Comparative lca of two approaches with different emphasis on energy or material recovery for a municipal solid waste management system in gipuzkoa

Comparative LCA of two approaches with different emphasis on energy
or material recovery for a Municipal Solid Waste Management System in
Gipuzkoa
DOI:10.1016/j.rser.2015.06.021
G. Buenoa
*, I. Latasab
, P.J. Lozanob
a
Department of Electronics Engineering, Faculty of Engineering, University of the Basque
Country UPV/EHU, Alameda Urquijo s/n, 48013 Bilbao, Spain.
b
Department of Geography, Prehistory and Archaeology, Faculty of Arts, University of the
Basque Country UPV/EHU, Tomás y Valiente s/n, 01006 Vitoria-Gasteiz, Spain.
*Corresponding author. gorka.bueno@ehu.eus T: +34 94 601 41 34; F: +34 94 601 42 59
Abstract
Two alternative approaches for an integrated municipal solid waste management system
(MSW-MS) have been confronted in the province of Gipuzkoa, in the north of Spain, during
the last decade. While one of them prioritizes energy recovery from mixed residual waste
in an incineration plant, the other approach gives precedence to material recovery of
separately collected waste. Which system would present a lower environmental impact
and be more desirable from a sustainability perspective? Answering this question is
hindered by the fact that recovered energy and materials are not directly comparable or
directly substitutable with each other.
Based on the powerful framework provided by life cycle assessment (LCA) methodology,
this work performs a comparative LCA of overall environmental impacts of these two
alternative approaches, showing that comparisons of alternative systems in terms of direct
energy recovery or direct material recovery should be avoided in favor of other indicators
already proposed in the LCA framework, such as the Cumulative Energy Demand category
from Ecoinvent, or the Global Warming Potential and the Abiotic Resources Depletion
categories from the CML 2001 method.
Applying the LCA framework, this work shows that when a high share of waste is collected
separately, and processes assumed in the background system are adequately
characterized, especially the production of the electricity mix, then prioritizing material
recovery provides better results even in environmental categories tightly related to fossil
energy consumption, such as the global warming potential impact category.
Keywords
Life cycle assessment (LCA); Municipal solid waste (MSW); Material recovery; Energy
recovery; Waste management.
Abbreviations
acid Acidification impact category from CML 2001 method
ard Abiotic Resource Depletion impact category from CML 2001 method
Page 1

eutro Eutrophication impact category from CML 2001 method
GHG Greenhouse Gas
gw Global Warming impact category from CML 2001 method
htox Human Toxicity impact category from CML 2001 method
ILCD International Reference Life Cycle Data System
ISO International Organization for Standardization
LCA Life Cycle Assessment
LCA-IWM LCA Tools for the Development of Integrated Waste Management
MBP mechanical biological pre-treatment
MSW Municipal Solid Waste
MSW-MS Municipal Solid Waste Management Systems
P Product
PE Primary energy demand
ph-tox Photo-oxidant Formation impact category from CML 2001 method
RM Resource material demand
SC Separate Collection
WFD Waste Framework Directive
WP waste prevention
WtE Waste-to-energy, incineration plant with energy recovery
1. Introduction
The aim of integrated Municipal Solid Waste Management Systems (MSW-MS) is to give
an adequate treatment to collected waste with a minimum environmental impact under
affordable costs. These systems comprise all the treatment and processing steps
underwent by collected fractions of municipal solid waste (MSW) generated in a specific
area, from temporary storage and collection through final disposal of secondary fluxes
generated in processing plants. In order to improve sustainability and minimize impacts,
some waste treatments—such as incineration or anaerobic digestion—aim at recovering
energy from waste, while others are focused on preparing the waste for material recovery.
In fact, integrated MSW-MS normally combine different kinds of material and energy
recovery.
1.1 Waste management strategies in Gipuzkoa
Local administrations in Spain have been redefining their municipal waste-management
systems for more than a decade. On the one hand, they are obliged to comply with
European Directives regarding minimum recovery and recycling rates for packaging
wastes and closure of landfills; on the other hand, many administrations have to face up to
the saturation of landfill sites. This is the case, for example, in the Basque province of
Gipuzkoa, where 64% of all MSW generated in 2012 was derived to landfills. This figure,
actually, is similar to the values registered in nearby provinces and regions in Spain, as
can be checked in table 1, which shows the percentages of MSW derived to final
treatments that year in the three Basque provinces and Spain. There, treatment of MSW
has been mainly based in landfilling and to a much lesser degree in energy recovery;
material recovery, on the other hand, has remained below 40% for many years [1-4].
Page 2

Table 1. Final treatments of MSW in 2012 in Gipuzkoa and nearby regions (other Basque
Provinces and Spain).
Final treatment Gipuzkoa Bizkaia Araba Spain
Landfilling 64% 28% 63% 63%
Energy recovery 0% 36% 2% 10%
Material recycling 29% 36% 34% 17%
Composting 7% <1% 1% 10%
With a population of 731 thousand inhabitants in 2013, Gipuzkoa is administratively
divided into eight municipality commonwealths. Historically, municipality commonwealths
are the administrative bodies that have been in charge of the collection and treatment of
municipal waste, especially through its disposal to controlled landfills. Figure 1 shows the
trend of MSW generation in Gipuzkoa between 2000 and 2013, altogether with planning
objectives established by the provincial administration in 2008 (DdP-2008 Strategy, for
year 2016 [5]) and in 2012 (EDDdP-2012 revision Strategy [1], for 2016 and 2020).
Figure 1. Historical evolution of the MSW flux in Gipuzkoa, and planning objectives
established by the DdP-2008 Strategy (for year 2016) and those established by the
EDDdP-2012 revision Strategy (for years 2016 and 2020). Broken lines are eye guides.
Source: [1,5,6]
MSW generation in Gipuzkoa increased since 2000 until 2006, when a peak of 411
thousand metric tons was generated. During that period around 80% of the MSW was
mixed residual wastes derived to landfills, as most of the waste was not separately
collected—from 15.3% in 2000 up to 25.5% in 2006. In order to reduce environmental
Page 3

impacts related to such a big waste flux being derived to landfill sites, during those years
the provincial administration made a strong commitment to energy recovery of the mixed
residual waste. This commitment was materialized in the DdP-2008 Strategy, approved in
the beginning of 2008. This planning projected a progressive increase in waste generation
and recycling until 2016. According to it, in that year 57% of the generated waste would be
separately collected and 53.3% could be recycled [5]. Most of the resting mixed residual
waste (213 thousand metric tons, annually) would be incinerated with energy recovery.
This strategy would have required the installation of at least one new incineration plant in
Gipuzkoa, although up to three new plants were eventually considered [5,7]. It must be
emphasized that the DdP-2008 Strategy was established previous to the approval of the
European Waste Framework Directive (WFD), which sets a minimum target of 50% for re-
use and recycling of MSW by 2020 [8]. That target could be tightly achieved inside the
DdP-2008 Strategy by 2016, but some serious problems arise when the evolution of MSW
generation in Gipuzkoa after 2006 is considered.
Since 2007 the MSW flux generated in Gipuzkoa has diminished steadily, as can be
checked in figure 1. This reduction in waste generation seems to be due, partially at least,
to a social context more sensible every year with recycling, re-use and environmental
impacts derived from landfilling, as the decline started before the economy got into
recession by the end of 2008. At that moment, MSW generation in Gipuzkoa had already
diminished by 15% when compared to 2006 levels. By 2013 the reduction was 22%, and
35% less than the forecast for 2016.
After the approval of the DdP-2008 Strategy and the WFD in 2008, some municipalities
boosted an alternative approach in order to avoid the installation of any new incineration
facility in the province. This alternative strategy was mainly based on a strong commitment
to separate collection of household wastes, which would allow for the separate recovery of
each material fraction, and thus minimizing the need for final disposal to landfills and
incineration. A change in the provincial government in 2011 allowed a further
implementation of this alternative approach. The new provincial government revised the
DdP-2008 Strategy in 2012, which materialized in an updated waste management
planning for the period 2012-2016, the EDDdP-2012 revision Strategy [1]. This updated
planning took into account the new waste generation trend after the 2006 peak, and
reformulated separate collection and recycling targets for years 2016 and 2020, improving
the targets imposed by the WFD: by the end of the decade 76% of MSW generated in
Gipuzkoa would be separately collected, which could boost materials recycling well over
70%. In the new planning, by 2020 the residual fraction would be reduced down to 77
thousand metric tons annually, or 36% of the flux that in the previous planning was
supposed to be needed to feed the new incineration facility, 213 metric tons. Under these
circumstances of more ambitious recycling targets and less MSW generation the
economical viability of the incineration facility would be seriously jeopardized, as its
functioning would diverge too much from full capacity [9].
In the context of this socio-political debate—not exempt of understandable economic
conflicts, as waste management demands a significant part of every municipal budget,
even in times of economic turndown—, social agents and decision-makers from Gipuzkoa
have addressed our research group with questions such as the following, to be answered
from a technical and scientific point of view: Which kind of recovery has to be given
precedence in a waste-management system—energy or material recovery? Which is the
significance of separate collection in an integrated MSW-MS such as the one to be
implemented in Gipuzkoa?
Page 4

1.2 Objectives of the study
The framework necessary to answer those previous questions is already settled in the
WFD, which establishes, through its waste hierarchy, a legally binding priority order for
waste management in the EU [8]. Prevention and preparing for re-use rank at the top of
the hierarchy, followed by different kinds of material and energy recovery. This hierarchy is
not arbitrary, as the WFD states that potential deviations from it—and the choice among
alternatives at each hierarchy level—have to be justified by life cycle thinking of the overall
impacts. This is often achieved by the application of Life Cycle Assessment (LCA), which
is a preferred and standardized scientific approach for life cycle thinking. The basic
framework for LCA is provided by the ISO 14040 and 14044:2006 standards [10,11].
Handbooks are available for its application [12], along with an international reference guide
[13] and a guidance for its application in waste management [14], where a number of
models have been developed during the last two decades [15]. The use of these models
abounds in the literature, and they are especially suited for the assessment of integrated
MSW-MS that may combine energy and material recovery from waste. De Feo and
Malvano [16] use the WISARD LCA tool in selecting the best MSW-MS for the Campania
Region, in Southern Italy. Bovea et al. [17] make use of SimaPro7 for the assessment of
alternatives in the Spanish town of Castellón de la Plana. Pire et al. [18] carry out an LCA
for a future MSW-MS in the Setúbal peninsula, in the Portuguese region of Lisbon, using
the Umberto 5.5 software. Tunesi [19] uses the WRATE modeling tool for the assessment
of different energy recovery strategies in England. Slagstad and Brattebø [20] use
EASEWASTE to assess different alternatives for waste management in a new urban
settlement in the city of Trondheim, in central Norway. Song et al. [21] use SimaPro7 for
the assessment of environmental performance of MSW-MS in Macau, China. Bernstad
and la Cour Jansen [22] compare different alternatives for the integrated management of
household food waste in the area of Augustenborg, Southern Sweden, using the
EASEWASTE LCA-tool. Eriksson et al. [23] study different MSW-MS for the Swedish
municipalities of Uppsala, Stockholm and Älvdalen, using the ORWARE model. Merrild et
al. [24] assess recycling versus incineration in waste management systems in Denmark,
by modeling in EASYWASTE. Nadzirah Othman et al. [25] review six life cycle
assessments of integrated MSW-MS in Asian countries that combine both energy and
material recovery approaches.
The main objective in this work is to determine which integrated MSW-MS may cause in
the province of Gipuzkoa a lower environmental impact and be more desirable from a
sustainability perspective—either a management system that prioritizes energy recovery
from mixed residual waste in an incineration facility, or another one that gives precedence
to material recovery of separately collected waste. In order to compare these two
alternative approaches, this work carries out a comparative LCA of these two alternatives,
to be implemented in a generic municipality commonwealth. The modeling of this generic
municipality commonwealth is based on the present context of Gipuzkoa, and its detailed
characterization is performed in the following section, Materials and methods. We believe
that the quantitative assessment of environmental impact indicators in a generic
municipality commonwealth allows drawing some important qualitative conclusions that
may be valid not only for the whole province of Gipuzkoa, but also for other provinces or
regions with a similar socio-economic situation and waste treatment conditions in Spain,
as shown in table 1.
Some methodological choices may have important consequences when performing a
comparative LCA of alternative waste-management systems. Gentil et al. [15] reviewed the
Page 5

importance of technical assumptions related to the definition of the functional unit, system
boundaries, and energy and process modeling in LCA models, concluding that making
different choices may lead to contradictory results. Other important factors may also have
important effects when assessing the environmental impact of waste-management
systems, such as considering different waste prevention strategies, different collection
systems, or different spreading levels of separate collection. Regarding to waste
prevention (WP), Gentil et al. [26] evaluated several measures for municipal waste
management; Slagstad and Brattebø [20], on the other hand, quantified WP potential to
reduce household waste generation in circa 17% for a new urban settlement in Norway.
Other studies have centered on the influence of different collection systems, altogether
with different treatment options [17,20,27]. The spreading of separate collection is also
analyzed in some comparative LCA studies [16,28-32], but with quite different ranges
under consideration: while Buttol et al. [29] assumed very limited variations in separate
collection, Rigamonti et al. [30] considered a range from 35% up to 60%, Calabrò [31] from
15% up to 50%, Consonni et al. [32] from 35% to 65%, and De Feo and Malvano [16] from
35% up to 80%. But other studies do not consider any increase in separate collection, e.g.
Cimpan and Wenzel [33] and Belboom [34] when comparing different pretreatments of
residual waste, or Koci and Trecakova [35] when comparing different treatments of mixed
residual waste. Similarly, the possibility to increase separate collection is absent in other
studies that compare different technologies for incineration [36,37], that compare final
disposal to landfill versus incineration [38], different ways for energy recovery [39,40], or
that compare material versus energy recovery [24].
Taking all this into account, it is also an objective of this work to check the importance of
the spreading of separate collection of MSW on the overall environmental balance of
integrated MSW-MS, along with other factors such as the presence of waste prevention
strategies, and the adequate characterization of the electricity mix generation in the
background process.
This work also aims to demonstrate that the LCA methodology framework provides a set of
indicators, such as the Cumulative Energy Demand category from Ecoinvent, or the Global
Warming Potential and the Abiotic Resources Depletion categories from the CML 2001
method, that allow to assess and compare life-cycle material and energetic consumption in
systems of very different nature that involve energy fluxes and material resources not
directly comparable or directly substitutable with each other.
2. Materials and methods
2.1 Goal and scope definition
As this study is centered in the proper accounting of different environmental impacts when
comparing systems, the attributional modeling principle has been chosen for this
comparative LCA, and the system expansion/substitution approach has been considered
for solving multifunctionality (Situation C1 in [13]).
The comparative LCA is carried out with the LCA-IWM tool [41]. The assessment tool of
LCA-IWM allows comparing different scenarios, based on the LCA methodology,
considering all waste management steps, from temporary storage through final disposal of
secondary fluxes generated in previous treatments, such as recycling, incineration or
Page 6

composting. This tool was specially designed for planning and optimizing waste-
management systems in areas that still require much effort to be adjusted to the state-of-
the-art in Europe, as is the case in Southern European countries, and particularly in Spain.
The general diagrams of the two integrated MSW-MS modeled with the LCA-IWM
assessment tool in this work are shown superimposed in figure 2, with the corresponding
divergences between them in fluxes and processing steps.
Figure 2. Material flux diagram of the two integrated MSW-MS considered in this work.
Our model considers five different waste flows separately collected: biowaste, glass,
metals, plastics, and paper and cardboard, with the specific compositions assumed in the
LCA-IWM tool by default—for every parameter not specified from now on, LCA-IWM
default data should be assumed. The percentages of separately collected fractions are
specified in table 2, and resemble those of Gipuzkoa in 2011 [1]. A sixth primary flow
corresponds to the residual waste collected in mixed form, of which almost 70% is
biowaste [42]. One of the key differences between the two systems considered affects the
treatment of this residual flow. On one hand, in the system that prioritizes material
recovery, this mixed residual flow is transported to an aerobic mechanical biological pre-
Page 7

treatment (MBP), where the organic fraction is stabilized, the high caloric fraction is
recovered for its combustion in cement kilns, and the resulting secondary residual waste is
left ready for its safe disposal to landfill. On the other hand, in the alternative that
prioritizes energy recovery, the residual flow is directed to an incineration plant, and the
ashes and the slag there produced are also landfilled, as the Basque legislation does not
allow for its use as gravel for road construction or similar.
Table 2. Waste fractions considered in the functional unit.
Waste fraction System with incineration System with aerobic MBP
Mixed residual waste,
of which 70% is bioresidue
75% 25%
Separately collected waste,
of which:
25% 75%
Paper & Cardboard 24% 24%
Glass 11% 11%
Metals 5% 5%
Plastics 15% 15%
Biowaste 45% 45%
TOTAL 100% 100%
Historically, MSW management systems in Spain have been reliant on the disposal to
landfills of not separately collected mixed wastes. In 2006, as much as 80% of household
wastes in Gipuzkoa were collected this way [5]. In the nearby province of Bizkaia a similar
percentage was reached in 2013 [2]. In parallel, it is well known that small sized
incineration plants are seriously handicapped because of lower electric efficiencies due to
scale effects, higher specific consumption of auxiliaries, and more conservative design
conditions and less sophisticated configurations, as economic constraints are tighter in
them [43]. Incineration plants perform better if incoming waste fluxes are bigger. As they
normally recover energy from mixed wastes that cannot be recycled, administrations do
not find much incentive to broaden selective collection schemes that reduce incoming
waste fluxes to incineration plants and may jeopardize their viability. This is specially the
case in Gipuzkoa, where annual household waste generation barely exceeds 300.000
metric tons. On the contrary, systems that prioritize material recycling should always try to
extend separate collection schemes, as only separately collected waste can be most
satisfactorily recycled. Coherent with this reasoning, our modeling assumes different
separate collection levels for each system: 25% in the system with the incineration plant,
and 75% in the system with the aerobic MBP.
The two alternative MSW-MS analyzed in this work give service to a population of
100,000 inhabitants living in 25,000 households in an area of 1,000 km2
and generating an
annual waste flux of 50,000 metric tons when no waste prevention strategies are put into
action. These and other characteristics of the functional unit are gathered in table 3.
Page 8

Table 3. Characteristics of the functional unit, and of processes that diverge from default
options in the LCA-IWM assessment tool.
Data input to the LCA-IWM assessment tool
Population 100,000 inhabitants
Area 1,000 km2
Number of households 25,000
Waste generation 50,000 metric tons/year
Reduction due to Waste Prevention no waste prevention (0%); 20%
Temporary Storage
Recycled materials
Mixed residual waste
80 L sacks
1,100 L plastic bins
Collection & Transport
Recycled materials
Mixed residual waste
Fist pick-up distance
Average distance from sector to facilities
150 days/year (biowaste)
100 days/year (others)
310 days/year (as in Bilbao [44] or Donostia [45])
7.5 km
10 km
Efficiency of incineration plant 25%
Electricity mix profiles considered 211 g CO2/kWh (high penetration of renewables);
498 g CO2/kWh (mainly fossil generation)
At this point, an adequate definition of the functional unit is crucial. Several problems
related to the definition of the functional unit arise when performing a comparative LCA of
structurally different waste-management systems.
One of these problems is to solve the allocation of impacts and benefits of different
systems that are intrinsically multifunctional, while maintaining the comparability of the
systems through a common functional unit to all of them. Along with the waste
management service, these integrated systems allow for the recovery of different recycled
materials and energy carriers. But as these recoveries are complementary to the waste
management service, which is the common function to all systems, the functional unit of
the systems compared in this work is defined as a service: the collection and treatment of
all household waste in the defined area in one year. Once the functional unit is defined this
way, the multifunctionality problem can be solved by system expansion/subtraction. This
process is thoroughly explained in Appendix A.
2.2 Waste prevention derived from the broadening of selective collection in Gipuzkoa
Recent experience in several municipalities of Gipuzkoa shows that the substitution of
kerbside collection of mixed residual waste by door to door collection of the different
fractions—including a very small residual fraction—may significantly reduce the total flux of
the waste to be managed by the system. This is the case, for example, of Hernani, a town
of 19,300 inhabitants where the implantation of door to door collection altogether with the
Page 9

promotion of home and district composting and campaigns to raise public awareness has
led to a stable reduction of 28.6% in total generated municipal solid waste (Figure 3; [46-
48]).
Figure 3. Evolution of municipal solid waste generated in Hernani (19,300 inhabitants,
Gipuzkoa) in 2009, 2010 and 2011 before and after the implantation of door to door
collection in May 2010.
This work compares two management systems with different levels of separate collection
(SC), and thus that implement waste prevention strategies up to different levels. This
would be an example of waste prevention as a result of different system dynamics [49].
If the functional unit of the systems under comparison is defined as the one that provides
the service for collection and treatment of all household waste in a given area and year,
then comparability of different waste-management systems is guaranteed only as long as
prevented waste generation remains equal in all systems. Otherwise, the comparison must
account for the avoided impacts in those systems that prevent more waste generation.
Ways to solve this problem have been proposed [36,50,51]. Basically, these works
propose to consider the managed waste flux as the sum of the collected and treated
wastes plus a virtual flux corresponding to the prevented waste. The burdens associated
to the prevented waste should be accounted, in that case, as avoided burdens of the
Page 10

waste-management system because of waste prevention. But this approach is not exempt
from problems [26]. It requires the quantification of a dematerialized flux [50], and entails
abandoning the “zero-burden assumption”, as upstream burdens carried about by
prevented and dematerialized waste should be accounted. As this approach complicates
significantly our comparative LCA, this work does without considering any virtual flux
associated with prevented waste, but always keeping in mind that an accounting error of
avoided burdens is being committed in favor of those systems with less ambitious
prevention strategies.
2.3 Characterization of background and foreground processes
Electricity produced from waste, e.g. in incineration plants with energy recovery, is credited
in our comparative LCA with the corresponding avoided burdens from power generation in
the background system. Thus, electricity generation may cause a huge impact on the net
environmental balance of the waste-management system. When crediting these avoided
burdens, comparative LCAs in the literature often consider national and local electricity
mixes with a very high penetration of fossil fuels [29,35,37,38]. In some studies the
electricity mix of the background system is not even characterized much farther than as
strongly based on fossil, and thus giving way to important avoided burdens [52].
But LCA is often applied to systems that are being projected for the near future [29,36,53]
or although already functioning, that are not expected to be dismantled soon [23]. If
Attributional LCA is applied for the modeling of future systems [54], it has to take into
account data from background processes as they are forecast to be in the future, when the
system under study is supposed to be put into operation. In our study the new incineration
plant in one of the alternative systems would start operation not before 2015, and would
not finish its pay-off period until 2030 [9], being probably in operation by the middle of the
century. Taking into account that the European Commission plans that, due to fossil energy
depletion and fight against climate change, the European power sector should reduce its
GHG emissions between 54% and 68% in 2030 and between 93% and 99% by 2050 [55],
the average production of electricity to be considered in the background system cannot be
carbon intensive.
Actually, Spain has already reduced its electricity mix emissions level from 430 g CO2/kWh
in 2000 [56] down 236 g CO2/kWh in 2013 [57], and will probably reduce it further during
the next decade, well below 200 g CO2/kWh. Following this trend, our comparative LCA
will consider for the background system an emissions level of 211 g CO2/kWh,
corresponding to an electricity mix with a high penetration of renewables. In order to
perform a sensitivity analysis of these avoided burdens, our study will also consider
another electricity mix, much more dependent on fossil fuels, with an emissions level of
498 g CO2/kWh. These two electricity mixes are characterized in the Ecoinvent-2000
database [58] and can be used by the LCA-IWM assessment tool.
The need to correctly address the average process is also applicable to products obtained
from material recovery. When assessing forecast systems, the LCA practitioner should
also take into account that the production technologies of paper, plastics, ferrous metals,
aluminum and organic fertilizers—which are displaced by compost—will probably reduce
their burdens in the future, e.g. as it has occurred with the production of nitrogenous
fertilizers, where using best available techniques may significantly reduce N2O emissions
and energy demand [22].
Page 11

Also, sufficient information has to be provided about the assessed processes for energy
and material recovery. In the case of our comparative LCA, these processes are those
modeled by the LCA-IWM assessment tool, and characterized in its documentation [59]:
• The incineration plant is equipped with grate firing and flue gas cleaning
(electrostatic precipitator for dust and fly ashes; acid flue gas scrubbing for removal
of HCl, HF and heavy metals; neutral SO2-scrubbing facility with suspended
Ca(OH)2; filters with activated carbon for removal of dioxines/furanes; and Selective
Catalytic Reduction for denitrification). The Waste-to-Energy plant (WtE) produces
only electricity, as climatic conditions in Gipuzkoa would not guarantee sufficient
heat demand from a CHP plant [60]. A thermoelectric efficiency of 25% has been
supposed, so that the incineration plant reaches the R1 status of the WFD [61].
• For the recycling of plastics, it is assumed that plastics and composites separately
collected are composed by the following seven fractions: HDPE, PET, LDPE film,
mixed plastics, liquid beverage cartons, other composites, and contaminants (11%).
These fractions are sorted in a Material Recovery Facility, and transported to
recycling facilities. Recycled HDPE substitutes primary HDPE for multi-layered
bottles (1:1 basis). Recycled PET substitutes primary PET for three-layered bottles
(1:1). Recycled LDPE film substitutes primary LDPE for sacs (1:1). Mixed plastics
are recycled into plastic pickets, which replace wood pickets (1:1 basis). Liquid
beverage cartons are recycled into pulp that substitutes primary pulp for domestic
paper (1:1). Rejects of sorting processes and some composites are incinerated if
the system has an incineration plant, otherwise they are landfilled.
• Recycling of metals. To reprocess steel from scrap, first it is sorted to remove
contaminants, so that it can be melted and recast. Tinplate is electrolytically de-
tined to produce steel. Reprocessing of aluminum, which is much less energy
intensive than its production from virgin materials, requires sorting and then melting
in a furnace. Our model assumes that metals are sorted in a Material Recovery
Facility and transported to recycling facilities, where tinplate steel is recycled into
secondary steel, substituting primary steel in a 1:1 basis; aluminum is recycled into
secondary aluminum, which substitutes primary aluminum in a 1:1 basis. Rejects of
sorting processes (5%) are landfilled or incinerated.
• Related to recycling of paper and cardboard, following LCA-IWM, our model
assumes that 1 kg of recycled pulp replaces 1 kg of primary pulp, and that
cardboard is recycled into cardboard. 2% rejects are derived to incineration if
available; otherwise they are landfilled.
• Different subfractions of glass (green, brown, clear, mixed glass) are cleaned and
crushed into broken glass in a Material Recovery Facility and transported to a
recycling facility. Rejects (3%) of cleaning and crushing processes are landfilled or
incinerated. Clean broken glass is recycled into glass, assuming that 1 kg replaces
1.19 kg of raw materials.
• The modeled landfill is equipped with gas and leachate collection systems. The
collected gas is utilized for energy production, and leachate is treated before
discharge.
• The composting process of the biowaste is modeled by the LCA-IWM tool assuming
Page 12

the operation of a fully encapsulated composting plant with a first stage of intensive
composting in a box system, and a subsequent maturation step in enclosed
windrows. Obtaining high quality compost is not a problem when the biowaste is
separately collected. Its application brings positive effects in form of nutrient and
organic carbon supply, along with carbon sequestration. Our modeling assumes
default parameters from the LCA-IWM tool, which imply the substitution of mineral
fertilizers in a 1:1 basis (based on the nutrient content), the substitution of peat—
which is considered a fossil resource—for introduction of organic matter to the soil,
and carbon sequestration equivalent to 8.2% of the carbon present in final compost.
3. Results and discussion
In this section we present the results of the comparative life-cycle assessment of the two
alternative integrated MSW-MS whose characteristics have been previously detailed.
These results are gathered in table 4. The scenario labeled as A25 models the system in
which 25% of waste is separately collected and the other 75% of mixed residual waste is
treated in a WtE plant. The scenario labeled as B75 models the system in which 75% of
waste is separately collected, and the other 25% of mixed residual waste is subjected to
aerobic mechanical biological pretreatment and subsequent disposal of nonrecyclable inert
materials to landfill. Scenarios A25 and B75 are modeled assuming a power system in the
background with a high penetration of renewables (emissions level of 211 g CO2/kWh).
These two basic scenarios are complemented with other three in which some of the
simulation conditions are modified in order to perform sensitivity analysis of some
significant parameters:
• In order to check the relevance of waste prevention and recycling derived from the
increase of selective collection, scenario B25 resembles scenario B75 but where
just 25% of waste is separately collected, and there is no reduction in waste
generation due to prevention.
• In order to check the relevance of the electricity mix assumed in the background,
A25C and B25C scenarios model the systems considered in scenarios A25 and
B25, but assuming a power system in the background that is carbon intensive (498
g CO2/kWh).
Table 4 shows the five scenarios analyzed, with the parameters that differentiate each one,
and their modeling results for six significant impact categories. These categories are those
assessed by the LCA-IWM tool following the CML 2001 method [12], and they are
identified as the most significant when comparing waste-management systems. The first
two, abiotic resource depletion (ard, measured in Mg Sb eq) and global warming potential
(gw, measured in GgCO2 eq) are very good indicators of cumulative material resource
consumption (ard) and cumulative fossil energy demand (gw), representing very good
indicators of global energy and material recovery. The other four impact categories
analyzed are: human toxicity (htox, measured in kg 1,4-Dichlorobenzene-eq), photo-
oxidant formation (ph-tox, measured in kg Ethene-eq), acidification (acid, measured in kg
SO2 eq) and eutrophication (eutro, measured in kg PO4 eq). Quantities of annual waste
derived to landfills are also gathered in table 4 for each scenario, measured in metric tons.
Page 13

Table 4. Parameter characterization and results of significant impact categories for five scenarios analyzed (A25, A25C, B25, B25C,
B75), organized in four comparative pairs (A25C-B25C, A25C-A25, B25-B75, A25-B75) with the changing parameters in each pair
in bold type.
Scenario Mixed
residual waste
treatment
Separate
Collection
(%)
Reduction due to
waste prevention
(WP, %)
Electricity mix
(g CO2/kWh)
Abiotic Resource
Depletion
(ard, Mg Sb-eq)
Global Warming
Potential
(gw, Gg CO2 eq)
Human Toxicity
(htox, kg 1,4-
Dichlorobenzene-eq)
Photo-oxidant
formation (ph-tox
kg Ethene-eq)
Acidification
(acid, kg SO2eq)
Eutrophication
(eutro, kg PO4eq)
Waste
Landfilled
(tonnes)
A25C Incineration 25% No 498 –88.7 –9.56 –2.34 –4.76 –105 –946 7,790
B25C Aerobic MBP 25% No 498 –50.2 –5.04 1.44 –2.04 –56.8 2,352 27,175
A25C Incineration 25% No 498 –88.7 –9.56 –2.34 –4.76 –105 –946 7,790
A25 Incineration 25% No 211 –54.1 –4.76 –1.60 –3.15 –64.5 353 7,790
B25 Aerobic MBP 25% No 211 –45.0 –4.32 1.55 –1.80 –50.7 2,549 27,175
B75 Aerobic MBP 75% Yes, 20% 211 –85.7 –11.09 –1.79 –6.55 –139 907 9,939
A25 Incineration 25% No 211 –54.1 –4.76 –1.60 –3.15 –64.5 353 7,790
B75 Aerobic MBP 75% Yes, 20% 211 –85.7 –11.09 –1.79 –6.55 –139 907 9,939
Page 14

The first pair of scenarios shown in table 4 (scenarios A25C-B25C) compare impact
categories in both waste-management systems when separate collection is 25%, and a
carbon intensive electricity mix is assumed in the background. The life cycle assessment
provides better results (more negative) in all impact categories for scenario A25C, showing
that it is environmentally more beneficial to incinerate the mixed residual waste than to
inertize and dispose of it to landfill when just 25% of all generated household waste is
separately collected.
The Spanish power sector is undergoing a decarbonization process that will strengthen in
the coming decades. Hence it seems more adequate to assume an electricity mix for the
background system less reliant on fossil fuels than that considered in scenarios A25C-
B25C. The second pair of scenarios compared in table 4 (A25C-A25) allows a sensitivity
analysis of the electricity mix in the background. The comparison shows the consequence
of reducing the electricity emissions from 498 down to 211 g CO2/kWh in the system with
the WtE plant: all environmental impacts remain beneficial due to important avoided
burdens, but they are significantly reduced, from 32% (htox) up to 50% (gw).
Another factor that has to be considered when comparing the two alternative integrated
MSW-MS is the possibility to increase separate collection. Rigamonti et al. [30] state that
the optimum share for separate collection may be around 50% due to contaminations; but
assuring high efficiencies in the separate collection of each fraction would locate the
optimum well over 60%. Actually, Slagstad and Brattebø [20] consider in their comparative
assessment for a new urban settlement a feasible sorting efficiency of 70% for food waste,
and between 70% and 90% for all other waste fluxes. In our case, the third pair of
scenarios compared in table 4 (B25-B75) perform a sensitivity analysis of the spreading of
separate collection, comparing impact categories when it is 25% and 75% in the
management system that derives the mixed residual waste to aerobic MBP. The results
show important improvements in all impact categories. This is due to the increased
avoided burdens that are accounted when tripling separate collection, and thus material
recovery. The improvement is significant even in the global warming potential category,
directly linked to fossil energy consumption (increase of 156%). It has to be added that this
modeling underestimates the environmental benefit of increasing separate collection, as
our modeling does not assign avoided burdens to a waste prevention that is estimated in
20%.
Direct energy recovery from waste is an environmental improvement when performed in a
waste-management system. But the expansion of separate collection schemes provides
environmental benefits through expanded material recovery that may overwhelm those
derived from energy recovery. A better result from direct material recovery (e.g. recycling)
when compared with direct energy recovery (e.g. incineration) is confirmed by other works
[27,62,63], and supports the fact that the former is located higher in the waste hierarchy
[8]. This point is confirmed by the last pair of scenarios compared in table 4 (A25-B75),
where the waste-management system with an incineration plant that separately collects
just 25% of all household waste is compared with the system that separately collects 75%
for material recovery, and derives to aerobic MBP the mixed residual waste. This second
system (scenario B75) behaves better in all environmental categories except
eutrophication, in which the gap between the two systems is nevertheless significantly
reduced with respect to results when separate collection is 25% in both systems (A25C-
B25C).
Giving priority to material recycling over direct energy recovery improves material
Page 15

recovery, and therefore scenario B75 shows a better environmental impact in the Abiotic
resource depletion category (–85.7 Gg Sb eq) than scenario A25 (–54.1 Gg Sb eq). But
results show that overall energy recovery is also improved when material recovery is
prioritized: scenario B75 shows a better result in the global warming potential category
(–11.09 Gg CO2 eq), closely related to fossil fuels consumption, than scenario A25
(–4.76 Gg CO2 eq). This is due to the fact that important quantities of energy are required
to produce materials that can be substituted by recycled products. This energy
consumption is avoided with material recovery, and actually exceeds direct energy
recovery form waste in the considered systems. This is shown in figure 4, which details the
partial contribution of each management stage and treatment process to the net
environmental impact in scenarios A25 and B75.
Figure 4. Comparison of significant impact categories of scenarios A25 (energy recovery
from 75% mixed residual waste, material recovery from 25% separately collected waste)
and B75 (material recovery from 75% separately collected waste, aerobic MBP of 25%
mixed residual waste), broken down into partial contributions in each category from waste
management stages and treatment processes that make up both systems.
Figure 4 shows the importance of the avoided burdens in material recovery from the
Page 16

separately collected plastics, paper, glass and metals residues. The avoided burdens are
especially important for material recovery from plastics residues in the categories of abiotic
resource depletion and eutrophication; for recovery from glass in human toxicity; and for
recovery from paper in photo-oxidant formation and acidification. Avoided burdens due to
recovery from metals seem to be less important in the category of human toxicity, but are
comparatively significant in all other categories.
Credits for the avoided burdens in material recovery are also important in the system with
incineration, but these are less significant than in the modeled system with aerobic MBP of
the mixed residual waste. Actually, most of the credits come from the recovery of materials
separately collected, and therefore they keep approximately proportional to the share of
separate collection in total waste collection. The increase of avoided burdens carried out
by the increase of the share of separate collection in one system (B75) more than
compensates for the credits gained in the other system when those residues are
incinerated as part of the mixed residual fraction (A25). Those credits, besides, are limited
to the abiotic resource depletion and human toxicity categories, and to the avoided
burdens from the aerobic MBP—inexistent in the system with WtE plant—and also limited
to the impact categories of human toxicity and eutrophication.
Composting biowaste provides some significant environmental credits, especially in the
categories of global warming and human toxicity. Inasmuch as composting of biowaste is
not free of some emissions, especially of ammonia [64], those reflect with a significant
impact in the category of eutrophication, and with a much lesser extent in the categories of
photo-oxidant formation and acidification. Composting brings about with it some
environmental impacts that would be inexistent in a management system where most of
the biowaste is incinerated. Nevertheless, assessment tools do not normally consider
some environmental benefits of composting e.g. improvement of soil health, fertility and
water retention capacity, and reduced pesticide consumption [14]. In addition, other
alternatives to the aerobic processing of biowaste to produce compost could be also
considered as alternatives to biowaste incineration, such as anaerobic digestion, which,
besides, allows for the direct recovery of energy by means of biogas production, along with
other material recoveries (digestate). The consideration of these alternatives falls out of
the scope of this paper, but other studies have already addressed a more beneficial net
balance of anaerobic treatments when compared with composting [27]. Nevertheless,
composting is credited as a very suitable biowaste treatment option for European Southern
regions [65].
Another important environmental impact of the waste-management systems under
analysis is the disposal to landfill of final waste fluxes, mainly rejected materials in
recycling plants, and slag and ashes from incineration. Although these secondary wastes
generated in incineration plants are not statistically reported as part of the municipal waste
data collected in Europe [66], in many countries landfilling is inseparable from incineration
if the complete life-cycle of municipal wastes is considered. This is well known, for
example, in land-scarce and incineration-intensive Singapore, where the spread of
separate collection of municipal waste is addressed as a key approach to reduce the need
of almost saturated landfills for the disposal of slag and ashes generated in incineration
plants [67].
Final waste fluxes disposed of to landfill in each scenario are gathered in the last column
in table 4. While the system with incineration and 25% of separate collection (scenarios
A25, A25C) manages annually 50,000 metric tons of waste and derives to landfill
Page 17

7,790 metric tons, the system without incineration under the same conditions for separate
collection (B25) derives to landfill 27,175 metric tons of final residues. From this
comparison we may conclude that incineration is a viable strategy to reduce the flux of
final waste derived to landfill; but not the only strategy. When waste prevention and the
spreading of separate collection are implemented in our model, the system without
incineration (scenario B75) derives just 9,939 metric tons to landfill, which supposes a
reduction of 63.4%.
4. Conclusions
This work performs a comparative analysis of two alternative approaches for an integrated
MSW-MS to be implemented in the Basque province of Gipuzkoa (Spain). These
alternatives place different emphasis on energy or material recovery from waste,
significantly complicating their overall environmental assessment. In order to solve this
problem, LCA methodology provides a powerful framework for the overall sustainability
assessment of systems that combine different levels of energy and material recovery.
The comparative LCA of the two systems (results in table 4) shows that, when separate
collection is limited to 25%, the system with the incineration plant provides much better
environmental results in all impact categories, especially if an electricity mix very
dependent on fossil fuels is assumed for the background system. But the results change
drastically if the comparison is performed considering that separate collection reaches
75%. This level of separate collection is supported by evidence in municipalities of the
province of Gipuzkoa like Hernani, where the increase of separate collection up to 80%, in
conjunction with other waste prevention strategies, has also carried with it important
reductions in household waste generation. Under these conditions the system that
emphasizes separate collection and material recovery obtains better results in all impact
categories but eutrophication, when compared to the system with the WtE plant. The
improvement is especially significant in the category of abiotic resource consumption
(+58%), and in the category of global warming potential (+132% better).
The breakdown of each category result into partial contributions from waste management
stages and treatment processes shows the importance of the avoided burdens in material
recovery from the separately collected plastics, paper, glass and metals. Under the
conditions assumed in this work for the functional unit operating in Gipuzkoa, it can be
concluded that separately collecting a high share of waste—which thereby can be derived
to recycling processes for material recovery—provides better environmental results than
deriving it as a mixed residue to an incineration plant where energy is recovered in the
form of electricity. These superior environmental results are obtained even in impact
categories tightly related to fossil energy consumption, such as the global warming
potential category. The only impact category in which the system with the incineration plant
performs better is eutrophication, due to ammonia emissions in composting of biowaste.
Besides, both systems generate similar final fluxes to landfill: 7,790 metric tons in the
system with the incineration plant, vs. 9,939 metric tons in the system without incineration.
This shows that spreading separate collection and promoting waste prevention may be
such a good strategy as well as incinerating mixed residual waste in order to reduce the
quantity of residues finally derived to landfill.
Page 18

Appendix A. System expansion to determine avoided burdens
System expansion/subtraction is performed to solve the allocation of impacts and benefits
of different systems that are intrinsically multifunctional. It is performed as follows.
Figure A.1(a) shows a diagram of waste-management system i for the treatment of waste
Wi (the Service that determines the functional unit); the system also produces a series of
complementary products (Pj,i), and causes some specific impacts. In our study we perform
a screening LCA in which we focus on abiotic resource depletion (ardi) and global warming
potential (gwi) impact categories, as they are considered to show the following trend of
most important environmental impact categories [56]. RMi is the resource material demand
for the functioning of system i, and PEi corresponds to primary energy demand, which is
analogous to the Cumulative Energy Demand impact assessment method implemented in
the Ecoinvent database [68].
Multifunctionality is solved by system expansion [69]. In a first step, system expansion is
performed in all compared systems until all expanded systems produce identical quantities
of common products and services. Such system expansion is performed in each system
for each product Pj, making use of the corresponding production blocks for each product
(figure A.1(b)), in which production inputs and corresponding impacts are recorded. In
coherence with the attributional modeling principle, average processes in the background
system are considered for their characterization. Secondly, production outputs and inputs
related to all coproducts complementary to the main service provided by the waste-
management system are subtracted from all expanded systems, using again the average
processes in the background system. These two steps can be condensed in just one step
in which production of every complementary coproduct is subtracted in each system using
the energy and material input demand and environmental impacts that correspond for the
production of each complementary product in the background system; the net result is
shown schematically in figure A.1(c).
Figure A.1. (a) System i for treatment of waste Wi, which also produces a series of
Page 19

complementary products (Pj,i), and causes some specific impacts ardi and gwi; (b)
Production system of product j to be considered in expanded systems, which requires of
resource materials (RMPj) and primary energy (PEPj), and causes impacts (ardPj, gwPj); (c)
Waste-management system i in which complementary coproducts and corresponding
inputs and impacts have been subtracted.
The multifunctionality problem is solved in attributional LCA by the accounting as avoided
burdens of those impacts associated with the production in the background system, with
some specific average processes, of the products substituted by the complementary
coproducts. This way, a correct characterization of these average processes is critical;
actually, these avoided burdens are so important that net environmental impacts are
usually negative in most systems and for most indicators: the net environmental balance of
the waste-management system results to be beneficial due to the substitution of other
more harmful ways to produce the coproducts in the background system complementary
to the waste management service.
When systems expansion/substitution is performed in order to solve the multifunctionality
problem, with the crediting of avoided burdens, it is not fair to compare different waste-
management systems in terms of direct energy generation or direct material recovery.
When different systems (figure A.1(a)) are credited with the avoided burdens associated to
the production of the coproducts in each system (figure A.1(b)), the resultant systems that
we are actually comparing through the LCA neither produce energy nor recover materials
(figure A.1(c)), and consequently it is not adequate to compare those systems in terms of
directly generated electricity, or of quantities of recycled materials. At best, a fair
comparison of produced coproducts should be made through the expanded systems; but
the result is previously known: all expanded systems under comparison must provide
exactly the same coproducts—altogether with the service of the functional unit—, as that is
actually the condition imposed to solve the multifunctionality problem, indispensable to
allow a fair comparison of environmental impacts. A similar argument is applicable when
we refer to efficiency, e.g. of electricity generation. The efficiency of a waste-management
system with an incineration plant that presents a thermoelectric efficiency of 25% is not
better than that of an expanded system that lacks incineration plants, as the efficiency of
the latter is precisely the one of the background system, i.e. a power system with highly
optimized units [43].
From the previous reasoning, however, we may not conclude that energy and material
recovery is neither considered nor quantified in comparative LCA. Indeed, they are
accounted through the avoided burdens linked to the production of the materials and
energy substituted by the coproducts, and thus credited to the systems. As shown in
figure A.1(c), the substitution of material Pj with a recycled material in system i is credited
with a negative impact –ardPj,i in the field of abiotic resource depletion, and a negative
impact –gwPj,i in the field of global warming, e.g. due to the avoided consumption of fossil
fuels needed to obtain product Pj in the background system.
When these avoided burdens are credited, after subtraction, they also appear among the
inputs to the compared systems. System i is credited with a negative input of resource
materials (–RMPj,i) and primary energy (–PEPj,i) due to the avoided consumption of
materials and energy otherwise required to obtain the product/material Pj, substituted by a
particular recovered coproduct. For the case of primary energy, the term PEi–ΣjPEPj,i
Page 20

corresponds to the net primary energy demand of system i subtracted the coproducts —
which is analogous to applying the Cumulative Energy Demand impact assessment
method implemented in the Ecoinvent database [68]—. RMi–ΣjRMPj,i corresponds to the
net resource material demand for the functioning of system i, subtracted the coproducts.
Net material and energy demands may be negative in this calculation, as they correspond
to a subtracted system that is credited with some avoided burdens, and those may be
significant. This negative net input flux of energy and materials, however, should not be
interpreted as a net positive output flux, as we are considering subtracted (differential)
systems. Its effect in the overall balance is normally reflected through the monetization [14]
of energy and materials recovered by the waste-management system, which, through their
market values, internalize the primary energy and resource materials required for their
production or fabrication in the background system [23,70].
Acknowledgement
This Research was supported by the Provincial Government of Gipuzkoa (R&D Research
Contract 2012.0485, “Hiri hondakinei buruzko txostena, haien tratamendu eta kudeaketa
Gipuzkoako Lurrandean”).
References
[1] Diputación Foral de Gipuzkoa. AURRERAPEN DOKUMENTUAREN
GARAPENERAKO ESTRATEGIA 2008 – 2016 ESTRATEGIA DE DESARROLLO DEL
DOCUMENTO DE PROGRESO 2008 – 2016, 2012,
http://www.gipuzkoaberri.net/WAS/CORP/DPDOficinaPrensaDigitalWEB/descarga.do?
1211101171160660730431041161151211061151101051160660700431041161051101081
16066055043116119105106115066054 (accessed 20.05.2014).
[2] Diputación Foral de Bizkaia. Observatorio Permanente de Residuos Urbanos del
Territorio Histórico de Bizkaia. Datos de residuos. Años 2012, 2013, 2014.
http://www.bizkaia.net/home2/Temas/DetalleTema.asp?
Tem_Codigo=7709&idioma=CA&bnetmobile=0&dpto_biz=9&codpath_biz=9|351|7709
(accessed 20.05.2014).
[3] Mijangos F. Urban-rural duality and waste management. Klimagune Workshop 2014
Opportunities and challenges for rural areas in the context of climate change,
http://www.bc3research.org/klimagune/images/stories/workshop/2014/ponencias/KW2014
_Fernando_Mijangos.pdf (accessed 17.03.2015)
[4] Eurostat. Municipal waste Database. http://appsso.eurostat.ec.europa.eu/nui/show.do?
dataset=env_wasmun&lang=en (accessed 17.03.2015)
[5] Diputación Foral de Gipuzkoa. PIGRUG 2002-2016, Documento de Progreso 2008-
2016. 2008.
http://www4.gipuzkoa.net/medioambiente/dpro/doc/es/01Documento_de_Progreso_CAST.
pdf (accessed 20.05.2014).
Page 21

[6] Consorcio de Residuos de Gipuzkoa. Tablas de datos de los Residuos Urbanos de
Gipuzkoa. http://www.ghk.eus/es/datos/gipuzkoa/gipuzkoa-2013 (accessed 16.01.2015).
[7] Muñoz I, Rieradevall J, Doménech X, Milà L. LCA application to integrated waste
management planning in Gipuzkoa (Spain), Int J Life Cycle Assess, 2004;9:272-80,
http://dx.doi.org/10.1007/BF02978603.
[8] European Commission. Directive 2008/98/EC of the European Parliament and of the
Council of 19 November 2008 on waste and repealing certain Directives. Off J Eur Union L
2008;312.
[9] European Integrated Pollution Prevention and Control Bureau (EIPPCB). Best Available
Techniques (BAT) reference document for Waste Incineration; 2006.
[10] ISO 14040:2006. Environmental management – life cycle assessment – principles and
framework. CEN (European Committee for Standardisation), Brussels.
[11] ISO 14044:2006. Environmental management – life cycle assessment – requirements
and guidelines. CEN (European Committee for Standardisation), Brussels.
[12] Guinée JB, Gorrée M, Heijungs R, Huppes G, Kleijn R, de Koning A. Handbook on life
cycle assessment: operational guide to the ISO standards. Dordrecht, The Netherlands:
Kluwer Academic Publisher; 2002.
[13] EC-JRC-IES. International reference life cycle data system (ILCD) handbook. General
guide for life cycle assessment—detailed guidance. 1st ed., European Commission–Joint
Research Centre–Institute for Environment and Sustainability; 2010.
[14] EC-JRC-IES. Supporting Environmentally Sound Decisions for Waste Management -
A technical guide to Life Cycle Thinking (LCT) and Life Cycle Assessment (LCA) for waste
experts and LCA practitioners. 1st ed., European Commission–Joint Research Centre–
Institute for Environment and Sustainability; 2011.
[15] Gentil EC, Damgaard A, Hauschild M, Finnveden G, Eriksson O, Thorneloe S et al..
Models for waste life cycle assessment: Review of technical assumptions. Waste Manage
2010; 30:2636-48, http://dx.doi.org/10.1016/j.wasman.2010.06.004.
[16] De Feo G, Malvano C. The use of LCA in selecting the best MSW management
system. Waste Manage 2009;29:1901-15,
http://dx.doi.org/10.1016/j.wasman.2008.12.021.
[17] Bovea MD, Ibáñez-Forés V, Gallardo A, Colomer-Mendoza FJ. Environmental
assessment of alternative municipal solid waste management strategies. A Spanish case
study. Waste Manage 2010;30:2383-95, http://dx.doi.org/10.1016/j.wasman.2010.03.001.
[18] Pires A, Chang N, Martinho G. Reliability-based life cycle assessment for future solid
waste management alternatives in Portugal. Int J Life Cycle Assess 2011;16: 316-37,
http://dx.doi.org/10.1007/s11367-011-0269-7.
[19] Tunesi S. LCA of local strategies for energy recovery from waste in England, applied
Page 22

to a large municipal flow. Waste Manage 2011;31:561-71,
[20] Slagstad H, Brattebø H. LCA for household waste management when planning a new
urban settlement. Waste Manage 2012;32:1482-90,
[21] Song Q, Wang Z, Li J. Environmental performance of municipal solid waste strategies
based on LCA method: a case study of Macau. J Cleaner Prod 2013;57:92-100,
http://dx.doi.org/10.1016/j.jclepro.2013.04.042.
[22] Bernstad A, la Cour Jansen J. Review of comparative LCAs of food waste
management systems – Current status and potential improvements. Waste Manage
2012;32:2439-55, http://dx.doi.org/10.1016/j.wasman.2012.07.023.
[23] Eriksson O, Carlsson Reich M, Frostell B, Björklund A, Assefa G, Sundqvist JO et al.
Municipal solid waste management from a systems perspective. J Cleaner Prod
2005;13:241-52, http://dx.doi.org/10.1016/j.jclepro.2004.02.018.
[24] Merrild H, Larsen AW, Christensen TH. Assessing recycling versus incineration of key
materials in municipal waste: The importance of efficient energy recovery and transport
distances. Waste Manage 2012;32: 1009-18,
[25] Nadzirah Othman S, Zainon Noor Z, Halilu Abba A, O. Yusuf R, Ariffin Abu Hassan M.
Review on life cycle assessment of integrated solid waste management in some Asian
countries. J Cleaner Prod 2013;41:251-262,
[26] Gentil EC, Gallo D, Christensen TH. Environmental evaluation of municipal waste
prevention. Waste Manage 2011;31:2371-9,
[27] Bernstad A, la Cour Jansen J. A life cycle approach to the management of household
food waste – A Swedish full-scale case study. Waste Manage 2011;32:1879-96,
[28] Koroneos CJ, Nanaki EA. Integrated solid waste management and energy production
- a life cycle assessment approach: the case study of the city of Thessaloniki. J Cleaner
Prod 2012;27:141-50, http://dx.doi.org/10.1016/j.jclepro.2012.01.010.
[29] Buttol P, Masoni P, Bonoli A, Goldoni S, Belladonna V, Cavazzuti C. LCA of integrated
MSW management systems: Case study of the Bologna District. Waste Manage
[30] Rigamonti L, Grosso M, Giugliano M. Life cycle assessment for optimising the level of
separated collection in integrated MSW management systems. Waste Manage
[31] Calabrò PS. Greenhouse gases emission from municipal waste management: The
role of separate collection. Waste Manage 2009;29:2178-87,
Page 23

[32] Consonni S, Giugliano M, Massarutto A, Ragazzi M, Saccani C. Material and energy
recovery in integrated waste management systems: Project overview and main results.
Waste Manage 2011;31:2057-65, http://dx.doi.org/10.1016/j.wasman.2011.04.016.
[33] Cimpan C, Wenzel H. Energy implications of mechanical and mechanical–biological
treatment compared to direct waste-to-energy. Waste Manage 2013;33:1648-58,
[34] Belboom S, Digneffe JM, Renzoni R, Germain A, Léonard A. Comparing technologies
for municipal solid waste management using life cycle assessment methodology: a Belgian
case study. Int J Life Cycle Assess 2013;18:1513-23, http://dx.doi.org/10.1007/s11367-
013-0603-3.
[35] Koci V, Trecakova T. Mixed municipal waste management in the Czech Republic from
the point of view of the LCA method. Int J Life Cycle Assess 2011;16:113–24,
http://dx.doi.org/10.1007/s11367-011-0251-4.
[36] Scipioni A, Mazzi A, Niero M, Boatto T. LCA to choose among alternative design
solutions: The case study of a new Italian incineration line. Waste Manage 2009;29:2462-
74, http://dx.doi.org/10.1016/j.wasman.2009.04.007.
[37] Wittmaier M, Langer S, Sawilla B. Possibilities and limitations of life cycle assessment
(LCA) in the development of waste utilization systems – Applied examples for a region in
Northern Germany. Waste Manage 2009;29:1732-38,
[38] Assamoi B, Lawryshyn Y. The environmental comparison of landfilling vs. incineration
of MSW accounting for waste diversion. Waste Manage 2012;32:1019-30,
[39] Burnley S, Phillips R, Coleman T, Rampling T. Energy implications of the thermal
recovery of biodegradable municipal waste materials in the United Kingdom. Waste
Manage 2011;31:1949-59, http://dx.doi.org/10.1016/j.wasman.2011.04.015.
[40] Fruergaard T, Astrup T. Optimal utilization of waste-to-energy in an LCA perspective.
Waste Manage 2011;31: 572-82, http://dx.doi.org/10.1016/j.wasman.2010.09.009.
[41] den Boer J, den Boer E, Jager J. LCA-IWM: A decision support tool for sustainability
assessment of waste management systems. Waste Manage 2007;27:1032-45,
[42] ALTAIR Ingeniería. Documento 1/2: Metodología y caracterización de la fracción resto
de los residuos domésticos generados en hogares y comercios, y la fracción resto de los
residuos comerciales que se depositan en el mismo contenedor, para el Territorio Histórico
de Gipuzkoa, 2013.
http://www4.gipuzkoa.net/MedioAmbiente/gipuzkoaingurumena/adj/documentacion/CARA
CTERIZACION%202012-2013.pdf (accessed 20.05.2014).
[43] Consonni S, Viganò F. Material and energy recovery in integrated waste management
systems: The potential for energy recovery. Waste Manage 2011;31:2074-84,
Page 24

[44] Household collection schedule in Bilbao http://www.bilbao.net/cs/Satellite?
c=BIO_Servicio_FA&cid=3007556277&language=es&pageid=3000094417&pagename=Bil
baonet%2FBIO_Servicio_FA
%2FBIO_Servicio&anclaServ=aB3&rutaCatServ=3003446956 (accessed 20.05.2014).
[45] Household collection schedule in Donosita-San Sebastián
http://jokogarbia.donostia.org/es/errefusa/ (accessed 20.05.2014).
[46] Ayuntamiento de Hernani. Datos de recogida 2010.
http://www.hernani.net/images/stories/zerbitzuak/Atez_ate/2010eko_datuak.pdf (accessed
20.05.2014).
[47] Ayuntamiento de Hernani. Datos de recogida 2011.
http://www.hernani.net/images/stories/zerbitzuak/Atez_ate/2011KO_BILKETAREN__DATU
AK.pdf (accessed 20.05.2014).
[48] Mancomunidad de SanMarkos. Datos-oficiales-municipios09 (Hernani).xls, personal
communication, 2013.
[49] Ekvall T, Assefa G, Björklund A, Eriksson O, Finnveden G. What life-cycle assessment
does and does not do in assessments of waste management. Waste Manage
[50] Cleary J. The incorporation of waste prevention activities into life cycle assessments
of municipal solid waste management systems: methodological issues. Int J Life Cycle
Assess 2010;15:579–89, http://dx.doi.org/10.1007/s11367-010-0186-1.
[51] Nessi S, Rigamonti L, Grosso M. Discussion on methods to include prevention
activities in waste management LCA. Int J Life Cycle Assess 2013;18:1358-73,
http://dx.doi.org/10.1007/s11367-013-0570-8.
[52] Hong J, Li X, Zhaojie C. Life cycle assessment of four municipal solid waste
management scenarios in China. Waste Manage 2010;30:2362-69,
[53] Papageorgiou A, Barton JR, Karagiannidis A. Assessment of the greenhouse effect
impact of technologies used for energy recovery from municipal waste: A case for England.
J Environ Manage 2009;90:2999-3012, http://dx.doi.org/10.1016/j.jenvman.2009.04.012.
[54] Ekvall T, Tillman A, Molander S. Normative ethics and methodology for life cycle
assessment. J Cleaner Prod 2005;13:1225-34,
[55] European Commission. Communication from the Commission to the European
parliament, the Council, the European Economic and social committee and the Committee
of the regions: a roadmap for moving to a competitive low-carbon economy in 2050,
European commission SEC(2011)288 final.
[56] CO2Scorecard, 2014. <http://www.co2scorecard.org/> (accessed 20.05.2014).
[57] Red Eléctrica de España. El sistema eléctrico español, AVANCE DEL INFORME
2013. 2013.
Page 25

http://www.ree.es/sites/default/files/downloadable/avance_informe_sistema_electrico_201
3.pdf (accessed 20.05.2014).
[58] Swiss Centre for Life Cycle Inventories. ECOINVENT-2000 Data V1.01 (2003), LCI of
electricity supply mix in European countries, http://www.ecoinvent.org/database/ (accessed
20.05.2014).
[59] den Boer E, den Boer J, Jager J, Rodrigo J, Meneses M, Castells F et al. Deliverable
Report on D3.1 and D3.2: Environmental Sustainability Criteria and Indicators for waste
management (Work Package 3) The Use of Life Cycle Assessment Tool for the
Development of Integrated Waste Management Strategies for Cities and Regions with
Rapid Growing Economies LCA-IWM, 2005.
[60] Reimann DO. CEWEP Energy Report III (Status 2007-2010) Results of Specific Data
for Energy, R1 Plant Efficiency Factor and NCV of 314 European Waste-to-Energy (WtE)
Plants, Bamberg, Germany, 2012. http://www.cewep.eu/m_1069 (accessed 20.05.2014).
[61] Gobierno Vasco. Proyecto técnico y estudio de impacto ambiental del Centro de
Gestión de Residuos de Gipuzkoa. 2009. http://www.ingurumena.ejgv.euskadi.net/r49-
6172/es/contenidos/informe_estudio/gipuzkoako_hondakin_kudeaketa/es_doc/inicio.html
(accessed 20.05.2014).
[62] Arafat HA, Jijakli K, Ahsan A. Environmental performance and energy recovery
potential of five processes for municipal solid waste treatment. J Cleaner Prod Available
2013, http://dx.doi.org/10.1016/j.jclepro.2013.11.071.
[63] Fricke K, Bahr T, Bidlingmaier W, Springer C. Energy efficiency of substance and
energy recovery of selected waste fractions. Waste Manage 2011;31:644-48,
[64] Saer A, Lansing S, Davitt NH, Graves RE. Life cycle assessment of a food waste
composting system: environmental impact hotspots. J Cleaner Prod 2013;52:234-44,
[65] Cadena E, Colón J, Artola A, Sánchez A, Font X. Environmental impact of two aerobic
composting technologies using life cycle assessment. Int J Life Cycle Assess 2009;14:401-
10, http://dx.doi.org/10.1007/s11367-009-0107-3.
[66] Eurostat, 2012. Guidance on municipal waste data collection, November 2012,
Eurostat – Unit E3 – Environment and forestry, Guidance on municipal waste data
collection November–2012, WASTE WG 5.2 b(2012), 2012.
http://epp.eurostat.ec.europa.eu/portal/page/portal/waste/documents/Municipal_waste_statistics_gui
dance.pdf (accessed 20.05.2014).
[67] Khoo HH, Tan LLZ, Tan RBH. Projecting the environmental profile of Singapore’s
landfill activities: Comparisons of present and future scenarios based on LCA. Waste
Manage 2012;32:890-900, http://dx.doi.org/10.1016/j.wasman.2011.12.010.
[68] Frischknecht R, Jungbluth N, Althaus H, Bauer C, Doka G, Dones R. et al.
Implementation of life cycle impact assessment methods: data v2.0. Ecoinvent report no.
3. Swiss centre for Life Cycle Inventories. Dübendorf, Switzerland; 2007.
Page 26

[69] Weidema B. Avoiding Co-Product Allocation in Life-Cycle Assessment. J Ind Ecol
2001;11:4-33.
[70] Massarutto A, de Carli A, Graffi M. Material and energy recovery in integrated waste
management systems: A life-cycle costing approach. Waste Manage 2011;31:2102-11,
Page 27

Comparative lca of two approaches with different emphasis on energy or material recovery for a municipal solid waste management system in gipuzkoa

Recommended

Recommended

More Related Content

What's hot

What's hot (20)

Similar to Comparative lca of two approaches with different emphasis on energy or material recovery for a municipal solid waste management system in gipuzkoa

Similar to Comparative lca of two approaches with different emphasis on energy or material recovery for a municipal solid waste management system in gipuzkoa (20)

More from Zero Zabor ingurumen babeserako elkartea

More from Zero Zabor ingurumen babeserako elkartea (20)

Recently uploaded

Recently uploaded (20)

Comparative lca of two approaches with different emphasis on energy or material recovery for a municipal solid waste management system in gipuzkoa