Snow et al. Fuel Processing Technology 254 (2024) 108031 20 References [1] Plastics Europe, Plastics – the Facts 2022 (2022) Available from: https://plasticse urope.org/knowledge-hub/plastics-the-facts-2022/. [2] EU Directive 2018/850 of the European Parliament and of the Council of 30 May 2018 amending Directive 1999/31/EC on the landfill of waste. https://eur-lex.eu ropa.eu/legal-content/EN/TXT/?uri=celex%3A32018L0850, 2018. [3] EU Communication from the commission - The European Green Deal. https:// eur-lex.europa.eu/legal-content/EN/TXT/?uri=COM:2019:640:FIN, 2019. [4] Z. Manˇ zuch, et al., Chemical Recycling of Polymeric Materials from Waste in the Circular Economy-Final Report Prepared for the European Chemicals Agency (Issue August). https://echa.europa.eu/documents/10162/1459379/chem_re cycling_final_report_en.pdf/887c4182-8327-e197-0bc4-17a5d608de6e, 2021. [5] PlasticsEurope PlasticsEurope’s position on Recycled Content for plastics packaging under the review of the Directive 94/62/EC on Packaging and Packaging Waste (PPWD). https://plasticseurope.org/wp-content/uploads/2021 /11/Plastics_Europe_position_on_PPWD_MRC_final-5.pdf, 2021. [6] Z.O.G. Schyns, M.P. Shaver, Mechanical recycling of packaging plastics: a review, Macromol. Rapid Commun. 42 (3) (2021) 2000415, https://doi.org/10.1002/ marc.202000415. [7] J.M. Garcia, M.L. Robertson, The future of plastics recycling, Science 358 (6365) (2017) 870–872, https://doi.org/10.1126/science.aaq0324. [8] D.J. Gardner, Y. Han, L. Wang, Wood–plastic composite technology, Current Forestry Rep. 1 (3) (2015) 139–150, https://doi.org/10.1007/s40725-015-00166. [9] A. Demirbas, Pyrolysis of municipal plastic wastes for recovery of gasoline-range hydrocarbons, J. Anal. Appl. Pyrolysis 72 (1) (2004) 97–102, https://doi.org/ 10.1016/j.jaap.2004.03.001. [10] A. Buekens, Introduction to Feedstock Recycling of Plastics, in Feedstock Recycling and Pyrolysis of Waste Plastics, 2006, pp. 1–41, https://doi.org/ 10.1002/0470021543.ch1. [11] V. Sinha, M.R. Patel, J.V. Patel, Pet waste management by chemical recycling: a review, J. Polym. Environ. 18 (1) (2010) 8–25, https://doi.org/10.1007/s10924008-0106-7. [12] J. Snow, et al., Dechlorination during pyrolysis of plastics: effect of municipal plastic waste composition, Fuel Process. Technol. 248 (2023) 107823, https:// doi.org/10.1016/j.fuproc.2023.107823. [13] M. Fukushima, et al., Study on dechlorination technology for municipal waste plastics containing polyvinyl chloride and polyethylene terephthalate, J. Mater. Cycles Waste Manage. 12 (2) (2010) 108–122, https://doi.org/10.1007/s10163010-0279-8. [14] M.J. Antal, M. Grønli, The art, science, and technology of charcoal production, Ind. Eng. Chem. Res. 42 (8) (2003) 1619–1640, https://doi.org/10.1021/ ie0207919. [15] M. Kusenberg, et al., A comprehensive experimental investigation of plastic waste pyrolysis oil quality and its dependence on the plastic waste composition, Fuel Process. Technol. 227 (2022) 107090, https://doi.org/10.1016/j. fuproc.2021.107090. [16] N. Alanazi, F. Adam, M. Nagu, Organochloride contamination in a refinery naphtha hydrotreater unit, Mater. Perform. 56 (10) (2017) 1–5. [17] J.A. Moulijn, A.E. van Diepen, F. Kapteijn, Catalyst deactivation: is it predictable?: what to do? Appl. Catal. A Gen. 212 (1) (2001) 3–16, https://doi. org/10.1016/S0926-860X(00)00842-5. [18] M. Kusenberg, et al., Opportunities and challenges for the application of postconsumer plastic waste pyrolysis oils as steam cracker feedstocks: to decontaminate or not to decontaminate? Waste Manag. 138 (2022) 83–115, https://doi.org/10.1016/j.wasman.2021.11.009. [19] R. Palos, et al., Waste refinery: the valorization of waste plastics and end-of-life tires in refinery units. A review, Energy & Fuels 35 (5) (2021) 3529–3557, https://doi.org/10.1021/acs.energyfuels.0c03918. [20] M. Kusenberg, et al., Towards high-quality petrochemical feedstocks from mixed plastic packaging waste via advanced recycling: the past, present and future, Fuel Process. Technol. 238 (2022) 107474, https://doi.org/10.1016/j. fuproc.2022.107474. [21] S.D. Anuar Sharuddin, et al., A review on pyrolysis of plastic wastes, Energy Convers. Manag. 115 (2016) 308–326, https://doi.org/10.1016/j. enconman.2016.02.037. [22] S.M. Al-Salem, et al., A review on thermal and catalytic pyrolysis of plastic solid waste (PSW), J. Environ. Manag. 197 (2017) 177–198, https://doi.org/10.1016/ j.jenvman.2017.03.084. [23] R. Miandad, et al., Catalytic pyrolysis of plastic waste: a review, Process Saf. Environ. Prot. 102 (2016) 822–838, https://doi.org/10.1016/j. psep.2016.06.022. [24] Y. Peng, et al., A review on catalytic pyrolysis of plastic wastes to high-value products, Energy Convers. Manag. 254 (2022) 115243, https://doi.org/10.1016/ j.enconman.2022.115243. [25] M.S. Qureshi, et al., Pyrolysis of plastic waste: Opportunities and challenges, J. Anal. Appl. Pyrolysis 152 (2020) 104804, https://doi.org/10.1016/j. jaap.2020.104804. [26] B. Kunwar, et al., Plastics to fuel: a review, Renew. Sust. Energ. Rev. 54 (2016) 421–428, https://doi.org/10.1016/j.rser.2015.10.015. [27] T. Xayachak, et al., Pyrolysis for plastic waste management: an engineering perspective, J. Environ. Chem. Eng. 10 (6) (2022) 108865, https://doi.org/ 10.1016/j.jece.2022.108865. [28] K. Murthy, R.J. Shetty, K. Shiva, Plastic waste conversion to fuel: a review on pyrolysis process and influence of operating parameters, in: Energy Sources, Part A: Recovery, Utilization, and Environmental Effects, 2020, pp. 1–21, https://doi. org/10.1080/15567036.2020.1818892. [29] X. Yang, et al., Pyrolysis and dehalogenation of plastics from waste electrical and electronic equipment (WEEE): a review, Waste Manag. 33 (2) (2013) 462–473, https://doi.org/10.1016/j.wasman.2012.07.025. [30] Y. Shen, et al., Waste-to-energy: Dehalogenation of plastic-containing wastes, Waste Manag. 49 (2016) 287–303, https://doi.org/10.1016/j. wasman.2015.12.024. [31] C. Ma, et al., Chemical recycling of brominated flame retarded plastics from ewaste for clean fuels production: a review, Renew. Sust. Energ. Rev. 61 (2016) 433–450, https://doi.org/10.1016/j.rser.2016.04.020. [32] S. Liu, et al., Plastic waste to fuels by hydrocracking at mild conditions, Sci. Adv. 7 (17) (2021) eabf8283, https://doi.org/10.1126/sciadv.abf8283. [33] M.S. Abbas-Abadi, et al., Challenges and opportunities of light olefin production via thermal and catalytic pyrolysis of end-of-life polyolefins: Towards full recyclability, Prog. Energy Combust. Sci. 96 (2023) 101046, https://doi.org/ 10.1016/j.pecs.2022.101046. [34] M. Zeller, et al., Chemical recycling of mixed plastic wastes by pyrolysis – pilot scale investigations, Chem. Ing. Tech. 93 (11) (2021) 1763–1770, https://doi. org/10.1002/cite.202100102. [35] A. Rohatgi, WebPlotDigitizer. https://automeris.io/WebPlotDigitizer, 2022. [36] Y. Chen, et al., Improving bromine fixation in co-pyrolysis of non-metallic fractions of waste printed circuit boards with Bayer red mud, Sci. Total Environ. 639 (2018) 1553–1559, https://doi.org/10.1016/j.scitotenv.2018.05.269. [37] R.-X. Yang, et al., Thermochemical conversion of plastic waste into fuels, chemicals, and value-added materials: a critical review and outlooks, ChemSusChem 15 (11) (2022) e202200171, https://doi.org/10.1002/ cssc.202200171. [38] M. Brebu, et al., The role of temperature program and catalytic system on the quality of acrylonitrile-butadiene-styrene degradation oil, J. Anal. Appl. Pyrolysis 63 (1) (2002) 43–57, https://doi.org/10.1016/S0165-2370(01)00140-1. [39] S. Shaw, Halogenated flame retardants: do the fire safety benefits justify the risks? Rev. Environ. Health 25 (4) (2010) 261–306, https://doi.org/10.1515/ REVEH.2010.25.4.261. [40] E. Hansen, et al., Hazardous substances in plastic materials, Danish Technol. Inst. (2013). Available from: https://www.byggemiljo.no/wp-content/uploads/2014/ 10/72_ta3017.pdf. [41] J. Yu, et al., Thermal degradation of PVC: a review, Waste Manag. 48 (2016) 300–314, https://doi.org/10.1016/j.wasman.2015.11.041. [42] R. Miranda, et al., Vacuum pyrolysis of commingled plastics containing PVC II. Product analysis, Polym. Degrad. Stab. 73 (1) (2001) 47–67, https://doi.org/ 10.1016/S0141-3910(01)00066-0. [43] F. Barontini, V. Cozzani, Formation of hydrogen bromide and organobrominated compounds in the thermal degradation of electronic boards, J. Anal. Appl. Pyrolysis 77 (1) (2006) 41–55, https://doi.org/10.1016/j.jaap.2006.01.003. [44] E.E. Lewis, M.A. Naylor, Pyrolysis of Polytetrafluoroethylene, J. Am. Chem. Soc. 69 (8) (1947) 1968–1970, https://doi.org/10.1021/ja01200a039. [45] G. Puts, P. Crouse, B. Ameduri, Thermal Degradation and Pyrolysis of Polytetrafluoroethylene, in Handbook of Fluoropolymer Science and Technology, 2014, pp. 81–104, https://doi.org/10.1002/9781118850220.ch5. [46] E. Meissner, A. Wr´ oblewska, E. Milchert, Technological parameters of pyrolysis of waste polytetrafluoroethylene, Polym. Degrad. Stab. 83 (1) (2004) 163–172, https://doi.org/10.1016/S0141-3910(03)00259-3. [47] C.M. Simon, W. Kaminsky, Chemical recycling of polytetrafluoroethylene by pyrolysis, Polym. Degrad. Stab. 62 (1) (1998) 1–7, https://doi.org/10.1016/ S0141-3910(97)00097-9. [48] W. Kaminsky, Chemical recycling of plastics by fluidized bed pyrolysis, Fuel Communications 8 (2021) 100023, https://doi.org/10.1016/j. jfueco.2021.100023. [49] A. Bezuidenhoudt, P.W. Sonnendecker, P.L. Crouse, Temperature and pressure effects on the product distribution of PTFE pyrolysis by means of qualitative, inline FTIR analysis, Polym. Degrad. Stab. 142 (2017) 79–88, https://doi.org/ 10.1016/j.polymdegradstab. 2017.05.025. [50] W. Ma, S. Rotter, Overview on the Chlorine Origin of MSW and Cl-Originated Corrosion During MSW & RDF Combustion Process, in: 2008 2nd International Conference on Bioinformatics and Biomedical Engineering, 2008, pp. 4255–4258, https://doi.org/10.1109/ICBBE.2008.564. [51] H.C. Genuino, et al., Pyrolysis of mixed plastic waste (DKR-350): effect of washing pre-treatment and fate of chlorine, Fuel Process. Technol. 233 (2022) 107304, https://doi.org/10.1016/j.fuproc.2022.107304. [52] B. Indrawan, P. Prawisudha, K. Yoshikawa, Combustion characteristics of chlorine-free solid fuel produced from municipal solid waste by hydrothermal processing, Energies 5 (11) (2012) 4446–4461, https://doi.org/10.3390/ en5114446. [53] N. Borsodi, et al., Hydrocarbons obtained by pyrolysis of contaminated waste plastics, in: 45th International Petroleum Conference, Bratislava, Slovak Republic, 2011. [54] F.-R. Xiu, et al., Co-treatment of PVC and used LCD panels in low-temperature subcritical water: Enhanced dechlorination and mechanism, Process Saf. Environ. Prot. 151 (2021) 10–19, https://doi.org/10.1016/j.psep.2021.05.001. [55] S.M. Shin, T. Yoshioka, A. Okuwaki, Dehydrochlorination behavior of rigid PVC pellet in NaOH solutions at elevated temperature, Polym. Degrad. Stab. 61 (2) (1998) 349–353, https://doi.org/10.1016/S0141-3910(97)00221-8. [56] T. Yoshioka, et al., Dechlorination behaviour of flexible poly(vinyl chloride) in NaOH/EG solution, Polym. Degrad. Stab. 93 (10) (2008) 1822–1825, https://doi. org/10.1016/j.polymdegradstab. 2008.07.009. J. Snow et al. Fuel Processing Technology 254 (2024) 108031 21 [57] J. Poerschmann, et al., Hydrothermal carbonization of poly(vinyl chloride), Chemosphere 119 (2015) 682–689, https://doi.org/10.1016/j. chemosphere.2014.07.058. [58] F.-R. Xiu, Y. Qi, F.-S. Zhang, Co-treatment of waste printed circuit boards and polyvinyl chloride by subcritical water oxidation: Removal of brominated flame retardants and recovery of Cu and Pb, Chem. Eng. J. 237 (2014) 242–249, https://doi.org/10.1016/j.cej.2013.10.026. [59] Y. Wang, et al., Low chlorine oil production through fast pyrolysis of mixed plastics combined with hydrothermal dechlorination pretreatment, Process. Saf. Environ. Prot. 149 (2021) 105–114, https://doi.org/10.1016/j. psep.2020.10.023. [60] T. Yoshinaga, et al., Alkaline dechlorination of poly(vinyl chloride) in organic solvents under mild conditions, Polym. Degrad. Stab. 86 (3) (2004) 541–547, https://doi.org/10.1016/j.polymdegradstab.2004.06.008. [61] B. Lv, et al., Dechlorination and oxidation for waste poly(vinylidene chloride) by hydrothermal catalytic oxidation on Pd/AC catalyst, Polym. Degrad. Stab. 94 (7) (2009) 1047–1052, https://doi.org/10.1016/j.polymdegradstab. 2009.04.004. [62] G. Gandon-Ros, et al., Dechlorination of polyvinyl chloride electric wires by hydrothermal treatment using K2CO3 in subcritical water, Waste Manag. 102 (2020) 204–211, https://doi.org/10.1016/j.wasman.2019.10.050. [63] C.-C. Zhang, F.-S. Zhang, Enhanced dehalogenation and coupled recovery of complex electronic display housing plastics by sub/supercritical CO2, J. Hazard. Mater. 382 (2020) 121140, https://doi.org/10.1016/j.jhazmat.2019.121140. [64] A.M. Hapipi, et al., Dechlorination of polyvinyl chloride under superheated steam with catalysts and adsorbents, Energy Fuel 32 (7) (2018) 7792–7799, https://doi. org/10.1021/acs.energyfuels.8b00838. [65] K. Hashimoto, et al., Hydrothermal dechlorination of PVC in the presence of ammonia, J. Mater. Sci. 43 (7) (2008) 2457–2462, https://doi.org/10.1007/ s10853-007-2015-x. [66] A. Soler, J.A. Conesa, N. Ortu˜ no, Application of Subcritical Water to Dechlorinate polyvinyl Chloride Electric Wires, Energies 11 (10) (2018) 2612, https://doi.org/ 10.3390/en11102612. [67] H. Nishibata, M.A. Uddin, Y. Kato, Simultaneous degradation and dechlorination of poly (vinyl chloride) by a combination of superheated steam and CaO catalyst/ adsorbent, Polym. Degrad. Stab. 179 (2020) 109225, https://doi.org/10.1016/j. polymdegradstab.2020.109225. [68] Y. Qi, et al., A novel treatment method of PVC-medical waste by near-critical methanol: Dechlorination and additives recovery, Waste Manag. 80 (2018) 1–9, https://doi.org/10.1016/j.wasman.2018.08.052. [69] T. Yoshioka, et al., Dechlorination of poly(vinylidene chloride) in NaOH/ethylene glycol as a function of NaOH concentration, temperature, and solvent, Polym. Degrad. Stab. 93 (10) (2008) 1979–1984, https://doi.org/10.1016/j. polymdegradstab. 2008.06.008. [70] K. Kusakabe, et al., Dechlorination of polyvinyl Chloride via Solvothermal Treatment with Glycerol, Processes 10 (10) (2022) 2047, https://doi.org/ 10.3390/pr10102047. [71] D. Ma, et al., Insight into chlorine evolution during hydrothermal carbonization of medical waste model, J. Hazard. Mater. 380 (2019) 120847, https://doi.org/ 10.1016/j.jhazmat.2019.120847. [72] P. Zhao, et al., Dechlorination of PVC wastes by hydrothermal treatment using alkaline additives, Environ. Technol. 39 (8) (2018) 977–985, https://doi.org/ 10.1080/09593330.2017.1317841. [73] P. Prawisudha, et al., Dechlorination behavior of mixed plastic waste by employing hydrothermal process and limestone additive, J. Environ. Sci. Eng. 5 (4) (2011). [74] D. Ma, et al., Dechlorination of polyvinyl chloride by hydrothermal treatment with cupric ion, Process Saf. Environ. Prot. 146 (2021) 108–117, https://doi.org/ 10.1016/j.psep.2020.08.040. [75] S. Saeki, et al., Effect of additives on dechlorination of PVC by mechanochemical treatment, J. Mater. Cycles Waste Manag. 3 (1) (2001) 20–23, https://doi.org/ 10.1007/s10163-000-0035-6. [76] W. Tongamp, et al., Simultaneous treatment of PVC and oyster-shell wastes by mechanochemical means, Waste Manag. 28 (3) (2008) 484–488, https://doi.org/ 10.1016/j.wasman.2007.01.022. [77] J. Lu, et al., Practical dechlorination of polyvinyl chloride wastes in NaOH/ ethylene glycol using an up-scale ball mill reactor and validation by discrete element method simulations, Waste Manag. 99 (2019) 31–41, https://doi.org/ 10.1016/j.wasman.2019.08.034. [78] T. Kameda, et al., Ball Mill-Assisted Dechlorination of flexible and rigid Poly (vinyl chloride) in NaOH/EG solution, Ind. Eng. Chem. Res. 47 (22) (2008) 8619–8624, https://doi.org/10.1021/ie8006819. [79] M. Bal´ aˇ z, et al., Mechanochemical dechlorination of PVC by utilizing eggshell waste, Acta Phys. Pol. A 126 (4) (2014) 884–887, https://doi.org/10.12693/ APhysPolA.126.884. [80] T. Inoue, et al., Dechlorination of polyvinyl chloride by its grinding with KOH and NaOH, Adv. Powder Technol. 16 (1) (2005) 27–34, https://doi.org/10.1163/ 1568552053166638. [81] J. Lu, et al., Practical dehalogenation of automobile shredder residue in NaOH/ ethylene glycol with an up-scale ball mill reactor, J. Material Cycles Waste Manag. 22 (5) (2020) 1620–1629, https://doi.org/10.1007/s10163-020-01052-z. [82] M. Bal´ aˇ z, et al., Simultaneous valorization of polyvinyl chloride and eggshell wastes by a semi-industrial mechanochemical approach, Environ. Res. 170 (2019) 332–336, https://doi.org/10.1016/j.envres.2018.12.005. [83] F. Osada, J. Yana, Deplasticization and dechlorination of flexible polyvinyl chloride in NaOH solution by microwave heating, J. Material Cycles Waste Manag. 12 (3) (2010) 245–253, https://doi.org/10.1007/s10163-010-0294-9. [84] F. Osada, T. Yoshioka, Dechlorination of polyvinyl chloride in NaOH/ethylene glycol solution by microwave heating, J. Mater. Cycles Waste Manag. 11 (1) (2009) 19–22, https://doi.org/10.1007/s10163-008-0213-5. [85] S. Moriwaki, et al., Dehydrochlorination of poly(vinyl chloride) by microwave irradiation, Appl. Therm. Eng. 26 (7) (2006) 745–750, https://doi.org/10.1016/j. applthermaleng.2005.09.001. [86] Z. Liu, et al., Dechlorination of organochloride waste mixture by microwave irradiation before forming solid recovered fuel, Waste Manag. 62 (2017) 118–124, https://doi.org/10.1016/j.wasman.2016.11.022. [87] M. Hori, et al., Selective Dechlorination of polyvinyl Chloride by Microwave Irradiation, Asian Pacific Confed. Chem. Eng. Congress Prog. Abstracts 2004 (2004) 857, https://doi.org/10.11491/apcche.2004.0.857.0. [88] T. Kameda, et al., Chemical modification of poly(vinyl chloride) by nucleophilic substitution, Polym. Degrad. Stab. 94 (1) (2009) 107–112, https://doi.org/ 10.1016/j.polymdegradstab. 2008.10.006. [89] J. Lu, et al., Study on characterization of pyrolysis and hydrolysis products of poly (vinyl chloride) waste, J. Appl. Polym. Sci. 90 (12) (2003) 3252–3259, https:// doi.org/10.1002/app.12984. [90] Y. Takeshita, et al., Basic study on treatment of waste polyvinyl chloride plastics by hydrothermal decomposition in subcritical and supercritical regions, J. Supercrit. Fluids 31 (2) (2004) 185–193, https://doi.org/10.1016/j. supflu.2003.10.006. [91] T. Moriya, H. Enomoto, Characteristics of polyethylene cracking in supercritical water compared to thermal cracking, Polym. Degrad. Stab. 65 (3) (1999) 373–386, https://doi.org/10.1016/S0141-3910(99)00026-9. [92] Y. Sato, et al., Decomposition of Ployvinylchloride using Supercritical Water, Jpn. J. Appl. Phys. 37 (11R) (1998) 6270, https://doi.org/10.1143/JJAP.37.6270. [93] A. Kub´ atov´ a, A.J.M. Lagadec, S.B. Hawthorne, Dechlorination of Lindane, Dieldrin, Tetrachloroethane, Trichloroethene, and PVC in Subcritical Water, Environ. Sci. Technol. 36 (6) (2002) 1337–1343, https://doi.org/10.1021/ es011186k. [94] Y. Nagai, et al., Direct observation of polyvinylchloride degradation in water at temperatures up to 500◦C and at pressures up to 700 MPa, J. Appl. Polym. Sci. 106 (2) (2007) 1075–1086, https://doi.org/10.1002/app.26790. [95] J. Lu, S. Ma, J. Gao, Study on the Pressurized Hydrolysis Dechlorination of PVC, Energy Fuel 16 (5) (2002) 1251–1255, https://doi.org/10.1021/ef020048t. [96] P. Zhao, et al., The study of nickel effect on the hydrothermal dechlorination of PVC, J. Clean. Prod. 152 (2017) 38–46, https://doi.org/10.1016/j. jclepro.2017.03.101. [97] P. Prawisudha, T. Namioka, K. Yoshikawa, Coal alternative fuel production from municipal solid wastes employing hydrothermal treatment, Appl. Energy 90 (1) (2012) 298–304, https://doi.org/10.1016/j.apenergy.2011.03.021. [98] P. Bal´ aˇ z, et al., Hallmarks of mechanochemistry: from nanoparticles to technology, Chem. Soc. Rev. 42 (18) (2013) 7571–7637, https://doi.org/ 10.1039/C3CS35468G. [99] G. Cagnetta, et al., Mechanochemical destruction of halogenated organic pollutants: a critical review, J. Hazard. Mater. 313 (2016) 85–102, https://doi. org/10.1016/j.jhazmat.2016.03.076. [100] X. Guo, et al., A review of mechanochemistry applications in waste management, Waste Manag. 30 (1) (2010) 4–10, https://doi.org/10.1016/j. wasman.2009.08.017. [101] J. Zhou, T.-G. Hsu, J. Wang, Mechanochemical Degradation and Recycling of Synthetic Polymers, Angew. Chem. Int. Ed. n/a(n/a) (2023) e202300768, https:// doi.org/10.1002/anie.202300768. [102] X. Hu, et al., Microwave-assisted pyrolysis of waste plastics for their resource reuse: a technical review, Carbon Resour. Conver. 6 (3) (2023) 215–228, https:// doi.org/10.1016/j.crcon.2023.03.002. [103] H. Bockhorn, et al., Environmental engineering: Stepwise pyrolysis of plastic waste, Chem. Eng. Sci. 54 (15) (1999) 3043–3051, https://doi.org/10.1016/ S0009-2509(98)00385-6. [104] R. Miranda, et al., Vacuum pyrolysis of PVC I. Kinetic study, Polym. Degrad. Stab. 64 (1) (1999) 127–144, https://doi.org/10.1016/S0141-3910(98)00186-4. [105] S. Xin, et al., Thermochemical emission and transformation of chlorinated paraffins in inert and oxidizing atmospheres, Chemosphere 185 (2017) 899–906, https://doi.org/10.1016/j.chemosphere.2017.07.019. [106] Y. Masuda, et al., Pyrolysis study of poly(vinyl chloride)–metal oxide mixtures: Quantitative product analysis and the chlorine fixing ability of metal oxides, J. Anal. Appl. Pyrolysis 77 (2) (2006) 159–168, https://doi.org/10.1016/j. jaap.2006.03.001. [107] J. Yanik, et al., The catalytic effect of Red Mud on the degradation of poly (vinyl chloride) containing polymer mixture into fuel oil, Polym. Degrad. Stab. 73 (2) (2001) 335–346, https://doi.org/10.1016/S0141-3910(01)00095-7. [108] Q. Zhou, et al., Catalytic degradation and dechlorination of PVC-containing mixed plastics via Al–Mg composite oxide catalysts, Fuel 83 (13) (2004) 1727–1732, https://doi.org/10.1016/j.fuel.2004.02.015. [109] Y. Tang, et al., Hydrogen-Rich and Clean fuel Gas Production from Co-pyrolysis of Biomass and Plastic Blends with CaO Additive, ACS Omega 7 (41) (2022) 36468–36478, https://doi.org/10.1021/acsomega.2c04279. [110] N. Sophonrat, et al., Ex situ catalytic pyrolysis of a mixture of polyvinyl chloride and cellulose using calcium oxide for hcl adsorption and catalytic reforming of the pyrolysis products, Ind. Eng. Chem. Res. 58 (31) (2019) 13960–13970, https:// doi.org/10.1021/acs.iecr.9b02299. [111] Y. Sakata, et al., Development of a catalytic dehalogenation (Cl, Br) process for municipal waste plastic-derived oil, J. Material Cycles Waste Manag. 5 (2) (2003) 113–124, https://doi.org/10.1007/s10163-003-0092-8. J. Snow et al. Fuel Processing Technology 254 (2024) 108031 22 [112] T. Bhaskar, et al., Prevention of chlorinated hydrocarbons formation during pyrolysis of PVC or PVDC mixed plastics, Green Chem. 8 (8) (2006) 697–700, https://doi.org/10.1039/B603037H. [113] H. Bockhorn, A. Hornung, U. Hornung, Stepwise pyrolysis for raw material recovery from plastic waste, J. Anal. Appl. Pyrolysis 46 (1) (1998) 1–13, https:// doi.org/10.1016/S0165-2370(98)00066-7. [114] J. Hub´ aˇ cek, et al., Dechlorination during pyrolysis of plastics: the potential of stepwise pyrolysis in combination with metal sorbents, Fuel Process. Technol. 231 (2022) 107226, https://doi.org/10.1016/j.fuproc.2022.107226. [115] J. Lei, et al., Investigation on thermal dechlorination and catalytic pyrolysis in a continuous process for liquid fuel recovery from mixed plastic wastes, J. Material Cycles Waste Manag. 20 (1) (2018) 137–146, https://doi.org/10.1007/s10163016-0555-3. [116] K.-B. Park, et al., Separate two-step and continuous two-stage pyrolysis of a waste plastic mixture to produce a chlorine-depleted oil, Energy 244 (2022) 122583, https://doi.org/10.1016/j.energy.2021.122583. [117] B.B. Perez-Martinez, et al., Dechlorination of plastic-rich fractions rejected from waste electric and electronic equipment recycling plants by means of stepwise pyrolysis for valorization, WIT Trans. Ecol. Environ. 254 (2021) 81–90 (1784664499). [118] A. L´ opez, et al., Dechlorination of fuels in pyrolysis of PVC containing plastic wastes, Fuel Process. Technol. 92 (2) (2011) 253–260, https://doi.org/10.1016/j. fuproc.2010.05.008. [119] K.-B. Park, et al., Production of clean oil with low levels of chlorine and olefins in a continuous two-stage pyrolysis of a mixture of waste low-density polyethylene and polyvinyl chloride, Energy 157 (2018) 402–411, https://doi.org/10.1016/j. energy.2018.05.182. [120] T. Bhaskar, et al., Pyrolysis studies of PP/PE/PS/PVC/HIPS-Br plastics mixed with PET and dehalogenation (Br, Cl) of the liquid products, J. Anal. Appl. Pyrolysis 72 (1) (2004) 27–33, https://doi.org/10.1016/j.jaap.2004.01.005. [121] M. Brebu, et al., Removal of nitrogen, bromine, and chlorine from PP/PE/PS/ PVC/ABS–Br pyrolysis liquid products using Feand Ca-based catalysts, Polym. Degrad. Stab. 87 (2) (2005) 225–230, https://doi.org/10.1016/j. polymdegradstab. 2004.08.008. [122] B. Fekhar, L. Gombor, N. Miskolczi, Pyrolysis of chlorine contaminated municipal plastic waste: In-situ upgrading of pyrolysis oils by Ni/ZSM-5, Ni/SAPO-11, red mud and ca(OH)2 containing catalysts, J. Energy Inst. 92 (5) (2019) 1270–1283, https://doi.org/10.1016/j.joei.2018.10.007. [123] M.-H. Cho, S.-H. Jung, J.-S. Kim, Pyrolysis of mixed plastic wastes for the recovery of benzene, toluene, and xylene (BTX) aromatics in a fluidized bed and chlorine removal by applying various additives, Energy Fuel 24 (2) (2010) 1389–1395, https://doi.org/10.1021/ef901127v. [124] N. Miskolczi, F. Ates¸, N. Borsodi, Comparison of real waste (MSW and MPW) pyrolysis in batch reactor over different catalysts. Part II: Contaminants, char and pyrolysis oil properties, Bioresour. Technol. 144 (2013) 370–379, https://doi. org/10.1016/j.biortech.2013.06.109. [125] N. Miskolczi, F. Ates¸, Thermo-catalytic co-pyrolysis of recovered heavy oil and municipal plastic wastes, J. Anal. Appl. Pyrolysis 117 (2016) 273–281, https:// doi.org/10.1016/j.jaap.2015.11.005. [126] A. L´ opez, et al., Catalytic pyrolysis of plastic wastes with two different types of catalysts: ZSM-5 zeolite and Red Mud, Appl. Catal. B Environ. 104 (3) (2011) 211–219, https://doi.org/10.1016/j.apcatb.2011.03.030. [127] A. Lopez-Urionabarrenechea, et al., Catalytic stepwise pyrolysis of packaging plastic waste, J. Anal. Appl. Pyrolysis 96 (2012) 54–62, https://doi.org/10.1016/ j.jaap.2012.03.004. [128] Y.-S. Jeong, K.-B. Park, J.-S. Kim, Kinetics and characteristics of activator-assisted pyrolysis of municipal waste plastic and chlorine removal using hot filter filled with absorbents, Energy 238 (2022) 121814, https://doi.org/10.1016/j. energy.2021.121814. [129] J. Yanik, M.A. Uddin, Y. Sakata, The effect of red mud on the liquefaction of waste plastics in heavy vacuum gas oil, Energy Fuel 15 (1) (2001) 163–169, https://doi. org/10.1021/ef0001080. [130] T. Karayıldırım, et al., Degradation of PVC Containing Mixtures in the Presence of HCl Fixators, J. Polym. Environ. 13 (4) (2005) 365–374, https://doi.org/ 10.1007/s10924-005-5531-2. [131] R. Miranda, et al., Vacuum pyrolysis of PVCII: product analysis, Polym. Degrad. Stab. 66 (1) (1999) 107–125, https://doi.org/10.1016/S0141-3910(99)00060-9. [132] Z. Yuan, et al., Synergistic effect and chlorine-release behaviors during copyrolysis of LLDPE, PP, and PVC, ACS Omega 5 (20) (2020) 11291–11298, https://doi.org/10.1021/acsomega.9b04116. [133] I. Coralli, et al., Secondary reactions in the analysis of microplastics by analytical pyrolysis, J. Anal. Appl. Pyrolysis 161 (2022) 105377, https://doi.org/10.1016/j. jaap.2021.105377. [134] Z. Cz´ eg´ eny, E. Jakab, M. Blazs´ o, Thermal Decomposition of Polymer Mixtures Containing Poly(vinyl chloride), Macromol. Mater. Eng. 287 (4) (2002) 277–284, https://doi.org/10.1002/1439-2054(20020401)287:4<277::AIDMAME277>3.0.CO;2-#. [135] Z. Cz´ eg´ eny, et al., Thermal decomposition of polymer mixtures of PVC, PET and ABS containing brominated flame retardant: Formation of chlorinated and brominated organic compounds, J. Anal. Appl. Pyrolysis 96 (2012) 69–77, https://doi.org/10.1016/j.jaap.2012.03.006. [136] T. Bhaskar, et al., Comparison of thermal degradation products from real municipal waste plastic and model mixed plastics, J. Anal. Appl. Pyrolysis 70 (2) (2003) 579–587, https://doi.org/10.1016/S0165-2370(03)00027-5. [137] K. Kulesza, K. German, Chlorinated pyrolysis products of co-pyrolysis of poly (vinyl chloride) and poly(ethylene terephthalate), J. Anal. Appl. Pyrolysis 67 (1) (2003) 123–134, https://doi.org/10.1016/S0165-2370(02)00057-8. [138] ˙ I. Çit, et al., Comparative pyrolysis of polyolefins (PP and LDPE) and PET, Polym. Bull. 64 (8) (2010) 817–834, https://doi.org/10.1007/s00289-009-0225-x. [139] N. Sophonrat, et al., Stepwise pyrolysis of mixed plastics and paper for separation of oxygenated and hydrocarbon condensates, Appl. Energy 229 (2018) 314–325, https://doi.org/10.1016/j.apenergy.2018.08.006. [140] J.N. Hahladakis, et al., An overview of chemical additives present in plastics: Migration, release, fate and environmental impact during their use, disposal and recycling, J. Hazard. Mater. 344 (2018) 179–199, https://doi.org/10.1016/j. jhazmat.2017.10.014. [141] X.-G. Zheng, et al., Dehydrochlorination of PVC Materials at High Temperature, Energy Fuel 17 (4) (2003) 896–900, https://doi.org/10.1021/ef020131g. [142] W. Li, et al., Comparative study on pyrolysis behaviors and chlorine release of pure PVC polymer and commercial PVC plastics, Fuel 340 (2023) 127555, https://doi.org/10.1016/j.fuel.2023.127555. [143] G. Yuan, et al., High efficiency chlorine removal from polyvinyl chloride (PVC) pyrolysis with a gas–liquid fluidized bed reactor, Waste Manag. 34 (6) (2014) 1045–1050, https://doi.org/10.1016/j.wasman.2013.08.021. [144] A. L´ opez, et al., Pyrolysis of municipal plastic wastes: Influence of raw material composition, Waste Manag. 30 (4) (2010) 620–627, https://doi.org/10.1016/j. wasman.2009.10.014. [145] S. Kumagai, et al., Impact of Common Plastics on Cellulose Pyrolysis, Energy Fuel 33 (7) (2019) 6837–6841, https://doi.org/10.1021/acs.energyfuels.9b01376. [146] H. Kuramochi, et al., HCl emission during co-pyrolysis of demolition wood with a small amount of PVC film and the effect of wood constituents on HCl emission reduction, Fuel 87 (13) (2008) 3155–3157, https://doi.org/10.1016/j. fuel.2008.03.021. [147] G. Jiang, et al., Understanding the Dechlorination of Chlorinated Hydrocarbons in the Pyrolysis of mixed Plastics, ACS Sustain. Chem. Eng. 9 (4) (2021) 1576–1589, https://doi.org/10.1021/acssuschemeng.0c06461. [148] A. Fontana, et al., Municipal waste pyrolysis (2): chlorine capture by addition of calcium and sodium-based sorbents, Erd¨ ol Erdgas Kohle (2001) 117. https://www.osti.gov/etdeweb/biblio/20195001#fullrecord. [149] A. L´ opez, et al., Deactivation and regeneration of ZSM-5 zeolite in catalytic pyrolysis of plastic wastes, Waste Manag. 31 (8) (2011) 1852–1858, https://doi. org/10.1016/j.wasman.2011.04.004. [150] J. Scheirs, Overview of commercial pyrolysis processes for waste plastics, in: Feedstock recycling and pyrolysis of waste plastics: converting waste plastics into diesel and other fuels, 2006, pp. 381–433, https://doi.org/10.1002/0470021543. ch15. [151] J. Scherzer, A.J. Gruia, Hydrocracking Science and Technology, CRC Press, 1996 (9781482233889). [152] B. Coq, G. Ferrat, F. Figueras, Conversion of chlorobenzene over palladium and rhodium catalysts of widely varying dispersion, J. Catal. 101 (2) (1986) 434–445, https://doi.org/10.1016/0021-9517(86)90271-X. [153] F.J. Urbano, J.M. Marinas, Hydrogenolysis of organohalogen compounds over palladium supported catalysts, J. Mol. Catal. A Chem. 173 (1) (2001) 329–345, https://doi.org/10.1016/S1381-1169(01)00157-1. [154] M.A. Aramendia, et al., Influence of the reaction conditions and catalytic properties on the liquid-phase hydrodechlorination of chlorobenzene over palladium-supported catalysts: activity and deactivation, J. Catal. 187 (2) (1999) 392–399, https://doi.org/10.1006/jcat.1999.2632. [155] B. Coq, et al., Conversion under Hydrogen of Dichlorodifluoromethane over Supported Palladium Catalysts, J. Catal. 141 (1) (1993) 21–33, https://doi.org/ 10.1006/jcat.1993.1115. [156] E.J. Creyghton, et al., Vapour-phase hydrodehalogenation of chlorobenzene over platinum/H-BEA zeolite, Appl. Catal. A Gen. 128 (2) (1995) 275–288, https:// doi.org/10.1016/0926-860X(95)00080-1. [157] F. Gioia, F. Murena, Simultaneous catalytic hydroprocessing of chlorine-, nitrogen-, and Sulphur-containing aromatic compounds, J. Hazard. Mater. 57 (1) (1998) 177–192, https://doi.org/10.1016/S0304-3894(97)00082-4. [158] F. Murena, F. Gioia, Catalytic hydroprocessing of chlorobenzene–pyridine mixtures, J. Hazard. Mater. 60 (3) (1998) 271–285, https://doi.org/10.1016/ S0304-3894(98)00151-4. [159] T. Karayıldırım, et al., Conversion of plastics/HVGO mixtures to fuels by two-step processing, Fuel Process. Technol. 73 (1) (2001) 23–35, https://doi.org/10.1016/ S0378-3820(01)00192-8. [160] S. Miller, N. Shah, G. Huffman, Production of premium oil products from waste plastic by pyrolysis and hydroprocessing, in: Feedstock Recycling and Pyrolysis of Waste Plastics: Converting Waste Plastics into Diesel and Other Fuels, 2006, pp. 345–361, https://doi.org/10.1002/0470021543.ch13. [161] M. Akimoto, et al., Hydrothermal Dechlorination and Denitrogenation of Municipal-Waste-Plastics-Derived fuel Oil under Suband Supercritical Conditions, Ind. Eng. Chem. Res. 41 (22) (2002) 5393–5400, https://doi.org/ 10.1021/ie020338x. [162] N. Lingaiah, et al., Catalytic dehydrochlorination of chloro-organic compounds from PVC containing waste plastics derived fuel oil over FeCl/SiO catalyst, Green Chem. 3 (2) (2001) 74–75, https://doi.org/10.1039/B009471O. [163] N. Lingaiah, et al., Catalytic dechlorination of chloroorganic compounds from PVC-containing mixed plastic-derived oil, Appl. Catal. A Gen. 207 (1) (2001) 79–84, https://doi.org/10.1016/S0926-860X(00)00656-6. [164] N. Lingaiah, et al., Removal of organic chlorine compounds by catalytic dehydrochlorination for the refinement of municipal waste plastic derived oil, J. Snow et al. Fuel Processing Technology 254 (2024) 108031 23 Fuel 80 (13) (2001) 1901–1905, https://doi.org/10.1016/S0016-2361(01) 00046-1. [165] K.-R. Hwang, et al., Catalytic cracking of chlorinated heavy wax from pyrolysis of plastic wastes to low carbon-range fuels: Catalyst effect on properties of liquid products and dechlorination, J. Anal. Appl. Pyrolysis 155 (2021) 105090, https:// doi.org/10.1016/j.jaap.2021.105090. [166] A. Lopez-Urionabarrenechea, et al., Upgrading of chlorinated oils coming from pyrolysis of plastic waste, Fuel Process. Technol. 137 (2015) 229–239, https:// doi.org/10.1016/j.fuproc.2015.04.015. [167] S. Wang, et al., Drop-in fuel production with plastic waste pyrolysis oil over catalytic separation, Fuel 305 (2021) 121440, https://doi.org/10.1016/j. fuel.2021.121440. [168] L. Ye, T. Li, L. Hong, Co-pyrolysis of Fe3O4-poly(vinyl chloride) (PVC) mixtures: Mitigation of chlorine emissions during PVC recycling, Waste Manag. 126 (2021) 832–842, https://doi.org/10.1016/j.wasman.2021.04.021. [169] P. Hennebert, Concentrations of brominated flame retardants in plastics of electrical and electronic equipment, vehicles, construction, textiles and non-food packaging: a review of occurrence and management, Detritus 12 (2020) 34–50, https://doi.org/10.31025/2611-4135/2020.13997. [170] M. Alaee, et al., An overview of commercially used brominated flame retardants, their applications, their use patterns in different countries/regions and possible modes of release, Environ. Int. 29 (6) (2003) 683–689, https://doi.org/10.1016/ S0160-4120(03)00121-1. [171] S.-J. Chen, et al., Brominated flame retardants in children’s toys: concentration, composition, and children’s exposure and risk assessment, Environ. Sci. Technol. 43 (11) (2009) 4200–4206, https://doi.org/10.1021/es9004834. [172] O.T. Fatunsin, et al., Children’s exposure to hazardous brominated flame retardants in plastic toys, Sci. Total Environ. 720 (2020) 137623, https://doi.org/ 10.1016/j.scitotenv.2020.137623. [173] J. Samsonek, F. Puype, Occurrence of brominated flame retardants in black thermo cups and selected kitchen utensils purchased on the European market, Food Additives Contamin. A 30 (11) (2013) 1976–1986, https://doi.org/ 10.1080/19440049.2013.829246. [174] R. Paseiro-Cerrato, et al., Brominated flame retardants (BFRs) in contaminated food contact articles: identification using DART-HRMS and GC-MS, Food Additives Contamin. A 38 (2) (2021) 350–359, https://doi.org/10.1080/ 19440049.2020.1853250. [175] A. Turner, M. Filella, Bromine in plastic consumer products – evidence for the widespread recycling of electronic waste, Sci. Total Environ. 601-602 (2017) 374–379, https://doi.org/10.1016/j.scitotenv.2017.05.173. [176] F. Puype, et al., Evidence of waste electrical and electronic equipment (WEEE) relevant substances in polymeric food-contact articles sold on the European market, Food Additives Contamin. A 32 (3) (2015) 410–426, https://doi.org/ 10.1080/19440049.2015.1009499. [177] M.A.-E. Abdallah, et al., Hexabromocyclododecane in polystyrene packaging: a downside of recycling? Chemosphere 199 (2018) 612–616, https://doi.org/ 10.1016/j.chemosphere.2018.02.084. [178] J. Lyu, et al., Efficient bromine removal and metal recovery from waste printed circuit boards smelting flue dust by a two-stage leaching process, J. Clean. Prod. 322 (2021) 129054, https://doi.org/10.1016/j.jclepro.2021.129054. [179] Y. Shen, Effect of chemical pretreatment on pyrolysis of non-metallic fraction recycled from waste printed circuit boards, Waste Manag. 76 (2018) 537–543, https://doi.org/10.1016/j.wasman.2018.02.036. [180] Y. Chen, et al., New insights into the debromination mechanism of non-metallic fractions of waste printed circuit boards via alkaline-enhanced subcritical water route, Resour. Conserv. Recycl. 165 (2021) 105227, https://doi.org/10.1016/j. resconrec.2020.105227. [181] M. Xing, F.-S. Zhang, Degradation of brominated epoxy resin and metal recovery from waste printed circuit boards through batch sub/supercritical water treatments, Chem. Eng. J. 219 (2013) 131–136, https://doi.org/10.1016/j. cej.2012.12.066. [182] Y. Wang, F.-S. Zhang, Degradation of brominated flame retardant in computer housing plastic by supercritical fluids, J. Hazard. Mater. 205-206 (2012) 156–163, https://doi.org/10.1016/j.jhazmat.2011.12.055. [183] F.-R. Xiu, et al., Application of critical water-alcohol composite medium to treat waste printed circuit boards: Oil phase products characteristic and debromination, J. Hazard. Mater. 344 (2018) 333–342, https://doi.org/10.1016/ j.jhazmat.2017.10.033. [184] J.A. Onwudili, P.T. Williams, Alkaline reforming of brominated fire-retardant plastics: Fate of bromine and antimony, Chemosphere 74 (6) (2009) 787–796, https://doi.org/10.1016/j.chemosphere.2008.10.029. [185] M. Brebu, et al., Alkaline hydrothermal treatment of brominated high impact polystyrene (HIPS-Br) for bromine and bromine-free plastic recovery, Chemosphere 64 (6) (2006) 1021–1025, https://doi.org/10.1016/j. chemosphere.2006.02.036. [186] M.A. Uddin, et al., Debromination of flame retardant high impact polystyrene (HIPS-Br) by hydrothermal treatment and recovery of bromine free plastics, Green Chem. 5 (2) (2003) 260–263, https://doi.org/10.1039/B206704H. [187] T. Kameda, et al., Efficient dehalogenation of automobile shredder residue in NaOH/ethylene glycol using a ball mill, Chemosphere 74 (2) (2009) 287–292, https://doi.org/10.1016/j.chemosphere.2008.09.009. [188] G. Grause, et al., A novel process for the removal of bromine from styrene polymers containing brominated flame retardant, Polym. Degrad. Stab. 112 (2015) 86–93, https://doi.org/10.1016/j.polymdegradstab. 2014.12.017. [189] S. Lu, et al., Mechanochemical dehalogenation of brominated flame retardants and preliminary application for recycling BFR-containing plastic waste, J. Environ. Chem. Eng. 11 (3) (2023) 109916, https://doi.org/10.1016/j. jece.2023.109916. [190] G. Cagnetta, et al., Mechanochemical pre-treatment for viable recycling of plastic waste containing haloorganics, Waste Manag. 75 (2018) 181–186, https://doi. org/10.1016/j.wasman.2018.02.008. [191] R. Wang, et al., Mechanochemical degradation of brominated flame retardants in waste printed circuit boards by Ball Milling, J. Hazard. Mater. 385 (2020) 121509, https://doi.org/10.1016/j.jhazmat.2019.121509. [192] F.-R. Xiu, et al., A novel management strategy for removal and degradation of polybrominated diphenyl ethers (PBDEs) in waste printed circuit boards, Waste Manag. 100 (2019) 191–198, https://doi.org/10.1016/j.wasman.2019.09.022. [193] F.-R. Xiu, et al., A novel treatment of waste printed circuit boards by lowtemperature near-critical aqueous ammonia: Debromination and preparation of nitrogen-containing fine chemicals, Waste Manag. 84 (2019) 355–363, https:// doi.org/10.1016/j.wasman.2018.12.010. [194] L. Zhan, et al., Leaching behavior of Sb and Br from E-waste flame retardant plastics, Chemosphere 245 (2020) 125684, https://doi.org/10.1016/j. chemosphere.2019.125684. [195] M. Xing, et al., Swelling-enhanced catalytic degradation of brominated epoxy resin in waste printed circuit boards by subcritical acetic acid under mild conditions, Waste Manag. 102 (2020) 464–473, https://doi.org/10.1016/j. wasman.2019.11.011. [196] A.M. Altwaiq, M. Wolf, R. van Eldik, Extraction of brominated flame retardants from polymeric waste material using different solvents and supercritical carbon dioxide, Anal. Chim. Acta 491 (1) (2003) 111–123, https://doi.org/10.1016/ S0003-2670(03)00785-2. [197] P. Evangelopoulos, et al., Reduction of brominated flame retardants (BFRs) in plastics from waste electrical and electronic equipment (WEEE) by solvent extraction and the influence on their thermal decomposition, Waste Manag. 94 (2019) 165–171, https://doi.org/10.1016/j.wasman.2018.06.018. [198] F. Vilaplana, et al., Analysis of brominated flame retardants in styrenic polymers: Comparison of the extraction efficiency of ultrasonication, microwave-assisted extraction and pressurised liquid extraction, J. Chromatogr. A 1196-1197 (2008) 139–146, https://doi.org/10.1016/j.chroma.2008.05.001. [199] A. Kousaiti, et al., Assessment of tetrabromobisphenol-A (TBBPA) content in plastic waste recovered from WEEE, J. Hazard. Mater. 390 (2020) 121641, https://doi.org/10.1016/j.jhazmat.2019.121641. [200] F. Vilaplana, A. Ribes-Greus, S. Karlsson, Microwave-assisted extraction for qualitative and quantitative determination of brominated flame retardants in styrenic plastic fractions from waste electrical and electronic equipment (WEEE), Talanta 78 (1) (2009) 33–39, https://doi.org/10.1016/j.talanta.2008.10.038. [201] T. Bhaskar, et al., Enhanced debromination of brominated flame retardant plastics under microwave irradiation, Green Chem. 10 (7) (2008) 739–742, https://doi. org/10.1039/B807370H. [202] M.A. Charitopoulou, L. Papadopoulou, D.S. Achilias, Microwave-assisted extraction as an effective method for the debromination of brominated flame retarded polymeric blends with a composition that simulates the plastic part of waste electric and electronic equipment (WEEE), Sustain. Chem. Pharm. 29 (2022) 100790, https://doi.org/10.1016/j.scp.2022.100790. [203] J. Zhu, et al., An energy-saving and environment-friendly technology for debromination of plastic waste: Novel models of heat transfer and movement behavior of bromine, J. Hazard. Mater. 421 (2022) 126814, https://doi.org/ 10.1016/j.jhazmat.2021.126814. [204] Y. Qin, et al., Debromination process of Br-containing PS of E-wastes and reuse with virgin PS, J. Hazard. Mater. 431 (2022) 128526, https://doi.org/10.1016/j. jhazmat.2022.128526. [205] A. Soler, J.A. Conesa, N. Ortu˜ no, Emissions of brominated compounds and polycyclic aromatic hydrocarbons during pyrolysis of E-waste debrominated in subcritical water, Chemosphere 186 (2017) 167–176, https://doi.org/10.1016/j. chemosphere.2017.07.146. [206] C.-C. Zhang, F.-S. Zhang, Removal of brominated flame retardant from electrical and electronic waste plastic by solvothermal technique, J. Hazard. Mater. 221222 (2012) 193–198, https://doi.org/10.1016/j.jhazmat.2012.04.033. [207] K. Li, Z. Xu, Application of supercritical water to decompose brominated epoxy resin and environmental friendly recovery of metals from waste memory module, Environ. Sci. Technol. 49 (3) (2015) 1761–1767, https://doi.org/10.1021/ es504644b. [208] K. Liu, Z. Zhang, F.-S. Zhang, Direct extraction of palladium and silver from waste printed circuit boards powder by supercritical fluids oxidation-extraction process, J. Hazard. Mater. 318 (2016) 216–223, https://doi.org/10.1016/j. jhazmat.2016.07.005. [209] R. Gao, et al., Catalytic effect and mechanism of coexisting copper on conversion of organics during pyrolysis of waste printed circuit boards, J. Hazard. Mater. 403 (2021) 123465, https://doi.org/10.1016/j.jhazmat.2020.123465. [210] J. Liu, et al., Mechanistic insights into catalysis of in-situ iron on pyrolysis of waste printed circuit boards: Comparative study of kinetics, products, and reaction mechanism, J. Hazard. Mater. 431 (2022) 128612, https://doi.org/ 10.1016/j.jhazmat.2022.128612. [211] J. Liu, et al., Catalytic effect and mechanism of in-situ metals on pyrolysis of FR4 printed circuit boards: Insights from kinetics and products, Chemosphere 280 (2021) 130804, https://doi.org/10.1016/j.chemosphere.2021.130804. [212] M. Brebu, et al., The effect of PVC and/or PET on thermal degradation of polymer mixtures containing brominated ABS, Fuel 83 (14) (2004) 2021–2028, https:// doi.org/10.1016/j.fuel.2004.04.011. [213] T. Bhaskar, et al., Effect of poly(ethylene terephthalate) on the pyrolysis of brominated flame retardant containing high impact polystyrene and catalytic J. Snow et al. Fuel Processing Technology 254 (2024) 108031 24 debromination of the liquid products, J. Anal. Appl. Pyrolysis 71 (2) (2004) 765–777, https://doi.org/10.1016/j.jaap.2003.10.006. [214] J. Simon, et al., Thermal analysis of Sb2O3 /Organohalide-based flame retardants including atomic absorption detection of the evolved species, J. Therm. Anal. 25 (1) (1982) 57–77, https://doi.org/10.1007/BF01913054. [215] T. Bhaskar, et al., Effect of Sb2O3 in brominated heating impact polystyrene (HIPS-Br) on thermal degradation and debromination by iron oxide carbon composite catalyst (Fe-C), Appl. Catal. B Environ. 43 (3) (2003) 229–241, https:// doi.org/10.1016/S0926-3373(02)00306-5. [216] W.J. Hall, P.T. Williams, Pyrolysis of brominated feedstock plastic in a fluidised bed reactor, J. Anal. Appl. Pyrolysis 77 (1) (2006) 75–82, https://doi.org/ 10.1016/j.jaap.2006.01.006. [217] T. Bhaskar, et al., Controlled pyrolysis of polyethylene/polypropylene/ polystyrene mixed plastics with high impact polystyrene containing flame retardant: effect of decabromo diphenylethane (DDE), Polym. Degrad. Stab. 92 (2) (2007) 211–221, https://doi.org/10.1016/j.polymdegradstab. 2006.11.011. [218] E. Jakab, et al., Thermal decomposition of flame-retarded high-impact polystyrene, J. Anal. Appl. Pyrolysis 68-69 (2003) 83–99, https://doi.org/ 10.1016/S0165-2370(03)00075-5. [219] P. Shaohong, et al., Controlled Pyrolysis of Waste TV Housing Plastic Added Brominated Flame Retardants. in 2011 International Conference on Computer Distributed Control and Intelligent Environmental Monitoring, 2011. [220] G. Grause, et al., Effect of heating rate on the pyrolysis of high-impact polystyrene containing brominated flame retardants: fate of brominated flame retardants, J. Material Cycles Waste Manag. 14 (3) (2012) 259–265, https://doi.org/ 10.1007/s10163-012-0067-8. [221] C. Ma, et al., The behavior of heteroatom compounds during the pyrolysis of waste computer casing plastic under various heating conditions, J. Clean. Prod. 219 (2019) 461–470, https://doi.org/10.1016/j.jclepro.2019.02.100. [222] S. Peng, et al., Debromination of flame-retarded TV housing plastic waste, J. Material Cycles Waste Manag. 12 (2) (2010) 103–107, https://doi.org/ 10.1007/s10163-010-0278-9. [223] C. Ma, T. Kamo, Two-stage catalytic pyrolysis and debromination of printed circuit boards: effect of zero-valent Fe and Ni metals, J. Anal. Appl. Pyrolysis 134 (2018) 614–620, https://doi.org/10.1016/j.jaap.2018.08.012. [224] T. Bhaskar, et al., Thermal degradation of ABS-Br mixed with PP and catalytic debromination by iron oxide carbon composite catalyst (Fe–C), Green Chem. 4 (6) (2002) 603–606, https://doi.org/10.1039/B206487A. [225] M.A. Uddin, et al., Dehydrohalogenation during pyrolysis of brominated flame retardant containing high impact polystyrene (HIPS-Br) mixed with polyvinylchloride (PVC), Fuel 81 (14) (2002) 1819–1825, https://doi.org/ 10.1016/S0016-2361(02)00109-6. [226] T. Bhaskar, et al., Novel calcium based sorbent (Ca-C) for the dehalogenation (Br, Cl) process during halogenated mixed plastic (PP/PE/PS/PVC and HIPS-Br) pyrolysis, Green Chem. 4 (4) (2002) 372–375, https://doi.org/10.1039/ B203745A. [227] J. Wang, et al., NaOH–KOH capillary action enhances the pyrolysis and debromination of waste printed circuit boards, ACS Sustain. Chem. Eng. 9 (50) (2021) 17164–17173, https://doi.org/10.1021/acssuschemeng.1c06962. [228] J. Sun, et al., Study of the transference rules for bromine in waste printed circuit boards during microwave-induced pyrolysis, J. Air Waste Manage. Assoc. 61 (5) (2011) 535–542, https://doi.org/10.3155/1047-3289.61.5.535. [229] S. Oleszek, et al., Mitigation of bromine-containing products during pyrolysis of polycarbonate-based tetrabromobisphenol a in the presence of copper(I) oxide, J. Hazard. Mater. 409 (2021) 124972, https://doi.org/10.1016/j. jhazmat.2020.124972. [230] H. Wu, et al., Fuel oil production from two-stage pyrolysis-catalytic reforming of brominated high impact polystyrene using zeolite and iron oxide loaded zeolite catalysts, Open J. Ecol. 5 (04) (2015) 136, https://doi.org/10.4236/ oje.2015.54012. [231] M. Brebu, et al., Thermal degradation of PE and PS mixed with ABS-Br and debromination of pyrolysis oil by Feand Ca-based catalysts, Polym. Degrad. Stab. 84 (3) (2004) 459–467, https://doi.org/10.1016/j.polymdegradstab. 2004.02.003. [232] C. Ma, T. Kamo, Enhanced debromination by Fe particles during the catalytic pyrolysis of non-metallic fractions of printed circuit boards over ZSM-5 and Ni/ SiO2-Al2O3 catalyst, J. Anal. Appl. Pyrolysis 138 (2019) 170–177, https://doi. org/10.1016/j.jaap.2018.12.021. [233] W.J. Hall, P.T. Williams, Removal of organobromine compounds from the pyrolysis oils of flame retarded plastics using zeolite catalysts, J. Anal. Appl. Pyrolysis 81 (2) (2008) 139–147, https://doi.org/10.1016/j.jaap.2007.09.008. [234] C. Ma, T. Kamo, Effect of steam-iron reaction on product characteristics and debromination during pyrolysis of epoxy-printed circuit boards, J. Hazard. Mater. 379 (2019) 120803, https://doi.org/10.1016/j.jhazmat.2019.120803. [235] N.M.M. Mitan, et al., Individual and simultaneous degradation of brominated high impact polystyrene and brominated acrylonitrile-butadiene-styrene and removal of heteroelements (Br, N, and O) from degradation oil by multiphase catalytic systems, J. Material Cycles Waste Manag. 9 (1) (2007) 56–61, https:// doi.org/10.1007/s10163-006-0159-4. [236] S.-H. Jung, S.-J. Kim, J.-S. Kim, Thermal degradation of acrylonitrile–butadiene–styrene (ABS) containing flame retardants using a fluidized bed reactor: the effects of Ca-based additives on halogen removal, Fuel Process. Technol. 96 (2012) 265–270, https://doi.org/10.1016/j. fuproc.2011.12.039. [237] S.-H. Jung, S.-J. Kim, J.-S. Kim, Fast pyrolysis of a waste fraction of high impact polystyrene (HIPS) containing brominated flame retardants in a fluidized bed reactor: the effects of various Ca-based additives (CaO, ca(OH)2 and oyster shells) on the removal of bromine, Fuel 95 (2012) 514–520, https://doi.org/10.1016/j. fuel.2011.11.048. [238] C. Ma, et al., Catalytic pyrolysis of flame retarded high impact polystyrene over various solid acid catalysts, Fuel Process. Technol. 155 (2017) 32–41, https://doi. org/10.1016/j.fuproc.2016.01.018. [239] W.J. Hall, et al., Thermal Processing of toxic Flame-Retarded Polymers using a Waste Fluidized Catalytic Cracker (FCC) Catalyst, Energy Fuel 22 (3) (2008) 1691–1697, https://doi.org/10.1021/ef800043g. [240] C. Ma, et al., Pyrolysis-catalytic upgrading of brominated high impact polystyrene over Fe and Ni modified catalysts: Influence of HZSM-5 and MCM-41 catalysts, Polym. Degrad. Stab. 146 (2017) 1–12, https://doi.org/10.1016/j. polymdegradstab. 2017.09.005. [241] C. Ma, et al., Influence of Zeolites and Mesoporous Catalysts on Catalytic Pyrolysis of Brominated Acrylonitrile–Butadiene–Styrene (Br-ABS), Energy Fuel 30 (6) (2016) 4635–4643, https://doi.org/10.1021/acs.energyfuels.6b00460. [242] C. Ma, et al., Influence of Fe based ZSM-5 catalysts on the vapor intermediates from the pyrolysis of brominated acrylonitrile-butadiene-styrene copolymer (BrABS), Fuel 230 (2018) 390–396, https://doi.org/10.1016/j.fuel.2018.05.077. [243] O. Terakado, R. Ohhashi, M. Hirasawa, Thermal degradation study of tetrabromobisphenol a under the presence metal oxide: Comparison of bromine fixation ability, J. Anal. Appl. Pyrolysis 91 (2) (2011) 303–309, https://doi.org/ 10.1016/j.jaap.2011.03.006. [244] O. Terakado, R. Ohhashi, M. Hirasawa, Bromine fixation by metal oxide in pyrolysis of printed circuit board containing brominated flame retardant, J. Anal. Appl. Pyrolysis 103 (2013) 216–221, https://doi.org/10.1016/j. jaap.2012.10.022. [245] M. Blazs´ o, Z. Cz´ eg´ eny, C. Csoma, Pyrolysis and debromination of flame retarded polymers of electronic scrap studied by analytical pyrolysis, J. Anal. Appl. Pyrolysis 64 (2) (2002) 249–261, https://doi.org/10.1016/S0165-2370(02) 00035-9. [246] S. Kumagai, et al., Thermal decomposition of tetrabromobisphenol-A containing printed circuit boards in the presence of calcium hydroxide, J. Material Cycles Waste Manag. 19 (1) (2017) 282–293, https://doi.org/10.1007/s10163-0150417-4. [247] R. Gao, et al., In-situ debromination mechanism based on self-activation and catalysis of ca(OH)2 during pyrolysis of waste printed circuit boards, J. Hazard. Mater. 392 (2020) 122447, https://doi.org/10.1016/j.jhazmat.2020.122447. [248] Z. Ye, et al., The debrominated and lightweight oil generated from two stage pyrolysis of WPCBs by using compound chemical additives, Process Saf. Environ. Prot. 116 (2018) 654–662, https://doi.org/10.1016/j.psep.2018.03.025. [249] Y. Chen, et al., Enhanced bromine fixation and tar lightweighting in co-pyrolysis of non-metallic fractions of waste printed circuit boards with Bayer red mud, Waste Manag. 162 (2023) 72–82, https://doi.org/10.1016/j. wasman.2023.03.010. [250] T. Zhang, et al., Microwave-assisted catalytic pyrolysis of waste printed circuit boards, and migration and distribution of bromine, J. Hazard. Mater. 402 (2021) 123749, https://doi.org/10.1016/j.jhazmat.2020.123749. [251] Y.-M. Kim, et al., Catalytic co-pyrolysis of epoxy-printed circuit board and plastics over HZSM-5 and HY, J. Clean. Prod. 168 (2017) 366–374, https://doi.org/ 10.1016/j.jclepro.2017.08.224. [252] T.U. Han, Y.-K. Park, Y.-M. Kim, High-quality oil production via the catalytic conversion of printed circuit boards, J. Clean. Prod. 296 (2021) 126614, https:// doi.org/10.1016/j.jclepro.2021.126614. [253] Y.-K. Park, et al., Debrominated high quality oil production by the two-step catalytic pyrolysis of phenolic printed circuit boards (PPCB) using natural clays and HY, J. Hazard. Mater. 367 (2019) 50–58, https://doi.org/10.1016/j. jhazmat.2018.12.040. [254] N. Kathongthung, C. Khaobang, C. Areeprasert, An investigation on debromination performance of sulfur impregnated catalyst for e-waste pyrolysis process, in: 2021 Budapest 19th International Conference on Agricultural, Chemical, Biological & Environmental Sciences (ACBES-21), 2021. http://uruae. org/siteadmin/upload/4833EAP0621104.pdf. [255] L. Ali, et al., Removal of Bromine from the non-metallic fraction in printed circuit board via its Co-pyrolysis with alumina, Waste Manag. 137 (2022) 283–293, https://doi.org/10.1016/j.wasman.2021.11.025. [256] L. Rosi, M. Bartoli, M. Frediani, Microwave assisted pyrolysis of halogenated plastics recovered from waste computers, Waste Manag. 73 (2018) 511–522, https://doi.org/10.1016/j.wasman.2017.04.037. [257] Z. Huang, J. Zhu, J. Ruan, A novel technology of vacuum low-temperature pyrolysis with NVZI for the high-efficiency debromination of resin particles from waste printed circuit boards, Resour. Conserv. Recycl. 188 (2023) 106711, https://doi.org/10.1016/j.resconrec.2022.106711. [258] S.Y. Li, et al., Thermal debromination of waste printed circuit boards by ironbased catalyst, Adv. Mater. Res. 881-883 (2014) 589–593, https://doi.org/ 10.4028/www.scientific.net/AMR.881-883.589. [259] Z. Ye, et al., Improvement of pyrolysis oil obtained from co-pyrolysis of WPCBs and compound additive during two stage pyrolysis, J. Anal. Appl. Pyrolysis 135 (2018) 415–421, https://doi.org/10.1016/j.jaap.2018.06.011. [260] S.-X. Shi, et al., Innovative method for removing bromine in waste printed circuit boards: Ultrafine milling and porous media loaded debromination agent, Adv. Powder Technol. 33 (8) (2022) 103662, https://doi.org/10.1016/j. apt.2022.103662. [261] J. Zhu, et al., Catalytic debromination of waste brominated resin by co-pyrolysis with Pd-containing spent automotive catalysts, Resour. Conserv. Recycl. 188 (2023) 106721, https://doi.org/10.1016/j.resconrec.2022.106721. J. Snow et al. Fuel Processing Technology 254 (2024) 108031 25 [262] M.A. Charitopoulou, et al., Catalytic pyrolysis of polymers with brominated flame-retardants originating in waste electric and electronic equipment (WEEE) using various catalysts, Sustain. Chem. Pharm. 26 (2022) 100612, https://doi. org/10.1016/j.scp.2022.100612. [263] C. Vasile, et al., Feedstock recycling from the printed circuit boards of used computers, Energy Fuel 22 (3) (2008) 1658–1665, https://doi.org/10.1021/ ef700659t. [264] C. Ma, et al., Production of BTX via Catalytic Fast Pyrolysis of Printed Circuit Boards and Waste Tires using Hierarchical ZSM-5 Zeolites and Biochar, ACS Sustain. Chem. Eng. 10 (45) (2022) 14775–14782, https://doi.org/10.1021/ acssuschemeng.2c04031. [265] Y. Wang, et al., The effects of activated Al2O3 on the recycling of light oil from the catalytic pyrolysis of waste printed circuit boards, Process Saf. Environ. Prot. 98 (2015) 276–284, https://doi.org/10.1016/j.psep.2015.07.007. [266] T. Chen, et al., Catalytic performance and debromination of Fe–Ni bimetallic MCM-41 catalyst for the two-stage pyrolysis of waste computer casing plastic, Chemosphere 248 (2020) 125964, https://doi.org/10.1016/j. chemosphere.2020.125964. [267] W. Chen, et al., Co-pyrolysis of waste printed circuit boards with iron compounds for Br-fixing and material recovery, Environ. Sci. Pollut. Res. 28 (45) (2021) 64642–64651, https://doi.org/10.1007/s11356-021-15506-w. [268] L. Ali, et al., Separation of bromine and hydrocarbons from polymeric constituents in e-waste through thermal treatment with calcium hydroxide, Sep. Purif. Technol. 307 (2023) 122836, https://doi.org/10.1016/j. seppur.2022.122836. [269] C. Ma, et al., Effect of polypropylene on the pyrolysis of flame retarded high impact polystyrene, Fuel Process. Technol. 135 (2015) 150–156, https://doi.org/ 10.1016/j.fuproc.2014.12.011. [270] A. Hornung, et al., Polypropylene as a reductive agent for dehalogenation of brominated organic compounds, J. Clean. Prod. 13 (5) (2005) 525–530, https:// doi.org/10.1016/j.jclepro.2003.09.001. [271] M. Altarawneh, et al., Co-pyrolysis of polyethylene with products from thermal decomposition of brominated flame retardants, Chemosphere 254 (2020) 126766, https://doi.org/10.1016/j.chemosphere.2020.126766. [272] M. Brebu, E. Jakab, Y. Sakata, Effect of flame retardants and Sb2O3 synergist on the thermal decomposition of high-impact polystyrene and on its debromination by ammonia treatment, J. Anal. Appl. Pyrolysis 79 (1) (2007) 346–352, https:// doi.org/10.1016/j.jaap.2007.02.003. [273] D. S´ anchez-Rodríguez, C. Ma, T. Kamo, Inhibition effect of amine compounds derived from hardening agents on the extraction of hydrogen bromide by water from the pyrolysis oil of brominated printed circuit boards, J. Clean. Prod. 265 (2020) 121811, https://doi.org/10.1016/j.jclepro.2020.121811. [274] X. Ma, et al., Comparative study on catalytic hydrodehalogenation of halogenated aromatic compounds over Pd/C and Raney Ni catalysts, Sci. Rep. 6 (1) (2016) 25068, https://doi.org/10.1038/srep25068. [275] K.V. Murthy, P.M. Patterson, M.A. Keane, C–X bond reactivity in the catalytic hydrodehalogenation of haloarenes over unsupported and silica supported Ni, J. Mol. Catal. A Chem. 225 (2) (2005) 149–160, https://doi.org/10.1016/j. molcata.2004.08.038. [276] C. Vasile, et al., Feedstock recycling from plastics and thermosets fractions of used computers. II. Pyrolysis oil upgrading, Fuel 86 (4) (2007) 477–485, https://doi. org/10.1016/j.fuel.2006.08.010. [277] A. Hornung, et al., Detoxification of brominated pyrolysis oils, J. Anal. Appl. Pyrolysis 70 (2) (2003) 723–733, https://doi.org/10.1016/S0165-2370(03) 00049-4. [278] A.I. Balabanovich, et al., Pyrolysis study of halogen-containing aromatics reflecting reactions with polypropylene in a posttreatment decontamination process, Environ. Sci. Technol. 39 (14) (2005) 5469–5474, https://doi.org/ 10.1021/es0500106. [279] H. Zhou, Z. Tian, C. Liao, Corrosion behaviour characterisation of 316L stainless steel and Inconel 625 in supercritical water containing hydrochloric acid and high oxygen, Corros. Eng. Sci. Technol. 57 (7) (2022) 640–647, https://doi.org/ 10.1080/1478422X.2022.2112931. [280] P. Kritzer, Corrosion in high-temperature and supercritical water and aqueous solutions: a review, J. Supercrit. Fluids 29 (1) (2004) 1–29, https://doi.org/ 10.1016/S0896-8446(03)00031-7. [281] H.E. Toraman, et al., Detailed compositional characterization of plastic waste pyrolysis oil by comprehensive two-dimensional gas-chromatography coupled to multiple detectors, J. Chromatogr. A 1359 (2014) 237–246, https://doi.org/ 10.1016/j.chroma.2014.07.017. [282] S. Pongstabodee, N. Kunachitpimol, S. Damronglerd, Combination of three-stage sink–float method and selective flotation technique for separation of mixed postconsumer plastic waste, Waste Manag. 28 (3) (2008) 475–483, https://doi.org/ 10.1016/j.wasman.2007.03.005. [283] J. Oenema, et al., Review on the pyrolysis products and thermal decomposition mechanisms of polyurethanes, J. Anal. Appl. Pyrolysis 168 (2022) 105723, https://doi.org/10.1016/j.jaap.2022.105723. [284] M.W. Nolte, B.H. Shanks, A perspective on catalytic strategies for deoxygenation in biomass pyrolysis, Energ. Technol. 5 (1) (2017) 7–18, https://doi.org/ 10.1002/ente.201600096. [285] A. Saraeian, M.W. Nolte, B.H. Shanks, Deoxygenation of biomass pyrolysis vapors: improving clarity on the fate of carbon, Renew. Sust. Energ. Rev. 104 (2019) 262–280, https://doi.org/10.1016/j.rser.2019.01.037. [286] K.G. Kalogiannis, et al., First pilot scale study of basic vs acidic catalysts in biomass pyrolysis: Deoxygenation mechanisms and catalyst deactivation, Appl. Catal. B Environ. 238 (2018) 346–357, https://doi.org/10.1016/j. apcatb.2018.07.016. [287] G. Zhou, et al., Direct upgrading of fast pyrolysis lignin vapor over the HZSM-5 catalyst, Green Chem. 18 (7) (2016) 1965–1975, https://doi.org/10.1039/ C5GC01976A. [288] Y. Xue, P. Johnston, X. Bai, Effect of catalyst contact mode and gas atmosphere during catalytic pyrolysis of waste plastics, Energy Convers. Manag. 142 (2017) 441–451, https://doi.org/10.1016/j.enconman.2017.03.071. [289] T.L. Marker, et al., Integrated hydropyrolysis and hydroconversion (IH2®) for the direct production of gasoline and diesel fuels or blending components from biomass, part 2: continuous testing, Environ. Prog. Sustain. Energy 33 (3) (2014) 762–768, https://doi.org/10.1002/ep.11906. [290] A. Eschenbacher, et al., Performance of mesoporous HZSM-5 and Silicalite-1 coated mesoporous HZSM-5 catalysts for deoxygenation of straw fast pyrolysis vapors, J. Anal. Appl. Pyrolysis 145 (2020) 104712, https://doi.org/10.1016/j. jaap.2019.104712. [291] A. Eschenbacher, et al., Highly selective conversion of mixed polyolefins to valuable base chemicals using phosphorus-modified and steam-treated mesoporous HZSM-5 zeolite with minimal carbon footprint, Appl. Catal. B Environ. 309 (2022) 121251, https://doi.org/10.1016/j.apcatb.2022.121251. [292] A. Okuwaki, et al., The Liquefaction of Plastic Containers and Packaging in Japan, in Feedstock Recycling and Pyrolysis of Waste Plastics, 2006, pp. 663–708, https://doi.org/10.1002/0470021543.ch26. [293] S. Kumagai, et al., Aromatic hydrocarbon selectivity as a function of CaO basicity and aging during CaO-catalyzed PET pyrolysis using tandem μ -reactor-GC/MS, Chem. Eng. J. 332 (2018) 169–173, https://doi.org/10.1016/j.cej.2017.09.077. [294] M. Brebu, et al., Catalytic degradation of acrylonitrile−butadiene−styrene into fuel oil 1. The effect of iron oxides on the distribution of nitrogen-containing compounds, Energy Fuel 15 (3) (2001) 559–564, https://doi.org/10.1021/ ef000124x. [295] N. Miskolczi, et al., Fuels by pyrolysis of waste plastics from agricultural and packaging sectors in a pilot scale reactor, Fuel Process. Technol. 90 (7) (2009) 1032–1040, https://doi.org/10.1016/j.fuproc.2009.04.019. [296] N. Miskolczi, C. Wu, P.T. Williams, Fuels by waste plastics using activated carbon, MCM-41, HZSM-5 and their mixture, MATEC Web Conf. 49 (2016) 05001, https://doi.org/10.1051/matecconf/20164905001. [297] C. Witpathomwong, et al., Improving light olefins and light oil production using Ru/MCM-48 in catalytic pyrolysis of waste tire, Energy Procedia 9 (2011) 245–251, https://doi.org/10.1016/j.egypro.2011.09.026. [298] M. Goto, M. Sasaki, T. Hirose, Reactions of polymers in supercritical fluids for chemical recycling of waste plastics, J. Mater. Sci. 41 (5) (2006) 1509–1515, https://doi.org/10.1007/s10853-006-4615-2. [299] W.-T. Chen, K. Jin, N.-H. Linda Wang, Use of Supercritical Water for the Liquefaction of Polypropylene into Oil, ACS Sustain. Chem. Eng. 7 (4) (2019) 3749–3758, https://doi.org/10.1021/acssuschemeng.8b03841. [300] M. Watanabe, et al., Polyethylene conversion in supercritical water, J. Supercrit. Fluids 13 (1) (1998) 247–252, https://doi.org/10.1016/S0896-8446(98)00058-8. [301] T. Helmer Pedersen, F. Conti, Improving the circular economy via hydrothermal processing of high-density waste plastics, Waste Manag. 68 (2017) 24–31, https://doi.org/10.1016/j.wasman.2017.06.002. [302] M.S. Seshasayee, P.E. Savage, Oil from plastic via hydrothermal liquefaction: production and characterization, Appl. Energy 278 (2020) 115673, https://doi. org/10.1016/j.apenergy.2020.115673. [303] D.T. Chen, et al., Depolymerization of tire and natural rubber using supercritical fluids, J. Hazard. Mater. 44 (1) (1995) 53–60, https://doi.org/10.1016/03043894(95)00047-X. [304] Y. Liu, et al., The resource utilization of ABS plastic waste with subcritical and supercritical water treatment, Int. J. Hydrog. Energy 44 (30) (2019) 15758–15765, https://doi.org/10.1016/j.ijhydene.2018.08.012. [305] L. Zhang, et al., Hydrothermal conversion of scrap tire to liquid fuel, Chem. Eng. J. 285 (2016) 157–163, https://doi.org/10.1016/j.cej.2015.10.001. [306] S.J. Pathak, A. Gangal, V. Prabu, Direct liquefaction of discarded printer cartridge plastics and its Kinetic Modelling, Fuel Process. Technol. 228 (2022) 107147, https://doi.org/10.1016/j.fuproc.2021.107147. [307] X. Zhao, et al., Hydrothermal Treatment of E-Waste Plastics for Tertiary Recycling: Product Slate and Decomposition Mechanisms, ACS Sustain. Chem. Eng. 7 (1) (2019) 1464–1473, https://doi.org/10.1021/acssuschemeng.8b05147. [308] T. Funazukuri, T. Takanashi, N. Wakao, Supercritical extraction of used automotive tire with water, J. Chem. Eng. Jpn 20 (1) (1987) 23–27, https://doi. org/10.1252/jcej.20.23. [309] Z. Cheng, et al., Transformation of nitrogen, sulfur and chlorine during waste tire pyrolysis, J. Anal. Appl. Pyrolysis 153 (2021) 104987, https://doi.org/10.1016/j. jaap.2020.104987. [310] I. Hita, et al., Opportunities and barriers for producing high quality fuels from the pyrolysis of scrap tires, Renew. Sust. Energ. Rev. 56 (2016) 745–759, https://doi. org/10.1016/j.rser.2015.11.081. [311] J. Walendziewski, M. Steininger, Thermal and catalytic conversion of waste polyolefines, Catal. Today 65 (2) (2001) 323–330, https://doi.org/10.1016/ S0920-5861(00)00568-X. [312] M.Z. Stummann, et al., A perspective on catalytic hydropyrolysis of biomass, Renew. Sust. Energ. Rev. 143 (2021) 110960, https://doi.org/10.1016/j. rser.2021.110960. [313] W. Chen, et al., Aromatic hydrocarbons production and synergistic effect of plastics and biomass via one-pot catalytic co-hydropyrolysis on HZSM-5, J. Anal. Appl. Pyrolysis 147 (2020) 104800, https://doi.org/10.1016/j. jaap.2020.104800. J. Snow et al.