-otl

aA

DD

77/

810

2

+ihfjh

a*GMFTSH

9 NBSI 8-7597-06-259-879 )detnirp( NBSI 5-8597-06-259-879 )fdp( NSSI 4394-9971 )detnirp( NSSI 2494-9971 )fdp(

ytisrevinU otlaA

gnireenignE fo loohcS tnemnorivnE tliuB fo tnemtrapeD

if.otlaa.www

+ SSENISUB YMONOCE

+ TRA

+ NGISED ERUTCETIHCRA

+ ECNEICS

YGOLONHCET

REVOSSORC

LAROTCOD SNOITATRESSID

obl

haD

anel

eH

txe

tnoc

hsi

nniF

eht

ni sn

oitp

o tne

mega

nam

etsa

w gn

isse

ssa

ni se

hcao

rppa

des

ab e

lcyc

efi

L y

tisr

evi

nU

otla

A

8102

tnemnorivnE tliuB fo tnemtrapeD

sehcaorppa desab elcyc efiL etsaw gnissessa ni

eht ni snoitpo tnemeganam txetnoc hsinniF

oblhaD aneleH

LAROTCOD SNOITATRESSID

seires noitacilbup ytisrevinU otlaASNOITATRESSID LAROTCOD 77 / 8102

ni sehcaorppa desab elcyc efiL ni snoitpo tnemeganam etsaw gnissessa

txetnoc hsinniF eht

oblhaD aneleH

fo rotcoD fo eerged eht rof detelpmoc noitatressid larotcod A eht fo noissimrep eht htiw ,dednefed eb ot )ygolonhceT( ecneicS

dleh noitanimaxe cilbup a ta ,gnireenignE fo loohcS ytisrevinU otlaA enuJ 8 no gnidliuB SAUT ytisrevinU otlaA fo 1UT llah erutcel eht ta

.21 ta 8102

ytisrevinU otlaA gnireenignE fo loohcS

tnemnorivnE tliuB fo tnemtrapeD gnireenignE latnemnorivnE

rosseforp gnisivrepuS gnireenignE fo loohcS ytisrevinU otlaA ,iravroS anaaJ rosseforP

rosivda sisehT

EKYS etutitsnI tnemnorivnE hsinniF ,äläppeS iryJ rosseforP

srenimaxe yranimilerP ynamreG ,nilreB UT ,reniebkniF saihttaM rosseforP kramneD ,UTD ,tneruaL sixelA rosseforP etaicossA

tnenoppO

nedewS ,elväG fo ytisrevinU ,nosskirE alO rosseforP

seires noitacilbup ytisrevinU otlaASNOITATRESSID LAROTCOD 77 / 8102

© 8102 oblhaD aneleH

NBSI 8-7597-06-259-879 )detnirp( NBSI 5-8597-06-259-879 )fdp( NSSI 4394-9971 )detnirp( NSSI 2494-9971 )fdp(

:NBSI:NRU/if.nru//:ptth 5-8597-06-259-879

yO aifarginU iknisleH 8102

dnalniF

tcartsbA otlaA 67000-IF ,00011 xoB .O.P ,ytisrevinU otlaA if.otlaa.www

rohtuA oblhaD aneleH

noitatressid larotcod eht fo emaN txetnoc hsinniF eht ni snoitpo tnemeganam etsaw gnissessa ni sehcaorppa desab elcyc efiL

rehsilbuP gnireenignE fo loohcS

tinU tnemnorivnE tliuB fo tnemtrapeD

seireS seires noitacilbup ytisrevinU otlaA SNOITATRESSID LAROTCOD 77 / 8102

hcraeser fo dleiF tnemeganam etsaW

dettimbus tpircsunaM 8102 yaM 2 ecnefed eht fo etaD 8102 enuJ 8

)etad( detnarg hsilbup ot noissimreP 8102 hcraM 91 egaugnaL hsilgnE

hpargonoM noitatressid elcitrA noitatressid yassE

tcartsbAdna ecracs gnimoceb era secruoser larutan elbawener-non ,gniworg era etsaw fo stnuomA

sti esaercni ot dna smelborp eseht elkcat oT .demusnoc-revo gnieb era secruoser elbawener ot yhcrareih etsaw eht dna seigetarts etsaw detnemelpmi sah eporuE ssenevititepmoc cimonoce

si yhcrareih eht morf noitaiveD .yrevocer etsaw esimixam dna noitareneg etsaw esiminim stnemssessa desab )CL( elcyc efiL .emoctuo latnemnorivne tseb eht hcaer ot dedeen fi elbatpecca etsaw dna efil-fo-dne elbaniatsus tsom eht yllacimonoce dna yllatnemnorivne dnfi ot deriuqer era etsaw gnidulcni smetsys MW dna swofl etsaw hsinniF ,yduts siht nI .snoitulos )MW( tnemeganam elcyc efil laicos ,)ACL( stnemssessa elcyc efil htiw dessessa erew rotces tserof eht morf stcudorp ro

ehT .)AFM( sesylana wofl lairetam dna )CCLE( gnitsoc elcyc efil latnemnorivne ,)CCLS( gnitsoc stnemssessa cimonoce dna latnemnorivne desab-CL morf stluser rehtehw etaulave ot saw mia

sehcaorppa fo noitanibmoc eht rehtehw tuo dnfi ot dna dnalniF redro yhcrareih etsaw eht troppus gnidnatsrednu eht evorpmi ot loot a sa noitasilamroN .gnikam-noisiced rof eulav dedda sedivorp cfiiceps-esac eht morf stluser ehT .deiduts osla saw stluser )AICL( tnemssessa tcapmi elcyc efil fo

MW fo redro yhcrareih etsaw eht troppus ylraelc ton did stnemssessa cimonoce dna latnemnorivne nI .yrevocer ygrene htiw noitarenicni ro gnilcycer naht esrow demrofrep gnillfidnaL .snoitpo

ygrenE .gnigakcap elbasuer naht retteb demrofrep gnigakcap elbalcycer fo esu eht ,tsartnoc eht ,gnilcycer naht stcapmi latnemnorivne llarevo rewol decudorp etsaw repapswen fo yrevocer

ot eud taht dewohs CCLS eht ,noitidda nI .revewoh ,seirogetac tcapmi laudividni ni deirav stluser denibmoc ehT .gnilcycer naht evisnepxe erom saw yrevocer ygrene ,stsoc latnemnorivne rehgih sti

,smetsys MW fo ecnamrofrep eht gnivorpmi rof slaitnetop delaever sehcaorppa desab-CL fo esu rof snoitcarf tnereffid fo laitnetop eht defiitnedi CCLE dna ACL ,AFM fo noitanibmoc eht .g.e

etsaw noitilomed dna noitcurtsnoc fo metsys MW eht fo ecnamrofrep llarevo eht gnivorpmi rof tegrat gnilcycer %07 eht gniveihca taht detacidni tnemssessa tnioj eht ,yllanoitiddA .)WD&C( yllanoiger taht detacidni stluser ehT .metsys eht ni stnemevorpmi rojam seriuqer 0202 ni WD&C saw stluser AICL fo gnidnatsrednu ehT .eporuE nihtiw deriuqer eb yam stegrat gnilcycer gnireffid

noitacinummoc eht ni sehcaorppa CL eht no noitamrofni gnidulcni yb devorpmi tnetxe emos ot seulav ecnerefer noitasilamron tnereffid fo esu eht oslA .srenoititcarp-non ACL rof lairetam

ygrene ro lairetam eht ot eud dediova stcapmi ehT .ytilibadnatsrednu eht no tcapmi dah desaercni lanoitarepo eht ni segnahc tub ,gnilledom MW ni srotcaf yek eht fo eno era etsaw fo yrevocer

ytilibacilppa eht timil dna gnilledom siht etacilpmoc erutuf eht tuoba ytniatrecnu dna tnemnorivne evah thgim slairetam yradnoces fo esu eht ni dna rotces ygrene eht ni segnahC .stluser eht fo

era stnemssessa cimetsys dna MW fo ytilibaniatsus latnemnorivne eht no stceffe gnitciflnoc .deriuqer

sdrowyeK etsaw ,noitasilamron ,sisylana wofl lairetam ,gnitsoc elcyc efil ,tnemssessa elcyc efil doow ,etsaw noitilomed dna noitcurtsnoc ,etsaw dilos lapicinum ,yhcrareih

)detnirp( NBSI 8-7597-06-259-879 )fdp( NBSI 5-8597-06-259-879

)detnirp( NSSI 4394-9971 )fdp( NSSI 2494-9971

rehsilbup fo noitacoL iknisleH gnitnirp fo noitacoL iknisleH raeY 8102

segaP 531 nru :NBSI:NRU/fi.nru//:ptth 5-8597-06-259-879

ämletsiviiT otlaA 67000 ,00011 LP ,otsipoily-otlaA if.otlaa.www

äjikeT oblhaD aneleH

imin najriksötiäV assitsketnok nemouS niitnioivra nejotheothiavotlouhetäj tavatsimytsehäl tesiajhopiraaknilE

ajisiakluJ uluokaekrok nedieteitiröönisnI

ökkiskY sotial nötsiräpmy nutennekaR

ajraS seires noitacilbup ytisrevinU otlaA SNOITATRESSID LAROTCOD 77 / 8102

alasumiktuT otlouhetäJ

mvp neskutiojrikisäK 8102.50.20 äviäpsötiäV 8102.60.80

äviäpsimätnöym navulusiakluJ 8102.30.91 ileiK itnalgnE

aifargonoM ajriksötiävilekkitrA ajriksötiäveessE

ämletsiviiTajoravnonnoul nikaivutuisuu aj tavenekkuin taravnonnoul tamottamutuisuu ,tavavsak täräämetäJ no appooruE iskesimatnarap nyvykuliaplik aj iskesimesiaktar neimlegno nediäN .aakiil naatetuluk

naamiomiskam aj nenimytnys neettäj naamiominim näätiryp alloj ,naikrareihetäj tunitaal .mm no es soj ,atekiop naadiov neskytsejräjajisute nollouhetäj ile atsaikrareihetäJ .nenimätnydöyh

netsunnatsuk aj nötsiräpmY .iskesimaasnaakia nusiaktar naahrap atlannak nötsiräpmy neeprat aisiajhopiraaknile naativrat iskesimätyöl nejotheothiav nollouhetäj neipmivätsek atlannak

ätiettäj aivelo nisiärep atsiettout nirotkesästem ,aisialamous ätiäre ässöyt ässäT .äimletenem ,)ACL( aitnioivrairaaknile neättyäk niitioivra äimletsejräjotlouhetäj aj ajotrivetäj äivätläsis

netsillötsiräpmy ,)CCLS( aitnioivra netsunnatsukiraaknile netsillannuksiethy ilo aneettiovaT .)AFM( aisyylana-atrivilaairetam aj )CCLE( aitnioivra netsunnatsukiraaknile

naikrareihetäj tuletsakratsunnatsuk aj -ötsiräpmy tesiajhopiraaknile oktavekut äättivles .aoketneskötääp allamekut aovraäsil nenimätsidhy neimletenem okout aj assemouS atsimattaduon

änävääsil ättyyvätterrämmy netsolut nACL anesillodham aitniosilamron niittiktut iskupoL tävie teskolutsutukiavsunnatsuk aj -ötsiräpmy assiskumiktutsuapat ässiydheT .änämletenem

tammeruus tavilo teskutukiavötsiräpmy neskutiojisakkiapotaaK .aaikrareihetäj itsävles teenekut niuk tämmeneip neskuakkap nävättetärreik attum ,nesimätnydöyhaigrene iat neskytärreik niuk

tavilo teskutukiavötsiräpmy nesimätnydöyhaigrene neettäjithelamonaS .nävättetyäkneelleduu iskäsiL .tavilethiav niknetiuk teskolut neikkoulsutukiavötsiräpmy ire ,neskytärreik niuk tämmeneip

navelo netsunnatsukötsiräpmy nesimätnydöyhaigrene ittioso itnioivra netsutukiavsunnatsuk itsajlap nenimätsidhy neimletenem netsiajhopiraaknilE .neskytärreik niuk tammaekrok

,isyylana-atrivilaairetam ällämätsidhy .mm ;atiethoksunnarap neimletsejräjotlouhetäj aiskuusillodham aisiathokeajetäj aatsinnut niitiov uletsakratsunnatsuk aj itnioivrairaaknile

,niittetioso iskäsiL .aiskutukiav nämletsejräjotlouhetäj neettäjukrup aj -sunnekar okok äätneneip äättyllede 0202 neetouv sutuvaas neettiovatsytärreik % 07 neettäjukrup aj -sunnekar ätte

nediettiovatsytärreik neivelethiav itsesilleeula tavittioso teskoluT .ässämletsejräj okok aiskunnarap nirääm niassoj niittytsyp ättyyvätterrämmy netsolut-ACL .naajis netsialappooruesiely neeprat

tesialire söyM .niilaairetamätnitseiv teskuavuk neimletenem ällämättiil naamatnarap asseskunnillam neimletsejräjotlouhetäJ .neetyyvätterrämmy tavittukiav täjiketitniosilamron

nesimätnydöyhaigrene aj -ilaairetam teskutelo tavo ätiöjiketsuumraväpe äipmivättikrem suumraväpe aj teskotuum tavuthapat ässötsiräpmyatnimioT .ätsiötsääsötsääp atsisillodham

aj irotkesaigrenE .äöttyäk netsolut tavattiojar aj atsunnillam tavattuekiav atseduusiavelut aiskutukiav aisiakkatsav allo iov alliskotuum nötyäk nedienia-akaarsytärreik

.ajetnioivrasutukiav äisimeetsys naativrat netoj ,neetyyvätsekötsiräpmy neimletsejräjotlouhetäj

tanasniavA ,isyylana-atrivilaairetam ,atneksalsunnatsukiraaknile ,itnioivrairaaknile uup ,etäjukrup aj -sunnekar ,etäjatnuksydhy ,aikrareihetäj ,itniosilamron

)utteniap( NBSI 8-7597-06-259-879 )fdp( NBSI 5-8597-06-259-879

)utteniap( NSSI 4394-9971 )fdp( NSSI 2494-9971

akkiapusiakluJ iknisleH akkiaponiaP iknisleH isouV 8102

äräämuviS 531 nru :NBSI:NRU/fi.nru//:ptth 5-8597-06-259-879

1

Acknowledgements

This has been a long journey and a learning process during which I have worked at the Finn-ish Environment Institute SYKE in several research projects with a whole bunch of inspiring, interesting and lovely people. I greatly appreciate the contribution of all my colleagues in-volved in these projects and especially of all the co-authors of the scientific papers. The finan-cial support granted for the projects primarily by Business Finland (previously Tekes, Finnish Funding Agency for Technology and Innovation) and numerous companies from the waste management sector of Finland is gratefully acknowledged. I want to express my gratitude to my institute and directors for giving me the opportunity to work with extremely interesting research questions and tolerating the too long process of producing this thesis. Moreover, I am grateful to Adj. Prof. Timo Assmuth for hiring me to SYKE in the first place and for lead-ing me to the path of waste management research. Having struggled with the work quite some years, with some years break in between, I got a new sparkle from Prof. Jaana Sorvari who encouraged me to dive into the work again and finalise the thesis. The motivation and support from Jaana and my superior Prof. Jyri Seppälä, who at this time started as my thesis advisor, made me believe in my possibility to complete the work. I am grateful for their inspiring and patient support in writing the sum-mary. Furthermore, I want to thank my long time colleague, Sirkka Koskela, for providing tutorial co-operation in our research projects and a model for a never ending interest in de-veloping oneself and learning new things. I gratefully acknowledge the time taken by the preliminary examiners, Assoc. Prof. Alexis Laurent and Prof. Martin Finkbeiner, to review the manuscript of my thesis. I appreciate the valuable and constructive comments they gave, they helped me significantly improve my work. The working atmosphere could not have been better than within the Center for Sustainable Consumption and Production at SYKE. Words are not enough to express my gratitude for the empathy I have received from my colleagues during the past three years of heavy and painful private life. I want to hug you all for making my life somewhat lighter and turning work into fun. Special thanks go to my closest colleagues Tuuli Myllymaa and Hanna Salmenperä for being open for inspiring discussions on wastes and offering shoulders to lean on. I am indebted to all my dear friends and my great family network of sisters and brothers, their children and grand-children. Knowing you are there gives me confidence in being able to confront everything in life. I am especially grateful to Margareta and Markku, family

2

friends for almost twenty years, for sharing the joys and griefs of life, providing care and laughter and many-sided discussions about everything in life. I also send my warmest thoughts to my parents, who always supported the education of their children. My father Vil-ho, who never had the possibility to go to school for more than four years himself, would have been extremely proud of me. Finally, I want to express my love and gratitude to the dearest of all, my son Johan and his girlfriend Linda, for reminding me of the power of passion in whatever you do, for taking me out to be alive and for sharing their young energy with me. I dedicate this thesis to my late husband Kim, who believed in me when I didn’t, and who kept on rolling and filling my life with music and movement during the 23 years we had to-gether. Kim, I made it. Helsinki, 28 April 2018 Helena Dahlbo

3

Contents

1. Introduction ............................................................................................................................................ 9

1.1 Development of waste management ..................................................................................................... 9

1.2 Life cycle assessment and waste management ................................................................................... 12 1.2.1 Environmental life cycle assessment as a support to waste policy ................................................. 12 1.2.2 Life cycle assessment methodology for evaluating waste hierarchy .............................................. 13 1.2.3 Life cycle assessment studies on waste management .................................................................... 18

1.3 Broadening life cycle assessment with cost and material assessments ............................................... 21

1.4 Aims, focus and structure of the thesis ................................................................................................ 24

2. Materials and methods .......................................................................................................................... 26

2.1 Case studies.......................................................................................................................................... 26

2.2 System descriptions and modelling assumptions ................................................................................. 28

2.3 Data sources and methods for life cycle assessment ........................................................................... 31

2.4 Life cycle costing and material flow analysis ....................................................................................... 33

3. Results and discussion ............................................................................................................................ 34

3.1 Environmental and economic impacts on different waste hierarchy levels ......................................... 34

3.2 Added value for decision-making of combining life cycle based approaches ...................................... 37

3.3 Using normalisation to improve the understanding of LCIA results ..................................................... 40

3.4 Uncertainties of the assessments......................................................................................................... 41

4. Conclusions and recommendations for future research ......................................................................... 45

References ...................................................................................................................................................... 49

5

List of Abbreviations

CCB Corrugated cardboard

C&DW Construction and demolition waste (C&DW) arises from activities such as the construction and total or partial demolition of buildings and civil infrastructure, road planning and maintenance. Different definitions are applied throughout the EU. C&DW can consist of numerous materials, including concrete, bricks, gypsum, wood, glass, metals, plastic, solvents, asbestos and excavated soil. In this study rocks, gravel, soil or other earth-based materials were excluded from C&DW, since they are also excluded from the 70% recycling target given for C&DW in the Waste Framework Directive (2008/98/EC).

ELCC Environmental life cycle costing (ELCC) summarises all costs associated with the life cycle of a product that are directly covered by one or more of the actors in that life cycle. These costs must relate to real money flows. Environmental impacts are not monetised, but ELCC is complemented with a life cycle assess-ment (LCA), using equivalent system boundaries and functional units. (Hunkeler et al., 2008).

FU A functional unit (FU) provides a reference to which all the studied systems inputs and outputs are related. It gives a quantitative description of the primary function or service that the analysed system provides. (European Commission, 2011; ISO, 2006a; Laurent et al., 2014b).

GHG Greenhouse gas

HDPE High-density polyethylene

LCA A life cycle assessment (LCA) is a quantitative tool that helps in implementing life cycle thinking (LCT). An LCA is used for analysing and assessing environ-mental impacts of a material, product or service during its entire life cycle. The full life cycle of a product is taken into account.

LCC Conventional life cycle costing (LCC) is based on purely economic evaluation, considering various stages in the life cycle. However, LCC does not always con-sider the complete life cycle, for example end-of-life operations are not includ-ed. LCC generally includes conventional costs associated with a product that are borne directly by a given actor (Hunkeler et al., 2008).

LCI A life cycle inventory (LCI) is the second component of a life cycle assessment in which an analysis is made of the environmental interventions associated with the processes required for that functional product unit.

6

LCIA A life cycle impact assessment (LCIA) is the third component of a life cycle as-sessment in which the data gathered in the inventory analysis are interpreted and assessed in terms of their environmental impact potential.

LCT Life cycle thinking (LCT) means a qualitative or quantitative assessment of up-stream supply chains, downstream processes, and the use of products in terms of environmental, social, and economic considerations. It is applied to analysing the performance of, and improvement options for, both goods and services. An LCT accounts for the full life cycle of a product or service.

MFA Material flow analysis (MFA) is a tool to analyse the transformation, transporta-tion, or storage of materials within a defined system. (Brunner and Rechberger, 2004).

MSW Municipal solid waste (MSW) refers to waste generated in households and waste comparable to household waste generated in production, especially in the service industries. The general common feature of municipal waste is that it is generated in the consumption of final products in communities and is covered by municipal waste management systems. (Statistics Finland, 2015)

SFA Substance flow analysis (SFA) is a specific type of MFA tool used for tracing the flow of a selected chemical (or group of substances) through a defined system.

SLCC Social life cycle costing (SLCC) looks at the costs associated with the entire life cycle of a product or a service from the viewpoint of the whole society. Various cost items can be distinguished. The most conventional costs are direct costs, i.e. the costs of investments, labour, energy, and so on. SLCC also includes all external costs associated with environmental impacts of the product’s or ser-vice’s life cycle. The stakeholders of SLCC include governments and other public bodies not directly concerned with the product system. Therefore the costs do not include transfer payments, such as subsidies and taxes, because they are in-ternal to the system. (Hunkeler et al., 2008; Paper II).

SWMS Solid waste management system

WM Waste management

WtE Waste to energy

7

List of Publications

This doctoral dissertation consists of the present summary and of the following papers which are referred to in the text by their numerals:

1. Dahlbo, H., Koskela, S., Laukka, J., Myllymaa, T., Jouttijärvi, T., Melanen, M., Tenhunen, J. 2005. Life cycle inventory analyses for five waste management options for discarded news-paper. Waste Management & Research, 22(4): 291-303.

2. Dahlbo, H., Ollikainen, M., Peltola, S., Myllymaa, T. & Melanen, M. 2007. Combining eco-logical and economic assessment of options for newspaper waste management. Resources, Conservation and Recycling 51:42-63.

3. Dahlbo, H., Koskela, S., Pihkola, H., Nors, M., Federley, M. & Seppälä, J. 2013. Compari-son of different normalised LCIA results and their feasibility in communication. International Journal of Life Cycle Assessment 18(4):850-860.

4. Dahlbo, H., Bachér, J., Lähtinen, K., Jouttijärvi, T., Suoheimo, P., Mattila, T., Sironen, S., Myllymaa, T., Saramäki, K. 2015. Construction and demolition waste management – holistic assessment of environmental performance. Journal of Cleaner Production 107:333-341.

5. Koskela, S., Dahlbo, H., Judl, J., Korhonen, M.-R. & Niininen, M. 2014. Reusable plastic crate or recyclable cardboard box? A comparison of two delivery systems. Journal of Cleaner Production 69(3):83-90.

8

Author’s Contribution

Publication 1: H. Dahlbo and M. Melanen planned the study. H. Dahlbo was responsible for writing the paper. J. Laukka and T. Myllymaa performed the life cycle inventories with guidance from S. Koskela. The other authors commented on the paper.

Publication 2: H. Dahlbo was responsible for the life cycle impact assessments. M. Ol-likainen and S. Peltola were in charge of the development and calculations concerning the social life cycle cost concept. The paper was written jointly by H. Dahlbo, M. Ollikainen and S. Peltola. M. Melanen provided comments on the paper.

Publication 3: H. Dahlbo and S. Koskela initiated and planned the paper. H. Dahlbo pro-duced the LCIA and analyses on normalisation factors. H. Pihkola contributed by providing support on communication issues. The other authors provided comments on the paper. The conclusions were written jointly by H. Dahlbo and S. Koskela.

Publication 4: H. Dahlbo initiated, planned and structured the paper with support from J. Bachér, K. Lähtinen and T. Mattila. H. Dahlbo was responsible for conducting the LCA and preparing the paper. Other authors contributed by providing results from the other assess-ments and comments on the paper. The conclusions were written jointly by H. Dahlbo, J. Bachér and K. Lähtinen.

Publication 5: S. Koskela was responsible for the overall paper. H. Dahlbo was responsible for considering the end-of-life aspects of the studied products. J. Judl conducted the life cycle modelling. The other authors commented on the paper.

9

1. Introduction

1.1 Development of waste management

In early human history, people generated little waste and reused and repaired what they could. As city populations grew, waste management (WM) systems became necessary. The first known European waste statute was given in Athens in 500 BC. Regulations required waste to be deposited at least one and a half kilometres outside the city limits. (Barbalace, 2003; Nygård, 2016). With growing cities, waste began piling up and threatening human health due to the presence of rats and other pests as well contaminated water supplies that led to the transmission of diseases. During the 14th century, orders and restrictions began to emerge for taking better care of wastes. The Black Death that killed approximately 25 million people in Europe in the mid-14th century, was one factor that affected European WM and sanitation. At least partly due to this, in 1388, the English parliament forbade waste dispersal in public waterways and ditches. In 1400 the garbage piles outside Paris gates were been reported to have been so high that they affected the city’s defences (Barbalace, 2003; Nygård, 2016). The oldest street sanitation statutes in Finland were given in the 13th century in Sweden. They embodied orders on the property owners’ responsibility for keeping the streets clean and sanctions for dereliction of these duties. (Nygård, 2016). More organised waste collection and disposal systems aiming to maintain public health were launched in the 18th century, when the Industrial Revolution gave rise to rapid growth of both production and amounts of waste (Barbalace, 2003; Nygård, 2016). Since then, WM has developed towards material and resource management, which takes into consideration envi-ronmental aspects in addition to health and safety aspects. Sophisticated technologies are now applied for the sorting, separating, processing, recovering and treating various types of waste and the valuable components it contains. Due to growing the population and increasing consumption and production, the amount of waste is constantly increasing both in Europe and globally. Simultaneously, non-renewable natural resources are becoming scarce, while renewable resources are being over-consumed. Several attempts have been made and indicators developed to assess and predict the sustain-ability of the human use of global natural resources. For example, the Ecological Footprint measures the supply of nature (i.e., the planet’s biologically productive land areas including forests, pastures, cropland and fisheries), and the demand on nature (i.e., the productive area

Introduction

10

required to provide the renewable resources humanity is using and to absorb its waste) (Global Footprint Network, 2016). According to the 2014 official Ecological Footprint as-sessments on the world and national level, humanity demanded the resources and services of 1.5 planets in 2010 (Mancini et al., 2016; WWF International et al., 2014). In 2009 Rockström et al. proposed a new approach to global sustainability in which they de-fined planetary boundaries within which humanity could operate safely. They identified nine planetary boundaries and proposed quantifications for seven of them, namely climate change, ocean acidification, stratospheric ozone, the biogeochemical nitrogen (N) cycle and phosphorus (P) cycle, global freshwater use, land system change and the rate at which biolog-ical diversity is lost. The two additional planetary boundaries are chemical pollution and at-mospheric aerosol loading. According to Rockström et al. (2009), transgressing one or more of the boundaries may be deleterious due to the risk of crossing thresholds that will trigger non-linear, abrupt environmental changes within continental- to planetary-scale systems. Rockström et al. (2009) estimated that humanity has already transgressed three planetary boundaries for: climate change, the rate of biodiversity loss, and changes to the global nitro-gen cycle. In a more recent study, Steffen et al. (2015) concluded that anthropogenic pertur-bation levels of four of the Earth system processes (climate change, biosphere integrity (earli-er biodiversity loss), biogeochemical flows, and land-system change) already now exceed the proposed planetary boundary. To fight against the current, unacceptable development Europe has implemented waste strat-egies focusing on minimising waste generation and maximising the recovery of waste. In July 1975 the Council Directive on waste (75/442/EEC) addressed the need for the nine member states to support appropriate actions for waste prevention, recycling and processing of waste, the extraction of raw materials and possibly of energy from waste and any other process for the re-use of waste. The first Community Strategy for Waste Management (SEC (89) 934 Fi-nal 1989) was based on this directive and established the hierarchical system for WM, under which waste prevention and minimisation have the highest priority followed by recycling and disposal. In 1996 the Community Strategy for Waste Management was published (COM (96) 399). In the Sixth Environment Action Programme of the European Union (the EU’s ten-year envi-ronmental policy programme, 2002-2012), the European Commission defined the sustaina-ble use of natural resources and management of wastes as one of four priority areas. This led to the development of two thematic strategies on resources and waste management, namely 1) the Thematic strategy on the prevention and recycling of waste (COM (2003) 301 final, COM (2005) 666 final) (Salhofer et al., 2008), and 2) the Thematic strategy on the sustaina-ble use of natural resources (COM (2005) 670 final). The strict five-level priority order for WM was introduced in EU legislation in 2008 with the Waste Framework Directive (WFD, 2008/98/EC), which gives waste prevention and re-use (or preparing for re-use) the highest priority, recycling taking the third place and thermal treatment with energy recovery and disposal representing the least favourable options (Fig. 1).

Introduction

11

Figure 1. The five-level waste hierarchy according to the EU Waste Framework Directive (2008/98/EC).

According to the WFD, when applying the waste hierarchy, member states must take measures to promote the options that deliver the best overall environmental outcome. This may require specific waste streams to depart from the waste hierarchy where this is justified for reasons of technical feasibility, economic viability and environmental protection, among other things. (WFD 2008/98/EC). Recently, van Ewijk and Stegemann (2016) challenged the waste hierarchy by concluding that in its current form the hierarchy is an insufficient foundation for waste and resource policy to achieve absolute reductions in material throughput. Suggested improvements include the adoption of a value-based conception of waste and related collection practices, more strin-gent and targeted policies towards the least desirable options such as landfill, the specifica-tion of WM targets based on dematerialisation ambitions, and the use of the waste hierarchy within a resource productivity-oriented framework. The concept of the Circular Economy highlighted especially by the Ellen McArthur Founda-tion (2012; 2013; 2014) has during recent years strongly emphasised the need for waste pre-vention through business models that do not base their value creation on producing materials but rather on fulfilling the needs of customers by providing services. Such development re-quires system level innovations. However, while developing these, we continuously generate waste that needs sustainable solutions. The circular management of resources and waste has been emphasized both in the bio-economy strategy of the European Commission (2012) and in the EU Circular Economy Package (European Commission, 2017a) as a means to increase resource efficiency. Especially for biomass and bio-based waste streams, cascading utilisation by recovering residues and waste as materials for as many rounds as possible and finally re-covering them as energy, has been raised as way to improve sustainability and extend bio-mass availability in a system (Vis et al., 2016).

Introduction

12

1.2 Life cycle assessment and waste management

1.2.1 Environmental life cycle assessment as a support to waste policy

In order to assess the most sustainable solutions for WM and end-of-life options for prod-ucts, methods are needed which enable the consideration of various impacts and the benefits from reuse, recycling and recovery. According to the WFD, assessment methods based on life cycle thinking (LCT) should be used to justify decisions that are not in accordance with the waste hierarchy (WFD, 2008/98/EC). LCT aims to take into account overall environmental impacts by considering the range of impacts occurring throughout the life time of a product. LCT is applied in a variety of approaches, most strictly, however, in life cycle assessment (LCA). LCA is a standardised tool for quantifying the life-cycle impacts of a product or service by assessing the emissions, resources consumed and pressures on health and the environ-ment that can be attributed to the subject of assessment (ISO, 2006a; 2006b). It takes into account the entire life cycle – from the extraction of natural resources through material pro-cessing, manufacturing, distribution to use, and finally the re-use, recycling, thermal treat-ment with energy recovery and the disposal of the remaining waste at the end-of-life phase. It is important to note that reducing the environmental impact of one life cycle phase may lead to a greater environmental impact in another phase. An apparent benefit of a WM option can therefore be levelled out if not thoroughly evaluated. LCA provides a comprehensive view of the processes and impacts involved in the product system to be assessed and prevents prob-lem-shifting between system compartments. In the thematic strategies developed by the EU on resources and WM (see Chapter 1.1) the life cycle approach was emphasised as an important part of the work for prevention and recy-cling of waste and the sustainable use of resources. With the help and opportunities offered by LCAs, in the recent decades, the basis for EU policy on waste and resource use has changed from focussing on the amounts of waste and resources and shifted towards the po-tential environmental impacts they cause. An LCA based impact assessment has recently been carried out (European Commission, 2014; 2015) to back the European Commission proposal for reviewing recycling and other waste-related legislation in the EU Waste Framework Directive (2008/98/EC), the Landfill Directive (1999/31/EC) and the Packaging and Packaging Waste Directive (94/62/EC). The proposal was first introduced by the European Commission in July 2015 with the aim of help-ing turn Europe into a circular economy, and to boost recycling, secure access to raw materi-als and create jobs and economic growth. This was done by setting ambitious targets and adding key provisions on the instruments required to achieve and to monitor them (Europe-an Commission, 2014). The proposal was included in the Circular Economy Package, which was adopted in December 2015, and is being implemented according to the Circular Econo-my action plan (European Commission, 2017a).

Introduction

13

1.2.2 Life cycle assessment methodology for evaluating waste hierarchy

The LCA methodology for WM studies has been developed to take into consideration the spe-cific issues concerning WM. LCA widens the scope of analysis from the WM itself (waste con-nected processes) to the links with other sectors such as manufacturing, energy production and agriculture (Björklund et al., 2010). Thus, the quantification of benefits from waste recy-cling and recovery becomes possible and relevant. Some of the earliest documentation of LCA methodology for WM was provided by Clift et al. (2000) and Ekvall and Finnveden (2000). Recent LCA guidelines have been developed by the Directorate General Environment and the Joint research Centre (JRC) to support environmentally sound decision-making in WM. These guidelines are tailored to the needs of different target audiences and partly focus on specific waste streams, e.g. biowaste and construction and demolition waste (C&DW) (Euro-pean Commission, 2011; Manfredi et al., 2011). Additionally, Laurent et al. (2014a) have pre-sented recommendations on implementing LCA for solid waste management systems (SWMS) based on their review and findings from over 200 LCA studies performed on these systems. The overall framework for LCA is standardised (ISO, 2006a; 2006b) and well established as a four-step process, which is initiated with a goal and scope definition, followed by a life cycle inventory (LCI), life cycle impact assessment (LCIA) and interpretation (Fig. 2). The assess-ment proceeds iteratively, returning when necessary, from one step to the preceding one/ones to adjust the model.

Figure 2. The framework of life cycle assessment (ISO, 2006a).

Goal and scope definition

The goal definition describes the purpose of the study and the decision process it provides support for. In LCAs for WM, for example, the goal may be related to the determination of whether recycling or thermal treatment with energy recovery is environmentally preferable for a certain waste fraction, or whether one energy recovery technique is preferable over an-

Introduction

14

other or whether one flue gas cleaning technology generates lower environmental impacts from waste incineration than another (Hauschild and Barlaz, 2010). During the scope definition, the object of the LCA, i.e. the exact system to be analysed, is identified and defined in detail. One key aim is to define the functional unit (FU) of the sys-tem. FU provides the reference to which all the system inputs and outputs are related. It gives a quantitative description of the primary function or service that the analysed system pro-vides (European Commission, 2011; ISO, 2006a; Laurent et al., 2014b). The same FU shall be used for all alternative systems compared in the LCA. For the LCA of a WM system, the FU could be defined as, for example, the quantity of waste to be managed during a certain period of time (e.g. tonnes/year), composition of the waste, duration of the WM service or emission limits (Hauschild and Barlaz, 2010). In their review on LCAs for SWMS, Laurent et al. (2014b) showed that most LCA studies used the management or treatment of 1 tonne of waste as the FU. The authors considered this in-sufficient since specification of the type of waste considered should be included in order to give a precise picture of the function of the system (Laurent et al., 2014b). Composition of waste may vary in different regions and locations due to e.g. housing and consumption pat-terns, different practices and efficiencies in source separation of waste, differences in collec-tion systems, availability of composting and several other aspects. The waste composition affects the performance of WM, hence the composition and the practical implementation of the SWMS studied should be defined and described in the scope definition (Laurent et al., 2014b). Also other, geographically varying factors affect the environmental performance of WM pro-cesses and systems. These factors include the natural conditions, such as climate that can greatly affect the possibilities of e.g. arranging various collection systems. Institutional fac-tors such as legislation, e.g. on the quality of waste-derived materials, provide requirements for the processing of waste and this way affect the performance of WM. Economic factors such as existence and location of potential markets for waste or recovered products from waste will determine logistical requirements of the system. Technical factors such as availa-bility of waste treatment facilities and infrastructure for waste transports are crucial for WM strategies. All these issues should be considered when defining the goal and scope of the LCA study on WM. (European Commission, 2011). Boundaries of the studied product or service system need to be defined on the basis of the goals of the assessment. An overall system of consumption, production and WM includes numerous elements, different products, processes and flows. Product specific LCAs include the whole life cycle of a product from resources extraction, production to consumption and end-of-life of the product. From the point of view of WM both the upstream (preceding the WM) and downstream (use of waste-derived materials or products) processes can be includ-ed in these assessments. The goal of a product LCA can be to identify the product life cycle phases that most influence the overall product performance, or whether recycling or thermal treatment with energy recovery of the final product can improve the overall performance, among others.

Introduction

15

When focusing on a WM system, i.e. the end-of-life phase, the starting point for modelling is usually waste generation, the point where products are discarded as waste. Waste comes into the system with zero-burden, assuming that the same kind of waste will be generated no mat-ter how the WM is organised. This assumption is in compliance with the LCA definition which allows users to disregard parts that are identical between compared systems. However, if one of the systems would include a waste minimisation strategy that affects waste genera-tion, the production and use phase of the products would need to be included in the model as well, making the model much more complex. (Björklund et al., 2010; Ekvall et al., 2007). The foreground of a WM system covers the actual WM processes, i.e. collection, separation, processing, recovery and treatment, whereas the background system covers processes outside WM (Clift et al., 2000). WM systems most commonly provide several functions, such as managing waste and producing recycled materials. Recycled materials can be considered by-products of waste treatment, and a share of the overall impacts of the system needs to be al-located to them. However, the ISO standard recommends avoiding allocation whenever pos-sible, e.g., by expanding the product system to include additional functions related to the by-products (ISO, 2006b). Thus, all interventions can be attributed to the product system and credits can be given for the production of by-products, assuming that their production re-places other ways of producing similar products, such as using virgin materials. Nevertheless, it is not always easy to determine what product or production method the recycled products replace and there may even be several correct choices for a certain case (Björklund et al., 2010). The review by Laurent et al. (2014b) discovered that 75% of the LCA studies on SWMSs applied system expansion in order to credit waste material or energy recovery. How-ever, the reporting of the type of material or energy substituted and the substitution rate used was found defective (Laurent et al., 2014b). LCA modelling of a WM system usually ends at the point of substitution, hence downstream processes are not included in the assessment (Björklund et al., 2010; Clift et al., 2000; European Commission, 2011). The change-oriented, or consequential LCI modelling aims at identifying the consequences of a decision in the foreground system on other processes and systems of the economy and builds the to-be-analysed system around these consequences. These consequences include processes that are assumed to be affected due to a specific decision. In contrast, attributional modelling depicts the system as it can be observed, linking single processes to the flow of matter, energy, and services (i.e. the existing supply-chain). (European Commission, 2010a). WM modelling sometimes combines these two types of modelling. Life cycle inventory (LCI) The inventory phase focuses on data collection and system modelling. Inventory analysis is the most laborious and important phase of LCA. The compilation of inventory data relates the inputs and outputs of different processes to the FU of the study (European Commission, 2011; Hauschild and Barlaz, 2010). Two types of inventory data are usually required for an LCA. Process specific primary data on the input materials, energy, water, chemicals, waste, wastewater and gaseous emissions are needed for the foreground system, whereas non-specific, generic, secondary data is sufficient for the background system (Clift et al., 2000).

Introduction

16

In some cases conclusions can already be drawn on the basis of the results of the inventory phase (Seppälä, 2003). However, an LCIA (Fig. 2) is often needed for interpreting the results of the inventory. The consideration of environmental effects as a consequence of environ-mental interventions provides additional information which is not covered by the inventory step (Seppälä, 2003). Life cycle impact assessment (LCIA) An LCIA is typically divided into five phases: 1) selection of impact categories, category indi-cators and characterisation models, 2) classification of LCI results, 3) calculation of category indicator results, i.e. characterisation, 4) normalisation and 5) weighting. The first three phases are mandatory and two last phases are optional according to the ISO standard (ISO, 2006a; 2006b). The LCIA is still under development and apart from the global impact categories, global warming and stratospheric ozone depletion, no consensus has been reached on how to model the impacts (Hauschild and Barlaz, 2010). Land use impacts and their assessment is one of the developing topics. Baan et al. (2012) developed a methodology for assessing the impacts on biodiversity, focusing on occupational impacts, quantified as biodiversity damage poten-tial (BDP). According to Baan et al. (2012) the reported characterisation factors for BDP can be used to approximate biodiversity impacts in LCA studies not focusing on land manage-ment practices. For studies focusing on land management practices, more detailed and site-dependent assessments are required (Baan et al., 2012). Land use impacts may not be rele-vant in all assessments of an SWMS, but could be important when focusing on biomass use (Laurent et al., 2014b). In recent years the need for and definition of reference situations for the assessment of land use impacts has been discussed by and Cao et al. (2017) and Soima-kallio et al. (2015), among others. Even on the global warming impacts various issues are still under discussion, such as the inclusion of methane oxidation impacts into the assessment, which has been challenged by Muñoz and Schmidt (2016). Two approaches, midpoint and endpoint modelling, exist in LCIA modelling. In the most common midpoint approach impacts are modelled at a midpoint level in the environmental mechanism between emissions and damage, but as close as possible to the areas of protection in the causality web (Goedkoop and Spriensma, 2001; Hauschild and Barlaz, 2010; Seppälä, 1999; 2003). Commonly assessed impact categories include climate change, ozone depletion, human toxicity, particulate matter, ionizing radiation, photochemical ozone formation, acidi-fication, eutrophication, ecotoxicity, land use and resource depletion (European Commission, 2010a). In endpoint modeling the impacts are modelled all the way down to the areas of pro-tection, such as human health, ecosystem quality, and resource depletion, using the best available environmental models (Hauschild and Barlaz, 2010; Steen,1999a; 1999b). Most recent LCIA methods, such as the ReCiPe Mid/Endpoint methodology (Goedkoop et al., 2009; Wegener Sleeswijk et al., 2008) combine the two approaches and give the user options to select which one to use.

Introduction

17

In the classification phase of midpoint modelling, the emissions and resources derived from an LCI (i.e. interventions) are assigned to impact categories according to their ability to con-tribute to the environmental problem each category describes. In the characterisation phase interventions are converted into impact category indicators using characterisation factors that reflect the pressure that one unit of emission or resource consumed generates in the con-text of each impact category. The result of this is the characterised impact profile of the prod-uct, i.e. its LCIA result. (European Commission, 2010b). Characterisation is considered the last mandatory step of the LCIA due to its scientific char-acter, but an LCIA can be continued with one or two optional steps, normalisation and weighting. Normalised LCIA results are generated to translate the impact category scores into more understandable and presentable figures and to place them in an adequate environmen-tal context (e.g. Laurent et al., 2011; Pizzol et al., 2017; Wegener Sleeswijk et al., 2008). Nor-malisation is performed by dividing the LCIA results by the normalisation basis separately for each impact category. Regionally defined reference values are most commonly used as the basis for normalisation (Laurent et al., 2011; Lautier et al., 2010; Pizzol et al., 2017; Wegener Sleeswijk et al., 2008). However, the normalisation basis can be chosen based of various di-mensions, such as system bases, spatial or temporal scaling or even regional consumption (Breedveld et al., 1999). The normalisation reference needs to be complete in respect to the analysed system (Pizzol et al., 2017). Biased results can be obtained if a substance which is part of the LCI of a product system and a main contributor to the characterised results of a given impact is not part of the LCI used to calculate the normalisation reference. This would mean the normalisation reference would be underestimated and the normalised results would be overestimated (Heijungs et al., 2007; Pizzol et al., 2017). Normalised LCIA results of the different impact categories show to which impacts the ana-lysed system mainly contributes. The relevance of the impact categories relative to each other cannot be judged on the basis of these results. To directly compare or sum up results across impact categories, weighting is needed. In weighting, the normalised LCIA results for the different impact categories are each multiplied with a weighting factor expressing the relative importance of the category. Weighting factors can be generated with several methods, which Huppes and van Oers (2011) classify into non-monetised methods, such as panel and dis-tance to target method, and monetised methods, such as the damage prevention cost method. Pizzol et al. (2017) provide some recommendations on the generation of weighting factors, such as suggesting that panels comprising affected stakeholders should be preferred instead of expert panels unless the relevant stakeholders delegate the task to other experts or repre-sentatives. For monetary valuation methods, Pizzol et al. (2017) recommend using observed preferences (market prices) whenever possible. The normalised and weighted LCIA results can subsequently be added up across all impact categories or areas of protection to create one single score for the environmental impact of a product or scenario (European Commission, 2010b). Interpretation The interpretation of LCA results is performed against the goals of the assessment and within the restrictions arising from the scope of the study and the assumptions made during the

Introduction

18

modelling process. The interpretation includes identifying significant issues that have the largest affect or relevance to the overall environmental impacts. Findings in the impact as-sessment and sensitivity and contribution analysis, which are performed as part of the inter-pretation, help to identify the key processes, elementary flows, impact categories, modelling choices and assumptions of the system. This way the reliability of the final results can be as-sessed. (European Commission, 2010a; European Commission, 2011; Hauschild and Barlaz, 2010). Due to the aggregation of emissions over time and space in the inventory phase, the impacts calculated are the sum of contributions from emissions that may be released over a period of several years in different locations. Thus, it is difficult to interpret these impacts in terms of effects on the environment and instead they should be considered as environmental sustain-ability indicators that can be used for process optimisation (Hauschild and Barlaz, 2010). In WM LCAs the assumptions made concerning processes which have been avoided often determine the overall results. In the interpretation phase it is therefore important to consider the effects of different alternatives on the processes avoided. The final phase of the interpretation is drawing of conclusions and identification of limita-tions of the LCA, and development of recommendations by integrating the outcome of the elements of the interpretation phase, and summarising the main findings from the earlier phases of the LCA. This is done in accordance with the goal definition and the intended ap-plications of the results. Typically, the conclusions state whether the questions that were posed in the formulation of the goal definition can be answered by the LCA, i.e. whether sig-nificant differences exist between alternatives and what role the issues of the sensitivity anal-ysis play regarding such differences. The limitations of the study within the given goal and scope of the LCA study must be reported. These can include items such as the incompleteness of elementary flows with relevance to impact categories, the limited time-representativeness, or pre-selection of climate change impacts only for carbon footprint studies. (European Commission, 2010a; European Commission, 2011).

1.2.3 Life cycle assessment studies on waste management

LCAs have been applied for assessments of WM most commonly to do with municipal solid waste (MSW) management since from the early 1990s (Björklund et al., 2010; Laurent et al., 2014a). The different levels of waste hierarchy have been compared in numerous studies on various WM systems and parts of it, as well as on different waste fractions and recovery tech-nologies. However, the first priority, waste prevention, has seldom been assessed in an LCA. Yet some examples can be mentioned, such as the study by Cleary (2014), who assessed the life cycle impacts of five types of waste prevention measures, and the study by Arushanyan et al. (2017), who included waste prevention in their overall model of the Swedish WM system. LCAs of products with different end-of-life options have been performed to challenge the second priority level of the waste hierarchy, reuse (or preparation for reuse), and in product specific studies such as the UK Environment Agency (2008) study on nappies, Lighthart and Ansems (2007) study on drinking systems, Raugei et al.’s (2009) on drums for chemical transport, as well as Sørensen and Wenzel’s (2014) study on bedpans and Unger and Landis’

Introduction

19

(2014) on dental burs, which all compared reuse versus disposal for a product requiring washing in the reuse option. Numerous LCA studies have compared the three lowest levels of the waste hierarchy, namely recycling, thermal treatment with energy recovery and landfilling in different WM systems. Several meta-analyses have been performed to find out whether the results can be general-ised for decision support. These meta-analyses include reviews by Bernstad and la Cour Jan-sen (2012) on food waste LCAs, Björklund and Finnveden (2005) on LCAs for various frac-tions of MSW, Cleary (2009) on mixed waste LCAs, Laurent et al. (2014a; 2014b) on all LCA studies on plastic, paper, organic and mixed waste fractions of MSW, Lazarevic et al. (2010) on plastic waste, Michaud et al. (2010) on LCAs for various fractions of MSW, Morris et al. (2013) on source separated organics and Villanueva and Wenzel (2007) on paper and card-board waste. In addition, Bassi et al. (2017) compared the environmental performance of household WM in seven European countries. Most LCAs on SWMSs have been done in Eu-rope, where the number of studies grew considerably after the acceptance of the EU Waste Framework Directive in 2008 (Laurent et al., 2014a). However, in recent years LCAs on WM issues in countries such as China have been published increasingly. These include studies by Havukainen et al. (2017) and Liu et al. (2017) on MSW management, among others. Looking more specifically at paper WM, Villanueva and Wenzel (2007) conducted a review of nine paper and cardboard LCA studies with altogether 73 scenarios with different manage-ment options for waste paper. They found the outcome of individual LCA studies significantly depended on fifteen key assumptions including assumptions made regarding the use of wood saved by recycling, the efficiency of energy recovery technology modelled and the type of en-ergy production substituted with waste based energy (Villanueva and Wenzel, 2007). Merrild et al. (2012; 2008) further assessed the impacts of different choices such as system boundary choices, and the significance of technical data when modelling the environmental impact of recycling and incineration of waste paper. Kim and Song (2014) applied the concept of tem-porary biomass carbon storage in their comparison of wood waste recycling in particle boards to energy recovery in combined heat and power production (CHP) plants. Wood waste is currently mostly used for energy production. The environmental benefits of the cascading utilisation of wood compared to the primary use of wood and non-cascading uses of secondary wood have been assessed by Höglmeier et al. (2014; 2015). Additionally, the contribution of wood polymer composites (WPC) to the cascading utilisation of wood has been assessed by Teuber et al. (2016). The increased use of sawmill residues and other wood waste in WPCs instead of using the waste for energy purposes could increase biomass cascad-ing and utilisation (Vis et al., 2016). In recent years, concern has grown about the sustainable management of the increasing amounts of C&DW, and consequently the number of published LCA studies on C&DW has grown (Bovea and Powell, 2016; Butera et al., 2015; Diyamandoglu and Fortuna, 2015; Hoss-ain et al., 2017; Martinez et al., 2013; Mercante et al., 2012; Ortiz et al., 2010). Hossain et al. (2017) applied an LCA to evaluate the performance of C&DW management systems which were equipped with different sorting technologies. Butera et al. (2015) studied mineral, source separated C&DW which was either utilised in road construction as a substitute for

Introduction

20

natural aggregates, or landfilled. Bovea and Powell (2016) performed a review of studies ap-plying LCA to C&DW management systems, but also included studies on the performance of recycling plants for different C&DW fractions and comparisons of natural materials versus materials manufactured from recycled C&DW. However, most studies analysed the whole life cycle of buildings or construction elements, including the end-of-life stage, and not the C&DW management as such (Bovea and Powell, 2016). In Finland LCAs on WM have been carried out and increasingly published over the past 10 years (e.g., FCG, 2010; Kiviranta and Tanskanen, 2009; Manninen et al., 2016; Moliis et al., 2012; Myllymaa et al., 2008; Niskanen, 2012; Sundström et al., 2014). The increase is caused due to changes in the operational environment of WM. Before 2007, only one incineration plant for MSW existed in Finland, and most of the mixed MSW was landfilled. Source separa-tion and material recovery options were available for biowaste, paper, cardboard, metal and glass. In some regions cardboard, paper and plastics were collected as energy waste to be co-combusted as REF (recovered fuel) together with other fuels. From 2007 onwards, eight new waste-to-energy power plants for mixed MSW have started operation and this has helped in diverting waste away from landfills. However, at the same time, the recycling rate for MSW has remained low, being now at the level of 41% (Laaksonen et al., 2017). The introduction of increased capacity and further possibilities to incinerate mixed MSW have led to questions concerning the benefits of material recovery of combustible fractions of MSW and have hence increased the need for LCA studies. Several LCAs have focused on biowaste, and compared the environmental impacts of some or all of the following alternatives: composting, anaerobic digestion, thermal treatment with energy recovery and biofuel production (FCG, 2010; Manninen et al., 2016; Myllymaa et al., 2008; Sundström et al., 2014; Virtavuori, 2009). The majority of the Finnish LCAs on bio-waste have only assessed climate change impacts (FCG, 2010; Virtavuori, 2009). However, Sundström et al. (2014) also addressed eutrophication impacts. The separate collection and recovery of different packaging material waste flows has been assessed in Finland primarily from the point of view of the potential and environmental im-pacts of expanding the separate collection to small properties and sparsely populated areas (Kiviranta and Tanskanen, 2009; Moliis et al., 2012; Niskanen, 2017). Finland has long traditions in recycling paper waste to be used as a resource for paper pro-duction. In contrast, waste wood has been considered preliminarily a source of energy. This arises from the fact that virgin wood resources in the country are vast and the need for recov-ering waste wood as a material has not been confronted in practice. In recent years recycling options for wood waste have been of interest due to the wide use of wood packaging and the high share of wood in, for example, C&DW in Finland. Manninen et al. (2015) compared the environmental impact of energy recovery of wood waste and the current applications for re-cycled wood waste, namely composites and particle boards. The use of wood residues or wastes in WPCs has also been studied by Judl et al. (2016) and Turku et al. (2017), for exam-ple.

Introduction

21

In recent years scientific discussion concerning the role of forests and forestry on climate change has been vivid and is also connected to the management of wood-based products and waste. The issue of wood utilisation and the environmental benefits from it is complex, as was shown by Soimakallio et al. (2016), who recently analysed climate change mitigation challenges for the Finnish wood utilisation. However, thermal treatment with energy recov-ery was considered the end-of-life option for paper instead of recycling, based on the fact that paper fibres can only be recycled a limited number of times (Soimakallio et al., 2016).

1.3 Broadening life cycle assessment with cost and material assessments

An LCA can be used for decision-support both in industry, and by municipalities, policymak-ers and regulatory authorities. The information provided by carrying out an LCA must often be combined with other types of information, environmental and economic, for example, to reach a fully informed decision (Hauschild and Barlaz, 2010). To strengthen the LCA as a tool and eventually increase its usefulness in supporting sustainable decisions, it has been argued that there is a need to expand the ISO LCA framework through integration and linking it to other concepts and methods (Ekvall et al., 2007; Jeswani et al., 2010; Kloepffer, 2008). Finnveden et al. (2007) performed an overview of methods and approaches that could be used for supporting WM decisions at different levels in society. The tools included approach-es for both environmental and economic assessments such as the LCA, as well as life cycle costing (LCC), environmental impact assessments (EIA), strategic environmental assess-ments (SEA), cost-benefit analyses (CBA), risk assessments, material flow accounting or analyses (MFA), substance flow analyses (SFA) and environmental management systems (EMS). In addition, Jeswani et al. (2010) performed an overview of some approaches and methods available for deepening and broadening the LCA to also cover other sustainability issues than environmental sustainability. The methods are classified according to their focus or level of assessment (micro, meso or macro) and according to the sustainability dimensions they are able to cover (Table 1). Jeswani et al. (2010) concluded that there is no ‘‘one size fits all’’ solu-tion to integrating different life cycle related concepts for better sustainability assessments. Instead options for broadening the LCA depend on the field of application and the users, as well as their requirements and the goal and scope of an investigation.

Introduction

22

Table 1. Methods to assess environmental, economic and social sustainability (Jeswani et al., 2010).

Assessment method Focus/level Sustainability dimension Procedural method (assessment framework) Environmental Impact Assessment (EIA)

Project (micro) Environmental and social

Strategic Environmental Assessment (SEA)

Policy (meso/macro) Environmental and social

Sustainability Assessment (SA) Policy (macro) Environmental, economic and social Multi-Criteria Decision Analysis (MCDA)

Policy/project (micro/meso/macro) Decision-support tool which can in-clude environmental, economic and social dimensions

Analytical method Material Flow Analysis (MFA) Policy, plan (macro) Environment (natural resources) Substance Flow Analysis (SFA) Specific substance (macro) Environment (natural resources) Energy/Exergy Analysis (EA) Process, Product/service (micro) Environment (natural resources) Environmental Extended Input Output Analysis (EIOA)/Hybrid LCA

Policy, Product/service (meso/macro) Environment

Risk Analysis (RA/ERA/HERA) Chemicals/Projects (micro) Environmental and health impacts Life Cycle Costing (LCC) Product/Service (micro) Economics Cost-Benefit Analysis (CBA) Policy/Project (micro/meso/macro) Economics (includes cost of environ-

mental and social impacts) Eco-Efficiency (EE) Analysis Product/Service (micro) Integration of environmental and

economic Social Life Cycle Assessment (SLCA) Product (micro) Social

LCC has gained support for its use as a complementary tool to LCA for sustainability assess-ments for example by Kloepffer (2008), who proposed to complement LCAs with LCC and a social life cycle assessment (SLCA) to generate a life cycle sustainability assessment (LCSA). The system boundaries of the three assessments need to be consistent and the best solution would be the use of one identical LCI for all three components. Economic issues often play a decisive role in decision-making, hence it is important to include economic aspects in as-sessments along with environmental aspects. In addition, a social impact assessment could be considered important, at least in less developed countries. According to Kloepffer (2008) it seems also that the societal LCI would be much more demanding in regard to regional reso-lution than LCAs or LCC. For environmental life cycle costing (ELCC), a code of practice has been developed (Swarr, 2011). ELCC summarises all costs associated with the life cycle of a product that are directly covered by one or more of the actors in the life cycle (e.g., supplier, producer, user/consumer and those involved at the end-of-life). These costs must relate to real money flows (Hunkeler et al., 2008). ELCC is performed on a basis analogous to LCA, including the definition of an FU and similar system boundaries in both LCAs and ELCC. The costs are complemented with an LCA, using equivalent system boundaries and FUs. (Hunkeler et al., 2008). Social life cycle costing (SLCC) focuses on the costs associated with the life cycle of a product or a service defined from the viewpoint of the whole society. Furthermore, the stakeholders of SLCC include governments and other public bodies not directly concerned with the product system. The most conventional costs are direct costs, such as the costs of investments, labour and energy. SLCC also includes all external costs associated with the environmental impacts throughout the product’s or service’s life cycle. The costs do not include subsidies or taxes, because they are internal to the system and are levelled out in the overall results. (Hunkeler et al., 2008; Paper II).

Introduction

23

Developing a consistent data set for LCC can be challenging because the data comes from many different sources and may be business sensitive (Swarr et al., 2011). Recently, Mar-tinez-Sanchez et al. (2015) developed a consistent cost model for LCC including cost items for all key technologies within modern WM systems. Cost items were classified as: (1) budget costs, (2) transfers and (3) externality costs. The cost model allows calculation of convention-al LCC, ELCC and SLCC. (Martinez-Sanchez et al., 2015). A number of studies have assessed both environmental and economic impacts of WM by combining LCAs with LCC, ELCC (Carlsson Reich, 2005), SLCC, CBA (Ahamed et al., 2016) or other economic assessments such as input-output analyses (Ferrão et al., 2014). However, Martinez-Sanchez et al. (2015) found very few examples in the literature of really combining economic and environmental aspects in assessing WM systems. In Eriksson et al. (2005) different combinations of incineration, materials recycling of sepa-rated plastic and cardboard containers, and biological treatment (anaerobic digestion and composting) of biodegradable waste, were studied and compared to landfilling. The evalua-tion covered the use of energy resources, the environmental impacts and the financial and environmental costs. Cleary (2009) performed a review of LCA studies on mixed waste treatment and found that eight of the twenty LCAs incorporated financial LCCs focusing on waste treatment activities. Five of the twenty studies undertook an environmental LCC, attributing economic values to the environmental impacts of each waste treatment scenario (Cleary, 2009). Packaging waste has been in focus of several studies combining environmental and economic assessments. In Finland, Moliis et al. (2012) studied the environmental impacts and costs of a separate collection system for packaging waste in sparsely populated Northern Finland. Ferreira et al. (2014a; 2014b; 2017a; 2017b) assessed the costs and benefits of recycling pack-aging waste from the perspective of local authorities in Portugal, Belgium and Italy. Pires et al. (2017) assessed the environmental impacts and costs to the WM company of packaging waste collection systems in Portugal. Larsen et al. (2010) assessed the environmental and economic impacts of an MSW management system with alternative collection systems for recyclables (paper, glass, metal and plastic packaging) in Aarhus. The economic assessment focused on the WM expenses for the municipality and included the costs of the collection equipment, as well as actual collection and treatment costs. (Larsen et al., 2010). In recent years the use of material flow analyses (MFA) and substance flow analyses (SFA) as supporting tools for LCA have also emerged (Arena and Di Gregorio, 2014; Rochat et al., 2013; Sevigné-Itoiz et al., 2014; Wäger et al., 2011). MFA is a systematic assessment of the flows and stocks of materials (goods and substances) within a system defined in space and time. It connects the sources, the pathways and the intermediate and final sinks of a material. MFA is a descriptive approach that provides snapshots of parts of the physical economy (Reuter et al., 2005). SFA is a type of MFA that focuses on a specific substance within a sys-tem. MFA and SFA can be applied in different types of studies in different ways. For example, the internal flows, stocks and losses of a company can be calculated with MFA. Examples of

Introduction

24

applying SFA to WM systems can be found in Döberl et al. (2002), and examples of SFA ap-plied to specific treatment methods can be found in Morf et al. (2005). The combination of MFA and SFA with an environmental assessment method such as LCA appears to be an at-tractive tool-box for comparing alternative WM technologies and scenarios, and supporting WM decisions on both strategic and operating levels (Arena and Di Gregorio, 2014). These types of MFA and SFA are the focus in this study. MFAs and SFAs have been combined with LCAs for determining the cause-effect chains made up of physical flows in order to properly assess the market links between raw materials and waste (Sevigne-Itoiz et al., 2015) or to transparently report waste flows, as well as their connections and the impacts of process changes on them in a WM system (Tunesi et al., 2016). Arena and Di Gregorio (2014) and Tunesi et al. (2016) integrated SFA and MFA com-bined with LCA results for planning an MSW system. Turner et al. (2016) combined an MFA and LCA to assess the performance of an MSW man-agement system of a local authority and to compare it to alternative systems to assess the potential effectiveness of different waste policy measures. The approach focused on green-house gas (GHG) emissions, and the most efficient ways to reduce them. Padeyanda et al. (2016) applied an MFA and LCA to assess the potential environmental impacts of food WM practices in different recycling facilities for Daejeon Metropolitan City in Korea in order to find sustainable solutions for food WM. Sevigne-Itoiz et al. (2015) applied a consequential LCA to GHG emissions and combined it with an MFA to assess the paper and cardboard re-cycling system in Spain.

1.4 Aims, focus and structure of the thesis

This thesis examines various Finnish waste flows and WM systems from the viewpoint of their environmental and in some cases also economic impacts. The thesis comprises of case studies which were chosen based on the fact that they all consid-ered wastes or products from the forest sector. Wood-based products and wood waste are crucial for Finland due to the vast forest resources and high economic importance of the for-est industry. Issues connected to the forest sector and wood-based products are actively dis-cussed and developed due to the need in Europe to move towards a sustainable bioeconomy to ensure sufficient supplies of raw materials, energy and industrial products under condi-tions of decreasing fossil carbon resources. The management of waste flows originating from forestry products, such as waste paper, cardboard and wood waste, and different options for them are therefore of interest. One of the case studies focuses on construction and demolition waste (C&DW), one of the priority waste flows in the EU Commissions Circular Economy package (European Commis-sion, 2017a). At the national, but also the European level, C&DW is a waste flow of major concern, due to the large and growing volumes and low recycling rates. Different levels of waste hierarchy are assessed and compared in the case studies presented in the thesis papers. The case studies were analysed using life cycle based approaches.

Introduction

25

The thesis aims to answer the following research questions (RQ) by focusing on specific case studies on newspaper, C&DW and delivery packaging: RQ1: Do the results of the life cycle based approaches that generate information on the envi-ronmental and/or economic impacts support the waste hierarchy order of WM options in the Finnish context? RQ2: Does the combination of life cycle based approaches provide added value for decision-making for Finnish WM systems? RQ3: How can normalisation be used for improving the understanding of LCIA results? The analysis of the waste hierarchy levels proceeds from the lowest hierarchy levels, i.e., landfilling, thermal treatment with energy recovery and recycling, towards the second highest level, i.e., reuse (or preparing for reuse) (Table 2, Fig. 1). The focus is on the WM phase, i.e. the options available when products or materials are about to be discarded. For this reason, product design is beyond the scope of this study and the first level of the waste hierarchy, i.e., waste prevention, is not addressed. The materials and methods used in this thesis are presented in Chapter 2. In Chapter 3 the results from the papers constituting this thesis are presented and discussed from the perspec-tive of the research questions defined above. Finally, conclusions are drawn and recommen-dations for future research are given in Chapter 4.

26

2. Materials and methods

2.1 Case studies

The papers comprising this thesis are based on results of the following four case studies with varying focus and objectives and using varying life cycle based approaches (Table 2): 1) WM options for discarded newspaper in the Helsinki Metropolitan Area (Papers I and II) 2) Communication of LCIA results concerning newspapers and digital books for non-LCA practitioners (Paper III) 3) Current level and future insights into the environmental impacts of C&DW management system (Paper IV) 4) Comparison of a reusable and recyclable bread transport packaging (Paper V) The WM options included in the case studies were landfilling, thermal treatment with energy recovery and recycling (Papers I, II, III, IV). Paper V evaluated options at the higher levels of waste hierarchy, namely recycling and reuse (Table 2). Methodologically the study began from the life cycle inventory (LCI)(Paper I), then broad-ened LCI to include a life cycle impact assessment (LCIA)(Papers II – V) and combined LCIA with social or environmental life cycle costing (SLCC, ELCC)(Papers II and IV). Furthermore, an LCA and ELCC assessment was complemented by an MFA (Paper IV). An LCA was used for the environmental impact assessment in all papers (Table 2). Process specific primary data were used for the inventory of the foreground system when available. Generic data from databases were used for the background system (Table 2, Fig. 3).

Materials and methods

27

Table 2. Overview of the papers constituting the thesis. LCI = life cycle inventory, LCIA = life cycle impact assess-ment, SLCC = social life cycle costing, LCA = life cycle assessment, ELCC = environmental life cycle costing, MFA = material flow analysis, NP = newspaper, WM = waste management, CCB = corrugated cardboard).

Paper num-ber

Research ques-tions (RQ) dealt with1)

WM options included

Waste hierarchy levels evaluated

Approaches used

Description of data2)

I RQ1 Landfilling, ener-gy recovery, recycling

5, 4, 3 LCI Primary data: NP manufac-turing, collection and sorting of discarded NP. Generic data: forestry, waste recovery and treat-ment

II RQ1, RQ2 Landfilling, ener-gy recovery, recycling

5, 4, 3 LCIA, SLCC Primary data: NP manufac-turing, collection and sorting of discarded NP. Generic data: forestry, waste recovery and treat-ment

III RQ3 Landfilling, energy recovery, recycling

5, 3 LCA Primary data: Printing. Generic data: Pulp and paper production, WM

IV RQ1, RQ2 Landfilling, ener-gy recovery, recycling

5, 4, 3 LCA, ELCC, MFA

Generic data: the whole system

V RQ1 Recycling, reuse 3, 2 LCA Primary data: CCB box manufacturing and recy-cling, delivery system. Generic data: plastic crate manufacturing, recycling and energy recovery

1) Research questions: RQ1. Do the results of the life cycle based approaches that generate information on the environ-mental and/or economic impacts support the waste hierarchy order of WM options in the Finnish context? RQ2. Does the combination of life cycle based approaches provide added value for decision-making for Finnish WM systems? RQ3. How can normalisation be used for improving the understanding of LCIA results?

2) Primary data = measured/reported data from the processes involved in the system, generic data = data from literature and generic databases.

Materials and methods

28

Figure 3. General overview of elements identified in a combined production, consumption and waste manage-ment system of several products. The case study systems were defined to include parts of the overall system case specifically.

2.2 System descriptions and modelling assumptions

The papers included in this study considered products or fractions of MSW and C&DW streams (Table 3). The studies focused on one or several materials within these waste streams. Hazardous wastes were excluded. The objectives of the studies varied (Table 3), from very general aims such as generating in-formation on the impacts of different WM options for the development of waste policy on paper waste (Papers I and II) to very specific goals such as comparing the environmental im-pacts of a delivery system using either reusable high-density polyethylene (HDPE) plastic crates or recyclable corrugated cardboard (CCB) boxes for product transportation (Paper V). The FUs for the assessments were defined to meet the objectives of the studies (Table 3). The most significant assumptions made for the assessments are related to the calculation of the benefits of material or energy recovery through avoided processes, i.e. processes that could be compensated for by waste recovery (Table 3). These processes were determined case specifically, based on aspects such as the level of assessment (regional, national), or the char-acteristics of the waste based material or fuel, knowledge about current practices and fore-seeable alternatives for recovery. Different methods for calculating the benefits of recycling and their consequences were tested using open-loop allocation (ISO, 2000), monetary alloca-tion and the system expansion approach (Paper V). The results did not differ significantly from each other (less than 1%). Hence, the system expansion approach applying avoided pro-cesses recommended by ISO standard (ISO, 2006a) was chosen.

Materials and methods

29

Table 3. Basic features and modelling assumptions for the systems studied. CCB = corrugated cardboard, C&DW = construction and demolition waste, FU = functional unit, LCIA = life cycle impact assessment, MSW = municipal solid waste, SLCC = social life cycle costing, SRF = solid recovered fuel, WM = waste management.

Paper Waste flows / products studied and FU used

Objectives Scenarios Assumptions on processes avoided by material or energy recovery of waste

I Discarded newspapers collected separately or within mixed MSW; FU: One tonne of newspa-pers delivered to consum-ers (76% to separate collection, 21% discarded into mixed waste, 3% to small-scale recycling)

To provide information based on inventory re-sults on the emissions of different paper WM op-tions.

5 scenarios includ-ing closed-loop recycling1) or ener-gy recovery (co-combustion or incineration) for newspaper, and landfilling or ener-gy recovery for mixed waste; Additionally 4 different rates for separate collection (0, 45, 82, 97%)

Material recovery: use of virgin wood in newsprint manufactur-ing Energy recovery: the average Finnish energy production + use of coal and natural gas as fuels

II As for Paper I

To provide information based on LCIA and SLCC on different WM options for the development of waste policy on paper waste.

As for Paper I As for Paper I

III Newspaper, digital photo-book FU: One tonne of the prod-uct

To assess the differences between Finnish and European reference val-ues for normalisation, and the impacts these have on the interpretation of normalised results.

None Material recovery of newspa-per: use of virgin wood in newsprint manufacturing Energy recovery: average Finn-ish energy production No avoided processes for the photobooks

IV C&DW2); FU: One tonne of C&DW generated

To provide a general view of the environmental and economic impacts of C&DW management and the key issues influencing these. To assess the Finnish C&DW management against the EU recycling target 70% by 2020.

3 scenarios includ-ing recycling of metals and miner-als, and energy recovery of wood and SRF3) produced from miscellaneous fraction

Material recovery of metals: use of virgin ore in iron and aluminium production; Material recovery of concrete and minerals: use of gravel; Energy recovery of wood: heavy fuel oil in a heat boiler; Energy recovery of SRF: natural gas in a co-combustion plant

V Plastic crate, CCB box; FU: Delivery of 8 loaves of bread in one crate/box (One delivery requires 1/700 of a plastic crate or 1 CCB box)

To compare the environ-mental impacts of a deliv-ery system using either reusable plastic crates or recyclable CCB boxes for product transportation.

2 scenarios with different delivery packaging includ-ing recycling (20%) and energy recov-ery (80%) for the plastic crate, and recycling for the CCB box

Material recovery of plastic crate: production of impreg-nated wood; Material recovery of CCB box: use of virgin fibres (fluting) for core-board manufacturing; Energy recovery of plastic crate: average separate heat production

1) Closed-loop recycling of separately collected newspapers for newsprint manufacturing 2) C&DW constituted of five main fractions: metals, concrete & mineral, wood, miscellaneous and an unsorted mixed frac-tion 3) SRF consisted of 80% bio-based materials (e.g. cardboard and wood) and 20% plastics.

Materials and methods

30

Figure 4. General flow chart of a product system for assessing end-of-life options for product life cycle perfor-mance with a special focus on waste management options. Applied in Papers I, II, III and V with modifications.

The system boundaries were defined for an LCA case-specifically to include only the process-es and flows relevant for the assessment (Fig. 4 and 5). Paper I evaluated emissions and Papers II, III and V evaluated the impacts of various end-of-life options for a product life cycle performance, where the processes of the whole product life cycle were included (Fig. 4). Two types of approaches were distinguished: 1) one product with one or several end-of-life options (Papers I, II and III), and 2) several products performing the same function, but having different end-of-life options (Paper V). The study on C&DW in Paper IV represents an assessment of a WM system (Fig. 5). The common Finnish C&DW management system began with the waste generated and sorted at construction, demolition and renovation sites and ended at the point where outputs such as material or waste based fuel were recovered into production. The basic flows and description of the system produced by implementing the MFA was expanded for an LCA and ELCC by adding flows (mainly fuels and energy) needed to maintain the system and processes that could be avoided by recovering the outputs.

Materials and methods

31

Figure 5. General flow chart of a product system for a waste management LCA, studied in Paper IV.

2.3 Data sources and methods for life cycle assessment

The inventory data used in the case studies were compiled from several sources depending on the goal and scope of the study and the system assessed (Table 4). The approaches used for the LCAs differed in the studies (Table 5). In Paper I only emission profiles based on LCIs were assessed, whereas an LCIA was carried out to the normalisation phase in Papers III and V. In Paper II an LCIA was carried out even further and was extended to weighting. Different methods and approaches were used for assessing the life cycle envi-ronmental impacts (Table 5). Paper II compared the LCIA results produced by three different methods, whereas in Papers III and V, and LCIA was performed using one method. The impact categories included in the LCIA were selected based on the objectives of the stud-ies, availability of data and the relevance of the impact categories to the studied product sys-tems. Several impact categories were assessed in Papers II, III and V, whereas in Paper IV only climate change impacts were assessed (Table 5). In Paper IV climate change impacts were considered sufficient for providing a general picture of the environmental and economic impacts of C&DW management and the key issues influencing these.

Materials and methods

32

Table 4. Data sources for the case studies included in the thesis. If not separately indicated, the cost data originat-ed from the same source as the environmental data. CCB = corrugated cardboard, C&DW = construction and demolition waste, NP = newspaper.

Paper Processes / life cycle phases Data sources Paper I Forestry operations KCL-ECO Database

Pulp and paper mill: NP manufactur-ing, recycling, energy production

Process data from the pulp and paper mill (Paper I)

Separate collection and pretreatment of NP

Data from operators (Paper I)

Landfilling of NP Generic data from literature (Paper I) Cost data from operator (Paper I)

Thermal treatment with energy re-covery of NP

Planning data for co-combustion; Litera-ture data for incineration (Paper I)

Energy production avoided Generic data from literature (Paper I)

Paper II As for Paper I As for Paper I Paper III Pulp and paper production KCL-ECO Database

Printing Data from operators (Paper III) End-of-life scenarios for NP, including

material and energy production avoided

Generic data from literature (Paper III)

Paper IV Quality and quantity of waste gener-ated

For all processes/phases: Expert evaluations (Paper IV) Generic data from literature (Paper IV) Cost data from operator (Paper IV)

Distribution of waste fractions Quality of outputs from C&DW man-

agement Energy production avoided Material production avoided

Paper V Plastic crate manufacturing Ecoinvent v. 2.2 Database (Paper V) Plastic crate delivery Data from operators (Paper V) Plastic crate end-of-life, including ma-

terial and energy production avoided Generic data from literature (Paper V)

CCB manufacturing Data from operators (Paper V) CCB delivery Data from operators on the current system

(Paper V) CCB end-of-life, including material

production avoided Data from operators, based on assump-

tions on the recycling process (Paper V)

Table 5. Comparison of the methodologies used for the life cycle assessments (LCAs) in the papers. LIC = life cycle inventory, LCIA = life cycle impact assessment, DAIA = Decision Analysis Impact Assessment, EPS = Envi-ronmental priority system.

Paper Final phase of the LCA1)

LCIA method-ology used

Modelling approach of the LCIA meth-odology

Number of impact categories assessed

Reference values for normalisation

Weighting method used in the LCIA model

I LCI None II LCIA (weighting) DAIA2) midpoint 5 Finnish Panel

Eco-indicator 993)

endpoint 11 European Panel

EPS 20004) endpoint 16 None Monetary III LCIA (normalisation) ReCiPe5) midpoint 7 European,

Finnish con-sumption, Finnish produc-tion

IV LCIA (characterisa-tion)

ReciPe midpoint 1 None

V LCIA (normalisation) ReCiPe midpoint 6 European 1) see figure 2, chapter 1.3; 2) Seppälä, 1999; 2003; 3) Goedkoop and Spriensma, 2001; 4) Steen, 1999a; 1999b; 5) version 1.08, December 2012; Goedkoop et al., 2009

Materials and methods

33

2.4 Life cycle costing and material flow analysis

To strengthen the LCA, an economic assessment was combined with it in Paper II (Table 2). In Paper IV the LCA was combined with both an MFA and an economic assessment. The cost calculations and environmental impact assessments were performed separately, but they were both applied to the system defined for the LCA as the basis for calculations. Some ex-ceptions (explained below), were made case specifically to the boundaries. MFA results on the treatment lines of different waste fractions were used as the basis for the LCA system def-inition in Paper IV. The economic assessment was performed using SLCC in Paper II and ELCC in was used in Paper IV. SLCC looks at the costs from the point of view of society, hence SLCC was preferred for analysing the costs of waste policy alternatives for paper waste, for which the responsibil-ity of management is shared by society (i.e. municipalities) and industry (Paper II). The regu-lations on producer responsibility (Government decree on separate collection and recycling of wastepaper, 528/2013) have obliged the forest industry to organise the collection and re-covery of wastepaper but left the municipalities with the responsibility for non-recovered wastepaper. Here the analysis must combine the costs and benefits accruing to multiple ac-tors, firms and consumers. In contrast, in the case of C&DW (Paper IV) the municipalities are only responsible for the C&DW generated by households, and the majority of C&DW is taken care of by companies. Here the viewpoint of a single enterprise can be used in the cost calcu-lations, as is done in ELCC, since the company both bears the costs and receives the credits from the treatment and recovery of C&DW. The SLCC (Paper II) covered the costs associated with the life cycle of a newspaper within the boundaries of the LCA: e.g., direct costs of investments, labour and energy (Paper II). One exception was made compared to the LCA: the costs of increased communication efforts were included, although the environmental impacts were not included in the LCA, since they were assumed to be low compared to the overall impacts. The ranking for newspaper waste man-agement options from the LCIA was compared to that produced by SLCC. Additionally, the external costs associated with environmental impacts related to climate change and depletion of fossil fuels were added and compared to the LCIA results of the corresponding impact cat-egories. The ELCC assessment (Paper IV) was implemented in a C&DW management company. The timescale of the analysis was one year of C&DW management system operations and the as-sessment included direct internal costs and capital, following the guidelines of ELCC (Swarr et al., 2011). A material flow analysis (MFA, Paper IV) was applied to produce a description of the waste flows within the system and to quantify the inputs and outputs of the C&DW system. Material and energy recovery rates were calculated from the outputs of sorting and separation, i.e. the inputs to the recovery processes. The MFA was performed using the STAN 2.0 software (Cencic and Rechberger, 2008).

34

3. Results and discussion

3.1 Environmental and economic impacts on different waste hierarchy levels

Different levels of waste hierarchy and their emissions or environmental and/or economic impacts were assessed especially for two systems, namely for newspaper production and WM (Papers I and II) and for a bread delivery system focusing on packaging (Paper V). The re-sults showed that moving up in the waste hierarchy does not necessarily improve the envi-ronmental or economic performance of the system studied. Of the two studied cases, only the one concerning the WM of discarded newspaper (Papers I and II) focused on comparing the hierarchy levels for the identical waste fraction. The evalua-tion was performed in three phases: 1) using emission profiles from the LCI (Paper I), 2) us-ing LCIA results obtained via three LCIA methods (Paper II) and 3) using additionally eco-nomic results from SLCC (Paper II). The LCI emissions profile results supported the waste hierarchy by showing lower emissions for recycling and thermal treatment with energy re-covery than for landfilling. Different emission parameters slightly changed the preference order, nevertheless landfilling clearly showed the highest emissions. However, inventory data refers to inputs and outputs of the system, not environmental impacts, hence the environ-mental performance of different options cannot yet be interpreted on the basis of the LCI results (ISO, 2006b). From the LCI phase, the study was continued with an LCIA (Paper II). The LCIA results showed that increasing the energy recovery of newspaper waste in a co-combustion plant resulted in a better performance compared to recycling if fossil fuels were compensated for. However, different LCIA models and different impact category results gave slightly different results. Especially, the impacts related to land occupation and conversion, which are included in the Eco-Indicator 99 LCIA methodology (in the damage to ecosystem quality category), showed that recycling outperformed energy recovery. Nevertheless, none of the models ade-quately assessed the biodiversity effects of forestry. Hence, a highly important impact of wood-based products was underestimated in the study. In contrast to the LCIA results, the SLCC (without environmental costs) showed a preference, i.e. lowest costs, of landfilling fol-lowed by recycling (combined with incineration). Several LCA studies on paper WM (e.g. Merrild et al., 2008; 2012; Villanueva and Wenzel, 2007) have found recycling to perform environmentally better than energy recovery. Addi-tionally, based on their review of LCAs for SWMSs Laurent et al. (2014a) concluded that with the exception of landfilling, there was no definite agreement on the best performing waste

Results and discussion

35

treatment technology, but most studies tended to favour recycling over landfilling and ther-mal treatment for paper. However, they also concluded, that the strong dependence of each SWMS on its context or local conditions prevents generalisation of LCA results, which is cur-rently symbolised by the waste hierarchy (Laurent et al., 2014a). Contradicting the studies by Merrild et al. (2008; 2012) and Villanueva and Wenzel (2007) this study indicated that recy-cling of newspaper generated fewer environmental benefits than its energy recovery. Com-mon to Finland, the mill studied is an integrated pulp and paper mill, where the interconnec-tions between processes and flows are complex and may affect the product system in unex-pected ways when changing WM options. Recycling could have been improved by making changes in the paper mill phase, such as using grid electricity with low emissions for power generation instead of fossil fuels. The case study on delivery packaging (Paper V) compared the impacts of reusable or recycla-ble packaging on the performance of the delivery system. The waste hierarchy specifies reuse on a higher level of priority than recycling and several studies comparing packaging systems have concluded that reusable plastic containers outperform single-use packaging (e.g. Levi et al., 2011; Raugei et al., 2009; Singh et al., 2006). This assessment showed that with the spe-cific delivery system studied, the recyclable packaging performed better than the reusable packaging. A prerequisite for this result was that the recycled product had an efficient recy-cling system, where virgin material was compensated with recycled material. In Raugei et al. (2009) a comparison of transport of a batch of chemicals by means of single-use disposable fibre drums versus reusable steel drums resulted in the reuse of drums 200 times with lower environmental impacts than the single-use fibre drums. However, unlike the recyclable deliv-ery boxes in Paper V, the fibre drums were landfilled. The results of the delivery packaging study are very case specific. Transportation was shown to dominate in all impact categories and changes in the weights of the crates/boxes and the modes of transport, for example, may change the results. The results underline that material properties such as reusability or recyclability cannot solely be used as the basis for choosing materials for a specific use. In contrast, a broader view needs to be obtained using a systemic assessment. This conclusion is also supported by Sørensen and Wenzel (2014), who found bedpans made of disposable plastics to perform better than reusable steel products. The assessment of the benefits gained by waste recovery is crucial in LCAs for WM. For the material recovery options in both the newspaper study and the bread delivery packaging study, the impacts of the wood saved by recycling were assessed conventionally, i.e., without considering the potential increase of carbon sequestration in forests or an alternative use for the saved wood. The impacts of wood harvesting on the forest carbon balance have been studied and discussed, for example by Soimakallio et al. (2016). However, scientific consen-sus is yet to be found on how to incorporate CO2 removal by forests into product specific as-sessments. Assumptions on the alternative use of wood saved by recycling have been identified as one of the key factors for the environmental impacts concerning the recycling of paper in compari-son to its thermal treatment with energy recovery by Villanueva and Wenzel (2007), for ex-ample, and Merrild et al. (2012). Energy use of the saved wood to compensate fossil fuels has

Results and discussion

36

been assumed when modelling paper recycling by Merrild et al. (2012). However, the energy use of saved wood was not included in the LCAs of this study. The reasoning for this was that the wood for newsprint manufacturing is obtained mainly from forest management opera-tions, i.e. from thinning. If the demand for wood for paper manufacturing decreases, the out-come may be fewer forest management activities leaving the thin trees (i.e. the pulp wood) in the forest. Hence, no alternative use would exist. In the current situation, however, energy use would be a relevant assumption, since the use of forest-based bioenergy is constantly growing. Wood-based energy is primarily (80% in 2014) produced using residues from the forest industry and from silvicultural and harvesting operations (Koponen et al., 2015). Additionally, new products are being developed from wood fibres to compensate for the de-creasing use of wood in paper manufacturing. A recent example of such product development are wood based textiles (Michud and Rissanen, 2014). Such textiles could in the long run re-place the production and use of cotton and this production could in the future be assumed as an alternative for the wood saved by paper recycling. The European Commission’s Circular Economy Package also calls for new wood-based products and for the cascading use of re-newable resources (European Commission, 2017a). Biorefineries can offer practical solutions for generating a wide range of products for food, material and energy applications (Metsä Group, 2017). The cascading use of wood requires that the biomass is converted into prod-ucts, which can be recycled or used for energy production after their first utilisation stage (Vis et al., 2016). The consideration of cascading would probably have changed the results of this study, but to verify this, the outcome would need to be assessed. The benefits of the waste to energy (WtE) options studied in Paper II were assessed assuming that the energy produced from waste based fuels would compensate fossil fuels, which sub-stantially affected and improved the performances of the WtE options. In the comparison of MSW management for seven countries, Bassi et al. (2017) concluded that WtE plants can lead to environmental loads or savings, depending on the energy efficiency of the plant and on the composition of the energy being substituted. However, on the one hand, the use of renewable resources for energy production is increasing (Bioeconomy, 2017), and with this development the benefits gained from WtE will decrease and probably improve the environmental perfor-mance of recycling in the future. On the other hand, in the long run, in the circular economy the use of secondary materials in production should become a default and this would proba-bly reduce the benefits gained by recycling. Hence, it is challenging to predict the effects of these changes on the environmental performance of different WM options without systemic assessments. These conclusions are supported by the results by Arushanyan et al. (2017), who assessed several future scenarios for the Swedish WM system. They concluded that by recycling and recovering materials and energy the WM system would continue to contribute to the production of materials, energy carriers and other products replacing virgin produc-tion. However, this would lead to lower environmental benefits in the more sustainable sce-narios, where the surrounding energy and transportation systems are less harmful to the en-vironment and less waste is produced (Arushanyan et al., 2017).

Results and discussion

37

3.2 Added value for decision-making of combining life cycle based approach-es

All the case studies presented here were research oriented, and did not directly aim at pro-ducing results for specific decision-making situations. However, in general it can be claimed that in the contemporary society, decision-making is still heavily based on costs. Environ-mental aspects are gaining more weight, but it is clear that very seldom do environmental values alone determine decision-making. Adding an economic analysis to the environmental assessment, as was done in Papers II and IV, increases the usability of the LCA results. The joint assessment of environmental and economic impacts of the newspaper WM options (Paper II) showed that including both dimensions in the assessment is crucial for providing a basis for making sustainable decisions. The differences in the environmental impacts be-tween material and energy recovery were relatively small. Concentrating solely on the eco-nomic aspects would lead to the environmentally worst alternative, whereas the environmen-tally best solution would result in the highest costs. When adding environmental costs to the SLCC calculations, the results did not change crucially, but showed that both options that included higher recycling (Cases 2a and 3a) had the second lowest costs (Table 6). This way SLCC provided an important additional aspect for the decision-maker by showing that when environmental costs are considered in the assessment, the thermal treatment with energy recovery of separately collected waste paper is more expensive than recycling (Table 6). Larsen et al. (2010) assessed the environmental and economic impacts of an MSW manage-ment system with alternative collection systems for recyclables (paper, glass, metal and plas-tic packaging) in Aarhus and obtained similar results on the economic impacts but contradic-tory results on the environmental impacts. The results by Larsen et al. (2010) showed that enhancing recycling instead of incineration was recommendable due to improved perfor-mance in several environmental impact categories. In addition, the costs for the municipality from waste collection and treatment were reduced by increasing recycling, mainly because the high cost of incineration was avoided. However, the results were considered to be sensi-tive to assumptions made about the quality and potential utilisation of the recyclables and the energy substituted by the WtE. (Larsen et al., 2010).

Table 6. The ranking of the newspaper (NP) waste management alternatives with the life cycle impact assessment (LCIA) results produced by DAIA (Decision Analysis Impact Assessment), Eco-indicator 99 and EPS 2000 (Envi-ronmental priority system) (I = lowest environmental impacts, V = highest environmental impacts), with the social life cycle costing (SLCC) results without environmental costs, and with the SLCC results including environmental costs (related to CO2 emissions and resource scarcity) (I = lowest costs, V = highest costs) (modified from Paper II).

Waste manage-ment alternative

Ranking with LCIA results Ranking with SLCC results DAIA, impact value / t of NP

Eco-indicator 99, ecopoints / t of NP

EPS 2000, ELUs1) / t of NP

Excluding envi-ronmental costs, million €

Including envi-ronmental costs, million €

Case 1 7.31 (V) 36.1 (V) 278 (V) 2.55 (I) 4.15 (I) Case 2a 6.20 (III) 33.8 (III) 248 (II) 3.54 (III) 4.81 (II) Case 2b 3.73 (I) 28.0 (I) 228 (I) 5.11 (IV) 5.95 (IV) Case 3a 6.74 (IV) 35.4 (IV) 261 (III) 3.49 (II) 4.81 (II) Case 3b 5.32 (II) 33.1 (II) 268 (IV) 7.16 (V) 8.15 (V) 1) ELUs = environmental load units

In the C&DW management system assessment (Paper IV) an MFA and ELCC were combined with an LCA to provide a better understanding of the studied system. The MFA provided the

Results and discussion

38

basis for an LCA and ELCC. Similarly, Sevigne-Itoiz et al. (2015) combined an MFA with a consequential LCA of GHG emissions to assess the paper and cardboard recycling system in Spain. They concluded that it is necessary to have methodologies that properly map all mate-rial flows as a first step in determining the potential environmental impacts to facilitate the accurate accounting of GHG for decision making (Sevigne-Itoiz et al., 2015). The overall assessment of the C&DW management system showed that both environmental and economic benefits can be gained with the current system. Together the different method-ologies gave a diversified view of the potentials that different waste fractions have in improv-ing the C&DW management (Table 7). The rate of material or energy recovery calculated us-ing the MFA shows the current recovery situation for the specific waste fraction, while the climate change impacts from an LCA indicate the environmental performance and the eco-nomic profits from ELCC describe the economic performance of treatment of the specific fractions. The volume of each fraction reflects the potential that the fraction provides for im-proving the performance of the overall system (Table 7).

Table 7. Material and energy recovery rates (material flow analysis, MFA, results), climate change impacts (LCA results) and economic profits (environmental life cycle costing, ELCC, results) of the treatment of different waste fractions in the current construction and demolition waste (C&DW) management system. The results are compa-rable only within one methodology. (Modified from Paper IV).

Source separated waste fraction

MFA1) LCA2) ELCC3) Volume of the fraction Material

recovery Energy re-covery Climate change impacts Profits

Metal ++ - + ++ small Concrete and mineral ++ - - - large

Wood - ++ ++ + large Miscellaneous waste + + + + medium Mixed waste - - - - medium

Scaling (based on the results of MFA, LCA and ELCC): 1) MFA: ++ = maximal material/energy recovery, + = medium material/energy recovery, - = low material/energy recovery. 2) LCA: ++ = minimum climate change impact, + = medium climate change impact, - = high climate change impact. 3) ELCC: ++ = maximal profits, + = medium profits, - = low profits.

Combining the different approaches indicated ways to achieve the recycling target of 70% set by the EU for 2020. At the time of the study, the recycling rate for C&DW was way beyond the target. Assessment of the effects of assumed changes in the waste composition on the recycling rate (Table 8) showed that an increase in the concrete and mineral fraction (Scenar-io 2) would improve the material recovery rate but would generate a reduction in the overall climate benefits and economic profits of the system. In contrast, an increase in the wood frac-tion (Scenario 1) would reduce the material recovery rate but increase the climate benefits and economic profits (Table 8). Hence, changes in waste composition will not be adequate for achieving the recycling target for C&DW. According to these results, the 70% recycling target could not be attained unless the share of wood in the C&DW can be reduced or recycling technologies for low-quality wood can be de-veloped. According to a recent study on C&DW, the share of wood had decreased to 15% in 2013 due to recession and a reduction in building wooden one-family houses (Salmenperä et al., 2016). In a better economic situation, an increase is again foreseeable. Since the share of wood is exceptionally high in the Finnish C&DW (36%) compared to other EU countries (2-

Results and discussion

39

4%, Monier et al., 2011), the results indicate that regional differences in waste composition may support arguments for differing recycling targets in different European regions or alter-natively focusing the targets on materials, not the overall waste flow. Similarly, Turner et al. (2016) combined LCA and MFA to compare the performance of exist-ing MSW management system and alternative systems formed using different waste policy measures. The recycling rates calculated from the MFA results suggested that national recy-cling targets were unlikely to be met even if the assessed policies were optimally implement-ed. Turner et al. (2016) considered it likely that for the targets to be met, the investigated policies would need to be combined with additional policies that aim at reducing waste gen-eration.

Table 8. The overall recovery rate, material and energy recovery rates, rate of utilisation in landfills, disposal rate and the climate change impact reductions and profits for the typical status, and scenarios 1 and 2 for construction and demolition waste (C&DW) management (modified from Paper IV).

Typical status Scenario 1 Scenario 2 Overall recovery rate1), % 73 70 66 Material recovery, % 38 27 44 Energy recovery, % 35 43 22 Utilisation at landfills, % 6 5 7 Disposal, % 21 25 26 Climate change impact reduc-tion, kg CO2-eq/t of C&DW

360 390 210

Profit, €/t of C&DW 81 79 48 1) material and energy recovery; excluding utilisation at landfills

As a whole, the overall performance of the C&DW management system could be significantly improved by enhancing the sorting of the mixed and miscellaneous waste fractions and de-veloping the current recovery practices especially for the wood waste and mineral waste frac-tions. This finding is supported by Manninen et al. (2015) who concluded that environmen-tally, thermal treatment with energy recovery outperformed the currently most common re-cycling concepts for wood waste, namely the manufacture of composites and particle boards. As for the mineral fraction Butera et al. (2015) also concluded that the current recycling in road construction is preferable to landfilling, but does not generate environmental benefits. Hence, for both wood and mineral waste there is clearly a need for developing environmen-tally more beneficial recycling concepts. Research and development (e.g. ongoing Horizon 2020 project HISER; HISER, 2017) on the possibilities to increase recycling of the various materials, e.g. wood, in C&DW may lead to new solutions. Nevertheless, before implementation, new solutions need to be assessed for their environmental and economic impacts. Considering several rounds of recovery for wood waste, as is the case in cascading use of wood waste, benefits from both material and energy recovery could be gained. According to Vis et al. (2016), the construction and demolition sec-tor is a questionable source of wood for cascading primarily due to the growing material complexity and use of different wood materials. There is more potential in the construction sector, for which research is ongoing to increase wood recyclability and recycled products (Vis et al., 2016). Although the environmental and economic assessments produced somewhat controversial results in the case studies, their joint use deepened the understanding of the systems (Papers

Results and discussion

40

II and IV) and showed that striving for the maximum environmental improvement and min-imum cost may produce a good compromise between ecological and economic aspects (Paper II). Using different methodologies and analysing their results in different combinations can reveal various potentials for improving the environmental or economic performance of the system overall. Although the application of several methodologies increases the amount of information generated, this needs not be confusing for the decision-making. The results can be broken down and combined in various ways in order to ease their interpretation and to highlight different aspects. A common problem for all the methodologies especially in the C&DW case (Paper IV) was the lack of proper data on the treated waste, the quality and quantity of outputs and the technol-ogy and costs of the processes. The results show that in the future, basic information needs to be produced on the characteristics and management of C&DW in order to support the gener-ation of reliable and accurate information on the possibilities and benefits of increasing waste recycling.

3.3 Using normalisation to improve the understanding of LCIA results

The understandability of LCIA results is crucial to their usability and value for decision-making. Normalisation was used to show the contribution of the studied system to the im-pacts of a reference system and hereby improve the understandability of the LCIA results (Papers III and V). The use of different normalisation reference values and their impact on the understanding of LCA results by LCA non-practitioners were assessed for newspapers and digital photo books (Paper III). The use of different reference values for normalisation (Paper III) produced slightly different LCIA results for the studied products. LCA non-practitioners considered that the consump-tion-based national reference values produced somewhat more easily understandable LCIA results than the EU-based reference values. In their survey on the state of the art of normali-sation and weighting approaches in LCA, Pizzol et al. (2017) concluded that consumption-based normalisation references were preferable to production-based ones because of better consistency with the geographical scope of the LCA studies. However, they highlighted that consumption-based normalisation references are in practice almost entirely unavailable to-day (Pizzol et al., 2017). The fact that the harmfulness of the impacts in one category relative to impacts in another category cannot be judged on the basis of the normalised results was difficult for LCA non-practitioners to comprehend (Paper III). Generally, even after normalisation, the LCIA re-sults were considered difficult to understand. The complexity of the processes and systems they describe is one reason behind this. Descriptions of the methods and assumptions used were considered important information for understanding the methods and their restrictions and to support the interpretation. Furthermore, Pizzol et al. (2017) recommended communi-cating normalised results clearly by reporting units and explaining their meaning, as these may not be easily understandable to audiences beyond LCA experts.

Results and discussion

41

The study showed that normalised LCIA results can be feasible in business to business com-munication since they provide information on the contribution of the studied system to the overall environmental impacts of the reference system (Paper III). This helps to focus the system development and technical improvements to mitigate the relevant environmental problems. Descriptions of the methods and assumptions used are important for supporting the interpretation. However, it is challenging to keep the information simple and at the same time scientifically valid. These results highlight the complexity of LCIA results and the challenges in their interpreta-tion. The results support the approach of ISO standards in that characterisation should be the last mandatory step of an LCIA due to the fact that normalisation brings subjectivity into the results and their interpretation. However, subjectivity can be entailed in the assessment also in the earlier phases of LCA, for example in the LCI phase via the selection of input data and assumptions for modelling the system as well as in the LCIA phase via the selection of impact categories. Therefore, transparency is a key requirement for reporting LCA and other assess-ments. The challenge of presenting results from LC-based assessment in a comprehensive and un-derstandable way has been recognised by Finkbeiner et al. (2010), for example, who suggest-ed more sophisticated communication tools such as the “Life Cycle Sustainability Dashboard” and the “Life Cycle Sustainability Triangle” for communicating to both experts and non-expert stakeholders. To improve the communication of LCIA results to experts, decision-makers and the public, the European Commission DG Environment and European Commis-sion's Joint Research Centre (JRC IES) have worked on developing a harmonised methodol-ogy for the calculation of a Product Environmental Footprint (European Commission, 2016). Communication to business partners, consumers, stakeholders, e.g., public administration, NGOs and/or investors was tested in the Product Environmental Footprint Pilots (European Commission, 2017b). Different ways to communicate on environmental impacts to target audiences, such as performance (improvement) labels, barcodes, information available in shops, instruction manuals, product declarations, consumer receipts, websites, applications, campaigns and sustainability rankings were used during the testing (European Commission, 2017b).

3.4 Uncertainties of the assessments

The assessments included in the study contain a number of uncertainties related to the data used, the assumptions concerning, e.g., avoided impacts and the environmental impact cate-gories. The data sources used in the environmental assessments are summarised in Table 4 (Chapter 2.3). In general, the environmental data used for assessing C&DW management included the highest uncertainty, starting from the volume and quality of C&DW, distribution of the waste into various fractions, processing of the fractions, the quality and recyclability of the outputs and the recycling processes. Due to these uncertainties the climate change impacts can be over- or underestimated. Sensitivity checks were not performed since the range of parameter values could not be estimated. Hence, the results of this assessment are rough approximates

Results and discussion

42

and can only be utilised for generating a better understanding of the overall C&DW manage-ment chain and the critical aspects affecting the performance of the chain. In contrast to the C&DW study, parts of the data for the other cases were specific to the pro-cesses and process chains under study. For Papers I and II the data concerning newsprint manufacturing, recycling and energy production at the pulp and paper mill were obtained from the mill, and hence can be considered reliable. Additionally, the data on separate collec-tion and pre-treatment of paper waste prior to recycling consisted of process specific data from operators. In contrast, data on the WtE options were very uncertain, due to the fact that the plants were not yet operating at the time of the study. This together with the assumptions on fuels and energy production compensated for by the waste-based energy generates high uncertainties in the results. For Paper III, database data and generic data from the literature were mostly used. For the printing phase, data were compiled from operators. The data on environmental releases and extractions behind the normalisation reference values used in the study were of different ac-curacy and extensiveness. The Finnish reference values were generated from the environ-mentally extended input-output (EE-IO) model of the Finnish economy, where the emissions were compiled for each industry from the national emission inventories or calculated on the basis of activity information (Seppälä et al., 2009). For the ReCiPe reference values for Eu-rope, on the other hand, the emissions of several substances were based on estimations and calculations (Wegener Sleeswijk et al., 2008). These differences would weaken the compara-bility of the normalised LCIA results produced with the different reference values. However, they would not affect the stakeholders’ perception of the results, since here the focus was not on comparing, but on understanding the results obtained with different normalisation refer-ences. In the bread delivery study (Paper V) data on the current delivery system using plastic crates and on CCB manufacturing were obtained from the operators. However, databases were used for plastic crate manufacturing and its end-of-life options. Since the use of CCB in the system was an option for the future, assumptions were made regarding its recycling. However, the data for the recycling process were obtained from the operator, but it is uncertain, whether the CCB would in practice be recycled in the modelled process. Economic data is always difficult to obtain, but in the newspaper case study (Paper II) relia-ble and comprehensive data on costs were obtained from operators similarly to the data on emissions and use of resources. For the options including thermal treatment with energy re-covery the cost data were more uncertain, since the facilities were not yet operating. In con-trast, in the C&DW management system study (Paper IV), cost data was received from just one company. Yet the data gained was comprehensive for the treatment chain of this compa-ny and therefore helped in improving the understanding of the C&DW management chain. The assessment of environmental impacts was not comprehensive and sufficient in all case studies. In the newspaper case studies (Papers II and III), the biodiversity impacts of paper recycling were not included in any of the LCIA models used, although they are of great im-portance in connection with wood based products. Additionally, only a restricted number of

Results and discussion

43

impact categories were analysed in the other case studies (Papers III, IV and V), which were selected based on data availability and the relevance to the studied product systems. For the newspaper and digital photobook LCA (Paper III) the impact categories related to human toxicity or ecotoxicity were left out primarily due to the incompleteness of inventory data. Inventory data from generic databases included toxic emissions, but the process-specific data for the core processes did not. For the C&DW management system (Paper IV) only climate change impacts were assessed due to the lack of data on other than energy use within the treatment chains. In comparing the reusable and recyclable delivery boxes (Paper V), one important category missing from the selected impact categories was land use, for which no data were obtained and hence it was not considered. Restricted impact categories may limit the possibility to interpret the environmental impacts of the system from the LCA results alone. Hence, care should be taken in drawing any conclusions. When assessing waste containing wood-based products, the inclusion of the positive impacts on biodiversity or carbon sequestration arising from saving wood by recycling would proba-bly affect the results of this study. However, the models used in this study did not consider these elements. The impacts of forestry on biodiversity may result from complicated and wide-ranging chains of consequences, such as increased recycling reducing the demand and price of fibre wood, possibly lowering the value of forest land, and consequently giving rise to interest among forest owners to protect their forest (if there are incentives to promote pro-tection). The biodiversity impacts from such a chain could be more extensive than could be evaluated on the basis of the use of wood alone. Carbon sequestration of forests is an addi-tional important part of the life cycle of fibre-based products. However, at the time of con-ducting the case study, no scientific consensus existed on how to incorporate CO2 removal by forests into product specific LCAs. Therefore a conventional approach was applied and no benefit from carbon sequestration was allocated to wood-based products. Landfills as carbon sinks are an additional issue concerning wood-based products, such as newspaper and cardboard. The calculations in Papers I, II, III and IV included only emissions generated from the degradation of waste in landfills. However, studies have shown that the degradation of paper can be extremely slow and environmental conditions in landfills and the presence of lignin can limit the decomposition (Barlaz, 2006; Ximenes et al., 2008). It has been suggested that the role of landfills as a sink of biotic CO2 in connection with paper waste, for example, should be included in LCAs as benefits for the waste remaining non-degraded in the landfill (Björklund et al., 2010; Christensen et al., 2009). Some of the LCA-models designed for assessing WM options include carbon sequestration in landfills, but some do not (Gentil et al., 2010). In practice, the option of landfilling organic waste is becom-ing non-available due to the legislative restrictions taking place in many countries in Europe. Theoretically, however, considering benefits for landfilled paper waste acting as a CO2 sink would improve the environmental impacts of landfilling. Most of the assessments in this study included assumptions made concerning the possibili-ties for waste based energy or materials to compensate for other types of energy and materi-als. However, uncertainty about the future complicates this type of modelling and therefore, the applicability of the results is limited. The operational environments are constantly chang-ing, and currently, several strong drivers such as the Circular Economy Package (European

Results and discussion

44

Commission, 2017a) and the Finnish Bioeconomy Strategy: Sustainable growth from the bio-economy (Bioeconomy, 2017), are driving towards major changes in the European and Finn-ish economies. For example, in the future, fuels used for energy production will increasingly be bio-based fuels or other fuels with low emissions. Hence, compensating energy production with waste-based energy will in the future be less beneficial to the environment. Since the study on discarded newspapers (Papers I and II), the operational environment of WtE sys-tems has changed significantly in Finland. Since 2007 eight incineration plants have started operating, which are more efficient in comparison to the modelled incineration plant, pro-ducing between 20% and 40% electricity, the rest being heat. In comparison, the incineration plant used in Papers I and II generated 3.5% electricity and 96.5% heat. Hence, the evalua-tion performed in the case study does not reflect the situation today concerning the waste incineration option. Finally, the comparisons of environmental impacts of reusable and recyclable products (e.g. Raugei et al., 2009; Sørensen and Wenzel, 2014) are in general limited by the fact that they do not consider the production avoided by the reuse of the product. The same limitation ap-plies to the LCA in Paper V. Here, the reusable packaging (plastic crate) is reused 700 times and the modelling allocates the environmental burden of the plastic crate production equally for each 700 uses. Hence, one delivery (the FU used) only receives 1/700 of the burden of the crate production. Nevertheless, with each reuse, the production of one new packaging could additionally be avoided. The modelling of this avoided production and the implications of adding it to the assessment faces challenges, since the avoided product would not be identical to the one reused, since the durability would presumably be lower. Additionally, the model-ling should extend to include the end-of-life phase of the avoided product, whether it would be recycled, or used for energy recovery or landfilling. The impacts of such additions cannot be predicted but need to be studied.

45

4. Conclusions and recommendations for fu-ture research

The reviews of LCA studies on MSW or fractions of it have indicated that the performance of different WM options varies due to differences in the waste composition, system boundaries used in the modelling, technologies within the system and substituted energies among other things, but also very much due to the systems context and local specificities. Similar con-straints can be recognised for C&DW management, since building materials and solutions and practices for the demolition of buildings and treatment of waste fractions vary in differ-ent regions. Hence, although several studies already exist, there is still need for national, re-gional and municipal level assessments of the environmental performance of WM options representing various levels of the waste hierarchy. In this study, various Finnish WM flows and systems were assessed. The primary objectives of the assessments varied. Common to all the assessments was that they included wastes or products from the forest sector, which is one of the most important economic sectors in Fin-land. The case studies were analysed using life cycle based approaches such as LCA, SLCC, ELCC and MFAs. The research aimed at evaluating whether the environmental and economic results of the life cycle based approaches supported the waste hierarchy order of WM options in the Finnish context (RQ1). Moreover the aim was to assess whether the combination of life cycle based approaches provides added value for decision-making concerning Finnish WM systems (RQ2). Finally, the use of normalisation was analysed to improve the understanding of LCA results (RQ3). The environmental or economic impacts assessment results of the case studies included in this thesis did not give clear support to the waste hierarchy order of WM options (RQ1). In case studies where landfilling was one of the evaluated options, the environmental perfor-mance of this lowest hierarchy level was worse than that of recycling or thermal treatment with energy recovery. Otherwise, the results were less unambiguous. Although reuse is on a higher level in the waste hierarchy than recycling, the delivery system using a recyclable de-livery packaging (cardboard) and utilising an efficient recycling system generated lower envi-ronmental impacts than the one using a reusable one (plastic crate). Hence decisions on which material to choose for a specific use cannot be made on the basis of the material itself, but need to be supported by system level assessments. For the newspaper waste, the WtE option showed lower overall environmental impacts than recycling, but the results varied in individual impact categories. The lower costs favoured recycling of newspaper over thermal treatment with energy recovery.

Conclusions and recommendations for future research

46

In the modelling of the case studies, not all aspects related to wood and forestry were satis-factorily covered. The benefits of wood saved by recycling through carbon sequestration in forests or alternative use of the saved wood were not included in the assessments due to the lack of scientific consensus on methodologies, among other reasons. The current national and international policies that strive towards a bioeconomy and circular economy, aim at reduc-ing the use of non-renewable resources and using renewable resources more efficiently, e.g. by cascading the use of biomass. If cascading were considered the results would probably be different. In order to verify the results, the outcome would need to be systematically assessed. Modelling biogenic processes to understand the consequences of the increasing and more efficient use of renewable resources is highly important. In Finland, and other regions with ample forest resources, methods are needed for combining WM systems for wood-based waste with forestry, especially the balances of carbon in forests and the use of wood. All the case studies were research oriented, and did not directly aim at producing results for specific decision-making situations. However, in general, LCA results can be used in deci-sion-making. Environmental aspects are gaining more weight, but very seldom do environ-mental values alone constitute a basis for decision-making. This research showed that the combined use of different life cycle based approaches can increase the usability of the results by deepening the understanding of the systems and of the critical aspects for improving the overall performance (RQ2). For the newspaper WM system the SLCC showed that when envi-ronmental costs are considered in the assessment, the thermal treatment with energy recov-ery of separately collected waste paper is definitely more expensive than recycling. For C&DW management the combination of MFA, LCA and ELCC identified the potential of dif-ferent waste fractions for improving the overall performance of the WM system. Various frac-tions contribute to the performance differently, for example material recovery of the mineral fraction generated few or no environmental benefits, whereas energy recovery of wood gen-erated high environmental benefits. Environmentally and economically more beneficial recy-cling concepts should be developed for both fractions to increase the recycling rate of C&DW. However, systemic environmental and economic assessments should be connected to the development of recycling solutions in order to ensure improvements in both impacts. Addi-tionally, the joint assessment of recovery rates (by MFA), climate impacts (by LCA) and costs (by ELCC) of scenarios with varying waste compositions enabled the evaluation of the ability to achieve the 70% recycling target for C&DW in 2020. The results indicate that regional dif-ferences in waste composition may support arguments for differing recycling targets in dif-ferent European regions or alternatively focusing the targets on materials, not the overall waste flow. The understandability of LCIA results is essential to their usability for decision-making. The use of different normalisation reference values was shown to have some impact on the under-standability of the LCIA results, yet overall the results remained complex to understand for an LCA non-expert (RQ3). The inclusion of documentation on the methods and assumptions used were considered important for understanding the methods and their restrictions and to support the interpretation. The challenge of communicating LCA results has been recognised broadly and to tackle it, the European Commission together with industry is developing a harmonised Product Environmental Footprint methodology. This methodology could be a

Conclusions and recommendations for future research

47

way to improve and ease the communication of LCIA results for non LCA-practitioners by using unified terminology, consistent calculations and clear output. All the applied approaches suffered from inadequacy of data. Databases are available for pro-cess data for environmental assessments on a general level, but region specific public data are also needed. The results of this study show that basic information needs to be generated es-pecially on the amounts, quality, processing and management of C&DW in order to support comprehensive assessment of the possibilities for increasing waste recycling and maximising the benefits. Data for economic assessments is in general more challenging to obtain than environmental data. Recently a cost model including cost data on several WM processes has been developed. However, the applicability of this data in Finland is still to be assessed. Shortcomings in all the adopted approaches were recognised and hence the outcomes of the case studies are not unambiguous. LCAs and other life cycle based studies are always simpli-fications of reality and include many assumptions, which can be decisive for the outcome. In the modelling of WM systems the assumptions made on avoided impacts potentially obtained through the material and energy recovery of waste are key factors, but uncertainty about the future complicates making the assumptions. The energy sector especially is facing constant changes resulting from the increasing use of bio-based fuels and other cleaner forms of ener-gy production. Hence, in the long run the benefits gained from compensating for fossil fuel based energy with energy recovered from waste will decrease and the performance of WtE options will decline. Simultaneously, in the circular economy the use of secondary materials in production is increasing and causing a reduction in the benefits of recycling obtained pri-marily from compensating the use of virgin materials. Hence, predicting the effects of such changes on the environmental performance of various WM systems is challenging and calls for comprehensive systemic assessments.

49

References

Ahamed, A., Yin, K., Ng, B.,J.,H., Ren, F., Chang, V.W.-C. & Wang, J.-Y. 2016. Life cycle assessment of the present and proposed food waste management technologies from environmental and eco-nomic impact perspectives. J Clean Prod 131:607-614.

Arena, U. & Di Gregorio, F. 2014. A waste management planning based on substance flow analysis.

Resour Conserv Recycl 85:54–66. Arushanyan, Y., Björklund, A., Eriksson, O., Finnveden, G., Ljunggren Söderman, M., Sundqvist, J.-O.

& Stenmarck, Å. 2017. Environmental Assessment of Possible Future Waste Management Sce-narios. Energies 10:247-274.

Baan, L., Alkemade, R. & Koellner, T. 2012. Land use impacts on biodiversity in LCA: a global ap-

proach. Int. J. Life Cycle Assess. 18:1216–1230. Barbalace, K. 2003.The History of Waste. EnvironmentalChemistry.com.

http://EnvironmentalChemistry.com/yogi/environmental/wastehistory.html. [Accessed 1.11.2017]

Barlaz, M.A. 2006. Forest products decomposition in municipal solid waste landfills. Waste Manag

26:321–333. Bassi, S.A., Christensen, T.H. & Damgaard, A. 2017. Environmental performance of household waste

management in Europe - An example of 7 countries. Waste Manag 69:545–557. Bernstad, A. & la Cour Jansen, J. 2012. Review of comparative LCAs of food waste management sys-

tems – Current status and potential improvements. Waste Manag 32:2439–2455. Bioeconomy 2017. Finnish bioeconomy strategy 2014. http://www.bioeconomy.fi/facts-and-

contacts/finnish-bioeconomy-strategy/?lang=en [Accessed 31.10.2017] Björklund, A. & Finnveden, G. 2005. Recycling revisited – life cycle comparisons of global warming

impact and total energy use of waste management strategies. Resour Conserv Recycl 44:309-371.

Björklund, A., Finnveden, G. & Roth, L. 2010. Application of LCA in waste management. In Christen-

sen, T. H. (Ed.), Solid waste technology & management, Chapter 3.2, John Wiley & Sons, Ltd, Chichester (ISBN: 978-1-405-17517-3).

Bovea, M.D & Powell, J.C. 2016. Developments in life cycle assessment applied to evaluate the envi-

ronmental performance of construction and demolition wastes. Waste Manag 50:151–172. Breedveld, L., Lafleur, M. & Blonk, H. 1999. A framework for actualizing normalisation data in LCA:

experiences in the Netherlands. Int J Life Cycle Assess 4:213–220.

References

50

Brunner, P. H. & Rechberger, H. 2004. Practical handbook of material flow analysis. Boca Raton, FL, USA: CRC.

Butera, S., Christensen, T.H. & Astrup, T.F. 2015. Life cycle assessment of construction and demolition

waste management. Waste Manag 44:196–205. Cao, V., Margni, M., Favis, B.D. & Deschênes, L. 2017. Choice of land reference situation in life cycle

impact assessment. Int J Life Cycle Assess 22:1220–1231. Carlsson Reich, M. 2005. Economic assessment of municipal waste management systems—case stud-

ies using a combination of life cycle assessment (LCA) and life cycle costing (LCC). J Clean Prod 13:253–263.

Cencic, O. & Rechberger, H. 2008. Material Flow Analysis with Software STAN. J Environ Eng Manage

18:3-7. Christensen, T.H., Gentil, E., Boldrin, A., Larsen, A.W., Weidema, B.P. & Hauschild, M. 2009. C bal-

ance, carbon dioxide emissions and global warming potentials in LCA-modelling of waste man-agement systems. Waste Manag Res 27:707–715.

Cleary, J. 2009. Life cycle assessments of municipal solid waste management systems: A comparative

analysis of selected peer-reviewed literature. Environ Int 35: 1256–1266. Cleary, J. 2014. A life cycle assessment of residential waste management and prevention. Int J Life

Cycle Assess 19:1607–1622. Clift, R., Doig, A. & Finnveden, G. 2000. The application of life cycle assessment to integrated solid

waste management. Part 1, Methodology. Trans IChemE 78, Part B: 279-287. Consoli, F., Allen, D., Boustead, I., Fava, J., Franklin, W., Jensen, A.A., de Oude, N., Parrish,

R.,Perriman, R., Postlethwaite, D., Quay, B., Séguin, J. & Vigon, B. (Eds.) 1993. Guidelines for lifecycle assessment: A ’Code of Practice’. Society of Environmental Toxicology and Chemistry (SETAC), Brussels.

Diyamandoglu, V. & Fortuna. L.M. 2015. Deconstruction of wood-framed houses: Material recovery

and environmental impact. Resour Conserv Recycl 100:21-30. Döberl, G., Huber, R., Brunner P.H., Eder, M., Pierrard, R., Schönbäck, W., Frühwirth, W. & Hutterer,

H. 2002. Long-term assessment of waste management options – a new, integrated and goal-oriented approach. Waste Manag Res 20:311–327.

Ekvall, T. & Finnveden, G. 2000. The application of life cycle assessment to integrated solid waste

management. Part 2 – Perspectives on energy and material recovery from paper. Trans IChemE, Part B: Process Safety and Environmental Protection, 78: 288-294.

Ekvall, T., Assefa, G., Björklund, A., Eriksson, O. & Finnveden, G. 2007. What life-cycle assessment

does and does not do in assessments of waste management. Waste Manag 27:989–996. Ellen McArthur Foundation 2012. Towards the Circular Economy Vol. 1: an economic and business

rationale for an accelerated transition. https://www.ellenmacarthurfoundation.org/publications [Accessed 15.5.2016]

References

51

Ellen McArthur Foundation 2013. Towards the Circular Economy Vol. 2: Opportunities for the con-

sumer goods sector. https://www.ellenmacarthurfoundation.org/publications [Accessed 15.5.2016]

Ellen McArthur Foundation 2014. Towards the Circular Economy Vol. 3: Accelerating the scale-up

across global supply chains. https://www.ellenmacarthurfoundation.org/publications [Accessed 15.5.2016]

Eriksson, O., Carlsson Reich, M., Frostell, B., Björklund, A., Assefa, G., Sundqvist, J.-O., Granath, J.,

Baky, A., Thyselius, L. 2005. Municipal solid waste management from a systems perspective. J Clean Prod 13:241–52.

European Commission 2010a. European Commission - Joint Research Centre - Institute for Envi-

ronment and Sustainability 2010a. International Reference Life Cycle Data System (ILCD) Handbook - General guide for Life Cycle Assessment - Detailed guidance. EUR 24708 EN - 2010. European Union. http://eplca.jrc.ec.europa.eu/?page_id=86 [Accessed 14.9.2015].

European Commission 2010b. European Commission - Joint Research Centre - Institute for Environ-

ment and Sustainability 2010b. International Reference Life Cycle Data System (ILCD) Hand-book - Framework and requirements for Life Cycle Impact Assessment (LCIA) models and indi-cators. EUR 24586 EN - 2010. European Union. http://eplca.jrc.ec.europa.eu/?page_id=86 [Accessed 14.9.2015].

European Commission 2011. European Commission - Joint Research Centre – Institute for Environ-

ment and Sustainability 2011. Supporting Environmentally Sound Decisions for Waste Man-agement – A technical guide to Life Cycle Thinking (LCT) and Life Cycle Assessment (LCA) for waste experts and LCA practitioners. EUR 24916 EN –2011. Publications Office of the European Union, Luxembourg, LU. http://publications.jrc.ec.europa.eu/repository/handle/JRC65850 [Accessed 7.7.2017]

European Commission 2012. Innovating for sustainable growth: a bioeconomy for Europe. Brussels.

https://www.eea.europa.eu/policy-documents/innovating-for-sustainable-growth-a. [Accessed 18.6.2016]

European Commission 2014. Commission staff working document, Impact assessment. Accompanying

the document: Proposal for reviewing the European waste management targets. SWD(2014) 208. http://eur-lex.europa.eu/legal-content/EN/TXT/?uri=CELEX:52014SC0207 . [Accessed 18.6.2016]

European Commission 2015. Commission staff working document, Additional analysis to complement

the impact assessment SWD (2014) 208 supporting the review of EU waste management tar-gets. SWD(2015)0259 final. http://eur-lex.europa.eu/legal-content/EN/TXT/?uri=CELEX:52015SC0259 [Accessed 18.6.2016]

European Commission 2016. Environmental footprint pilot phase. Policy background.

http://ec.europa.eu/environment/eussd/smgp/policy_footprint.htm [Accessed 26.9.2016] European Commission 2017a. Circular Economy. http://ec.europa.eu/environment/circular-

economy/index_en.htm [Accessed 7.7.2017]

References

52

European Commission 2017b. Environmental footprint pilot phase. The environmental foot-print pi-lots. http://ec.europa.eu/environment/eussd/smgp/ef_pilots.htm [Accessed 30.10.2017]

Van Ewijk, S. & Stegemann, J.A. 2016. Limitations of the waste hierarchy for achieving absolute reduc-

tions in material throughput. J Clean Prod 132:122-128. http://dx.doi.org/10.1016/j.jclepro.2014.11.051.

FCG 2010. Finnish Consulting Group, Veera Sevander. Biojätteen erilliskeräyksen elinkaariarvio. Lop-

puraportti. (LCA on the separate collection of biowaste. Final report). Commis-sioned by Rosk’n Roll Oy Ab.

Ferrão, P., Ribeiro, P., Rodrigues, J., Marques, A., Preto, M., Amaral, M., Domingos, T., Lopes, A. &

Costa, I. 2014. Environmental, economic and social costs and benefits of a packaging waste management system: A Portuguese case study. Resour Conserv Recycl 85:67– 78.

Ferreira, S., Cabral, M., da Cruz, N.F., Marques, R.C. 2014a. Economic and environmental impacts of

the recycling system in Portugal. J Clean Prod 79:219-230. Ferreira, S., Cabral, M., da Cruz, N.F., Simões, P., Marques, R.C. 2014b. Life cycle assessment of a

packaging waste recycling system in Portugal. Waste Manag 34:1725–1735. Ferreira, S., Cabral, M., da Cruz, N.F., Simões, P., Marques, R.C 2017a.The costs and benefits of pack-

aging waste management systems in Europe: the perspective of local authorities. J Environ Plan Manag 5:773-791.

Ferreira, S., Cabral, M., De Jaeger, S., da Cruz, N.F., Simões, P., Marques, R.C. 2017b. Life cycle as-

sessment and valuation of the packaging waste recycling system in Belgium. J Mater Cycles Waste Manag 19:144–154.

Finkbeiner, M., Schau, E.M., Lehmann, A. & Traverso, M. 2010. Towards life cycle sustainability as-

sessment. Sustainability 2:3309-3322. Finnveden, G., Björklund, A., Ekvall, T. & Moberg, Å. 2007. Environmental and economic assessment

methods for waste management decision-support: possibilities and limitations. Waste Manag Res 25:263–269.

Gentil, E., Damgaard, A., Hauschild, M., Finnveden, G., Eriksson, O., Thorneloe, S., Kaplan, P.O., Bar-

laz, M., Muller, O., Matsui, Y., Ii, R. & Christensen, T.H. 2010. Models for waste life cycle as-sessment: Review of technical assumptions. Waste Manag 30:2636–2648.

Global Footprint Network 2016. http://www.footprintnetwork.org/en/index.php/GFN/ [Accessed

15.5.2016] Goedkoop, M. & Spriensma, R. 2001. The Eco-indicator 99—a damage oriented method for life cycle

impact assessment. Methodology Report. 3rd ed. 22.6.2001. Amersfoort, The Netherlands: PRé Consultants B.V.

Goedkoop, M.J., Heijungs, R., Huijbregts, M., De Schryver, A., Struijs, J. & Van Zelm, R. 2009. ReCiPe

2008, A life cycle impact assessment method which comprises harmonised category indicators at the midpoint and the endpoint level. First edition (revised). Report I: Characterisation; Feb-ruary 2013, http://www.lcia-recipe.net. [Accessed 23.9.2016].

References

53

Hauschild, M.Z. & Barlaz, M. 2010. LCA in waste management: Introduction to principle and method. In Christensen, T. H. (Ed.), Solid waste technology & management, Chapter 3.1, John Wiley & Sons, Ltd, Chichester (ISBN: 978-1-405-17517-3).

Havukainen, J., Zhan, M., Dong, J., Liikanen, M., Deviatkin, I., Li, X. & Horttanainen, M. 2017. Envi-

ronmental impact assessment of municipal solid waste management incorporating mechanical treatment of waste and incineration in Hangzhou, China. J Clean Prod 141:453-461.

Heijungs, R., Guinee, J., Kleijn, R. & Rovers, V. 2007. Bias in normalization: causes, consequences,

detection and remedies. Int J Life Cycle Assess 12:211–216. HISER 2017. HISER project. http://www.hiserproject.eu/ Hossain, M.U., Wu, Z. & Poon, C.S. 2017. Comparative environmental evaluation of construction waste

management through different waste sorting systems in Hong Kong. Waste Manag 69:325–335. Hunkeler, D., Lichtenvort, K. & Rebitzer, G. (Eds.) 2008. Enivronmental life cycle costing. Society of

Environmental Toxicology and Chemistry (SETAC), USA. ISBN-13:978-1-880611-83-8. Huppes, G. & van Oers, L. 2011. Background review of existing weighting approaches in Life Cycle

Impact Assessment (LCIA). EUR 24997 EN – 2011. European Union. http://eplca.jrc.ec.europa.eu/?page_id=547

Höglmeier, K., Weber-Blaschke, G. & Richter, K. 2014. Utilization of recovered wood in cascades ver-

sus utilization of primary wood—a comparison with life cycle assessment using system expan-sion. Int J Life Cycle Assess 19:1755–1766

Höglmeier, K., Steubing, B., Weber-Blaschke, G. & Richter, K. 2015. LCA-based optimization of wood

utilization under special consideration of a cascading use of wood. J Environ Manag 152:158-170.

ISO (International Organization for Standardization) 2000. ISO/TR 14049: Environmental Manage-

ment - Life Cycle Assessment – Examples of application of ISO 14041 to goal and scope defini-tion and inventory analysis. Technical report. International Organization for Standardization, Geneva.

ISO (International Organization for Standardization) 2006a. ISO 14040: Environmental Management

- Life Cycle Assessment - Principles and Framework. International Organization for Standardi-zation, Geneva.

ISO (International Organization for Standardization) 2006b. ISO 14044. Environmental Management

- Life Cycle Assessment - Requirements and Guidelines. International Organi-zation for Stand-ardization, Geneva.

Jeswani, H.K., Azapagic, A., Schepelmann, P. & Ritthoff, M. 2010. Options for broadening and deepen-

ing the LCA approaches. J Clean Prod 18:120–127. Judl, J., Manninen, K., Sahimaa, O., Salminen, J., Eskelinen, H., Dahlbo, H. & Väntsi, O. 2016. Wood-

polymer composite - a step towards circular economy? Presented at the SUM 2016 – 3rd Sym-posium on Urban Mining and Circular Economy, 24.5.2016.

References

54

Kim, M.H. & Song, H.B. 2014. Analysis of the global warming potential for wood waste recycling sys-tems. J Clean Prod 69:199-207.

Kiviranta, M. & Tanskanen, N. 2009. Pääkaupunkiseudun keräyskartongin ympäristövaikutusten elin-

kaariarviointi. (Life cycle assesment of discarded cardboard in the Helsinki Metropolitan Area). YTV:n julkaisuja 21/2009. Commisionned by YTV Helsinki Metropolitan Area Council. (In Finnish, with an English abstract).

Kloepffer, W. 2008. Life cycle sustainability assessment of products (with comments by Helias A. Udo

de Haes, p. 95). Int J Life Cycle Assess 13:89-95. Koponen, K., Sokka, L., Pingoud, K., Koljonen, T., Alakangas, E., Sipilä, K., Salminen, O., Sievänen, R.,

Ilvesniemi, H., Routa, J., Ikonen, T. & Asikainen, A. 2015. Sustainability of forest energy in Northern Europe. VTT Technology 237. https://jukuri.luke.fi/handle/10024/520492 [Accessed 26.9.2016]

Laaksonen, J., Merilehto, K., Pietarinen, A., Salmenperä, H., 2017. Valtakunnallinen jätesuunnitelma

vuoteen 2023. Taustaraportti (The national waste management plan to 2023. Background re-port). The Finnish Environment 3/2017. Ministry of the Environment. http://urn.fi/URN:ISBN:978-952-11-4734-0 (In Finnish, with an English abstract)

Larsen, A.W., Merrild, H., Møller, J. & Christensen, T.H. 2010. Waste collection systems for recycla-

bles: An environmental and economic assessment for the municipality of Aarhus (Denmark). Waste Manag 30:744–754.

Laurent, A., Olsen ,S.I. & Hauschild, M.Z. 2011. Normalization in EDIP97 and EDIP2003: updated

European inventory for 2004 and guidance towards a consistent use in practice. Int J Life Cycle Assess 16:401–409.

Laurent, A., Bakas, I., Clavreul, J., Bernstad, A., Niero, M., Gentil, E., Hauschild, M.Z. & Christensen,

T.H. 2014a. Review of LCA studies of solid waste management systems – Part I: Lessons learned and perspectives. Waste Manag 34:573–588.

Laurent, A., Clavreul, J., Bernstad, A., Bakas, I., Niero, M., Gentil, E., Christensen, T.H. & Hauschild,

M.Z. 2014b. Review of LCA studies of solid waste management systems – Part II: Methodologi-cal guidance for a better practice. Waste Manag 34:589–606.

Lautier, A., Rosenbaum, R.K., Margni, M., Bare, J., Roy, P. & Deschênes, L. 2010. Development of

normalization factors for Canada and the United States and comparison with European factors. Sci Total Environ 409:33–42.

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. Re-sour Conserv Recycl 55:246-259.

LCA ReCiPe Midpoint Normalisation 2010. Revised version 1.1; 23 September 2010. http://www.lcia-

recipe.net/normalisation [Accessed 26.9.2016] 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 24:387-400.

References

55

Lighthart, T.N. & Ansems, A.M.M. 2007. Single Use Cups or Reusable (Coffee) Drinking Systems: an Environmental Comparison. TNO Built Environment and Geosciences Report. 2006-A-R0246(E)/B.

Liu, G., Hao, Y, Dong, L., Yang, Z., Zhanga, Y. & Ulgiati, S. 2017. An emergy-LCA analysis of municipal

solid waste management. Resour Conserv Recycl 120:131–143. Mancini, M. S., Galli, A., Niccoluccia, V., Linc, D., Bastianonia, S., Wackernagel, M. & Marchettinia, N.

2016. Ecological Footprint: Refining the carbon Footprint calculation . Ecol In-dic1:390–403. Manfredi, S., Pant, R., Pennington, D.W. & Versmann, A. 2011. Supporting environmentally sound

decisions for waste management with LCT and LCA. Int J Life Cycle Assess 16:937–939. Manninen, K., Judl, J. & Myllymaa, T. 2015. Rakentamisen puujätteiden ja puupakkausjätteiden käsit-

telyvaihtoehtojen elinkaarenaikaiset ympäristövaikutukset (Life cycle environ-mental impacts of different construction wood waste and wood packaging waste processing methods). Reports of the Ministry of the Environment 29/2015. http://hdl.handle.net/10138/159224 (In Finnish, with an English abstract).

Manninen, K., Judl, J. & Myllymaa, T. 2016. Mädätyksen, energiahyödyntämisen ja biopolttoaineen

tuotannon elinkaarenaikaiset ympäristövaikutukset. Biojätteen ja paistorasvan käsitte-lyvaihtoehdot.(The life cycle environmental impact of anaerobic digestion, energy uti-lisation and the production of biofuel – The processing alternatives of biodegradable waste and cooking fat). Reports of the Ministry of the Environment 23/2016. https://helda.helsinki.fi/handle/10138/159940 (In Finnish, with an English abstract).

Martinez, E., Nuñez, Y. & Sobaberas, E. 2013. End of life of buildings: three alternatives, two scenarios.

A case study. Int J Life Cycle Assess 18:1082-1088. Martinez-Sanchez, V., Kromann, M.A. & Astrup, T.F. 2015. Life cycle costing of waste management

systems: Overview, calculation principles and case studies. Waste Manag 36:343–355. Mercante, I.T., Bovea, M.D., Ibáñez-Forés, V. & Arena, A.P. 2012. Life cycle assessment of construction

and demolition waste management systems: a Spanish case study. Int J Life Cycle Assess 17:232-241.

Merrild, H., Damgaard, A. & Christensen, T. 2008. Life cycle assessment of waste paper management:

the importance of technology data and system boundaries in assessing recycling and incinera-tion. Resour Conserv Recycl 52:1391-1398.

Merrild, H., Larsen, A.W. & Christensen, T. 2012. Assessing recycling versus incineration of key mate-

rials in municipal waste: The importance of efficient energy recovery and transport distances. Waste Manag 32:1009–1018.

Metsä Group 2017. The next-generation bioproduct mill. http://bioproductmill.com/. [Accessed

19.9.2017] Michaud, J.-C., Farrant, L., Jan, O., Kjar, B. & Bakas, I. 2010. Environmental benefits of recycling –

2010 update. Final report. WRAP, Material change for a better environment. http://www.wrap.org.uk/sites/files/wrap/Environmental_benefits_of_recycling_2010_update.3b174d59.8816.pdf. [Accessed 3.7.2015]

References

56

Michud, A. & Rissanen, M. 2014. From cellulose to textile fiber and a ready products. http://fibic.fi/wp-content/uploads/2014/05/57_Fibic_Poster_FuBio_Cellulose_Anne-Michud_Marja-Rissanen_SHOK_Summit_20142.pdf. [Accessed 26.9.2016]

Moliis, K., Dahlbo, H., Retkin, R. & Myllymaa, T. 2012. Pohjois-Suomen pakkausjätteiden hyödyntä-

minen - Elinkaaren aikaiset ympäristö- ja kustannusvaikutukset. (Recovery of pack-aging waste in northern Finland – environmental and cost effects on a life-cycle basis). Reports of the Minis-try of the Environment 26/2012. http://hdl.handle.net/10138/41395 (In Finnish, with an Eng-lish abstract).

Monier, V., Mudgal, S., Hestin, M., Trarieux, M., Mimid, S. 2011. Service contract on management of

construction and demolition waste – SR1. Final Report Task 2. European Com-mission (DG ENV), A project under the Framework contract ENV.G.4/FRA/2008/0112. http://ec.europa.eu/environment/waste/studies/mixed_waste.htm#deliverables.[Accessed 26.9.2017]

Morf, L., Tremp, J., Gloor, R., Huber, Y., Stengele, M. & Zennegg, M. 2005. Brominated flame retard-

ants in waste electrical and electronic equipment: Substance flows in a recycling plant. Environ Sci Technol 39:8691–8699.

Morris, J., Matthews, S. & Morawski, C. 2013. Review and meta-analysis of 82 studies on end-of-life

management methods for source separated organics. Waste Manag 33:545–551. Muñoz, I. & Schmidt, J.H. 2016. Methane oxidation, biogenic carbon, and the IPCC’s emission metrics.

Proposal for a consistent greenhouse-gas accounting. Int J Life Cycle Assess 8:1069–1075. Myllymaa, T., Moliis, K., Tohka, A., Rantanen, P., Ollikainen, M. & Dahlbo, H. 2008. Environmental

Loads and Costs of Waste Recycling and Treatment. Inventory report. Reports of the Finnish Environment Institute 28. http://hdl.handle.net/10138/39792 [Accessed 23.9.2016]

Niskanen, A. 2012. Landfill gas management at engineered landfills - Estimation and mitigation of

environmental impacts. Lappeenranta, Acta Universitatis Lappeenrantaensis 506. Diss. Lap-peenranta University of Technology. ISBN 978-952-265-359-8 (PDF), ISSN 1456-4491.

Niskanen, A. 2017. Personal communication, emails and discussions on LCAs performed on biowaste,

cardboard, metal and plastic collection and recovery for several Finnish waste management companies by LCA Consulting (unpublished studies). Several contacts during 2016-2017.

Nygård, H. 2016. Kuopasta kiertotalouteen. Suomen yhdyskuntajätehuollon historia. (From hole to

circular economy. The history of Finnish municipal solid waste management. JLY - Finnish Sol-id Waste Association. ISBN 978-952-5239-01-0.

Ortiz, O., Pasqualino, J.C. & Castells, F. 2010. Environmental performance of construction waste:

comparing three scenarios from a case study in Catalonia, Spain. Waste Manag 30:646-654. Padeyanda, Y., Jang, Y.-C., Ko,Y. & Yi, S. 2016. Evaluation of environmental impacts of food waste

management by material flow analysis (MFA) and life cycle assessment (LCA). J Mater Cycles Waste Manag 18:493–508.

References

57

Pires, A., Sargedas, J., Miguel, M., Pina, J., & Martinho, G. 2017. A case study of packaging waste col-lection systems in Portugal – Part II: Environmental and economic analysis. Waste Manag 61:108–116.

Pizzol, M., Laurent, A., Sala, S., Weidema, B., Verones, F. & Koffler, C. 2017. Normalisation and

weighting in life cycle assessment: quo vadis? Int J Life Cycle Assess 6:853–866. 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 22:443–450. Reuter, M.A., Heiskanen, K., Boin, U., van Schaik, A., Verhoef, E., Yang, Y. & Georgalli, G., 2005. The

Metrics of Material and Metal Ecology. Elsevier. Rochat, D., Blinder, C.R., Diaz, J. & Jolliet, O. 2013. Combining material flow analysis, life-cycle as-

sessment, and multi-attribute utility theory. J Ind Ecol 17:642–55. Rockström, J., Steffen, W., Noone, K., Persson, A., Chapin, F.S.I., Lambin, E., Lenton, T.M., Scheffer,

M., Folke, C., Schellnhuber, H.J., Nykvist, B., De Wit, C.A., Hughes, T., Van der Leeuw, S., Rodhe, H., S€orlin, S., Snyder, P.K., Costanza, R., Svedin, U., Falkenmark, M., Karlberg, L., Co-rell, R.W., Fabry, V.J., Hansen, J., Walker, B., Liverman, D., Richardson, K., Crutzen, P. & Foley, J. 2009. Planetary boundaries: exploring the safe operating space for humanity. Ecol Soc 14(2):32. [online] URL: http://www.ecologyandsociety.org/vol14/iss2/art32/

Salhofer, S., Obersteiner, G., Schneider, F. & Lebersorger, S. 2008. Potentials for the prevention of

municipal solid waste. Waste Manag 28:245–259. Salmenperä, H., Sahimaa, O., Kautto, P., Dahlbo, H., Haavisto, T., Wahlström, M., Bachér, J., Laine-

Ylijoki, J., Vahvelainen, S. & Espo, J. 2016. Reaching the waste recycling targets requires signifi-cant actions. Policy brief 20/2016. Perspectives into topical issues in society and ways to support political decision making. Governments analysis, assessment and research activities. http://tietokayttoon.fi/julkaisu?pubid=15102

Seppälä, J. 1999. Decision analysis as a tool for life cycle impact assessment. In: Klöpffer, W. & Hutz-

inger, O. (Eds.). LCA Documents 4. Eco-Informa Press, Bayreuth (Also published in The Finnish Environment 123. Finnish Environment Institute, Helsinki).

Seppälä, J. 2003. Life cycle impact assessment based on decision analysis. Systems Analysis Laborato-

ry Research Reports A86. Doctoral Dissertation. Espoo, Finland: Helsinki University of Tech-nology, Department of Engineering Physics and Mathematics.

Sevigne-Itoiz, E., Gasol, C.M., Rieradevall, J. & Gabarrell, X. 2015. Methodology of supporting deci-

sion-making of waste management with material flow analysis (MFA) and consequential life cy-cle assessment (CLCA): case study of waste paper recycling. J Clean Prod 105:253-262.

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 vegeta-bles. Packag Technol Sci 19:279-293.

Soimakallio,S., Saikku,L., Valsta, L. & Pingoud, K. 2016. Climate Change Mitigation Challenge for

Wood Utilization – The Case of Finland. Environ Sci Technol 50:5127–5134.

References

58

Soimakallio, S., Cowie, A., Brandão, M., Finnveden, G., Ekvall, T., Erlandsson, M., Koponen, K. & Karlsson, P.-E. 2015. Attributional life cycle assessment: is a land-use baseline necessary? Int J Life Cycle Assess 20:1364–1375.

Sørensen, B.L. & Wenzel, H. 2014. Life cycle assessment of alternative bedpans e a case of comparing

disposable and reusable devices. J Clean Prod 83:70-79. Statistics Finland 2015. Official Statistics of Finland (OSF): Waste statistics [e-publication].

ISSN=2323-5314. Helsinki: Statistics Finland. http://tilastokeskus.fi/til/jate/kas_en.html [Ac-cessed 19.5.2015]

Steen, B. 1999a. A systematic approach to environmental priority strategies in product development

(EPS). Version 2000—general system characteristics. Gothenburg: Chalmers Univer-sity of Technology.CPMReport 1999a:4.

Steen, B. 1999b. A systematic approach to environmental priority strategies in product development

(EPS). Version 2000—models and data of the default method. Gothenburg: Chalmers University of Technology. CPM Report 1999b:5.

Steffen,W., Richardson, K., Rockström,J., Cornell, S.E., Fetzer, I., Bennett, E.M., Biggs,R., Carpenter,

S.R., de Vries, W., de Wit, C.A., Folke, C., Gerten, D., Heinke, J., Mace, G.M., Persson, L.M.,Ramanathan, V., Reyers, B., Sörlin, S. 2015. Planetary boundaries: Guiding human devel-opment on a changing planet. Science 347, 6223.

Sundström, H., Kaila, J., Moliis, K., Dahlbo, H. & Myllymaa, T. 2014. Elinkaarimallin kehittäminen

HSY:n jätehuollolle. Vaihe I: biojätteen käsittely. (Developing LCA-model for HSY Waste Man-agement. Step I: management of biowaste.) (In Finnish).

Swarr, T., Hunkeler, D., Klöpffer, W., Pesonen, H.-L., Ciroth, A., Brent, A. & Pagan, R. 2011. Environ-

mental life-cycle costing: a code of practice. Int J Life Cycle Assess. 16:389-391. Teuber, L., Osburg, V.-S., Toporowski, W., Militz, H. & Krause, A. 2016. Wood polymer composites

and their contribution to cascading utilization. J Clean Prod 110:9-15. Tunesi, S., Baroni, S., & Boarini, S. 2016. Waste flow analysis and life cycle assessment of integrated

waste management systems as planning tools: Application to optimise the system of the City of Bologna. Waste Manag Res 9:933–946.

Turku, I., Keskisaari, A., Kärki, T., Puurtinen, A. & Marttila, P. 2017. Characterization of wood plastic

composites manufactured from recycled plastic blends. Composite Structures 161:469–476. Turner, D.A., Williams, I.D. & Kemp, S. 2016. Combined material flow analysis and life cycle assess-

ment as a support tool for solid waste management decision making. J Clean Prod 129:234-248. UK Environment Agency 2008. An updated lifecycle assessment study for disposable and reusable

nappies. Science Report – SC010018/SR2. https://www.gov.uk/government/publications/an-updated-lifecycle-assessment-for-disposable-and-reusable-nappies. [Accessed 26.9.2017]

Unger, S. & Landis, A.E. 2014. Comparative Life Cycle Assessment of Reused versus Disposable Dental

Burs. Int J Life Cycle Assess 19:1623–1631.

References

59

Villanueva, A. & Wenzel, H. 2007. Paper waste – Recycling, incineration or landfilling? A review of existing life cycle assessments. Waste Manag 27:29–46.

Vis M., Mantau, U. & Allen, B. (Eds.) 2016. Study on the optimised cascading use of wood. No

394/PP/ENT/RCH/14/7689. Final report. Brussels. https://publications.europa.eu/en/publication-detail/-/publication/04c3a181-4e3d-11e6-89bd-01aa75ed71a1

Wegener Sleeswijk, A., Van Oers, L.F.C.M., Guinée, J.B., Struijs, J. & Huijbregts, M.A.J. 2008. Nor-

malization in product life cycle assessment: An LCA of the global and European economic sys-tems in the year 2000. Sci Total Environ 390:227-240.

Virtavuori, V. 2009. Biojätteen käsittelyvaihtoehdot pääkaupunkiseudulla – Kasvihuonekaasupäästö-

jen vertailu. (Appropriate treatment processes for biowaste in the Helsinki Metro-politan Area - benchmarking of greenhouse gas emissions). Publications of the YTV Helsinki Metropolitan Ar-ea Council 8/2009. (In Finnish, with an English abstract).

WWF International, Zoological Society of London, Global Footprint Network & Water Footprint Net-

work 2014. Living Planet Report 2014. WWF, Gland, Switzerland. ISBN 978-2-940443-87-1. http://wwf.panda.org/about_our_earth/all_publications/living_planet_report/ [Accessed 15.5.2016]

Wäger, P.A., Hischier, R.& Eugster, M. 2011. Environmental impacts of the Swiss collection and recov-

ery systems for Waste Electrical and Electronic Equipment (WEEE): a follow-up. Sci Total Envi-ron 429:1746–56.

Ximenes, F.A., Gardner, F.A. & Cowie, A.L. 2008. The decomposition of wood products in landfills in

Sydney, Australia. Waste Manag 28:2344-2354.

-otl

aA

DD

77/

810

2

+ihfjh

a*GMFTSH

9 NBSI 8-7597-06-259-879 )detnirp( NBSI 5-8597-06-259-879 )fdp( NSSI 4394-9971 )detnirp( NSSI 2494-9971 )fdp(

ytisrevinU otlaA

gnireenignE fo loohcS tnemnorivnE tliuB fo tnemtrapeD

if.otlaa.www

+ SSENISUB YMONOCE

+ TRA

+ NGISED ERUTCETIHCRA

+ ECNEICS

YGOLONHCET

REVOSSORC

LAROTCOD SNOITATRESSID

obl

haD

anel

eH

txe

tnoc

hsi

nniF

eht

ni sn

oitp

o tne

mega

nam

etsa

w gn

isse

ssa

ni se

hcao

rppa

des

ab e

lcyc

efi

L y

tisr

evi

nU

otla

A

8102

tnemnorivnE tliuB fo tnemtrapeD

sehcaorppa desab elcyc efiL etsaw gnissessa ni

eht ni snoitpo tnemeganam txetnoc hsinniF

oblhaD aneleH

LAROTCOD SNOITATRESSID