Life cycle assessment of alternative technologies for municipal solid waste and plastic solid waste management in the Greater London area

Life cycle assessment of alternative technologies for municipal solid waste and plastic solid waste management in the Greater London area

Chemical Engineering Journal 244 (2014) 391–402 Contents lists available at ScienceDirect Chemical Engineering Journal journal homepage: www.elsevie...

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Chemical Engineering Journal 244 (2014) 391–402

Contents lists available at ScienceDirect

Chemical Engineering Journal journal homepage: www.elsevier.com/locate/cej

Life cycle assessment of alternative technologies for municipal solid waste and plastic solid waste management in the Greater London area S.M. Al-Salem a,1, S. Evangelisti b, P. Lettieri b,⇑ a b

Polymeric Products Enhancement and Customization Program, Petroleum Research Centre, Kuwait Institute for Scientific Research (KISR), P.O. Box 24885, Safat 13109, Kuwait Department of Chemical Engineering, University College London (UCL), Torrington Place, London WC1E 7JE, UK

h i g h l i g h t s  We investigated environmental impact of the current municipal waste management in Greater London.  We analysed different advanced thermo-chemical technologies for plastic solid waste treatment.  Choice of technologies depends on market’s ability to take-in the petrochemical products.  Recycling textile and paper could bring the largest improvement for the environment.  Substitution of primary aggregates with IBAs has a significant impact in terms of GHG savings.

a r t i c l e

i n f o

Article history: Received 7 October 2013 Received in revised form 20 January 2014 Accepted 22 January 2014 Available online 31 January 2014 Keywords: Life cycle assessment Plastic solid waste Pyrolysis Hydrogenation Petrochemicals Waste management

a b s t r a c t The aim of this study is to present the results of a life cycle assessment performed for different scenarios reflecting the management, treatment and handling of plastic solid waste as a fraction of municipal solid waste (MSW) in the Greater London area. The study is divided in two parts: Part I comprises a LCA on the current MSW management strategy adopted in the Greater London area. This includes a materials recovery route via a dry Materials Recovery Facility (MRF) and an energy recovery route (incineration unit (IU) with combined heat and power). Part II investigates two alternative thermo-chemical treatment (TCT) technologies for the management of plastic solid waste (PSW): a low temperature pyrolysis (LTP) reactor and a hydrogenation reactor (VCC). The LTP process recovers valuable chemicals and petrochemicals (e.g., gases (C3–C4), liquid fractions (naphtha), waxes (atmospheric residue, AR) and heat in the form of steam), whilst the hydrogenation process produces syncrude and e-gas which is comparable to natural gas. The system expansion methodology was applied to the scenarios developed. A sensitivity analysis investigated different degrees of material recovery and reprocessing for the substitution of the relevant conventional processes. At the same time the study tackled the impact on the environment of introducing TCT units with the aim of petrochemicals’ recovery. Results showed that the current waste management system is more environmentally friendly compared with the landfill scenario for all the impact categories investigated. Moreover, the employment of the alternative TCT technologies investigated depends on the market’s ability to take-in the petrochemical by-products hence replacing their conventional production. Scenarios including pyrolysis appeared to be more environmentally friendly in terms of Greenhouse gas emissions when compared with hydrocracking, while the reverse was true for the eutrophication potential category. Ó 2014 Elsevier B.V. All rights reserved.

1. Introduction Municipal solid waste (MSW) is by far the most heterogeneous of all refuse, and is a direct result of activities in urban environments. ⇑ Corresponding author. Address: Department of Chemical Engineering, University College London (UCL), Room 312, Roberts Building, Torrington Place, London WC1E 7JE, UK. Tel.: +44 (0) 20 7679 7867; fax: +44 (0) 20 7383 2348. E-mail address: [email protected] (P. Lettieri). 1 This work was performed while undertaking a PhD in Chemical Engineering at University College London. http://dx.doi.org/10.1016/j.cej.2014.01.066 1385-8947/Ó 2014 Elsevier B.V. All rights reserved.

MSW consists of an organic fraction (wet waste: kitchen waste, food waste, straw, garden trimmings, sawdust, etc.) and a non-organic fraction (dry waste: glass, plastics, metals, ash, atone and bricks, etc.). The properties of waste differ immensely depending on many factors, such as the area of collection (rural, urban, industrial or commercial), seasonal variations and recycling levels [1]. In the UK, studies show that plastic solid waste (PSW) constitutes 7% of the final waste stream [2], whereas in the US, PSW found in MSW has increased from 11% in 2002 to 12.1% in 2007 [3]. The current economic climate gives PSW a new perspective

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Nomenclature AR AP CCGT CHP EfW EP GWP IBA ILCD ISO IU LCA

atmospheric residue acidification potential combined cycle gas turbine combined heat and power energy from waste eutrophication potential global warming potential incineration bottom ash International reference Life Cycle Data system International Standard Organization incineration unit life cycle assessment

as a sustainable feedstock. Since polymers have a high calorific value (comparable to heavy and gas oil), treating PSW thermochemically is a more preferable route to recovery and reprocessing processes (i.e., recovery through Material Recovery Facilities (MRF)), incineration processes and the conventional route of landfilling. Since thermo-chemical treatment (TCT) processes recover a number of valuable products and petrochemicals (e.g., gases, liquid fractions, waxes, syncrude, e-gas and steam), utilising PSW as a feedstock for such processes on an industrial scale warrants investigation and further study. Plastic solid waste produced in the UK capital is typically collected by different councils, boroughs, waste authorities and contractors. PSW ends up in different transfer stations (TSs) which distribute the waste to the relevant processing lines. Landfill is the most common method of MSW disposal, and it accounted for 49% of MSW disposal in UK in 2009 [4]. Many countries – including the UK – have established rules to limit the amount of MSW sent to landfill, and have set drivers to establish more environmental sustainable technologies for dealing with the treatment of MSW. As a result, different strategies have been proposed to reduce the environmental impacts of the waste management process. These strategies include direct incineration, MSW sorting, gasification, anaerobic digestion, MSW derived ethanol production [5–11]. These methods need to be investigated in order to identify the most appropriate strategies for different municipalities. One of the most important criteria to inform decision making on the most sustainable option for waste management is the evaluation of the environmental impacts. The objective of this paper is to investigate alternative technologies for the treatment of the plastic solid waste contained in the MSW in order to improve the environmental impact of the waste management system in the London area. In order to do this, we first report the results of a life cycle assessment (LCA) study on the current waste management strategy adopted in Greater London for the treatment of MSW (Part I). In Part II, three different scenarios for the treatment of PSW as fraction of MSW are then investigated: the first scenario reflects the current strategy employed for waste management in London, and comprises a materials recovery route via a dry materials recovery facility and reprocessing process, while the second and third scenarios include two different TCT industrial scale technologies for the treatment of PSW. The TCT units chosen are a low temperature pyrolysis (LTP) reactor working under BPÒ technology and a Vea Combi-Cracking (VCCÒ) hydrogenation reactor. 2. Methodology 2.1. Life cycle assessment (LCA) LCA is considered one of the most developed environmental assessment tools to evaluate the performance of technologies

LPT MRF MSW PA POCP PSW SA SF TCT tpa TS VCC

low temperature pyrolysis reactor Material Recovery Facility municipal solid waste primary aggregates photochemical ozone creation potential plastic solid waste secondary aggregates substitution factor thermo chemical treatment tonnes per annum transfer station Vea Combi-Cracking hydrogenation reactor

when the location of the activity is already defined [12]. Life cycle thinking approach is fundamental for many European directives, such as the ‘‘Thematic strategy on the prevention and recycling of waste’’ and the ‘‘Waste framework directive’’. The ISO 14040 life cycle assessment framework is employed in this study to conduct the environmental impact assessment of the different scenarios developed [13]. The ILCD (International reference Life Cycle Data system) handbook is used as a guide internationally recognized to conduct the LCA [14]. The study takes into consideration the following impact categories: global warming potential (GWP) as an indicator of Greenhouse effect; acidification potential (AP) as an indicator of acid rain phenomenon; photochemical ozone creation potential (POCP) as an indicator of photo-smog creation and eutrophication potential (EP) as an indicator of over fertilization of water and soil as defined by CML [15]. These categories have been chosen given their environmental significance and the fact that they are internationally accepted in accordance with the recommendations of ISO 14044 [13]. The impact categories listed above are based upon a distinct identifiable environmental mechanism and they ensure that the results are robust enough to form a basis for further consideration or decisions [14,16]. The time horizon of the assessment is 100 years. Life cycle assessment modelling is undertaken using the widely-used software GaBi 5 (PE International), implementing the PE database [17]. In the definition of LCA, the term ‘product’ includes not only material products but can also include service systems, for example waste management system [18]. In LCA, a multifunctional process is defined as an activity that fulfills more than one function, such as a waste management process dealing with waste and generating energy [19]. It is then necessary to find a rational basis for allocating the environmental burdens between the processes. The problem of allocation in LCA has been the topic of much debate [12,20]. The ISO standards [13] recommend that the environmental benefits of recovered resources should be accounted for by broadening the system boundaries to include the avoided burdens of conventional production [21]. The same approach is recommended for product labelling provided that it can be proved that the recovered material or energy is actually put to the use claimed [22]. This approach is also applied in the work presented here. Following the methodological approach of Clift et al. [12] for Integrated Waste Management (IWM), a pragmatic distinction between Foreground and Background is made in this study, considering the first as ‘the set of processes whose selection or mode of operation is affected directly by decisions based on the study’ and the second as ‘all other processes which interact with the Foreground, usually by supplying or receiving material or energy’. This work is defined as an ‘‘attributional LCA with system expansion’’ [23]. The general understanding is that the aim of consequential LCA is to model environmental consequences of

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marginal changes, whereas attributional LCA study describes the current situation [24], and ‘‘using average data from environmental impact connected to electricity and heat production is suitable in attributive assessments, whereas marginal production is justified when a consequential approach is taken’’ [24]. In this work average data is used for both energy production and material substitution. 2.2. Goal and scope definition This LCA study is divided in two parts. The goal of the first part is to analyse the current waste management system in the Greater London. It includes the treatment of MSW generated by four boroughs of the Greater London, i.e., Greenwich, Lewisham, Westminster and Bromley, and MSW generated by the City of Exeter in Devon which is transported and treated in Greater London. The MSW generated in these areas is currently sent part to a MRF in Greenwich and part to an incineration unit (IU) in Lewisham (Greater London). As a reference scenario, an engineered landfill plant with electricity recovery is assumed; this is done because landfill is still the primary destination for MSW in the UK. The landfill is assumed to be located at an average distance amongst the different generation points. The functional unit considered for this part of the study is 552,141 tpa, which is equal to the combined throughput from the MRF and the IU. Different substitution factors are considered for the material recovered in the MRF station and reprocessed before entering again the market. The goal of the second part is to investigate three different alternatives for the treatment of the total PSW sent to the MRF station by the locations considered: scenario 1 includes recycling of plastic through a MRF plant and reprocessing before being sold to the market; whilst scenarios 2 and 3 incorporate a pyrolysis and a hydrogenation reactor, respectively. The functional unit for this part of the study is a defined quantity of PSW: 1000 tonnes per annum (tpa). This represents the capacity of an average TCT plant for plastic solid waste [25]. Different recycling rates are considered for the substitution of recycled products with virgin or conventional materials for the three processes analysed. As suggested by Laurent el al. [26], four major classes of functional unit can be used in a waste management study: unitary functional unit, defined by a unitary measure; generation based functional unit, defined by the waste generated in a delimited region for a specific period of time; input-based functional unit defined by the waste amounts entering a given facility, and output-based functional unit, defined by the waste by-products, i.e. the amount of recovered energy (steam or electricity) or recycled material. In this study we have used a generation based functional unit to assess the environmental impact of the current MSW management in the case study area (consistent with the specific purpose of Part I), and an input-based functional unit to assess different scenarios for the treatment of the PSW fraction (consistent with the purpose of Part II). Figs. 1–3 show the overall system investigated in the first part (Figs. 1 and 2) and in the second part of the work (Fig. 3) indicating the foreground system (resulting in direct burdens), and the background system (resulting in indirect burdens). In particular, following the study of Perugini et al. [25], it is assumed that materials recycled in the MRF and reprocessed can substitute the production of virgin materials; power production and bottom ashes generated by the IU can substitute electricity from the grid and secondary aggregates for road construction respectively; petrochemical products and steam produced from the LTP plant can substitute commercial products and electricity from the grid; chemical products from the VCC can substitute commercial products. The total life cycle inventory in this study is reported as the sum of the following Eq. (1) [6]:

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Direct burdens ðresulting from the ForegroundÞ þ Indirect burdens ðresulting from the BackgroundÞ  Avoided burdens ðresulting from the displaced productsÞ ð1Þ In the first part of the work, six different combinations are considered for the recovery of the bottom ashes from IU and their substitution with primary and secondary aggregates (see Table 1). In the second part of the study, nine different combinations are considered for the treatment of PSW (via MRF, LTP or VCC) for the three different levels of material recovery substitution in the system expansion (see Table 2). The electrical production is displaced using average data (i.e., UK electricity mix). The electricity mix in the UK consists of a number of contributing sources that result in what is known as the average mix. These sources and their contribution are described in Table 3. The combinations studied in the first part of the work are also compared to a landfill scenario (reference scenario) in which all the waste is directly sent to landfill from its point of generation. In the UK, the proximity principle has been described and implemented in many facilities involving energy from waste (EfW) treatment, and it has been part of incineration units operation and other EfW schemes [1,27,28]. The principle is concerned with treating the waste as close as possible to its point of origin and in this study the proximity principle was considered in the development of the overall system studied. The TCT units are all assumed to be on the same location as the MRF station to avoid the extra travelling distance to deliver the plastic feed to the LTP or hydrogenation unit. By developing the overall system thus, the proximity principle is met. 3. Life cycle inventory: management of MSW in Greater London 3.1. Waste breakdown and flows The UK Office for National Statistics [29] published in their 2009 report the amount of co-mingled waste produced by each individual in the country (0.495 tpa of total waste per resident). This value was considered in this study in order to determine the total amount of waste produced in each of the five areas considered in this study (Greenwich, Lewisham, Exeter, Westminster and Bromley) and served by the MRF station and the CHP IU plant. Table 4 indicates the population of the area, the total amount of waste and the breakdown of waste fractions considered in this study. The percentage of dry waste was taken as 46%, of which plastics make up 7% of the total amount [2]. The plastics breakdown (%) considered in this study is shown in Table 5 in accordance to the UK plastic breakdown [30]. The materials recovery facility processes the dry fraction generated from the Boroughs of Greenwich and Lewisham, as well as the City of Exeter in Devon. This amounts to 137,303 tpa, which was considered to be the MRF throughput in this analysis. The incineration unit processes co-mingled waste generated from the Boroughs of Greenwich, Lewisham, Westminster and Bromley. In addition, the IU feed stream also includes 30,000 tpa of collected waste from the Greater London area (GLA). In this study, the latter was assumed to be organic waste. Furthermore, no glass or metals were assumed to enter the IU feed stream and hence the glass, metals packaging and white goods fractions generated from Bromley and Westminster (39,468 tpa) were excluded from the analysis. This assumption was made due to the fact that the MRF receives only dry waste from three points of origin only (Greenwich, Lewisham and Exeter). Therefore, the IU throughput feed consists of the wet waste fractions generated by Greenwich, Lewisham,

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MSW_B1

MSW_B2

MSW_B3

MSW_B4

MSW_B5

TS_B1

TS_B2

TS_B3

TS_B4

TS_B5

MSW_GLA

Consumables Emissions to air, water, soil

Water Energy

Transport

Fuel

IBA

IU

Steam

Power

Landfill

Acronym: MSW:municipal Solid Waste TS: Transfer Station B: Borough IU: Incineration Unit IBA: Inceneration Bottom Ash MRF: Material Recovering Facility GLA: Greater London Area

MRF

Recycled Aggregates

Recovered Products

Landfill Foreground

Process Steam from natural gas plant

Electricity Grid

Primary/Secon dary Aggregates

Background Production of Virgin

Fig. 1. MRF + IU scenario and system boundaries – Part I.

Bromley and Westminster and it was also assumed that the wet fraction of Exeter is sent to the IU plant. This is consistent with the maximum capacity of the IU plant (420,000 tpa) considered. This also services the integration strategy undertaken in this work, by delivering the waste from similar points of origin to the unit operation lines. Therefore, the IU throughput amounts to 414,838 tpa. These assumptions relate this work to the current practices of waste management undertaken in the areas studied. The specific methodological aspects are described in detail in the following sections. Mass and energy calculations with respect to each process and activity are reported, including the products and energy displaced. 3.2. Materials Recovery Facility (MRF) station The purpose of a Materials Recovery Facility is to separate comingled materials into their individual material streams and prepare them for sale into the commodity markets after reprocessing [31]. The Greenwich MRF station is considered in this study as the route for dry recyclables in the overall system developed and a 146,000 tpa maximum capacity of the station is assumed. The electricity and diesel consumption of the MRF station production line depend on the amount of dry waste fed into its processing line. The MRF consumes diesel in its different processing stages and diesel consumption is calculated with respect to the total MRF throughput. In the MRF station, dry recyclables are recovered so that, after adequate reprocessing, can be sold, including plastics, glass, steel. This follows the common practice undertaken in the British capital. Different substitution factors are assumed for the reprocessed products from the MRF. When a reprocessed material does not substitute the virgin product, this is assumed to be sent to landfill.

The products separated in the MRF station and reprocessed are assumed to substitute virgin materials, produced otherwise by specific production technologies. Table 6 show the technologies considered and the associated environmental burdens used to evaluate the total avoided environmental impact of the system studied. 3.3. Incineration unit implemented Energy recovered from waste incineration was considered in this study. The IU plant was assumed to be located in London within the Borough of Lewisham. The plant operates an incineration unit based on the mass-burn process. The maximum capacity of the plant is assumed to be 420,000 tpa of waste and the feed to the IU consists of co-mingled waste that originates from the Boroughs of Westminster, Bromley, Lewisham, Greenwich and a collection delivery from all of the GLA. The electricity is generated using a 35 MW steam turbine generator and the electrical current is transformed for national grid exporting. The combined heat and power (CHP) plant was assumed to produce both electricity for the national grid and steam to export to nearby industrial users. The electricity and heat generation efficiencies were taken for the standard CHP plant as reported by Murphy and McKeogh [7]: the value of gelec was taken as 18%, whilst gHeat equals 50%. The maximum seasonal efficiency of domestic boilers (gBoiler) was taken in this study as 92% [32]. Bottom ash is produced from incineration as a result of the combustion process and is the largest residue resulting from incineration processes. It differs in composition, but consists mainly of aggregate (80%), organics (5%) and other trace amounts, and it is used mainly in infrastructure projects [33]. In Switzerland, bottom ash is mainly landfilled, however, in countries like Denmark and

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MSW_B2

MSW_B1

MSW_B3

MSW_B4

MSW_B5

MSW_GLA

Consumables

Water

Emissions to air, water, soil

Transport

Energy Fuel

Engineered Landfill

Power

Foreground

Electricity Grid

Background

Fig. 2. Landfill scenario and system boundaries – reference scenario.

PSW Consumables

Water

Emissions to air, water, soil

Energy Fuel

LTP

MRF

VCC Acronym: PSW: Plastic Solid Waste LTP: Low Temperature Pyrolysis VCC: Vea Combi-Cracking hydrogenation reactor

Recovered Products

Landfill

Petrochemicals

Landfill

Steam

Landfill

Products & Chemicals

Foreground Background Production of Virgin

Commercial Production

Process Steam from Natural gas

Commercial Production

Fig. 3. PSW scenarios and system boundaries – Part II.

Sweden it is used as road fillers [33]. In this study, bottom ash was estimated as 20% of the IU throughput [34] and it was assumed to be split into two fractions. A fraction was assumed to be recycled as an aggregate for the construction sector, and the rest was landfilled. Two different commercial processes were assumed to be

substituted by the recycled fraction of the incineration bottom ash (IBA): primary aggregate production (from virgin material) from marine sand deposit and secondary aggregate production from recycled material (i.e., asphalt). Table 7 shows the environmental impact factors used for this assessment [35].

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S.M. Al-Salem et al. / Chemical Engineering Journal 244 (2014) 391–402 Table 1 Six combinations for the treatment of MSW with bottom ash treated as PA and SA (Part I of the study).

MRF

Substitution factor (%)

IBA substituted as secondary aggregate

IBA substituted as primary aggregate

100 50 0

1 (Subst Fact 1_SA) 2 (Subst Fact 0.5_SA) 3 (Subst Fact 0_SA)

4 (Subst Fact 1_PA) 5 (Subst Fact 0.5_PA) 6 (Subst Fact 0_PA)

Table 2 Nine combinations analysed for the treatment of PSW with three different degree of material recovery (Part II of the study). Substitution factor (%)

MRF–PSW recycling

LTP

VCC

0 50 100

1 (MRF_1) 2 (MRF_0.5) 3 (MRF_0)

4 (LTP_1) 5 (LTP_0.5) 6 (LTP_0)

7 (VCC_1) 8 (VCC_0.5) 9 (VCC_0)

Table 3 Sources for the average UK electricity mix [17]. Mix of sources and their contribution Source

%

Natural gas Hard coal Nuclear Hydro energy Wind Fuel oil Biogas Electricity from waste Biomass Coal gases

44 28 18 2 2 1 2 1 1 1

1,134,560

Waste generation (tonnes/year)

561,600

Plastic (dense + film) (%) Textiles (%) Glass (%) Metal packaging (%) White goods/metal (%) Fines (%) Paper and cardboard (%) Organic waste (%)

7 3 7 3 5 3 18 54

Dry fraction (% of total waste generated)

46

Table 5 Polymer by type in the case study area. Plastic amount (tonnes/year) Polyethylene (%) (LDPE + HDPE) Polypropylene (PP) (%) Polyvinyl chloride (PVC) (%) Polystyrene (PS) (%) Rest (%)

3.4. Transfer stations (TS) considered Generated waste from each individual point of origin (borough, city, etc.) is typically transferred to a large collection center or a depot known as a transfer station (TS). This activity is considered as a part of the development of the system studied in the first part of this work. The electrical and diesel consumption for the transfer station were calculated. 3.5. Transport

Table 4 Total amount of waste and waste fraction breakdown considered in the study. Population

turbine. An average efficiency of 90% is assumed and the emission factors considered per kW hth of steam produced are shown in Table 6.

39,311 37.5 (24.3 + 13.2) 18.5 18.8 6.3 18.9

The UK average mix of technologies for electricity production was considered in this analysis to assess the avoided burdens due to the electricity produced by the IU [17]. Steam from a combine cycle gas turbine plant (CCGT) using natural gas is assumed to be substituted by the production of steam in the incineration unit (and in the LTP plant as showed later in Section 4.1.1). A CCGT plant generates electricity via a gas turbine and waste heat is used to make steam to generate additional electricity via a steam

The transportation distances vary between the different units considered in this work. Transportation contributes to the total environmental burden in terms of airborne pollutants and these include CO, NOx, hydrocarbons (HC), PM10, CO2 is also considered as a part of the transportation load activity. The contribution of CO2 is the result of the diesel consumed by the trucks that transport the waste from the TSs to the MRF and the IU sites. The transportation burden is considered a direct emission that is added to the final burdens evaluation. It was assumed that each borough manages ten diesel engine trucks, which were assumed to have a capacity of 40 tonnes (maximum). Each truck operates with a maximum payload (40 tonnes), i.e., full capacity, as described by the Volvo Truck Co. [36] and the number of trips required by all trucks was calculated as shown in Eq. (2).

NTR ¼

WG ðtonnes=weekÞ MPL ðtonne=truckÞ  NT ðtruck=weekÞ

ð2Þ

where NTR is the number of trips required for each truck to and from the TS, WG is the amount of waste transferred by each truck (tonne/week), MPL is the maximum truck payload (40 tonnes/truck) and NT is the number of trucks managed by each borough for TS activities (10 trucks/week). In order to calculate the amount of CO, NOx and hydrocarbons (HC) emitted from the truck exhaust, the Euro 4 fuel category was assumed for the truck engine emissions and the conversion factors were taken from the engine specifications described by Volvo Truck Co. [36]. 3.6. Engineered landfill Few countries have reached a zero reliance on MSW landfill disposal. Sweden is a prime example, where household municipal and combustible wastes are strictly prohibited from being landfilled, and countries like Germany, Belgium and France are approaching such a status, whilst in the UK, almost 50% of the waste is still landfilled. The landfill scenario considers recovery of electricity, whereby the landfill gas is collected from and split into two fractions: the first is flared and the second one is fed to a CHP process. The transportation to the landfilling site is also accounted for in this study following the same approach described previously. The landfill site assumed was the Basillond site in Essex, with an average distance of 36.4 miles from the areas considered in the study.

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S.M. Al-Salem et al. / Chemical Engineering Journal 244 (2014) 391–402 Table 6 Avoided impact associated with the substitution of commercial products by recovered products. Products

GWP (kg CO2eq/kg product)

AP (kg SO2eq/kg product)

POCP (kg C2H4eq/kg product)

EP (kg phosphateeq/kg product)

Commercial products

Plastics Glass Metal packaging White goods and metal scrap Textiles Paper and cardboard Fines Electricity Steam

1.85 0.562 1.13 1.13 5.14 0.658 0.524 0.56 0.25

5.75  103 2.64  103 3.38  103 3.38  103 23.2  103 3.78  103 3.74  103 1.9  103 0.12  103

4.76  104 9.17  104 5.56  104 5.56  104 18.8  104 4.07  104 3.54  104 1.1  104 0.23  104

1.49  103 2.4  104 3.6  104 3.6  104 6.26  103 2.12  103 1.73  103 1.6  104 0.2  104

Virgin LLDPE production Commercial white glass for packaging Steel section Steel section Textile refinement (cotton) Corrugated board base paper Paper, newsprint UK electricity mix Steam from CCGT natural gas plant

Table 7 Avoided impacts associated with the substitution of primary and secondary aggregates by recycled IBA. Avoided environmental impacts

Primary aggregates (marine sand)

Secondary aggregates (asphalt)

GWP (kg CO2 eq/kg aggregate) AP (kg SO2 eq/kg aggregate) POCP (kg C2H4eq/kg aggregate)

34.24 0.606 9.3  102 4.84  102

2.42 12.13  103 8  107 7.06  104

EP (kg PO3 4 eq/kg aggregate)

4. Life cycle inventory: alternative treatments for the plastic solid waste fraction 4.1. Thermo-chemical treatment (TCT) technologies incorporated in the system developed In this study two TCT industrial technologies (the BPÒ LTP and VCCÒ hydrogenation processes) are incorporated in the overall system developed. In order to compare each technology with the conventional waste treatment routes employed in London, each TCT technology is assigned to a different scenario. The first scenario (baseline) includes the MRF to reflect the current waste management treatment adopted in London for the PSW; the second includes the LTP process; while the third scenario includes the VCC process. The feed to each scenario is assumed to be a fraction of the plastics recovered by the MRF. 4.1.1. Low temperature pyrolysis technology The pyrolysis technology incorporated in this work is the BPÒ LTP or polymer cracking technology, as described previously by Tukker et al. [37] and Perugini et al. [25]. This pyrolysis technology was commissioned by BPÒ in a pilot scale and is also known as BP cracking technology [38]. The process accepts dry plastics as indicated by the feed criteria described in Table 8. It is assumed that the unit receives a plastic feed of 1000 tpa diverted from the PSW generally sent to the MRF station and which mainly consists of polyolefins (PE + PP) (83%) [37]. It is very important to satisfy the chlorine content in the pyrolysis reactor by not exceeding the PVC amount being fed to the unit. The plastics breakdown previously shown in Table 5 is considered in this analysis. Plastics (namely polyolefins) undergo a certain treatment, mainly concerned with size and chlorine content reduction, to meet with the requirements of the BPÒ polymer cracking (pyrolysis) process. Introducing a pyrolysis reactor provides the option of recovering a number of valuable chemicals (considered in this study), including rich gases and tars (heavy waxes and liquids). These chemicals can substitute a number of petrochemicals and in a consequential order include, propane (C3) and butane (C4), atmospheric residue (AR), naphtha and heat (energy) in the form of steam [25]. Table 9 summarizes the inputs and outputs of the BP LTP technology. The processes to produce the input materials

Table 8 LTP reactor feed criteria. Polymer type input

% Of the feed mix (%)

Amount fed in this study (tpa)

Polyolefins (PE + PP) Polyvinyl chloride (PVC) Polystyrene (PS)

>83 <2 <15

830 20 150

are considered in terms of their GWP, AP, POCP and EP contribution to the overall system and are considered as an indirect burden. The petrol chemical by-products are also considered in the system expansion and the processes to produce the equivalent amount of conventional products are considered as an avoided emission in this study. Table 10 also indicates the commercial products which are assumed to be substituted by the LTP process outputs. 4.1.2. Veba-Combi Cracking hydrogenation By definition, hydrogenation is the process of molecular cracking into highly reactive free radicals which are saturated with hydrogen as they form. The process integrated here and previously described by Tukker et al. [37] is known as the VCC process. The main criterion of polyolefin feed is concerned with the PVC content (610%) and in this study the VCC unit has a PVC content of 10%. The feed in such processes is typically sent to a depolymerizing unit to produce a light top product (consisting of 71 wt% C5+, with a boiling range of 400 °C and non-condensable (C3–C4) gases) and a heavy bottom product. The main product of this process is the syncrude produced, which can replace crude oil in a 1:1 ratio. The plastics breakdown considered is similar to that for the LTP process. The process is assumed to have a throughput feed of 1000 tpa and Table 11 summarizes the process throughput and the plastics breakdown considered in this study. Table 12 shows the main input materials and chemicals produced by the VCC process and their amounts per unit of waste treated. The commercial products assumed to be substituted by the VCC chemical by-products are: crude oil (substituted by syncrude); natural gas (substituted by e-gas); and HCl (substituted by the HCl from the VCC process). Solid waste, residue and CaCl2 are assumed to be sent to landfill. 5. Results 5.1. Life cycle impact assessment of the current waste management strategy in the Greater London area In this section, the focus is on the environmental impacts (GWP, AP, POCP and EP) for the scenarios studied. It is important to note that the results presented here are often negative; this means that environmental interventions can be avoided and the impact of the studied scenario is reduced. This argument strengthens the case for

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Table 9 Summary of inputs and outputs considered in this study for the LTP process and their amount requested for unit of waste treated [25].

Table 12 Summary of inputs and outputs considered in this study for the VCC process and their amount requested for unit of waste treated [25].

Input materials

Amount required

Input materials

Amount required

Sand CaO Water Naphtha Electric energy

0.0085 kg/kg Feed 0.046 kg/kg Feed 0.002 m3/kg Feed 0.131 MJ/tonne Feed 0.212 MJ/tonne Feed

Steam Electric energy Natural gas CaO Hydrogen

0.112 MJ/kg Feed 0.96 MJ/kg Feed 4.62 MJ/kg Feed 0.001 kg/kg Feed 0.011 kg/kg Feed

Output products

Amount obtained

Output products

Amount obtained

Gases (C3–C4) Liquid (Naphtha) Wax (AR) CaO CaCl2 Steam Sand and coke Waxy filter (deposit)

0.147 kg/kg Feed 0.265 MJ/kg Feed 0.448 kg/kg Feed 0.04 kg/kg Feed 0.017 kg/kg Feed 1.48 MJ/kg Feed 0.076 kg/kg Feed 0.046 kg/kg Feed

Syncrude e-Gas HCl CaCl2 Solid waste Residue

0.822 kg/kg Feed 0.09 kg/kg Feed 0.005 kg/kg Feed 0.0041 kg/kg Feed 0.05 kg/kg Feed 0.066 kg/kg Feed

Table 10 Summary of the substituted products assumed for the LTP by-products. Output products

Substituted products

Gases (C3–C4) Liquid (Naphtha) Wax (AR) CaO CaCl2 Steam

LPG (C3/C4 compounds) Naphta Vacuum residue CaO CaCl2 Steam from natural gas

viewing waste as a resource rather than a burden on the urban environment, society and the industrial community. Figs. 4–6 show the results of the first part of study. Fig. 4 shows the total environmental impacts of the six combinations investigated plus the reference scenario (landfill). The most environmentally friendly combinations are obtained when the recycled fraction of the IBA is assumed to substitute primary aggregates (indicated as PA in the figures, while SA indicates secondary aggregates), as per combination 4–6 (see Table 1). The amount of avoided burdens in these combinations is substantial in a way that the influence of the substitution factor for the recycled product from the MRF is negligible. This is because the burdens associated with the production of primary aggregates (i.e., from marine sand) are much bigger than the burdens associated with the production of aggregates from secondary materials (i.e., 16 times larger for the GWP). However, all scenarios display a lower impact for all impact categories when compared with the engineered landfill. In terms of GWP, the combinations with SA (when the recycled fraction of IBA is assumed to substitute secondary aggregates) show a positive impact for every substitution factor assumed, and the total amount of CO2 eq becomes larger at lower substitution factors. Fig. 5 shows a contribution analysis to identify the hotspots in all six combinations investigated, in order to find the process which contributes the most to the overall GWP and AP values. The contribution analysis is shown for two of the four impact categories analysed (GWP and AP), as they are considered the most significant and better understood impact categories in waste Table 11 VCC unit feed criteria. Polymer type

Amount (tpa)

% Of the VCC feed

Plastics mix (VCC feed) PO (PP + PE) PVC

1000 900 100

100 90 10

management studies. For the combinations with PA, the avoided burdens from the IU are the main contributors to the (negative) final impact for both the categories (GWP and AP). On the other hand, when SA are considered, the contribution of the IU is positive for the GWP – and it varies from 95% to 70%, while it is again negative for the AP – from 38% to 90%. In terms of GWP, this is mainly due to the CO2 content of the flue gas which is emitted in the atmosphere after the combustion of the waste in the IU, while in terms of AP the electricity which is produced in the CHP allows substituting a considerable quantity of burdens from the electricity mix (hence a negative impact). The contribution of the transfer station is totally negligible as it is shown in Fig. 5, while the transport is negligible in all the contribution except for the AP impact in the reference scenario. This is mainly due to the SOx emissions associated with the transport of the waste from the point of generation to the landfill which is strictly dependent by the distance assumed in this study. The MRF contributes positively (with a negative value) to the final impacts, thanks to the recycled products that are reprocessed and sold as substitutes of commercial products. Fig. 6 shows the detailed results of the contribution analysis for every category of products recycled through the MRF station. In terms of AP, the recycling of the paper contributes for 35% to the total (avoided) impact associated with the MRF, while the textile contributes for 30% and plastic for 18%. In terms of GWP, instead, the recycling of plastic and textile contribute by nearly 85% to the total impact associated with the MRF station. For both the indicators, the burdens associated with the production of electricity and diesel supplied to the TS is negligible. 5.2. Life cycle impact assessment of PSW management scenarios Figs. 7–9 show the results of the second part of the study. Fig. 7 illustrates the total impacts for the nine combinations investigated for the treatment of 1000 tpa of PSW. The values are all positive, except for the following scenarios and impact categories: the VCC scenario with a substitution factor equal to 1 used for all the byproducts, for the AP, and EP impact categories; for the VCC scenario with substitution factor 0.5 for the EP impact category; and for the LTP scenario with substitution factor 1 for the POCP impact category. A positive value means that the amount of direct and indirect burdens associated with the foreground and background processes are bigger than the avoided burdens. In terms of AP, the MRF scenario is the best environmental solution when the substitution factor is in the range 0.5–1. If the market conditions are unfavorable and less than 50% of recycled products can be sold as substitutes of virgin materials, the most preferable treatment is the VCC process, assuming a substitution factor for its by-products of 100%. The

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Impact Indicator AP [kg SO2-Equiv.]x10(5); EP [kg Phosphate-Equiv.]x10(5) GWP [kg CO2-Equiv.]x10(8)]; POCP [kg Ethene-Equiv.]x10(4)

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50 0 -50 -100 -150 -200 -250 -300 -350 -400 Acidification Potential

Subst Fact 1_SA

Eutrophication Potential Global Warming Potential

Subst Fact 0.5_SA

Photochem. Ozone Creation Potential

Subst Fact 0_SA

Subst Fact 1_PA

Fig. 4. Environmental impacts for the six scenarios investigated in Part I (showing sensitivity analysis on the substitution factor applied to primary (PA) and secondary (SA) aggregates). Landfill is shown as a baseline (see Table 1).

100% 80%

Relative Contribution (%)

60% 40% 20% 0% -20% -40% -60% -80% -100% AP

GWP

Sub Fact 1_SA

AP

GWP

Subst Fact 0.5_SA

IU

AP

GWP

AP

GWP

Subst Fact 0_SA Sub Fact 1_PA

MRF

TS

Transport

AP

GWP

Subst Fact 0.5_PA

AP

GWP

AP

Subst Fact 0_PA

GWP

Landfill

Landfill

Fig. 5. Hot spot analysis for GWP and AP indicators for the six scenarios + landfill.

latter is the most environmentally friendly solution also in terms of EP, but not in terms of GWP and POCP. For the GWP, the best solution is the MRF (SF = 1), while the LTP process (SF = 1) shows a slightly higher impact. Comparing the three alternative scenarios for a substitution factor less than 50%, the best solution is LTP, then MRF and VCC is the last. In terms of POCP, the VCC process has the highest impact, almost 5 times greater than the LTP process, mainly due to the supply of natural gas to the process. Fig. 8 shows the contribution analysis for the nine combinations analysed. For the MRF station, the burdens are mainly arising from the process (supply of electricity and diesel) when the substitution

factor is 1, while the landfilling of the unsold products is the main contributor in the other two combinations (SF = 0.5–0). The avoided burdens are almost negligible for these two combinations. For the LTP and VCC processes with SF equal to 1, the avoided burdens due to by-products substitution contribute for 30% and 10% for the AP and GWP impact for the first scenario, and for 50% and 40% for the second scenario. For both technologies where the SF = 0, the impact of landfilling of the unsold by-products is almost negligible. Finally, Fig. 9 shows the depletion of fossil fuel for the nine combinations. The worst results are obtained for the VCC technology, mainly due to the supply of natural gas to the process. The best

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Relative Contribution (%)

20% 0% -20% -40% -60% -80% -100% AP

GWP Sub Fact 1_SA

MRF

Avoided Plastic

Avoided Metal

Avoided Glass

Avoided Paper

Avoided Textile

Landfill

100

80

60

40

20

0

-20

-40 AP EP GWP POCP AP EP GWP POCP AP EP GWP POCP AP EP GWP POCP AP EP GWP POCP AP EP GWP POCP AP EP GWP POCP AP EP GWP POCP AP EP GWP POCP

Impact Indicator AP [kg SO2-Equiv.]x10(2); EP [kg Phosphate-Equiv.]x10(2) GWP [kg CO2-Equiv.]x10(5)]; POCP [kg Ethene-Equiv.]x10(6)

Fig. 6. Hot spot analysis for MRF when substitution factor is 1.

MRF_1

MRF_0.5

MRF_0

LTP_1

LTP_0.5

LTP_0

VCC_1

VCC_0.5

VCC_0

Fig. 7. Environmental impacts for the nine scenarios investigated in the second part on PSW management (see Table 2).

values are instead obtained for the LTP scenario, which shows a negative value (savings of fossil resources) for the first two combinations (SF 1 and 0.5). This is mainly due to the avoided emissions associated with the production of naphtha.

6. Discussion The model developed in this study is based on the current waste management system in the Greater London area chosen as a case study. Several assumptions are made to allow modelling a complex system as the waste treatment. About the MRF, the process flow diagram built in GaBi does not include a specific re-processing unit for each recycled material. There is no a general consensus in the waste management expertise sector on which specific process is going to be offset by the recycling of specific materials from waste. This depends on several factors, such as the quality of the MSW collected, the collection method, and the separation method used at the MRF and the reprocessing process implemented [39]. Despite this, the uncertainty of the reprocessing process and the quality of the recycled material is

taken into account in this study by varying the degree of the substitution factor for the recycled materials. There is no general consensus in the relevant literature on which is the suitable product that can be substituted by reprocessed incineration bottom ash. In UK the primary aggregates’ market is dominant compared with the secondary aggregates’ market and it is plausible to assume that IBAs will replace primary resources for road construction. This study is based on average data. A more specific investigation is needed to identify the actual route of IBA substitution.

7. Conclusions This study presents the results of an attributional life cycle assessment with system expansion performed on different scenarios that reflect the current municipal solid waste management and waste plastics treatment in Greater London. Possible futuristic scenarios have also been analysed for the management of plastic solid waste (PSW). At present, the waste produced in Grater London is sent to a dry Materials Recovery Facility (MRF) and to an incineration

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100%

Relative Contribution (%)

80% 60% 40% 20% 0% -20% -40% -60% AP

GWP

MRF_1

AP

GWP

MRF_0.5

AP

GWP

MRF_0

AP

GWP

LTP_1

Process

AP

GWP

LTP_0.5

Avoided

AP

GWP

LTP_0

AP

GWP

VCC_1

AP

GWP

VCC_0.5

AP

GWP

VCC_0

Landfill

Fig. 8. Hot spot analysis for GWP and AP indicators for the nine scenarios.

2.5E+08

Depletion of fossil resources (MJ)

2.0E+08

1.5E+08

1.0E+08

5.0E+07

0.0E+00

-5.0E+07 MRF_1

MRF_0.5

MRF_0

LTP_1

LTP_0.5

LTP_0

VCC_1

VCC_0.5

VCC_0

Fig. 9. Depletion of fossil resources in the three treatment technologies for PSW.

unit (IU) with combined heat and power production. This conventional route of waste treatment is analysed in the first part of this study and is compared with a reference landfill scenario. In the second part of the study, two alternative technologies for the treatment of the PSW (namely pyrolysis and hydrogenation)are investigated. A number of conclusions can be drawn from this study:  The first part of the study showed that the actual waste management system in Greater London has a lower environmental impact compared with the reference scenario (landfill).  The MRF station allows recycling of different kind of materials. In particular, increasing the amount of recycled and reprocessed textile and paper would significantly improve the environmental performance. This is because the burdens associated with the average production of textile are considerably high.

 The results show that substitution of primary aggregates with the incineration bottom ash may have a significant impact in terms of environmental protection – especially in terms of GHG. Moreover primary aggregates substitution shows a saving seventy times greater than secondary aggregates.  The scenarios including hydrogenation (VCC) result in the highest savings in terms of EP. This is mainly due to the avoided impact for the naphtha production.  In terms of GWP, including an MRF provides the best solution for the treatment of PSW, when the substitution factor is 1; this is because of the avoided emissions associated with the production of virgin plastic. If the substitution factor is 0.5 or less, the LTP becomes the best environmental technology. Hence, decision makers should strongly consider the potential of the market to absorb reprocessed products when defining future waste management strategies.

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However, improved waste management strategies would require implementing governmental incentives to overcome the current corporation tax imposed on energy from waste (EfW). In particular, corporation tax on pyrolysis and hydrogenation currently represents a serious obstacle preventing investors to view plastics as a profitable feedstock. If such incentives were implemented in the future, EfW technologies would be considered as both green and profitable technologies and they would be developed in the UK. Acknowledgments This research work was conducted with sponsorship from the Kuwait Institute for Scientific Research (KISR). The authors confirm no restrictions in using the GaBi software. References [1] L. Yassin, P. Lettieri, S.R.J. Simons, A. Germanà, Techno-economic performance of energy-from-waste fluidized bed combustion and gasification processes in the UK context, Chem. Eng. J. 146 (3) (2009) 315–327. [2] J. Parfitt, Analysis of Household Waste Composition and Factors Driving Waste Increases. Waste and Resources Action Programme (WRAP) for Strategy Unit, Government Cabinet Office, London, England, UK, 2002. [3] USEPA, Municipal Solid Waste in the United States: 2007 Facts and Figure, United States Environmental Protection Agency, 2008. [4] EUROSTAT, Municipal Waste Statistics, 2011. Available from: . [5] F. Cherubini, S. Bargigli, S. Ulgiati, Life cycle assessment (LCA) of waste management strategies: landfilling, sorting plant and incineration, Energy 34 (2009) 2116–2123. [6] Y. Kalogo, S. Habibi, H.L. MacLean, S.V. Joshi, Environmental implications of municipal solid waste-derived ethanol, Environ. Sci. Technol. 41 (2007) 35–41. [7] J.D. Murphy, E. McKeogh, Technical, economic and environmental analysis of energy production from municipal solid waste, Renew. Energy 29 (2004) 1043–1057. [8] D. Özeler, Ü. Yetisß, G.N. Demirer, Life cycle assessment of municipal solid waste management methods: Ankara case study, Environ. Int. 32 (2006) 405–411. [9] M.A. Bozorgirad, H. Zhang, K.R. Haapala, Environmental impact and cost assessment of incineration and ethanol production as municipal solid waste management strategies, Int. J. Life Cycle Assess. 18 (2013) 1502–1512. [10] S.N.M. Menikpura, S.H. Gheewala, S. Bonnet, Framework for life cycle sustainability assessment of municipal solid waste management systems with an application to a case study in Thailand, Waste Manage. Res. 30 (2012) 708–719. [11] P. Nuss, K.H. Gardner, S. Bringezu, Environmental implications and costs of municipal solid waste-derived ethylene, J. Ind. Ecol. 17 (6) (2013) 912–925. [12] R. Clift, A. Doig, G. Finnveden, The application of life cycle assessment to integrated waste management, Part 1 – methodology, Process Safety Environ. Protect. 78 (2000) 279–287. [13] International Standardization Organization (ISO-14040), Environmental Management: Life Cycle Assessment/Principles and Framework, 2006. [14] European Commission, Joint Research Centre, Institute of Environment and Sustainability, International Reference Life Cycle Data System Handbook, Publication Office of the European Union, Luxemburg, 2012.

[15] J.B. Guinée, M. Gorrèe, R. Heijungs, et al., Life Cycle Assessment an Operational Guide to the ISO Standards, Centre of Environmental Science, Leiden, 2001. [16] H.K. Stranddorf, L. Hoffmann, A. Schmidt, Impact Categories, Normalization and Weighting in LCA, vol. 98, Danish Ministry of Environment, Environmental Protection Agency, Environmental News, 2005. [17] PE International, Gabi Version 5 Software Built-in Database (DB) Version 5.56, 2012. Available from: . [18] G. Finnveden, Methodological aspects of life cycle assessment of integrated solid waste management systems, Res. Constr. Rec. 26 (1999) 173–187. [19] T. Ekvall, G. Finnveden, Allocation in ISO 14041 – a critical review, J. Clean. Prod. 9 (2001) 197–208. [20] R. Heijungs, J.B. Guinée, Allocation and ‘‘what-if’’ scenarios in life cycle assessment of waste management systems, Waste Manage. 27 (2007) 997– 1005. [21] O. Eriksson, G. Finnveden, T. Ekvall, A. Björklund, Life cycle assessment of fuels for district heating: a comparison of waste incineration, biomass and natural gas combustion, Energy Policy 35 (2007) 1346–1362. [22] BSI, Specification for the Assessment of the Life Cycle Greenhouse GasEmissions of Goods and Services (PAS 2050), 2011, British Standards Institution, London, 2011. [23] S. Evangelist, P. Lettieri, D. Borello, R. Clift, Life cycle assessment of energy from waste via anaerobic digestion: a UK case study, Waste Manage. 24 (1) (2013) 226–237. [24] A. Bernstad, J. la Cour Jansen, A life cycle approach to the management of household food waste – a Swedish full-scale case study, Waste Manage. 31 (8) (2011) 1879–1896. [25] F. Perugini, M.L. Mastellone, U. Arena, A life cycle assessment of mechanical and feedstock recycling options for management of plastic packaging wastes, Environ. Prog. 24 (2005) 137–154. [26] A. Laurent, J. Clavreul, A. Bernstad, et al., Review of LCA studies of solid waste management systems – Part II: methodological guidance for a better practice, Waste Manage. (2013). [27] SELCHP, 2010. Available at: , (accessed 19.01.12). [28] DEC – Department of Energy and Climate Change, Offshore Wind Expansion Biggest Ambition in the World, Press Notice, 2010. [29] ONS – Office for National Statistics, Environment. Household Waste Recycling Rate in England Rises to 35%, 2009. [30] ADAS UK Ltd., Energy Audit of the Kerbside Recycling Services, The London Borough of Camden, 2008. [31] WRAP, An Analysis of MSW MRF Capacity in UK, in: Business Growth Report, 2007. [32] Energy Efficiency. Best Practice in Housing Domestic Condensing Boilers – The Benefits and the Myths, 2003. [33] Waste Management World, Wet or Dry Separation: Management of Bottom Ash in, Europe, 2012. [34] T. Rand, J. Haukohl, U. Marxen, Municipal Solid Waste Incineration: Requirements for a Successful Project, in: World Bank Technical Report, 2000. [35] A. Korre, S. Durucan, Aggregates Industry Life Cycle Assessment Model; Modelling Tools and Case Studies, in: Final Report, EVA 025 Project. Prepared for WRAP, 2007. [36] Volvo Truck Co., Emissions from Volvo trucks, 2008, Reg no. 2064008003, date: 21/5/2008. [37] A. Tukker, H. de Groot, L. Simons, S. Wiegersma, Chemical Recycling of Plastic Waste: PVC and Other Resins, in: Final Report, European Commission, DG III, 1999. [38] E.A. Williams, P.T. Williams, The pyrolysis of individual plastics and plastic mixture in a fixed bed reactor, J. Chem. Technol. Biotechnol. 70 (1997) 9–20. [39] M.A. Reuter, C. Hudson, A. van Schaik, K. Heiskanen, C. Meskers, C. Hagelüken. UNEP – Metal Recycling: Opportunities, Limits, Infrastructure, in: A Report of the Working Group on the Global Metal Flows to the International Resource Panel, 2013.