Post on 20-Nov-2018
SUPPLEMENTARY DATA
Life cycle assessment of conventional and advanced two-stage
energy-from-waste technologies for methane production
C. Tagliaferri1,2, S. Evangelisti1, R. Clift3, P. Lettieri1*
C. Chapman2, Richard Taylor2
1Department of Chemical Engineering, University College London, Torrington Place London WC1E
7JE, UK.
2Advanced Plasma Power (APP), Unit B2, Marston Gate, South Marston Business Park, Swindon,
SN3 4DE, UK.
3Centre for Environmental Strategy, The University of Surrey, Guildford, Surrey, GU2 7XH, UK
*Corresponding author: Email: p.lettieri@ucl.ac.uk; Phone: +44 (0)20 7679 7867
S.1. Life Cycle Assessment Methodology
The Global Warming Potential (GWP) characterises and calculates the impact of greenhouse gases
based on the extent to which these gases enhance radiative forcing. GWP values for specific gases,
developed by the Intergovernmental Panel on Climate Change (IPCC), express the cumulative
radiative forcing over a given time period following emission in terms of the quantity of carbon
dioxide giving the same effect (IPCC, 2006). Following common convention, for example in the
Kyoto Protocol, the 100-year values have been used here. Carbon dioxide from biogenic carbon is
sometimes excluded from the comparison in the GWP (Christensen et al., 2009) because it forms part
of the renewable carbon cycle, theoretically removed from the atmosphere in succeeding products.
However, in this study carbon dioxide emissions from biogenic carbon are included in the estimates
for the Global Warming Potential (GWP) because the assessment is based on existing waste streams
with defined carbon content so that the production of the materials in the waste does not enter the
analysis. Therefore the total carbon content of the waste is considered, with no distinction between
biogenic and non-biogenic carbon.
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The Acidification Potential (AP) quantifies the impact of acid substances and their precursors such as
SO2, NOx and HCl. Rain, fog and snow trap the atmospheric pollutants and lead to environmental
damage such as fish mortality, leaching of toxic metals from soil and rocks, and damage to forests and
to buildings and monuments. Abiotic Depletion (ADP) addresses the problem of the diminishing pool
of resources, focussing on the depletion of non-living resources such as iron ore, crude oil, etc. When
considering fossil energy resources, the measurement unit of abiotic depletion is MJ; whereas when
the depletion of element is quantified, the measurement unit of abiotic depletion is kg of Sb eq. The
Eutrophication Potential (EP) includes all pollutants that promote microbiological growth leading to
oxygen consumption, such as “algal blooms”. Nitrogen and phosphorus are the two main nutrients
implicated in eutrophication: they can cause shifts of species composition and increased biological
productivity. The Ozone Layer Depletion Potential (ODP) quantifies the thinning of the stratospheric
ozone. Chlorinated and bromated substances increase the rate of ozone destruction. Finally, the Fresh
Water Aquatic Ecotoxicity Potential (FAETP) assess the toxic effects of polluting compounds to
water life based on both the inherent toxicity of a compound and the potential human exposure
(Guinée, 2002).
In this study, the so called ‘zero burden approach’ is used (Buttol et al., 2007; Coleman et al., 2003):
the waste life- cycle starts from the moment when the material becomes waste, through the treatment
processes until the material ceases to be waste and becomes an emission into air, soil or water, inert
material in a landfill, or a useful product. Whether or not and how considering the biogenic carbon has
been a much debated subject in literature (Ballinger et al., 2009; Biobased Products Working Group,
2010; Cherubini et al., 2011; CHERUBINI et al., 2011; Christensen et al., 2009; Levasseur et al.,
2013; McDougall et al., 2001; Schmidt et al., 2007). Some authors (Cherubini et al., 2011;
CHERUBINI et al., 2011; Christensen et al., 2009; Levasseur et al., 2013; McDougall et al., 2001;
Schmidt et al., 2007) do not consider the impact of the emitted biogenic CO2 because it forms part of
the renewable carbon cycle. On the other hand, Finnveden (Finnveden et al., 2000) reports that it
might be inappropriate to exclude the biogenic carbon from the GWP in the case processes dealing
with biogenic carbon in different ways are compared. In this study carbon dioxide emissions from
biogenic carbon are included in the estimates for GWP, because following the zero burden approach,
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our boundary stars when the waste becomes waste and we do not account for its previous life-cycle.
Furthermore, in a comparative analysis based on a defined waste stream amounts the biogenic carbon
would count the same in all scenarios and therefore it would not add any more information to the
results.
S.2. Life Cycle Inventory
S.2.1. Process descriptions and Life Cycle assessment assumptions
S.2.1.1. System expansion
Ferrous material is assumed to be substituted at a 1 to 0.51 rate and the recovered material is assumed
to be recycled by electric furnace processing, as reported by the Worldsteel LCA Methodology report
(World steel association, 2011). Non-ferrous metal is assumed to be substituted at a 1 to 0.6 rate. The
recovered aluminium is assumed to be recycled by clean scrap melting and casting, as reported in the
Environmental profile report for the Aluminium Industry (European Aluminium Association, 2013).
A sensitivity analysis on the substitution ratio of the metals recovered has been performed but the
variation in the results was negligible and the results have not been reported.
S.2.1.2. Transport
For scenarios 1-2 MSW is assumed to be transported from transfer station to the processing plants (50
km distance) but the transport of the mechanically separated fraction to landfill/incineration is not
considered. The environmental burden due to the use of truck has been allocated to the direct burden
of the mechanical treatement plant and the diesel production has been allocated to the indirect burden
of this plant.
For scenario 3-4, source separated biodegradable waste is assumed to be transported from kerbside to
the AD plant (50 km distance) and the residual waste is assumed to be transported from transfer
station to incineration/landfill (50 km distance). The environmental burdens due to the use of truck
and production of diesel have been allocated to the section of incineration/landfill for the 75 %
(amount of residual waste) and for the 35% (amount of source separated waste) to the digester and
pre-treatment section of the anaerobic digestion plant.
For scenarios 5 MSW is assumed to be transported from transfer station to the advacend thermal
treatment process (50 km distance).
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All waste is assumed to be transported in an EURO 4-22t payload truck. The burden due to transport
(including the diesel production) is always less than 5% of the total environmental burden of the
processes, therefore, no sensitivity analysis has been performed on this parameter.
S.2.1.3. Incineration
Waste incineration is modelled according to average data for UK waste-to-energy plants taken from
the database of GaBi 5.0 software (Thinkstep, 2015). Two different incineration models are used,
respectively with wet and dry flue gas treatment (FGT). Different NOx-removal technologies are used
to represent the application of different FGT systems in Europe; the data from GaBi represent
averages over a number of European incinerators. The system includes generation of steam to produce
electricity and heat.
S.2.1.4. Landfill
The inventory data for landfilling with electricity recovery were based on the GaBi database
(Thinkstep, 2015). The data set represent a typical municipal waste landfill with surface and basic
sealing, meeting European limits for emissions. The site operations include landfill gas treatment,
leachate treatment, sludge treatment and deposition. Part of the landfill gas is assumed to be flared
(22%), part of it to be used for electricity production (28%) and the rest emitted to the environment
(50%). All manufacturing processes of the sealing materials, as well energy requirements for the site,
were included within the system.
S.2.2. Advanced thermal treatment: dual stage gasification and plasma process
Avoided burdens are allocated to the production of Plasmarok as it can be used as substitute of
aggregate materials (Korre and Durucan, 2009). We assume that Plasmarok substitutes the production
of primary aggregates from crushed rock as this is the most important source of primary aggregates in
England (Mankelow et al., 2011). We assume that the oxygen supplied to the process is produced
using the technology of cryogenic separation of air. Average UK data are applied (Thinkstep, 2015).
The inventory for the activated carbon used to remove the APC residue is reported in Noijuntira et al.
(Noijuntira and Kittisupakorn, 2010).
A wet scrubbing system is used in the dual stage advanced process to further cool and clean the
syngas from acid and alkali compounds. As the physical and chemical composition of the liquid
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effluents of this process do not exceed the limits reported in the WID directive (European Parliament,
2000), we assume that water effluents are treated in standard waste water treatment plants (Thinkstep,
2015).
The inventory and the environmental burden of all chemicals used in the process (such as nitrogen,
sodium hypochlorite, urea etc.) are reported in GaBi 6 LCA software (Thinkstep, 2015) and in
Ecoinvent (Swiss Centre for Life Cycle Inventories, 2014). The production of Bio-SNG is considered
according to Table 1.
S.2.3. Anaerobic Digestion of centrally separated waste
Six operations are identified in the AD process: i) pre-treatment; ii) anaerobic digestion; iii) water and
acid compounds removal; iv) upgrading of the biogas; v) disposal of digestate to incineration. The
characteristics of each section and the assumptions constituting the LCA models and the inventory
data are specified below.
Pre-treatment and anaerobic digestion phase (i and ii). After the mechanical separation of the
MSW, the biodegradable centrally separated waste (the composition of the biodegradable part
is reported in literature (Banks et al., 2011)) enters the pre-treatment section where the waste
undertakes maceration and hygienization. Further details of this phase (i.e. pre-treatment
steps, vessel dimensions, etc.) are reported in literature (Banks et al., 2011; Berglund and
Börjesson, 2006; Evangelisti et al., 2014). The AD phase is assumed to be using a continuous,
single-stage, mixed tank mesophilic reactor operating at a temperature of 35 ˚C in wet
condition given its broad application (Berglund and Börjesson, 2006; Evangelisti et al., 2014;
Monnet, 2003; Severn Wye Energy Agency, 2009). The majority of AD plants from centrally
separated biodegradable waste actually in use operate under this condition. Monson et al.
(Monson et al., 2007) reported that 73% of the anaerobic digesters treating centrally separated
MSW analyzed in their report use a wet system as well as 90% operate in mesophilic
temperature range. The yield of biogas produced during this phase decreases if the organic
fraction of centrally separated municipal solid waste is used instead of the source separated
fraction. This is due to the higher content of plastic and impurity of the organic matter
supplied to the digester and therefore a lower composition of volatile solid on which the yield
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of methane depends (Laraia, 2002). The yield of biogas is assumed to be 0.079 Nm3 per kg of
the centrally separated organic fraction (Monson et al., 2007). The model accounts also for
methane losses occurring in the digester (Berglund and Börjesson, 2006; Boldrin et al., 2011;
Dalemo et al., 1997; Fruergaard and Astrup, 2011), which are assumed to be 3 %.
Cleaning (water and H2S removal), up-grading and off gas flaring phase (iii and iv). The gas
has to be cleaned and upgraded before grid injection. The cleaning of the biogas includes
removal of H2S and water that can cause damages to the subsequent upgrading unit and to the
grid pipes. The most common method for H2S removal from the crude biogas is through the
reaction of H2S with metal oxides (Hagen and Polman, 2001; Persson, 2003). In this study
H2S is assumed to be removed in a desulphuriser unit with a catalytic bed of ZNO, which is
placed at the digester plant, where the biogas is produced. Water is assumed to be adsorbed
on silica gel in the upgrading unit in the pre-treatment phase. To achieve the strict regulation
limits set in the GMRS (GSMR, 1996), the raw methane is assumed to be upgraded in a
pressure swing adsorption system. Petersson et al. (Petersson, A. Wellinger, 2009) reported
that the PSA electricity consumption including gas compression to 7 barg is 0.25 Kwh/Nm 3 of
raw biogas. Persson et al. (Persson et al., 2006) assumed instead that the electricity
requirement for PSA was 0.5-0.6 Kwh/m3 of upgraded gas, not accounting for high pressure
compression (those are the figures reported by the owners), whereas Persson (Persson, 2003)
reported the same values as Persson et al. (Persson et al., 2006) for electricity requirements
not accounting for compression and specified that the electricity needs to be increased to the
values of 0.8-0.88 a Kwh/m3 of upgraded gas if compression is considered as well. The latter
value has been used in the LCA model. This value is in line with the value reported in the
Ecoinvent database (Swiss Centre for Life Cycle Inventories, 2014) for raw gas upgrading
(CH methane, 96% from biogas purification). The model accounts also for a 3% methane
losses (the amount of methane not recovered from the raw biogas, hence the methane content
of the off gas of the PSA system) occurring in the upgrading system (Patterson et al., 2011;
Persson et al., 2006; Petersson, A. Wellinger, 2009). The PSA model does not include the
production of a combustile stream for electricity production, as in literature this layout is not
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reported to be used for AD systems. The off gas of the PSA system is assumed to be flared
before emission to environment.
Digestate use (v). When anaerobic digestion is used to treat centrally separated organic wastes
the low quality of the digestate prevents from the use as fertiliser and it has to be disposed of
either by thermal treatment or landfill (Department of the Environment Heritage and Local
Government, 2006; Monson et al., 2007). In this study, the digestate produced in scenarios 1-
2 is assumed to be incinerated according to Ecoinvent database (Swiss Centre for Life Cycle
Inventories, 2014).
A sensitivity analysis on main parameters regarding the AD process has not been reported here as
significant results have already been published in Evangelisti et al. (Evangelisti et al., 2014).
S.2.4. Anaerobic Digestion of source separated waste and additional disposal
The whole digestate is assumed to be separated in liquor and fibre as standard practice (Wrap, 2012)
and the separation method is assumed to be physical (Wrap, 2010). The liquor is normally separated
using a separator or centrifuge to remove coarse fibres. The fibres represent 20 % of the total digestate
whereas the liquor is usually the 80% in weight of the total digestate (Wrap, 2012). To calculate the
nutrient of the liquor after dewatering the values for nutrient partition between liquid and solid phases
as reported in Lukehurst et al. (Lukehurst et al., 2010) are used. The electricity requirements for
dewatering are taken from Wiliams et al. (Wiliams and Esteves, 2011).
We assume that the liquor separated from the whole digestate in the dewatering section is used as
fertilizer whereas the fibres are sent to incineration as inert material (as reported in Wrap (Wrap,
2012)). The LCA model accounts for the burden associated with the use of liquor as fertilizer.
Organic fertilizers coming for example from the anaerobic digestion can be used to improve the
nutrient content of soils and therefore avoiding the use of chemical fertilizers (the use in agriculture of
fertilisers with high available nitrogen content, i.e. digestate, is anyway, affected by restrictions
according to the Nitrogen Vulnerable Zones as reported by the European Parliament (European
Parliament, 1991)).
We assume that the nutrients are not lost during the anaerobic digestion phase; thus all the nutrients of
the bio-degradable waste (N, K and P) remain in the whole digestate (Evangelisti et al., 2014; Møller
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et al., 2009). The nutrient content of the fertilizers is calculated based on the amounts of N, P 2O5 and
K2O for N, P and K, respectively (Defra, 2010), 2010). The distinction between readily availability
and crop availability of nutrient (Defra, 2012) is used to calculate the avoided chemical fertilizers and
the emissions due to fertilizer spreading.
The amount of readily available nutrients assumed in this study is reported in Wrap (Wrap, 2011)
(80% of the N content of the digestate is readily available and 100% of K2O and P2O5). Defra (Defra,
2010) reported the chemical fertilizers usually employed for N, K and P. The N readily available
content of the digestate is assumed to substitute the chemical fertilizer of ammonium sulphate, the
K2O readily available content of the digestate is assumed to substitute the chemical fertilizer of
potassium chloride and the readily available content of P2O5 of the digestate is assumed to substitute
the chemical fertilizer of the superphosphate. The results have been calculated equalling the amount
of nutrients readily available in the digestate to the amount of chemical fertilizer needed. (i.e. 1 kg of
N readily available in the digestate equal 1 kg of the chemical fertilizer NH4(SO2) substituted).
In the LCA model we have also accounted for the emission due to the organic fertilizers when those
are on the soil.
During the application of both chemical and organic fertilizers part of the nutrients is dispersed into
environment and is not crop available. This means that some amount of the readily available nutrient
might be lost as air emission (run off or evaporation), water leakage (leaching) or might not be readily
absorbed by the plants because chemically bound in a form of not easy uptake from plants. The
amount of the nutrients really uptaken by the crops is defined as the nutrient crop availability. Bruun
et al. (Bruun et al., 2006) reported the emission coefficients due to the use of the digestate as fertilizer.
In particular those emissions represented the difference between normal agricultural practice only
using inorganic fertilizers and use of digestate supplemented with inorganic fertilizers, according to
the Danish legislation. Those coefficients have already been used in some recent works (Boldrin et al.,
2011; Evangelisti et al., 2014; Møller et al., 2009). Nitrogen emissions from organic fertilizers are
higher than those of chemical fertilizers for two reasons: i) chemical fertilizers are given to the plants
when and where they need them and this reduces N evaporation; ii) not all the readily available N is
absorbed by the plants. The emissions associated with the spreading of fertilizers, either chemical or
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organic, are highly variable and depend strongly on the soil and weather conditions, spreading
practice and crop practice. Therefore, it is possible that the emission coefficients reported in Bruun et
al. (Bruun et al., 2006) do not exactly mirror the UK situation but as far as the author’s knowledge
these data are the only available. Wrap (Wrap, 2012) reported that they are undertaking studies on the
emission coefficients for digestate spreading applied to UK situation but no data have already been
released. New data on evaporation, leaching and loss of fertilisers might influence the results.
In the LCA model we also account for the part of the carbon of the feed waste not released as biogas.
Part of the carbon content of the liquor used as fertilizer is sequestered in the soil and not released to
the atmosphere as CO2 during the timeframe considered (Møller et al., 2009). This has been accounted
as an actual removal of CO2 from atmosphere and therefore as a negative contribution to the global
warming.
S.3.UK future energy scenarios: marginal electricity and natural gas
The marginal energy supply (in particular electricity supply), is reported to strongly affect the results
of an LCA analysis (Kløverpris et al., 2008; Moora and Lahtvee, 2009) and hence, a study of the
environmental burden of the technologies analysed have been performed according to different
marginal energy technologies.
The UK energy mixes (electricity mix and natural gas mix) are evolving towards renewables. National
Grid (National Grid, 2014) has foreseen possible future energy scenarios for the UK and has
undertaken a detailed analysis to 2035 for each scenario. Four scenarios have been identified: i) gone
green; ii) slow progression; iii) no progression; iv) low carbon life. Those scenarios are used by UK
National Grid ‘as a reference point for a range of modelling activities including network analysis to
identify potential gas and electricity network investment requirements in the future’. A brief
description is reported here.
i) ‘Gone Green is a future where more money is available, with strong policy and regulation
and new environmental targets. The economy is growing, and environmental
sustainability is not restrained by financial limitations as more money is available at both
an investment level for energy infrastructure and at a domestic level via disposable
income’.
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ii) ‘Slow Progression is a future where less money is available compared to Gone Green, but
with similar strong policy and regulation and new targets. Economic recovery is slower in
this scenario than in Gone Green and Low Carbon Life, resulting in investor uncertainty.
Financial constraints lead to difficult political decisions’.
iii) ‘No Progression is a future where there is less money available and less emphasis on
sustainability. There is slower economic recovery in this scenario, meaning less money is
available at both a government and consumer level. Government policy and regulation
remains the same as today, and no new targets are introduced’.
iv) ‘Low Carbon Life is a future where more money is available and there is less emphasis on
sustainability. There is higher economic growth. Society has more disposable income
which results in higher uptake of electric vehicles, and more renewable generation at a
local level. Government policy is focused on the long term’.
Based on those four possibilities, National Grid (National Grid, 2014) reports the mix of technologies
used in the UK to produce electricity and natural gas each year till 2035.
Table S1 reports some of the data published by National Grid (National Grid, 2014) regarding the
future energy mix. As an example, only few years (13-14, 20-21, 25-26 and 35-36) are reported for
the gone green scenario.
National Grid (National Grid, 2014) infers that the imports generic regarding the natural gas (see
Table S1) can be any mixture of the LNG and continental gas, ranging from all LNG to all continental
gas. In the LCA model, it is assumed that the generic imports of natural gas are completely supplied
by LNG as this is the most probable option.
The marginal technologies for the foreseen energy scenarios are modelled using Gabi database
(Thinkstep, 2015).
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2526
UK electricity grid mix[%] 2013/14 2020/21 2025/26 2035/36
Nuclear 0.181992 0.1750140.11747
7 0.163797
Coal 0.38821 0.0519680.03726
2 0.004331Gas 0.214257 0.313289 0.28323 0.127476
CCS Coal 0 00.00825
4 0.065028CCS Gas 0 0 0 0.073832Imports 0.05169 0.107407 0.03405 0.008322
Wind 0.088052 0.224610.37546
4 0.398648
Solar 0.005019 0.0172080.02462
3 0.036622Biomass 0.031754 0.065839 0.06928 0.056445
Other Renewables 0.018442 0.0226610.02466
9 0.030114
Hydro/Pumped Storage/Marine 0.020399 0.0218430.02553
7 0.035256
Oil/Others 0.000187 0.0001610.00015
3 0.000129Natural gas UK mix
[%] UKCS 2013 2020 2025 2035
Norway 0.421694 0.4570970.35211
4 0.156376
Shale 0.354716 0.320640.29230
4 0.286783
Biomethane 0 00.05792
8 0.247799CBM 9.88E-05 0.007522 0.01815 0.052077
Continent 0 0.0081370.01181
3 0.013523
LNG 0.106354 0.0323120.01924
3 0.007681
Import Generic 0.117137 0.0394530.03979
1 0.045552
Demand 0 0.1348390.20865
7 0.190209
Table S1: Foreseen energy mixes in the UK according to National Grid (National Grid, 2014).
Normalization factor EU25+3, year 2000. CML, IPCC, ReCiPe (person equivalents) CML2001 - Apr. 2013, Abiotic Depletion (ADP elements) 6040000CML2001 - Apr. 2013, Abiotic Depletion (ADP fossil) 3.51E+13CML2001 - Apr. 2013, Acidification Potential (AP) 1.68E+10CML2001 - Apr. 2013, Eutrophication Potential (EP) 1.85E+10CML2001 - Apr. 2013, Freshwater Aquatic Ecotoxicity Pot. (FAETP inf.) 2.09E+11CML2001 - Apr. 2013, Global Warming Potential (GWP 100 years) 5.21E+12
14
407
408409
2728
CML2001 - Apr. 2013, Global Warming Potential, excl biogenic carbon (GWP 100 years) 5.21E+12CML2001 - Apr. 2013, Human Toxicity Potential (HTP inf.) 5E+11CML2001 - Apr. 2013, Marine Aquatic Ecotoxicity Pot. (MAETP inf.) 4.45E+13CML2001 - Apr. 2013, Ozone Layer Depletion Potential (ODP, steady state) 10200000CML2001 - Apr. 2013, Photochem. Ozone Creation Potential (POCP) 1.73E+09CML2001 - Apr. 2013, Terrestric Ecotoxicity Potential (TETP inf.) 1.16E+11
Table S2: Impact values used for normalisation. The normalisation is done based on CML, IPCC, ReCiPe (region equivalents), EU25+3, year 2000 (Thinkstep, 2015).
Normalised results
S.1 S.2 S.3 S.4 S.5
Abiotic Depletion (ADP elements) [kg Sb-Equiv.] -7.00449E-14 4.92193E-15 -4.4064E-13 -1.25636E-143.45E-
14
Abiotic Depletion -6.66487E-14 -2.714E-14 -6.5113E-14 -1.34569E-14-9.1E-
14
Acidification Potential -3.3403E-14 -1.6067E-14 -9.8073E-15 1.31538E-141.95E-
14
Eutrophication Potential 1.98249E-14 3.4767E-14 2.48966E-14 4.02271E-144.21E-
15
Freshwater Aquatic Ecotoxicity Pot. 1.09476E-13 1.10582E-13 -1.0807E-14 -9.49733E-152.27E-
14
Global Warming Potential 9.32546E-14 1.71507E-13 1.04598E-13 1.87681E-131.38E-
13
Global Warming Potential, excl biogenic carbon -3.53443E-14 6.27702E-14 -2.4001E-14 8.81235E-149.23E-
15
Human Toxicity Potential -8.12672E-14 -7.6677E-14 -8.4066E-15 -4.02202E-15-6.3E-
14
Marine Aquatic Ecotoxicity Pot. -2.68149E-12 -2.9049E-12 5.1791E-13 2.16172E-13-2.7E-
12
Ozone Layer Depletion Potential 2.7674E-16 1.7921E-16 2.95602E-16 4.02731E-179.02E-
17
Photochem. Ozone Creation Potential -2.4393E-14 5.28661E-14 -1.6495E-14 7.07699E-14 -1E-14
Terrestric Ecotoxicity Potential -7.73728E-16 3.2548E-15 -2.1158E-16 3.73001E-151.32E-
15
Table S3: Normalised results.
2013-2
014
2016-2
017
2019-2
020
2022-2
023
2025-2
026
2028-2
029
2031-2
032
2034-2
035
0.0E+01.0E-12.0E-13.0E-14.0E-15.0E-16.0E-17.0E-18.0E-19.0E-11.0E+0
Slow progression
S.1 S.5 S.3
Glo
bal W
arm
ing
Pote
ntia
l [kg
CO
2-E
quiv
.]
2013-2
014
2016-2
017
2019-2
020
2022-2
023
2025-2
026
2028-2
029
2031-2
032
2034-2
035
0.0E+01.0E-12.0E-13.0E-14.0E-15.0E-16.0E-17.0E-18.0E-19.0E-11.0E+0
Low carbon life
S.1 S.5 S.3
Glo
bal W
arm
ing
Pote
ntia
l [kg
CO
2-E
quiv
.]
15
410411
412
413
414
2930
Figure S.1. Fossil GWPs of S.1, S.3 and S.5 for future electricity and natural gas UK mix according to
the a) slow progression scenario; b) Low carbon life scenario. Results are reported for 1 MJ of
methane as functional unit.
20
13-201
4
2016-2
017
2019-2
020
2022-2
023
2025-2
026
2028-2
029
2031-2
032
2034-2
035
0.0E+01.0E-12.0E-13.0E-14.0E-15.0E-16.0E-17.0E-18.0E-1
Slow progression
S.1 S.5 S.3
Glo
bal W
arm
ing
Pote
ntia
l [kg
CO
2-E
quiv
.]
20
13-201
4
2016-2
017
2019-2
020
2022-2
023
2025-2
026
2028-2
029
2031-2
032
2034-2
035
0.0E+01.0E-12.0E-13.0E-14.0E-15.0E-16.0E-17.0E-18.0E-1
Low carbon life
S.1 S.5 S.3G
loba
l War
min
g Po
tent
ial [
kg C
O2-
Equ
iv.]
Figure S.2. GWPs of S.1, S.3 and S.5 for future electricity and natural gas UK mix according to the a)
slow progression scenario; b) Low carbon life scenario. Results are reported for 1 kg of MSW as
functional unit.
16
415
416
417
418
419
420
421
422
423
3132