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Reusable plastic crate or recyclable cardboard box? A comparison of two delivery systems
Journal of Cleaner Production 69 (2014) 83e90
Contents lists available at ScienceDirect
Journal of Cleaner Production journal homepage: www.elsevier.com/locate/jclepro
Reusable plastic crate or recyclable cardboard box? A comparison of two delivery systems Sirkka Koskela a, *, Helena Dahlbo a, Jáchym Judl a, Marja-Riitta Korhonen a, Mervi Niininen b a b
Finnish Environment Institute SYKE, Centre for Sustainable Consumption and Production, PO Box 140, FI-00251 Helsinki, Finland Stora Enso Oyj, Research Centre Imatra, FI-55800 Imatra, Finland
a r t i c l e i n f o
a b s t r a c t
Article history: Received 10 January 2013 Received in revised form 9 January 2014 Accepted 12 January 2014 Available online 27 January 2014
During a product’s entire life cycle the significance of packaging varies in terms of environmental impacts. From the perspective of companies which manufacture packaging or packaging has an important role in their value chain it can be a relevant issue to focus on in their efforts to improve the environmental performance of their activities. The aim of this study was to compare the life cycle environmental impacts of a real product (bread) delivery system using either reusable HPDE plastic crates or recyclable corrugated cardboard (CCB) boxes for product transportation. In this paper we focused on the delivery systems (not the delivered product) covering the manufacturing of the crates/boxes, their use, the delivery routes from bakery to retailers and waste management/recycling of the crates/boxes. As a result we concluded that the recyclable CCB box system was a more environmentally friendly option than the reusable HPDE plastic crate system in all the studied impact categories based on the defined boundaries and assumptions. Transportation played a very important role in the environmental impacts of the analysed systems. Therefore, changes, e.g. in the weights of products and their secondary packaging or the transportation distances could affect the results considerably. Ó 2014 Published by Elsevier Ltd.
Keywords: Packaging Transportation Delivery Life cycle assessment Corrugated cardboard Plastic
1. Introduction During a product’s entire life cycle the significance of packaging varies in terms of environmental impacts. Especially with foodstuff, manufacturing of the product itself is much more resource and energy intensive than the manufacturing of its packaging (Jungbluth et al., 2000). However, from the perspective of companies which manufacture packaging or packaging has an important role in their value chain, it can be a relevant issue to focus on in their efforts to improve the environmental performance of their activities. Emissions from the production stage of packaging are not the only aspects to be considered. In delivery systems, upstream processes, transportation in the distribution network and waste management issues must also be taken into account in order to assess environmental impacts holistically. Many industrialised countries have policy frameworks and measures aiming to minimize packaging waste and their environmental impacts (e.g. Sonneweld, 2000). The measures vary from
* Corresponding author. Tel.: þ358 400 148 811; fax: þ358 9 5490 2491. E-mail address: sirkka.koskela@ymparisto.fi (S. Koskela). 0959-6526/$ e see front matter Ó 2014 Published by Elsevier Ltd. http://dx.doi.org/10.1016/j.jclepro.2014.01.045
strict regulations imposed by governments to voluntary agreements between stakeholders. According to the waste hierarchy given in the EU Waste Framework Directive (2008/98/EC), the first priority of waste management is to prevent waste from being generated. Also the European Parliament and Council Directive on packaging and packaging waste (94/62/EC, amended by the Directive 2004/12/EC) contains provisions on the prevention of packaging waste, on the reuse of packaging and on the recovery and recycling of packaging waste. Reuse of products is undoubtedly a good measure for preventing waste since it can lengthen the lifetime of a product significantly. However, when looking at the overall environmental impacts of the product system where the reusable product is included, the picture is more complex due to e.g. the transportation and washing needed in order to enable the product reuse. This emphasizes the need for comprehensive environmental assessments of product systems in order to support decision making when choosing between different types of packaging materials and products. At present, packaging is a necessary part of delivery systems. The basic packaging functions are transportation, storage and distribution (Oki and Sasaki, 2000). In general, the functions of packaging materials, such as prevention of contamination, protection
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against damages, preservation of contents and communications to the customer, are many and varied in extent and complexity (e.g., Oki and Sasaki, 2000; Pasquilo et al., 2011). In our study we focused on storage, loading and transportation functions. The transport of goods demands particular container properties related to e.g. shape, weight, hygiene, handling and labelling (Stiftung Initiative Mehrweg, 2009). Transportation itself can be a very important factor in climate change impacts caused over the life cycles of packaging (e.g. Andersson and Ohlsson, 1999; Sim et al., 2007; Pretty et al., 2005; Meisterling et al., 2009). The importance of transport modelling in LCA has been well known for long (Jørgensen et al., 1996). According to Jørgensen et al. (1996), transport contributes to 5e15% of the major environmental impacts of a life cycle of a product. According to Madival et al. (2009), transport may contribute significantly to the environmental impacts of agricultural products and Gunady et al. (2012) emphasise especially the effects of a long distance transport. In recent decades, there has been an ongoing trend to find new materials based on biomass or renewable resources to replace nonrenewable materials, e.g. petroleum-based plastics (e.g. Madival et al., 2009). Several current policies (e.g. “Thematic Strategy on the Sustainable Use of Natural Resources COM/2005/670, An EU Strategy for Biofuels” COM/2006/34, A resource-efficient Europe e Flagship initiative under the Europe 2020 Strategy COM/2011/21), which aim to achieve a more resource-efficient economy, support the increased use of materials made from renewable resources. Products made from renewable materials are not automatically a better choice over ones made from non-renewable materials, since the whole supply chain from extraction to end-of-life must be considered. There can be aspects, e.g. reusing or recycling of packaging, which could change the ranking. Many comparison studies of packaging systems have been accomplished (e.g. Ross and Evans, 2003; Lee and Xu, 2004; Singh et al., 2006; Raugei et al., 2009). In all of them reusable plastic containers proved to be a better choice compared to single-use packaging. This bears out the general conception that reuse is always better than recycling. But is it true in every case? In our study we compared the life cycle environmental impacts of a real life delivery system using HDPE plastic crates or CCB boxes for transportation of the delivery product. We assessed the impacts of two delivery systems, one using a crate made of non-renewable plastic and a box made of renewable CCB. Both materials have advantages and disadvantages in terms of environmental impacts. Plastic crates are durable and washable, hence they can be reused several hundred times (in our study approximately 700 times) before finally being recovered as material for new plastic products or as energy. CCB boxes can be designed to be strong but light, and although they can only be used once, they can be recovered and used in the production of new fibre products or as energy. Levi et al. (2011) compared plastic containers and corrugated boxes to each other in Italian fruit distribution. They concluded that emissions from the manufacturing of corrugated box were greater than those from manufacturing plastic crates and the importance of transportation was identified in the environmental impacts of the distribution systems. The study of Stiftung Initiative Mehrweg (2009) presents the results of a comparison of fruit delivery systems in some European countries and Singh et al. (2006) in Northern American market finding the plastic container system better than the CCB system. However, these studies cannot be compared to our study as such due to several differences in the modelling assumptions. The greatest differences existed in e.g. the material composition of the crates/boxes, their weights, the number of circulations and transportation parameters. Additionally, end-of-life phases deviated from each other for both plastic crates and CCB boxes.
The scope of the comparison is not the use of secondary packaging needed for the delivery system, but the delivery system of packaged bread using CBB and HDPE plastic crates. The aim was to compare the life cycle environmental impacts of a real delivery system using either reusable HDPE plastic crates or recyclable CCB boxes for product transportation. The delivered product was toast bread which is a light weighted packed daily foodstuff delivered to the whole Finland. The results do not include the processes related to bread baking and its upstream, because the delivered product is not in the focus of this study. The weight of bread is, however, taken account in the impacts of transportation. The study was implemented in cooperation with the leading bakery company in Finland (VAASAN Oy) and with a global manufacturer of biomaterials, paper, packaging and wood products (Stora Enso Oyj). Both provided data and valuable insights from a business perspective for the study. A critical review of the study was conducted by the Swedish Environmental Research Institute (IVL). 2. Materials and methods 2.1. Life cycle assessment and data sources In order to achieve more sustainable production patterns, the environmental implications of the whole supply chain of products (both goods and services), their use, and waste management (ILCD, 2010) must be considered. Life cycle assessment (LCA) studies thereby help to avoid resolving one environmental problem while creating others, avoiding so called “shifting of burdens”. Life cycle assessment (LCA) is a method for integrating the environmental impacts of a studied product or a service over the whole value chain. It is an internationally standardized method (ISO 14040, 14044) with comprehensive guidelines (ILCD, 2010). In full LCA all processes and flows are followed from cradle-to-grave (i.e. from resource extraction to waste disposal) taking into account all relevant environmental impact categories. The goal of the study was to compare the life cycle environmental impacts of two different product systems for bread delivery from the bakery to consumers. The main difference in the systems was the type of material used for the delivery crates, either plastic or CCB, which generated differences in, among others, manufacturing and transportation (Fig. 1). The product systems (referred to as plastic crate system and CCB box system) included the life cycles of manufacturing of the crates/boxes from virgin materials and the delivery system of bread. The study tried to establish which container material would be more favourable from an environmental perspective in this specific distribution system. The weight of one plastic crate is 1.450 g with inside dimensions of 560 360 125 mm. It is made of high-density polyethylene (HDPE). The CCB box weighs 190 g and its dimensions are 540 330 110 mm. The bread delivered is toast bread. The weight of an average loaf of bread is 340 g (2.720 g in one crate/box). The weight of one plastic bag used for the bread packaging is 2 g (16 g in one crate/box). The different dimensions of a crate/box indicate slightly different capacities. However, all the crates/boxes hold the same load, 8 loaves of bread, therefore they perform the same function in the studied systems. Collected inventory data consisted of primary data from the participating companies, e.g. data related to the manufacturing of CCB boxes, transportation distances and modes and the washing process for crates. The washing of crates, but not the consumption of tap water (as a resource), was included in our assessment. The washing process also requires energy for heating the water and for the washing process, impacts of which are included in the study. The washing mainly removes dust from the crates and the detergents used for washing do not include phosphorus. Generic data
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Fig. 1. The studied product systems. Life cycle impacts of bread baking were excluded from the results shown in this paper.
from the ecoinvent v. 2.2 database were used for the manufacturing of plastic crates, washing chemicals and energy production. Emissions from transportation were calculated according to the Finnish Lipasto database (VTT, 2009) and the ecoinvent v2.2 database. Data for recovery processes in waste management were based on studies published by Korhonen and Dahlbo (2007) and Myllymaa et al. (2008). The impact categories analysed in the life cycle impact assessment (LCIA) were selected based on data availability and their relevance to the studied product systems: climate change, terrestrial acidification, photochemical oxidant formation, particulate matter formation, and fossil depletion. These impact categories indicate well the impacts of transportation and energy use for manufacturing processes. In addition, fossil depletion is connected to plastic production. Freshwater eutrophication was selected because cardboard production causes nutrient releases into water. Land use issues are very important aspects considering impacts of renewable resources. We did not obtain data for land use, therefore it is not considered in our study. Additionally, consensus of the methods for calculating land use has not been reached. For calculating the environmental impacts of the systems, characterisation and normalisation factors (European reference area) were taken from the ReCiPe (2011) midpoint (hierarchist) method, which has been developed for life cycle impact assessments (Goedkoop et al., 2009). ReCiPe was selected, because it’s one of the most up-to-date LCIA methods, currently. It is also under the continuous development. The data in this study are subject to uncertainties, which are very common in all LCA studies related to lacking specific data and model uncertainties (see e.g. Heijungs and Huijbregts, 2004; Guo and Murphy, 2012; Mattila et al., 2012). The model uncertainties are similar in both systems, but uncertainties related to parameters partly differ. No uncertainty analysis was conducted in this study, but three sensitivity analyses were done. Since transportation was
found to be the main contributor to the overall environmental impacts of the system, a sensitivity analysis was performed for transport distances. Two additional sensitivity analyses concerned the number of uses of the plastic crates and different allocation methods for calculating the benefits of CCB recycling. 2.2. Description of the compared product systems In life cycle comparisons, the boundaries of product systems must be consistent (ILCD, 2010). Therefore the plastic crate and CCB box systems had exactly the same boundaries including many equal components such as delivery routes from bakery to retailers and primary bread packaging (plastic bag). The systems differed in the manufacturing of crates/boxes, their use, transportation impacts in delivery (crate collection and take-back) and waste management/ recycling of the crates/boxes (Fig. 1). The function of the studied systems was to distribute bread from bakery to consumers. The functional unit of the product systems was 8 loaves of bread delivered in one crate/box. However, the study focused on the comparison of the two different crate/box materials and unit processes related to them, hence bread baking with its upstream were excluded from the results presented in this paper. 2.3. Transport modelling In plastic crate system the plastic crates were manufactured in Finland and transported to Tallinn, Estonia, where the bread was baked. In CCB box system the sheets of CCB were manufactured in Latvia where they were also cut into individual box sheets known as CCB blanks and transported to Tallinn. At the bakery, the boxes were assembled from the blanks in the box forming machine. Packed bread was then transported in crates or boxes from Tallinn (via Helsinki) to the main distribution centre in Eastern Finland. From there it was delivered to local distribution centres and then to
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Table 1 Transport distances and modes in bread delivery and crate/box collection. Load
Route
Distance (km)
Mode of transport
Hdpe granulate CCB blanks Crate
Production site / crate manufacturing site Production site (Latvia) / bakery Production site / bakery
Bread
Bakery / port of Tallinn Tallinn / Helsinki Helsinki / main distr. Centre Main distr. Centre / local distr. Centre Local distr. Centre / retail Same distances and transport modes as for bread delivery Same distances and transport modes as for bread delivery Retail / incineration/recycling Retail / recycling
140 310 100 80 10 10 80 130 333 162
Lorry 16e32 t (ecoinvent) Lorry 16e32 t (ecoinvent) Semi-trailer combination (hwy) RORO ship Semi-trailer combination (urb) Semi-trailer combination (urb) RORO ship Semi-trailer combination (hwy) Full trailer combination (hwy) Heavy delivery lorry (deli)
150 250
Heavy delivery lorry (deli) Heavy delivery lorry (deli)
Crate backhaul Empty runs (CCB case) Obsolete crate collection CCB collection
Note: symbols hwy, urb and deli stand for highway, urban and delivery driving conditions, respectively. The delivery driving mode has a defined highway mileage share which accounts for 30%.
retailers. From retailers empty crates were transported back to the main distribution centre for washing and then back to the bakery in Tallinn where they were reused. Obsolete plastic crates were collected and transported to energy recovery and recycling plants. Empty CCB boxes were collected and transported to recycling plants (Table 1). 2.3.1. Road freight transport Typically, transport impacts in LCA are calculated so that a mass of transported load (in tonnes) is multiplied by a distance (in kilometres) and this value (in tonne-kilometres) is linked to an inventory of a unit process representing the transport of 1 t-km using a specific mode of transport. LCI of road freight transport unit processes in major LCI databases, such as the ecoinvent v2.2 (Ecoinvent, 2010), are inventoried for trucks of a specific tonnage, for an average annual mileage and an average load (Spielmann and Scholz, 2005). The average values usually represent the situation in Europe. Although commonly used, these datasets are not suitable for every case study. In some products volume is the limiting factor for transportation, known as volume-limited. These are typically light products large in volume, such as empty plastic crates or other light three-dimensional objects. After the first LCIA calculations it became clear that the studied system is sensitive to the way transport is accounted for. According to the ILCD Guidelines (2010), more specific transport modelling had to be implemented. For this study the Finnish calculation system for traffic exhaust emissions and energy consumption, Lipasto (VTT, 2009) was chosen as the most appropriate methodology and database. Lipasto
enables calculation of the unit emission profile and fuel consumption for defined trucks of a specific load, which gives an advantage over generic LCI databases by providing more casespecific results. Emission data for a 2010 fleet average were used in the calculation. In order to account for the upstream environmental impacts of transportation we combined Lipasto with ecoinvent v2.2 (production and distribution of diesel fuel, manufacturing, maintenance and end-of-life of a vehicle and construction, maintenance and end-of-life of road infrastructure). 2.3.2. Sea transport The calculation of specific RORO (roll-on/roll-off) ship emissions is not applicable for trailers of different loads. Thus there was no difference when accounting for sea transport between the compared systems. The only difference is a justified assumption that 20% of otherwise empty trucks on the inbound route of CCB box system are utilised by another product system. These are therefore excluded from the system boundary. Unit emissions of the RORO ship were treated in the same way as those of the vehicles. They were combined with ecoinvent processes for fuel production, barge manufacture, maintenance and end-of-life. 2.3.3. Calculation approach The Lipasto methodology (VTT, 2009) features an equation (Eq. (1)) which was used in the calculation.
ex ¼
ea þ
eb ea lx lc
1 lx
(1)
Table 2 Weights of loads for distribution and return transportation. Vehicle type
Palettes per vehicle
Specific load (in tonnes) HDPE crate system Distribution
Semi-trailer combination Gross vehicle mass 40t Payload capacity 25t Full trailer combination Gross vehicle mass 60t Payload capacity 40t Heavy delivery lorry Gross vehicle mass 15t Pay load capacity 9t
CCB box system Return and collection
Distribution
Return and collection
26 Palettes
8.75
3.02
6.12
Empty
41 Palettes
13.79
4.76
9.66
Empty
3.15
1.08
2.21
Empty (return) 4.5 (collection)
Approx. 9 pallets
Note: distribution means the traffic from bakery to retailers (loaded with bread and crates/boxes). In the return and collection trips only empty crates (collected CCB boxes for recycling) are transported. The same degree of loading as for return and collection trips in the plastic crate system was used in crate transport from the production site to the bakery.
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The symbol ex in the equation above represents tonne-kilometre emissions (g/t-km) of a vehicle with user-defined payload lx (t). The symbol lc stands for payload capacity (t) of the vehicle; eb represents vehicle-kilometre emissions (g/v-km) of the vehicle when transporting lc and ea represents vehicle-kilometre emissions (g/vkm) of the vehicle when empty. The values of ea and eb were obtained from the Lipasto database. Inventory included emissions of CO, HC, NOx, PM, CH4, N2O, NH3, SO2 and CO2 as well as fuel consumption of the vehicle. The maximum load of each truck was calculated as the sum of the masses of crates, bread and bread packaging. Transportation EUR pallets were assumed to be an integral part of a truck and therefore were not accounted for as part of the load. It was calculated that one EUR-palette fits 80 crates/boxes. Based on that figure, specific loads of different trucks were calculated (Table 2). These were used for calculating unit emission profiles in urban and highway driving mode. Unit emissions for delivery driving were calculated for the last step in the bread distribution. The selected driving mode for each route is specified in Table 1. We calculated LCI of 1 tonne-km for each truck loaded to its maximum volume with the transported product (packed breads) in the transportation containers (CCB box, plastic crate). We calculated LCIA results by multiplying this LCI by the mass of transportation containers only. 2.4. Benefits from reuse, recycling and energy recovery After a fibre-based product has been used, it can be recycled. There are several ways of calculating the benefits of recycling. We studied how the impacts of climate change would vary using openloop allocation (ISO/TR 14049), monetary allocation and the system expansion approach to the CCB Box System. The results did not differ significantly from each other (less than 1%). We chose to use the system expansion (or avoided emissions) approach for this comparison study. In Finland, CCB is typically recycled for use in coreboard manufacturing. In practice, coreboard is always manufactured from recycled fibres, and virgin fibres are used only in minor quantities or not at all. However, in our study, we assumed that if virgin fibres were used they would be virgin fluting, which due to its fibre properties is the most suitable for coreboard manufacturing. Thus the emissions avoided by recycling were calculated as the difference between the emissions of manufacturing coreboard with virgin fluting and the emissions of manufacturing coreboard using recycled CCB. The raw materials for coreboard used in the calculations were 83% recovered fibre and 17% virgin fibre (data obtained from the participating forest company). The virgin fibre input is needed to compensate the loss of fibre strength during recycling. The plastic crates can also be recovered as material or energy at the end-of-life phase. When the crates were returned to the main distribution centre for washing, obsolete and broken crates were separated from the reusable ones. Approximately 20% of the obsolete crates were recovered as material in the production of plastic profiles, which is a process using e.g. HDPE as raw material (based on information from the bakery company). Plastic profiles can be used e.g. in patio constructions instead of impregnated wood. Therefore the production of impregnated wood was considered to be the process that can be avoided by plastic recycling (Korhonen and Dahlbo, 2007). In addition, around 80% of the obsolete crates were recovered as energy using in a boiler. The type of boiler used could not be specified; hence we assumed that the combustion of one kg of plastic produced 33 MJ of heat (based on the lowest heating value of plastic waste, Statistics Finland, 2011). This heat was assumed to replace separate production of
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heat for which an emission factor of 62.77 kg CO2/GJ was used (average emission factor for separate production of heat, Motiva, 2004). In our product system, the benefits of reusing plastic crates were considered by only including the amount of emissions that can be associated with one circulation of the crate. The number of circulations of one crate was calculated from the average lifetime of one crate (13.75 years), the number of circulations of one crate per year (61.54) and the duration of one circulation (4.87 days) (data obtained from the participating bakery company). For the specific route modelled in our system the circulation of a crate took one day longer than for the average route. Taking this into consideration, we estimated 700 circulations for each plastic crate. Hence we assumed that the distribution of 8 loaves of bread delivered in one crate represented 1/700 of the emissions of manufacturing one plastic crate. Likewise, the benefits of material and energy recovered at the end-of-life of the crate were assessed for 1/700 of a crate. 3. Results and discussion 3.1. Environmental impacts 3.1.1. Overall impacts This LCA study showed that for delivering 8 loaves of bread in one container, CCB box system was a more environmentally friendly option than plastic crate system in all impact categories based on the defined boundaries and assumptions (Fig. 2, Table 3). 3.1.2. Transportation In the delivery phase of the studied systems, transportation was the most significant contributor to all studied environmental impacts. For example the overall climate change impact related to transportation was 1.144 kg CO2-eq in plastic crate system and 0.757 kg CO2-eq in CCB box system. Distances, modes of transportation and particularly load were the most important factors in terms of transportation impacts. The greatest differences in the impacts of transportation between the systems were caused by the different weights of the crates/boxes and by the circulations of plastic crates. It should be noted that the delivery network in the studied systems covered the whole of Finland where the distances were very long. Local bakeries would decrease the amount of transportation, but such a decentralised system would have different impacts and the overall outcome cannot be evaluated without a comprehensive analysis. In the use phase (i.e. delivery network) CCB boxes performed better, because plastic crates need washing after every use which causes impacts on the environment. However, the significance of washing on the total impacts was very low. 3.1.3. Manufacturing of crates/boxes The environmental impacts of manufacturing one HDPE plastic crate were higher than those of one CCB box, but the fact that crates were reused hundreds of times decreased the impacts significantly (Fig. 3). As a result of the reuse of crates, the impacts from manufacturing a plastic crate were lower than those of a CCB box. In our study, however, the recycling of CCB to coreboard production changed the overall impacts more in favour of CCB. Decreasing, or increasing, the number of times the plastic crates are used had only a very small impact on the overall results of the system (unless the number of uses is extremely low). For illustrating this, the results of climate change were calculated with the number of uses of the crates ranging from 10 to 800. The results showed that for a range of uses from 10 to 100 times the impacts of manufacturing decreased notably, but there after the differences
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Fig. 2. Climate change (CC), terrestrial acidification (TA), freshwater eutrophication (FE), photochemical oxidant formation (POF), particulate matter formation (PMF) and fossil depletion (FD) according to the life cycle stages of the systems analysed and not including impacts of bread baking with its upstream. FU ¼ 8 loaves of bread delivered in one crate/ box. Note: benefits of recovery of plastic crates by incineration are not visible due to the minor relevance.
were not significant (Fig. 3). In the range of hundreds of uses the impacts from crate manufacturing per one use were minor.
Table 3 Environmental impacts of the analysed systems. Impact category
Unit
Plastic crate system
CCB box system
Climate change Terrestrial acidification Freshwater eutrophication Photochemical oxidant formation Particulate matter formation Fossil depletion
kg CO2-eq kg SO2 eq kg PO4 eq kg NMVOC
1.18Eþ0 7.90E3 4.15E4 1.02E2
8.76E1 6.24E3 3.79E4 7.40E3
kg PM10 eq kg oil eq
3.14E3 4.29E1
2.37E3 3.14E1
Therefore the manufacturing of the crate played a very small role in the system as a whole.
3.1.4. End-of-life phases The benefits of the recovery of plastic crates were very low (per functional unit) with the defined end-of-life assumptions (20% material recovery, 80% energy recovery based on the current situation) (Fig. 2) and therefore they are not visible as negative values in the results. The optimal end-of-life phase for plastic would be 100% material recovery to compensate for virgin plastic as has been shown by, e.g. Lazarevic et al. (2010). However, since only 1/700 of a plastic crate (which is equal to 2.07 g) was allocated per a functional unit, changes in the benefits would not affect the overall result of our study.
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Fig. 3. Climate change impacts of crate manufacturing (3.99 kg CO2-eq/crate) allocated per number of circulations.
The benefits of recycling CCB boxes are shown as negative values originating from the avoided emissions of using recycled CCB instead of virgin fibre in coreboard manufacturing (Fig. 2). Also in the future, demand for recycled corrugated boxes is expected to be high, due to continuous demand for recycled fibres. Corrugated boxes from retailers and industry are very often collected separately, which ensure the high purity and quality of fibres. CCB recycling into coreboard manufacturing generated the greatest benefits in the climate change impact category (Fig. 2). Without the benefit (0.106 kg CO2 -eq) the result for CCB box system would be 12% higher, but still lower than climate impact of plastic
CC
TA POF PMF FD 0.00E+00
2.00E-13
4.00E-13
6.00E-13
normalised results corrugated cardboard box (CCB) plastic crate Fig. 4. Normalised results of impact categories. FU ¼ 8 loaves of bread delivered in one crate/box.
crate system. Carbon sequestration in forests is an important part of the life cycle of fibre-based products. Taking this into consideration would increase the benefits of recycling in CCB Box System. There are models attempting to link carbon sequestration in forests to product carbon footprints but they contain high uncertainties. No scientific consensus has yet been reached on how to incorporate CO2 removal by forests into product specific assessments. Therefore a conventional approach was applied in this study and no benefit from carbon sequestration was allocated to CCB boxes.
3.2. Normalised results Normalisation is a means to present LCIA results with more comprehensible values than the impact category indicator scores and this also made it possible to make comparisons between the contributions of impact categories to the reference system (e.g. Dahlbo et al., 2013). In this study, European reference values were used for each impact category. Since weighting between impact categories was not done, the normalised impact category results can be considered to be of equal importance from the perspective of environmental protection within the region considered in the normalisation reference values (Europe) and the harmfulness of one category over another cannot be evaluated (Dahlbo et al., 2013). The normalised results indicated that plastic crate system made a greater contribution to the impacts in Europe than CCB box system (Fig. 4). For both systems, the lowest contribution was to climate change impacts and the highest to fossil depletion. For CCB box system the contribution to the particulate matter formation was higher than to the terrestrial acidification and the photochemical oxidant formation. In contrast, for plastic crate system the contributions to particulate matter formation and photochemical oxidant formation were higher than the contribution to terrestrial acidification.
Fig. 5. Sensitivity analysis of transport distances on freshwater eutrophication.
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3.3. Sensitivity analysis on transport distances Transport clearly dominates the overall results for each of the analysed impact category. Thus we applied a sensitivity analysis on transport distances in order to analyse the robustness of our findings. We varied all distances except the ferry transport, which would in any case remain the same. Two alternative scenarios were calculated, one for distances half shorter and one for distances twice longer. Due to linearity of the model the main conclusions did not change for most impact categories. Except for freshwater eutrophication (Fig. 5), where the contribution of the CCB manufacturing became more significant for the shorter distances scenario and it outbalanced lower transport impacts caused by lighter CCB box, compared to the plastic crate. It is important to notice that the impacts for shorter distances are more uncertain than ones for longer distances, because transport equipment would likely differ from the assumptions made in the study (e.g. smaller trucks for local delivery). 4. Conclusions In this LCA comparison study two bread delivery systems were examined. In the first system the product was delivered in HDPE plastic crates and in the second one in CCB boxes. As a result we can conclude that the CCB box system was a more environmentally friendly option than the plastic crate system in all studied impact categories based on the defined boundaries and assumptions. Our study proved that a conclusion on which delivery system has more favourable environmental impacts cannot be made based on the container material only. Before decision making, the whole delivery system including all necessary processes must be assessed. In general, a long-lasting, reusable product is considered to be a better choice, but as our study proved a recycled product can also be a good option but it requires a profitable and effective recycling system. Transportation played a very important role in the environmental impacts of the analysed systems. However, changes, e.g. in the weights of products and their secondary package or the transportation distances could affect the results considerably. Regardless of the size of the distribution area it is important to develop logistics and also vehicles further in order to decrease environmental impacts of delivery systems. Acknowledgements This study was conducted as part of the MMEA research programme managed by the Cluster for Energy and Environment and funded by TEKES e the Finnish Funding Agency for Technology and Innovation, Stora Enso Oyj and the Finnish Environment Institute (SYKE). A critical review of the study was conducted by the Swedish Environmental Research Institute (IVL). References Andersson, K., Ohlsson, T., 1999. Life cycle assessment of bread produced on different scales. Int. J. Life Cycle Assess. 4 (1), 25e40. Dahlbo, H., Koskela, S., Pihkola, H., Nors, M., Federley, M., Seppälä, J., 2013. Comparison of different normalised LCIA results and their feasibility in communication. Int J Life Cycle Assess. 18, 850e860. http://dx.doi.org/10.1007/s11367012-0498-4. Ecoinvent Database v.2.2, 2010. Swiss Centre for Life Cycle Inventories. http://www. ecoinvent.ch. Goedkoop, M.J., Heijungs, R., Huijbregts, M., De Schryver, A., Struijs, J., Van Zelm, R., 6 January 2009. ReCiPe 2009. A Life Cycle Impact Assessment Method Which Comprises Harmonised Category Indicators at The Midpoint and The Endpoint Level, first ed. Report I: Characterisation. http://www.lcia-recipe.net (accessed 20.05.11). Gunady, M.G.A., Biswas, W., Solah, V.A., James, A.P., June 2012. Evaluating the global warming potential of the fresh produce supply chain for strawberries, romaine/
cos lettuces (Lactuca sativa), and button mushrooms (Agaricus bisporus) in Western Australia using life cycle assessment (LCA). J. Clean. Prod. 28, 81e87. Guo, M., Murphy, R.J., 2012. LCA data quality: sensitivity and uncertainty analysis. Sci. Total Environ. 435e436, 230e243. Heijungs, R., Huijbregts, M.A.J., 2004. A review of approaches to treat uncertainty in LCA. In: Proceedings of the Second Biennial Meeting of IEMSs, Complexity and Integrated Resources Management, 14e17 June 2004, Osnabrück, Germany, pp. 332e339. ILCD Handbook. International Reference Life Cycle Data System, 2010. European Commission, Joint Research Centre. http://lct.jrc.ec.europa.eu/pdf-directory/ ILCD-Handbook-General-guide-for-LCA-DETAIL-online-12March2010.pdf (accessed 04.12.12). ISO 14040, 2006. Environmental Management e Life Cycle Assessment e Principles and Framework. International Organization for Standardization. ISO 14044, 2006. Environmental Management e Life Cycle Assessment e Requirements and Guidelines. International Organization for Standardization. Jungbluth, N., Tietje, O., Scholz, R.W., 2000. Food purchases: impacts from the consumers’ point of view investigated with a modular LCA. Int. J. Life Cycle Assess. 5 (3), 134e142. Jørgensen, A.-M., Ywema, P., Frees, N., Exner, S., Bracke, R., 1996. Transportation in LCA. Int. J. Life Cycle Assess. 1 (4), 218e220. Korhonen, M.-R., Dahlbo, H., 2007. Reducing greenhouse gas emissions by recycling plastics and textiles into products. In: The Finnish Environment, 30/2007. http://www.ymparisto.fi/default.asp?contentid¼249187&lan¼en (accessed 17.12.11). Lazarevic, D., Aoustin, E., Buclet, N., Brandt, N., 2010. Plastic waste management in the context of a European recycling society: comparing results and uncertainties in a life cycle perspective. Resour. Conserv. Recycl 55, 246e259. Lee, S.G., Xu, X., 2004. A simplified life cycle assessment of re-usable and single-use bulk transit packaging. Packag. Technol. Sci. 17, 67e83. Levi, M., Cortesi, S., Vezzoli, C., Salvia, G., 2011. A comparative life cycle assessment of disposable and reusable packaging for the distribution of Italian fruit and vegetables. Packag. Technol. Sci. 24, 387e400. Madival, S., Auras, R., Singh, S., Narayan, R., 2009. Assessment of the environmental profile of PLA, PET and PS clamshell containers using LCA methodology. J. Clean. Prod. 17, 1183e1194. Mattila, T., Leskinen, P., Soimakallio, S., Sironen, S., 2012. Uncertainty in environmentally conscious decision making: beer or wine? Int. J. Life Cycle Assess. 17 (6), 696e705. Meisterling, K., Samaras, C., Schweizer, V., 2009. Decisions to reduce greenhouse gases from agriculture and product transport: LCA case study of organic and conventional wheat. J. Clean. Prod. 17, 222e230. Motiva, 2004. Yksittäisen Kohteen CO2-päästöjen Laskentaohjeistus Sekä Käytettävät CO2-päästökertoimet (Guidance and CO2-emission factors for calculating CO2-emissions from a single object) (in Finnish). Myllymaa, T., Moliis, K., Tohka, A., Rantanen, P., Ollikainen, M., Dahlbo, H., 2008. Jätteiden Kierrätyksen Ja Polton Käsittelyketjujen Ympäristökuormitus Ja Kustannukset. Inventaarioraportti (Environmental loads and costs of waste recycling and incineration processes. Inventory report). Rep. Finnish Environ. Inst. 28/2008. Oki, Y., Sasaki, M., 2000. Social and environmental impacts of packaging (LCA and assessment of packaging functions). Packag. Technol. Sci. 13, 45e53. Pasquilo, J., Meneses, M., Castells, F., 2011. The carbon footprint and energy consumption of beverage packaging selection and disposal. J. Food Eng. 103, 357e 365. Pretty, J.N., Ball, A.S., Lang, T., Morison, J.I.L., 2005. Farm costs and food miles: an assessment of the full cost of the UK weekly food basket. Food Policy 30 (9), 1e 19. Raugei, M., Fullana-i-Palmer, P., Puig, R., Torres, A., 2009. A comparative life cycle assessment of single-use fibre drums versus reusable steel drums. Packag. Technol. Sci. 22, 443e450. ReCiPe, 2011 (accessed 04.12.11). http://www.lcia-recipe.net/. Ross, S., Evans, D., 2003. The environmental effect of reusing and recycling a plasticbased packaging system. J. Clean. Prod. 11, 561e571. Sim, S., Barry, M., Clift, R., Cowell, S.J., 2007. The relative importance of transport in determining an appropriate sustainability strategy for food sourcing. Int. J. Life Cycle Assess 12 (6), 422e431. Singh, S.P., Chonhenchob, V., Singh, J., 2006. Life cycle inventory and analysis of reusable plastic containers and display-ready corrugated containers used for packaging fresh fruits and vegetables. Packag. Technol. Sci. 19, 279e293. Sonneweld, K., 2000. The role of life cycle assessment as a decision support tool for packaging. Packag. Technol. Sci. 13, 55e61. Spielmann, M., Scholz, R.W., 2005. Life cycle inventories of transport services. Int. J. Life Cycle Assess 10 (1), 85e94. Stiftung Initiative Mehrweg, 2009. The Sustainability of Packaging Systems for Fruit and Vegetable Transport in Europe Based on Life-cycle Analysis e Update 2009. http://www.pe-international.com/company/newsroom/news/2009/news2009-detail/article/update-2009-the-sustainability-of-packaging-systems-forfruit-and-vegetable-transport-in-europe-ba/ (accessed 28.12.12). Statistics Finland, 2011. Fuel Classification and Emission Coefficients. http://www. stat.fi/tup/khkinv/khkaasut_polttoaineluokitus.html (accessed 17.12.11). VTT, 2009. LIPASTO e A Calculation System for Traffic Exhaust Emissions and Energy Consumption in Finland. http://lipasto.vtt.fi/yksikkopaastot/ tavaraliikennee/vesiliikennee/roroe.htm (accessed 17.05.11).
Contents lists available at ScienceDirect
Journal of Cleaner Production journal homepage: www.elsevier.com/locate/jclepro
Reusable plastic crate or recyclable cardboard box? A comparison of two delivery systems Sirkka Koskela a, *, Helena Dahlbo a, Jáchym Judl a, Marja-Riitta Korhonen a, Mervi Niininen b a b
Finnish Environment Institute SYKE, Centre for Sustainable Consumption and Production, PO Box 140, FI-00251 Helsinki, Finland Stora Enso Oyj, Research Centre Imatra, FI-55800 Imatra, Finland
a r t i c l e i n f o
a b s t r a c t
Article history: Received 10 January 2013 Received in revised form 9 January 2014 Accepted 12 January 2014 Available online 27 January 2014
During a product’s entire life cycle the significance of packaging varies in terms of environmental impacts. From the perspective of companies which manufacture packaging or packaging has an important role in their value chain it can be a relevant issue to focus on in their efforts to improve the environmental performance of their activities. The aim of this study was to compare the life cycle environmental impacts of a real product (bread) delivery system using either reusable HPDE plastic crates or recyclable corrugated cardboard (CCB) boxes for product transportation. In this paper we focused on the delivery systems (not the delivered product) covering the manufacturing of the crates/boxes, their use, the delivery routes from bakery to retailers and waste management/recycling of the crates/boxes. As a result we concluded that the recyclable CCB box system was a more environmentally friendly option than the reusable HPDE plastic crate system in all the studied impact categories based on the defined boundaries and assumptions. Transportation played a very important role in the environmental impacts of the analysed systems. Therefore, changes, e.g. in the weights of products and their secondary packaging or the transportation distances could affect the results considerably. Ó 2014 Published by Elsevier Ltd.
Keywords: Packaging Transportation Delivery Life cycle assessment Corrugated cardboard Plastic
1. Introduction During a product’s entire life cycle the significance of packaging varies in terms of environmental impacts. Especially with foodstuff, manufacturing of the product itself is much more resource and energy intensive than the manufacturing of its packaging (Jungbluth et al., 2000). However, from the perspective of companies which manufacture packaging or packaging has an important role in their value chain, it can be a relevant issue to focus on in their efforts to improve the environmental performance of their activities. Emissions from the production stage of packaging are not the only aspects to be considered. In delivery systems, upstream processes, transportation in the distribution network and waste management issues must also be taken into account in order to assess environmental impacts holistically. Many industrialised countries have policy frameworks and measures aiming to minimize packaging waste and their environmental impacts (e.g. Sonneweld, 2000). The measures vary from
* Corresponding author. Tel.: þ358 400 148 811; fax: þ358 9 5490 2491. E-mail address: sirkka.koskela@ymparisto.fi (S. Koskela). 0959-6526/$ e see front matter Ó 2014 Published by Elsevier Ltd. http://dx.doi.org/10.1016/j.jclepro.2014.01.045
strict regulations imposed by governments to voluntary agreements between stakeholders. According to the waste hierarchy given in the EU Waste Framework Directive (2008/98/EC), the first priority of waste management is to prevent waste from being generated. Also the European Parliament and Council Directive on packaging and packaging waste (94/62/EC, amended by the Directive 2004/12/EC) contains provisions on the prevention of packaging waste, on the reuse of packaging and on the recovery and recycling of packaging waste. Reuse of products is undoubtedly a good measure for preventing waste since it can lengthen the lifetime of a product significantly. However, when looking at the overall environmental impacts of the product system where the reusable product is included, the picture is more complex due to e.g. the transportation and washing needed in order to enable the product reuse. This emphasizes the need for comprehensive environmental assessments of product systems in order to support decision making when choosing between different types of packaging materials and products. At present, packaging is a necessary part of delivery systems. The basic packaging functions are transportation, storage and distribution (Oki and Sasaki, 2000). In general, the functions of packaging materials, such as prevention of contamination, protection
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against damages, preservation of contents and communications to the customer, are many and varied in extent and complexity (e.g., Oki and Sasaki, 2000; Pasquilo et al., 2011). In our study we focused on storage, loading and transportation functions. The transport of goods demands particular container properties related to e.g. shape, weight, hygiene, handling and labelling (Stiftung Initiative Mehrweg, 2009). Transportation itself can be a very important factor in climate change impacts caused over the life cycles of packaging (e.g. Andersson and Ohlsson, 1999; Sim et al., 2007; Pretty et al., 2005; Meisterling et al., 2009). The importance of transport modelling in LCA has been well known for long (Jørgensen et al., 1996). According to Jørgensen et al. (1996), transport contributes to 5e15% of the major environmental impacts of a life cycle of a product. According to Madival et al. (2009), transport may contribute significantly to the environmental impacts of agricultural products and Gunady et al. (2012) emphasise especially the effects of a long distance transport. In recent decades, there has been an ongoing trend to find new materials based on biomass or renewable resources to replace nonrenewable materials, e.g. petroleum-based plastics (e.g. Madival et al., 2009). Several current policies (e.g. “Thematic Strategy on the Sustainable Use of Natural Resources COM/2005/670, An EU Strategy for Biofuels” COM/2006/34, A resource-efficient Europe e Flagship initiative under the Europe 2020 Strategy COM/2011/21), which aim to achieve a more resource-efficient economy, support the increased use of materials made from renewable resources. Products made from renewable materials are not automatically a better choice over ones made from non-renewable materials, since the whole supply chain from extraction to end-of-life must be considered. There can be aspects, e.g. reusing or recycling of packaging, which could change the ranking. Many comparison studies of packaging systems have been accomplished (e.g. Ross and Evans, 2003; Lee and Xu, 2004; Singh et al., 2006; Raugei et al., 2009). In all of them reusable plastic containers proved to be a better choice compared to single-use packaging. This bears out the general conception that reuse is always better than recycling. But is it true in every case? In our study we compared the life cycle environmental impacts of a real life delivery system using HDPE plastic crates or CCB boxes for transportation of the delivery product. We assessed the impacts of two delivery systems, one using a crate made of non-renewable plastic and a box made of renewable CCB. Both materials have advantages and disadvantages in terms of environmental impacts. Plastic crates are durable and washable, hence they can be reused several hundred times (in our study approximately 700 times) before finally being recovered as material for new plastic products or as energy. CCB boxes can be designed to be strong but light, and although they can only be used once, they can be recovered and used in the production of new fibre products or as energy. Levi et al. (2011) compared plastic containers and corrugated boxes to each other in Italian fruit distribution. They concluded that emissions from the manufacturing of corrugated box were greater than those from manufacturing plastic crates and the importance of transportation was identified in the environmental impacts of the distribution systems. The study of Stiftung Initiative Mehrweg (2009) presents the results of a comparison of fruit delivery systems in some European countries and Singh et al. (2006) in Northern American market finding the plastic container system better than the CCB system. However, these studies cannot be compared to our study as such due to several differences in the modelling assumptions. The greatest differences existed in e.g. the material composition of the crates/boxes, their weights, the number of circulations and transportation parameters. Additionally, end-of-life phases deviated from each other for both plastic crates and CCB boxes.
The scope of the comparison is not the use of secondary packaging needed for the delivery system, but the delivery system of packaged bread using CBB and HDPE plastic crates. The aim was to compare the life cycle environmental impacts of a real delivery system using either reusable HDPE plastic crates or recyclable CCB boxes for product transportation. The delivered product was toast bread which is a light weighted packed daily foodstuff delivered to the whole Finland. The results do not include the processes related to bread baking and its upstream, because the delivered product is not in the focus of this study. The weight of bread is, however, taken account in the impacts of transportation. The study was implemented in cooperation with the leading bakery company in Finland (VAASAN Oy) and with a global manufacturer of biomaterials, paper, packaging and wood products (Stora Enso Oyj). Both provided data and valuable insights from a business perspective for the study. A critical review of the study was conducted by the Swedish Environmental Research Institute (IVL). 2. Materials and methods 2.1. Life cycle assessment and data sources In order to achieve more sustainable production patterns, the environmental implications of the whole supply chain of products (both goods and services), their use, and waste management (ILCD, 2010) must be considered. Life cycle assessment (LCA) studies thereby help to avoid resolving one environmental problem while creating others, avoiding so called “shifting of burdens”. Life cycle assessment (LCA) is a method for integrating the environmental impacts of a studied product or a service over the whole value chain. It is an internationally standardized method (ISO 14040, 14044) with comprehensive guidelines (ILCD, 2010). In full LCA all processes and flows are followed from cradle-to-grave (i.e. from resource extraction to waste disposal) taking into account all relevant environmental impact categories. The goal of the study was to compare the life cycle environmental impacts of two different product systems for bread delivery from the bakery to consumers. The main difference in the systems was the type of material used for the delivery crates, either plastic or CCB, which generated differences in, among others, manufacturing and transportation (Fig. 1). The product systems (referred to as plastic crate system and CCB box system) included the life cycles of manufacturing of the crates/boxes from virgin materials and the delivery system of bread. The study tried to establish which container material would be more favourable from an environmental perspective in this specific distribution system. The weight of one plastic crate is 1.450 g with inside dimensions of 560 360 125 mm. It is made of high-density polyethylene (HDPE). The CCB box weighs 190 g and its dimensions are 540 330 110 mm. The bread delivered is toast bread. The weight of an average loaf of bread is 340 g (2.720 g in one crate/box). The weight of one plastic bag used for the bread packaging is 2 g (16 g in one crate/box). The different dimensions of a crate/box indicate slightly different capacities. However, all the crates/boxes hold the same load, 8 loaves of bread, therefore they perform the same function in the studied systems. Collected inventory data consisted of primary data from the participating companies, e.g. data related to the manufacturing of CCB boxes, transportation distances and modes and the washing process for crates. The washing of crates, but not the consumption of tap water (as a resource), was included in our assessment. The washing process also requires energy for heating the water and for the washing process, impacts of which are included in the study. The washing mainly removes dust from the crates and the detergents used for washing do not include phosphorus. Generic data
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Fig. 1. The studied product systems. Life cycle impacts of bread baking were excluded from the results shown in this paper.
from the ecoinvent v. 2.2 database were used for the manufacturing of plastic crates, washing chemicals and energy production. Emissions from transportation were calculated according to the Finnish Lipasto database (VTT, 2009) and the ecoinvent v2.2 database. Data for recovery processes in waste management were based on studies published by Korhonen and Dahlbo (2007) and Myllymaa et al. (2008). The impact categories analysed in the life cycle impact assessment (LCIA) were selected based on data availability and their relevance to the studied product systems: climate change, terrestrial acidification, photochemical oxidant formation, particulate matter formation, and fossil depletion. These impact categories indicate well the impacts of transportation and energy use for manufacturing processes. In addition, fossil depletion is connected to plastic production. Freshwater eutrophication was selected because cardboard production causes nutrient releases into water. Land use issues are very important aspects considering impacts of renewable resources. We did not obtain data for land use, therefore it is not considered in our study. Additionally, consensus of the methods for calculating land use has not been reached. For calculating the environmental impacts of the systems, characterisation and normalisation factors (European reference area) were taken from the ReCiPe (2011) midpoint (hierarchist) method, which has been developed for life cycle impact assessments (Goedkoop et al., 2009). ReCiPe was selected, because it’s one of the most up-to-date LCIA methods, currently. It is also under the continuous development. The data in this study are subject to uncertainties, which are very common in all LCA studies related to lacking specific data and model uncertainties (see e.g. Heijungs and Huijbregts, 2004; Guo and Murphy, 2012; Mattila et al., 2012). The model uncertainties are similar in both systems, but uncertainties related to parameters partly differ. No uncertainty analysis was conducted in this study, but three sensitivity analyses were done. Since transportation was
found to be the main contributor to the overall environmental impacts of the system, a sensitivity analysis was performed for transport distances. Two additional sensitivity analyses concerned the number of uses of the plastic crates and different allocation methods for calculating the benefits of CCB recycling. 2.2. Description of the compared product systems In life cycle comparisons, the boundaries of product systems must be consistent (ILCD, 2010). Therefore the plastic crate and CCB box systems had exactly the same boundaries including many equal components such as delivery routes from bakery to retailers and primary bread packaging (plastic bag). The systems differed in the manufacturing of crates/boxes, their use, transportation impacts in delivery (crate collection and take-back) and waste management/ recycling of the crates/boxes (Fig. 1). The function of the studied systems was to distribute bread from bakery to consumers. The functional unit of the product systems was 8 loaves of bread delivered in one crate/box. However, the study focused on the comparison of the two different crate/box materials and unit processes related to them, hence bread baking with its upstream were excluded from the results presented in this paper. 2.3. Transport modelling In plastic crate system the plastic crates were manufactured in Finland and transported to Tallinn, Estonia, where the bread was baked. In CCB box system the sheets of CCB were manufactured in Latvia where they were also cut into individual box sheets known as CCB blanks and transported to Tallinn. At the bakery, the boxes were assembled from the blanks in the box forming machine. Packed bread was then transported in crates or boxes from Tallinn (via Helsinki) to the main distribution centre in Eastern Finland. From there it was delivered to local distribution centres and then to
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Table 1 Transport distances and modes in bread delivery and crate/box collection. Load
Route
Distance (km)
Mode of transport
Hdpe granulate CCB blanks Crate
Production site / crate manufacturing site Production site (Latvia) / bakery Production site / bakery
Bread
Bakery / port of Tallinn Tallinn / Helsinki Helsinki / main distr. Centre Main distr. Centre / local distr. Centre Local distr. Centre / retail Same distances and transport modes as for bread delivery Same distances and transport modes as for bread delivery Retail / incineration/recycling Retail / recycling
140 310 100 80 10 10 80 130 333 162
Lorry 16e32 t (ecoinvent) Lorry 16e32 t (ecoinvent) Semi-trailer combination (hwy) RORO ship Semi-trailer combination (urb) Semi-trailer combination (urb) RORO ship Semi-trailer combination (hwy) Full trailer combination (hwy) Heavy delivery lorry (deli)
150 250
Heavy delivery lorry (deli) Heavy delivery lorry (deli)
Crate backhaul Empty runs (CCB case) Obsolete crate collection CCB collection
Note: symbols hwy, urb and deli stand for highway, urban and delivery driving conditions, respectively. The delivery driving mode has a defined highway mileage share which accounts for 30%.
retailers. From retailers empty crates were transported back to the main distribution centre for washing and then back to the bakery in Tallinn where they were reused. Obsolete plastic crates were collected and transported to energy recovery and recycling plants. Empty CCB boxes were collected and transported to recycling plants (Table 1). 2.3.1. Road freight transport Typically, transport impacts in LCA are calculated so that a mass of transported load (in tonnes) is multiplied by a distance (in kilometres) and this value (in tonne-kilometres) is linked to an inventory of a unit process representing the transport of 1 t-km using a specific mode of transport. LCI of road freight transport unit processes in major LCI databases, such as the ecoinvent v2.2 (Ecoinvent, 2010), are inventoried for trucks of a specific tonnage, for an average annual mileage and an average load (Spielmann and Scholz, 2005). The average values usually represent the situation in Europe. Although commonly used, these datasets are not suitable for every case study. In some products volume is the limiting factor for transportation, known as volume-limited. These are typically light products large in volume, such as empty plastic crates or other light three-dimensional objects. After the first LCIA calculations it became clear that the studied system is sensitive to the way transport is accounted for. According to the ILCD Guidelines (2010), more specific transport modelling had to be implemented. For this study the Finnish calculation system for traffic exhaust emissions and energy consumption, Lipasto (VTT, 2009) was chosen as the most appropriate methodology and database. Lipasto
enables calculation of the unit emission profile and fuel consumption for defined trucks of a specific load, which gives an advantage over generic LCI databases by providing more casespecific results. Emission data for a 2010 fleet average were used in the calculation. In order to account for the upstream environmental impacts of transportation we combined Lipasto with ecoinvent v2.2 (production and distribution of diesel fuel, manufacturing, maintenance and end-of-life of a vehicle and construction, maintenance and end-of-life of road infrastructure). 2.3.2. Sea transport The calculation of specific RORO (roll-on/roll-off) ship emissions is not applicable for trailers of different loads. Thus there was no difference when accounting for sea transport between the compared systems. The only difference is a justified assumption that 20% of otherwise empty trucks on the inbound route of CCB box system are utilised by another product system. These are therefore excluded from the system boundary. Unit emissions of the RORO ship were treated in the same way as those of the vehicles. They were combined with ecoinvent processes for fuel production, barge manufacture, maintenance and end-of-life. 2.3.3. Calculation approach The Lipasto methodology (VTT, 2009) features an equation (Eq. (1)) which was used in the calculation.
ex ¼
ea þ
eb ea lx lc
1 lx
(1)
Table 2 Weights of loads for distribution and return transportation. Vehicle type
Palettes per vehicle
Specific load (in tonnes) HDPE crate system Distribution
Semi-trailer combination Gross vehicle mass 40t Payload capacity 25t Full trailer combination Gross vehicle mass 60t Payload capacity 40t Heavy delivery lorry Gross vehicle mass 15t Pay load capacity 9t
CCB box system Return and collection
Distribution
Return and collection
26 Palettes
8.75
3.02
6.12
Empty
41 Palettes
13.79
4.76
9.66
Empty
3.15
1.08
2.21
Empty (return) 4.5 (collection)
Approx. 9 pallets
Note: distribution means the traffic from bakery to retailers (loaded with bread and crates/boxes). In the return and collection trips only empty crates (collected CCB boxes for recycling) are transported. The same degree of loading as for return and collection trips in the plastic crate system was used in crate transport from the production site to the bakery.
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The symbol ex in the equation above represents tonne-kilometre emissions (g/t-km) of a vehicle with user-defined payload lx (t). The symbol lc stands for payload capacity (t) of the vehicle; eb represents vehicle-kilometre emissions (g/v-km) of the vehicle when transporting lc and ea represents vehicle-kilometre emissions (g/vkm) of the vehicle when empty. The values of ea and eb were obtained from the Lipasto database. Inventory included emissions of CO, HC, NOx, PM, CH4, N2O, NH3, SO2 and CO2 as well as fuel consumption of the vehicle. The maximum load of each truck was calculated as the sum of the masses of crates, bread and bread packaging. Transportation EUR pallets were assumed to be an integral part of a truck and therefore were not accounted for as part of the load. It was calculated that one EUR-palette fits 80 crates/boxes. Based on that figure, specific loads of different trucks were calculated (Table 2). These were used for calculating unit emission profiles in urban and highway driving mode. Unit emissions for delivery driving were calculated for the last step in the bread distribution. The selected driving mode for each route is specified in Table 1. We calculated LCI of 1 tonne-km for each truck loaded to its maximum volume with the transported product (packed breads) in the transportation containers (CCB box, plastic crate). We calculated LCIA results by multiplying this LCI by the mass of transportation containers only. 2.4. Benefits from reuse, recycling and energy recovery After a fibre-based product has been used, it can be recycled. There are several ways of calculating the benefits of recycling. We studied how the impacts of climate change would vary using openloop allocation (ISO/TR 14049), monetary allocation and the system expansion approach to the CCB Box System. The results did not differ significantly from each other (less than 1%). We chose to use the system expansion (or avoided emissions) approach for this comparison study. In Finland, CCB is typically recycled for use in coreboard manufacturing. In practice, coreboard is always manufactured from recycled fibres, and virgin fibres are used only in minor quantities or not at all. However, in our study, we assumed that if virgin fibres were used they would be virgin fluting, which due to its fibre properties is the most suitable for coreboard manufacturing. Thus the emissions avoided by recycling were calculated as the difference between the emissions of manufacturing coreboard with virgin fluting and the emissions of manufacturing coreboard using recycled CCB. The raw materials for coreboard used in the calculations were 83% recovered fibre and 17% virgin fibre (data obtained from the participating forest company). The virgin fibre input is needed to compensate the loss of fibre strength during recycling. The plastic crates can also be recovered as material or energy at the end-of-life phase. When the crates were returned to the main distribution centre for washing, obsolete and broken crates were separated from the reusable ones. Approximately 20% of the obsolete crates were recovered as material in the production of plastic profiles, which is a process using e.g. HDPE as raw material (based on information from the bakery company). Plastic profiles can be used e.g. in patio constructions instead of impregnated wood. Therefore the production of impregnated wood was considered to be the process that can be avoided by plastic recycling (Korhonen and Dahlbo, 2007). In addition, around 80% of the obsolete crates were recovered as energy using in a boiler. The type of boiler used could not be specified; hence we assumed that the combustion of one kg of plastic produced 33 MJ of heat (based on the lowest heating value of plastic waste, Statistics Finland, 2011). This heat was assumed to replace separate production of
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heat for which an emission factor of 62.77 kg CO2/GJ was used (average emission factor for separate production of heat, Motiva, 2004). In our product system, the benefits of reusing plastic crates were considered by only including the amount of emissions that can be associated with one circulation of the crate. The number of circulations of one crate was calculated from the average lifetime of one crate (13.75 years), the number of circulations of one crate per year (61.54) and the duration of one circulation (4.87 days) (data obtained from the participating bakery company). For the specific route modelled in our system the circulation of a crate took one day longer than for the average route. Taking this into consideration, we estimated 700 circulations for each plastic crate. Hence we assumed that the distribution of 8 loaves of bread delivered in one crate represented 1/700 of the emissions of manufacturing one plastic crate. Likewise, the benefits of material and energy recovered at the end-of-life of the crate were assessed for 1/700 of a crate. 3. Results and discussion 3.1. Environmental impacts 3.1.1. Overall impacts This LCA study showed that for delivering 8 loaves of bread in one container, CCB box system was a more environmentally friendly option than plastic crate system in all impact categories based on the defined boundaries and assumptions (Fig. 2, Table 3). 3.1.2. Transportation In the delivery phase of the studied systems, transportation was the most significant contributor to all studied environmental impacts. For example the overall climate change impact related to transportation was 1.144 kg CO2-eq in plastic crate system and 0.757 kg CO2-eq in CCB box system. Distances, modes of transportation and particularly load were the most important factors in terms of transportation impacts. The greatest differences in the impacts of transportation between the systems were caused by the different weights of the crates/boxes and by the circulations of plastic crates. It should be noted that the delivery network in the studied systems covered the whole of Finland where the distances were very long. Local bakeries would decrease the amount of transportation, but such a decentralised system would have different impacts and the overall outcome cannot be evaluated without a comprehensive analysis. In the use phase (i.e. delivery network) CCB boxes performed better, because plastic crates need washing after every use which causes impacts on the environment. However, the significance of washing on the total impacts was very low. 3.1.3. Manufacturing of crates/boxes The environmental impacts of manufacturing one HDPE plastic crate were higher than those of one CCB box, but the fact that crates were reused hundreds of times decreased the impacts significantly (Fig. 3). As a result of the reuse of crates, the impacts from manufacturing a plastic crate were lower than those of a CCB box. In our study, however, the recycling of CCB to coreboard production changed the overall impacts more in favour of CCB. Decreasing, or increasing, the number of times the plastic crates are used had only a very small impact on the overall results of the system (unless the number of uses is extremely low). For illustrating this, the results of climate change were calculated with the number of uses of the crates ranging from 10 to 800. The results showed that for a range of uses from 10 to 100 times the impacts of manufacturing decreased notably, but there after the differences
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Fig. 2. Climate change (CC), terrestrial acidification (TA), freshwater eutrophication (FE), photochemical oxidant formation (POF), particulate matter formation (PMF) and fossil depletion (FD) according to the life cycle stages of the systems analysed and not including impacts of bread baking with its upstream. FU ¼ 8 loaves of bread delivered in one crate/ box. Note: benefits of recovery of plastic crates by incineration are not visible due to the minor relevance.
were not significant (Fig. 3). In the range of hundreds of uses the impacts from crate manufacturing per one use were minor.
Table 3 Environmental impacts of the analysed systems. Impact category
Unit
Plastic crate system
CCB box system
Climate change Terrestrial acidification Freshwater eutrophication Photochemical oxidant formation Particulate matter formation Fossil depletion
kg CO2-eq kg SO2 eq kg PO4 eq kg NMVOC
1.18Eþ0 7.90E3 4.15E4 1.02E2
8.76E1 6.24E3 3.79E4 7.40E3
kg PM10 eq kg oil eq
3.14E3 4.29E1
2.37E3 3.14E1
Therefore the manufacturing of the crate played a very small role in the system as a whole.
3.1.4. End-of-life phases The benefits of the recovery of plastic crates were very low (per functional unit) with the defined end-of-life assumptions (20% material recovery, 80% energy recovery based on the current situation) (Fig. 2) and therefore they are not visible as negative values in the results. The optimal end-of-life phase for plastic would be 100% material recovery to compensate for virgin plastic as has been shown by, e.g. Lazarevic et al. (2010). However, since only 1/700 of a plastic crate (which is equal to 2.07 g) was allocated per a functional unit, changes in the benefits would not affect the overall result of our study.
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Fig. 3. Climate change impacts of crate manufacturing (3.99 kg CO2-eq/crate) allocated per number of circulations.
The benefits of recycling CCB boxes are shown as negative values originating from the avoided emissions of using recycled CCB instead of virgin fibre in coreboard manufacturing (Fig. 2). Also in the future, demand for recycled corrugated boxes is expected to be high, due to continuous demand for recycled fibres. Corrugated boxes from retailers and industry are very often collected separately, which ensure the high purity and quality of fibres. CCB recycling into coreboard manufacturing generated the greatest benefits in the climate change impact category (Fig. 2). Without the benefit (0.106 kg CO2 -eq) the result for CCB box system would be 12% higher, but still lower than climate impact of plastic
CC
TA POF PMF FD 0.00E+00
2.00E-13
4.00E-13
6.00E-13
normalised results corrugated cardboard box (CCB) plastic crate Fig. 4. Normalised results of impact categories. FU ¼ 8 loaves of bread delivered in one crate/box.
crate system. Carbon sequestration in forests is an important part of the life cycle of fibre-based products. Taking this into consideration would increase the benefits of recycling in CCB Box System. There are models attempting to link carbon sequestration in forests to product carbon footprints but they contain high uncertainties. No scientific consensus has yet been reached on how to incorporate CO2 removal by forests into product specific assessments. Therefore a conventional approach was applied in this study and no benefit from carbon sequestration was allocated to CCB boxes.
3.2. Normalised results Normalisation is a means to present LCIA results with more comprehensible values than the impact category indicator scores and this also made it possible to make comparisons between the contributions of impact categories to the reference system (e.g. Dahlbo et al., 2013). In this study, European reference values were used for each impact category. Since weighting between impact categories was not done, the normalised impact category results can be considered to be of equal importance from the perspective of environmental protection within the region considered in the normalisation reference values (Europe) and the harmfulness of one category over another cannot be evaluated (Dahlbo et al., 2013). The normalised results indicated that plastic crate system made a greater contribution to the impacts in Europe than CCB box system (Fig. 4). For both systems, the lowest contribution was to climate change impacts and the highest to fossil depletion. For CCB box system the contribution to the particulate matter formation was higher than to the terrestrial acidification and the photochemical oxidant formation. In contrast, for plastic crate system the contributions to particulate matter formation and photochemical oxidant formation were higher than the contribution to terrestrial acidification.
Fig. 5. Sensitivity analysis of transport distances on freshwater eutrophication.
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3.3. Sensitivity analysis on transport distances Transport clearly dominates the overall results for each of the analysed impact category. Thus we applied a sensitivity analysis on transport distances in order to analyse the robustness of our findings. We varied all distances except the ferry transport, which would in any case remain the same. Two alternative scenarios were calculated, one for distances half shorter and one for distances twice longer. Due to linearity of the model the main conclusions did not change for most impact categories. Except for freshwater eutrophication (Fig. 5), where the contribution of the CCB manufacturing became more significant for the shorter distances scenario and it outbalanced lower transport impacts caused by lighter CCB box, compared to the plastic crate. It is important to notice that the impacts for shorter distances are more uncertain than ones for longer distances, because transport equipment would likely differ from the assumptions made in the study (e.g. smaller trucks for local delivery). 4. Conclusions In this LCA comparison study two bread delivery systems were examined. In the first system the product was delivered in HDPE plastic crates and in the second one in CCB boxes. As a result we can conclude that the CCB box system was a more environmentally friendly option than the plastic crate system in all studied impact categories based on the defined boundaries and assumptions. Our study proved that a conclusion on which delivery system has more favourable environmental impacts cannot be made based on the container material only. Before decision making, the whole delivery system including all necessary processes must be assessed. In general, a long-lasting, reusable product is considered to be a better choice, but as our study proved a recycled product can also be a good option but it requires a profitable and effective recycling system. Transportation played a very important role in the environmental impacts of the analysed systems. However, changes, e.g. in the weights of products and their secondary package or the transportation distances could affect the results considerably. Regardless of the size of the distribution area it is important to develop logistics and also vehicles further in order to decrease environmental impacts of delivery systems. Acknowledgements This study was conducted as part of the MMEA research programme managed by the Cluster for Energy and Environment and funded by TEKES e the Finnish Funding Agency for Technology and Innovation, Stora Enso Oyj and the Finnish Environment Institute (SYKE). A critical review of the study was conducted by the Swedish Environmental Research Institute (IVL). References Andersson, K., Ohlsson, T., 1999. Life cycle assessment of bread produced on different scales. Int. J. Life Cycle Assess. 4 (1), 25e40. Dahlbo, H., Koskela, S., Pihkola, H., Nors, M., Federley, M., Seppälä, J., 2013. Comparison of different normalised LCIA results and their feasibility in communication. Int J Life Cycle Assess. 18, 850e860. http://dx.doi.org/10.1007/s11367012-0498-4. Ecoinvent Database v.2.2, 2010. Swiss Centre for Life Cycle Inventories. http://www. ecoinvent.ch. Goedkoop, M.J., Heijungs, R., Huijbregts, M., De Schryver, A., Struijs, J., Van Zelm, R., 6 January 2009. ReCiPe 2009. A Life Cycle Impact Assessment Method Which Comprises Harmonised Category Indicators at The Midpoint and The Endpoint Level, first ed. Report I: Characterisation. http://www.lcia-recipe.net (accessed 20.05.11). Gunady, M.G.A., Biswas, W., Solah, V.A., James, A.P., June 2012. Evaluating the global warming potential of the fresh produce supply chain for strawberries, romaine/
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