Environmental impacts of takeaway food containers

Abstract

The consumption of takeaway food is increasing worldwide. Single-use containers used for takeaway food represent a significant source of waste and environmental impacts due to their low recyclability. Consequently, it is important to identify the best available alternatives and improvement opportunities to reduce the environmental impacts of fast-food containers. For these purposes, this study estimates and compares for the first time the life cycle impacts of three most widely-used types of takeaway container: aluminium, polypropylene and extruded polystyrene. These are also compared to reusable polypropylene containers. The findings suggest that single-use polypropylene containers are the worst option for seven out of 12 impacts considered, including global warming potential. They are followed by the aluminium alternative with five highest impacts, including depletion of ozone layer and human toxicity. Overall, extruded polystyrene containers have the lowest impacts due to the lower material and electricity requirements in their manufacture. They are also the best option when compared to reused takeaway polypropylene containers, unless the latter are reused 3–39 times. The number of uses needed for the reusable “Tupperware” polypropylene food savers is even higher, ranging from 16 to 208 times, with terrestrial ecotoxicity being always higher than for extruded polystyrene, regardless of the number of uses. However, extruded polystyrene containers are currently not recycled and cannot be considered a sustainable option. If they were recycled in accordance with the European Union 2025 policy on waste packaging, most of their impacts would be reduced by >18%, while also reducing littering and negative effects on marine organisms. Most of the impacts of the other two types of container would also be reduced (>20%) through increased recycling. Implementing the European Union 2025 policy on recycling of waste packaging would reduce all the impacts by 2%–60%, including a 33% reduction in global warming potential. Based on 2025 million takeaway containers used annually in the European Union, the latter would save 61,700 t CO2 eq./yr, equivalent to the emissions of 55,000 light-duty vehicles. The outcomes of this study will be of interest to packaging manufacturers, food outlets, policy makers and consumers.

Link: https://www.sciencedirect.com/science/article/pii/S0959652618336230

Full text

Environmental impacts of takeaway food containers
Alejandro Gallego-Schmid
a
,
b
,
*
, Joan Manuel F. Mendoza
a
, Adisa Azapagic
a
a
Sustainable Industrial Systems, School of Chemical Engineering and Analytical Science, The University of Manchester, The Mill, Sackville Street,
Manchester, M13 9PL, UK
b
Tyndall Centre for Climate Change Research, School of Mechanical, Aerospace and Civil Engineering, The University of Manchester, HG1, Pariser Building,
Sackville Street, Manchester, M13 9PL, UK
article info
Article history:
Received 11 September 2018
Received in revised form
21 November 2018
Accepted 23 November 2018
Available online 24 November 2018
Keywords:
Aluminium
Global warming
Fast food
Packaging
Polypropylene
Extruded polystyrene
abstract
The consumption of takeaway food is increasing worldwide. Single-use containers used for takeaway
food represent a significant source of waste and environmental impacts due to their low recyclability.
Consequently, it is important to identify the best available alternatives and improvement opportunities
to reduce the environmental impacts of fast-food containers. For these purposes, this study estimates
and compares for the first time the life cycle impacts of three most widely-used types of takeaway
container: aluminium, polypropylene and extruded polystyrene. These are also compared to reusable
polypropylene containers. The findings suggest that single-use polypropylene containers are the worst
option for seven out of 12 impacts considered, including global warming potential. They are followed by
the aluminium alternative with five highest impacts, including depletion of ozone layer and human
toxicity. Overall, extruded polystyrene containers have the lowest impacts due to the lower material and
electricity requirements in their manufacture. They are also the best option when compared to reused
takeaway polypropylene containers, unless the latter are reused 3e39 times. The number of uses needed
for the reusable “Tupperware”polypropylene food savers is even higher, ranging from 16 to 208 times,
with terrestrial ecotoxicity being always higher than for extruded polystyrene, regardless of the number
of uses. However, extruded polystyrene containers are currently not recycled and cannot be considered a
sustainable option. If they were recycled in accordance with the European Union 2025 policy on waste
packaging, most of their impacts would be reduced by >18%, while also reducing littering and negative
effects on marine organisms. Most of the impacts of the other two types of container would also be
reduced (>20%) through increased recycling. Implementing the European Union 2025 policy on recycling
of waste packaging would reduce all the impacts by 2%e60%, including a 33% reduction in global
warming potential. Based on 2025 million takeaway containers used annually in the European Union, the
latter would save 61,700 t CO
2
eq./yr, equivalent to the emissions of 55,000 light-duty vehicles. The
outcomes of this study will be of interest to packaging manufacturers, food outlets, policy makers and
consumers.
©2018 Elsevier Ltd. All rights reserved.
1. Introduction
Consumption of fast food is increasing rapidly as a consequence
of modern lifestyles (Razza et al., 2009). Fast food is defined as the
sale of food and drink for immediate consumption after purchase
(Market Line, 2012), either at the food outlets or elsewhere (e.g.
home and work). If it is consumed away from the food outlet, it is
known as takeaway or take-out food. The takeaway food market
has been growing fast due to the convenience and competitive
pricing. According to TechNavio (2016), the global delivery and
takeaway food market, which was valued at $89 billion in 2015, is
expected to grow by 2.7% annually to over $102 billion by 2020.
The increasing importance of the takeaway food sector has
given rise to various sustainability concerns, including food safety
(Meldrum et al., 2009) and labour issues (Schmitt and Jones, 2013).
However, there are also important environmental sustainability
issues associated with the takeaway food supply chain. One of these
is the use of non-reusable containers with a low recyclability
*Corresponding author. Tyndall Centre for Climate Change Research, School of
Mechanical, Aerospace and Civil Engineering, The University of Manchester, HG1,
Pariser Building, Sackville Street, Manchester, M13 9PL, UK.
E-mail addresses: alejandro.gallegoschmid@gmail.com,alejandro.
gallegoschmid@manchester.ac.uk (A. Gallego-Schmid).
Contents lists available at ScienceDirect
Journal of Cleaner Production
journal homepage: www.elsevier.com/locate/jclepro
https://doi.org/10.1016/j.jclepro.2018.11.220
0959-6526/©2018 Elsevier Ltd. All rights reserved.
Journal of Cleaner Production 211 (2019) 417e427

potential (MacKerron, 2015). Based on previous studies (AFCMA,
2004;Alupro, 2016;Cassidy and Elyashiv-Barad, 2007), the esti-
mates in the current research suggests that over 7.5 billion
extruded polystyrene (EPS) containers are used annually in the USA
and more than 1.8 billion aluminium containers in the UK. This
results in the annual consumption of 58,500 t of EPS in USA and
13,680 t of aluminium in the UK. Taking into account their extrac-
tion and processing, this is equivalent to the emissions of 297 Mt
and 167Mt of CO
2
eq. per year, respectively.
The environmental concerns associated with the use of these
containers have been highlighted in several studies (Aarnio and
H€
am€
al€
ainen, 2008;Mason et al., 2004;Shokri et al., 2014). How-
ever, most focused on waste generation and management rather
than applying a full life cycle perspective. At present, there are no
life cycle assessment (LCA) studies of takeaway containers. Related
studies have analysed the environmental impacts associated with
the use of biodegradable materials (e.g. Madival et al., 2009;Razza
et al., 2009;Suwanmanee et al., 2013). Other studies considered
reusable cups (e.g. Potting and van der Harst, 2015;Vercalsteren
et al., 2010;Woods and Bakshi, 2014). Furthermore, Rieradevall
et al. (2000) discussed the implementation of eco-design criteria
to fast-food packaging.
One of the most critical environmental aspects related to single-
use takeaway containers is waste generation. In an attempt to
address this issue, consumers are being encouraged by some
groups (e.g. Takeout without, 2016;Tiffin Project Foundation, 2012)
to bring their own reusable food container from home and refuse
single-use containers used by fast-food outlets. Some companies
have also started to offer discounts to customers for bringing
reusable cups or selling its side dishes in reusable containers (e.g.
Mohan, 2010;Starbucks, 2016). In some countries, policy makers
are also considering taxation of single-use food containers. For
example, the UK government has recently issued a call for evidence
to justify the use of taxation aimed at reducing the demand for
single-use plastics, such as takeaway boxes (HM treasury, 2018).
However, while it is intuitively plausible that reusable food con-
tainers are environmentally better than their single-use counter-
parts, as far as the authors are aware, no study has yet quantified
the actual environmental implications of different takeaway food
containers and how they compare with their reusable alternatives.
Therefore, this paper assesses for the first time life cycle environ-
mental impacts of three types of container typically used by take-
away food outlets to serve hot food ealuminium, EPS and
polypropylene (PP) ein comparison with reused takeaway PP
containers and “Tupperware”food savers (also made of PP) used by
many consumers worldwide. The number of times the reusable
containers should be reused to balance out the impacts of single-
use containers is also considered.
The geographical area selected for this study is the European
Union (EU) because the number of takeaway outlets has increased
rapidly in recent years and is expected to grow even faster in the
future (Accorsi et al., 2014;Daedal Research, 2014). The UK, Ger-
many, France, Italy and Spain are among the top 13 world con-
sumers of takeaway food, with an expenditure of V9.9 billion in
2014, expected to increase by 17% by 2019 (Lago et al., 2011;Riera,
2015). The focus on the EU is also due to the European Commission
recently adopting an ambitious circular economy package, which
includes revised legislative proposals for packaging waste,
encouraging the recycling of aluminium and plastic packaging that
will also affect takeaway containers (European Commission,
2015a,b). Therefore, there is a need to quantify the impacts asso-
ciated with packaging in the takeaway food sector at EU level, as
well as the effect of future waste management legislation. Hence,
this paper also assesses the environmental effects of current and
future end-of-life scenarios for takeaway containers at EU level. The
results of this study will be of interest to EU policy makers, food
packaging manufacturers, fast-food outlets and consumers, helping
them to identify environmentally most sustainable food container
options and make informed choices.
2. Methods
The environmental impacts of the containers have been esti-
mated using LCA, which has been carried out in accordance with
the ISO 14040/44 (ISO, 2006a,b) guidelines. The next section de-
scribes the goal and scope of the study, followed by the inventory
data and the impact assessment methodology applied in the study.
2.1. Goal and scope of the study
The goals of this study are:
i) to estimate and compare the environmental impacts of three
commonly-used takeaway food containers: aluminium, EPS
and PP;
ii) to assess the environmental implications of reusing PP
takeaway containers and using reusable PP food savers
(Tupperware) instead of single-use containers; and
iii) to evaluate the environmental effects of different end-of-life
management options for the takeaway containers at the EU
level.
The functional unit for the first two goals of the study is defined
as “production, use and disposal of a container storing a meal for one
person”. The volume of the container considered is 670 ml. This
represents an average across the takeaway containers, determined
through own fieldwork and information provided by manufac-
turers. For the end-of-life options, the functional unit corresponds
to the “total number of takeaway containers used annually in the EU”.
The containers considered in the study are shown in Fig. 1.
Aluminium and EPS containers have been chosen for study as they
are used most commonly (Rieradevall et al., 2000). As shown in
Fig. 1A, in addition to the aluminium body, the aluminium
container also has a paper lid. In the EPS container, the body andthe
lid are integrated and made of a single material (Fig. 1B). Light-
weight PP containers are also used widely in the takeaway in-
dustry and they are more robust and durable than the other two
types (Fig. 1C). Consequently, even if it is usually considered a
single-use product, environmentally-conscious customers may opt
Fig. 1. Types of food containers considered in the study. A: Aluminium takeaway
container; B: Extruded polystyrene takeaway container; C: Polypropylene takeaway
container; D: Polypropylene reusable food saver (Tupperware). [Dimensions (mm):
Aluminium: 195x105x48; EPS: 170x133x75); PP (standard size: 160x105x48.].
A. Gallego-Schmid et al. / Journal of Cleaner Production 211 (2019) 417e427418

to reuse it. This is not possible for the EPS and aluminium con-
tainers as they are not sturdy enough and cannot be cleaned easily.
Nevertheless, to make PP containers comparable with the other
two types, they are assumed in the base case to be used only once
(study goal i)). The environmental implications of their reuse are
explored as part of the study goal ii), together with the use of
reusable Tupperware food savers. The latter have been selected for
study as they are used widely in Europe (Gallego-Schmid et al.,
2018). Tupperware food savers are commonly made of PP because
other common plastics have lower resistance to high temperatures
(Fellows and Axtell, 1993).
Although paper and cardboard takeaway containers are also
used, they are much less common as they are less suitable for ‘wet’
food (e.g. served with a sauce). Hence, they are not considered in
the study.
The scope of the study and the system boundaries are depicted
in Fig. 2. As shown, the following stages have been considered:
Raw materials:
o polypropylene (PP) and silicone;
o aluminium;
o polystyrene (PS); and
o cardboard, paper and polyethylene (PE) (for packaging and for
the lid of the aluminium container)
Production: thermoforming and extrusion for PP, silicone, PE
and PS, aluminium sheet rolling and metal stamping, PE coating
of paper and packaging
Use: manual and automatic washing (only for reusable
containers)
End-of-life: wastewater treatment and disposal of post-
consumer waste and;
Transport:
o packaging and raw materials to the production site;
o containers to food outlet; and
o production and end-of-life waste to waste management
facility.
For the reusable Tupperware food saver, transport of consumer
to and from the retailer to buy the container and to and from the
food outlet is excluded. The main reason is the uncertainty related
to consumer behaviour and to the allocation of the impacts to the
food container relative to other items and activities.
2.2. Inventory data
The inventory data are detailed in Table 1. These have been
calculated through direct measurements (weighing of components)
and primary production data obtained from major producers.
Fig. 2. System boundaries for the different types of container considered in the study. [PP: polypropylene; EPS: extruded polystyrene].
A. Gallego-Schmid et al. / Journal of Cleaner Production 211 (2019) 417e427 419

Ecoinvent v3.3 (Ecoinvent Centre, 2016) has been used as the pri-
mary source of background data and any data gaps have been filled
using the GaBi database (Thinkstep, 2016) and literature.
2.2.1. Raw materials and production of containers
The aluminium container is composed of the aluminium body
and a paper lid with a thin polyethylene coating on the internal
side. An average weight has been considered based on the samples
obtained from food outlets and different sellers in Europe (EAFA,
2016a). Ecoinvent data for sheet rolling for the production of the
aluminium foil and impact extrusion for shaping the container have
been used. Data for liquid packaging board have been used to
model the production of the paper lid (Vercalsteren et al., 2010;
Ligthart and Ansems, 2007;Rieradevall et al., 2000). The packaging
data for the containers have been obtained from an aluminium
container vendor (Papstar, 2016a). The containers are assumed to
be produced in China, the main producer of aluminium foil in the
World (Pani, 2015;Xie et al., 2011).
Polystyrene containers used in the takeaway industry are nor-
mally made of extruded polystyrene (Glenn et al., 2001;Marin
et al., 2004). The average weight of EPS containers has been esti-
mated based on the data obtained from food outlets and different
producers in Europe (FPA, 2016). The data for the EPS production
have been sourced from Ecoinvent while the amount of blowing
agent (butane) is from Ingrao et al. (2015). The packaging data for
the containers have been obtained from an EPS container vendor
(Papstar, 2016b). The production of EPS is assumed to be in Europe
because the leading producers are located in the Netherlands,
Luxemburg, Belgium or Germany (Plastics Europe, 2015; van der
Harst et al., 2014).
Table 1
Life cycle inventory data for takeaway containers.
Life cycle stage Aluminium
container
Extruded polystyrene
container
Polypropylene
container
Polypropylene food saver (Tupperware)
Raw materials
Polypropylene (g) ee 31.5 132.8
Aluminium (g) 7.6 eee
Polystyrene (g) e7.8 ee
Silicone (g) ee e 8.5
Paper (g) 6.6 eee
Polyethylene (lid) (g) 0.3 eee
Packaging
Cardboard (g) 1.5 1.8 2.1 7.7
Polyethylene (g) 0.06 0.07 0.09 0.3
Production
a
Extrusion: electricity (J) e28.1 57.5 258.4
Thermoforming: electricity (J) e28.0 112.6 506.1
Sheet rolling: electricity (J) 15.0 eee
Sheet rolling: heat (J) 14.5 eee
Metal stamping: electricity (J) 47.4 eee
Metal stamping: heat (J) 42.9 eee
Paper lid production: heat (J) 0.01 eee
Paper lid production: electricity (J) 4.2 eee
Packaging: heat (J) 0.05 0.06 0.07 0.3
Packaging: electricity (J) 0.7 0.9 1.1 3.8
Transport
Raw materials: transport to factory (kg
.
km) 2.4 1.5 5.0 22.4
Container: factory to Shanghai port (kg
.
km) 2.4 e5.0 22.4
Container: Shanghai to Rotterdam (kg
.
km) 314.3 e656.4 2911.2
Container: Rotterdam to Munich (kgt
.
km) 13.4 e27.9 123.8
Container: factory to Munich (kg
.
km) e4.8 ee
Container: Munich-retailer (kg
.
km) 2.4 1.5 5.0 22.4
End-of-life: transport waste treatment (kg
.
km) 0.8 0.5 1.7 7.5
Use (machine dishwashing)
Electricity (J) ee 57.5 57.5
Soap (g) ee 0.2 0.2
Salt (g) ee 0.2 0.2
Rising agent (g) ee 0.03 0.03
Water (L) ee 0.2 0.2
Use (manual washing)
Heat: electricity (J) ee 9.5 9.5
Heat: natural gas (J) ee 23.4 23.4
Heat: oil (J) ee 3.2 3.2
Heat: solar thermal (J) ee 1.0 1.0
Water (L) ee 1.4 1.4
Soap (g) ee 0.5 0.5
End-of-life waste management
Recycling: plastics (g) ee 3.4 14.4
Recycling: aluminium (g) 4.1 eee
Recycling: cardboard (g) 1.3 1.5 1.8 6.5
Incineration: plastics (g) e3.9 13.8 58.4
Incineration: cardboard (g) 3.8 0.2 0.2 0.6
Landfilling: plastics (g) 0.06 4.0 14.3 68.5
Landfilling: aluminium (g) 3.5 eee
Landfilling: cardboard (g) 3.3 0.1 0.2 0.5
a
Electricity mix in 2015: EU28 for EPS and China for the other three types of container. Heat: natural gas.
A. Gallego-Schmid et al. / Journal of Cleaner Production 211 (2019) 417e427420

For the PP containers (takeaway and reusable Tupperware food
savers), primary production data have been obtained from a major
producer, including the amount and type of raw and packaging
materials and a detailed description of production (Table 1). A sil-
icone rubber is also used in food-savers as part of the lid to ensure
tight closure. Both PP and silicone are extruded and thermoformed
to obtain the desired shape. According to Suwanmanee et al. (2013),
most of the manufacturing of plastic containers is located in South
East Asia. Therefore, the production of PP containers is considered
to be located in China and its electricity grid mix has been used in
LCA modelling.
2.2.2. Use
No impacts have been considered for the single-use takeaway
containers. For the reused PP takeaway containers and Tupperware
food savers, the impacts of their cleaning, either in a dishwasher or
by hand, have been considered. The inventory data for this are
based on Gallego-Schmid et al. (2018), recalculated for the size of
the container considered here (from 1100 to 670 ml).
2.2.3. End-of-life waste management
2.2.3.1. Single containers. The assumptions for end-of-life man-
agement of each type of the container can be found in Table 1. The
‘net scrap’approach has been applied for recycling (Bergsma and
Sevenster, 2013). This means that the system has been credited
only for the percentage of recycled material that exceeds the
recycled content in the original raw materials. For example,
aluminium is initially made up of 32% recycled and 68% virgin
metal, but 54% is recycled at the end-of-life of the container.
Therefore the system has been credited for recycling 22% of
aluminium (54% minus 32%) at the end of its life. The impact from
the recycling process are also included. The system has also been
credited for the electricity generated by incineration of waste ma-
terials. The EU electricity mix in year 2014 (ENTSO-E, 2016) has
been used for these purposes. Data for all waste and wastewater
treatments have been sourced from Ecoinvent, except for PP recy-
cling, which are from Schmidt (2012).
The following assumptions have been made for end-of-life of
different materials:
for PP, data for non-bottle rigid PP packaging treatment in the
EU28 have been considered: 44% incineration with energy re-
covery, 11% recycling and 45% for landfilling (Lh^
ote, 2011);
for aluminium, 54% of the aluminium has been considered be
recycled and 46% landfilled (Eurostat, 2016;EAFA, 2016b);
EPS containers are generally not recycled and are either land-
filled or incinerated (Belley et al., 2011;Davis and Song, 2006;
Ingrao et al., 2015). Considering the data for plastic packaging
waste treatment in Europe that is not recycled, it has been
assumed that 50% goes to landfilling and 50% to incineration
with energy recovery (Eurostat, 2016);
the strong bond between cellulose fibre board and the poly-
ethylene coating make the paper lids in the aluminium
container difficult to recycle (Mitchell et al., 2014). According to
Eurostat (2016), 54% of the paper packaging that is not recycled
is incinerated and 46% landfilled; these values have been
assumed here for the paper lid;
for the cardboard packaging of the containers, end-of-life
packaging data for the EU28 have been assumed: 85% recy-
cling, 8% incineration with energy recovery and 7% landfilling
(Eurostat, 2016); and
for the silicone (in Tupperware) and polyethylene (in pack-
aging), 100% landfilling has been assumed.
2.2.3.2. Waste management at the EU level. The European
Commission (2015b) has proposed that 75% of aluminium and
55% of plastic packaging waste should be recycled by 2025. This is
in line with the EU's approach to waste management based on the
waste hierarchy (European Parliament, 2008), which sets the
following priority order for waste: prevention, reuse, recycling,
recovery and, as the least preferred option, disposal (which in-
cludes landfilling and incineration without energy recovery). Based
on these, the following scenarios have been considered for the end-
of-life waste management of the takeaway containers:
current situation;
best case: implementation at the EU28 level of the best current
end-of-life waste management in an EU country according to
the EU waste hierarchy;
EU 2025 proposal: the implementation in EU28 of waste man-
agement proposed by the European Commission for 2025
(European Commission, 2015b); and
worst case: for comparison, the implementation at the EU28
level of the worst current end-of-life management in an EU
country according to the EU waste hierarchy.
The specific data for each container are described below and are
summarised in Table 2:
Aluminium containers: Two scenarios has been analysed with
the increase in the recycling rate from the current 54%e75% and
89%. The first increase is the EU's proposed goal for 2025
(European Commission, 2015b) and the second is the current
recycling rate of aluminium packaging in Germany, the highest
among all 10 EU countries with available data (Eurostat, 2016).
Finally, as the worst scenario, the current 22% recycling ratio of
the Czech Republic has been considered (Eurostat, 2016).
EPS containers: Two scenarios of 100% incineration with energy
recovery (currently in Austria and Sweden) and 100% landfilling
(Greece, Croatia, Cyprus, Lithuania and Malta) have been
assessed (Eurostat, 2016). A third scenario of 55% recycling and
45% incineration with energy recovery has been considered
based on the EU proposal for plastic packaging in 2025
(European Commission, 2015b).
PP containers: Similar to the aluminium containers, the best and
worst current performance at country level and the proposal for
the EU by 2025 have been considered as alternative scenarios.
According to the priority order of the EU waste hierarchy, Ger-
many has the best treatment ratios for non-bottle rigid PP
packaging (30% recycling, 69% incineration with energy recovery
and 1% landfilling) and the UK has the worst (2% recycling, 8%
incineration with energy recovery and 90% landfilling) out of the
five countries (France, Germany, Poland, Spain and UK) with
available data (Lh^
ote, 2011). The third scenario is the same as for
EPS: 55% recycling and 45% incineration by 2025, based on the
European Commission's (2015b) proposal.
Estimating the number of different types of container used in
the EU is a challenge as no specific data are available. Nevertheless,
based on data from previous studies (Lago et al., 2011;Riera, 2015),
estimates in this study
1
suggest that the total number of takeaway
food servings in the top five EU consumer countries (UK, Germany,
France, Italy and Spain) will reach 1916 million by 2025. For the rest
1
The total number of servings reported in the quoted studies is 1638 million for
the year 2014. They forecast an increase to 1760 million by 2019, a growth of 1.5%
per annum. The same annual increase rate has been applied in the current study for
the period 2019e2025, to obtain the total value of 1916 million in 2025.
A. Gallego-Schmid et al. / Journal of Cleaner Production 211 (2019) 417e427 421

of the EU countries, no data are available and hence some as-
sumptions have to be made to obtain the total number of con-
tainers in use. Taking a conservative approach, it has been supposed
that the remaining EU countries consume in total a third of the
takeaway meals consumed in the top five countries, i.e. a total of
639 million. Therefore, the total for the whole EU in 2025 has been
estimated at 2556 million takeaway food servings. Assuming an
equal share of the three types of container, the total number of each
type used annually in the EU is estimated at around 850 million.
However, this would imply that no other types of container are
used. Therefore, the analysis considers an exploratory range of
500e850 million units/yr, with an average of 675 million for each of
the three types of container. It should be noted that these figures
are estimates rather than necessarily the actual figures. Neverthe-
less, they represent best approximations available currently and are
aimed at understanding the magnitude of the impacts related to
the use of takeaway containers in the EU.
2.2.4. Transport
The following distances and transport means have been
considered:
For the raw materials and packaging, distances in Ecoinvent
have been assumed. If not specified in the database, a distance of
150 km to the factory in 16e32 t Euro 3 trucks has been
considered.
The aluminium and PP containers are shipped from China to
Europe by a transoceanic tanker, assuming a distance of
19,500 km from Shanghai to Rotterdam (World Ship Council,
2016). Afterwards, the containers are transported by 16e32 t
Euro 6 truck to a distribution centre located in Munich, the
central geographical location of Europe (829 km). Common
distances of 150 km have been assumed for transport of the
containers from the production site to the port in Shanghai
(truck 16e32 t Euro 3) and from a distribution centre to retailer
(16e32 t Euro 6 truck).
An average transport distance of 50 0 km has been assumed for
the EPS containers in Europe, from the factory to the distribution
centre in Munich (16e32 t Euro 6 truck).
For waste treatment, a distance of 50 km in a 16e32 t Euro 6
truck has been assumed for all the containers.
2.3. Impact assessment
GaBi v6.5 software (Thinkstep, 2016) has been used to model the
system. The environmental impacts have been calculated following
the CML 2001 (January 2016 version) mid-point impact assessment
method (Guin
ee et al., 2001). The following impacts are considered:
abiotic depletion potential of elements (ADP
e.
), abiotic depletion
potential of fossil resources (ADP
f.
), acidification potential (AP),
eutrophication potential (EP), global warming potential (GWP),
human toxicity potential (HTP), marine aquatic ecotoxicity poten-
tial (MAETP), freshwater aquatic ecotoxicity potential (FAETP),
ozone depletion potential (ODP), photochemical oxidants creation
potential (POCP), terrestrial ecotoxicity potential (TETP) and the
primary energy demand (PED).
3. Results and discussion
3.1. Non-reusable takeaway containers
As can be seen in Fig. 3, the aluminium container has the highest
impacts for five categories: ADP
e
, HTP, MAETP, ODP and TETP. These
are, respectively, between 2e23 and 4e28 times higher than for the
PP and EPS. The PP container is the most impactful for the other
seven impacts, largely due to the production of PP and its eventual
landfilling. These impacts are 2e3 times greater than for the
aluminium and 3e6 higher than for the EPS containers.
The results in Fig. 3 also show that the EPS is the best option
across the impact categories. Against the aluminium container, its
impacts are 7%e28 times lower and against the PP, 25% to six times
better. This is due to a lower amount of EPS needed to manufacture
the container compared to PP (four times) and less energy required
for the production of EPS in comparison with aluminium. These
findings go against the ongoing debate on the negative impacts of
EPS containers and their ban in India, China, Taiwan and in several
cities in the USA and UK, based on concerns about food health and
safety, inefficient recycling, low degradability and contribution to
marine pollution (Barnes et al., 2011). However, these bans remain
controversial and some of them have been revoked through law
suits by packaging manufacturers (Bapasola, 2015;Mueller, 2015).
Regarding health and safety, Cohen et al. (2002) found that the risk
of migration of styrene from EPS to packaged food are quite low and
of no concern. These findings are in agreement with the European
Food Safety Authority (EFSA) which stated that EPS is safe for use
for contact with food (European Commission, 2011). In relation to
recyclability, manufacturers argue that EPS is technically recy-
clable; however, in practice, due to low cost-effectiveness, the
recycling rate is negligible (Razza et al., 2015). Because of its
lightness eEPS is 95% air by weight evast amounts need to be
collected and compressed, or “densified,”before being shipped to a
recycler (MacKerron, 2015). Therefore, large investments in com-
pactors and logistical systems are necessary to achieve significant
percentages of recycling. Nevertheless, the potential benefits of
(theoretically) increasing the recycling rate of EPS at the EU level
are considered in section 0 to gauge the magnitude of environ-
mental savings that could be achieved. Finally, the low degrad-
ability and contribution to marine debris are clearly related. The
EPS containers, due to their lightness, can easily be blown away,
contributing to urban and riverine litter (Rubio, 2014). Owing to
their low degradability, EPS waste remains in the environment and
can end up in marine environments. For example, Fok and Cheung
(2015) found that 92% of the microplastics collected on 25 beaches
along the Hong Kong coastline was EPS. Aquatic organisms can
ingest these microplastics or become entangled in larger parts of
plastic (Stefatos et al., 1999;Sutherland et al., 2010) with
Table 2
End-of-life scenarios for the takeaway containers in the European Union (EU28)
a
.
Current situation Best case Worst case EU 2025 proposal
Aluminium R: 54%, L: 46%
b
R: 89%, L: 11%
b
R: 22%, L: 78%
b
R: 75%, L: 15%
d
PP R: 11%, I: 44%, L: 45%
c
R: 30%, I: 69%, L: 1%
c
R: 2%, I: 8%, L: 90%
c
R: 55%, I: 45%
d
EPS I: 50%, L: 50%
b
I: 100%
b
L: 100%
b
R: 55%, I: 45%
d
a
R: recycling, I: incineration with energy recovery, L: landfilling.
b
Eurostat (2016).
c
Lh^
ote (2011).
d
European Commission (2015b).
A. Gallego-Schmid et al. / Journal of Cleaner Production 211 (2019) 417e427422

detrimental consequences for marine life. Thus, despite its lower
life cycle environmental impacts relative to the other containers,
EPS cannot be considered a sustainable packaging option unless its
end-of-life management can be improved significantly.
3.1.1. Contribution analysis and improvement opportunities
As indicated in Fig. 3, the extraction and refining of aluminium
are the main contributors (>48%) to 11 of the 12 impacts of the
aluminium container. These are mainly associated with the gen-
eration of electricity used in the refining process (ADP
e
, ADP
f
,AP,
EP, GWP, PED) and the emissions of carbon monoxide (POCP),
hydrogen fluoride (MAETP), polycyclic aromatic hydrocarbons
(HTP) and heavy metals (FETP) from the aluminium extraction and
refining process. The production process of the PE-coated paper lid
is the main contributor to ODP (86%) due to the emissions of halon
1301 in the coating process. The stamping of the aluminium foil to
obtain the desired shape contributes 24% to both aquatic toxicities,
mainly due to the electricity consumed in the process. Finally, end-
of-life treatment has a positive impact on all impact categories
(between 2 and 23% reduction), mainly due to the system credits
associated with aluminium recycling. For further details on the life
cycle stages that contribute most to the impacts of the aluminium
container, see Table S1 the Supporting Information (SI).
The production of EPS in the raw materials stage contributes
more than half of ADP
f
, AP, GWP, POCP and PED of the EPS
container. It also has a significant contribution (>20%) to HTP,
MAETP, ODP and TETP. The consumption of fuel oil and electricity
and the emissions of CO
2
,SO
2
and non-methane volatile organic
compounds (VOCs) are the main cause of these impacts. The
extrusion and thermoforming processes in the production stage are
the main contributors to six of the impact categories (>33%), mainly
due to electricity consumption. Transport has no significant influ-
ence in any category (<9%). Finally, the end-of-life stage contributes
significantly to EP (62%) and FAETP (46%) and, to a lesser extent, to
HTP (28%), GWP (20%) and MAETP (16%). The increase of organic
carbon in the leachates associated with EPS landfilling is what leads
to the high EP and the emissions of CO
2
while heavy metals due to
incineration contribute to GWP, FAETP, HTP and MAETP.
The raw materials stage is also the main contributor (>43%) to
ADP
f
, PED, POCP and GWP of PP containers, mainly due to the
consumption of fuel oil and the emissions of VOCs and CO
2
in the PP
production process. The manufacture of the container is the main
contributor (>42%) to ADP
e
, AP and TETP. All these impacts are
mainly related to the generation of the electricity consumed in the
extrusion and thermoforming processes. Transport only contrib-
utes significantly (>35%) to ADP
e.
due to the metals used in the
battery of the trucks and ODP, related to halogenated gases used as
fire suppressants in pipelines. In the end-of-life stage, only land-
filling contributes significantly to EP (69%), FAETP (56%), HTP (39%)
and MAETP (37%). However, this stage also has a positive effect on
seven categories (5%e21%) due to the systems credits for PP
incineration and recycling.
In summary, the findings suggest that the raw materials,
container production and end-of-life contribute significantly to
most impacts. To address these, different improvement opportu-
nities could implemented. In the raw materials stage, these include
product light-weighting and increasing the recycled content in the
containers. However, manufacturers claim that the container
thickness has already been reduced significantly and further light-
weighting can compromise the function of the containers (EAA,
2016;Ingrao et al., 2015;Tupperware, 2016). Regarding the
increased recycled content, recycled plastics are not commonly
used in food packaging because of concerns about food safety and
hygiene standards (Mudgal et al., 2011). This restriction does not
apply to aluminium, but there have not been significant changes in
the amount of recycled aluminium (30%) in recent years (EAA,
2013;EAA, 2008).
There are fewer improvement opportunities in the manufacture
of containers. Production processes, such as plastics extrusion and
thermoforming or aluminium sheet rolling and stamping, are
Fig. 3. Life cycle environmental impacts of single-use containers. [All impacts expressed per 670 ml container. The values on top of each bar represent the total impact after the
recycling credits and for relevant impacts should be multiplied by the factor shown in brackets to obtain the original values. EPS: extruded polystyrene. PP: polypropylene. ADP
e
:
abiotic depletion potential of elements, ADP
f
: abiotic depletion potential of fossil resources, AP: acidification potential, EP: eutrophication potential, GWP: global warming potential,
HTP: human toxicity potential, MAETP: marine aquatic ecotoxicity potential, FAETP: freshwater aquatic ecotoxicity potential, ODP: ozone layer depletion potential, POCP:
photochemical ozone creation potential, TETP: terrestrial ecotoxicity potential, PED: primary energy demand and DCB: dichlorobenzene.].
A. Gallego-Schmid et al. / Journal of Cleaner Production 211 (2019) 417e427 423

mature and no significant advancement is expected in the future.
However, the end-of-life stage is where the most feasible oppor-
tunities for improvements are expected in the short to medium
terms for the three types of containers. These are considered for
each container at EU level in section 0. Prior to that, the impacts of
the three non-reusable containers are compared to the reusable PP
containers (takeaway and Tupperware).
3.2. Reusable vs non-reusable containers
In order to compare reusable and non-reusable containers, the
concept of “transition point”has been applied. The transition point
is defined as the point where a system starts to perform better than
the system with which it is being compared (Ligthart and Ansems,
2007). This can be determined by varying certain variables of in-
terest. In this study, the variable of interest is the number of times
both types of PP container (takeaway and Tupperware) have to be
reused to balance out the impacts of the single-use aluminium and
EPS containers.
Table 3 shows the transition points. It can be seen that the
Tupperware container has to be reused eight times to equal the
impacts of the aluminium containers for half of the impacts cate-
gories considered, 11 times for GWP and up to 16 to balance out all
the impacts. The number of reuses needed increases even more
with respect to the EPS containers. The Tupperware should be
reused 18 times to equal the GWP of EPS and 24 times to balance
out half of the impacts categories. As Tupperware containers can be
reused on average 43 times before they need to be discarded
(Gallego-Schmid et al., 2018;Harnoto, 2013), they could balance out
all the impacts of the aluminium container and most of the EPS
container. The only exception for the latter are ADP
e
, for which 208
reuses would be required and TETP, which can never be better than
for the EPS container. The reason for these two exceptions is the
high impacts from the use of electricity to heat the water for
washing the reusable containers.
The results are more favourable for the reused PP takeaway
containers which match all the impacts of the aluminium option
only after four reuses, with GWP requiring only three reuses. They
also balance out most of the impacts of the EPS containers after
nine reuses; for GWP, this reduces to four times. This level of reuse
is more realistic for consumers than the number of uses required
for the Tupperware. Furthermore, this type of plastic container is
already in use by food outlets, making it easier for businesses to
encourage their reuse, while reducing the total cost of containers
they need to purchase, improving their environmental credentials
at no extra cost and securing customer loyalty.
However, there are various practical and legal obstacles to a
widespread use of reusable containers. One of these is the incon-
venience requiring consumers to carry the container to the food
outlet and to clean it afterwards. Only committed or incentivised
(e.g. food discounts) consumers could achieve the number of reuses
necessary to balance out the impacts of single-use containers,
especially those made of EPS. Another problem is the portion sizes
ethis can be dealt with more easily if the reused containers are
provided by the food outlet but less so if consumers bring their own
reusable containers which would invariably be of different sizes.
Another issue is related to possible legal challenges, for example,
due to food poisoning. This could be either due to food contami-
nation at the outlet or due to the inadequate cleaning of the con-
tainers by the consumer, but it would be difficult to prove who is
responsible. Thus, these and other related barriers should be
considered as part of future research to identify the best way
forward.
In addition, other improvement opportunities should also be
considered. One of these includes better end-of-life management of
single-use waste containers. This is discussed further in the next
section, with the focus on the EU level to determine the magnitude
of potential environmental improvements.
3.3. End-of-life scenarios at the EU level
As indicated in Fig. 4 and Table S5 in the SI, implementing the EU
proposal for recycling 75% of aluminium and 55% of plastic pack-
aging waste by 2025 (Table 2) would reduce all the impacts by 2%
(ODP) to 60% (EP and FAETP). For GWP, the reduction would amount
to 33%, saving 61,700 t CO
2
eq./yr if a total of 2025 million/yr of
containers are assumed with an equal share of each type (45,700
and 77,300 t CO
2
eq. for 1500 and 2550 million/yr, respectively).
These savings are equivalent to the greenhouse gas emissions
generated annually by around 55,000 (40,500-69,000) light-duty
vehicles (Winkler et al., 2014).
The implementation of the “Best”scenario (highest recycling
rates practiced currently in some EU countries; see Table 2) would
improve seven impact categories on the EU 2025 proposal, but the
differences between the two scenarios are relatively small (3e18
percentage points). However the EU 2025 scenario would lead to
higher reductions in APD
f
, AP, GWP, POCP and PED. Finally, the
implementation of the worst scenario (lowest recycling in some EU
countries) would increase all impacts by 2% (GWP) to 62% (EP)
compared to the current situation.
Fig. 4 also shows the contribution of different waste manage-
ment options to the total impacts for each scenario. It can be seen
that EPS containers generally contribute the least to the total im-
pacts. Aluminium containers are the main cause of ADP
e
, HTP, ODP,
Table 3
Number of uses of polypropylene (PP) reusable containers needed to equal the impacts of single-use containers (aluminium and extruded polystyrene (EPS)).
Impact
a
PP food saver (Tupperware) vs aluminium PP food saver (Tupperware) vs EPS Reusable takeaway PP vs aluminium Reusable takeaway PP vs EPS
ADP
e.
3 208 1 32
ADP
f.
16 18 4 4
AP 8 29 2 7
EP 14 18 3 4
FAETP 12 39 3 9
GWP 11 18 3 4
HTP 2 37 1 9
MAETP 4 24 1 6
ODP 1 27 1 3
POCP 9 16 2 4
TETP 8 e
b
2e
b
PED 10 19 3 5
a
For the impacts nomenclature, see Fig. 3.
b
EPS container performs always better than the reusable containers for TETP.
A. Gallego-Schmid et al. / Journal of Cleaner Production 211 (2019) 417e427424

MAETP and TETP for all scenarios while PP containers contribute
most to ADP
f
, AP, GWP, POCP and PED; for EP and FAETP, the highest
contributor depends on the scenario.
With reference to the specific types of container, increasing the
recycling rate of aluminium would result in significant environ-
mental improvements (Table S6 in the SI). For every 10% increase in
the amount of aluminium recycled, the impacts would be reduced
by 6%e19%, with the highest reductions found for TETP (14%) and
HTP (19%). These savings are due to the credits for avoiding the use
of virgin aluminium and the reduced amount of waste landfilled.
For the GWP, increasing the current EU aluminium recycling rate
from 54% to 75%, as per the EU 2025 proposal, would save 11,600 t
CO
2
eq. per 675 million of aluminium containers annually, repre-
senting a reduction of 23% on the current situation. Increasing the
recycling to 89%, the best current recycling rate in the EU, would
reduce the impact by 38%, saving 19,400 t CO
2
eq./yr. On the other
hand, if aluminium is recycled in all EU countries at the same rate as
in the country with the lowest recycling rate (22%), the impacts
would increase by 19%e60% compared to the current situation,
with an increase in GWP of 35%.
In the case of EPS, landfilling all containers in the worst case or
incinerating them in the best scenario shows no significant effect
on the impacts compared to the current situation (Table S7 in the
SI). The only exceptions are EP, ODP and GWP. For the EP, 100%
incineration reduces the impact by 69%, mainly because of the
avoidance of the organic carbon in landfill leachates. Incineration
also reduces ODP (by 15%) due tothe credits for the avoided natural
gas displaced by the heat recovered from the incinerator. This is in
turn associated with the reduced use of halogens as fire suppres-
sants in gas pipelines. For the rest of the impacts, 100% incineration
improves the current situation by 3%e7%, except for FAETP, HTP and
TETP which are not affected by the changes in the rate of inciner-
ation (or landfilling). Finally, in the 100% landfilling scenario, GWP
is reduced by 18%, principally due to the reduction in CO
2
emissions
from EPS incineration.
Implementing the EU 2025 proposal of 55% EPS recycling and
45% incineration would improve the impacts by more than 18% for
ten of the categories considered. The highest reductions are found
for ADP
f.
, GWP, POCP and PED, the categories most influenced by
the production of EPS, benefiting from the credits for the avoided
virgin material. GWP would be reduced by 9600 t CO
2
eq./yr.
However, as discussed in section 3.1, achieving these recycling rates
will require large investment in compactors and logistical systems
as well as consumer engagement.
Finally, if in the worst case 90% of PP containers are landfilled, as
opposed to 45% at present, all the impacts but GWP are increased
(Table S8), with EP, FAETP and ODP being most affected (>31%).
GWP is reduced by 9% (8800 t CO
2
eq.) because of the reduced
emissions of CO
2
from incineration. The best case and the EU 2025
scenarios each lead to significant reductions (>52%) in EP, FAETP,
HTP, MAETP, related to the avoidance of landfilling and leachates of
heavy metals and organic carbon. The best case, with a higher
percentage of incineration (69%), has the lowest ADP
e.
, ODP and
TETP, mainly due to the credits for energy recovery. However, the
reductions in GWP are minimal (<1%) because of the CO
2
emissions
from incineration of plastics. The EU 2025 scenario, with a signifi-
cant percentage of recycling (55%), would lead to the highest re-
ductions (>38%) in ADP
f.
, EP, GWP, POCP and PED, mainly due to the
credits for the avoidance of virgin PP. For the GWP, the total
reduction of 40,500 t CO
2
eq. per 675 million containers is equiv-
alent to avoiding the annual emissions of 36,500 light-duty
vehicles.
4. Conclusions
This study has presented for the first time life cycle environ-
mental impacts of most-commonly used takeaway containers:
aluminium, EPS and PP. The results suggest that the use of
aluminium containers leads to the highest depletion of elements
and ozone layer as well as human, marine and terrestrial toxicities.
The PP container is the worst alternative for the other seven impact
categories considered.
The best option among the three is the EPS container with the
lowest impacts across the 12 categories. Against the aluminium
container, its impacts are 7%e28 times lower and against the PP,
25% to six times better. The EPS is also the best option when
compared to reusable takeaway PP containers, unless these are
reused 3e39 times, depending on the impact. The number of uses
for the reusable PP Tupperware food savers is even higher, ranging
from 16 to 208 times, with terrestrial ecotoxicity being always
Fig. 4. Annual environmental impacts of aluminium (Al), extruded polystyrene (EPS) and polypropylene (PP) containers in the EU28 considering different end-of-life scenarios.
[Basis: 675 million/yr of each type of container (total of 2025 million), with the error bars showing the range from 500 to 850 million/yr (total 1500e2550 million). R: recycling; L:
landfilling; I: Incineration with energy recovery. For the description of the scenarios, see Table 2 and for the impacts nomenclature, see Fig. 3.].
A. Gallego-Schmid et al. / Journal of Cleaner Production 211 (2019) 417e427 425

higher than for the EPS, regardless of the number of uses. Therefore,
these LCA findings show clearly that single-use plastic containers
are not necessarily the worst option environmentally as the going
debates would suggest.
However, EPS containers cause other environmental impacts
which cannot be assessed through LCA, including littering and
negative effects on marine organisms. These impacts could be alle-
viated by recycling of EPS which, although technically possible, is
negligible in the EU due to high costs. This goes against the circular
economy principles that the EU is trying to apply to packaging so that
future efforts should be focused on improving end-of-life manage-
ment of EPS containers to reduce their impacts on the environment.
The results of this research show that reductions in most impacts
greater than 18% would be possible if EPS were to be recycled in
accordance with the EU 2025 policy on waste packaging.
The impacts would also be reduced if the recycling rates for
aluminium and PP containers were increased to 75% and 55%,
respectively, as proposed by the EU. For the former, every 10% in-
crease in recycling would decrease the impacts by 6%e19%. In the
case of the PP containers, the savings would be greater than 38% for
fossil fuels, primary energy, eutrophication, photochemical oxidants
and greenhouse gases. The total reduction in the latter would
amount to 41,000 t CO
2
eq. per 675 million containers used annually.
Considering the total number of single-use containers of 2025
million used annually in the EU, implementing the EU 2020 pro-
posal for recycling of plastic packaging waste would reduce all the
impacts by 2%e60%. For GWP, the reduction would amount to 33%,
saving 61,700 t CO
2
eq./yr, equivalent to the greenhouse gas emis-
sions generated annually by 55,000 light-duty vehicles.
Future studies should assess the impacts of other materials used
for takeaway containers, such as cardboard or other types of plas-
tics (e.g. polyethylene terephthalate or polylactic acid). Further
research should also consider how principles of eco-design and
circular economy could be applied to improve the environmental
performance of fast-food containers. Studies of consumer accep-
tance of reusable containers would also help towards a more sus-
tainable use of takeaway food containers. Future work should also
work towards identifying most appropriate fiscal and other policy
instruments and incentives for reducing the use of throw-away
containers as well as for increasing their reuse and recycling.
Acknowledgments
This work was funded by the UK Engineering and Physical Sci-
ences Research Council (EPSRC, Gr. no. EEP/K011820/1) and the
Sustainable Consumption Institute at the University of Manchester.
The authors gratefully acknowledge this funding.
Appendix A. Supplementary data
Supplementary data to this article can be found online at
https://doi.org/10.1016/j.jclepro.2018.11.220.
References
Aarnio, T., H€
am€
al€
ainen, A., 2008. Challenges in packaging waste management: a
case study in the fast food industry. Resour. Conserv. Recycl. 52 (4), 612e621.
Accorsi, R., Cascini, A., Cholette, S., Manzini, R., Mora, C., 2014. Economic and
environmental assessment of reusable plastic containers: a food catering sup-
ply chain case study. Int. J. Prod. Econ. 152, 88e101.
AFCMA, 2004. Aluminium foil containers. Available at: http://www.afcma.org/
uploads/downloads/welcome_to_aluminum.pdf. (Accessed 10 September
2018).
Alupro, 2016. Aluminium packaging: foil containers. Available at: http://www.
alupro.org.uk/aluminium-packaging-foil-containers/. (Accessed 10 September
2018).
Bapasola, A., 2015. Posting the bans: can we ban EPS packaging? Available at: http://
www.isonomia.co.uk/?p¼4196. (Accessed 10 September 2018).
Barnes, M., Chan-Halbrendt, C., Zhang, Q., Abejon, N., 2011. Consumer preference
and willingness to pay for non-plastic food containers in Honolulu, USA.
J. Environ. Protect. 2, 1264e1273.
Belley, C., Michaud, R., Cl
ement, 
E., Samson, R., M
enard, J.F., 2011. Comparative Life
Cycle Assessment Report of Food Packaging Products. Report of CIRAIG (Inter-
university Research Centre for the Life Cycle of Products, Processes and Ser-
vices), Montreal (Canada), p. 47.
Bergsma, G., Sevenster, M., 2013. End-of-life Best Approach for Allocating Recycling
Benefits in LCAs of Metal Packaging. CE Delft, Delft (The Netherlands), p. 25.
Cassidy, K., Elyashiv-Barad, S., 2007. US FDA's revised consumption factor for
polystyrene used in food-contact applications. Food Addit. Contam. 24 (9),
1026e1031.
Cohen, J.T., Carlson, G., Charnley, G., Coggon, D., Delzell, E., Graham, J.D., Greim, H.,
Krewski, D., Medinsky, M., Monson, R., Paustenbach, D., Petersen, B.,
Rappaport, S., Rhomberg, L., Barry Ryan, P., Thompson, K., 2002.
A comprehensive evaluation of the potential health risks associated with
occupational and environmental exposure to styrene. J. Toxicol. Environ. Health
B Crit. Rev. 5 (1e2), 1e263.
Daedal Research, 2014. Global Takeaway Food Delivery Market - Focus on Online
Channel (2014-2019). Daedal Research, New Delhi (India), p. 97.
Davis, G., Song, J.H., 2006. Biodegradable packaging based on raw materials from
crops and their impact on waste management. Ind. Crop. Prod. 23 (2), 147e161.
EAA, 2008. Environmental Profile Report for the European Aluminium Industry.
Data for Year 2005. European Aluminium Association, Brussels (Belgium), p. 73.
EAA, 2013. Environmental Profile Report for the European Aluminium Industry.
Data for Year 2010. European Aluminium Association, Brussels (Belgium), p. 78.
EAA, 2016. Aluminium foil. Available at: http://european-aluminium.eu/about-
aluminium/aluminium-in-use/packaging/. (Accessed 31 January 2016).
EAFA, 2016a. Container group members of the European aluminium association.
Available at: http://www.alufoil.org/container-group.html. (Accessed 10
September 2018).
EAFA, 2016b. Recycling &recovery of aluminium foil containing packaging. Avail-
able at: http://www.alufoil.org/recycling-recovery-new.html. (Accessed 10
September 2018).
Ecoinvent Centre, 2016. Ecoinvent Dataset V 3.3. Swiss Centre for Life Cycle In-
ventories. Ecoinvent, Dübendorf (Switzerland).
ENTSO-E, 2016. ENTSO-E Yearly Statistics &Adequacy Retrospect 2014. European
Network of Transmission System Operators for Electricity, Brussels (Belgium),
p. 68.
European Commission, 2011. Commission Regulation (EU) No 10/2011 on plastic
materials and articles intended to come into contact with food as monomer. Off.
J. Eur. Union. L 12, 1e89, 15.01.2011.
European Commission, 2015a. Communication from the Commission from to the
European Parliament, the Council and the Committee of the Regions. Closing
the Loop - an EU Action Plan for the Circular Economy. European Commission,
Brussels (Belgium), p. 22.
European Commission, 2015b. Proposal for a Directive of the European Parliament
and to the Council Amending Directive 94/62/EC on Packaging and Packaging
Waste. European Commission, Brussels (Belgium), p. 16.
European Parliament, 2008. Directive 20 08/98/EC of the European parliament and
the Council of 19 November 2008 on waste and repealing certain directives. Off.
J. Eur. Union L 312, 3e30, 22.11.2008.
Eurostat, 2016. Packaging waste statistics. Available in: http://appsso.eurostat.ec.
europa.eu/nui/show.do?dataset¼env_waspac&lang¼en. (Accessed 21 January
2017).
Fellows, P., Axtell, B., 1993. Appropriate Food Packaging. TOOL Publications,
Amsterdam (The Netherlands), p. 138.
FPA, 2016. Members of food packaging association. Available at: http://
foodservicepackaging.org.uk/members/. (Accessed 10 September 2018).
Fok, L., Cheung, P.K., 2015. Hong Kong at the Pearl River Estuary: a hotspot of
microplastic pollution. Mar. Pollut. Bulletin. Mar. Pollut. Bull. 99 (1e2), 112e118 .
Gallego-Schmid, A., Mendoza, J.M.F., Azapagic, A., 2018. Improving the environ-
mental sustainability of reusable food containers in Europe. Sci. Total Environ.
628, 979e989.
Glenn, G.M., Orts, W.J., Nobes, G.A.R., 2001. Starch, fiber and CaCO
3
effects on the
physical properties of foams made by a baking process. Ind. Crop. Prod. 14 (3),
201e212.
Guin
ee, J.B., Gorr
ee, M., Heijungs, R., Huppes, G., Kleijn, G.R., van Oers, R.L.,
Wegener, L., Sleeswijk, A., Suh, S., de Haes, H.A. Udo, de Bruijn, H., van
Duin, H.R., Huijbregts, M.A.J., 2001. Life Cycle Assessment, an Operational Guide
to the ISO Standards. Part 2a: Guide. Kluwer Academic Publishers, Dordrecht
(the Netherlands).
Harnoto, M.F., 2013. A Comparative Life Cycle Assessment of Compostable and
Reusable Takeout Clamshells at the University of California, Berkeley. University
of California, Berkeley (USA), p. 24.
HM treasury, 2018. Tackling the Plastic Problem. Using the Tax System or Charges to
Address Single-use Plastic Waste. HM treasury, London (UK), p. 22.
Ingrao, C., Lo Giudice, A., Bacenetti, J., Mousavi Khaneghah, A., Sant'Ana, A.D.S.,
Rana, R., Siracusa, V., 2015. Foamy polystyrene trays for fresh-meat packaging:
life-cycle inventory data collection and environmental impact assessment. Food
Res. Int. 76 (3), 418e426.
ISO, 2006a. ISO14040:2006. Environmental Management - Life Cycle Assessment -
Principles and Framework. ISO standards, Geneva (Switzerland), p. 20.
ISO, 2006b. ISO14044:2006 Environmental Management - Life Cycle Assessment e
Requirements and Guidelines. ISO standards, Geneva (Switzerland), p. 46.
A. Gallego-Schmid et al. / Journal of Cleaner Production 211 (2019) 417e427426

Lago, J.A., Rodríguez, M., Lamas, A., 2011. El consumo de comida r
apida. Situaci
on en
el mundo y acercamiento auton
omico. EAE Business School, Madrid (Spain),
p. 38.
Lh^
ote, S., 2011. The Recycling and Recovery of Polyolefin Waste in Europe. Contri-
bution to Identiplast 2011 Congress, Madrid (Spain), 03rd October 2011.
Ligthart, T.N., Ansems, A.M.M., 2007. Single Use Cups or Reusable (Coffee) Drinking
Systems: an Environmental Comparison. TNO Built Environment and Geo-
sciences, Delft (The Netherlands), p. 158.
MacKerron, 2015. Waste and Opportunity 2015: Environmental Progress and
Challenges in Food, Beverage, and Consumer Goods Packaging. The Natural
Resources Defence Council and As You Sow, Oakland (USA), p. 62.
Madival, S., Auras, R., Singh, S.P., Narayan, R., 2009. Assessment of the environ-
mental profile of PLA, PET and PS clamshell container using LCA methodology.
J. Clean. Prod. 17 (13), 1183e1194.
Marin, R., Paparian, M., Moulton-Patterson, L., Peace, C., Mul
e, R., Washington, C.,
Leary, M., 2004. Use and Disposal of Polystyrene in California. Integrated Waste
Management Board, Sacramento (USA), p. 25.
Market Line, 2012. Fast Food in the United Kingdom. Market Line, London (UK),
p. 36.
Mason, I.G., Oberender, A., Brooking, A.K., 2004. Source separation and potential re-
use of resource residuals at a university campus. Resour. Conserv. Recycl. 40 (2),
155e172.
Meldrum, R.J., Little, C.L., Sagoo, S., Mithani, V., McLauchlin, J., de Pinna, E., 2009.
Assessment of the microbiological safety of salad vegetables and sauces from
kebab take-away restaurants in the United Kingdom. Food Microbiol. 26 (6),
573e577.
Mitchell, J., Vandeperre, L., Dvorak, R., Kosior, E., Tarverdi, K., Cheeseman, C., 2014.
Recycling disposable cups into paper plastic composites. Waste Manag. 34 (11),
2113e2119.
Mohan, A.M., 2010. KFC's sustainable sides container is 'sogood. Available at: http://
www.greenerpackage.com/reusability/kfc%E2%80%99s_sustainable_sides_
container_sogood. (Accessed 10 September 2018).
Mudgal, S., Lyons, L., Bain, J., Dias, D., Faninger, T., Johansson, L., Dolley, P., Shields, L.,
Bowyer, C., 2011. Plastic Waste in the Environment. BIO Intelligence Service and
AEA Technology report, Paris, p. 171.
Mueller, B., 2015. Judge strikes down New York city's ban on foam food containers.
Available at: http://www.nytimes.com/2015/09/23/nyregion/judge-strikes-
down-new-york-citys-ban-on-foam-food-containers.html. (Accessed 10
September 2018).
Pani, B.S., 2015. The economics of SME aluminium foil production. Available at:
http://blog.alcircle.com/2015/05/18/economics-sme-aluminium-foil-
production/. . (Accessed 10 September 2018).
Papstar, 2016a. Recipientes con tapas para insertar de cart
on con laminaci
on de PE
cuadrado. Available at: http://www.papstar-shop.es/Envases-para-alimentos/
Envases-de-aluminio/Recipientes-con-tapas-para-insertar-de-carton-con-
lamination-de-PE-cuadrado-0-7.htm?shop¼papstar&SessionId¼&a¼article&
ProdNr¼14517&t¼200 02&c¼20039&p¼20 039. (Accessed 10 September 2018).
Papstar, 2016b. Envases para comida con tapa-bisagra incorporada de P.S.E. sin
compartimento. Available at: http://www.papstar-shop.es/Envases-para-
alimentos/Enva ses-para-comida-para-llevar-y-Hamburguesas/Envases-para-
comida-con-tapabisagra-inc orporada-de-P-S-E-sin-compartimentos-7.htm?
shop¼papstar&a¼articleProdNr¼12043&t¼20002&c¼20335&p¼20335.
(Accessed 10 September 2018).
Potting, J., van der Harst, E., 2015. Facility arrangements and the environmental
performance of disposable and reusable cups. Int. J. Life Cycle Assess. 20 (8),
114 3e1154 .
Razza, F., Fieschi, M., Innocenti, F.D., Bastioli, C., 2009. Compostable cutlery and
waste management: an LCA approach. Waste Manag. 29 (4), 1424e1433.
Razza, F., Innocenti, F.D., Dobon, A., Aliaga, C., S
anchez, C., Hortal, M., 2015. Envi-
ronmental profile of a bio-based and biodegradable foamed packaging
prototype in comparison with the current benchmark. J. Clean. Prod. 102 (1),
493e500.
Riera, M., 2015. El gasto de comida r
apida en Espa~
na. Situaci
on internacional,
evoluci
on esperada y revisi
on de la situaci
on nacional y auton
omica. EAE
Business School, Madrid (Spain), p. 35.
Rieradevall, J., Domenech, X., Bala, A., Gazulla, C., 2000. Ecodise ~
no de envases. El
sector de la comida r
apida. Elisava Edicions, Barcelona (Spain), p. 132.
Rubio, M.R., 2014. Foam of contention: dealing with polystyrene wastes. Available
at: http://www.isonomia.co.uk/?p¼2898. (Accessed 31 January 2017).
Schmidt, J.H., 2012. Plastberegner.dk-LCA tool for plastics converters in Denmark.
In: Documentation of the Tool and Database, vol. 2. -0 LCA consultants, Aalborg
(Denmark), p. 126.
Schmitt, J., Jones, J., 2013. Slow Progress for Fast-food Workers. Available at: http://
http://cepr.net/documents/publications/fast-food-workers-2013-08.pdf.
(Accessed 31 January 2017).
Shokri, A., Oglethorpe, D., Nabhani, F., 2014. Evaluating sustainability in the UK fast
food supply chain: review of dimensions, awareness and practice. J. Manuf.
Technol. Manag. 25 (8), 1224e1244.
Starbucks, 2016. Starbucks Global Responsibility Report 2015. Starbucks, Seattle
(USA), p. 14.
Stefatos, A., Charalampakis, M., Papatheodorou, G., Ferentinos, G., 1999. Marine
debris on the seafloor of the Mediterranean Sea: examples from two enclosed
gulfs in Western Greece. Mar. Pollut. Bull. 36, 389e393.
Sutherland, W.J., Clout, M., C^
ot
e, I.M., Daszak, P., Depledge, M.H., Fellman, L.,
Fleishman, E., Garthwaite, R., Gibbons, D.W., De Lurio, J., Impey, A.J., Lickorish, F.,
Lindenmayer, D., Madgwick, J., Margerison, C., Maynard, T., Peck, L.S., Pretty, J.,
Prior, S., Redford, K.H., Scharlemann, J.P.W., Spalding, M., Watkinson, A.R., 2010.
A horizon scan of global conservation issues for 2010. Trends Ecol. Evol. 25,1e7.
Suwanmanee, U., Varabuntoonvit, V., Chaiwutthinan, P., Tajan, M., Mungcharoen, T.,
Leejarkpai, T., 2013. Life cycle assessment of single use thermoform boxes made
from polystyrene (PS), polylactic acid, (PLA), and PLA/starch: cradle to con-
sumer gate. Int. J. Life Cycle Assess. 18 (2), 401e417.
Takeout without, 2016. Takeout without campaign. Available at: http://
takeoutwithout.org/. . (Accessed 10 September 2018).
TechNavio, 2016. Global Delivery and Takeaway Food Market 2016-2020. TechNavio,
Elmhurst (USA), p. 64.
Thinkstep, 2016. GaBi 7.3 Database. Thinkstep, Leinfelden-Echterdingen (Germany).
Tiffin Project Foundation, 2012. The Tiffin Project. Available at: http://
thetiffinproject.com/. (Accessed 10 September 2018).
Tupperware, 2016. 2013-2015 Sustainability Report. Tupperware Brands, Orlando
(USA), p. 39.
van der Harst, E., Potting, J., Kroeze, C., 2014. Multiple data sets and modelling
choices in a comparative LCA of disposable beverage cups. Sci. Total Environ.
494e495, 129e143.
Vercalsteren, A., Spirinckx, C., Geerken, T., 2010. Life cycle assessment and eco-
efficiency analysis of drinking cups used at public events. Int. J. Life Cycle
Assess. 15, 221e230.
Winkler, S.L., Wallington, T.J., Maas, H., Hass, H., 2014. Light-duty vehicle CO
2
targets
consistent with 450 ppm CO
2
stabilization. Environ. Sci. Technol. 48 (11),
6453e6460.
Woods, L., Bakshi, B.R., 2014. Reusable vs. disposable cups revisited: guidance in life
cycle comparisons addressing scenario, model, and parameter uncertainties for
the US consumer. Int. J. Life Cycle Assess. 19 (4), 931e940.
World Ship Council, 2016. Top 50 world containers ports. Available at: http://www.
worldshipping.org/about-the-industry/global-trade/top-50-world-container-
ports. (Accessed 31 January 2017).
Xie, M., Li, L., Qiao, Q., Sun, Q., Sun, T., 2011. A comparative study on milk packaging
using life cycle assessment: from PA-PE-Al laminate and polyethylene in China.
J. Clean. Prod. 19 (17e18), 2100e210 6.
A. Gallego-Schmid et al. / Journal of Cleaner Production 211 (2019) 417e427 427