Sustainability assessment of industrial waste treatment processes: The case of automotive shredder residue

Resources, Conservation and Recycling 69 (2012) 17–28

Contents lists available at SciVerse ScienceDirect

Resources, Conservation and Recycling journal homepage: www.elsevier.com/locate/resconrec

Sustainability assessment of industrial waste treatment processes: The case of automotive shredder residue Isabel Vermeulen a,∗ , Chantal Block a , Jo Van Caneghem a , Wim Dewulf b,c , Subhas K. Sikdar d , Carlo Vandecasteele a a

University of Leuven, Department of Chemical Engineering, Willem De Croylaan 46, 3001 Heverlee, Belgium Group T, Leuven Engineering College, KU Leuven Association, Andreas Vesaliusstraat 13, 3000 Leuven, Belgium c University of Leuven, Department of Mechanical Engineering, Celestijnenlaan 300a, 3001 Heverlee, Belgium d National Risk Management Research Laboratory, Office of Research and Development, U.S. Environmental Protection Agency, West Martin Luther King Drive 26, 45268 OH, United States b

a r t i c l e

i n f o

Article history: Received 17 April 2012 Received in revised form 13 August 2012 Accepted 28 August 2012 Keywords: Sustainability indicators Automotive shredder residue Recycling Energy recovery

a b s t r a c t To date numerous environmental, economic and societal indicators have been applied to evaluate and compare the sustainability of products and processes. This study presents a set of ad hoc sustainability indicators suitable for assessing and comparing processes for the treatment of industrial waste streams and for allowing to address efficiently all aspects of sustainability. This set consists of the following indicators: energy intensity, material intensity, water consumption, land use, global warming, human toxicity and treatment cost. The application of these indicators to industrial waste treatment processes is discussed in depth. A distinction is made between direct contributions to sustainability, occurring at the process level itself, and indirect contributions related to the production of auxiliaries and the recovery of end products. The proposed sustainability assessment method is applied to treatment processes for automotive shredder residue (ASR), a complex and heterogeneous waste stream with hazardous characteristics. Although different strategies for recycling and valorization of ASR were developed, with some of them already commercialized, large quantities of ASR are still commonly landfilled. This study concludes that for ASR the most sustainable alternative to the present landfill practice, both in short and long term perspective, consists of recycling combined with energetic valorization of the residual fraction. © 2012 Elsevier B.V. All rights reserved.

1. Introduction Analysis of industrial waste treatment processes should enable to address all sustainability aspects that are of importance to the respective stakeholders (European Commission Environment, 2011). Often, several treatment options and/or strategies are technically feasible for one specific waste stream. In that case, besides technical considerations, an in depth analysis should be conducted to achieve the optimal selection (Stehlik, 2009). Numerous sets of indicators have been suggested from a life cycle assessment (LCA) perspective and are commonly applied to evaluate and compare products and industrial processes (Azapagic and Perdan, 2000). Studies concerning the impact of (industrial) waste treatment processes are usually limited to an environmental impact assessment (Astrup et al., 2009; Boughton and Horvath, 2006; Ciacci et al., 2010; Mendes et al., 2004; Tarantini et al., 2007; Vos et al., 2007) and do not address sustainability as a

∗ Corresponding author. Tel.: +32 16 322353; fax: +32 16 322991. E-mail address: [email protected] (I. Vermeulen). 0921-3449/$ – see front matter © 2012 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.resconrec.2012.08.010

whole. Conceptual variations in the definition of sustainability often hamper the proper implementation of this concept (Sikdar, 2012). While it is impossible to define sustainability in an absolute sense, relative gain or loss of sustainability (over time or between alternatives) can be determined through the practical use of sustainability indicators (Azapagic and Perdan, 2000; Martins et al., 2007; Schwarz et al., 2002; Sikdar, 2003). The development and/or selection of a sufficient and relevant set of sustainability indicators can enhance the implementation of this concept in the assessment of industrial waste treatment processes. Within the context of sustainability evaluation, most commonly encountered methods are those that address eco-efficiency issues, as for instance proposed by the World Bussiness Council for Sustainable Development (Verfaillie and Bidwell, 2000), the Canadian National Round Table on the Environment and the Economy (NTREE, 2001) and BASF (Salling et al., 2002). A number of papers (Azapagic and Perdan, 2000; Martins et al., 2007; Schwarz et al., 2002; Sikdar, 2003, 2009, 2012) propose the use of three-dimensional (3D) sustainability indicators, if necessary complemented with two- or one-dimensional (2D or 1D) sustainability indicators. The method for sustainability evaluation,

18

I. Vermeulen et al. / Resources, Conservation and Recycling 69 (2012) 17–28

Table 1 Reuse and recycling/recovery targets, set by EU Directive 2000/53/EC, expressed in % of an ELV’s mass. Target

Current (%)

By 2015 (5)

Reuse and recycling Reuse and recovery

80 85

85 95

proposed by Sikdar (2003, 2009,2012), aims at using a limited set of 3D, all or not complemented with 2D, sustainability indicators that are easily quantifiable and can be deduced from readily available data. This evaluation method allows to obtain information on 3 dimensions of sustainability without necessitating specialized, 1D, indicators on each dimensions for which data are often not available. Applying this methodology, the present paper proposes a set of sustainability indicators, to assess, evaluate and compare the sustainability of industrial waste treatment processes. This set consists of the following indicators: energy intensity, material intensity, water consumption, land use, global warming, human toxicity and treatment cost. These indicators are specifically selected to cover the most important problems that can be encountered during industrial waste treatment (e.g. loss of resources and energy, consumption of water, land use, etc.) and to quantify these problems in terms of sustainability. This way, it is possible to determine which of the proposed or applied treatment strategies can be considered as ‘most sustainable’ for a certain industrial waste stream. Moreover, the application of these indicators to industrial waste treatment processes is discussed in depth, making a clear distinction between direct and indirect contributions to sustainability. Direct contributions occur at the process level itself, i.e. energetic and non-energetic resource consumption, land use, water consumption and emissions to air and water. Indirect contributions occur upstream or downstream of the process and are mainly related to the production of auxiliaries, and the recovery of end products. Auxiliaries have caused a certain impact on each of the indicators during their production; recovery of end products can avoid a certain impact as they replace newly processed products. The inventories of these indirect processes and their corresponding sustainability impacts, have been summarized in so-called impact factors. These impact factors are defined as the impact of an indirect contribution (e.g. NH3 , NaOH, cement, electricity, plastics, metals, etc.) on a certain sustainability indicator, expressed per unit of energy, per unit of mass or per unit of mass times unit of distance. For any given indirect contribution, an impact factor for each sustainability indicator can be calculated. Making a distinction between direct and indirect contributions, limits the input and output inventory to the level of the process itself, thus facilitating the comprehension of the results and the identification of bottlenecks/strengths of the process. The proposed sustainability assessment method is applied to the most commonly proposed and commercially applied treatment processes for automotive shredder residue (ASR). ASR can be defined as the 15–25% of an end-of-life vehicle (ELV) mass, remaining after de-pollution, dismantling, shredding of the ELV, and subsequent removal of ferrous and non-ferrous metals (Simic and Dimitrijevic, 2012; Vermeulen et al., 2011). As it is a complex and heterogeneous waste stream, large quantities of ASR are still commonly landfilled in Europe and throughout the world. This fraction is believed to further increase in the future as the amounts of plastics used in vehicles are increasing at the expense of metals and efficient separation of plastics from ASR is thus for not common practice (Passarini et al., 2012). To limit this otherwise growing waste stream, Europe has imposed very stringent targets regarding reuse, recovery and recycling of ELVs in the EU Directive 2000/53/EC (Council of the European Union, 2000). Present and future targets in terms of ELV mass have been given in Table 1.

Using the most recent Eurostat data (Eurostat, 2012), it can be concluded that substantial efforts still need to be done in the EU15 to reach the ELV reuse and recycling/recovery-rate (Council of the European Union, 2000). In 2009, the average ELV reuse and recycling-rate in the EU-15 amounted to 81.9% and the average ELV reuse and recovery-rate to 84.8%. A further increase of these rates by 2015 can be accomplished by focusing on two routes: increased recycling and/or increased recovery of ASR. Assuming that about 10% of an ELV is removed in view of reuse during de-pollution and dismantling, and on average 70% of the present metals (ferrous and non-ferrous) can be separated from the shredded fraction in view of recycling, the residual ASR fraction will represent 20% of the original ELV mass. This way the ELV targets of 2015 can be re-calculated in terms of ASR fractions: • Recycling of ASR ≥ 25% (5% of an ELV mass). • Recovery of ASR ≤ 75% (15% of an ELV mass). • Landfill of ASR ≤ 25% (5% of an ELV mass). Alternative treatment options to landfill of ASR, include: recovery of different materials by use of post shredder technologies (PSTs) in view of recycling and incineration of ASR with energy recovery (waste-to-energy) and thermo-chemical treatment of ASR (pyrolysis, gasification) (Boughton and Horvath, 2006; Ciacci et al., 2010; Srogi, 2008). A combination of different methods will be inevitable for a complete treatment of ASR that meets the European targets. To date, only very few papers have addressed the evaluation of ASR treatment processes (Boughton and Horvath, 2006; Ciacci et al., 2010; Duval and Maclean, 2007; Passarini et al., 2012) and depending on specific assumptions, e.g. the applied indicators, system boundaries, etc., their conclusions differ (Vermeulen et al., 2011). Agreement exists on the fact that landfill should be seen as the least preferred option. Moreover, these studies are limited to an environmental impact assessment (Boughton and Horvath, 2006; Ciacci et al., 2010; Passarini et al., 2012) or an economical assessment (Duval and Maclean, 2007), but none of them address sustainability as a whole, which is the goal of the present study. 2. Sustainability indicators After intensive literature screening (Azapagic and Perdan, 2000; Martins et al., 2007; Goedkoop et al., 2009; Guinée et al., 2002; Schwarz et al., 2002; Sikdar, 2003, 2009, 2012) and elaborate discussion with actors in the field, the following set of sustainability indicators is proposed: • Energy intensity: the net amount of energy consumed or recovered due to the processing of the functional unit of 1 metric tonne (t) of industrial waste. It is expressed in GJ/t of industrial waste and is a measure for the energy demand of a process. • Material intensity: the net amount of non-energetic or mineral resources consumed or recovered due to the processing of the functional unit of 1 t of industrial waste. It is expressed in kg Feeq./t of industrial waste and is a measure for the mineral resource demand of a process. • Water consumption: the net amount of water consumed or recovered due to the processing of the functional unit of 1 t of industrial waste. It is expressed in m3 /t of industrial waste and is a measure for the water demand of a process. • Land use: the land area occupied due to the processing of the functional unit of 1 t of industrial waste. It is expressed in m2 y/t industrial waste and reflects the damage to ecosystems due to the effects of occupying land.

I. Vermeulen et al. / Resources, Conservation and Recycling 69 (2012) 17–28

19

Table 2 Overall equations of the proposed sustainability indicators for industrial waste treatment processes. Indicator

Unit

Energy intensity (EI)

GJ/t industrial waste

Equation

 

Eres,i +



i

Material intensity (MI)

kg Fe-eq./t industrial waste



j

RDres,i × mres,i +

i

Water consumption (WC)

Land use (LU)

m3 /t industrial waste

m2 a/t industrial waste

i  i 

Global warming (GW)

kg CO2 -eq./t industrial waste i 

Human toxicity (HT)

Treatment cost (TC)

kg C6 H4 Cl2 -eq./t industrial waste D /t industrial waste

i 

IEI,mat,j × mmat,j +

Wi +

 j

Ai × ti +



IEI,car,k × Ecar,k +

k

IMI,mat,j × mmat,j +

j

IWC,mat,j × mmat,j +











IMI,car,k × Ecar,k +

k

IWC,car,k × Ecar,k +





k

ILU,mat,j × mmat,j +

j

GWPi × mem,i + HTPi × mem,i +



j 

ILU,car,k × Ecar,k +

k

IGW,mat,j × mmat,j +

IHT,mat,j × mmat,j +



k

j

IEI,transp,l × dl

l



IWC,transp,l × dl

l 

ILU,transp,l × dl

IGW,car,k × Ecar,k +

IHT,car,k × Ecar,k +

Ci × mi

IMI,transp,l × dl (10)

l

l

k

(9)



l

(11)

(12)

IGW,transp,l × dl (13)

IHT,transp,l × dl

(14)

l

(15)

i

• Global warming: the net amount of greenhouse gases (e.g. CO2 , CH4 , N2 O) emitted due to the processing of the functional unit of 1 t of industrial waste. It is expressed in kg CO2 -eq./t of industrial waste and is a measure for the potential contribution of different GHG emissions to the greenhouse effect. • Human toxicity: the net amount of toxic substances (e.g. heavy metals, PCDD/Fs, etc.) emitted due to the processing of the functional unit of 1 t of industrial waste. It is expressed in kg C6 H4 Cl2 -eq./t industrial waste and is a measure for the potential risk of emitted toxic substances to human health. • Treatment cost: the gate fee for the processing of the functional unit of 1 t of industrial waste and expressed in D /t industrial waste. In this indicator taxes are included, while possible revenues due to the recovery of products are not.

is made between direct and indirect contributions of a process to the sustainability indicators. A brief description of the sustainability assessment of the direct contributions is given in Section 2.1. This way of calculating is also generally described in literature (Guinée et al., 2002; Goedkoop et al., 2009). The sustainability assessment of the indirect contributions of a process, introducing the concept of impact factors, is discussed in Section 2.2. The overall equations to calculate the proposed sustainability indicators for industrial waste treatment processes (taking into account both direct and indirect contributions) are listed in Table 2. Section 2.3 elaborates on the dimensionality of the different sustainability indicators, indicating that all three dimensions of sustainability are effectively addressed.

This set of indicators was specifically selected in order to address efficiently and without overlap all sustainability aspects that are of importance to stakeholders of industrial waste treatment processes. It consists of the most commonly proposed and used sustainability indicators for decision making: material intensity, energy intensity and water consumption (Azapagic and Perdan, 2000; Martins et al., 2007; Schwarz et al., 2002; Sikdar, 2003). Land use was introduced as landfilling of industrial waste generally serves as reference scenario. This basic set of sustainability indicators was complemented with the following indicators: global warming, human toxicity and treatment cost. Global warming and human toxicity were adopted from the field of environmental impact assessment and were selected as relevant indicators for industrial activities. Industrial activities in Flanders (Belgium) account for 54% of the total impact on global warming and for 81% of the total impact on human toxicity; for other environmental indicators such as acidification and photochemical oxidant formation the contribution of industry is substantially smaller: 28% and 33%, respectively (Verbinnen et al., 2010). Furthermore, global warming receives substantial attention from policymakers and society, and human toxicity is of importance to reflect the possible hazardous nature of waste treatment processes. Ecotoxicity was omitted to avoid overlap as this indicator will give a similar trend as human toxicity. Finally, treatment cost was introduced in order to ensure the economic feasibility of the proposed treatment processes, since waste processing remains to be an economically driven activity. All these seven indicators are quantifiable, simple, unambiguous, useful to management decisions, robust and reproducible, which is important for their practical use (Schwartz et al., 2002; Martins et al., 2007). As already mentioned (Section 1) distinction

This section describes how the contribution of activities occurring at the process level itself (e.g. consumption of raw materials, emissions, etc.) is assessed by the different impact indicators. An inventory of inputs (energetic and non-energetic resources, water, land use, cost) and outputs (emissions, recovered end products) of the waste treatment process is required for this purpose. To assess the direct contribution to energy intensity, the energy inputs to the process are aggregated. All possible energetic resources (natural gas, fuel oil and coal, but also biomass, etc.) are included (Eq. (1)).

2.1. Direct contributions

Energy intensity =



Eres,i

(1)

i

where Eres,i = energy of resource i, expressed in GJ/t industrial waste To assess the direct contribution to material intensity, masses of mineral resources are multiplied by a characterization factor for mineral resource depletion and aggregated (Eq. (2)). This characterization factor for mineral resource depletion, as proposed by the ReCiPe method (midpoint, hierarchist perspective), is based on the surplus cost for society due to extraction of the mineral resource (Goedkoop et al., 2009). Material intensity =



RDres,i × mres,i

(2)

i

where RDres,i = characterization factor for mineral resource depletion of resource I (ReCiPe, midpoint, hierarchist perspective), expressed in kg Fe-eq./kg, mres,i = mass of resource i, expressed in kg/t industrial waste

20

I. Vermeulen et al. / Resources, Conservation and Recycling 69 (2012) 17–28

To assess the direct contribution to water consumption, the volumes of water consumed are aggregated. No distinction is made based on the origin of the water (Eq. (3)). Water, such as cooling water, not actually consumed in a process, is not included in this indicator.

direct contribution and will thus not be discussed again in Section 2.2.

Water consumption =

As stated before, indirect contributions to sustainability occur upstream or downstream of the process and are mainly related to the production of auxiliaries and the recovery of end products. Indeed, auxiliaries of waste treatment processes, such as cement for stabilization of hazardous waste or calcium oxide (CaO) introduced in the flue gas cleaning of waste incinerators, have caused emissions and land occupation, used energy and materials, etc. when they were processed. Three different types of indirect contributions are distinguished: energy carriers, materials (both can be consumed or recovered), and transport. Inventories of energy generation (expressed per unit of energy), of the processing of materials (expressed per unit of mass) and of the transportation of materials (expressed per unit of mass times unit of distance) are listed in the Eco-Invent database and Simapro (Goedkoop et al., 2012). By applying Eqs. (1)–(6) to these inventories, so-called impact factors can be determined, giving the indirect contribution on a sustainability indicator expressed per unit of mass (for materials), per unit of energy (for energy carriers) or per unit of mass times unit distance (for transportation), respectively. Impact factors can be determined for each indirect contribution on each of the considered sustainability indicators, except for ‘treatment cost’ as it only consists of a direct contribution (Section 2.1). For instance, the impact factor for the indicator ‘energy intensity’ for cement consumption (used for the stabilization of hazardous waste before landfill), is calculated by applying Eq. (1) to the inventory of energetic resources of the production of 1 kg cement (Goedkoop et al., 2012; Kellenberger et al., 2007). This impact factor will thus give the impact of the production of 1 kg cement on the indicator ‘energy intensity’. Multiplying this impact factor with the amount of cement consumed in a specific waste treatment strategy, gives the actual impact of the cement consumption in that specific strategy on ‘energy intensity’. This way of addressing the problem has the advantage to limit the inventory to the level of the process itself (amounts of auxiliaries and end products) and will thus facilitate comprehension of the results and possible identification of bottlenecks within the process. For the sustainability assessment of the indirect contributions, the following terms are added to the Eqs. (1)–(6), discussed in Section 2.1:



Wi

(3)

i

where Wi = volume of water type i, expressed in m3 /t industrial waste To assess the direct contribution to land use, the occupied area (expressed in m2 ) is multiplied by the time of occupation (expressed in years) and aggregated. For simplicity and as this is usually not done for midpoint assessment methods, no distinction or weighing is applied to the different land types (Eq. (4)) (Guinée et al., 2002; Goedkoop et al., 2009). Land use =



Ai × ti

(4)

i

where Ai = amount of area of land type i that is occupied, expressed in m2 /t industrial waste ti = time of occupation of land type i, expressed in years (y) To assess the direct contribution to global warming, the greenhouse gas emissions (e.g. CO2 , CH4 , N2 O, etc.) are multiplied by a characterization factor for global warming and aggregated (Eq. (5)). Global warming potentials (GWPs) for an impact period of 100 years, as proposed by IPCC, are generally accepted as the baseline midpoint characterization method for quantifying global warming. Climate change =



GWPi × mem,i

(5)

i

where GWPi = global warming potential of a substance i, expressed in kg CO2 -eq./kg, mem,i = mass of an emitted substance i, expressed in kg/t industrial waste To assess the direct contribution to human toxicity, the mass of emitted toxic substances (e.g. heavy metals, PCDD/Fs, etc.) is multiplied by a characterization factor for human toxicity and aggregated (Eq. (6)). The midpoint characterization factors proposed by the ReCiPe methodology (midpoint, hierarchist perspective) are used. These are based on a distribution of the toxic substances in accordance with the USES 2.0 model (exposure factor) and the potential risk as a result of a certain exposure to the toxic substance, quantified by the ADI (acceptable daily intake) of the toxic substance (damage factor) (Goedkoop et al., 2009). Human toxicity =



HTPi × mem,i

(6)

i

where HTPi = human toxicity potential of a substance i, expressed in kg C6 H4 Cl2 -eq./kg, mem,i = mass of an emitted substance i, expressed in kg/t industrial waste Treatment cost is defined as the gate fee for 1 t of industrial waste, including the imposed taxes. Gate fees and taxes of the different treatment processes within a specific waste treatment strategy are aggregated to obtain the overall treatment cost per t of industrial waste. Treatment cost =



Ci × mi

(7)

i

where Ci = the gate fee of a waste treatment process i, expressed in D /kg industrial waste, mi = mass fraction of the industrial waste stream treated by a treatment process i, expressed in kg industrial waste/t industrial waste Possible costs and revenues due to the consumption of auxiliaries and the recovery of end products are not separately taken into account. This way, the indicator ‘treatment cost’ consists only of a

2.2. Indirect contributions

 j

Ix,mat,j × mmat,j +

 k

Ix,car,k × Ecar,k +



Ix,transp,l × dl

(8)

l

where Ix,mat,j = impact factor of material j on the sustainability indicator x, mmat,j = mass of material j, expressed in kg/t industrial waste, Ix,car,k = impact factor of an energy carrier k on the sustainability indicator x, Ecar,k = amount of energy of an energy carrier k, expressed in MJ/t industrial waste, Ix,transp,l = impact factor of a transportation modus l on the sustainability indicator x, dl = transport distance of a transportation modus l, expressed in tkm/t industrial waste Depending on the sustainability indicator ‘x’, impact factors are denominated as IEI (energy intensity), IMI (material intensity) IWC (water consumption), ILU (land use), IGW (global warming) and IHT (human toxicity). In some industrial waste treatment processes, end products such as energy carriers (e.g. electricity, steam) and/or materials (e.g. metals, plastics) are recovered. As these recovered end products replace newly processed products, the impacts associated with the production of these newly processed products can be avoided. Amounts of recovered end products appear in this case as negative values in the inventory, meaning that mmat,j or Ecar,k (Eqs. (8) and

I. Vermeulen et al. / Resources, Conservation and Recycling 69 (2012) 17–28

economic

environmental Global warming Water consumpon

Energy intensity Material intensity Land use Treatment cost

Human toxicity

societal Fig. 1. Dimensionality of the proposed set of sustainability indicators.

(9)–(14)) become negative values. This way the avoided (negative) impacts are subtracted from the actual (positive) impacts. If the avoided impacts exceed the actual impacts, the value of the indicator calculated using Eqs. (9)–(14) becomes negative. A net negative value for a sustainability indicator means that a general improvement of the sustainability aspects covered by that specific indicator can be achieved by the waste treatment strategy. 2.3. Dimensionality of the sustainability indicators Depending on the number of dimensions of sustainability (environmental, economic, societal) incorporated in an indicator, it can be denominated as one dimensional (1D), two dimensional (2D) or three dimensional (3D). Fig. 1 presents graphically the dimensionality of the proposed set of sustainability indicators, defined for Europe. It has to be remarked that local conditions can influence the dimensionality of an indicator. For instance when water is scarce in a certain region, water consumption will also reflect a societal dimension, whereas this is not the case for e.g. Europe. Fig. 1 shows that all dimensions of sustainability are well covered, even when certain 3D indicators would be redefined as 2D ones or vice versa. In general preference should be given to a limited set of 3D sustainability indicators, where necessary complemented with 2D and 1D indicators (Sikdar, 2003). The most commonly proposed 3D indicators: energy intensity and material intensity were included in the proposed set (Schwarz et al., 2002; Martins et al., 2007; Sikdar, 2003). As landfilling of industrial waste is generally taken as reference scenario, land use, another 3D indicator, was added. This set of three 3D indicators was complemented with four 2D indicators: water consumption, as water is an important and even vital resource to industry and society, global warming, which receives a lot of attention from policymakers and society, human toxicity, selected due to the possible hazardous characteristics of industrial waste streams, and treatment cost, selected as waste processing is economically driven. Energy intensity can be seen as a measure for the energy demand of a process. Energy is a prime driver for economic growth, its generation is generally associated with an environmental impact and, as energy sources are limited, energy consumption affects future generations (Sikdar, 2003). Therefore, energy intensity is identified as a 3D sustainability indicator. Material intensity can be considered as a measure for the mineral resource demand of a process. Material use has a direct environmental impact, for instance during mining or processing of raw materials; materials are associated with value creation (economic) and excessive material depletion can have an intergenerational impact (societal) (Sikdar, 2003). Therefore,

21

material intensity is identified as a 3D sustainability indicator. Land use reflects the damage to ecosystems due to land occupation. Land use has a direct environmental impact, because it affects eco-systems. The economic value of land is important in today’s society, especially in densely populated regions as Europe. Competition between different occupational land uses and subsequent shortage of land, are important societal aspects. Therefore, land use is identified as a 3D sustainability indicator. In sparsely populated areas however, where there is no land shortage, the indicator can shift from 3D to 2D. Water consumption can be considered as a measure for the water demand of a process. The residuals from water works can constitute an environmental problem and water shortage can directly affect eco-systems (Sikdar, 2003), Water is also an important and even vital resource, the indicator has thus an economic dimension and in case of water scarcity, also a societal dimensions. However, as in Europe water is not scarce, the societal dimension is not included for the indicator in this study and the indicator is thus defined as a 2D (eco-efficiency) indicator. Global warming reflects the potential contribution of different GHG emissions from the process to the greenhouse effect. On the one hand, GHG emissions have an impact on climate change, thus affecting the environment. On the other hand, the global warming indicator has also an economic dimension i.e. GHG emissions control regulations such as carbon trading. Global warming is thus a 2D (eco-efficiency) indicator. Human toxicity reflects the potential risk of emitted toxic substances to human health. Toxic substances have a direct environmental impact and, as this indicator reflects the effect of these toxic substances on human health, the societal dimension is also clear. When toxic substances are emitted in such amounts that legal limits are exceeded, this indicator may also have an economic dimension. In the present study however it has been assumed that the considered industrial waste treatment processes are feasible within legal limitations. Therefore, human toxicity is identified as a 2D (socio-ecologic) indicator. Treatment cost reflects on the one hand the type of technology used and its possible economic value creation, and on the other hand also the general affordability of a treatment strategy, which is a societal aspect (Sikdar, 2003). Therefore, treatment cost is identified as a 2D (socio-economic) indicator. 3. Inventory of ASR treatment strategies 3.1. ASR waste ASR is generally defined as the 15–25% of an ELV mass, remaining after de-pollution, dismantling, shredding of the ELV, and subsequent removal of ferrous and non-ferrous metals (Vermeulen et al., 2011). ASR is a complex and heterogeneous waste stream, classified as hazardous according to the list of hazardous wastes 2000-532-EEC (Chapter 1910, Annex of the European Directive 91-689-EEC on hazardous waste). The treatment of 1 t of ASR is used as reference flow for the sustainability assessment. Table 3 gives the composition range and average composition of ASR, used in the present study. These values are based on extensive literature search (Börjeson et al., 2000; Boughton, 2007; Vos et al., 2007; Mancini et al., 2010; Morselli et al., 2010; Nourreddine, 2007; Roy and Chaala, 2001; Vermeulen et al., 2011). 3.2. ASR treatment strategies Despite the potential of recovering materials and energy from ASR, a number of barriers exist on legislative, technical and economic level (Forton et al., 2006). Therefore, large quantities of ASR are still commonly landfilled in Europe and throughout the

22

I. Vermeulen et al. / Resources, Conservation and Recycling 69 (2012) 17–28

Table 3 Composition range and average composition of ASR. Parameter

Unit

ASR composition range

Average ASR composition

C N S Cal. value As Cd Cr Cu Hg Ni Pb Zn

kg/t kg/t kg/t GJ/t ppm ppm ppm ppm ppm ppm ppm ppm

279–626 8.8–45 1.9–5.6 13–29 1.2–70 2.0–86 17–7000 27–16,600 0.2–14 54–4,000 94–7000 1430–14,100

410 19 3.7 17 30 34 1120 4910 4.1 734 2610 8260

world (Eurostat, 2012; Ferrão et al., 2006; Morselli et al., 2010; Nourredine, 2007). As discussed in the introduction (Section 1), the average ELV reuse and recovery rate in the EU-15 amounts to about 85% of an ELVs mass. In terms of ASR, this means that on average 75% of the ASR is landfilled in the EU-15. In countries like Finland, Denmark and France even 95%, 91.5% and 89.5% of the ASR was landfilled respectively in 2009 (Eurostat, 2012). Therefore, complete landfill of ASR has been considered as reference scenario in the present study, as this will represent the worst case scenario. As already mentioned before, alternative treatment processes for ASR include: recovery of different materials (mainly plastics) by use of PSTs in view of recycling, incineration with energy recovery and thermo-chemical treatment. Generally, a combination of different treatment processes will be inevitable for a complete treatment of ASR that meets the European targets: 85% “reuse and recycling” and 95% “reuse and recovery” of an ELVs mass by the year 2015 (Directive 2000/53/EC). Thermo-chemical treatment (e.g. pyrolysis and gasification) of ASR has been omitted from the present discussion, as it is an emerging technology still requiring substantial research and only little data concerning commercial thermo-chemical processing of ASR is available (Mancini et al., 2010; Donaj et al., 2011). One of the aspects that still require extensive research is the end-product treatment, in particular of the chars (Al-Salem et al., 2010; Donaj et al., 2011; Vermeulen et al., 2011). Other ASR treatment strategies that are proposed at times consist of injection in a blast furnace or direct incorporation into manufactured products (Vermeulen et al., 2011; Robson and Goodhead, 2003). Considered ASR treatment strategies of the present study are: (i) landfilling (reference scenario), (ii) recycling combined with landfilling, (iii) energy recovery with landfilling of the ashes and (iv) recycling combined with energy recovery. This study is based on input and output data derived from processes applied on commercial scale, as will be discussed further. The objective of this assessment is purposefully comparative among the alternative processes, and in this sense will yield processes that are more (or less) sustainable in contrast to the reference process of landfilling. The sustainability assessment focuses on the actual waste treatment strategy, setting the system boundary at the physical boundaries of each of the considered ASR treatment strategies. Consequently, the system boundary begins with the arrival of ASR at the ASR treatment plant (i.e. a residual material landfill, an Argonne PST plant or a fluidized bed combustor of the SLECO type, depending on the treatment strategies) and closes after the complete treatment and final disposal of ASR and all residual fractions. This is in accordance with the principle of excluding equivalent activities as given in the ISO guidelines, as all life cycle phases prior to the actual treatment of the ASR are common and hence assumed not to be affected by the choice of ASR treatment process and can thus be excluded from the assessment. The most important direct and indirect contributors to sustainability (e.g.

emissions, consumption/recovery of energy and products, etc.), selected according to their relevance on sustainability, are summarized in Table 4 for the different treatment strategies. The time frame of the inventory of processes is typically set at 100 y. However, emissions of landfill sites can occur over time frames substantially larger than 100 y, so that for landfills timeframes of 60,000 y are often used (Frischknecht et al., 2007). Both short (100 y) and long (60,000 y) term inventory timeframes are included in the present assessment. These two inventory timeframes are only of importance for landfilling of the waste or residual fractions, because emissions of landfilling will occur spread over a long time, while for all other waste treatment processes the emissions do not occur over such long time frames. 3.2.1. Landfilling The considered landfill type for ASR disposal is a “residual material landfill”, as industrial waste streams, low in organic carbon, such as ASR, are typically disposed of on such sites (Doka, 2007; Ciacci et al., 2010). Time of occupation was taken to be 35 y: 5 y for construction and 30 y for operation, as also suggested by Doka (2007). In this type of landfill, landfill gas is usually not captured because of the very low production typically encountered. Nevertheless, based on a literature search it was estimated that on short term basis (100 y) about 3% of the carbon incorporated in ASR will degrade, 70% of which is emitted as methane and 30% as carbon dioxide (Eriksson and Finnveden, 2009). On long term basis (60,000 year) it is expected that about 65% of the carbon is leached out (Doka, 2007); and it is estimated that the remaining 35% of carbon will eventually be degraded, 70% of which is emitted as methane and 30% as carbon dioxide. Next to the production of biogas, leaching of heavy metals into ground and surface water due to infiltration of rainwater can be considered as an important emission source. Leaching of heavy metals from residual material landfills was estimated based on the composition of ASR, reported in Table 3, and short and long term transfer coefficients into water, reported by Doka (2007). Based on preliminary estimates of impacts, the most important heavy metals that are leached out are: As, Cr(VI), Ni, Sb, Se, Tl, V and Zn. Industrial waste that is directly landfilled in residual landfills, may require additional solidification. It is assumed that the ASR is solidified prior to landfilling by addition of cement and water in the proportion waste-cement-water of 50%-20%-30% as proposed by Doka (2007). The most important indirect contributions to the different impact indicators of landfilling are due to transport of material during construction and to the solidification of the waste. They are also included in Table 4. 3.2.2. Recycling combined with landfilling Recovery of all recyclable material from ASR in one single step is not feasible due to its heterogeneity (Vermeulen et al., 2011). PSTs mainly concentrate on the recovery of residual metals (both ferrous and non-ferrous) and plastics that can be recovered with sufficient quality. The remaining fraction can be incinerated in view of energy recovery or landfilled. Data from the PST of Argonne National laboratory, as described by Gallon and Binder (2006), are used for the present inventory. This technology enables the recovery of 7.2% of the original ASR input as metals (steel, aluminium, copper and brass) and 21.2% as plastics (PET, PUR-foam, PP, low grade PS, ABS, PC and butadiene) in view of recycling; an additional 13.3% is evaporated moisture during processing. The exact amounts of recovered materials per t of treated ASR along with the related consumption of energy and of auxiliaries are detailed in Table 4. The residual fraction, amounting to 58.3% of the original ASR input, mainly consists of fines and is generally processed by incineration with energy recovery and/or landfilling (Gallon and Binder, 2006). In the present study, this residual fraction is assumed to be either

Table 4 Inventory of the considered ASR treatment strategies. Parameter

Landfill

Recycle + landfill

Incineration

Recycle + incineration

Short term

Long term

Short term

Long term

Short term

Long term

Short term

Long term

m2 a/t m3 /t D /t

6.25 0.6 106

6.25 0.6 106

3.6 0.35 161

3.6 0.35 161

0.8 0.6 133

0.8 0.6 133

0.5 0.3 177

0.5 0.3 177

kg/t kg/t kg/t kg/t kg/t kg/t kg/t kg/t kg/t kg/t

13.5 11.5

158 134

9.15 8.07

111 94

1504

1504

1057

1057

1.58 × 10−5 7.00 × 10−6 1.28 × 10−6 2.90 × 10−5 5.37 × 10−5 1.50 × 10−4 5.95 × 10−5 4.20 × 10−4

1.58 × 10−5 7.00 × 10−6 1.28 × 10−6 2.90 × 10−5 5.37 × 10−5 1.50 × 10−4 5.95 × 10−5 4.20 × 10−4

3.78 × 10−6 8.18 × 10−7 2.20 × 10 × −7 3.56 × 10−5 5.68 × 10−6 3.68 × 10−5 3.12 × 10−5 3.45 × 10−4

3.78 × 10−6 8;18 × 10−7 2.20 × 10 × −7 3.56 × 10−5 5.68 × 10−6 3.68 × 10−5 3.12 × 10−5 3.45 × 10−4

kg/t kg/t kg/t kg/t kg/t kg/t kg/t kg/t

2.97 × 10−2 6.73 × 10−2 4.44 × 10−4 2.38 × 10−2 5.64 × 10−4 1.15 × 10−6 4.96 × 10−5 1.69 × 10−4

0 7.69 × 10−3 3.29 × 10−5 1.43 × 10−2 4.61 × 10−4 0 0 3.56 × 10−5

0 3.20 × 10−2 1.97 × 10−2 4.04 × 10−2 1.31 × 10−3 0 0 2.14 × 10−2

0 4.48 × 10−3 1.29 × 10−5 8.33 × 10−2 2.69 × 10−4 0 0 2.07 × 10−5

0 1.58 × 10−2 1.14 × 10−2 2.34 × 10−2 7.57 × 10−4 0 0 1.24 × 10−2

−8.5 −4.6

−8.5 −4.6

1.2

1.2

16.8 50.4

16.8 50.4

−3.6 −1.8 −5.5 −28.6 0.7 −1.0 −124 9.8 29.4 −6.5 −19.4 −0.8 −1.9 −42.9 −17.0 0.6 −48.6 11.3

GJ/t GJe/t kg/t kg/t kg/t kg/t kg/t kg/t kg/t kg/t kg/t kg/t kg/t kg/t kg/t kg/t kg/t

400

tkm/t

154

2.97 × 10−2 2.80 × 10−1 2.66 × 10−1 6.74 × 10−2 1.60 × 10−3 6.89 × 10−4 1.56 × 10−2 1.01 × 10−1

400

154

7.16 × 10−3 2.40 × 10−4 1.08 × 10−4 2.20 × 10−2 6.16 × 10−4 0 1.72 × 10−5 1.39 × 10−4

7.16 × 10−3 9.97 × 10−4 6.50 × 10−2 6.23 × 10−2 1.75 × 10−3 0 5.39 × 10−3 8.37 × 10−2

0.5 −5.5 −28.6

0.5 −5.5 −28.6

−1.0 −124

−1.0 −124

233 −6.5 −19.4 −0.8 −1.9 −42.9 −17.0

233 −6.5 −19.4 −0.8 −1.9 −42.9 −17.0

−30.3

−30.3

1.0 −31.6

1.0 −31.6

−3.6 −1.8 −5.5 −28.6 0.7 −1.0 −124 9.8 29.4 −6.5 −19.4 −0.8 −1.9 −42.9 −17.0 0.6 −48.6

89.7

89.7

19.4

19.4

11.3

I. Vermeulen et al. / Resources, Conservation and Recycling 69 (2012) 17–28

Direct contributions Land Water Treatment cost Emissions to air CO2 CH4 As Cd Cr(VI) Cu Hg Ni Pb Zn Emissions to water As Cr(VI) Ni Sb Se Tl V Zn Indirect contributions Energy Electricity ABS Aluminium Ammonia Brass Butadiene Calcium oxide Cement Copper Low grade PS PC PET PP PUR-foam Sodium hydroxide Steel Transport Transport

Unit

23

24

I. Vermeulen et al. / Resources, Conservation and Recycling 69 (2012) 17–28

completely landfilled (this section) or completely incinerated with energy recovery (Section 3.2.4). This way, the influence on the overall sustainability of these different treatment approaches of the residual fraction can be compared. The composition of the residual fraction was estimated based on composition data of different ASR fractions reported in previous work of the authors (Vermeulen et al., 2011). The inventory for landfilling of the residual fraction, as given in Table 4, is determined similarly to the inventory in Section 3.2.1. 3.2.3. Energy recovery with landfilling of the ashes In the EU-15, on average 3% of an ELV’s mass is treated by energy recovery (Eurostat, 2012). ASR can for instance be used as calorific waste stream (calorific value > 9 MJ/kg) for co-incineration with wastewater treatment (WWT) sludge, as performed in the SLECO fluidized bed combustor (FBC) at the Indaver site in Antwerp, Belgium. Data of a full scale application of energy recovery from ASR, reported in previous work (Vermeulen et al., 2012) are used in the present paper. The considered fluidized bed combustor is of the ROWITEC internal rotating type. Temperatures in the sand bed are on average 720 ◦ C and those in the freeboard are on average 920 ◦ C. Energy from the flue gas is recovered as electricity, with a net energy recovery of 27% of the energetic input. This percentage is quite elevated considering that the FBC produces only electricity (Pavlas et al., 2011). Next to this, also ferrous metals are recovered from the waste material, before the actual incineration. Calcium oxide (CaO), sodium hydroxide (NaOH) and ammonia (NH3 ) are introduced in the flue gas cleaning system to remove PCDD/Fs, heavy metals, acid gases and NOx emissions from the flue gas. Waste water originating from the flue gas cleaning is completely recirculated, leading to a zero wastewater discharge. About 12.6% of the initial waste input is moistened and stabilized/solidified before being landfilled. With respect to the sustainability impact of energy recovery from ASR, the emissions of CO2 and of heavy metals (As, Cd, Cr(VI), Cu, Hg, Ni, Pb, Zn) were identified as the most important emissions into air (Table 4). The inventory for landfilling of the ashes, as given in Table 4, is determined similarly to the inventory in Section 3.2.1. 3.2.4. Recycling combined with energy recovery Analogously to Section 3.2.2, data from the PST of Argonne National Laboratory are used for the inventory of the recycling of different materials from ASR, as described by Gallon and Binder (2006) (Section 3.2.2). This technology enables the recovery of 7.2% of the original ASR input as metals (steel, aluminium, copper and brass) and 21.2% as plastics (PET, PUR-foam, PP, low grade PS, ABS, PC and butadiene); an additional 13.3% of the original ASR input is evaporated moisture during processing. For this treatment strategy, it was assumed that the entire residual fraction after material recovery (58.3% of the original ASR input), was incinerated with energy recovery in the SLECO FBC. The composition of this residual fraction was estimated based on data of different ASR fractions reported in previous work (Vermeulen et al., 2011). The inventory for energetic valorization of the residual fraction, as given in Table 4, is determined similarly to the inventory in Section 3.2.3. 4. Application of sustainability indicators to the ASR treatment strategies By applying Eqs. (9)-(15) (Table 2) to the data of the inventory (Table 4), a sustainability assessment for each of the considered ASR treatment strategies was performed. To conduct this assessment, impact factors for the most relevant energy carriers, products and transportation modes were determined using the Eco-Invent database and Simapro (Section 2.2). The impact factors used in the present study, are summarized in Table 5. Because heavy metal

leaching from landfilling (of residual fractions and by-products) are different on short (100 years) and long (60,000 years) term perspective, two impact factors are given for the indicator human toxicity, one on short and one on long term perspective. Table 5 gives the results of the sustainability assessment of the considered ASR treatment strategies as calculated using Eqs. (9) to (15) (Table 1). The higher the indicator, the less sustainable the treatment strategy for that indicator; the lower the indicator, the more sustainable the treatment strategy for that indicator. Negative values represent a general improvement of sustainability and can be attributed to the recovery of end products (i.e. energy carriers and/or materials). To compare the sustainability of the different treatment strategies, the figures obtained for the strategy “landfill” were taken as reference, setting them at 100% for each indicator and normalizing the data of the other strategies to this reference strategy. These normalized data are presented in Fig. 2. It appears from Table 6 and Fig. 2, that “landfill” is the least sustainable strategy for the treatment of ASR, as it has the highest impact for most of the considered sustainability indicators. Only for the sustainability indicators global warming (on short term perspective) and treatment cost is the contribution of landfill not the highest. The treatment strategy “recycle + energy recovery” (recycling combined with energy recovery) gives for most sustainability indicators the lowest impact, with the exception of the indicators global warming (on short term perspective) and treatment cost. It can thus be considered as the most sustainable treatment method for ASR. Considering energy intensity, it is clear that “landfill” is the only treatment strategy that actually consumes energy. The contribution of this treatment strategy to energy intensity can mainly be attributed to the indirect impact due to the consumption of cement for solidification of the waste. The other treatment strategies all have negative values for this indicator, implying that more energy can be recovered than is consumed. The strategy “recycle + energy recovery” avoids only slightly more energy consumption than the strategy “energy recovery”, but both avoid two times more energy consumption than the combination of “recycle + landfill”. Also for the indicator material intensity, “landfill” is the only treatment strategy with a net consumption of non-energetic resources, the most important contribution being the indirect impact due to the consumption of cement. As expected, the strategies where materials are recovered in view of recycling (“recycle + landfill” and “recycle + energy recovery”) show the lowest impact on material intensity, saving 408 and 438 kg Fe-eq./t ASR, respectively (Table 6). The treatment strategy “recycle + energy recovery” scores better than “recycle + landfill”, due to the additional pre-treatment of the ASR before co-incineration in the FBC. This pre-treatment also explains the saving of 48.2 kg Fe-eq./t ASR for the treatment strategy “energy recovery” (Table 6). Landfilling of ASR is the only treatment strategy with a net consumption of water (1.7 m3 /t ASR) and thus the only treatment strategy with a positive contribution to the indicator water consumption. Water is mainly consumed during solidification of the waste via direct water consumption and indirectly due to the consumption of cement. The other treatment strategies consume less water than the amount saved through the recovery of end products (materials and/or energy carriers). Indeed, water consumption related to the processing of metals and plastics and to the generation of energy from raw materials is avoided in these treatment strategies. The treatment strategy “recycle + energy recovery” has the highest water saving of about 8 m3 of water per t of ASR that would be consumed if the ASR were not treated. Also for the indicator land use, “landfill” is obviously the least preferable option, as the process itself occupies large surfaces of land. All other treatment strategies reduce land occupation due to

I. Vermeulen et al. / Resources, Conservation and Recycling 69 (2012) 17–28

25

Table 5 Impact factors of the considered products. Product

Energy intensity IEI

Material intensity IMI

Water consumption IWC

3

Land use ILU

2

Global warming IGW

Human toxicity IHT Short term

Long term

Unit Electricity

MJ/MJe 3.38 × 100

kg/MJe 1.75 × 10−3

m /MJe 1.20 × 10−3

m a/MJe 2.38 × 10−3

kg CO2 -eq./MJe 1.48 × 10−1

kg C6 H4 Cl2 -eq./MJe 6.40 × 10−3

kg C6 H4 Cl2 -eq./MJe 9.55 × 10−2

Unit Cement Steel Aluminium Copper Brass PET PUR-foam PP Low grade PS ABS PC Butadiene Cao NaOH NH3

MJ/kg 3.34 × 100 3.27 × 101 1.75 × 102 3.14 × 101 4.61 × 101 5.09 × 101 6.57 × 101 2.83 × 101 4.43 × 101 5.86 × 101 9.49 × 101 4.78 × 101 4.89 × 100 2.43 × 101 4.39 × 101

kg/kg 5.60 × 10−3 1.29 × 100 2.30 × 10−1 5.27 × 101 2.38 × 101 2.02 × 10−1 6.36 × 10−2 1.39 × 10−3 1.07 × 10−2 1.24 × 10−2 3.02 × 10−3 1.70 × 10−3 1.51 × 10−3 8.48 × 10−2 7.09 × 10−2

m3 /kg 2.39 × 10−3 1.26 × 10−2 2.98 × 10−2 1.38 × 10−1 6.40 × 10−2 1.43 × 10−2 1.19 × 10−1 4.71 × 10−3 8.77 × 10−3 1.08 × 10−2 1.33 × 10−2 9.66 × 10−3 6.60 × 10−2 1.42 × 10−2 4.02 × 10−3

m2 a/kg 5.16 × 10−3 5.83 × 10−2 1.17 × 10−1 6.31 × 10−1 2.14 × 10−1 6.18 × 10−2 1.84 × 10−2 1.65 × 10−4 3.00 × 10−4 3.65 × 10−3 7.04 × 10−4 3.52 × 10−4 6.46 × 10−4 3.48 × 10−2 1.85 × 10−2

kg CO2 -eq./kg 7.60 × 10−1 1.51 × 100 1.08 × 101 1.34 × 100 2.45 × 100 2.89 × 100 4.84 × 100 1.97 × 100 3.50 × 100 4.39 × 100 7.78 × 100 3.91 × 100 9.76 × 101 1.10 × 100 2.10 × 100

kg C6 H4 Cl2 -eq./kg 2.20 × 10−2 1.23 × 10−1 8.44 × 10−1 1.17 × 102 1.87 × 101 1.70 × 10−1 2.17 × 10−1 7.66 × 10−3 1.83 × 10−2 5.39 × 10−2 5.49 × 10−1 1.52 × 10−2 4.00 × 10−3 3.84 × 10−1 2.05 × 10−1

kg C6 H4 Cl2 -eq./kg 4.66 × 10−2 5.75 × 10−1 3.65 × 100 4.06 × 102 8.34 × 101 9.58 × 10−1 5.24 × 10−1 1.72 × 10−2 4.40 × 10−2 1.30 × 10−1 5.96 × 10−1 3.84 × 10−2 1.21 × 10−2 1.11 × 100 3.86 × 10−1

Unit Transport lorry [28 t]

MJ/tkm 2.95 × 100

kg/tkm 8.73 × 10−3

m3 /tkm 7.51 × 10−4

m2 a/tkm 2.76 × 10−3

kg CO2 -eq./tkm 1.94 × 10−1

kg C6 H4 Cl2 -eq./tkm 8.74 × 10−3

kg C6 H4 Cl2 -eq./tkm 2.18 × 10−2

the indirect savings related to the recovery of end products (materials and/or energy carriers). The excavation of raw materials needed for the production of metals and plastics as well as for energy generation, along with the landfill of residual fractions of these processes, indeed require large surfaces of land. These impacts can be avoided when materials and energy are recovered in a waste treatment process. The treatment strategy “recycle + energy recovery” has the lowest impact on land use, avoiding 15 m2 y of land occupation per t of treated ASR, about four times more than for the treatment strategy “recycle + landfill”. For the indicator global warming, the outcome is different for the short and long term approach. On short term perspective the order of preference for the different treatment strategies is: “recycle + landfill” > “recycle + energy recovery” > “landfill” > “energy recovery”. On long term perspective this order of preference changes to: “recycle + energy recovery” > “energy recovery” > “recycle + landfill” > “landfill”. This can be explained by the fact that it was assumed that on short term (100 years) only 3% of the carbon contained in ASR is degraded, while on long term (60,000 years) about 35% of the carbon will ultimately be

degraded, giving 70% of methane and 30% of carbon dioxide. As the GWP for methane is 25 times higher than that of carbon dioxide, the estimated carbon contribution to global warming for landfilling of ASR in long term perspective will be substantially higher than the one for the strategies with energy recovery, although during incineration almost 100% of the carbon is emitted as carbon dioxide. Next to the direct contribution of landfilling, also the indirect contribution due to the consumption of cement (stabilization of the waste) has a substantial impact, accounting for about half of the total impact on global warming for the scenario “landfill”. For the indicator human toxicity, the short and long term approach also lead to a different outcome. Although the order of preference for the different treatment strategies remains the same in both cases, the difference in impact between the strategy “landfill” and the other strategies is much higher for the long term perspective. This can be explained by two facts: on the one hand the impact for a residual landfill increases due to an increase in leaching of heavy metals (Table 4), together with larger impact factors for the indirect contributions (Table 5); and on the other hand the avoided impact increases for the other treatment strategies due to

Fig. 2. Sustainability assessment of ASR treatment strategies, graphically presented with landfill set at 1 (100%).

106 161 133 177 472 −675 12 −812 3844 1614 841 −325

533 −2617 −383 −3000

Short term kg C6 H4 Cl2 -eq./t ASR Long term kg CO2 -eq./t ASR

637 −641 841 −325 8.7 −3.6 −11.5 −14.6 3.6 −408 −48.2 −438 1.8 −13.1 −24.6 −26 Landfill Recycle + landfill Energy recovery Recycle + energy recovery

1.7 −4.3 −5.2 −7.8

Short term kg CO2 -eq./t ASR m2 y/t ASR m3 /t ASR kg Fe-eq./t ASR GJ/t ASR Unit

Land use Material intensity Energy intensity

Water consumption

energy and material recovery. The difference in impact between the strategy “recycle + landfill” and “landfill” can be attributed to the recovery of materials; the lower impact of “recycle + energy recovery” compared to that of “recycle + landfill” can be attributed to the recovery of energy. For the indicator treatment cost the order of preference for the different treatment strategies differs significantly from that for the other sustainability indicators. Average treatment costs for the different ASR treatment processes have been estimated using literature data: landfill costs on average D 106/t industrial waste, recycling D 100/t industrial waste and incineration D 120/t industrial waste (GHK-Bio, 2006; Kanari et al., 2003; Moakly et al., 2010; OVAM, 2012). When calculating the total treatment cost of a treatment strategy, the costs of the separate treatment steps have to be aggregated. For instance, the treatment cost of the strategy recycle + landfill consists of the sum of the cost for recycling of 1 t of the total ASR fraction and the cost for landfilling of 580 kg of the residual fraction. This amounts to the treatment cost of D 161/t industrial waste, for the treatment strategy recycle + landfill (Tables 4 and 6). From the results presented in Table 6 and Fig. 2, it appears that “landfill” is the cheapest treatment strategy (resulting in the lowest impact), followed by “energy recovery” and “recycle + landfill”; the treatment strategy “recycle + energy recovery” is the most expensive.

5. Discussion

Treatment strategy

Table 6 Sustainability assessment of ASR treatment strategies.

Global warming

Human toxicity

Long term kg C6 H4 Cl2 -eq./t ASR

D /t ASR

I. Vermeulen et al. / Resources, Conservation and Recycling 69 (2012) 17–28

Treatment cost

26

Comparing the results of the present sustainability assessment, with these of previously executed evaluations of ASR treatment processes (Boughton and Horvath, 2006; Ciacci et al., 2010; Duval and Maclean, 2007; Passarini et al., 2012), some similarities as well as differences appear. Boughton and Horvath (2006), Ciacci et al. (2010) and Passarini et al. (2012) all compare environmental aspects of the same waste treatment strategies for ASR: thermo-chemical treatment, energy recovery, material recovery and landfill of ASR. While Boughton and Horvath (2006) conclude that energy recovery from ASR is the most advantageous treatment option, Ciacci et al. (2010) and Passarini et al. (2012) conclude that material recovery and thermo-chemical treatment are the most promising treatment strategies to reach the European ELV targets. All these studies use an LCA based methodology, Boughton and Horvath (2006) used the CML methodology, Ciacci et al. (2010) and Passarini et al. (2012) used the Ecoindicator 99 methodology. Duval and Maclean (2007), on the other hand, compare environmental and economic aspects of the ASR treatment strategy currently used in Canada, with a scenario based on an increased recycling of ASR. They conclude that in spite of the environmental benefits, the added costs of the recycling scenario make it unprofitable. It is clear that the conclusions of these previously executed studies differ and largely depend on their basic assumptions and system boundaries. Agreement exists only on the fact that landfill of ASR should be the least preferred option. In contrast to these studies, the present paper provides a sufficient and relevant set of indicators to assess and compare sustainability of industrial waste treatment options, not only the environmental and/or economical impact. As also stated in Section 1, the applied methodology allows to obtain information on sustainability as a whole, without necessitating the use of specialized, 1D, indicators on each of the dimensions for which data are often not available. The novelty of this research consists of the selection of this specific set, covering all important issues of industrial waste treatment and making them quantifiable in terms of sustainability, and of the in depth discussion of the application of these

I. Vermeulen et al. / Resources, Conservation and Recycling 69 (2012) 17–28

indicators, making a clear distinction between direct and indirect contributions. The application of the proposed sustainability assessment method on ASR as a case (Sections 3 and 4), clearly proves the applicability of this method. All required data to define the sustainability indicators were available for each of the considered treatment strategies (Section 3, Table 2). Only for the indicator treatment cost no case specific data could be found, so that averaged literature values were used. Making a distinction between direct and indirect contributions to sustainability make the results easily interpretable and enable the identification of the bottlenecks/strengths of the considered treatment strategies with regard to sustainability (Section 4, Table 6, Fig. 2). From the present study it appears that for landfilling, the consumption of cement for solidification of the waste is the most important contributor to the overall sustainability, while for the other ASR treatment strategies the avoided sustainability impacts due to the recovery of products and/or energy carriers is determinant for the overall sustainability. It must however be remarked that the results and conclusions of the evaluation of ASR treatment strategies are valid within certain limits. Except for treatment costs, all used inventory data were derived from specific ASR treatment processes, i.e. a residual material landfill using cement for waste solidification, an Argonne PST plant, a fluidized bed combustor of the SLECO type. Other landfill types may use other auxiliaries than cement, other PST plants may recover less or lower grade materials than assumed in the present study or other energy recovery plants may recover less or lower grade energy carriers. This means that for these processes, input and output data, and thus results may slightly deviate from these in the present study. This set of sustainability indicators has been selected specifically to evaluate industrial waste treatment processes in Europe (Section 2.3); for other regions or case specific studies, some adjustments to the currently proposed set can be necessary. Finally, it must be remarked that the goal of this study is to provide an overall sustainability assessment method of industrial waste treatment processes. By using a limited set sustainability indicators, information is obtained on 3 dimensions of sustainability without necessitating the use of specialized, 1D, indicators on each of the dimensions for which often data are not available. Additional research within the field of specialized 1D indicators, especially social and economic indicators, can only be encouraged so that social or economic evaluations can be performed more easily.

6. Conclusions To date, numerous indicators are applied to evaluate and compare products and processes. The present study proposes a set of seven sustainability indicators (energy intensity, material intensity, water consumption, land use, global warming, human toxicity, treatment cost) suitable for assessing and comparing industrial waste treatment processes in terms of sustainability. The practical application of these indicators is discussed in depth. Distinction is made between direct contribution of a process to sustainability indicators on process level itself, and indirect contributions, occurring upstream or downstream of the process and mainly related to production of auxiliaries and recovery of end products. The sustainability assessment method is applied to the most commonly proposed and commercially applied treatment processes of ASR. It could be concluded that: • The set of indicators was found to address efficiently all aspects of sustainability, without overlap, and to satisfy all required characteristics: quantifiable, simple, unambiguous, useful to

•

•

•

•

27

management decisions, robust and reproducible. They are thus suitable for practical use. From the application to ASR treatment processes, it appeared that the treatment strategy recycling combined with energy recovery was the most sustainable option, whereas landfilling the least. The indirect sustainability impact due to the consumption of cement was found to be an important contributor to the general sustainability impact of landfill. The avoided sustainability impacts due to the recovery of energy carriers and products were found to be determinant for the sustainability impacts of all other treatment strategies. The most sustainable option, recycling combined with energy recovery, also enables to reach the European targets set by 2015, reaching an 88.4% reuse and recycling rate and 98.5% reuse and recovery rate. The proposed set of sustainability indicators is developed based on European conditions. The dimension of an indicator can be influenced by local conditions in other regions.

References Al-Salem SM, Lettieri P, Baeyens J. The valorization of plastic solid waste (PSW) by primary to quaternary routes: from re-use to energy and chemicals. Progress in Energy and Combustion Science 2010;36:103–29. Astrup T, Moller J, Fruergaard T. Incineration and co-combustion of waste: accounting of greenhouse gases and global warming contributions. Waste Management and Research 2009;27:789–99. Azapagic A, Perdan S. Indicators of sustainable development for industry. Process Safety and Environment Protection 2000;78(B4):243–61. Börjeson L, Löfvenius G, Hjelt M. Characterisation of automotive shredder residues from two shredding facilities with different refining processes in Sweden. Waste Management and Research 2000;18:358–66. Boughton B, Horvath A. Environmental assessment of shredder residue management. Resources, Conservation and Recycling 2006;47:1–25. Boughton B. Evaluation of shredder residue as cement manufacturing feedstock. Resources, Conservation and Recycling 2007;51:621–42. Ciacci L, Morselli L, Passarini F, Santini A, Vassura I. A comparison among different automotive shredder residue treatment processes. The International Journal of Life Cycle Assessment 2010;15:896–906. Council of the European Union. Council directive 2000/53/EC of 18 September 2000 on end-of-life vehicles. Official Journal of the European Communities 2000;L269:34–49. Duval D, Maclean HL. The role of product information in automotive plastics recycling: a financial and life cycle assessment. The Journal of Cleaner Production 2007;15:1158–68. Doka G. Life Cycle Inventories of Waste Treatment Services, Ecoinvent Rep No. 13. Dübendorf: Swiss Centre for Life Cycle Inventories; 2007. Donaj P, Blasiak W, Yang W, Forsgren C. Conversion of microwave pyrolysed ASR’s char using high temperature agents. The Journal of Hazardous Materials 2011;185:472–81. Eriksson O, Finnveden G. Plastic waste as a fuel–CO2 -neutral or not. Energy and Environmental Science 2009;2:907–14. European Commission Environment. The story behind the strategy – EU waste policy. http://ec.europa.eu/environment/waste/pdf/story book.pdf; 2011 (accessed 22.12.2011). Eurostat. End-of-life vehicles: reuse, recycling and recovery, totals. http://epp. eurostat.ec.europa.eu/portal/page/portal/waste/data/wastestreams/elvs; 2012 (accessed 23.06.2012). Ferrão P, Nazareth P, Amaral J. Strategies for meeting EU end-of-life vehicle reuse/recovery targets. Journal of Industrial Ecology 2006;10:77–93. Forton OT, Harder MK, Moles NR. Value from shredder waste: ongoing limitations in the UK. Resources, Conservation and Recycling 2006;46:104–13. Frischknecht R, Jungbluth N, Althaus HJ, Doka G, Heck T, Hellweg S, et al. Overview and methodology, Ecoinvent rep no. 1. Dübendorf: Swiss Centre for Life Cycle Inventories; 2007. Gallon N, Binder M. Life cycle inventory (LCI) of argonne’s process for recycling shredder residue. Leinfelden-Echterdingen: PE Europe GmbH by order of Argonne National Laboratory, Vehicle Recycling Partnership and American Plastics Council; http://www.es.anl.gov/Energy systems/CRADA Team/ 2006 publications/Final%20Report ANL%20LCA%20report 09292006.pdf; (accessed 23.06.2012). GHK-Bio Intelligence Service. A study to examine the benefits of the end of life vehicles directive and the costs and benefits of a revision of the 2015 targets for recycling, re-use and recovery under the ELV Directive, Final Report J2232. Birmingham: DG Environment; 2006. Goedkoop M, Heijungs R, Huijbregts MAJ, De Schryver A, Struijs J, Van Zelm R. ReCiPe 2008. A life cycle impact assessment method which comprises harmonised category indicators at the midpoint and the endpoint level, 1st ed., Report I: Characterisation. The Netherlands; 2009. Goedkoop M, De Schryver A, Oele M, de Roest D, Vieira M, Durksz S. SimaPro 7 Tutorial, PRé Consultants. www.pre.nl. (accessed 23.03.2012).

28

I. Vermeulen et al. / Resources, Conservation and Recycling 69 (2012) 17–28

Guinée BJ, Gorrée M, Heijungs R, Huppes G, Kleijn R, de Koning A, et al. Handbook on life cycle assessment—operational guide to the ISO standards. Dordrecht: Kluwer Academic Publishers; 2002. Kanari N, Pineau JL, Shallari S. End-of-life vehicle recycling in the European union. Journal of Management 2003;8:15–9. Kellenberger D, Althaus HJ, Künniger T, Lehman M, Thalmann P. Life cycle inventories of building products, Final Rep Ecoinvent data v2.0 No. 7. Dübendorf: Swiss Centre for Life cycle inventories; 2007. Mancini G, Tamma R, Viotti P. Thermal process of fluff: preliminary tests on a full scale treatment plant. Waste Management 2010;30:1670–82. Martins A, Mata T, Costa C, Sikdar S. Framework for sustainability metrics. Industrial and Engineering Chemistry Research 2007;46:2962–73. Mendes MR, Aramaki T, Hanaki K. Comparison of the environmental impact of incineration and landfilling in São Paulo City as determined by LCA. Resources, Conservation and Recycling 2004;41:47–63. Moakly J, Weller M, Zelic M. An evaluation of shredder waste treatments in Denmark. Bachelor thesis at Worcester Polytechnic Institute, USA; 2010. Morselli L, Santini A, Passarini F, Vassura I. Automotive shredder residue (ASR) characterization from valuable management. Waste Management 2010;30:2228–34. Nourreddine M. Recycling of auto shredder residue. Journal of Hazardous Materials 2007;A139:481–90. OVAM. Tarieven en capaciteiten voor storten en verbranden – Actualisatie tot 2009. Evolutie en prognose (Dutch article). Mechelen: Danny W, http://www.ovam.be (accessed 23.09.2012). NTREE. National round table on the environment and the economy. Eco-efficiency indicators workbook. Ottawa: Renouf Publishing; 2001. Pavlas M, Tous M, Klimek P, Bebar L. Waste incineration with production of clean and reliable energy. Clean Technologies and Environmental Policy 2011;13:595–605. Passarini F, Ciacci L, Santini A, Vassura I, Morselli L. Auto shredder residue LCA: implications of ASR composition evolution. The Journal of Cleaner Production 2012;23:28–36. Robson S, Goodhead TC. A process for incorporating automotive shredder residue into thermoplastic mouldings. Journal of Materials Processing Technology 2003;139:327–31. Roy C, Chaala A. Vacuum pyrolysis of automobile shredder residues. Resources, Conservation and Recycling 2001;32:1–27.

Salling P, Kicherer A, Dittrich-Krämer B, Wittlinger R. Eco-efficiency analysis by BASF: the method. The International Journal of Life Cycle Assessment 2002;7:1866–76. Schwarz J, Beloff B, Beaver E. Use of sustainability metrics to guide decision-making. Chemical Engineering Progress July 2002:58–63. Sikdar S. Sustainable development and sustainability metrics. AIChE Journal 2003;49(8):1928–32. Sikdar S. On aggregating multiple indicators into a single metric for sustainability. Clean Technologies and Environmental Policy 2009;11:157–61. Sikdar S. Measuring sustainability. Clean Technologies and Environmental Policy 2012;14:153–4. Simic V, Dimitrijevic B. Production planning for vehicle recycling factories in the EU legislative and global business environments. Resources, Conservation and Recycling 2012;60:78–88. Srogi K. An overview of current processes for the thermochemical treatment of automobile shredder residue. Clean Technologies and Environmental Policy 2008;10:235–44. Stehlik P. Efficient waste processing and waste to energy: challenge for the future. Clean Technologies and Environmental Policy 2009;11:7–9. Tarantini M, Buttol P, Maiorino L. An environmental LCA of alternative scenarios of urban sewage sludge treatment and disposal. Thermal Science 2007;11(3):153–64. Verbinnen B, Van Caneghem J, Vandecasteele C, Block C, Van Hooste H, Milieurapport Vlaanderen, Achtergronddocument 2010, Industrie. Vlaamse Milieumaatschappij, www.milieurapport.be (accessed 23.03.2012). Verfaillie HA, Bidwell R. Measuring eco-efficiency – a guide to reporting company performance. World Business Council for Sustainable Development; 2000. Vermeulen I, Van Caneghem J, Block C, Baeyens J, Vandecasteele C. Automotive shredder residue (ASR): reviewing its production from end-of-life vehicles (ELVs) and its recycling, energy or chemicals’ valorization. Journal of Hazardous Materials 2011;190:8–27. Vermeulen I, Van Caneghem J, Block C, Van Brecht A, Wauters G, Vandecasteele C. Energetic valorization of heavy ASR by co-incineration in a fluidized bed combustor. Clean Technologies and Environmental Policy 2012., http://dx.doi.org/10.1007/s10098-012-0474-5. De Vos S, Ligthart T, Görtzn J, Mulder E. LCA of thermal treatment of waste streams in the clinker production in Belgium. Apeldoorn: TNO Rep. Environ. Health Saf; 2007.