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Life cycle aspects of nanomaterials David Lazarevic and Göran Finnveden Environmental Strategies Research KTH - Royal Institute of Technology Stockholm, Sweden Title: Life cycle aspects of nanomaterials Authors: David Lazarevic and Göran Finnveden ISSN: 978-91-7501-821-8 TRITA‐INFRA‐FMS 2013:4 Division of Environmental Strategies Research Department of Sustainable Development, Environmental Science and Engineering School of Architecture and the Built Environment KTH - Royal Institute of Technology Stockholm, Sweden https://www.kth.se/abe/om-skolan/organisation/inst/see/om/avd/fms Printed by: US AB, Stockholm, 2013 FORWARD This report consists of two parts, a comprehensive compilation of published and ongoing research on nanomaterials from a lifecycle perspective, and an extended summary written in Swedish. The report is funded by the Swedish Governmental Commission charged with developing "A national action plan for the safe use and handling of nanomaterials" (Dir. 2012:89). The authors are grateful for the feedback received during this work from the investigator, Ethel Forsberg, and the commission secretary. The authors have received valuable comments during the work from Eva Hellsten, who has been the contact person in the investigation. The authors are responsible for the content of this report and the views expressed herein are those of the authors. The summary report in Swedish is also published by the Governmental Commission in their report. SUMMARY Nanotechnology and nanomaterials have been promoted as having the potential to bring benefits to many areas of research, and to positively contribute to sustainable development. As such, this rapidly growing field is increasing attracting investments from governments and businesses worldwide. At the same time, it is recognised that the application of nanomaterials may pose a risk to human health and the environment. The Swedish Government, therefore, released a Committee Directive (Dir. 2012:89) to produce a National Action Plan for the Safe Use and Handling of Nanomaterials. KTH was commissioned by the Governmental Commission, charged with developing this action plan, to review the current state of knowledge on the environmental aspects of nanomaterials from a life cycle perspective. The remit of this study was to: clarify the models best suited to highlight issues related to the safe use of nanomaterials; summarize the results of current life cycle research and difficulties in applying life cycle approaches to nanomaterials; identify ongoing research initiatives; propose priorities to achieve the level of knowledge required to understand risks and opportunities of nanomaterials and nano-products; provide suggestions for images to explain the importance of the life cycle perspective in the field of nanomaterials. There is a general consensus that the potential health and environmental risks of nanomaterial should be evaluated over their entire life cycle. This report reviews the literature on the application of life cycle assessment (LCA), risk assessment (RA) and substance flow analysis (SFA) to nanomaterials and nanoproducts. Whilst there is plenty of literature promoting the application of LCA, there are few studies that apply LCA to the area of nanotechnology. Twenty five LCA studies of nanomaterials were identified, including nanomaterial such as cadmium telluride, calcium carbonate, carbon black, carbon nanofibres, carbon nanotubes, nanoclay, nanoscale platinum-group metals, silica, silver, silicon, titanium and titanium oxide. Product systems studied include: auto-body panels, biopolymers, coatings, electronic displays, electronic sensors, lithium-ion batteries, photo voltaic systems, packaging and agriculture polymer films, nanomaterial production processes, textiles and wind turbine blades. These studies only looked at parts of the life cycle, with no quantitative studies addressing the impact of nanomaterials to human health and the environment from the cradle to the grave. Results from these studies showed the potential for a significant cumulative energy demand in the production of nanomaterials such as carbon nanotubes and carbon nanofibres. However, this is reduced when taking into consideration the small amounts of nanomaterials in products and the potential benefits during the use phase, such as weight reduction. It has been shown that the goal and scope definition is of vital importance to get meaningful results, as the different properties and functions of nanomaterials need to be considered when nano-enabled products are compared to conventional products. The life cycle inventories of current LCA studies cannot be classified as comprehensive as they often lack nanomaterial specific data related to the outputs of processes. Hence, populating life cycle inventory databases with nanomaterial specific information, such as size and shape, is of critical importance. Although the UNEP/SETAC framework for toxic impacts can in principle be used for specific impacts causes by nanoparticles, life cycle impact assessment methods currently lack characterisation factors for the release of nanoparticles indoors and outdoors. Hence, no LCA studies to date have considered the human toxicity and eco-toxicity of nanomaterials from a life cycle perspective with consideration of the nano-specific properties. There is a consensus that the RA framework is applicable to nanomaterials. However, many of the methodological steps with RA require further refinement or development. Although some RAs have been conducted for nanomaterial according to standard RA protocols, studies have concluded that, due to limited data and the presence of large uncertaitintites, it has not been possible to complete full RAs for regulatory decicion making. There is a lack of measured exposure data for nanomaterials, lack of validated exposure estimation models, extensive uncertainties when characterizing nanomaterials and a lack of (eco)toxicological studies in a variety of species. Hence, it is difficult to complete hazard identification, dose–response and exposure assessments for most nanomaterials. Two approached to RA from a life cycle perspective have been identified: ‘LC-based RA’ and ‘RA-complemented LCA’. RAcomplemented LCA combines life cycle and RA based methods and most publications and risk analysis frameworks utilise this method. SFA is occasionally applied prior to RA to estimate emissions, and has become the point of departure for the development of emission assessment methods. SFA traditionally uses mass as a measure of the stocks and flows of materials. Such an approach has been used to study the flow of nanomaterials such as nano titanium dioxide, nano zinc oxide, nanosilver and carbon nanotubes during waste incineration and landfilling of municipal solid waste and construction waste. Furthermore, particle flow analysis has been used to account for the relevant properties of nanomaterials such as particle size, and the processes that change the number of nanoparticles such as agglomeration and disassociation of particles into ions. This approach has been used to study the flows and stocks of nanosilver (in wound dressings, textiles and electronic circuitry) and nano titanium dioxide (in sunscreens, paints and selfcleaning cement) during the use phase. In light of this review of LCA, RA and SFA studies of nanomaterials, the following suggestions identify some potential ways forward:  Improved information concerning the use of engineered nanomaterials (ENMs)s. In order to assess risk, information is needed on the volumes society uses, in which applications, and in what forms.  Improved information on emissions is required in order to assess the risks of ENMs. As a first step, information is required on where emissions occur, which can be achieved through undertaking simplified SFAs of ENMs. Methods for this need to be developed where the reasonable worst-case assumptions can be made to assess whether further detailed analysis is required. Those who place a material on the market should be able to describe how the material will be disposed of or emitted to the environment.  In depth SFA in specific cases. These cases can be selected for several reasons: environmentally relevant ENMs, ENMs used in large quantities or ENMs that can be considered representative of larger groups and thus can be used to develop and verify the simplified models.  Measurements. SFA is based upon existing and available data which in turn need to come from actual measurements or model calculations, which in turn needs to be based on measurements. Examples of important situations where actual measurements are             required include exposure in the work environment, flows in waste water treatment plants and flows associated with recovery processes and other waste management activities. Methods for the characterization of nanoparticles. The properties of nanoparticles can change according to their shape and size. Nanoparticles need to be characterised in ways that are relevant for emission measurements, exposure analysis and toxic effects. Toxicological and eco-toxicological dose-response data are needed. Models for exposure analysis require further development and need to be adapted for nanoparticles. Environmental impact assessment methods in LCA require further development and need to be adapted for nanoparticles. As the methods for risk assessment of nanoparticles are developed, there is a need for LCA methodology to follow and adapt. Life-cycle inventory data for ENMs. LCA is heavily dependent on databases which have been developed over the past decade. However, these databases are limited with regards to ENM data. Life cycle inventory data is essential for the assessment of the potential benefits and impacts of ENMs in a life cycle perspective. Methods to develop life cycle data for emerging technologies. Nanotechnology is a field experiencing rapid development; this also applies to manufacturing processes and their environmental performance. International cooperation with a Swedish perspective. Much of the data and methods that are required for LCA should be developed in the context of international cooperation. However, it may be important to develop life cycle data for products manufactured in Sweden as some conditions may be country specific (for example, raw materials and energy). Furthermore, other processes such as waste management may have specific Swedish conditions. The collaboration of industry, governmental agencies and research. Much of the data which is required should be produced by industry. It is also important that governmental agencies and researcher are involved in such work to ensure credibility and transparency. Credible information to users. The safe use of ENMs and nanoproducts requires informed users. Labelling and other forms information is needed to be designed so that users in businesses, organizations, government agencies and consumers can make their own informed decisions. Avoid locking in a risk paradigm. Full risk assessments require copious amounts of data and take a significant amount of time to complete. It would be expensive and inefficient to complete risk assessments on every ENM and its specific application that is placed on the market. Hence, one must be able to make effective decisions about the safe use of ENMs without full risk assessments. Avoid a ‘material for material’ paradigm. The number of ENMs can be vast. In order to have effective processes, decisions can be taken without the complete data that is require for each individual material. Decisions can be made for groups of materials, or based on more simple criteria. Resources for research in several fields. There is need for research on data and methods that can be used for SFA, RA and LCA. Research is also needed on the use of ENMs, policy instruments and decision theory. SVENSK SAMMANFATTNING 1 INLEDNING .......................................................................................................................... I 2 ANVÄNDNING .................................................................................................................... I 3 LIVSCYKELPERSPEKTIV OCH METODER........................................................................... II 4 EMISSIONER AV NANOPARTIKLAR ............................................................................... VI 5 RISKBEDÖMNINGAR I LIVSCYKELPERSPEKTIV ........................................................... VIII 6 LIVSCYKELANALYSER AV NANOMATERIAL ................................................................ VIII 7 OM VAL AV METODER, BEGRÄNSNINGAR OCH UTVECKLINGSBEHOV ......................X 8 PÅGÅENDE FORSKNING................................................................................................. XII 9 REKOMMENDATIONER ................................................................................................. XIII 10 SLUTSATSER ................................................................................................................... XIV 11 REFERENSER ................................................................................................................... XVI 1 Inledning Nanomaterial kan definieras på olika sätt, men gemensamt för de flesta definitioner är att det handlar om material som innehåller partiklar som i någon dimension har en storlek på mellan 1 och 100 nm (se t.ex. European Commission, 2011). Partiklarna kan förekomma i fasta material, på en fast yta, i en gasfas eller suspenderade i en vätska. Jämfört med traditionella partiklar innebär den lilla storleken att ytan är mycket stor i förhållande till volymen. Nanopartiklar kan också ha annorlunda egenskaper än större partiklar och egenskaperna kan bestämmas av storleken och formen, inte bara av den kemiska sammansättningen. Nanopartiklar kan transporteras på annat sätt än större partiklar eller lösta ämnen. Storleken och formen på nanopartiklarna, och därmed egenskaperna, kan ändras under partiklarnas olika faser i sina livscykler, från tillverkning till slutanvändning och efter emission till naturen. Nanopartiklar kan förekomma naturligt och tillverkas. Nanomaterial har mött ett stort intresse och det finns förväntningar om innovationer inom många områden och stark tillväxt. Samtidigt finns det en oro att nanomaterial också kan vara miljö- och hälsofarliga. En del ämnen i nanomaterial har dokumenterade miljö- och/eller hälsorisker. Dessutom finns en oro att nanomaterial, genom sina speciella egenskaper, lättare kan exponera känsliga organismer och organ. 2 Användning Begreppet nanomaterial är vitt och täcker ett stort antal material och tillämpningar. Det finns ingen samlad offentlig statistik om användning av nanomaterial. I termer av marknadsvolym så är de viktigaste nanomaterialen enligt Europeiska kommissionen (European Commission, 2012) icke-metalliska oorganiska material (såsom kiseloxider, aluminiumoxid och titandioxid), kolbaserade nanomaterial (såsom kimrök (eng: carbon black) och ”kolnanorör”), metaller (t.ex. silver) och organiska makromolekyler och polymera material. Dessutom finns ett stort antal material som är under utveckling eller används i mindre kvantiteter. i Nanomaterial finns i en stor mängd produkter, från vardagliga konsumentvaror till högt specialiserade produkter inom biomedicinsk teknik och IKT (Informations och Kommunikationsteknologi). De största tillämpningarna av nanomaterial är i däck (kimrök, eng: carbon black) och i polymera material (huvudsakligen kiseloxid men också metaller), inom elektronik, kosmetika och biomedicinska tillämpningar (European Commission, 2012). Inom elektronik används nanomaterial bland annat som kiseldioxid vid tillverkning och bariumtitanat som används för kondensatorer. Inom kosmetika används bland andra nanomaterial kiseldioxid, titandioxid och zinkoxid. Inom biomedicin är guld och silver bland de viktigaste nanomaterialen (European Commission, 2012). Dessutom används ett stort antal nanomaterial i bland annat färger och bestrykningsmaterial, katalysatorer, solceller, bränsleceller osv. Användningsområden är som synes flera och det finns unika egenskaper som gör att nya funktioner och produkter kan utvecklas. I termer av kvantiteter av nanomaterial så dominerar ”carbon black” (9,6 miljoner ton per år) och kiseloxid (1,5 miljoner ton) (European Commission, 2012). Andra nanomaterial med signifikanta mängder är aluminiumoxid (200 000 ton), bariumtitanat (20 000 ton), titan dioxid (10 000 ton), ceriumoxid (10 000 ton) och zinkoxid (8 000 ton). Kolnanorör och kolnanofibrer marknadsförs i storleksordningen upp till några tusen ton. Försäljning av nanosilver uppskattas till 20 ton per år. Alla dessa uppgifter kommer från en rapport från Europeiska kommissionen (European Commission, 2012) som i sin tur refererar till rapporter från konsultföretag. Framtidens nanomaterial och dess användning kan förväntas utvecklas i en mängd olika riktningar. Exempel på intressanta områden är inom IKT och för läkemedelsdistribution. I dessa fall kan det handla om väldigt specifika material och tillämpningar. Man kan också tänka sig en utveckling mot mer förnybara nanomaterial exempelvis baserade på cellulosa. Produkter med nanopartiklar på ytan som katalysatorer kan få många tillämpningar. Kompositmaterial där nanofibrer ingår är ytterligare ett område som kan få en bred användning. Ur miljösynpunkt kan man notera att flera av dessa tillämpningar ingår i området miljöteknik, d.v.s. teknik som i ett livscykelperspektiv kan ge mindre miljöpåverkan, t ex i termer av minskade koldioxidutsläpp, än traditionella produkter. Det kan handla om energiteknik, katalysatorer och lättare material. 3 Livscykelperspektiv och metoder För att bedöma miljöpåverkan av produkter, kemikalier och material är ett livscykelperspektiv viktigt. Detta för att undvika att man missar viktiga aspekter eller väljer lösningar som innebär att man flyttar miljöproblem från en livscykelfas till en annan, eller från en plats eller tidsperiod till en annan, eller minskar ett hälso eller miljöproblem samtidigt som man skapar ett nytt. I den här rapporten används orden kemikalie, substans och ämne som synonymer. I kemikalielagstiftningen definieras en vara som ”ett föremål som under produktion får en särskild form, yta eller design, vilket i större utsträckning än dess kemiska sammansättning bestämmer dess funktion” (se t.ex. Kemikalieinspektionen, 2011). Ordet produkt används i den här rapporten såsom det används i samband med livscykelanalyser (se nedan) så att det omfattar både varor, kemiska produkter och tjänster. ii Ett livscykelperspektiv kan användas för både produkter, kemikalier och material. Livscykeln kan dock se lite olika ut beroende på vad det är man studerar. För kemikalier startar livscykeln antingen vid tillverkning, eller om det är ämne som finns naturligt, vid utvinning (Figur 1). Kemikalien kan användas i flera olika produkter. Varje produkt kan sedan genomgå tillverkning, användning, avfallshantering och eventuell återvinning. I varje fas kan utsläpp av kemikalien ske. Figur 1. Livscykelperspektiv för en substans som används i flera olika produkter. För en produkt startar livscykeln vid utvinning av de råvaror som behövs för tillverkning och användning av produkten. I livscykeln ingår sedan tillverkning av produkten och andra varor och tjänster som behövs i produktens livscykel, användning av produkten och avfallshantering (Figur 2). Figur 2. Livscykelperspektiv för en produkt. iii De olika livscykelperspektiven som beskrivs i figur 1 och 2 är också kopplade till olika metoder att bedöma miljö- och hälsoaspekter av olika system. Livscykeln i Figur 1 som berör substanser är kopplad till Substansflödesanalyser (SFA) och Riskbedömningar i livscykelperspektiv. I substansflödesanalyser studerar man en substans från att ämnet uppstår (antingen genom produktion eller utvinning). Sedan följer man flödet av substansen i samhället, hur och var den används och var ämnet emitteras till omgivningen (van der Voet, 2002). Med hjälp av substansflödesanalyser kan man identifiera dataluckor, alltså brist på information om emissioner. Förutom emissioner och dataluckor så kan även sänkor identifieras. Sänkor kan vara permanenta genom att ämnet förstörs eller tillfälliga. Ett exempel på en permanent sänka kan vara förbränning av organiska ämnen då ämnet förstörs. Ett exempel på en tillfällig sänka kan vara deponier där utlakningen av ett ämne kan ske långsamt men ändå vara större än noll. Substansflödesanalyser ger alltså information om emissioner. Däremot behandlas inte toxiska eller ekotoxiska effekter. Riskanalyser och riskbedömningar (eng Risk Assessment, RA) är termer som används i många olika sammanhang med lite olika mening. Riskbedömningar kopplat till kemiska substanser syftar till att bedöma miljö och/eller hälsorisker med ett visst ämne, antingen i en specifik exponeringssituation eller över hela substansens livscykel. Riskbedömningar kopplat till kemiska substanser innehåller både en exponeringsanalys och en effektanalys eller en dosresponsanalys (t.ex. Grieger et al, 2012). I exponeringsanalysen ingår att göra en analys av vilka grupper av människor eller vilka ekosystem som kan bli exponerade för substansen och i vilka halter. I exponeringsanalysen ingår då både uppgifter om emissioner och om spridning och omvandling av ämnet i miljön inklusive arbetsmiljö. I effektanalysen görs en analys av vilka effekter olika halter kan ge upphov till. Sedan vägs resultaten från exponerings- och effektanalysen ihop i en riskbedömning. Riskbedömningar av kemiska ämnen har bland annat reglerats på en Europeisk nivå i samband med REACH-lagstiftningen. Riskbedömningar kan göras i ett livscykelperspektiv, d.v.s. hänsyn tas till emissioner av ämnet i hela dess livscykel. Livscykeln i Figur 2 är kopplad till Livscykelanalyser (eng. life cycle assessment, LCA) som studerar den potentiella miljöpåverkan av en produkt från ”vaggan till graven”. Ordet ”produkt” ska tolkas brett så att det kan innehålla både varor och tjänster. För livscykelanalyser finns en internationell standard (ISO, 2006 a och b). Livscykelanalyser skiljer sig från substansflödesanalyser och riskbedömningar bland annat i att det som studeras inte är ett kemiskt ämne, utan en funktion som en produkt, en tjänst eller ett system uppfyller (Finnveden et al, 2009). En annan skillnad är att det man studerar inte bara är emissioner av ett ämne, utan ett brett spektrum av potentiellt miljöstörande ämnen. Vidare behandlas flera olika typer av miljöeffekter inklusive hälsoeffekter och resursanvändning. Ytterligare en skillnad är att en livscykelanalys studerar potentiell miljöpåverkan snarare än total. Detta beror bland annat på att man i en livscykelanalys (som ju har en produkt som utgångspunkt) bara studerar en mindre del av de totala utsläppen av ett ämne, nämligen den del som hör till den produkt (eller funktion) som man studerar i livscykelanalysen (Hauschild, 2005). I en riskbedömning (som har ett kemiskt ämne som utgångspunkt) kan man däremot inkludera samtliga emissioner av ämnet. Därmed finns möjligheter att uppskatta den totala eller absoluta risken för ämnet. De olika metoderna substansflöden, riskbedömningar i ett livscykelperspektiv och livscykelanalyser skiljer sig alltså åt på flera olika sätt. De studerar olika typer av objekt (SFA och RA studerar substanser och LCA produkter/funktioner). SFA och RA studerar ett ämne i iv taget medan LCA inkluderar flera ämnen och miljöproblem. SFA tittar bara på utsläpp av ett ämne medan RA också studerar risker med dessa ämnen. LCA studerar potentiell miljöpåverkan medan RA kan studera (absolut) miljöpåverkan/risk). Fast de tre metoderna alla kan ha ett livscykelperspektiv, så har de alltså olika syften och svarar på olika frågor. Om man är intresserad av var utsläpp av en kemikalie kan uppstå så är substansflödesanalyser ett bra metodval. Om man är intresserad av risker av ett specifikt ämne så är riskbedömningar det bästa valet. Om man vill studera potentiella för och nackdelar ur ett miljöperspektiv med en specifik produkt så är livscykelanalyser det bästa valet. Eftersom de olika metoderna är gjorda för att svara på olika frågor kan de inte lätt ersätta varandra, utan kompletterar varandra. Man kan också notera att de olika metoderna i viss mån bygger på varandra. Den information om emissioner som är ett resultat av en substansflödesanalys behövs också för att göra riskbedömningar i ett livscykelperspektiv och livscykelanalyser. För att göra riskbedömningar behövs modeller och data för exponeringsanalysen och effektanalysen. Dessa modeller och data kan också efter viss anpassning användas i livscykelanalyser. Riskbedömningar har utvecklats för kemiska ämnen och en viktig fråga är då om de också kan användas för nanomaterial. En viktig aspekt i det sammanhanget är att de toxiska effekterna av nanomaterial inte bara beror på den kemiska sammansättningen av materialet utan också kan bero på nanopartiklarnas storlek och form. Det innebär att när man ska karaktärisera dem så räcker det inte med sammansättningen utan det behövs även annan information. Det innebär också att de toxiska tester som används för kemiska ämnen kan behöva modifieras för nanopartiklar. Ytterligare en aspekt med nanomaterial som är speciell är att det kanske inte är koncentrationen (mätt som massa per volym) som är den mest relevanta parametern när toxiciteten ska bestämmas. Det har också föreslagits att antalet partiklar per volymsenhet eller yta per volymsenhet kan vara relevanta mått för att indikera risker med nanomaterial (Arvidsson, 2012). På motsvarande sätt kan de modeller som används för exponeringsanalysen av kemiska ämnen vara mindre relevanta eftersom nanopartiklar kan transporteras på andra sätt än kemiska substanser som är upplösta. Storleken och formen på nanopartiklarana kan också förändras vilket man kan behöva ta hänsyn till i exponeringsanalysen En slutsats är därför att även om det ramverk som utvecklats för riskbedömningar av kemiska ämnen är relevant också för nanopartiklar, så kan både de toxikologiska testerna och modellerna för exponeringsanalyser behöva modifieras och vidareutvecklas (Grieger et al, 2012, Praetorius et al, 2013). I substansflödesanalyser beskriver man flöden i termer av massa. Eftersom de toxiska egenskaperna hos nanopartiklar ibland kan vara mer relaterade till antalet partiklar och dess form, snarare än massan av partiklarna kan det vara mer relevant att arbeta med partikelflödesanalyser snarare än substansflödesanalyser (Arvidsson et al, 2011). Livscykelanalyser kan användas också för produkter som innehåller nanomaterial (Grieger et al, 2012). I de delar som analyserar potentiella toxiska effekter så får man dock samma problem som i riskbedömningarna, dvs att metoderna kan behöva modifieras och v vidareutvecklas för att fånga de aspekter som är specifika för nanomaterial. 4 Emissioner av nanopartiklar Det finns en bred enighet om att produktion, användning och avfallshantering av nanomaterial också leder till utsläpp. Det finns dock mycket begränsad information om dessa utsläpp (Gottschalk and Nowack, 2011). Utsläpp kan ske i alla faser, från produktion av nanomaterial och de produkter de finns i, till användning och avfallshantering. Utsläpp under produktion av nanomaterial kan ske både till luft och vatten. Sådana utsläpp är relevanta bland annat för att uppskatta risker i arbetsmiljön. Det finns dock begränsat med data. Vid modellering av utsläpp har man därför gjort olika antaganden. Man har antagit emissioner upp till något eller några procent, vid ”worst-case-scenarier” något högre (Gottschalk and Nowack, 2011). Samtidigt är det klart att vid noggrant kontrollerade produktionsprocesser kan utsläppen vara betydligt lägre. På motsvarande sätt kan man tänka sig utsläpp från tillverkning av produkter där nanomaterial ingår. Även här har man ibland antagit utsläpp på någon eller några procent (Gottschalk and Nowack, 2011), men det kan vara lägre vid kontrollerade processer och möjligen högre vid dåliga arbetsförhållanden. För många nanomaterial kan de största riskerna för utsläpp vara i samband med användningsfasen. Arvidsson et al (2011) har exempelvis studerat nanosilver med hjälp av partikelflödesanalyser med fokus på användningsfasen. Detta kan ses som ett intressant exempel på olika spridningsvägar och sänkor för nanomaterial. Nanosilver används framför allt i textilier, i sårförband och elektronik. Emissioner kopplat till användning av textilier uppskattas vara större än från sårförband samtidigt som det enligt Arvidsson et al (2011) är svårt att uppskatta emissionerna kopplat till användning av elektronik. Studier har visat att silverpartiklar kan frigöras vid tvättning av textilier. Hur mycket beror dock bland annat på hur mycket silver som finns i textilierna och det kan variera kraftigt. Silvret som frigörs vid tvättning hamnar i stor utsträckning i vattenreningsanläggningar där en stor del, men inte allt, kan förväntas hamna i slammet och resterande släppas ut med vattnet. Slammet kan sedan användas som täckningsmaterial på deponier eller användas på jord- eller skogsmark. Från slammet kan silvret lakas ut. Om det är på deponier kan frisättningsprocessen vara långsammare och lakvattnet kan fångas i lakvattenreningsprocesser och då eventuellt fastna i reningsverksslam igen. Arvidsson (2012) har analyserat några möjliga framtidsscenarier med en ökad användning av silver i textilier, Man finner då att halterna i slam kan bli högre än riskrelaterade riktvärden och om slammet används på jordbruksmark så kan halterna bli höga om hänsyn tas till risknivåer för maskar (Arvidsson, 2012). Halterna beror dock bland annat på hur mycket textilier med silver som används och också hur mycket silver som används i textilierna (Arvidsson, 2012). Slam kan också förbrännas. Vid förbränning kommer det mesta av silvret att hamna i bottenaskan men även i andra askfraktioner och endast en mindre del kan förväntas emitteras med rökgaserna (Mueller et al, 2013). De olika askorna kommer att deponeras eller eventuellt användas som konstruktionsmaterial. I båda fallen kan man förvänta sig långsamma utlakningsprocesser. vi För nanosilver som sitter i sårförband kommer en mindre del att lämna förbandet under användningen, men det mesta finns kvar i förbandet. Både för textilier och sårförband innebär användningen en direkt exponering av människor eftersom materialen ligger mot huden. Använda förband hamnar till stor del i brännbara avfallsfraktioner. Nanosilver i elektronikprodukter kommer antagligen i stor utsträckning att finnas kvar i produkterna efter användningsfasen. Om elektronikskrotet behandlas genom återvinning finns möjligheter att delar av silvret kan återvinnas. Det är dock väl känt att elektronikskrot inte bara behandlas genom avancerade återvinningsprocesser utan också i viss mån genom enklare och mer miljöfarliga processer i vissa utvecklingsländer (Robinson, 2009, Umair et al, 2013), varvid utsläpp kan ske. För andra nanomaterial kan emissioner ske på andra sätt. Titandioxid används t.ex. i solskyddsmedel. Då sker en direkt exponering av människor, men det sker också direkta emissioner till akvatiska miljöer i samband med att man badar med solskyddsmedel på kroppen (Arvidsson, 2012). Nano titandioxid kan också användas i färg och cement. Från dessa material sker dock emissioner i en långsammare takt. Grafen är ett nytt material som kan tänkas få många olika tillämpningar till exempel inom elektronik och som kompositmaterial. Användningen kan därför förväntas öka. Tillgängliga data indikerat att grafen kan ha miljöfarliga egenskaper (Arvidsson et al, 2013). Det finns dock väldigt lite information om potentiella emissioner av grafen både till yttre miljö och relaterat till arbetsmiljö (Arvidsson et al, 2013). Det är ett exempel på de databrister som finns för många nanomaterial. Dessa exempel illustrerar att utsläpp från nanomaterial kan ske på många olika sätt under livscykeln. Det kan ske både som nanopartiklar och som substanser lösta i vatten eller luft. Det kan ske under produktion, användning och i avfallshantering. Emissioner under användningsfasen kan ske direkt till naturen, t.ex. genom utlakning av fasader eller från kosmetiska produkter, eller via vattenreningsprocesser. Människor kan bli exponerade direkt, till exempel genom hudkontakt, i arbetsmiljön eller efter att utsläpp har skett till naturen. Avfallshantering kan ha en nyckelroll. Metaller kommer inte att förstöras under avfallshanteringen utan flyttas mellan olika former där deponier kan vara sänkor med långsam utlakning. Hastigheten för utlakningen kan dock bero på en mängd olika faktorer som löslighet, sammansättning och nedbrytningshastigheten för omgivande material, partikelstorlek o.s.v. För organiska material kan nanomaterialen destrueras under förbränning. Återvinning av produkter kan ske på många olika sätt och leda till att kretslopp sluts, men om det sker på dåligt kontrollerade sätt kan det leda till diffus spridning av farliga ämnen. Eftersom kompositmaterial, som är sammansatta av många olika material, ofta är svårare att återvinna kan en ökad användning av sådana leda till ökad förbränning och deponering. Att nanosilver valdes som exempel ovan beror dels på att silverjoner har miljöfarliga egenskaper, men också på att det i alla fall finns några studier tillgängliga om nanosilver. Annars är det tydligt att det för många nanomaterial och produkter saknas information, inte bara om användningen, utan också om vilka utsläpp som kan ske under olika livscykelfaser. Det saknas också ofta information om i vilka former emissionerna sker. Det är av betydelse både för exponerings- och effektanalyser om emissionerna av material sker i form av vii nanopartiklar, större partiklar eller om ämnet har lösts ivatten eller förångats. 5 Riskbedömningar i livscykelperspektiv Grieger et al (2012) gör en genomgång av riskbedömningar av nanomaterial. Man konstaterar att det finns ett antal studier som har försökt göra riskbedömningar av nanomaterial enligt gällande protokoll. Studierna behandlar bland annat nanosilver, titandioxid-partiklar och kolbaserade produkter som kolnanorör. Alla dessa har dock dragit slutsatsen att på grund av brist på data och stora osäkerheter har det inte varit möjligt att genomföra kompletta riskbedömningar för dessa nanomaterial. Resultaten måste därför betraktas som preliminära. Bland svårigheterna finns brist på mätta exponeringsdata, modeller för exponeringsanalyser, osäkerheter i karaktäriseringen av nanopartiklarna, tillämpligheten av olika tester, och brist på toxikologiska och ekotoxikologiska data. 6 Livscykelanalyser av nanomaterial I Figur 3 visas en förenklad bild av livscykeln av en produkt som innehåller nanomaterial. Den tunnare, mörka linjen innanför den tjockare representerar nanomaterialet i produktens livscykel och visar var emissioner av dessa kan ske. Vid utvinning av råvaror kan miljöpåverkan uppstå bland annat på grund av energiintensiv råvaruutvinning och associerade utsläpp, förlust av icke-förnybara råvaror och utsläpp av toxiska ämnen. Om det handlar om förnybara råvaror så kan bland annat markanvändningen leda till påverkan på biologisk mångfald. Under produktionsfasen kan miljöpåverkan uppstå på grund av energiintensiva tillverkningsprocesser för nanomaterial och associerade utsläpp och möjligen utsläpp av nanomaterial. Under användningsfasen kan utsläpp av nanomaterial ske, men produkter med nanomaterial kan också bidra till minskad miljöpåverkan jämfört med konventionell teknik. Under avfallshanteringen kan miljöpåverkan uppstå på grund av utsläpp från bland annat förbränning, återvinningsprocesser och deponering, men om material och/eller energi kan återvinnas kan det leda till minskad miljöbelastning då det kan ersätta annan produktion. Under det senaste decenniet har ett antal livscykelanalyser eller livscykelanalysliknande studier gjorts på produkter som innehåller nanomaterial. Både Gavankar et al (2012) och Hischier and Walser (2012) har nyligen publicerat översikter över gjorda studier. Sammanlagt handlar det om ca 30 studier. Många av de publicerade studierna är dock på olika sätt begränsade. En aspekt är att många av studierna är av typen ”vaggan till grind” där alltså produktionsprocesser ingår, men inte användning och avfallshantering. En annan begränsning är att många av studierna fokuserar på ett begränsat antal miljöeffekter, i första hand energianvändning och/eller växthusgaser. Däremot är toxiska effekter sämre behandlade. Båda dessa begränsningar är i stor utsträckning kopplade till brist på data och modeller, både för emissioner, exponeringsanalys och effektanalys av nanomaterial. Ett antagande som ibland görs är att potentiella effekter av nanomaterial kan modelleras som om materialet emitteras som om det har lösts upp i vatten (och därmed inte längre finns som nanopartiklar). viii Figur 3. En förenklad beskrivning av livscykeln hos produkt med nanomaterial och var emissioner kan uppstå. En viktig aspekt vid livscykelanalyser av nanomaterial är energiåtgång och därtill associerad miljöpåverkan vid tillverkning av nanopartiklarna. Denna kan vara svår att uppskatta bland annat därför att det ofta handlar om nya processer under utveckling. Det kan då vara svårt att skala upp från olika typer av pilotanläggningar till fullskaleprocesser för tillverkning av nanomaterial. Det kan också finnas olika tillverkningsprocesser med olika prestanda. Energianvändningen vid tillverkning av nanomaterial kan ibland vara signfikant. Ett exempel är tillverkning av kolnanopartiklar som kolnanorör och fullerener. Dessa är 2 till 100 gånger mer energikrävande per kg att tillverka än t ex aluminium, även med idealiserade produktionsmodeller (Kushnir and Sandén, 2008). Kolnanofibrer kan användas i kompositmaterial exempelvis tillsammans med polymerer och glasfibrer. Även dessa kompositmaterial kan vara mer energikrävande än stål att tillverka (Hischier and Walser, 2012). Betydelsen av livscykelperspektiv blir dock uppenbar om man tar hänsyn till att kompositmaterialen kan ersätta stål i till exempel fordon och då bidra till att dessa får en lägre vikt och därmed lägre bränsleförbrukning i användningsfasen. Sett över en bils livscykel kan då den totala energianvändningen bli lägre med kompositer som innehåller nanomaterial jämfört med stål, trots den högre energianvändningen i samband med tillverkningen (Hischier and Walser, 2012). Noteras kan dock att i den analysen ingick inte avfallsledet vilket skulle kunna påverka resultatet. Inte heller ingick att ett lättare material kanske inte leder till lättare fordon utan att man i stället stoppar in en större motor eller mer elektronik i fordonet, ix så att det i slutändan inte alls blir en minskning av bränsleanvändningen. För att fånga dessa aspekter krävs en bredare och mer fullständig livscykelanalys. En liknande situation kan vara fallet med nya litiumbatterier som innehåller nya typer av nanomaterial. Dessa kan vara mer energikrävande att producera. Men om de används exempelvis i fordon, kan det leda till betydligt större energivinster (Kushnir and Sandén, 2011). Tillgängligheten av litium över tiden kan dock vara en begränsande faktor för litiumbatterier (Vikström et al, 2013) vilket möjligen kan påverka miljöbelastningen från produktionen. Ett annat exempel där produktionen av nanomaterial kan vara av betydelse gäller nanosilver. Walser et al (2011) gjorde en studie av t-tröjor, tillverkade med och utan tillsatser av biocider i form av nanosilver eller triclosan. Studien visade att beroende på produktionsmetoden, kan utsläpp av växthusgaser under produktionen av nanosilverpartiklar vara signifikanta för en ttröjas hela livscykel (inklusive 100 tvättar) (Walser et al, 2011). Studien av Walser et al (2011) är också intressant eftersom den är en av få studier av produkter som innehåller nanomaterial som försöker bedöma även ekotoxiska effekter av dessa i ett livscykelperspektiv. Man fann i denna studie att varken silver- eller triclosanutsläpp från tvättning av t-tröjan stod för de största bidragen till hela produktens potentiella ekotoxiska effekter i akvatisk miljö. Man räknade då med en relativt hög avskiljning av nanosilver i vattenreningsverket. I studien ingick akvatisk ekotoxicitet, däremot inte terrester ekotoxicitet, t.ex. efter användning av silverhaltigt reningsverksslam. Utsläpp från gruvbrytning och gruvavfall från produktion av silver kunde däremot vara signifikanta. Annars uppstod de största ekotoxiska utsläppen i t-tröjans livscykel från produktion av t-tröjan och tvättning (inkluderande produktion och användning av tvättmedel och produktion av el för tvättning) (Walser et al, 2011). 7 Om val av metoder, begränsningar och utvecklingsbehov Enligt diskussionen i kapitel 3 så kan olika metoder användas för att belysa olika typer av frågeställningar. Lite förenklat kan vi dela in frågeställningarna i några huvudtyper och länka dessa till de metoder som beskrivits. En typ av frågor handlar om var i livscykeln de största emissionerna av nanomaterial kan ske. Dessa frågor kan i första hand besvaras av substansflödesanalyser (eller partikelflödesanalyser). Baserat på denna information kan man sedan belysa vilka grupper av människor (anställda eller konsumenter) som löper risk för direkt exponering och till vilka typer av miljöer emissionerna sker (vatten, mark, luft inomhus eller utomhus). Substansflödesanalyser kan också användas för att identifiera brist på data om utsläpp. Baserat på denna information kan man också analysera förändringar av utsläppen från olika typer av åtgärder för att minska risker. För att kunna göra substans- (eller partikelflödesanalyser) krävs kunskap om hur mycket av nanomaterialen som används i samhället och i vilka produkter. Som diskuterades ovan i kapitel 2 så finns det ingen samlad information om användning av olika nanomaterial, vare sig mängder eller i vilka produkter. För att kunna göra substansflödesanalyser behövs också information om utsläpp från x produktion, användning och avfallshanteringen. Dessa kan ofta uttryckas som emissionsfaktorer t.ex. från användningsfasen. För att uppskatta dessa behövs ofta mätningar och/eller beräkningar med vars hjälp emissioner kan uppskattas. Som vi såg i avsnitt 4 saknas idag ofta data även för potentiellt miljöfarliga nanomaterial vilket gör substans- (och partikelflödesanalyser) svåra att utföra. En andra typ av frågor handlar om risker med användning av nanomaterial. Det kan handla om att bedöma hur stora riskerna är med användning av ett specifikt nanomaterial, eller för att kunna bedöma vilka de största riskerna är med ett nanomaterial i ett livscykelperspektiv. För denna typ av frågor är riskbedömningar i ett livscykelperspektiv tänkta att användas. Som framgick ovan i kapitel 5 finns det idag stora svårigheter att göra kompletta riskbedömningar. Det handlar både om brist på data och brist på metoder (Gottschalk and Nowack, 2011, Grieger et al, 2012, Praetorius et al, 2013, Savolainen et al, 2010). Dels handlar det om uppgifter om hur och hur mycket nanomaterial som används och dels om emissioner av nanomaterial i olika exponeringssituationer. Men det behövs också vidareutveckling av metoder för exponeringsanalyser liksom de toxiska och ekotoxiska analyserna. Vid utveckling av exponeringsanalyser är det också viktigt att beakta olika situationer såsom arbetsmiljö, användning av produkter och efter utsläpp till vatten, luft och mark. Svårigheterna att göra riskbedömningar i livscykelperspektiv innebär att även om riskbedömningar utvecklas för att besvara frågor om risker som diskuterades ovan, så är det i praktiken svårt att använda dem för det syftet idag. Svårigheterna att göra riskbedömningar i ett livscykelperspektiv innebär också att även om ett nanomaterial kan leda till allvarliga miljöproblem, så kan det vara svårt att visa det i en riskbedömning. Detta illustrerar att för att få en säker användning av nanomaterial så kan man inte bara förlita sig på riskbedömningar som beslutsunderlag. Det kräver mycket data och tar tid. Ett riskparadigm där ett stort antal nanomaterial ska genomgå riskbedömningar blir därför dyrt och ineffektivt. Det behöver också utvecklas andra metoder som kan användas som beslutsunderlag, metoder som kan använda data och metoder som är mer lättillgängliga. Förutom att utveckla metoder och data för riskbedömningar så behöver det därför också utvecklas metoder som kan användas i stället för riskbedömningar vid reglering av nanomaterial. En parallell kan göras till kemikaliområdet där Miljömålsberedningen föreslog att ska kunna behandla och pröva grupper av ämnen utan att varje enskilt ämne genomgick en riskbedömning (Miljömålsberedningen, 2012). Det borde också kunna vara intressant även för nanomaterial. En tredje grupp av frågor handlar om att identifiera potentiella miljöproblem i ett livscykelperspektiv. För dessa frågor kan både substansflödesanalyser, riskbedömningar och livscykelanalyser vara användbara, under förutsättning att data och metoder finns tillgängliga. Man kan dock notera att de olika metoderna har olika ansatser och därmed möjlighet att identifiera olika typer av miljöproblem. Med substansflödesanalyser kan viktiga emissionspunkter i nanopartiklarnas livscykel identifieras. Med riskbedömningar kan viktiga risker i hanteringen av nanomaterial bedömas. Livscykelanalyser kan även bidra till att identifiera andra miljöeffekter än de som förknippas med utsläpp av det specifika nanomaterialet. Livscykelanalyser kan också utnyttjas för att identifiera de faser i livscykeln där viktiga naturresurser används, där utsläpp av växthusgaser sker, och där andra potentiella miljöproblem kan uppstå. xi Nanomaterial kan öppna upp nya möjligheter bland annat genom att man kan producera nya och lättare material, och genom användning i energi och miljötekniska sammanhang. Men användning av produkter med nanomaterial kan också leda till andra typer av miljöproblem. En fjärde grupp av frågor kan därför vara att identifiera potentiella för- och nackdelar med olika alternativa produkter, med eller utan nanomaterial, med avseende på olika miljöfrågor i ett livscykelperspektiv. För denna typ av frågor kan livscykelanalyser användas för att visa på möjligheter men också begränsningar med nanomaterial. I dagsläget begränsas möjligheten att fullt använda livscykelanalyser på produkter innehållande nanomaterial av brist på data för nanomaterialen (Gavankar et al, 2012, Hischier and Walser, 2012). Detta gäller till exempel data för energianvändning och emissioner från produktion av nanomaterial. Det gäller också brist på data om emissioner av nanomaterial som diskuterades ovan i samband med SFA och riskbedömningar liksom metoder och data för miljöpåverkansbedömningen av nanomaterial. Även för avfallshanteringen finns ofta brist på data hur produkter med nanomaterial kan hanteras och vad som händer med olika nanomaterial vid exempelvis deponerings och återvinningsprocesser. Bland annat dessa metodbegränsningar innebär dock att det sällan eller aldrig är möjligt att dra bestämda slutsatser om vilken produkt som är föredra. Rent generellt kan livscykelanalyser sällan besvara frågor av typen ”Är produkt A bättre än produkt B ur miljösynpunkt?” och detta gäller även produkter med nanomaterial. Mer specifika frågor av typen, ”kan den ökade energianvändningen under produktionsfasen av produkt A uppvägas av energibesparingar i användningsfasen jämfört med produkt B? är ofta mer lämpade för en LCA (Finnveden, 2000). En jämförelse mellan de två studierna av Arvidsson et al (2011) och Walser et al (2011) som diskuterades i kapitel 4 och 6 och som båda berör nanosilver illustrerar hur SFA och LCA kan användas, vilka typer av resultat man kan få och hur de kan komplettera varandra. Studien av Arvidsson visar bland annat hur ackumulering i jordbruksmark kan vara problematisk, en aspekt som är svår att fånga med en LCA som inte ser hela användningen av ett ämne utan bara den del som är associerad med en produkt, i studien av Walser et al (2011) en t-tröja. En LCA kan å andra sidan fånga upp andra aspekter såsom utsläpp av andra toxiska ämnen och andra miljöproblem, exempelvis klimatpåverkan från produktion av nanosilver, vilka inte alls studeras av en SFA. 8 Pågående forskning Utvecklingen av nanomaterial är snabb. Det har också skett en utveckling av metoder och data för olika typer av miljö- och riskbedömningar inom området. Det är signifikativt att de flesta referenser till denna rapport är från de allra senaste åren. Inte minst genom flera EUprojekt har kunskapsområdet utvecklats. Flera av de studier som redovisas kortfattat här är från pågående eller nyss avslutade EU-projekt. Exempel på sådana EU-projekt av relevans för denna rapport är Nanosustain och Prosuite som båda innehåller metodikutveckling för LCA och riskanalyser av nanomaterial. Projektet Nanopolytox berör några grupper av nanomaterial som används i polymerer och Nanohouse berör ytbehandlingsprodukter med nanomaterial. Projektet Nanovalid berör metoder för riskbedömningar och Licara metoder för LCA. Mer information om dessa och andra projekt finns i Lazarevic and Finnveden (2013) och på hemsidor med i de flesta fall adressen www.projektnamn.eu. Inom Sverige kan bland annat nämnas att Mistra kommer att starta ett forskningsprogram xii om nanomaterial under 2013. Inriktningen på detta är dock inte bestämt när detta skrivs. Genom de forskningsinsatser som nu pågår kommer kunskapsläget att förbättras under de närmaste åren. Man kan dock förvänta sig att många av de kunskapsluckor och utvecklingsbehov som nämnts ovan kommer att bestå. Detta bland annat därför att flera av forskningsprogrammen i första hand berör ett begränsat antal nanomaterial samtidigt som antalet nanomaterial är stort och det dessutom utvecklas nya. Även tillämpningar inom t.ex. biomedicin och IKT kan förväntas växa. Den forskning kring risker som pågår ger också nya frågeställningar. Det finns därför ett starkt behov av forskning och kunskapsuppbyggnad inom området livscykelaspekter av nanomaterial. 9 Rekommendationer För att kunna säkerställa en säker hantering av nanomaterial och för att kunna identifiera möjligheter i ett livscykelperspektiv krävs bättre data och analysmetoder. Nedan identifieras några vägar framåt:          Bättre information om användning av nanomaterial. För att kunna bedöma risker behövs information om vilka mängder samhället använder, i vilka tillämpningar och i vilka former. Bättre information om emissioner. För att kunna bedöma risker behövs information om var emissioner sker. Förenklade substansflödesanalyser behöver därför utföras på nanomaterial. Metoder för detta behöver tas fram där rimliga worst-case-antaganden kan göras för att bedöma om fördjupade analyser behöver göras. Den som sätter ett material på marknaden bör kunna beskriva hur den kommer att destrueras alternativt emitteras till naturen. Fördjupade substansflödesanalyser i vissa fall. Dessa fall kan väljas av flera olika skäl: miljömässigt problematiska nanomaterial, nanomaterial som används i stora mängder, eller nanomaterial som kan anses vara representativa för större grupper och därmed kan användas för att utveckla och verifiera de förenklade modellerna. Mätningar. Substansflödesanalyser bygger på att det finns data tillgängliga som i sin tur behöver komma från faktiska mätningar eller modellberäkningar, som i sin tur behöver bygga på mätningar. Exempel på viktiga situationer där faktiska mätningar behövs är för exponering i arbetsmiljö, exponering av konsumenter, flöden i vattenreningsanläggningar, flöden i samband med återvinningsprocesser och annan avfallshantering. Metoder för karaktärisering av nanopartiklar. Eftersom egenskaper hos nanomaterial kan förändras med partiklarnas form och storlek behöver de karaktäriseras på sätt som är relevanta för emissionsmätningar, exponeringsanalyser och toxiska effekter. Toxiska och ekotoxiska dos-responsdata behöver tas fram. Modeller för exponeringsanalyser behöver vidareutvecklas och anpassas för nanopartiklar. Metoder för miljöpåverkansbedömningen i livscykelanalyser behöver vidareutvecklas och anpassas för nanopartiklar. I takt med att metoder för riskbedömningar av nanopartiklar utvecklas behöver metodiken för livscykelanalyser följa efter och anpassas. Livscykelanalysdata för nanomaterial. Livscykelanalyser är starkt beroende av databaser och dessa har utvecklats under det senaste decenniet för traditionella xiii        material och tillverkningsprocesser. Det finns dock stora brister avseende nanomaterial. Livscykeldata är nödvändiga för att kunna bedöma potentiella för- och nackdelar med nanomaterial i livscykelperspektiv. Metoder att ta fram livscykeldata för nya teknologier. Nanoteknologi är ett område under stark utveckling. Det gäller även produktionsprocesser och miljöprestanda för dessa. Internationell samverkan, men med svenskt perspektiv. Mycket av de data och metoder som behöver tas fram bör ske i internationell samverkan. Det kan dock vara viktigt att ta fram livscykeldata för produkter som tillverkas i Sverige eftersom en del förhållanden kan vara specifika för Sverige (t.ex. råvaror och energimix). Också andra processer som till exempel avfallshantering kan ha specifika svenska förhållanden. Samverkan industri, myndigheter och forskning. Mycket av de data som behöver tas fram, behöver komma från industrin som har kunskap om tillverkningsprocesser etc. Utveckling av metoder behöver dock ske i samverkan med forskning och myndigheter. Trovärdig information till användare. En säker användning förutsätter informerade användare. Märkning och annan information behöver utformas så att användare i företag, organisationer, myndigheter och konsumenter kan fatta egna beslut. Undvik fastlåsning i ett riskparadigm. Fullständiga riskbedömningar kräver mycket data och tar tid. Det är dyrt och ineffektivt om det ska genomföras på ett stort antal nanomaterial. Man måste därför kunna fatta effektiva beslut om säker användning av nanomaterial utan fullständiga riskbedömningar. Undvik ett material för material-paradigm. Antalet nanomaterial kan vara stort. För att få effektiva processer måste beslut kunna fattas utan att fullständiga data finns tillgängligt för varje enskilt material. Beslut behöver kunna fattas för grupper av material, eller baserat på enklare kriterier. Resurser till forskning inom flera områden. Enligt ovan behöver det tas fram metoder och data inom flera områden för att kunna utveckla användningen av substansflödesanalyser, riskbedömningar och livscykelanalyser. För att detta ska ske behöver det också finnas resurser till forskning. 10 Slutsatser Det finns ett stort antal nanomaterial som används i ett stort antal produkter, exempelvis icke-metalliska oorganiska material (såsom kiseloxider, aluminiumoxid och titandioxid), kolbaserade nanomaterial (såsom kimrök (eng: carbon black) och ”kolnanorör”), metaller (t.ex. silver) och organiska makromolekyler och polymera material. Bland tillämpningar av nanomaterial finns i däck och i polymera material, inom elektronik, och kosmetika. Det finns också många specialiserade tillämpningar inom energiteknik, informations- och kommunikationsteknik och biomedicinska tillämpningar. Det finns ingen offentlig statistik tillgänglig om vilka material som används och i vilka produkter. Miljö- och hälsorisker är både förknippade med den kemiska sammansättningen av materialen, men också nanopartiklarnas storlek, form och egenskaper. Nanopartiklarna behöver därför klassas inte bara med avseende på kemisk sammansättning. Storlek och form är dock egenskaper som kan förändras under användning och efter utsläpp vilket försvårar bedömningar av miljö och hälsorisker. xiv För att bedöma miljöpåverkan av nanomaterial i livscykelperspektiv finns det tre huvudgrupper av metoder: substansflödesanalyser med vars hjälp utsläpp av nanomaterial över dess livscykel kan analyseras, riskbedömningar i livscykelperspektiv med vars hjälp risker för människor och miljö med användning av nanomaterial kan bedömas och livscykelanalyser med vars hjälp potentiell miljöpåverkan av en produkt som innehåller nanomaterial kan analyseras liksom potentiella för- och nackdelar med olika produkter med och utan nanomaterial. För alla dessa typer av metoder finns det exempel på mer eller mindre fullständiga fallstudier. För susbtansflödesanalyser krävs data om användning och emissioner av nanomaterial. Enstaka fallstudier, t.ex. för silver, pekar på att det finns risker att riskrelaterade riktvärden överskrids. Det finns dock stora kunskapsluckor avseende emissioner av nanomaterial under produktion, användning och avfallshantering även för ämnen med miljöfarliga egenskaper. För riskbedömningar behöver modeller för spridning och exponeringsanalyser för nanomaterial utvecklas, liksom dos-respons data för toxiska effekter. Fullständiga riskbedömningar i ett livscykelperspektiv av nanomaterial är svåra på grund av de brister i data och metoder som finns i dagsläget. Livscykelanalyser har gjorts på ett antal produkter innehållande nanomaterial. De är dock ofta begränsade i det att endast ett mindre antal miljöeffekter behandlas och/eller att bara delar av livscykeln analyserats. Produktion av nanomaterial kan ofta vara energikrävande. Dock kan i ett livscykelperspektiv användningen av nanomaterial leda till en minskad energianvändning som är större än den som orsakades av produktionen. Användningen av nanomaterial kan därför innebära viktiga möjligheter. Utvecklingen av nanomaterial är snabb. Det har också skett en utveckling av metoder och data de senare åren och det pågår bland annat flera EU-projekt vilket kommer att förbättra kunskapsläget. Många av de databrister och forskningsbehov som identifierats kommer dock sannolikt att finnas kvar. För att både kunna nå en säker användning av nanomaterial och att kunna utnyttja nanomaterialens möjligheter i ett livscykelperspektiv krävs bättre data och analysmetoder. Som exempel på vägar framåt kan nämnas bättre information om användning av nanomaterial, bättre information om emissioner av nanomaterial, fördjupade substansflödesanalyser av intressanta nanomaterial, mätningar i viktiga miljöer inklusive arbetsmiljöer och exponering av konsumenter, utveckling av metoder för karaktärisering av nanopartiklar, utveckling av modeller för exponeringsanalyser, framtagning av toxiska och ekotoxiska dos-responsdata, utveckling av livscykelanalysdata för nanomaterial, internationell samverkan med svenskt perspektiv, information till användare, utveckling av metoder som kan komplettera riskbedömningar och resurser till forskning. xv 11 Referenser Arvidsson, R. 2012. Contributions to Emission, Exposure and Risk Assessment of Nanomaterials. PhD Thesis, Chalmers University of Technology. Arvidsson, R., Molander, S., Sandén, B.A., 2011. Impacts of a Silver Coated Future: Particle Flow Analysis of Silver nanoparticles. Journal of Industrial Ecology 15, 844–854 Arvidsson, R., Molander, S., Sandén, B.A., 2013. Review of Potential Environmental and Health Risks of the Nanomaterial Graphene. Human and Ecological Risk Assessment, 19: 873– 887. European Commission, 2011. Commission Recommendation of 18 October 2011 on the defition of nanomaterial (2011/696/EU). Official Journal of the European Union L275/38L2 European Commission, 2012. Commission staff working paper. Types and uses of nanomaterials, including safety aspects SWD(2012) 288 final. Brussels. Finnveden, G., 2000. On the Limitations of Life Cycle Assessment and Environmental Systems Analysis Tools in General. Int. J. LCA, 5, 229-238. Finnveden, G., Hauschild, M.Z., Ekvall, T., Guinée, J., Heijungs, R., Hellweg, S., Koehler, A., Pennington, D., Suh, S., 2009. Recent developments in life cycle assessment. Journal of Environmental Management 91, 1–21. Gavankar, S., Suh, S., Keller, A.F., 2012. Life cycle assessment at nanoscale: review and recommendations. The International Journal of Life Cycle Assessment 17, 295–303. Gottschalk, F., Nowack, B., 2011. The release of engineered nanomaterials to the environment. Journal of Environmental Monitoring 13, 1145–1155. Grieger, K.D., Laurent, A., Miseljic, M., Christensen, F., Baun, A., Olsen, S.I., 2012. Analysis of current research addressing complementary use of life-cycle assessment and risk assessment for engineered nanomaterials: Have lessons been learned from previous experience with chemicals? Journal of Nanoparticle Research 14, 1– 23. Hauschild, M.Z., 2005. Assessing environmental impacts in a life cycle perspective. Environmental Science & Technology, 39, 81A-88A. Hischier, R., Walser, T., 2012. Life cycle assessment of engineered nanomaterials: State of the art and strategies to overcome existing gaps. Science of The Total Environment 425, 271–282. ISO, 2006a. ISO 14040 International Standard. In: Environmental Management – Life Cycle Assessment – Principles and Framework. International Organisation for Standardization, Geneva, Switzerland. ISO, 2006b. ISO 14044 International Standard. In: Environmental Management – Life Cycle Assessment – Requirements and Guidelines. International Organisation for Standardization, Geneva, Switzerland. Kemikalieinspektionen (2011): Kemikalier I varor. Strategier och styrmedel för att misnka riskerna med farliga ämnen I vardagen. Rapport nr 3/11. Kemikalieinspektionen. Kushnir, D., Sandén, B.A., 2008. Energy requirements of carbon nanoparticle production. Journal of Industrial Ecology 12, 360–375. Lazarevic, D. and Finnveden, G. 2013. Environmental aspects of nanomaterials in a life-cycle perspective. Avdelningen för miljöstrategisk analys, KTH, Stockholm. Under xvi bearbetning. www.kth.se/abe/fms Miljömålsberedningen, 2012. Minska riskerna med farliga ämnen. SOU 2012:38, Mueller, N.C., Buha, J., Wang, J., Ulrich, A., Nowack, B., 2013. Modeling the flows of engineered nanomaterials during waste handling. Environmental Science: Processes & Impacts 15, 251–259. Praetorius, A., Arvidsson, R., Molander, S., Scheringer, M. (2013): Facing complexity through informed simplifications: A research agenda for aquatic exposure assessment of nanoparticles. Environmental Sciences: Processes and Impacts, 15, 161-168. Robinson, B.H. (2009): E-waste: An assessment of global production and environmental impacts. Science of the Total Environment, 408, 183-191. Savolainen, K., Alenius, H., Norrpa, H., Pylkkänen, L., Tuomi, T. and Kasper, G. (2010) : Risk assessment of engineered nanomaterials and nanotechnologies – A review. Toxicology, 269, 92-104. Umair, S., Björklund, A. and Ekener Petersen, E. (2013): Social Life Cycle Inventory and Impact Assessment of Informal Recycling of Electronic ICT Waste in Pakistan. In Hilty, L., Aebischer, E., Andersson, G., Lohmann, W.: Proceedings of the First International Conference on Information and Communication Technologies for Sustainability ETH Zurich, February 14-16, 2013, 52-58. ETH Zurich, University of Zurich and Empa, Swiss Federal Laboratories for Materials Science and Technology Van der Voet, E., 2002. Substance flow analysis methodology, in: Ayres, R.U., Ayres, L.W. (Eds), A Handbook of Industrial Ecology. Edward Elgar, Cheltenham, UK. Walser, T., Demou, E., Lang, D.J., Hellweg, S., 2011. Prospective environmental life cycle assessment of nanosilver T-shirts. Environmental Science & Technology 45, 4570–4578. Vikström, H., Davidsson, S and Hööl, M., 2013. Lithium availability and future production outlooks. Applied energy, 110, 252-266. xvii CONTENTS ABBREVIATIONS ........................................................................................................................................ II LIST OF FIGURES ....................................................................................................................................... III LIST OF TABLES ......................................................................................................................................... III 1 INTRODUCTION ................................................................................................................................... 1 1.1 Background ...............................................................................................................................................................1 1.2 Overview of this study ..........................................................................................................................................1 2 NANOMATERIALS AND NANOTECHNOLOGY ................................................................................. 3 2.1 Defining nanomaterials and nanotechnology.............................................................................................3 2.1.1 Nanomaterials ...............................................................................................................................................3 2.1.2 Nanotechnology ...........................................................................................................................................4 2.2 Nanomaterials ..........................................................................................................................................................5 2.3 Applications ..............................................................................................................................................................8 2.1 Potential to contribute to sustainability and unintended consequences ........................................8 3 METHODS FOR INVESTIGATING NANOMATERIALS FROM LIFE CYCLE PERSPECTIVE ........... 11 3.1 Life cycle assessment ......................................................................................................................................... 12 3.2 Risk assessment .................................................................................................................................................... 15 3.3 Substance and particle flow analysis ........................................................................................................... 17 4 LIFE CYCLE ASSESSMENT OF NANOMATERIALS ........................................................................... 18 4.1 Research applying life cycle assessment to nanomaterials ................................................................ 18 4.1.1 Current research in the EU and OECD ............................................................................................... 18 4.1.2 Literature review ........................................................................................................................................ 19 4.1.3 Meta-analyses ............................................................................................................................................. 21 4.1.4 Results from selected case studies ..................................................................................................... 26 4.2 Obstacles to and strategies for the application of LCA to nanomaterials .................................... 32 4.2.1 Goal and scope definition ...................................................................................................................... 32 4.2.2 Life cycle inventory analysis (LCI) ........................................................................................................ 34 4.2.3 Life cycle impact assessment (LCIA) ................................................................................................... 37 5 RISK ASSESSMENT OF NANOMATERIALS ...................................................................................... 41 5.1 Risk assessment .................................................................................................................................................... 41 5.2 The complementarity of RA and LCA .......................................................................................................... 41 5.2.1 Life cycle based risk assessment ......................................................................................................... 42 5.2.2 Risk assessment complemented by life cycle assessment ........................................................ 42 5.2.3 A Stream lined approach ........................................................................................................................ 42 6 SUBSTANCE FLOW ANALYSIS OF NANOMATERIALS................................................................... 44 6.1 Research applying SFA and PFA to nanomaterials ................................................................................ 44 6.1.1 Literature review ........................................................................................................................................ 44 6.1.2 Results from selected case studies ..................................................................................................... 46 6.2 Potential life cycle release and exposure of nanomaterials ................................................................ 49 6.2.1 Production of nanomaterials and manufacture of nanoproducts ......................................... 49 6.2.2 Use phase ..................................................................................................................................................... 50 6.2.3 End-of-life phase ....................................................................................................................................... 51 7 COMMUNICATION OF A LIFE CYCLE APPROACH TO NANOMATERIALS ................................. 53 8 RECOMENDATIONS ........................................................................................................................... 54 9 CONCLUSIONS .................................................................................................................................... 56 10 REFERENCES ........................................................................................................................................ 57 APPENDIX A ............................................................................................................................................. 65 APPENDIX B ............................................................................................................................................. 69 i ABBREVIATIONS CED CNF CNT EC ENM ESA EU GWP ISO LCA LCC LCI LCIA MFA MWCNT OECD NP SFA SWCNT PFA UFP Cumulative Energy Demand Carbon Nanofibre Carbon Nanotubes European Commission Engineered Nanomaterial Environmental Systems Analysis European Union Global Warming Potential International Organisation for Standardisation Life Cycle Assessment Life Cycle Costing Life Cycle Inventory Life Cycle Impact Assessment Material Flow Analysis Multi-wall Carbon Nanotube Organisation for Economic Co-operation and Development Nanoparticle Substance Flow Analysis Single-wall Carbon Nanotube Particle Flow Analysis Ultra-Fine Particle ii LIST OF FIGURES Figure 1: Categorisation framework for nanomaterials ............................................................................ 5 Figure 2: Environmental systems analysis tools and their focus ......................................................... 11 Figure 3: Life cycle of a product system......................................................................................................... 12 Figure 4: Cycles for ENMs determined by the life cycles of nanoproducts ..................................... 14 Figure 5: Comparison of how RA and LCA perceive the term ‘life cycle’ ......................................... 16 Figure 6: Substance flow analysis model ...................................................................................................... 17 Figure 7: LCIA midpoint indicators for vapour-grown carbon nanotubes compared to aluminium, steel and polypropylene .............................................................................................................. 27 Figure 8: CED of polymer nanocomposites that provide equal stiffness to a steel component ....................................................................................................................................................................................... 28 Figure 9: Difference in CED of CNF reinforced PNCs compared to steel ......................................... 29 Figure 10: Cradle-to-grave climate footprint of biocidal T-shirts and a regular T-shirt ............. 30 Figure 11: Comparison of the freshwater toxicity for the life cycle of one T-shirt ....................... 31 Figure 12: Flow chart recommending the nanomaterial assessment path depending on the availability of data ................................................................................................................................................. 39 Figure 13: Flows of ENMs during waste disposal shown as a % of the total flow that enters the incineration/landfill system ............................................................................................................................... 48 Figure 14: Life cycle thinking and nanomaterials ....................................................................................... 53 LIST OF TABLES Table 1: Nanomaterials on the EU market ..................................................................................................... 7 Table 2: Applications of nanomaterials in different markets .................................................................. 9 Table 3: Environmental advantages in products for different nanotechnology sectors ............ 10 Table 4: Potential benefits and impacts of the use of nanomaterials ............................................... 10 Table 5: LCA studies of nanomaterials . ......................................................................................................... 19 Table 6: Summary of studies applying LCA studies to ENMs ................................................................ 23 Table 7: Cradle-to-gate energy requirement for various methods of CNT and CNF synthesis ....................................................................................................................................................................................... 25 Table 8: Nanomaterials with a high potential for future industrial applications ........................... 37 Table 9: Outcomes, strengths and weaknesses of LCA and RA ........................................................... 41 Table 10: Proposed Stepwise approach to LCT combined with RA ................................................... 43 Table 11: Case studies applying substance flow analysis to nanomateirals ................................... 45 Table 12: Current in-flow, stocks and emissions during the use phase for nanosilver applications in wound dressings, textiles, and electronic circuitry ..................................................... 46 Table 13: Current in-flow, stocks and emissions during the use phase for titanium dioxide nanoparticle applications in paint, sunscreen, and self-cleaning cement ....................................... 47 Table 14: Recovery of ENMs from PA and PP composites ..................................................................... 51 iii 1 INTRODUCTION 1.1 Background Nanotechnology and nanomaterials are increasingly seen for their potential to provide benefits to many areas of society. Consequentially, current and potential applications of nanomaterials are attracting increasing investments from businesses and governments worldwide (Royal Society 2004). Although still an emerging technology, nanotechnology has been labelled a key enabling technology, and applicable in almost all technological sectors (European Commission 2009a, 2004). There are high expectations as to the positive contribution nanotechnology can make to sustainable development. It has been suggested that nanotechnology has the potential to play a key role in addressing the UN’s Millennium Devilment Goals (Salamanca-Buentello et al., 2005; UNESCO, 2006), and it may increase environmental sustainability via energy technologies, water technologies, chemistry and green chemistry (Fleischer and Grunwald 2008). However, the history of technology shows the potentially harmful unintended consequences of technologies (Tenner 2001) (e.g., dichlorodiphenyltrichloroethane (DDT), chlorofluorocarbons and asbestos). As such, Maynard (2011, 31) suggests “It makes sense to assume that nanomaterials come with unanticipated risks”. Consequentially, as the pace of nanomaterial research, development and production has increased, so has the concern of the potential risk to health and the environment caused by the ubiquity of these materials. Although the development of methods to measure and test nanomaterials has progressed significantly, there remains significant knowledge and data gaps. This results in increased uncertainty when assessing the potential risk of nanomaterials throughout their life cycle. 1.2 Overview of this study In September 2012, the Swedish government released a Committee Directive to produce "A national action plan for the safe use and handling of nanomaterials" (Dir. 2012:89). The purpose of this action plan is that “... Sweden, in various ways, should exploit nanomaterial’s possibilities to meet economic, medical, technical and environmental challenges, whilst taking into account their health and environmental risks and their minimization” (Miljödepartementet 2012, 1). The intention of this action plan is to ensure that knowledge concerning nanomaterials being developed, coordinated and disseminated. The Committee Directive highlights the importance of a life cycle perspective, stating under the heading “A life cycle perspective”: “Faced with the government's standpoint on issues of importance to the development and use of nanomaterials at a national and international level, the availability of a comprehensive and broad basis that takes into account both possibilities with nanomaterials and their health and environmental risk from a lifecycle perspective is very valuable. An important starting point for such a health and environmental risk analysis must therefore be to review nanomaterials from a lifecycle perspective, which also includes the disposal and recycling of products containing nanomaterials. The 1 investigator will, if necessary, suggest measures that give the government a good basis for taking a position." (Miljödepartementet 2012, 6) In this context, this study has reviewed the current state of knowledge on the environmental aspects of nanomaterials in a life cycle perspective. The remit of this study was to: - clarify the types of models and methods that would be best suited to highlight issues relate to the safe use, safe to both human health and the environment, of nanomaterials from a life cycle perspective; - summarize the results of current life cycle research and difficulties, such as knowledge gaps and the lack of information sources specific to nanomaterials; - identify on-going research and other initiatives in Sweden, the European Union and internationally, which focus on the development of methodologies and data collection in order to illustrate the potential life cycle impacts of nanomaterials; - propose priorities, from a Swedish perspective, on what can be done with the current state of knowledge, and work which should be given priority in the short and long term, for Sweden to achieve the level of knowledge required to understand risks and opportunities of nanomaterials and nano-products; - provide suggestions for images to pedagogical explain the importance of the life cycle perspective in the Government's continuing work in the field of nanomaterials. 2 2 NANOMATERIALS AND NANOTECHNOLOGY 2.1 Defining nanomaterials and nanotechnology 2.1.1 Nanomaterials A nanometre (nm) is one billionth of a metre. To place this in context, a human hair is approximately 80,000 nm in width, a red blood cell is approximately 7,000 nm wide, and a water molecule is almost 0.3 nm across (Royal Society 2004). Broadly speaking, the term ‘nanomaterial’ refers to material with internal structures and/or external dimensions within the nanoscale (Lövestam et al., 2010, p. 6). The nanoscale has been reported to be between 1-100nm (ISO 2008; British Standards Institution 2007), 0.1100nm (Royal Society 2004), less than 100nm (O’Brien and Cummins 2010) or less than 500 nm (Handy et al., 2008). Lövestam et al. (2010) note that there is a general consensus that the definition of nanotechnology term should be pursued at a European or Global level. Hence, various international organisations and committees such as the International Organization for Standardization (ISO), the Organisation for Economic Co-operation and Development (OECD), the EU Scientific Committee on Emerging and Newly Identified Health Risks (SCENIHR), the EU Scientific Committee on Consumer Products (SCCP), and governmental institutions at the national level, have proposed definitions of nanomaterials (see Lövestam et al. (2010) for a detailed summary of these definitions). Nevertheless, the definition of nanomaterials have been the subject of intensive debate during recent years, as the term has obvious implications for regulation and policy1 (Lövestam et al., 2010). The definition of the term ‘nanomaterial’ as given by the ISO and the European Commission are presented below. International Organization for Standardization ISO (TS 80004-1) proposes the following definition for the term nanomaterial: “Material with any external dimension in the nanoscale or having internal structure or surface structure in the nanoscale. Note: This generic term is inclusive of nano-object and nanostructured material”2 European Commission The European Commission (EC) has adopted the following recommendation for the regulatory definition for nanomaterials which are set out in articles 2-4 of the Commission Recommendation of 18 October 2011 on the definition of nanomaterial (2011/696/EU): “ ‘Nanomaterial’ means a natural, incidental or manufactured material containing particles, in an unbound state or as an aggregate or as an agglomerate and where, for 1 For instance, the EU chemicals legislation REACH (Registration, Evaluation, Authorisation and Restriction of Chemicals) applies to chemical ‘substances’ on their own, in mixtures or in articles. Although REACH does not specifically refer to nanomaterials, REACH addresses chemical substances in any size, shape of physical form. Hence, the definition of a substance in REACH means that substances at the nanoscale are covered by REACH and its provisions apply to nanomaterials. 2 Nanoscale is referred to as “Size range from approximate 1nm to 100nm 3 50 % or more of the particles in the number size distribution, one or more external dimensions is in the size range 1 nm-100 nm. In specific cases and where warranted by concerns for the environment, health, safety or competitiveness the number size distribution threshold of 50 % may be replaced by a threshold between 1 and 50 %.” (European Commission 2011) This definition is based on scientific advice from the Scientific Committee on Emerging and Newly Identified Health Risks (SCENIHR 2010) and the Joint Research Centre (Lövestam et al., 2010). This report uses the term engineered nanomaterial (ENM), which is commonly defined as materials designed and produced to have structural features with at least one dimension of 100 nanometres or less (Oberdörster et al. 2005). 2.1.2 Nanotechnology Compared to the definition of the term ‘nanomaterial’, less focus has been placed on the definition of the term ‘nanotechnology’. Lövestam et al. (2010) note that this is due to the term only being of occasional practical use. Nanotechnology is a broad term which encompasses all nanoscale science, research, engineering and technology (Lloyd 2004). The European Commission (2004, 4) suggests that “Conceptually, nanotechnology refers to science and technology at the nanoscale of atoms and molecules, and to the scientific principles and new properties that can be understood and mastered when operating in this domain”. International Organization for Standardization ISO proposes the following definition of nanotechnology: “the application of scientific knowledge to manipulate and control matter in the nanoscale to make use of size- and structure- dependent properties and phenomena distinct from those associated with individual atoms or molecules or with bulk materials. Note: manipulate and control includes material synthesis” (ISO/DTS 80004-1) 4 2.2 Nanomaterials Bauer et al. (2008) note that ENMs can provide a vast range of functions and have material properties that shape a variety of products and services. ENMs are generally seen as having a great potential for providing benefits to sectors such as pharmaceuticals, potable water and water treatment, information and communication technologies (ICT), energy technologies, chemistry and green chemistry (Royal Society 2004; Fleischer and Grunwald 2008). This is largely due to the different material properties at the nanoscale, compared to materials at larger scales. Broadly speaking, ENMs differ from bulk materials due to two main reasons: relative surface area and quantum effects. Firstly, as particle size decreases a greater proportion of atoms can be found at the surface. Hence, nanoparticles have a greater surface area when compared to larger particles. This has the consequence of changing properties such a reactivity, strength and electrical characteristics. Secondly, quantum effects can affect the optical, electrical and magnetic behaviour of matter at the nanoscale. (Royal Society 2004) Various typologies have been developed to categorise nanomaterials. Foss Hansen et al. (2007) have developed a typology based on physical shape. This includes the categorises of I) bulk nanomaterials, II) materials that have nanostructure on the surface and III) materials that contain nanoparticles. This last category consists of several subcategories including, nanoparticles suspended in a solid, surface bound nanoparticles, airborne nanoparticles and nanoparticles suspended in a liquid. This categorisation is illustrated in Figure 1. Figure 1: Categorisation framework for nanomaterials (reproduced from Foss Hansen et al. (2007)) 5 Approximately 11.5 million tonnes of nanomaterials, with a market value of roughly 20 bn€, are produced and placed on the global market annually (European Commission 2012). Table 1 provides a list of nanomaterials currently on the European market. There is little quantitative data on the annual production of ENMs and although estimates of market size are perceived to be reliable, they still need to be taken with a degree of caution (European Commission 2012). The European Commission (2012) notes that carbon black (9.6 million t/year) and synthetic amorphous silica (1.5 million t/year) dominate the ENM market. Other ENMs produced in significant quantities include aluminium oxide (200,000 t/year), barium titanate (15,000 t/year), titanium dioxide (10,000 t/year), cerium dioxide (10,000 t/year), zinc oxide (8,000 t/year), iron oxides (2,500-3,000 t/year), zirconium dioxide (2,5003,000 t/year), carbon nanofibres (300-350 t/year), carbon nanotubes (200-250 t/year), silver (22 t/year) and platinum and palladium alloy (12 t/year). 6 Table 1: Nanomaterials on the EU market (Source: European Commission (2012)) Nanomaterial Inorganic non-metallic nanomaterials Synthetic amorphous silica (silicon dioxide) and similar substance Titanium dioxide Zinc oxide Aluminium oxide Aluminium hydroxides and aluminium oxo-hydroxides Iron oxides: diiron trioxide (ferric oxide, hematite) and triiron tetraoxide (ferrous-ferric oxide, magnetite) Cerium dioxide Zirconium dioxide Barium titanate Barium sulphate Strontium titanate Strontium carbonate Indium tin oxide Antimony tin oxide Calcium carbonate Aluminium nitride Silicon nitride Titanium nitride Titanium carbonitride Tungsten carbide Tungsten sulphide Metals and metal alloys Gold Silver Platinum and palladium alloy Nickel Cobalt Aluminium Zinc Manganese Molybdenum Tungsten Lanthanum Lithium Carbon-based nanomaterials Fullerenes Carbon nanotubes Carbon nanofibres Carbon black Graphene flakes Nanopolymers and dendrimers Polymer nanoparticles Polymer nanotubes, nanowires and nanorods Polyglycidylmethacrylate (PGMA) fibres Nanocellulose (fibrils and crystals) Nanostructured polymer-films Polyacrylonitrile nanostructures (PAN) Dendrimers Quantum Dots Nanoclays 7 2.3 Applications Compared to bulk materials, nanomaterials often display different chemical, physical, and biological properties; they behave differently even though possessing the same elemental or molecular composition. Nanomaterials have the potential to make every-day consumer products lighter, stronger, cleaner, less expensive, more efficient, more precise, or more aesthetic. (Lövestam et al. 2010) The Woodrow Wilson online nanotechnology consumer products inventory contains 1317 products3. Nanomaterials are used in a variety of product categories including health and fitness, home and garden, automotive, food and beverage, multifunctional products, electronics and computers, appliances and goods for children. Over 50% of products containing nanomaterial can be found within the health and fitness sector, including products such as cosmetics, clothing, personal care, sporting goods, sunscreen and filtration4. (Woodrow Wilson International Center for Scholars 2011) Table 2 details the various markets and applications of nanomaterials. 2.1 Potential to contribute to sustainability and unintended consequences The unique properties of nanomaterials are often associated with positive expectations in areas such as material and energy efficiency, pollution and waste reduction and sustainable development (Fleischer and Grunwald 2008; Bauer et al. 2008). Table 3 summarises some of the environmental advantages of nanotechnology in various sectors. However, environmental NGOs, such as Friends of the Earth, suggest that the “nanotechnology industry has over-promised and under delivered. Many of the claims made regarding nanotechnology’s environmental performance, and breakthroughs touted by companies claiming to be near market, are not matched by reality. Worse, the energy and environmental costs of the growing nano industry are far higher than expected.” Furthermore that warn that “… overall, this technology will come at a huge energy and broader environmental cost. Nanotechnology may ultimately facilitate the next wave of expansion of the global economy, deepening our reliance on fossil fuels and existing hazardous chemicals, while introducing a new generation of hazards.” (Illuminato and Miller 2010, 3–4) Gavankar et al. (2012, 296) note that at the nanoscale “materials of the same chemical composition but different particle-specific intrinsic and extrinsic factors may exhibit different behaviour and have different impacts on the environment and on human health.” Table 4 illustrates the potential impacts of the use of nanomaterials, highlighting the need to understand the environmental benefits and impacts of nanomaterials from a systems perspective. Hence, the claims regarding the potential for nanomaterials to contribute to sustainability require scrutiny. 3 th as of the 10 of March 2011 Som et al (2010) note this data should be used with caution. For instance, Dekkers et al. (2007) note that, at least from a Dutch perspective, it is possible that products on the market with the claim of ‘nano’ may neither contain nanomaterials nor be produced with nanotechnology, and not all products advertise the presence of nanomaterials in their products (as there has been no legal obligation to label products containing nanomaterials). 4 8 Table 2: Applications of nanomaterials in different markets (reproduced from Bauer et al. (2008)) Market Application C/P Automotive industry Lightweight construction Painting (fillers, base coat, clear coat) Catalysts Tyres (fillers) Sensors Wear protection for tools and machines Lubricant-free bearings C C C P C C C Construction materials Thermal insulation Flame retardants Surface-functionalized building materials for wood, stone, tiles. Façade coatings Groove mortar Surface-processed textiles Smart clothes Fuel cells Solar cells Batteries Capacitors Sun protection Lipsticks Skin creams Tooth paste Ceramic coatings for irons Odors catalysts Cleaner for glass, ceramic, floor, windows Fillers for paint systems Coating systems based on nanocomposites Impregnation of papers Switchable adhesives Magnetic fluids Data memory Displays Laser diodes Glass fibres Optical switches Filters (IR-blocking) Conductive, antistatic coatings Drug delivery systems Active agents Contrast medium Medical rapid tests Prostheses and implants Antimicrobial agents and coatings Agents in cancer therapy Package materials Storage life sensors Additives Clarification of fruit juices Ski wax Antifogging of glasses/goggles Antifouling coatings for ships/boats Reinforced tennis rackets and balls C C C C C/P C/P C P C C/P C/P C P P P P C P P C C/P C C C C P C C C C C C C C C C C C C C C C P C C/P P Engineering Construction Textile fabrics Energy Cosmetics Household Chemical industry Electronic industries Medicine Food and drinks Sports/outdoor C=component, P=product. 9 Table 3: Environmental advantages in products for different nanotechnology sectors (Reproduced from Bauer et al. (2008)) Sectors of nanotechnology Examples products Environmental advantages Nano electronic Electronic component, bioelectronic component Energy efficiency, speed data processing, replacement of silicon Nano optic Optoelectronic component Nano fabrication Nano structures for electronic components, ultra-thin layers of tools and components, Nanoparticles (as part) from new materials or new composites Higher data transfer rate, miniaturisation Energy efficiency, speed data processing, longer life time Nano chemistry, nanomaterials Nanobiotechnlogy Nano analytics Bio-based micro manufacturing of electronic components, bio sensors, bio catalyst, cellular engine Measuring instruments of quanta effects New mechanical, electrical, magnetically active, optical properties and therefore, unknown material functions less weight and volume, improvement of properties Medical early warning system energy efficiency Analysing nano structures Table 4: Potential benefits and impacts of the use of nanomaterials Nanomaterials and potential sustainability benefits Potential impacts Nanomaterials such as aluminium oxide, cerium oxide, zirconium oxide, perovskite, zeolites and precious metals (i.e., palladium, platinum and rhodium) can be used in catalytic conversion technologies in order to reduce unburned hydrocarbons, particulate matter and other emissions from cars and trucks. However, this would need to offset the potential negative environmental impacts that can occur during mining and production. Additionally, some materials are scarce leading to problems related to the scarcity of rare earth metals. Carbon nanotubes (CNTs) and aerogels can be used in nanocomposites in automotive applications to reduce weight and therefore increase fuel efficiency. However, the production of CNTs is energy intensive, which must be taken into consideration in any sustainability assessment. Biocides such as nanosilver (and to a lesser extent various metal oxides) are used in a wide variety of applications such as biocidal cleaning products, antimicrobial agents and coating, application to medical instrumentation and textiles. Nanosilver is an effective and antimicrobial treatment/coating and can reduce the application of hazardous substances (such as chlorine bleach) for similar purposes. However, silver is one of the most toxic metals to aquatic organisms (Luoma 2008). The increased use of nanosilver in consumer products, such as socks and other textiles, to reduce odour and/or kill bacteria my result in the increased flow of nanosilver to the environment. 10 3 METHODS FOR INVESTIGATING NANOMATERIALS FROM LIFE CYCLE PERSPECTIVE Environmental systems analysis (ESA) is a subfield of systems science (see Ackoff (1973) and Checkland (1999)) which aims at addressing environmental problems (Baumann and Tillman 2004). There are a number of ESA tools which differ in goal and scope (Finnveden and Moberg 2005; Finnveden et al. 2009). Tools which are used to investigate the environmental impact of products or substances include substance flow analysis (SFA), chemical risk assessment (RA), and life cycle assessment (LCA), see Figure 2. These tools incorporate the concept of a product or substance life cycle into their analysis. Figure 2: Environmental systems analysis tools and their focus (adapted from Finnveden and Moberg (2005)) In industry, the term ‘life cycle’ is generally understood as the life-span of either a material, a chemical or a product, covering its production, use and disposal (Som et al. 2009). Seager and Linkov (2008, 282) have noted that “It is now nearly universally accepted that the product life cycle is the proper perspective for thinking about materials, including nanomaterials”. However, the way the term ‘life cycle’ is used and perceived within different areas of expertise, such as LCA and RA, leads to different interpretations (Christensen and Olsen 2004). Below, Sections 3.1, 3.2 and 3.3 briefly introduce LCA, RA and SFA, respectively, and discuss how the ‘life cycle’ is perceived in each of these tools. 11 3.1 Life cycle assessment LCA assesses the potential environmental impacts of a product/service system over its life cycle. The term ‘life cycle’ includes the extraction and processing of raw materials, production, transportation and distribution, use, and end-of-life (re-use, recycling, recovery and final disposal) phases, see Figure 3. LCA is an accepted and internationally standardised tool (ISO 14040 – 14044), defined as the “compilation and evaluation of the inputs, outputs and the potential environmental impacts of a product system throughout its life cycle” (ISO 2006a, 2). The ISO standards lay down quality criteria for the design and execution of the LCA, as well as for the reporting of results, data, methods, assumptions and limitations. (Guinée, 2002). Part of this quality criteria is the need for a critical review, by a qualified expert or panel of experts, when an LCA is used to support comparative assertions. Figure 3: Life cycle of a product system (reproduced from UNEP (2007)) 12 LCA consists of four phases: goals and scope definition, life cycle inventory analysis (LCI), life cycle impact assessment (LCIA) and interpretation. - Goal and scope definition: The goals and scope definition of an LCA provides a description of the product system. According to ISO (2006b), the goal should state: “the intended application, the reasons for carrying out the study, the intended audience … and whether the results are intended to be used in comparative assertions intended to be disclosed to the public” (ISO 2006b, 7). - Life cycle inventory analysis (LCI): The LCI phase involves the compilation and quantification of data for all inputs (such as energy, water and materials usage) and outputs (such as air emissions, solid waste disposal, wastewater discharge) of all the processes in product/ service system throughout its life cycle. These data are related to the reference flow which is given by the functional unit (Hauschild 2005). - Life cycle impact assessment (LCIA): The LCIA phase translates the LCI input and output data into information about the system’s impact on the environment, human health and resources (Hauschild 2005). It is aimed at evaluating the significance of potential environmental impacts of the LCI phase (ISO 2006b). The LCIA phase consists of several steps: selection of impact categories, category indicators and characterisation models; classification; characterisation; normalisation; grouping; and weighting (see ISO (2006a) for detail about individual phases). According to ISO, mandatory elements of the LCIA phase are the selection of impact categories, category indicators and characterisation models, classification, and characterisation. Normalisation, grouping and weighting are optional. - Interpretation: The interpretation phase evaluates all the LCA results according to the defined goal and scope, which reach conclusions, explain limitations and provide recommendations. The interpretation phase should include a sensitivity and uncertainty analysis to qualify the results and conclusions of the study (Hauschild 2005). It is important to establish a demarcation between LCA and life cycle thinking (LCT). LCT is a concept that “seeks to identify possible improvements to goods and services in the form of lower environmental impacts and reduced use of resources across all life cycle stages. … The key aim …is to avoid burden shifting” (European Commission 2010), whereas LCA aims to describe the potential environmental impacts of a product/service system over its life cycle. 13 As early as 2004, the Royal Society (2004, 32) suggested that the “potential benefits of nanotechnologies should be assessed in terms of lifecycle assessment”. Furthermore, according to Grieger et al. (2012), there is a general consensus amongst scientists, researchers and regulatory agencies that the potential health and environmental risks of ENMs should be evaluated over there entire life cycle. In the context of nanotechnology, the term ‘life cycle’ can be used in reference to both ENMs and nanoproducts (Som et al. 2009). For instance, a nanoparticle can be incorporated into different ENMs which can then be used in different products. Hence, there may be a variety of use and end of life phases for nanomaterials, depending upon the products in which they are incorporated, as illustrated in Figure 4. Figure 4: Cycles for ENMs determined by the life cycles of nanoproducts (reproduced from Som et al. (2009)) 14 3.2 Risk assessment Although there are a multitude of definitions of risk, Renn (2008, 373) defines risk “an uncertain consequence of an event or an activity with regard to something that humans value” 5, the consequences of which can be either positive or negative depending upon the values people associate with them. Risk assessment (RA) has been the standard approach to assessing the potential risk of bulk chemicals. RA assesses the risk to human health and the environmental of a single substance at a particular point in a chemical’s life cycle or the total release of a substance from a chemical’s life cycle (Grieger et al. 2012). The term ‘life cycle’ covers all downstream uses of the chemical, from the manufacture of substance to its disposal or the preparations/articles containing the substance (Christensen and Olsen 2004). RA is often performed to identify whether any life-cycle stages pose a risk (Grieger et al. 2012). The difference between the conception of ‘life cycle’ in LCA and RA is illustrated in Figure 5. Whilst LCA assesses a range environmental impacts of a product system related to a functional unit from the cradle-to-grave, RA assesses the health and environmental risk of a single substance at a particular point in the substances life cycle. More specifically, RA is “the task of identifying and exploring, preferably in quantified terms, the types, intensities and likelihood of the (normally undesired) consequences related to a risk”. RA consists of four steps: - hazard identification: the mapping of a chemical’s inherent physico-chemical and biological properties required to provide a uniform basis for the evaluation of hazard potential. - dose-response assessment: quantitative estimation of the chemical concentration expected not to have an effect on human health or the structure and function of an ecosystem’s species. - exposure assessment: the application of generic and/or specific scenarios of exposure pathways for a chemical, resulting in the a predicted environmental concentration value for each scenario. - risk characterisation: compares the exposure of each exposed population with the appropriate derived no-effect level, compares the concentrations predicted in each environmental sphere with the predicted no-effect level, and assesses the likelihood and the severity of an event arising from the physico-chemical properties of the substance. There is a consensus that the RA framework is applicable to passive ENMs (SCENIHR 2009, 2010). However, many of the methodological steps within RA require further refinement or development for ENMs (Grieger et al. 2012). There has been a recent call for a complementary application of LCA and RA to ENMs (see Linkov and Seager (2011) Grieger et al. (2012) and Shatkin (2008)). In this context, two approached to RA from a life cycle perspective have been identified: ‘LC-based RA’ and ‘RA-complemented LCA’(Grieger et al. 2012) (See Chapter 5). 5 Original definition in Kates et al. (1985, p.21) 15 Figure 5: Comparison of how RA and LCA perceive the term ‘life cycle’ (reproduced from Grieger et al. (2012)) 16 3.3 Substance and particle flow analysis Several authors have suggested a substance life cycle approach to assess the emissions of ENMs (Lubick 2008; Sweet and Strohm 2006). Substance flow analysis (SFA) is a tool sometimes applied prior to RA in order to estimate emissions (van der Voet et al. 1999). Consequentially, SFA has become the point of departure for the development of emission assessment methods (Arvidsson 2012). SFA focus on the flows and stocks of materials, substances and particles of interest to society. Its overall goal is to quantify the flows and stocks of a substance and estimate the emissions from different life cycle stages, thus providing an input for policy relating to environmental pollution (van der Voet 2002). The core principle SFA, is based in the mass balance principle, derived from Lavoisier’s law of mass conservation (Lavoisier, 1789). Arvidsson (2012) notes that such an analysis is often based on product life cycles, which includes raw material extraction, production, use and end-of life, as illustrated in Figure 6. Flows between, and stocks within, the life cycle stages are quantified and measured as mass per unit time (i.e., tonnes/year) and mass only (i.e., tonnes), respectively (Arvidsson 2012). van der Voet (2002) suggests that SFA aims to provide relevant information for an overall management strategy with regard to one specific substance or group of substance. Arvidsson (2012) notes that emissions from society to the environment are of specific interest to SFA studies since the flows of some substances are of particular environmental importance. However, Arvidsson et al. (2011, 845) note that “there are strong indications, however, that mass may not be a relevant indicator of flow and stock magnitude, exposure, or toxic effects for the case of NPs”. Rather than using mass as a measure of the flows and stocks of nanomaterials, particle flow analysis (PFA) measures the flows and stocks of particles. This allows for relevant properties, such as particle size, to be accounted (Arvidsson et al. 2011). Furthermore, processes that change particle number (such as agglomeration, melting of particles, dissociation of particles into ions, and grinding) can be included into the analysis. Figure 6: Substance flow analysis model (reproduced from Arvidsson et al. (2012) 17 4 LIFE CYCLE ASSESSMENT OF NANOMATERIALS 4.1 Research applying life cycle assessment to nanomaterials 4.1.1 Current research in the EU and OECD European Union At the European level, there is a significant call to study ENMs from a life cycle perspective. For instance, the European Commission communication Towards a European Strategy for Nanotechnology states “… R&D also needs to take into account the impacts of nanotechnologies throughout their whole life cycle. For example, by using Life-Cycle Assessment Tools” (European Commission 2004). The European Commission’s Nanosciences and Nanotechnologies: An action plan for Europe 2005-2009. Second Implementation Report 2007-2009 (European Commission 2009b) notes that from a regulatory point of view, an urgent need is the improvement, development and validation of methods in the areas of “characterisation, exposure assessment, hazard identification, life cycle assessment and simulation.” (European Commission 2009b, 9). The Accompanying document to the Nanosciences and Nanotechnologies: An action plan for Europe 2005-2009 Second Implementation Report 2007-2009 (European Commission 2009c) suggests that there is a need to “further adjust, validate and harmonise currently available guidelines for the life cycle assessment of nanomaterials and nanotechnology-based products, building upon results from completed and ongoing activities. To develop hands-on guidance for simplified LCAs for SMEs.” (European Commission 2009c, 92). FP7 research projects which have some relationship to LCA of nanotechnology and ENMs have been identified (see Jovanovic and Cordella (2011), OECD (2011) and the OECD Database on Research into the Safety of Manufactured Nanomaterials6). These projects are outlined in Appendix A. Findings from the research programme PROSUITE (Walser et al. 2012, 2011; Hischier and Walser 2012) and NanoImpactNet (Som et al. 2010) feature prominently in Sections 4.1.2 and 4.1.3. Organisation for Economic Co-operation and Development Within the Organisation for Economic Co-operation and Development (OECD), two working parties have been established: the Working Party on Nanotechnology (WPN) and the Working Party on Manufactured Nanomaterials (WPMN). The OECD has recently released a summary of National Activities on Life Cycle Assessment of Nanomaterials (OECD 2011). This document has compiled information on OECD members national activities related to LCA of nanotechnology and ENMs which have been provided by delegations from the following countries: Austria, Finland, Germany, Korea, Poland, the United Kingdom, United States, the European Commission, as well as from the Business and Industry Advisory Committee to the OECD (BIAC). See OECD (2011) for a detailed summary of these research projects. 6 http://webnet.oecd.org/NANOMATERIALS/Pagelet/Front/Default.aspx 18 4.1.2 Literature review A comprehensive meta-analysis of the state-of-the-art of LCA research on ENMs is beyond the scope of this report. Whilst highlighting some of the current research results, our primary objective is to focus on the potentials and limitations of current research efforts in order to propose further research priorities. To this end, a non-exhaustive search of academic literature databases (Scopus, ScienceDirect) and an internet search for publications (such as those in scientific journals, conference proceedings, conference presentations, research reports and theses) was completed using the following combination of keywords: - nano + life cycle assessment - nano + “life cycle assessment” - nano + “life cycle” - nano + LCA The studies identified are highlighted in Table 5. Three meta-analyses of the LCA of ENMs can be found in the peer-reviewed literature: Hischier and Walser (2012), Gavankar et al. (2012) and Upadhyayula et al. (2012). Hischier and Walser (2012), Gavankar et al. (2012) reviewed all studies applying LCA to ENMs, whilst Upadhyayula et al. (2012) specifically focused on carbon nanotubes (CNTs) and carbon nanofibres (CNFs). The LCA studies considered by these meta-analyses can be found in Table 5. Table 5: LCA studies of nanomaterials. Publications Type Babaizadeh and Hassan (2013) †‡ Bauer et al. (2008) * J De Figueirêdo et al. (2012) J J † Fthenakis et al. (2008) ; † Fthenakis et al. (2009) Greijer et al. (2001) C C † J Griffiths and O’Byrne (2013) † Grubb and Bakshi (2008) ; † Grubb (2010) ; Grubb and Bakshi (2011a, †‡ 2011b) †‡ † Healy et al. (2008 *; 2006 ); † Isaacs et al. (2006, 2010) † Joshi (2008) J T C J Nanomaterial Focus of Study TiO2 Comparison of TiO2 coated class with float glass Ti, TiAl, Ti+TiAl Examine implications of life cycle thinking on nanotechnology (and nanoproduct) evaluation; 2 case studies Cellulose The comparison of two alternative nanowhiskers processes for the production of cellulose nanowhiskers Nanocrystaline-Si, Comparison of the cumulative energy nano CdTe, and demand for the production of PV nano-Ag PV systems systems using nanomaterials Nanocrystaline dye Identify the significant environmental (out of nano- TiO2 aspects of nanocrystaline dye sensitive and carbon black) solar cell system Multi walled carbon Identification and quantification of the nanotubes (MWCNT) environmental impact of MWCNT formation via catalytic chemical vapour deposition. TiO2 Evaluate the production processes for TiO2 J;C C;J J Life cycle phases E M U EOL O O O x O O O O O O x x O O x x O O O O O O x x O O x x Single walled carbon Environmental assessment of SWNCT O O x x nanotubes (SWCNT) production Nanoclay (ONMT, Comparison of nanoclay composite O O x x organically modified biopolymer with biobased polymers montmorillonite) † Reported in Hischier and Walser (2012); ‡ Reported in Gavankar et al. (2012); * Reported in Upadhyayula et al. (2012); C: Conference; J: Journal; T: Thesis; B: Book Chapter; R: Report. O: Included; x: Excluded 19 Table5 (Cont.): LCA studies of nanomaterials Publications Type † Khanna et al. (2007) ; Khanna, Zhang, et al. † (2008b) ; Khanna, Bakshi, et al. †‡ (2008a) * ‡ Köhler et al. (2008) Nanomaterial Focus of Study J B C CNFs Environmental burden of CNF synthesis J CNTs Life cycle phases E M U EOL O O x x Potential release of carbon nanotubes x O O O throughout the life cycle of textiles and lithium-ion batteries † Kushnir and Sandén (2008) J Fullerenes and CNT Implications for industrial scale O O x x ‡ * production †‡ Lloyd and Lave (2003) ; J Nanoclay-reinforced Replacing auto-body panels made of O O x x † Lloyd (2004) T polymer composites steel with those of polymer composites with aluminium †‡ Lloyd et al. (2005) ; J Nanoscale platinum- Evaluating reduction in non-renewable O O x x † Lloyd (2004) T group metal (PGM) resources like PGM via greater process particles control offered by nanotech Merugula et al. (2010) C CNTs (in reinforced Comparison of vapour-grown carbon O O O O wind turbine blades) nanofibre reinforced glass fibre epoxy matrix and glass fibre reinforced plastic ‡ Meyer et al. (2011) J Ag Identifying the life cycle hot spots via O O x x screening level LCA Moign et al. (2010) J Zirconium Comparison of spraying technologies O O x x nanopowder for the manufacture of yttria-stabilised zirconia †‡ Osterwalder et al. (2006) J Various oxide Energy comparison of wet and dry x O x x nanoparticles synthesis methods for oxide nanoparticle production †‡ Roes et al. (2007) J Polypropylene Compare environmental impact and O O O O Nanocomposite cost of polypropylene nanocomposite with conventional polypropylene in the use cases: i) packaging film, ii) agricultural film, iii) automotive body panel. † Roes et al. (2010) J SiO2, CaCo3, CNTs Compared the non-renewable energy O O O O WMCNTS, use of 23 nanocomposite materials organophilic with 3 conventional composite montmorillonite materials ‡ Şengül and Theis (2011) J QD photovoltaics LCA of a proposed type of O O x x nanophotovoltaic, quantum dot photovoltaic module † Singh et al. (2008) *; J CNTs Environmental Impact Assessment O O x x † Agboola (2005) T (EIA), via LCA method, of two methods for producing SWCNTs Steinfeldt, Gleich, et al. R Nanoelectronics Lighting – LEDs x x O x †‡ (2004a) ; R Nanomaterials/nano Chemical/paintings O O x x Steinfeldt, Petschow, et al., particles Chemical/plastics O O x x † (2004b) Electronics/displays O O x x † Walser et al. (2011) J Ag Comparison of the environmental O O O O benefits and impacts of nanosilver Tshirts with conventional T-shirts and Tshirts treated with triclosan † Reported in Hischier and Walser (2012); ‡ Reported in Gavankar et al. (2012); * Reported in Upadhyayula et al. (2012); C: Conference; J: Journal; T: Thesis; B: Book Chapter; R: Report. O: Included; x: Excluded 20 4.1.3 Meta-analyses Engineered nanomaterials Hischier and Walser (2012) and Gavankar et al. (2012) note that whilst there is plenty of literature promoting the application of LCA, studies applying LCA to the area of nanotechnology are ‘scarce’. Furthermore, these studies only looked at parts of the life cycle, with no quantitative studies addressing all life cycle phases. ENM studies in these meta-analyses included: cadmium telluride, calcium carbonate, carbon black, carbon nanofibres (CNFs), carbon nanotubes (CNTs), nanoclay, nanoscale platinumgroup metals, silica, silver, silicon, titanium and titanium oxide. Product systems studied included: auto-body panels, biopolymers, coatings, electronic displays, electronic sensors, lithium-ion batteries, photo voltaic systems, packaging and agriculture polymer films, ENM production processes, textiles and wind turbine blades. The general conclusions of these meta-studies can be summarised as follows: - Proper goal and scope definition is of “crucial importance in order to get meaningful results that take into account the different properties, especially for comparisons with traditional materials.” (Hischier and Walser 2012, 279) - The LCIs cannot be classified as comprehensive as they often lack ENM specific data related to the outputs of the processes (Hischier and Walser 2012). Hischier and Walser (2012, 279) also highlighted the “considerable variability of the (traditional) inventory items like energy input, material input, etc., … especially concerning the energy consumption for the production of the various engineered nanomaterials.” Hence, populating LCI databases with ENM specific information, such as size and shape, is of critical importance (Gavankar et al. 2012). The retention of as much nanospecific information as possible would then facilitate subsequence LCIA for ENMs (Gavankar et al. 2012). - Regarding LCIA, “there is a complete lack of characterization factors for release of nanoparticles indoors and outdoors. ... Only in exceptional cases are first approaches to examine e.g. the toxicity of the emissions to air and water reported. However, it is not always clear if nano-specific aspects were taken into account” (Hischier and Walser 2012, 279). A number of existing tools, such as USEtox (Rosenbaum et al. 2008), CALtox (McKone and Enoch 2002) and QSAR (Dudek et al. 2006; Puzyn et al. 2009, 2010), are available to quantitatively assess the fate, transport or toxicity of chemicals and bulk materials (Gavankar et al. 2012). The incorporation of additional information on ENM specific properties into these existing tools would allow the modelling capability of ENMs behaviour and impact in the environment. - Due to the lack of modelling techniques available for the critical, yet frequently omitted, use and end-of-life phases, the development of protocols and models are needed to enable a holistic assessment that takes into consideration ENM’s intrinsic properties (Gavankar et al. 2012). - In the absence of any empirical data, qualitative or screening LCAs should be performed (Gavankar et al. 2012). - LCA should be complemented with tools such as risk assessment when location specific parameters are critical for understanding the behaviour and impact of ENMs, as LCA may not be able to capture such context specific sensitivities (Gavankar et al. 2012). 21 Hischier and Walser (2012) and Gavankar et al. (2012) considered a diversity of ENMs and product systems, consequentially they could not compare the result of individual LCA studies. Instead, these studies investigated the how the LCA methodology was applied to each case; for instance, the function units considered, the life cycle stages considered, the environmental impact categories selected, the consideration of the ENM specific data in the LCI and LCIA phases, and data and methodological gaps. Hischier and Walser’s (2012) meta-analysis of the ‘Life cycle assessment of engineered nanomaterials’ aimed to 1) provide an overview of LCA studies in the area of ENMs, and 2) identify the shortcoming which contribute to delaying the comprehensive application of the LCA framework to ENMs, and 3) propose strategies to overcome these shortcomings. Although the authors noted a scarcity of studies applying the LCA approach to the area of nanotechnology, 17 studies7 were analysed. The authors categorised the studies by how they addressed the following issues: consideration of functional unit, consideration of system boundaries, production systems studied, LCI and LCIA. These results are summarised in Table 6. 7 Some studies having multiple publications 22 Table 6: Summary of studies applying LCA studies to ENMs as reported in Hischier and Walser (2012) Aspect Functional unit System boundaries Life cycle phases Production systems studies Resource extraction & production Use phase End of life Life cycle inventory Life cycle impact assessment Meta Study Results In terms of functional unit, two groups of studies were distinguished: - Weight based: Half the studies assessed the environmental impact of the production of a specified quantity (usually 1 kg) of ENM - Application/service based: The other half of the studies assessed the environmental impact of the specific application of the ENM, for instance 1kWh of electricity output from a nano solar cell system. In terms of life cycle phases addressed, two groups were identified: - Cradle-to-grave studies: considered all life cycle stages, from extraction of raw materials to the end-of-life phase. Six of the 17 studies considered all life cycle phases. - Cradle-to-gate studies: considered the resource phase and the production phase. The gates of these studies were considered as either the production site of the ENMs, or the production site of the product containing ENMs. Eleven of the 17 studies were cradle-to-gate studies. Three quarters of the studies compared the ENM with traditional materials. This comparison was divided into two groups. The first considered the specific application of the respective material and thus employed context related 2 functional unites (such as 1m of photovoltaic cell material). The other group assessed the production of the ENM without taking into consideration any contextual application. All studies included the extraction and production phases. Ten of the 17 studies included the use phase. Although there was typically not a lot of detail reported, such as the release of ENMs. The six studies that considered the end of life phase mostly assumed incineration in a municipal solid waste incinerator. However, traditional models for incineration were used which do not take into account the fate of ENMs as a separate flow. One study, Bauer et al. (2008), qualitatively described the potential release pathways during the end of life phase. Hence, ENMs were not quantitatively evaluated during the end of life phase. The majority of studies used publically available literature. Only in four out of the 17 studies were data taken from either actual production plant (pilot and commercial), theoretical calculations, or process simulations. All studies included detailed information on energy use, and many studies included information on material inputs. Whilst several studies reported emissions to air, these emissions were for ‘conventional’ flows. Furthermore, there was little detail on information regarding emissions to water and soil, be they ‘conventional’ or nano emissions. Only one study, Walser et al. (2011), covered the output of ENMs. Three studies considered were LCI or energy analysis studies (no LCIA performed). The remaining studies often reported one or two LCIA categories. LCIA categories linked to energy consumption such as global warming potential (GWP) were considered by the majority of the studies. Although the release of nanoparticles to air, water and soil are suspected of having potential impacts on human health and the environment, only Walser et al. (2011) considered the freshwater and seawater toxicity results for colloidal silver. Whilst several studies reported ecotoxicity, no LCIA methods contain characterisation factors for either the indoor or outdoor releases of nanoparticles. 23 Carbon nanotubes Upadhyayula et al. (2012) recently conducted a review of LCA of CNTs, analysing seven studies8. The aim of this research was to emphasise the role of LCA during the development of CNT products in order to mitigate potential impacts to human health and ecosystems over their life cycle. These studies considered CNTs and CNFs9. Five studies were cradle-to-gate and two studies were cradle-to-grave. For the cradle-to-gate studies, the functional units related to the production of 1kg (in one case 1 g) of CNTs via alternative production technologies. Table 7 illustrates the cradle-to-gate energy consumption for the four major synthesis routes for CNTs: electric arc discharge, laser ablation, chemical vapour deposition (CVD) and highpressure carbon monoxide (HIPCO) processes. CNT manufacture is energy intensive due to the processes involved in the preparation of raw materials (i.e., the need for ultrapure graphite, purification of gasses and purification of CNTs prior to use) and the high temperature requirements for synthesis processes (Upadhyayula et al. 2012). Table 7 does not account for the full impact of air emissions and waste stream discharges (Upadhyayula et al. 2012), because accurate models, and characterisation factors for human health and ecological impact, have yet to be developed (Khanna 2009). For instance, liquid waste from nanoproduct manufacture may potentially contain CNTs and other toxic materials, such as heavy metals, which require treatment that has not been included in these studies (Upadhyayula et al. 2012). Furthermore, solid waste from these manufacturing processes can potentially lead to the generation of solid hazardous waste due to the low recyclability of metal catalysis (Upadhyayula et al. 2012). The energy demand for a CNT product is dependent upon: a) quantity of CNTs used in the product, b) purity and type of CNTs needed and, c) specialised operational steps as demanded by the CNT product manufacturing process. Upadhyayula et al. (2012) infer that the potential environmental problems caused by CNT products can be attributed to two factors: the energy intensive manufacturing stage and the potential release of toxic air emissions and liquid waste discharges in the various life cycle stages. 8 Five of these LCAs where analysed by Gavankar et al. (2012) and six were analysed by Hischier and Walser (2012) Although the CNFs are not technical CNTs, both nanomaterials have a similar have similar production methods with comparable impacts. (Upadhyayula and colleagues 2012). 9 24 Table 7: Cradle-to-gate energy requirement for various methods of CNT and CNF synthesis (reproduced from Upadhyayula et al. (2012)) Synthesis method Arc discharge Material inputs Precursor Catalyst Pure Ni, Co, Y graphite Acid Nitric acid Laser ablation Pure graphite Ni, Co, Y Nitric acid CVD (Fixedbed) Hydrocarbons Fe, Ni, Co, Mo Mineral acids CVD (Fluidized bed) Methane CVD (Floating bed) Benzene CVD (VGCNF) Benzene HIPCO & CoMoCAT Carbon Monoxide Methane Ethylene Benzene Energy (MJ/kg) 4.6E+05 3.2E+05 2.2E+05 8.7E+07 9.6E+03 7.0E+05 9.2E+05 6.3E+05 SRCY (%) 4.5 4.5 50 - Fe, Ni, Co, Mo on Mineral metal oxides acids 8.5E+02 30 Fe, Ni, Co, Mo on Mineral metal oxides acids Ferrocene Mineral acids 4.8E+02 - 1.1E+04 8.0E+03 2.9E+03 50 50 70 Khana et al. 2008 Iron pentacarbonyl or Co and Mo 4.7E+05 1.6E+05 5.3E+08 8.1E+07 2.4E+07 5.8E+03 50 50 - Gutowski et al., 2010 Healy et al., 2008 Nikolaev et al., 1999 Brownikowski et al., 2001 Smalley et al., 2007 Kushnir & Sanden, 2008 Nitric acid 25 2.95 2.95 References Product characteristics Gutowski et al., 2010 Healy et al., 2008 Kushnir & Sanden, 2008 Gutowski et al., 2010 Kushnir & Sanden, 2008Ganter et al., 2009 Gutowski et al., 2010 Healy et al., 2008 (i) Structurally superior (ii) Low-level metal impurities (i) Structurally superior (ii) Low-level metal impurities (i) Low structural quality (ii) High-level metal impurity requiring intensive purification Kushnir & Sanden, 2008 (iii) Assumed purification yields up to 90% Kushnir & Sanden, 2008 (i) Assumed purification yields of 90% (ii) Less stringent purity requirements (i) Structurally superior (ii) High purification yields 4.1.4 Results from selected case studies Carbon nanofibres: Khanna et al (2008a) and Khanna and Bakshi (2009) Studies by Khanna et al. (2008a) and Khanna and Bakshi (2009) highlight the importance of considering the function of the nanoproduct with a specified context. Khanna et al. (2008a) evaluated the life cycle energy requirements (cumulative energy demand) and performed an LCIA for CNF synthesis using various hydrocarbon feedstocks and compared these results with those of traditional materials (steel, aluminium and polypropylene). An LCI was completed for vapour-grown carbon nanofibres (VGCNF) based upon on laboratory data and data available from literature sources. The authors show that the cradle-to-gate energy requirements of CNFs range from 2,872 MJ/kg for benzene feedstock to 10,925 MJ/kg for methane feedstock. This energy requirement is significantly greater when compared to conventional materials: steel 30MJ/kg, aluminium 218 MJ/kg and polypropylene 119 MJ/kg (Khanna et al. 2008a). The LCIA results for the midpoint indicators global warming potential, human toxicity potential, ozone layer depletion potential, photochemical oxidation potential, freshwater aquatic ecotoxicity potential, terrestrial ecotoxicity potential, acidification potential, acidification potential and eutrophication potential are illustrated in Figure 7. However, the LCIA did not consider the release and impact of nanoparticles on human and ecosystem species. This type of comparison can be useful for identifying hot-spots of environmental impact in the manufacture of CNFs, thus highlighting areas for improvement. Upadhyayula et al. (2012) note the findings of LCAs of CNT production highlight the need to refine CNT manufacturing by adapting synthesis techniques involving low temperatures, renewable feedstocks and recycled materials as catalytic supports. 26 Figure 7: LCIA midpoint indicators for vapour-grown carbon nanotubes compared to aluminium, steel and polypropylene (reproduced from Khanna et al. (2008a)) 27 However, the direct comparison of the CED of the production of 1 kg of CNFs to the production of 1kg of steel, aluminium or polypropylene does not reflect the actual replacement of steel, aluminium or polypropylene (Hischier and Walser 2012). The potential environmental impacts of specific nanoproducts need to be compared to conventional products (i.e., products manufactured from either steel, aluminium or polypropylene) which provide the same performance or utility (Khanna and Bakshi 2009). In this context, Khanna and Bakshi (2009, 2079) conducted such a study, with the aim of assessing the CED of CNF based polymer nanocomposites (PNCs) and their comparison with steel “on a functional unit basis of a standard plate and automotive body panel”. In this study, various CNF polymer nanocomposites (PNCs) (content varying between 0.6 to 15 vol.%) with and without glass fibres were analysed. Results from this study are illustrated in Figure 8. This comparison shows the dramatic reduction in CED due to the small amount of CNFs used to provide the same mechanical stiffness as a steel component. For example, to achieve a similar functionality (mechanical stiffness) to 1 kg of steel, a polypropylene CNF-glass fibre matrix containing 2.3% CNFs was necessary, resulting in a weight of 0.38 kg. Figure 8: CED of polymer nanocomposites that provide equal stiffness to a steel component (Reproduced from Khanna and Bakshi (2009)) 28 Furthermore, using the results produced by Khanna et al. (2008a), Hischier and Walser (2012) show the importance of considering the use phase. When the use phase is considered, the reduced weight of the CNF reinforced PNC component leads to a lower CED when compared to the steel component. Figure 9 highlights three types of comparisons of the CED of CNFs and steel, highlighting the importance of both the function unit and use phase in the analysis of ENMs. Point ① shows the CED for the production of 1 kg of CNF is almost 11 times higher than that of 1 kg of steel (Hischier and Walser 2012). Point ② shows the considerable reduction in the CED of the PNC component, although still higher than the steel component. Point ② shows the lighter weight of the PNC component influences the energy use during the use phase, and the PNC component has a lower CED than the steel component. “Difference in the CED value (expressed in GJ/kg) when compared with 1 kg of steel; i.e. a negative value indicates that the CED of the respective material is lower than the CED of 1 kg of steel – point ① on the level of pure materials, point ② on the level of equally stiff material, and point ③as part of a car, driving around 280000 km” (Hischier and Walser 2012, 276) Figure 9: Difference in CED of CNF reinforced PNCs compared to steel (Reproduced from Hischier and Walser (2012)) 29 Nanosilver T-Shirts: Walser et al. (2011) Walser et al. (2011) have compared the environmental impact of nanosilver T-shirts to conventional T-shirts with and without biocidal treatment. The authors investigated the environmental performance of two nanosilver production processes: commercialised flame spray pyrolysis (FSP) with melt-spun incorporation of silver nanoparticles, and plasma polymerization with silver cosputtering (PlaSpu) at the laboratory, pilot and commercial (estimated) scales. The T-shirts were compared via the functional unit of “being dressed with a biocidal polyester T-shirt for outdoor activities during one year in Switzerland (wearing it once a week)” (Walser et al. 2011, 4573). Figure 10 illustrates the life cycle CO2-equivilent emissions of a conventional T-shirt, a T-shirt treated with 22mg of triclosac (a bicoidal treatment), a T-shirt with 47mg of nano sized silvertricalciumphosphate (nanoAg-TCP) treated by the FSP process, and a T-shirt treated with 31mg of pure nanosilver from the PlaSpu process. For the conventional T-shift, triclosan Tshirt and FSP nano sliver coated T-shirt, the largest contribution to the carbon footprint was the use phase, which was assumed to be 100 washes. The PlaSpu process significantly increased the carbon footprint of the T-shirt. The production of nanosilver from the FSP process (0.21 kg of CO2-equivelnts) has no significant influence on the CO2-equivilent emissions during over the life cycle of the T-shirt. However, the PlaSpu process has a much greater influence on the carbon footprint of the Tshirt, even though this reduces with technological development (PlaSpu laboratory: 164.0 kg of CO2-equivelnts, PlaSpu pilot: 15.24 kg of CO2-equivelnts, and PlaSpu commercial: 5.14 kg of CO2-equivelnts). Figure 10: Cradle-to-grave climate footprint of biocidal T-shirts and a regular T-shirt (100 washings) (reproduced from Walser et al. (2011)) 30 This was the only LCA study to analyse the freshwater and seawater toxicity of Nanosilver. Compared to a conventional T-shirt, both silver and triclosan emissions are not considered relevant when taking the whole life cycle into consideration. The silver released from washing accounts for less than 1% of the overall freshwater toxicity of the nanosilver T-shirts. Figure 11 highlights the insignificant contribution of the release of silver over the T-shirts life cycle compared to freshwater toxicity associated with processes over the T-shirts life cycle. Figure 11: Comparison of the freshwater toxicity for the life cycle of one T-shirt (reproduce from Walser et al. (2011)) 31 4.2 Obstacles to and strategies for the application of LCA to nanomaterials 4.2.1 Goal and scope definition Hischier and Walser (2012) note that engineered ENMs may have specific functions and material properties that provide additional gains when used as a substitute for traditional materials. Hence, these additional functions and properties should be taken into consideration in the scope of the LCA. Due to these special material properties, Bauer at al. (2008, p 914) note that it “seems obvious that materials and services must be assessed in the context of a product or a functional purpose to quantify expected benefits also with regard to the entire life cycle.” Obstacles Functional unit ISO (2006b, 9) states that a key feature of LCA, separating it from other methods such as environmental impact assessment and risk assessment, is that it is a “relative approach based on a functional unit”. LCA relates the inputs and outputs of a system to the function that is provided: “This functional unit defines what is being studied. All subsequent analyses are then relative to that functional unit, as all inputs and outputs in the LCI and consequently the LCIA profile are related to the functional unit” (ISO 2006b, 7). Since LCA typically compares alternative ways of delivering the same function, it is important that the systems being compared actually provide the same function (Hauschild 2005). Typically, more than one function is provided by one of the systems being compared. In this case, methods such as allocation or systems expansion (see JRC (2010) for a detail description of these methods) can be followed to ensure that the systems being compared are equal. Several authors note that the choice of functional unit is an especially important consideration - a prerequisite - in order to perform a meaningful LCA of ENMs (Klöpffer et al. 2007; Hischier and Walser 2012; Bauer et al. 2008). The definition of the functional unit becomes more difficult when dealing with ENMs due to the plethora of functions and material properties which can be achieved at the nanoscale. As demonstrated in section 4.3.2, basing a functional unit on weight alone would only make sense if one is comparing alternative production processes to produce, for instance, one tonne of the same ENM with the same functionality. Since the potential sustainability benefits of ENMs are related to their interactions with other materials or components, once the goal of the study goes beyond the production of a specific ENM a functional unit based on weight is inappropriate (Hischier and Walser 2012). When the goal of the study involves the use of nanoproducts the functional unit should be defined in relation to the service provided by the product (the product performance during the use phase) (Hischier and Walser 2012). However, Klöpffer et al., (2007) note that nanoproducts fulfil functions that are quite new, leading to a difficulty to specify functional alternatives. 32 System boundaries The system boundaries determine what processes should be included in an LCA. ISO (2006b, 12) states that “Ideally, the product system should be modelled in such a manner that inputs and outputs at its boundary are elementary flows.” Decisions regarding what processes should be included or excluded from a study will ultimately have an influence on the result of the study. Hence, it is important that the models, assumptions and choices made should be transparent. Klöpffer et al. (2007) stress that LCA studies of ENMs and nanoproducts should address all life cycle stages. However, in their analysis of current research addressing complementary use of life-cycle assessment and risk assessment for ENMs, Grieger et al. (2012, 6) note that few “studies have encompassed the full life-cycle, and most of them focused on a cradle-to-gate study or on a specific LC stage. … Moreover, the majority of these have relied upon generic life-cycle impact databases or general literature in formulating the inventories and impact assessment criteria (i.e., excluding potential toxicological impacts of NMs).” Indeed, Gavankar et al. (2012) have noted that no LCAs of ENMs are compliant with the ISO standards as none of them have covered the complete life cycle of an ENM or product. Use phase Bauer et al. (2008, p. 914) note that a dissipative use of ENMs is characteristic for a large share of products, whereby the ENM enters the environment during its use. To this end, Gottschalk and Nowack (2011) have demonstrated that initial analytical and experimental studies have shown evidence for the release of ENMs from products such as textiles and paints. In their meta-analysis, Hischier and Walser (2012, 274) note that in studies which included the use phase there was limited information regarding how ENMs were considered. The authors noted that this is “an astonishing fact when one keeps in mind the various advantages, e.g. in view of the sustainability of ENMs highlighted in various studies about this new technology”. End of life A small group of nanoproducts are designed for a longer life, where the ENMs will enter the end of life phase of the product life cycle (Bauer et al. 2008). Bauer at al. (2008) note that little emphasis has been given to the end-of-life phase, which has often been disregarded due to data gaps. Although six of the studies reported in Hischier and Walser (2012) considered the end-of-life phase, the treatment of the ENM was omitted. Municipal solid waste incineration models used did not consider the amount or fate of ENMs as a separate flow (Hischier and Walser 2012). 33 Strategies Klöpffer et al. (2007) note that the goal and scope definition should reflect issues such as: the use of nanoproducts being in line with the product recommendation or predictions, and potential rebound effects resulting from the use of nanoproducts. To address these issues the use of sensitivity analysis is recommended in the interpretation phase. 4.2.2 Life cycle inventory analysis (LCI) Obstacles Ensuring the collection and use of complete, reliable, transparent and acceptable data, including the explanation of assumptions, are problems faced during the LCI phase of LCAs of ‘conventional’ product systems (Klöpffer et al. 2007). However, when LCA is applied to nanotechnologies and ENMs these problems are amplified (Klöpffer et al. 2007). As such, Hischier and Walser (2012) note the core issues of the LCI phase is the availability or adequate and comprehensive LCI data from ENMs. LCA of an emerging technology Klöpffer et al. (2007) suggest that it is difficult to apply a “full-spectrum” LCA to emerging technologies due to the lack of detailed knowledge regarding the inputs and output of the system. Nevertheless, there is a general trend to apply LCA to emerging technologies (e.g., solar, wind, bio-fuels) (Klöpffer et al. 2007). In the case of CNTs, Upadhyayula et al. (2012) note that the application of LCA is a challenging task because many of the technologies studied are still emerging; introducing a great degree of uncertainty and complexity into LCA. Obtaining accurate data for emerging technologies can be a challenge because data based on conceptual designs, and assumption about the scaling up of laboratory or pilot scale process may not accurately reflect industrial scale operations. Additionally, early prototypes may undergo several changes during product development and testing that can alter how a product is manufactured and used (Upadhyayula et al. 2012). Furthermore, Upadhyayula et al. (2012, 43) note that one of the greatest challenges when assessing nanoproducts is the “variable nature of manufacturing processes and how subtle differences in the resulting nanocomponents can affect the associated nanoproduct”. Data availability Klöpffer et al. (2007, 6) state that “The main problem with LCA of nanomaterials and nanoproducts is the lack of data and understanding in certain areas.” Studies analysed in Gavankar (2012) and Hischier and Walser (2012) show data on the input side covers energy inputs and in most cases material inputs, yet in the great majority of studies the output side is empty. Gavankar et al. (2012) note that presently available LCI databases, populated with material and product flows, do not distinguish between bulk materials and ENMs. Standard LCIs only require the quantity and chemical composition of material inputs. In some cases, additional characteristics are required for materials such as its isotope, stereo-isomer or valence (Klöpffer et al., 2007). Parameters likely to influence the toxicity of ENMs include the chemical composition, particle size, shape, aspect ratio, crystal structure, surface area, surface chemistry and charge, solubility, as well as adhesive properties and whether the ENM is in a pure form or in a composite (Klöpffer et al. 2007). Furthermore, it is important to know if 34 ENMs change their form during their life cycle, due to aging and other influences such as weather, mechanical stress/pressure, electromechanical fields or catalysis (Klöpffer et al. 2007). Indeed, Gavankar et al. (2012) highlight the lack of information on the production processes, suggesting that it will take substantial effort and time to build an understanding on the behaviour and impact of ENMs in order for the LCI and LCIA to be able to fully address their potential environmental impacts. Data confidentiality Klöpffer et al. (2007) highlight the difficulty in acquiring proprietary information from companies, especially from the producers of materials. In some cases, the exact composition of the ENM is confidential. Hence, the challenge lies in the coordination of a compromise between the interests of industry (confidentiality) and the interests of the LCA community (acquiring data at an appropriately aggregated level). Data quality Hischier and Walser (2012) note that for LCA studies of CNTs, data for some processes in some production technologies varied significantly. For instance, the CED for the production of high-pressure carbon monoxide differed by a factor of approximately 10,000, and up to almost 100,000 for the chemical vapour decomposition process. Additionally, Khanna et al. (2008) suggest that the industrial scale nano-LCA results may be gross overestimates due to expected increases in efficiency over time. Capital goods Klöpffer et al. (2007) highlight the large and energy-consuming capital equipment required to manufacture ENMs (such as lithography and ultra-clean rooms) which can rapidly become out-dated due to the fast pace of progress in the field. Hence, nanotechnology may be a case where the capital goods are required to be included in the system boundaries (Klöpffer et al. 2007). This leads to two issues, the collection of data required to model the impact of capital goods and the problem of allocation due to the multiple products/services than capital goods may provide (Klöpffer et al. 2007). Strategies A case study approach Emerging technologies do not lend themselves to analysis by a complete thorough LCA due to insufficient knowledge regarding the inputs and outputs of the system (Klöpffer et al. 2007). Joshi (2008, 487) highlights the need to “generate comprehensive, transparent, representative, and publicly available data for various process and material developments in nanotechnology that satisfy the data quality requirements outlined under ISO standards for LCA”. Concerning strategies to address the lack of information regarding LCI and LCIA data, Hischier and Walser (2012) identify two opposing strategies: ‘from back to front’, and ‘from font to back’. From back to front: addressing data gaps in the LCIA before “representative and comprehensive” LCIs are established. This strategy is not recommended as following such a 35 trajectory has been evaluated as ineffective (Hischier and Walser 2012). Gavankar et al. (2012) also highlight the critical importance of populating LCI databases with nano-specific information. Furthermore, Kuiken (2009) highlighted that the main obstacle for defining the environmental effects of specific ENMs is the slow advancement of metrology10. Without the ability to monitor emissions and conduct full scale ecosystem and human health effect studies on ENMs, full-scale LCAs cannot be properly performed. From front to back: Hischier and Walser (2012) recommend, as a first step, the collection of inventory data. The authors highlight the need for inventory datasets with a high level of representativeness of certain ENMs. Klöpffer et al. (2007) also suggest that a case-study based approach should be adopted. In following such an approach, LCI databases for specific ENMs and products can be populated with nanomaterial-related input and output flows of emissions and resources, supplementing them with information likely to influence the toxicity of ENMs (detailed above) which are necessary for the characterization step of the LICA (Gavankar et al. 2012). The Swiss Centre for Life Cycle Inventories, Ecoinvent, currently differentiates particulate matter (PM) in the following categories: PM greater than 10 µm (PM>10), PM from 10 to 2.5 µm (PM10) and PM smaller than 2.5 µm (PM2.5) (Frischknecht et al. 2004). Bauer et al. (2008) note that one option would be to differentiate UFPs, i.e. a category for PM 0.1. Two approaches to the selection of appropriate case studies can be identified. Klöpffer et al. (2007) recommend that the selection of case studies should prioritise criteria including the most toxic products, nature of dispersion, high volume production and fate and transport issues. Bauer et al. (2008) suggest the targeting ENMs with a high potential for future industrial application. The authors suggest that for the materials listed in Table 8, the most common production method should be identified and the LCI of these materials should be populated with ENM specific information. 10 The science of measurement; this underpins nanoscience and nanotechnologies as it allows the characterisation of materials not only in terms of dimensions but also in terms of attributes such as electrical properties and mass. (Royal Society 2004) 36 Table 8: Nanomaterials with a high potential for future industrial applications (Reproduced from Bauer et al. (2008, p.915) Material Category Carbon nanomaterials Nanocomposites based Metals and alloys Biological nanomaterials Nano polymers Nano glasses Nanoceramics Application Carbon black, carbon nanotubes, fullerenes, carbon nanofilms, etc. Polymer matrix nanocomposites, ceramic matrix nanocomposites, metal matrix filled with nanopolymer composites Ti, Ti-Al, Ti-transition metal alloy, Mg-Ni, Fe-Cu-Nb-Si-B alloy, Fetransition metal alloy, Al-transition metal alloy, Al, Mg, Al-Mg alloy, nanopowders of noble metals (Ag, Au, Pt, Pb) Protein-based materials, peptides, carbohydrates, virus particles, lipids, DNA, composites Metallic glasses, electrochromics, nanoporous glasses, nanochannel glass materials, photonic glasses, etc. Tungsten carbide, alumina, zirconia, titania, silica, zinc oxide, silicon nitride, magnesia, ferric oxide, ceria, hydroxyapatite (HAP), yttria, silicon carbide, boron nitride, TiC, amorphous silicon nitride, etc. An Input-output approach In order to take account of the issues such as capital goods, Klöpffer et al. (2007) suggest that input-output based LCA (IO-LCA) approaches, as input-output based models consider all elements in a product’s supply chain. This includes direct and indirect purchases required to produce a final product, including capital goods required to produce ENMs themselves and the capital goods required to produce chemical feedstock for ENMs. 4.2.3 Life cycle impact assessment (LCIA) Obstacles The meta-studies conducted by Hischier and Walser (2012) and Gavankar et al. (2012), indicate that the most common impact categories selected in LCA studies of ENMs related to energy consumption (CED and GWP). Due to their size and unique functionality, the properties of ENMs are different from their conventional counterparts (Oberdörster et al. 2005). This may lead to them exhibiting unconventional behaviour, leading to unexpected fate, transport, and toxicity mechanisms in human and ecological systems (Gavankar et al. 2012). Hence, it is difficult to address potential toxic impacts of ENMs on humans and the environment (Klöpffer et al. 2007). Klöpffer et al. (2007, 20) conclude that the UNEP/SETAC framework for toxic impacts “can, in principle, be used for specific impacts caused by nanoparticles and nanoproducts given that (nanomaterial-specific) fate, exposure and effects have been adequately identified” (emphasis added). However at this current point in time, Klöpffer et al. (2007) note that there are limits regarding its application, especially relating to the assessment of toxicity impacts (Klöpffer et al. 2007, 6). At the centre of the ENM discussion is risk associated with the possible release of ultra-fine particles (UFPs) and their potential impact on human health and the environment (Bauer et al. 37 2008). Characterisation of particle emissions is conventionally made according to their particle size (aerodynamic diameter): coarse particles (between 10 and 2.5 µm), fine particles (<2.5 µm) and UFP (<0.1 µm) (Englert 2004). In current impact assessment methods, outdoor emissions of particulate matter (PM) of size <10 and <2.5 μm are assessed, and their contribution to various impacts on the natural environment and human health (e.g., climate change, ozone depletion, acidification, eco-toxicity and human toxicity) is quantified (Hischier and Walser 2012). However, basic research concerning the toxicological effects of UFP and their risks for human health and the environment is still in its infancy and only few data on the safety of ENMs is available (Bauer et al. 2008). Since contribution of UFPs to the various impacts on human health and the environment is not known (Hischier and Walser 2012), LCIA methods such as CML 2001, Eco-Indicator 1999 or Impact 2002 do not cover toxicological effects of UFPs (Bauer et al. 2008). The current understanding of effect mechanisms, dose-response relationships, as well as transport and transformations in the environment may not be sufficient to ascertain a representative characterization of ENMs. (Klöpffer et al. 2007, 19). A significant “retooling” of existing tools, such as USEtox (Rosenbaum et al. 2008), CALtox (McKone and Enoch 2002) and QSAR (Dudek et al. 2006; Puzyn et al. 2009, 2010), would be required for them to account for the intrinsic and extrinsic factors that control the behaviour of ENMs (Gavankar et al. 2012). However, such work is underway. The modification of QSAR (Quantitative structure-activity relationship) to QNAR (quantitative nanostructure–activity relationship) (Fourches et al. 2010), to assess the biological effects of engineered nanoparticles based on their physical, chemical, and geometrical properties has recently been proposed (Gavankar et al. 2012). Additionally, Klöpffer et al. (2007) highlight increased dissipative use of scarce resources (such as indium, used in semiconductors) in nanotechnologies may lead to the need for reaching consensus on a framework for the characterization of abiotic resource depletion. Strategies Klöpffer et al. (2007) note that major efforts are needed (in terms of protocols and practical methodologies for toxicology studies, fate and transport studies and scaling approaches) in order to fully assess the potential risks and environmental impacts of nanoproducts and materials. The authors suggest that in order to assess the impacts from ENMs and nanoproducts, one should wait for the development of approaches used for the regulative RA of which can then be adapted for the comparative assessment of potential impacts in LCA. Klöpffer et al. (2007, p.19) note that the following types of studies still need to be undertaken for ENMs: - “Protocols and practical methodologies for toxicological studies; - Fate and transport studies; and - Scaling studies (i.e., how properties such as surface area, conductivity and magnetism change with the size of the nanomaterial).” Alternatively, Klöpffer et al., (2007) suggst one immediate action could be attempts to define categories of ENMs, based on currently available information, for the specific purpose of 38 LCIA, including categories, such as reactivity, degradability/fate and transport, and ecotoxicity vs. human toxicity. In this case, categorization should address: - Dispersive vs. non-dispersive uses - Chemical composition - Form and structure - Mobility of releases in the environment (air emissions, water release, waste, etc.) at each life cycle stage. Reactivity, fate and transport, and interactions with other sources of environmental impacts should also be addressed. (Klöpffer et al. 2007, 21) Qualitative Screening Approach The fact that there are significant data and methodology gaps does not mean that the impact of ENMs on human health and the environment should be ignored or omitted. Screening approachs have been suggested as an interim device to identify potentially significant issues and explore worst-case scenarios (where ENMs have an impact potential as high as that of the most toxic chemicals (Klöpffer et al. 2007; Gavankar et al. 2012). Scalability Gavankar et al. (2012) propose approach outlined in Figure 12 when faced with the limited data concerning ENMs. Figure 12: Flow chart recommending the nanomaterial assessment path depending on the availability of data (reproduced from Gavankar et al. (2012)) 39 In some cases, for instance some metal and metal oxide ENMs (see Auffan et al (2009)) scalability is known to exist. In these cases, the assessment of ENMs can be based on traditional characterization approaches for bulk materials. The LCI data for the conventional materials can be used in the place of the lacking ENM specific data, and the fate–transport and toxicity assessment approaches can be used in the LCIA (Gavankar et al. 2012). In cases where scalability cannot be established, (when particle size is below the threshold for conventional material properties to be applied), Gavankar et al. (2012, 300) suggest exploring any “empirical relationships, even if simplified, between ENM properties and their impact on human health and environment based on the existing literature, engineering, industrial, and other publicly available data”. Hence, empirical relationships may provide an early input for nano-specific assessments until specific data for ENM fate, transport, and toxicity become available (Gavankar et al. 2012). When no quantitative data exists, qualitative assessment can be performed based on the available information on the ENM and its release pathways. Bauer et al. (2008) and Gavankar et al. (2012) discuss the qualitative approach proposed by Reijnders (2006) which distinguishes between inherently non-dispersive11 and inherently dispersive nanoparticles12. Inherently dispersive nanoparticles may be further classified according to the likelihood of their dispersion in conjunction with their size range (Gavankar et al. 2012). Bauer et al. (2008, p. 916) suggest that “more in-depth toxicological studies about these materials are of crucial interest hence, as long as no such study results are known, these materials have to be avoided as far as possible”. 11 Such as coatings, textiles, ceramics, membranes, composite materials, glass products, prosthetic implants, antistatic packaging, cutting tools, industrial catalysts, a variety of electric and electronic devices including displays, batteries and fuel cells (Royal Society 2004) 12 Such as drugs, personal care products such as cosmetics, quantum dots and some pilot applications in environmental remediation (Royal Society 2004). 40 5 RISK ASSESSMENT OF NANOMATERIALS 5.1 Risk assessment Although some RAs have been conducted for ENMs according to standard RA protocols, Grieger et al. (2012) suggest that all have concluded, due to limited data and the presence of large uncertaitintites, it has not be possoible (based on the currently available information) to complete full RAs for regulatory decicion making. Hence, any results to date should be considered as pleminary results. The authors note that there is a lack of measured exposure data for ENMs, lack of validated exposure estimation models, extensive uncertainties within characterizing ENMs and a lack of (eco)toxicological studies in a variety of species. Hence, it is difficult to complete hazard identification, dose–response and exposure assessments for most ENMs (Grieger et al. 2012). 5.2 The complementarity of RA and LCA The European Commission’s Nanoscience and Nanotechnologies An action plan for 2005-2009 (European Commission 2005) suggests that “Risk assessment related to human health, the environment, consumer and workers should be responsibly integrated at all stages of the life cycle of the technology, starting at the point of conception and including R&D, manufacturing, distribution, use and disposal or recycling “(European Commission 2005, 10). Several authors have recommended the application of both LCA and RA to ENMs. On the one hand, it has been suggested that further efforts should be made for RA to consider life cycle concepts, and on the other hand, LCA should be more risk based when applied to ENMs (Sweet and Strohm 2006; Som et al. 2010). It has also be suggested that the risk of ENMs be analysed at each life cycle stage (Shatkin 2008). RA and LCA are tools that estimate the potential impact of a given substance or product, notwithstanding the fact that they vary in aim, scope, outcomes, strengths and weaknesses (Grieger et al. 2012). The difference in outcomes, strengths and weaknesses of these tools are shown in Table 9. Table 9: Outcomes, strengths and weaknesses of LCA and RA Outcomes Strengths weaknesses LCA Comparative basis Includes impacts from all life cycle stages - Includes a range of impact categories - Avoids in ‘burden shifting’, from one impact category to another and from one life cycle stage to another - requires substantial amounts of data - ineffective in handling uncertainties and lack of data - strong expert knowledge required - 41 RA - Absolute basis - the provision of an absolute assessment of the potential risk for specific settings - the use of worst-case evaluations to help ensure safety to a potential adverse effect - requires substantial amounts of data ineffective in handling uncertainties and lack of data strong expert knowledge required Similarities and differenced between LCA and RA have been identified by Grieger et al. (2012). Similarities include: providing a way of structuring, presenting and evaluating information for environmental decision-making in a life cycle perspective (although possessing different conceptions of life cycle) (Flemström et al. 2004), estimation of exposures and effects from emissions (Olsen and Christensen 2001), both contain methods to characterise uncertainty within the assessments (Evans et al. 2002), and both help in providing information to support decisions in situations of uncertainty (Evans et al. 2002). Differences include: LCA’s focus on the product/service system and RA’s focus on the emissions of a single substance (Christensen and Olsen 2004), different system boundaries and ‘life cycles’ are used (Christensen and Olsen 2004), the results of LCA are comparative whereas the results of RA are absolute (Grieger et al. 2012), and LCA covers a range of environmental impacts whereas RA primarily cover toxicological and (eco)toxicological impacts. Grieger et al. (2012) have identified two main approaches that have been proposed for combining LCA and RA for ENMs: life cycle-based risk assessment (LC-based RA) and risk assessment-complemented life cycle assessment (RA-complemented LCA). The authors note several research articles, frameworks and recommendations that apply these two approaches, most of them describing the use of the RA-complemented LCA. In addition, there are very few concrete case studies in the peer-reviewed literature that have tested and validated these approaches. 5.2.1 Life cycle based risk assessment LC-based RA is an approach that applies traditional RA in a life cycle perspective (Grieger et al. 2012). This is done to help “concentrate efforts where it is the most needed, i.e., at each life-cycle stage” (Grieger et al. 2012, 9), and is considered an extension of standard RA in a life cycle perceptive (as required by REACH). Wardak et al. (2008) recommends using LCA and RA methods based on scenario analyses with expert elicitation, specifically focusing on ENMs or nanoproducts during the use and disposal stages (Grieger et al. 2012). 5.2.2 Risk assessment complemented by life cycle assessment RA-complemented LCA consists of a conventional LCA (assessing the environment impact of a product) complemented with either a quantitative, semi-quantitative or qualitative RA which assesses the risks related to specific life cycle stages (Grieger et al. 2012). The authors note that this approach is the only approach that really combines life cycle and RA based methods for ENM risk. Grieger et al. (2012) note that most publications and risk analysis frameworks utilise this method. 5.2.3 A Stream lined approach Klöpffer et al. (2007) outline a screening approach, combining the use of LCA, RA and scenario analysis, to be used by industries (including SMEs) and stakeholders involved in the development of ENMs and nanoproducts in order to identify the main areas of concern in relation to the potential environmental impacts and to support go/no go decisions. Five steps where highlighted in this screening approach are illustrated in Table 10. 42 Table 10: Proposed Stepwise approach to LCT combined with RA Steps Purpose 1. Check for obvious - Compliance with harm health, safety and environmental regulation 2. Traditional LCA - Understanding without toxicity study burdens versus (focus on environmental benefits impacts) - If substantial benefits, then go forward 3. Toxicity and RA (or qualitative analysis) could include toxicity and risk questions 4. Combine LCA and RA 5. Scenario Analysis What Is Available - Usual assessment methods in industry What Is Missing - Analogies with - Some LCI data on existing materials nanomaterial - Confidential info production available to the “right” - Interface – to be people developed to deal - Software with easy towith fuzzy inputs use interface should this be sector - If material is not listed, or region- specific? use what’s similar - Find ways to make confidential information available within industry - What are the likely - Confidential - Hazard and exposure adverse risks that information available data (potential humans and other to the “right” people primary and organisms will be - Quantitative or fuzzy secondary exposed to at each life - Published information transformation into cycle stage is available unknown toxic - How structure of substances across life material influences cycle stages) behavior (surface, - Find ways to make area, shape, etc.) confidential information available within industry - To assess overall - No standard impacts over whole quantitative tool life cycle available to merge - Can evaluate impacts the data from the interaction of materials - To scale- up to - Lack of reasonable society-wide use upper and lower (consider issues such bounds for scaling as resource depletion) and impact estimations 43 6 SUBSTANCE FLOW ANALYSIS OF NANOMATERIALS 6.1 Research applying SFA and PFA to nanomaterials 6.1.1 Literature review Conducting a comprehensive meta-analysis of the state-of-the-art of SFA research on nanomaterials is beyond the scope of this report. Whilst highlighting some of the current research results, our primary objective is to focus on the potentials and limitations of current research efforts in order to propose research priorities. A search of academic literature databases (Scopus, ScienceDirect) and an internet search for publications (such as those in scientific journals, conference proceedings, conference presentations, research reports and theses) was conducted using the following combination of keywords - nano + substance flow analysis - nano + “substance flow analysis” - nano + SFA - nano + material flow analysis - nano + “material flow analysis” - nano + MFA - nano + particle flow analysis - nano + “particle flow analysis” - nano + PFA Table 11 highlights the peer reviewed scientific literature on nanotechnology and ENMs conducted from a SFA perspective. 44 Table 11: Case studies applying substance flow analysis to nanomateirals Author ENM Arvidsson et al. (2011) System Case Study Ag Generic Arvidsson et al. (2012a) Blaser et al. (2008) TiO2 Generic Ag Water EU Boxall et al. (2007) TiO2, ZnO, CeO2, Al2O3, SiO2, Au, Ag, C60 TiO2, ZnO, Ag, CNT, C60 TiO2, Ag, CNT TiO2, ZnO, Ag, CNT TiO2, Ag, CeO2 CeO2 Water, sludge, soil, air UK PFA of nanosilver emissions from dissipative and non-recyclable products (wound dressings and textiles). PFA of titanium dioxide nanoparticles from sunscreen, paint and cement. Dissolved silver (Ag) emissions from nano-silver containing biocidal products (textiles and plastics) were compared to the expected concentrations in the environment. Nanosilver is only responsible for a small share of the total dissolved silver flow in the environment, but did not consider any particulate emissions. Based on an assumption of 10% market penetrations of nanoproducts and the known usage of these products, concentrations of silver, aluminum oxide and fullerene were predicted to be in the ng/l in wastewaters, whereas nano-TiO2, silica, ZnO and hydroxyapatite were predicted to be in the µg/l range. Water, sludge, air, sediments, soils, groundwater CH, EU, USA Stochastic simulations of the release of all considered ENM to environmental and technical compartments during all life-cycle stages Water, air, soil CH Release of TiO2, Ag, CNT to environmental and technical compartments. Water, air, soil CH The flow of TiO2, ZnO, Ag, CNT during waste incineration and landfilling. Air, surface water Ireland The release of TiO2 from exterior paints, Ag from food packaging and CeO2 from fuel additives Air, Soil Generic The release of CeO2 from fuel additives Gottschalk et al. (2009) Mueller and Nowack (2008) Mueller et al. (2013) O’Brien and Cummins (2010) Park et al. (2008) Compartments 45 6.1.2 Results from selected case studies Release of nanosilver and nano titanium dioxide during the use phase Arvidsson et al. (2011) and Arvidsson et al. (2012a) are outcomes from the Swedish research programme NanoSphere: Centre for interaction and risk studies in Nano-Bio-Geo-Sociotechnosphere interfaces13. This research programme is a cooperation between twelve research groups from three faculties at University of Gothenburg and from Chalmers and funded by FORMAS (25 million SEK from 2010-2014). Arvidsson et al. (2011) and Arvidsson et al. (2012) applied PFA to estimate the global current emissions, and future emissions, of nanosilver (in wound dressings, textiles and nanosilver ink in electronic circuitry) and nano titanium dioxide (in sunscreen, paint and cement) to the environment. Both of these studies focused on the use phase, as the production phase emissions are highly dependent upon individual company’s management practice and there is poor knowledge on the fate of nanosilver during the waste-handling phase. Current and future production of nanosilver and nano titanium dioxide were analysed. The authors developed an exploratory scenario to assess the potential future development of nanosilver and nano titanium dioxide applications. The assumptions behind the nanosilver case were that application will reach 100% market share within the product group in question, the per capita in-flow to the use phase and stock of the product group will be equal to those found in today’s high-income regions, and the world population will increase to 10 billion people by 2050. The assumptions behind the nano titanium dioxide case were that the world average demand per capita for nano titanium dioxide applications are equal to the current demand in developed countries and that the world population will increase to 10 billion people by 2050. Nanosilver The current production of the nanosilver for wound dressings, textiles and electronic circuitry, was suggested to be 254 kg/year, < 4,700 kg/year, and < 4,700 kg/year, respectively. The current and future in-flow, stock and emissions of nanosilver particles are summarised in Table 12. Table 12: Current in-flow, stocks and emissions during the use phase for nanosilver applications in wound dressings, textiles, and electronic circuitry (adapted from Arvidsson et al. (2011)) Output parameter Current production Current stock Current emissions Explorative scenario production Explorative scenario stock Explorative scenario emissions 13 Symbol and unit particles/year Wound dressings 22 4.6 x 10 Textiles 23 < 8.5 x 10 Electronic circuitry 24 < 6.8 x 10 particles particles/year particles/year Insignificant 21 4.6 x 10 22 25 (1 x 10 , 1 x 10 ) Insignificant 23 < 8.5 x 10 28 32 (6 x 10 , 6 x 10 ) < 6.8 x 10 24 < 6.8 x 10 27 9 x 10 particles Insignificant Insignificant 9 x 10 particles/year (1 x 10 , 1 x 10 ) 21 24 http://www.nanosphere.gu.se/ 46 28 25 28 32 (6 x 10 , 6 x 10 ) 27 < 9 x 10 The highest inflow of nanosilver particles to the use phase occurs from their use in electronic circuitry followed by textiles and wound dressings. Due to the short product life of wound dressings, no stock of nanosilver is formed during the use phase. Likewise, there is no stock of nanosilver from its use in textiles, as it is assumed to be emitted during the first few washes. The life span of electronic circuitry was assumed to be 10 years, resulting in a stock of < 6.8 x 1025 nanosilver particles. Whilst the emissions for nanosilver from textiles reported in this study are very uncertain, the authors highlight the importance these emissions during both the current use phase and their potential future emissions. Nano titanium dioxide This study analysed the current and future emissions of titanium dioxide nanoparticles from the use phase of their applications in sunscreens, paints and self-cleaning cement. The current production of nano titanium dioxide used for paint and sunscreen was suggested to be, 29,000 kilotonnes/year and 72 kilotonnes/year, respectively. The production of nano titanium dioxide for self-cleaning cement was reported as being negligible. The current and future in-flow, stock and emissions of titanium dioxide nanoparticles are summarised in Table 13. Table 13: Current in-flow, stocks and emissions during the use phase for titanium dioxide nanoparticle applications in paint, sunscreen, and self-cleaning cement (adapted from Arvidsson et al. (2012)) Scenario Output parameter Symbol and unit Paint Current Inflow to use phase Stock in use Use phase emissions Inflow to use phase Stock in use Use phase emissions particles/year particles particles/year particles/year particles particles/year 1 x 10 26 1 x 10 19 1 x 10 25 7 x 10 26 7 x 10 19 8 x 10 Explorative scenario Sunscreen 25 25 2.6 x 10 Negligible 25 2.6 x 10 26 2 x 10 Negligible 26 2 x 10 Self-cleaning cement Negligible Negligible Negligible 27 9 x 10 28 9 x 10 27 < 9 x 10 The highest inflow of nano titanium dioxide particles to the use phase occurs from their use in sunscreen, even though the mass of nano titanium dioxide in paint applications is more than 400% greater than in sunscreen applications. This is a result of the much smaller nano titanium dioxide particles used in sunscreen applications compared to paint. Regarding the future inflows and stocks of nano titanium dioxide, the most important application is self-cleaning cement, due to the small particle size (almost equal to the size used in sunscreen) and the significant growth potential for this application. The authors note that majority of nano particles in paint and self-cleaning cement are not emitted during the use phase, but will be retaining in the materials when the pass through to the end of life phase. Due to the dissipative nature of sunscreen during the use phase, the authors note the importance of the inflow of sunscreen from an emissions perspective, which is the case for both the current and future scenario. 47 Flows of engineered nanomaterials during waste handling Muller et al. (2013) have used SFA to predict the flows of nano titanium dioxide, nano zinc oxide, nanosilver and carbon nanotubes during waste incineration and landfilling of municipal solid waste and construction waste in Switzerland. The inflows to the system consist of the direct deposit of construction waste to landfills for inert wastes, and the incineration of municipal solid waste and sewage sludge. The incineration process consists of a) burning under oxidisation conditions at around 1000°C, b) flue gas filtration (electrostatic precipitator of bag house filter), c) flue gas scrubbing, and d) waste water treatment for wastewaters from the cooling processes of the bottom ash, the scrubber, and possibly from the acid washing of the fly ash. Depending upon the physio-chemical properties of the ENMs, they may either be a) destroyed by oxidation, melting or volatolisation in the furnace or by dissolution/precipitation in the wastewater treatment plant or in the scrubber, or b) survive incineration be found in either the bottom ash or fly ash, be released into the air or the quench water. Figure 13 shows that the major flows for nano titanium oxide, nano zinc oxide and nanosilver from the incineration process go to landfill as bottom ash. The second most significant flow of ENMs to landfill was via the direct deposition of construction waste. For CNTs, 94% were combusted with insignificant amounts remaining in the system. Very small amounts of nano zinc oxide (< 5 t/year), nanosilver (< 5 t/year), and CNTs(< 100kg /year), were predicted to enter landfills. However, up to 150 t/year of for nano titanium dioxide was predicted to enter landfills. Figure 13: Flows of ENMs during waste disposal shown as a % of the total flow that enters the incineration/landfill system (Reproduced from Muller et al. (2013)) 48 6.2 Potential life cycle release and exposure of nanomaterials Som et al. (2010) note that the assessment of human and environmental exposure is closely linked to the release potential of ENMs during the different life cycle stages. To this end, it is important to know in what form (as ‘free’ ENMs, in an aggregated or agglomerated form, or integrated into a nanometre or micrometre sized material) and in what life cycle stage ENMs can be released. 6.2.1 Production of nanomaterials and manufacture of nanoproducts Som et al. (2010) and Gottschalk and Nowack (2011) note that the greatest likelihood of direct release and exposure to ENMs is during their manufacture. Direct release and exposure to ENMs already occurs and is mainly due to the production and handling of dry powders (see Bello et al. (2008), Han et al. (2008) and Fujitani et al. (2008)). Gottschalk and Nowack (2011) suggest that once the ENMs are released to indoor air it is likely that they will sooner or later enter into the environment. The authors also suggest ENMs may be directly released to the environment through open windows during the improper handling ENMs, from transport accidents and other types of spills. Concerning the production of ENMs, Gottschalk and Nowack (2011) note that recent studies on the direct release to the environment show uniform probability distributions ranging from 0 to 2% of the ENM produced. The authors note that generic worst-case scenario release coefficients for chemicals and the manufacturing process of such chemicals consider that 5% are released to the air, 6% to surface waters before reaching a sewage treatment plant and 0.01% to soils. However, depending on the production and maintenance procedures used, it may be possible that only a negligible release to the environment occurs when closed systems and solvent-free procedures are implemented and all waste from cleaning and maintenance is disposed of as special waste (Gottschalk and Nowack 2011). For the manufacture of nanoproducts, Gottschalk and Nowack (2011) note that recent studies on the direct release to the environment show uniform probability distributions ranging from 0 to 2% of the ENM produced. Generic worst-case scenario release coefficients for formulation of mixtures (not embedded into a matrix) considers that 2.5% are released to the air, 2% to surface waters before reaching a sewage treatment plant and 0.01% to soils (Gottschalk and Nowack 2011). The indirect release of ENMs during the production phase may be via untreated or treated water to rivers (Gottschalk and Nowack 2011). For instance, the production of fullerenes or carbon nanotubes results in the production of a greater proportion of waste that contains a variety of carbon-based structures whose characterisation is not yet available (Gottschalk and Nowack 2011). 49 6.2.2 Use phase The exposure of ENMs during the use phase can result from the intended or unintended release of nanoparticles (Som et al. 2009). The intended release of nanoparticles results from either point sources, such as the use of ENMs in groundwater remediation, or non-point sources, such as the use of ENMs in products such as sunscreens (Som et al. 2009; Gottschalk and Nowack 2011). Som et al. (2009, p, 166) suggest that the exposure of consumers to nonpoint sources could be estimated using “behavioural and anthropometric data, usage statistics, and from the prevalence and manner of integration of ENMs in different product categories”. Furthermore, the magnitude of ENM released via point sources is generally known. The unintended release of nanoparticles results from the use of ENM in products such as nanosilver in textiles (Som et al. 2009; Gottschalk and Nowack 2011). The release of ENMs from products during their use depends on several factors, including the amount of ENMs in the product, how the ENM is embedded in the product, the products life time, and the actual use of the product (Som et al. 2009; Gottschalk and Nowack 2011). Som et al. (2009) suggest that products that have a loose incorporation of ENMs or an intense use will most likely not contain any ENMs at the time of disposal. However, factors such as a low rate of use and strong fixation would increase the likelihood of ENMs entering the disposal phase. Hsu and Chein (2006) have shown the release of nano titanium dioxide from coatings on wood, polymers and tiles, with UV light contributing to an increased release of ENMs. Vorbau et al. (2009) have shown no significant release of nanoparticles from the abrasion of coatings containing nano zinc oxide, and that after abrasion the ENMs were still embedded in larger particles. Blaser et al. (2008) and Benn and Westerhoff (2008) have shown that nanosilver is released in ionic form from plastics and textiles, and as nanoparticles released from washing nanosilver containing textiles. Furthermore, the leaching of nano titanium dioxide to surface water from facades treated with nano titanium dioxide containing paints has been demonstrated (Kaegi et al. 2008). 50 6.2.3 End-of-life phase Although there is little information about the behaviour of ENMs during the end-of-life phase, it is assumed that there is a high risk that ENMs may be released to the environment during recycling or disposal (Royal Society 2004). It is likely that there will be unintentional releases of ENMs to the environment during either the recycling, incineration or landfill of ENMs or via wastewater treatment plants. Hence, incineration plants, landfills and wastewater treatment plants may be important sources for ENM releases to the air, water and soil. Incineration One pathway for ENMs to the environment is to the air via waste incineration plants (Gottschalk et al. 2009). Although modern incineration plants are equipped with multi stage flue gas cleaning systems (including electro filters, flue gas scrubbers, catalytic/NOx/furane/dioxin removal and possibly fabric filters), low concentrations of ENMs may be released to the air (Som et al. 2009). Burtscher et al. (2001) suggest that the concentration of particles less than 100nm is reduced by filters by 99.9% and in subsequent wet filtration by another 95%. Roes et al (2012) calculate that by 2020 approximately 0.5 kg of ENM in plastics will incinerated per ton of waste, equating to 1880 t/a of ENM entering Swiss waste incineration plants as nano-composites. The authors suggest that the concentrations of nano-objects found in the flue gas of waste containing nanocomposites would be 100-10 000 times higher than conventional waste, assuming no EMNs are destroying and all ENMs end up in the flue gas. Landfill Several authors (see Mueller and Nowack (2008) and Gottschalk et al (2010, 2009)) have shown a significant flow of ENMs to landfill, either via deposition of bottom or fly ash, from the incineration of wastewater sludge, or via the direct dumping of construction waste. Som et al. (2009) note that the degradation of nanoproducts containing ENMs in landfills is yet to be studied. Recycling The ability to recycle ENMs or nanostructured materials containing ENMs is uncertain. For some nanoproducts such as lithium batteries with a complete recycling system, no release of ENMs to the environment is expected (Som et al. 2009). A recent experimental study by Busquets-Fité et al. (2013) on the recovery of silicon dioxide, titanium dioxide, zinc oxide and WMCNTs from polyamide-6 (PA) and polypropylene (PP) composites has shown recovery rates of between 0 and 99%, as detailed in Table 14. Table 14: Recovery of ENMs from PA and PP composites ENM silicon dioxide (non-aged – aged) titanium dioxide (non-aged – aged) zinc oxide (non-aged – aged) MWCNT (non-aged – aged) PA 43-59% 60-59% 0-0% 50-45% PP 98-95% 97-96% 99-99% 97-80% However, currently ENMs are not recycled at significantly high rates and recycling process such ‘shredding’ may lead to the release of ENMs. 51 Wastewater treatment Gottschalk and Nowack (2011) suggest that one should expect at least some of the ENMs in wastewater to end up in freshwater. Furthermore, it should also be considered that ENMs may pass through several different technical compartments (the deposit of sludge from waste incineration plants to landfill and/or the incineration of biosolids from wastewater treatment plants) (Gottschalk and Nowack 2011). Arvidsson et al. (2012b) have assessed the risk of silver exposure to earthworms from applying sludge as fertilizer to agricultural land which contains nanosilver from clothing applications. They have shown that low concentrations of silver found in clothes pose an insignificant contribution to the total silver concentration in the waste water treatment plant studied and that the concentration of silver in sludge was below the natural level found in the earth’s upper crust. However, for high concentrations of silver in cloths it would be impossible to reach the long-term goal of having the same concentration of silver in the sludge as in Earth's upper crust. Furthermore, the authors suggest that for clothes with the highest silver concentration, there is a substantial risk that the concentration of silver in the soil would be toxic to earthworms, if the sludge were to be applied to agricultural land. 52 7 COMMUNICATION OF A LIFE CYCLE APPROACH TO NANOMATERIALS There exists a plethora of images used to represent the life cycle concept. They either are generic representations of the life cycle concept, or represent specific information related to the product/system service in question. These images generally belong to two categories, linear representations and cyclical representations. Almost all graphical representations of LCT have one aspect in common; they represent the connection between the life cycle phases of a product (resource extraction, manufacturing/production, transport, use and end-of-life). See Appendix B for generic, linear, general and specific life cycle images. These images are useful to convey the life cycle perspective, however one important aspect related to products containing ENMs that should be communicated is the release of ENMs to the environment during different life cycle stages. Figure 14 highlights the flows and potential releases of ENMs as a result of their incorporation into product life cycles. The blue line represents ENMs in the product life cycle and also the potential emissions of ENMs during a product life cycle. Figure 14: Life cycle thinking and nanomaterials 53 8 RECOMENDATIONS To ensure the safe handling of ENMs and to be able to identify opportunities in a life cycle perspective requires better data and analysis, but also more effective decision-making and policy instruments. The following suggestions identify some potential ways forward:            Improved information concerning the use of ENMs. In order to assess risk, information is needed on the volumes society uses, in which applications, and in what forms. Improved information on emissions is required in order to assess the risks of ENMs. As a first step, information is required on where emissions occur, which can be achieved through undertaking simplified SFAs of ENMs. Methods for this need to be developed where the reasonable worst-case assumptions can be made to assess whether further detailed analysis is required. Those who place a material on the market should be able to describe how the material will be disposed of or emitted to the environment. In depth SFA in specific cases. These cases can be selected for several reasons: environmentally relevant ENMs, ENMs used in large quantities or ENMs that can be considered representative of larger groups and thus can be used to develop and verify the simplified models. Measurements. SFA is based upon existing and available data which in turn need to come from actual measurements or model calculations, which in turn needs to be based on measurements. Examples of important situations where actual measurements are required include exposure in the work environment, flows in waste water treatment plants and flows associated with recovery processes and other waste management activities. Methods for the characterization of nanoparticles. The properties of nanoparticles can change according to their shape and size. Nanoparticles need to be characterised in ways that are relevant for emission measurements, exposure analysis and toxic effects. Toxicological and eco-toxicological dose-response data are needed. Models for exposure analysis require further development and need to be adapted for nanoparticles. Environmental impact assessment methods in LCA require further development and need to be adapted for nanoparticles. As the methods for risk assessment of nanoparticles are developed, there is a need for LCA methodology to follow and adapt. LCI data for ENMs. LCA is heavily dependent on databases which have been developed over the past decade. However, these databases are limited with regards to ENM data. Life cycle inventory data is essential for the assessment of the potential benefits and impacts of ENMs in a life cycle perspective. Methods to develop life cycle data for emerging technologies. Nanotechnology is a field experiencing rapid development; this also applies to manufacturing processes and their environmental performance. International cooperation with a Swedish perspective. Much of the data and methods that are required for LCA should be developed in the context of international cooperation. However, it may be important to develop life cycle data for 54      products manufactured in Sweden as some conditions may be country specific (for example, raw materials and energy). Furthermore, other processes such as waste management may have specific Swedish conditions. The collaboration of industry, governmental agencies and research. Much of the data which is required should be produced by industry. It is also important that governmental agencies and researcher are involved in such work to ensure credibility and transparency. Credible information to users. The safe use of ENMs and nanoproducts requires informed users. Labelling and other forms information is needed to be designed so that users in businesses, organizations, government agencies and consumers can make their own informed decisions. Avoid locking in a risk paradigm. Full risk assessments require copious amounts of data and take a significant amount of time to complete. It would be expensive and inefficient to complete risk assessments on every ENM and its specific application that is placed on the market. Hence, one must be able to make effective decisions about the safe use of ENMs without full risk assessments. Avoid a ‘material for material’ paradigm. The number of ENMs can be vast. In order to have effective processes, decisions can be taken without the complete data that is require for each individual material. Decisions can be made for groups of materials, or based on more simple criteria. Resources for research in several fields. There is need for research on data and methods that can be used for SFA, RA and LCA. Research is also needed on the use of ENMs, policy instruments and decision theory. 55 9 CONCLUSIONS ENMs are used in a growing number of products. Their application in products can be either inherently non-dispersive or inherently dispersive. Yet even ENMs in inherently nondispersive applications can be released to the environment during the life cycle of the product. The present assessment of the impacts, or benefits, of ENMs upon human health and the environmental is currently inadequate. One of the most important contributing causes for this inadequacy is the lack of data. Although there are a large variety of ENMs currently being used, there are no official statistics available on the amounts of ENMs currently used and products that contain ENMs. Environmental and health risks are both related to the chemical composition of the ENMs, but also the particles size, shape and properties. Hence, nanoparticles must be classified according to more than their mere chemical composition. There are major gaps in knowledge regarding the emission of ENMs and nanoparticles during production, use and disposal. Models for dispersion and exposure analysis for ENMs must be developed as well as dose-response data for toxic effects. Full RAs of ENMs are difficult because of the lack of data and methods available. LCAs have been completed for a number of products containing ENMs. ENMs are not considered in the LCIA, hence there is no information presented concerning the environmental impact on human health or the environment due to the release of ENMs. Production of ENMs can often be energy intensive. However, in a life cycle perspective, the use of ENMs may lead to reduced energy use that is greater than that caused by the production. To be able to both reach the safe use of ENMs and to exploit ENMs opportunities in a life cycle perspective requires better data and analysis but also effective instruments and decision-making. 56 10 REFERENCES Ackoff, R. 1973. Science in the systems age: beyond IE, OR, and MS. Operations Research 21(3 (May - Jun., 1973)): 661–671. Agboola, A.E. 2005. Development and model formulation of scalable carbon nanotube processes: HiPCO and CoMoCAT process models. Department of Chemical Engineering/Agricultural and Mechanical College. Louisiana State University. Arvidsson, R. 2012. Contributions to Emission, Exposure and Risk Assessment of Nanomaterials. Chalmers University of Technology. Arvidsson, R., S. Molander, and B.A. Sandén. 2011. Impacts of a Silver Coated Future: Particle Flow Analysis of Silver nanoparticles. Journal of Industrial Ecology 15(6): 844–854. Arvidsson, R., S. Molander, and B.A. Sandén. 2012a. Particle Flow Analysis: Exploring Potential Use Phase Emissions of Titanium Dioxide Nanoparticles from Sunscreen, Paint, and Cement. Journal of Industrial Ecology 16(16): 343–351. Arvidsson, R., S. Molander, and B.A. Sandén. 2012b. Assessing the Environmental Risks of Silver from Clothes in an Urban Area. Human and Ecological Risk Assessment: An International Journal Article Article in Press. Auffan, M., J. Rose, J.-Y. Bottero, G. V Lowry, J.-P. Jolivet, and M.R. Wiesner. 2009. Towards a definition of inorganic nanoparticles from an environmental, health and safety perspective. Nature Nanotechnology 4(10): Babaizadeh, H. and M. Hassan. 2013. Life cycle assessment of nano-sized titanium dioxide coating on residential windows. Construction and Building Materials 40(March): 314–321. Bauer, C., J. Buchgeister, R. Hischier, W.R. Poganietz, L. Schebek, and J. Warsen. 2008. Towards a framework for life cycle thinking in the assessment of nanotechnology. Journal of Cleaner Production 16(8): 910–926. Baumann, H. and A.M. Tillman. 2004. The Hitch Hiker’s Guide to LCA. Lund: Studentlitteratur. Bello, D., A. Hart, K. Ahn, M. Hallock, N. Yamamoto, E. Garcia, M.J. Ellenbecker, and B. Wardle. 2008. Particle exposure levels during CVD growth and subsequent handling of verticallyaligned carbon nanotube films. Carbon 46(6): 974–977. Benn, T. and P. Westerhoff. 2008. Nanoparticle silver released into water from commercially available sock fabrics. Environmental Science & Technology 42(11): 4133-4139. Blaser, S.A., M. Scheringer, M. MacLeod, and K. Hungerbühler. 2008. Estimation of cumulative aquatic exposure and risk due to silver: Contribution of nano-functionalized plastics and textiles. Science of The Total Environment 390(2): 396–409. British Standards Institution. 2007. Terminology for Nanomaterials. London. Burtscher, H., M. Zürcher, A. Kasper, and M. Brunner. 2001. Efficiency of flue gas cleaning in waste incineration for submicron particles. In Proc. Int. ETH Conf. on Nanoparticle Measurement., ed. A. Mayer. BUWAL. Busquets-Fité, M., E. Fernandez, G. Janer, G. Vilar, S.V.-C.R. Zanasca, C. Citterio, L. Mercante, and V. Puntes. 2013. Exploring release and recovery of nanomaterials from commercial polymeric nanocomposites. Journal of Physics: Conference Series 429(012048). Checkland, P. 1999. Systems thinking, systems practice: includes a 30-year retrospective. 57 Christensen, F.M. and S.I. Olsen. 2004. The potential role of life cycle assessment in regulation of chemicals in the European union. The International Journal of Life Cycle Assessment 9(5): 327–332. Dekkers, S., L.C.H. Prud’homme De Lodder, R. de Winter, A.J.A.M. Sips, and W.H. de Jong. 2007. Inventory of consumer products containing nanomaterials RIVM/SIR Advisory report 11124. Dudek, A., T. Arodz, and J. Galvez. 2006. Computational methods in developing quantitative structure-activity relationships (QSAR): a review. Combinatorial chemistry & high throughput screening 9(3), 213-228 Englert, N. 2004. Fine particles and human health--a review of epidemiological studies. Toxicology Letters 149(1-3): 235–42. European Commission. 2004. Towards a European strategy for nanotechnology. Brussels. European Commission. 2005. Nanosciences and nanotechnologies: An action plan for Europe 20052009. Brussels. European Commission. 2009a. Preparing for our future: Developing a comm on strategy for key enabling technologies in the EU COM(2009) 512 final. Brussels. European Commission. 2009b. Nanosciences and Nanotechnologies: An action plan for Europe 2005-2009. Second Implementation Report 2007-2009. Brussels. European Commission. 2009c. Accompanying document to the Nanosciences and Nanotechnologies: An action plan for Europe 2005-2009. Second Implementation Report 20072009. Brussels. European Commission. 2010. Life Cycle Thinking and Assessment - Our thinking - life cycle thinking. European Commission Institute for Environment and Sustainability. European Commission. 2011. Commission Recommendation of 18 October 2011 on the defition of nanomaterial (2011/696/EU). Official Journal of the European Union L275/38-L2. European Commission. 2012. COMMISSION STAFF WORKING PAPER Types and uses of nanomaterials, including safety aspects SWD(2012) 288 final. Brussels. Evans, J., P. Hofstetter, and T. McKone. 2002. Introduction to special issue on life cycle assessment and risk analysis. Risk Analysis Analysis 22(5): 819–820. Figueirêdo, M.C.B. de, M. de F. Rosa, C.M.L. Ugaya, M. de S.M. de Souza Filho, A.C.C. da Silva Braid, and L.F.L. de Melo. 2012. Life cycle assessment of cellulose nanowhiskers. Journal of Cleaner Production 35: 130–139. Finnveden, G., M.Z. Hauschild, T. Ekvall, J. Guinée, R. Heijungs, S. Hellweg, A. Koehler, D. Pennington, and S. Suh. 2009. Recent developments in life cycle assessment. Journal of Environmental Management 91(1): 1–21. Finnveden, G. and Å. Moberg. 2005. Environmental systems analysis tools–an overview. Journal of Cleaner Production 13(12): 1165–1173. Fleischer, T. and A. Grunwald. 2008. Making nanotechnology developments sustainable. A role for technology assessment? Journal of Cleaner Production 16(8): 889–898. Flemström, K., R. Carlson, and M. Erixon. 2004. Relationships between Life Cycle Assessment and Risk Assessment:– Potentials and Obstacles. Gothernburg. Foss Hansen, S., B.H. Larsen, S.I. Olsen, and A. Baun. 2007. Categorization framework to aid hazard identification of nanomaterials. Nanotoxicology 1(3): 243–250. 58 Fourches, D., D. Pu, C. Tassa, R. Weissleder, S.Y. Shaw, R.J. Mumper, and A. Tropsha. 2010. Quantitative nanostructure-activity relationship modeling. ACS Nano 4(10): 5703–12. Frischknecht, R., N. Jungbluth, H.-J. Althaus, G. Doka, R. Dones, T. Heck, S. Hellweg, et al.. 2004. The ecoinvent Database: Overview and Methodological Framework. The International Journal of Life Cycle Assessment 10(1): 3–9. Fthenakis, V., S. Gualtero, R. van der Meulen, and H.C. Kim. 2008. Comparative Life-cycle Analysis of Photovoltaics Based on Nano-materials: A Proposed Framework. In Materials Research Society Symposium Proceeding, 1041:R01–03. Fthenakis, V., H.C. Kim, S. Gualtero, and A. Bourtsalas. 2009. Nanomaterials in PV manufacture: Some life cycle environmental-and health-considerations. In Photovoltaic Specialists Conference (PVSC), 2009 34th IEEE, 2003–2008. IEEE. Fujitani, Y., T. Kobayashi, K. Arashidani, N. Kunugita, and K. Suemura. 2008. Measurement of the physical properties of aerosols in a fullerene factory for inhalation exposure assessment. Journal of Occupational and Environmental Hygiene 5(6): 380–9. Gavankar, S., S. Suh, and A.F. Keller. 2012. Life cycle assessment at nanoscale: review and recommendations. The International Journal of Life Cycle Assessment 17: 295–303. Gottschalk, F. and B. Nowack. 2011. The release of engineered nanomaterials to the environment. Journal of Environmental Monitoring 13(5): 1145–1155. Gottschalk, F., R.W. Scholz, and B. Nowack. 2010. Probabilistic material flow modeling for assessing the environmental exposure to compounds: Methodology and an application to engineered nano-TiO2 particles. Environmental Modelling & Software 25(3): 320–332. Gottschalk, F., T. Sonderer, R.W. Scholz, and B. Nowack. 2009. Modeled environmental concentrations of engineered nanomaterials (TiO2, ZnO, Ag, CNT, fullerenes) for different regions. Environmental Science & Technology 43(24): 9216–9222. Greijer, H., L. Karlson, S.E. Lindquist, and A. Hagfeldt. 2001. Environmental aspects of electricity generation from a nanocrystalline dye sensitized solar cell system. Renewable Energy 23(1): 27–39. Grieger, K.D., A. Laurent, M. Miseljic, F. Christensen, A. Baun, and S.I. Olsen. 2012. Analysis of current research addressing complementary use of life-cycle assessment and risk assessment for engineered nanomaterials: have lessons been learned from previous experience with chemicals? Journal of Nanoparticle Research 14(7): 1–23. Griffiths, O. and J. O’Byrne. 2013. Identifying the Largest Environmental Life Cycle Impacts during Carbon Nanotube Synthesis via Chemical Vapour Deposition. Journal of Cleaner Production 42: 180–189. Grubb, G.F. 2010. Improving the environmental performance of manufacturing systems via exergy, techno-ecological synergy, and optimization. The Ohio State University. Grubb, G.F. and B.R. Bakshi. 2008. Energetic and environmental evaluation of titanium dioxide nanoparticles. In Electronics and the Environment, 2008. ISEE 2008, 1–6. IEEE. Grubb, G.F. and B.R. Bakshi. 2011a. Life Cycle of Titanium Dioxide Nanoparticle Production. Journal of Industrial Ecology 15(1): 81–95. Grubb, G.F. and B.R. Bakshi. 2011b. Appreciating the role of thermodynamics in LCA improvement analysis via an application to titanium dioxide nanoparticles. Environmental Science & Technology 45(7): 3054–3061. 59 Han, H.J., E.J. Lee, J.H. Lee, K. Pyo So, Y. Hee Lee, G. Nam Bae, S.-B. Lee, J. Ho Ji, M.H. Cho, and I. Je Yu. 2008. Monitoring Multiwalled Carbon Nanotube Exposure in Carbon Nanotube Research Facility. Inhalation Toxicology 20(8): 741–749. Handy, R.D., F. von der Kammer, J.R. Lead, M. Hassellöv, R. Owen, and M. Crane. 2008. The ecotoxicology and chemistry of manufactured nanoparticles. Ecotoxicology (London, England) 17(4): 287–314. Hauschild, M.Z. 2005. Assessing environmental impacts in a life-cycle perspective. Environmental Science & Technology 39(4): 905–912. Healy, M.L., L.J. Dahlben, and J.A. Isaacs. 2008. Environmental Assessment of Single-Walled Carbon Nanotube Processes. Journal of Industrial Ecology 12(3): 376–393. Healy, M.L., A. Tanwani, and J.A. Isaacs. 2006. Economic and Environmental Tradeoffs in SWNT Production. NSTI-Nanotech, Nano Science and Technology Institute, Boston (MA, USA). Hischier, R. and T. Walser. 2012. Life cycle assessment of engineered nanomaterials: State of the art and strategies to overcome existing gaps. Science of The Total Environment 425(15 May 2012): 271–282. Hsu, L.-Y. and H.-M. Chein. 2006. Evaluation of nanoparticle emission for TiO2 nanopowder coating materials. Journal of Nanoparticle Research 9(1): 157–163. Illuminato, I. and G. Miller. 2010. Nanotechnology, climate and energy: over-heated promises and hot air? Isaacs, J., A. Tanwani, and M.L. Healy. 2006. Environmental Assessment of SWNT Production. In Proceedings of the 2006 IEEE International Symposium on Electronics & the Environment, 38 – 41. Scottsdale, USA: IEEE; 2006. Isaacs, J.A., A. Tanwani, M.L. Healy, and L.J. Dahlben. 2010. Economic assessment of single-walled carbon nanotube processes. Journal of Nanoparticle Research 12(2): 551–562. ISO. 2006a. Environmental Management - Life cycle assessment - Requirements and guidelines ISO 14044:2006. ISO. 2006b. Environmental Management - Life cycle assessment - Principles and Framework (ISO 14040:2006). ISO. 2008. ISO/TS 27687:2008 Nanotechnologies- Terminology and definitions for nano objects— nanoparticle, nanofibre and nanoplate. Joshi, S. 2008. Can Nanotechnology Improve the Sustainability of Biobased Products? The Case of Layered Silicate Biopolymer Nanocomposites. Journal of Industrial Ecology 12(3): 474–489. Jovanovic, A. and M. Cordella. 2011. Life cycle Assessment (LCA) & Risk Analysis in Nanomaterials related NMP projects. Special Brainstorming and Coordination Meeting March 2, 2001, Brussels. JRC. 2010. International Reference Life Cycle Data System (ILCD) Handbook. General guide for Life Cycle Assessment - Detailed guidance. Vol. First Edit. Ispra: Joint Research Centre, Institute for Environment and Sustainability, Eurpean Commission. Kaegi, R., A. Ulrich, B. Sinnet, R. Vonbank, A. Wichser, S. Zuleeg, H. Simmler, et al.. 2008. Synthetic TiO2 nanoparticle emission from exterior facades into the aquatic environment. Environmental Pollution 156(2): 233–239. 60 Khanna, V. 2009. Environmental and Risk Assessment at Multiple Scales with Application to Emerging Nanotechnologies. Ohio State University. Khanna, V. and B.R. Bakshi. 2009. Carbon nanofiber polymer composites: evaluation of life cycle energy use. Environmental Science & Technology 43(6): 2078–2084. Khanna, V., B.R. Bakshi, and L.J. Lee. 2007. Life cycle energy analysis and environmental life cycle assessment of carbon nanofibers production. In Electronics & the Environment, Proceedings of the 2007 IEEE International Symposium, 128–133. IEEE. Khanna, V., B.R. Bakshi, and L.J. Lee. 2008a. Carbon nanofiber production: Life Cycle Energy Consumption and Environmental Impact. Journal of Industrial Ecology 12(3): 394–410. Khanna, V., Y. Zhang, G. Grubb, and B.R. Bakshi. 2008b. Assessing the Life Cycle Environmental Implications of Nanomanufacturing: Opportunities and Challenges. In Nanoscience and Nanotechnology: Environmental and Health Impacts, ed. H Grassian, 19–42. New Jersey: John Wiley & Sons. Klöpffer, W., M.A. Curran, P. Frankl, R. Heijungs, A. Köhler, and S.I. Olsen. 2007. Nanotechnology and Life Cycle Assessment. A systems approach to Nanotechnology and the environment: Synthesis of Results Obtained at a Workshop Washington, DC 2–3 October 2006. tech. rep., European Commission, DG Research, jointly with the Woodrow Wilson International Center for Scholars. Köhler, A.R., C. Som, A. Helland, and F. Gottschalk. 2008. Studying the potential release of carbon nanotubes throughout the application life cycle. Journal of Cleaner Production 16(8): 927– 937. Kuiken, T. 2009. It’s Time to Move Forward on LCA of Nanomaterials. In Nanotechnology & Life Cycle Analysis Workshop; Chicago, IL (USA). Kushnir, D. and B.A. Sandén. 2008. Energy requirements of carbon nanoparticle production. Journal of Industrial Ecology 12(3): 360–375. Linkov, I. and T.P. Seager. 2011. Coupling multi-criteria decision analysis, life-cycle assessment, and risk assessment for emerging threats. Environmental Science & Technology 45(12): 5068– 5074. Lloyd, S.M. 2004. Using Life Cycle Assessment to inform nanotechnology research and development. Engineeing and Public Policy. Pittsburgh, Pennsylvania: Carnegie Mellon University, Carnegie Institute of Technology. Lloyd, S.M. and L.B. Lave. 2003. Life cycle economic and environmental implications of using nanocomposites in automobiles. Environmental Science & Technology 37(15): 3458–3466. Lloyd, S.M., L.B. Lave, and H.S. Matthews. 2005. Life cycle benefits of using nanotechnology to stabilize platinum-group metal particles in automotive catalysts. Environmental Science & Technology 39(5): 1384–1392. Lövestam, G., H. Rauscher, G. Roebben, B. Sokull Klüttgen, N. Gibson, J.-P. Putaud, and S. Hermann. 2010. Considerations on a definition of nanomaterial for regulatory purposes. Joint Research Centre (JRC) Reference Reports, 80004-1. Lubick, N. 2008. Risks of Nanotechnology Remain Uncertain. Environmental Science & Technology 42(6): 1821–1824. Luoma, S. 2008. Silver nanotechnologies and the environment: Old problems or new challenges? Washington, DC. 61 Maynard, A. 2011. Don’t define nanomaterials. Nature 475(7354), 31-31. McKone, T. and K. Enoch. 2002. CalTOX (registered trademark), a multimedia total exposure model spreadsheet user’sguide. version 4.0. Merugula, L.A., V. Khanna, and B.R. Bakshi. 2010. Comparative life cycle assessment: Reinforcing wind turbine blades with carbon nanofibers. In Proceedings of the 2010 IEEE International Symposium on Sustainable Systems and Technology, 1–6. IEEE, May. Meyer, D.E., M.A. Curran, and M.A. Gonzalez. 2011. An examination of silver nanoparticles in socks using screening-level life cycle assessment. Journal of Nanoparticle Research 13(1): 147–156. Miljödepartementet. 2012. Kommittédirektiv En nationell handlingsplan för säker användning och hantering av nanomaterial (Dir. 2012:89). Moign, A., A. Vardelle, N.J. Themelis, and J.G. Legoux. 2010. Life cycle assessment of using powder and liquid precursors in plasma spraying: The case of yttria-stabilized zirconia. Surface and Coatings Technology 205(2): 668–673. Mueller, N.C., J. Buha, J. Wang, A. Ulrich, and B. Nowack. 2013. Modeling the flows of engineered nanomaterials during waste handling. Environmental Science: Processes & Impacts 15: 251– 259. Mueller, N.C. and B. Nowack. 2008. Exposure modeling of engineered nanoparticles in the environment. Environmental Science & Technology 42(12): 4447–4453. O’Brien, N. and E. Cummins. 2010. Nano-scale pollutants: Fate in Irish surface and drinking water regulatory systems. Human and Ecological Risk Assessment 16(4): 847–872. Oberdörster, G., A. Maynard, K. Donaldson, V. Castranova, J. Fitzpatrick, K. Ausman, J. Carter, et al.. 2005. Principles for characterizing the potential human health effects from exposure to nanomaterials: elements of a screening strategy. Particle and Fibre Toxicology 2(1): 8. OECD. 2011. National Acitivities on Life Cycle Assessment of Nanomaterials. Paris. Olsen, S. and F. Christensen. 2001. Life cycle impact assessment and risk assessment of chemicals—a methodological comparison. Environmental Impact Assessment Review 21(4), 385-404. Osterwalder, N., C. Capello, K. Hungerbühler, and W.J. Stark. 2006. Energy consumption during nanoparticle production: How economic is dry synthesis? Journal of Nanoparticle Research 8(1): 1–9. Puzyn, T., A. Gajewicz, D. Leszczynska, and J. Leszczynski. 2010. Nanomaterials–the Next Great Challenge for Qsar Modelers. In Recent Advances in QSAR Studies Methods and Applications, ed. Tomasz Puzyn, Jerzy Leszczynski, and Mark T. Cronin. Dordrecht Heidelberg London New York: Puzyn, T., D. Leszczynska, and J. Leszczynski. 2009. Toward the Development of “Nano‐QSARs”: Advances and Challenges. Small 5(22): 2494–2509. Reijnders, L. 2006. Cleaner nanotechnology and hazard reduction of manufactured nanoparticles. Journal of Cleaner Production 14(2): 124–133. Renn, O. 2008. Risk governance: coping with uncertainty in a complex world. London: Earthscan. Roes, A.L., E. Marsili, E. Nieuwlaar, and M.K. Patel. 2007. Environmental and cost assessment of a polypropylene nanocomposite. Journal of Polymers and the Environment 15(3): 212–226. 62 Roes, A.L., L.B. Tabak, L. Shen, E. Nieuwlaar, and M.K. Patel. 2010. Influence of using nanoobjects as filler on functionality-based energy use of nanocomposites. Journal of Nanoparticle Research 12(6): 2011–2028. Roes, L., M.K. Patel, E. Worrell, and C. Ludwig. 2012. Preliminary evaluation of risks related to waste incineration of polymer nanocomposites. The Science of the Total Environment 417418: 76–86. Rosenbaum, R., T. Bachmann, L. Swirsky Gold, M.A.J. Huijbregts, O. Jolliet, R. Juraske, A. Koehler, et al.. 2008. USEtox—the UNEP-SETAC toxicity model: recommended characterisation factors for human toxicity and freshwater ecotoxicity in life cycle impact assessment. The International Journal of Life Cycle Assessment 13(7): 532–546. Royal Society. 2004. Nanoscience and nanotechnologies: opportunities and uncertainties. London. Salamanca-Buentello, F., D.L. Persad, E.B. Court, D.K. Martin, A.S. Daar, and P.A. Singer. 2005. Nanotechnology and the developing world. PLoS Medicine 2(5): 0383–0386. SCENIHR. 2009. Risk assessment of products of nanotechnologies. European Commission Health and Consumer Protection Directorate-General, DirectorateC— public health and risk assessment, C7—risk assessment. Brussels. SCENIHR. 2010. Scientific Basis for the Definition of the Term “Nanomaterial.” Seager, T.P. and I. Linkov. 2008. Coupling multicriteria decision analysis and life cycle assessment for nanomaterials. Journal of Industrial Ecology 12(3): 282–285. Şengül, H. and T.L. Theis. 2011. An environmental impact assessment of quantum dot photovoltaics (QDPV) from raw material acquisition through use. Journal of Cleaner Production 19(1): 21–31. Shatkin, J. 2008. Informing environmental decision making by combining life cycle assessment and risk analysis. Journal of Industrial Ecology 12(3): 278–281. Singh, A., H.H. Lou, R.W. Pike, A. Agboola, X. Li, J.R. Hopper, and C.L. Yaws. 2008. Environmental impact assessment for potential continuous processes for the production of carbon nanotubes. American Journal of Environmental Sciences 4(5): 522–534. Som, C., M. Berges, Q. Chaudhry, M. Dusinska, T.F. Fernandes, S.I. Olsen, and B. Nowack. 2010. The importance of life cycle concepts for the development of safe nanoproducts. Toxicology 269(2): 160–169. Som, C., N.C. Mueller, T. Sonderer, F. Gottschalk, R. Scholz, and B. Nowack. 2009. Exposure modeling of engineered nanoparticle. In Nanotechnology 2009: Fabrication, Particles, Characterization, MEMS, Electronics and Photonics. Cambridge, Massachusetts,: NSTI, Nanoscience and Technology Inst. Steinfeldt, M., A. von Gleich, and U. Petschow. 2004a. Nachhaltigkeitseffekte durch Herstellung und Anwendung nanotechnischer Produkte. Institut für ökologische Wirtschaftsforschung GmbH, Berlin (Germany). Steinfeldt, M., U. Petschow, R. Haum, and A. von Gleich. 2004b. Nanotechnology and sustainability. Schriftenreihe Des IÖW 167: 3. Sweet, L. and B. Strohm. 2006. Nanotechnology—life-cycle risk management. Human and Ecological Risk Assessment 12(3): 528–551. 63 Tenner, E. 2001. Nanotechnology and Unintended Consequences. In Societal Implications of Nanoscience and Nanotechnology, ed. Mihail C. Roco and William Sims Bainbridge, 241–246. Arlington, Virginia. UNEP. 2007. Life Cycle Management: A Business Guide to Sustainability. Paris. UNESCO. 2006. The Ethic and Politics of Nanotechnology. Paris. Upadhyayula, V.K.K., D.E. Meyer, M.A. Curran, and M.A. Gonzalez. 2012. Life cycle assessment as a tool to enhance the environmental performance of carbon nanotube products: a review. Journal of Cleaner Production 26(May 2012): 37–47. Voet, E. van der. 2002. Substance flow analysis methodology. In A Handbook of Industrial Ecology, ed. R U Ayres and L W Ayres. Cheltenham, UK: Edward Elgar. Voet, E. van der, L. van Oers, J.B. Guinée, and H.A. de Haes. 1999. Using SFA indicators to support environmental policy. Environmental Science and Pollution Research International 6(1): 49–58. Vorbau, M., L. Hillemann, and M. Stintz. 2009. Method for the characterization of the abrasion induced nanoparticle release p into air from surface coatings. Journal of Aerosol Science 40(3): 209–217. h Walser, T., E. Demou, D.J. Lang, and S. Hellweg. 2011. Prospective environmental life cycle assessment of nanosilver T-shirts. Environmental Science & Technology 45(10): 4570–4578. Walser, T., L.K. Limbach, R. Brogioli, E. Erismann, L. Flamigni, B. Hattendorf, M. Juchli, et al.. 2012. Persistence of engineered nanoparticles in a municipal solid-waste incineration plant. Nature Nanotechnology 7(8): 520–4. Woodrow Wilson International Center for Scholars. 2011. Project on Emerging Nanotechnologies. 64 Appendix A A.1 European Union FP7 Project NANOSUSTAIN Development of sustainable solutions for nanotechnologybased products based on hazard characterisation and LCA Duration 2010-05-01 2013-04-30 Link Description Life Cycle Related Publications http://www.nanosustain.eu WP4 - life cycle assessment and prospective technological assessment developing methods for extrapolation and scaling-up of processes of engineered nanoparticles; developing specific exposure models for engineered nanoparticles; assessing positive and negative effects on the environment during different life cycle stages of selected nanoproducts; developing criteria and guiding principles to foster the precautionary design of ENMs and guidelines for improved recyclability; testing these guidelines to explore new solutions for the sustainable use, recycling and final treatment of selected ENMs. PROSUITE Development and application of standardized methodology for the PROspective SUstaInability assessment of Technologies 2009-11-01 – 2013-10-31 http://prosuite.org The project goal is “to develop a coherent, scientifically sound, and broadly accepted methodology for the sustainability assessment of current and future technologies over their life cycle, applicable to different stages of maturity”. It is noted that the PROSUITE framework and software tools address the whole life cycle (from a technology’s use of primary materials through to the production and handling of wastes). - Steinfeldt, M. 2012. LCA case studies of nanotechnology based applications in the project. Conference on safe production and use of nanomaterials: Nanosafe 2012. Grenoble, France. 13 ‐ 15 November 2012. - Steinfeldt, M. 2013 Life Cycle Assessment of nanotechnology based applications. 2nd. QNano International Conference. Prague, Czech Republic. 27 February – 1 March 2013 - Steinfeldt, M. LCA case studies of nanotechnology-based applications in the project NanoSustain. Safety Issues and Regulatory Challenges of Nanomaterials. San Sebastian, Spain. rd th 3 - 4 May 2012. - Walser, et al. 2011. Prospective environmental life cycle assessment of nanosilver T-shirts. Environmental Science & Technology 45, 4570– 4578. - Walser, T; et al., 2012. Persistence of engineered nanoparticles in a municipal solid waste incineration plant. Nature Nanotechnology 2012, 7, 520–524 - Hischier, R &Walser, T, 2012. Life cycle assessment of engineered nanomaterials: State of the art and strategies to overcome existing Work packaging 6 contains four case studies. One focuses on nanotechnology and covers specific application such as: Antimicrobial nanoparticles and nanostructured particles in textiles with particular focus on occupational and consumer exposure. 65 Polymer nanocomposites as new engineering materials Electronic devices such as organic light emitting diodes, field effect transistors and organic photovoltaics. The main objective is the monitoring of the life cycle of three families of nanomaterials (carbon nanotubes, nanoclays and metal oxide nanoparticles) when embedded in selected polymeric hosts. - http://www.nanopolytox.eu NANOPOLYTOX Toxicological impact of nanomaterials derived from processing, weathering and recycling from polymer nanocomposites used in various industrial applications NANOHOUSE Life Cycle of Nanoparticlebased Products used in House Coating. 2010-05-012013-04-30 2010-01-012013-06-30 http://wwwnanohouse.cea.fr NANEX Development of Exposure Scenarios for Manufactured Nanomaterials ENFIRO Life Cycle Assessment of 2009-12-012010-11-30 http://nanex-project.eu 2009-09-012012-11-30 http://www.enfiro.eu Work Package 6 will study the influence of the processing and recycling of nanomaterials and the weathering of nanocomposites demonstrators, on the physical, chemical and toxicological properties of the nanofillers. Predictive models will be also developed that will be able to provide the needed information about the evolution of nanomaterials properties along their life cycle. Life Cycle Impact Assessment (LCIA) will be performed based on these predictive models. NanoHouse project covers the whole risk assessment by evaluating the exposure and the hazard. Through a combination of knowledge from Life Cycle Thinking and risk assessment, NanoHouse outlines a holistic and prospective overview on the potential Environmental Health and Safety (EHS) impacts of paints containing Engineered NanoParticles (ENPs) throughout all life stages of the paints. Objective4 of NANEX is to collect and review data on environmental release, risk management measures, and existing models for estimating environmental release and exposure during the various life cycle stages of MNMs for HARNs, mass-produced MNMs and specialised MNMs LCA case studies on the substitution options for specific brominated flame retardants, which includes nanoclaybased flame retardants in printed circuit boards. 66 gaps. Science of the Total Environment 2012, 425, 271-282 Martí Busquets-Fité et al. 2013. Exploring release and recovery of nanomaterials from commercial polymeric nanocomposites . Journal of Physics: Conference Series 429 012048 - Hischier, R &Walser, T, 2012. Life cycle assessment of engineered nanomaterials: State of the art and strategies to overcome existing gaps. Science of the Total Environment 2012, 425, 271-282 - Gottschalk, F., Nowack, B., 2011. The release of engineered nanomaterials to the environment. Journal of Environmental Monitoring 13, 1145– 1155. EnvironmentCompatible Flame Retardants: Prototypical Case Study Nano Impact Net European Network on the Health and Environmental Impact of Nanomaterials NANOMICEX Mitigation of risk and control of exposure in nanotechnology based inks and pigments NanoValid Development of reference methods for hazard identification, risk assessment and LCA of engineered nanomaterials LICARA Life cycle approach and 2008-04-012012-03-31 www.nanoimpactnet.eu 2012-04-012015-03-31 http://nanomicex.eu 2011-11-012015-10-31 http://www.nanovalid.eu NanoImpactNet was a multidisciplinary European network on the health and environmental impact of nanomaterials. NanoImpactNet established to create a scientific basis to ensure the safe and responsible development of engineered nanoparticles and nanotechnology-based materials and products, and to support the definition of regulatory measures and implementation of legislation in Europe One objective of NANOMICEX is the development of novel methods based on nanoparticle functionalization to reduce hazards caused by potential nanoparticle emissions during ink/pigment-based products life cycle. Work package 6 involves the Adaptive Streamlined Life Cycle/Risk Assessment of nanoparticle-based inks and pigments. The main objective of NanoValid is the development of new reference methods and certified reference materials, including methods for characterization, detection/quantification, dispersion control and labelling, as well as hazard identification, exposure and risk assessment of engineered nanmoaterials. Work package 4 will evaluate the potential of the methods to perform hazard and risk assessment and life cycle analyses of engineered nanmoaterials. 2012-10-012014-09-30 http://www.licara.eu The specific objectives of LICARA include: the development of a framework for LCA that properly addresses risks in data scarce situations, and the application of the life cycle approach in case studies. 67 - Som, C., Berges, M., Chaudhry, Q., Dusinska, M., Fernandes, T.F., Olsen, S.I., Nowack, B., 2010. The importance of life cycle concepts for the development of safe nanoproducts. Toxicology 269, 160– 169. human risk impact assessment, product stewardship and stakeholder risk/benefit communication of nanomaterials NanoCelluComp The development of very highperformance bioderived composite materials of cellulose nanofibres and polysaccharides SunPap Scale-up Nanoparticles in modern papermaking Work Package 4 Life cycle Analysis: Comparison of products based on nanomaterials with conventional products in order to illustrate their risks and benefits for the environment. Determination of the environmental and socio-economic impact of these products. 2011-03-012014-02-28 2009-07-012012-09-30 http://www.nanocellucomp .eu http://sunpap.vtt.fi The aim of NanoCelluComp is to develop a technology to utilise the high mechanical performance of cellulose nanofibres, obtained from food processing waste streams, combined with bioderived matrix materials, for the manufacture of 100% bio-derived high performance composite materials that will replace randomly oriented and unidirectional glass and carbon fibre reinforced plastics. The environmental sustainability benefits and risks will be quantified throughout the full product life cycle for selected products, where the new material may substitute for carbon fibre reinforced plastics and glass fibre reinforced plastics. Environmental health and safety issues will be considered for the full product life cycle of the selected products. SunPap aims to strengthen the European paper industry’s competitiveness by means of nanocellulose based processes to provide radical product performance improvements, new efficient manufacturing methods and the introduction of new added value functionalities. LCA was used to assess the environmental impact of Nano fibrillated cellulose coated board, compared to conventional board. 68 Hohenthal et al. 2012. Final assessment of nano enhanced new products. VTT APPENDIX B Figure B.1: European Commission, Joint Research Centre, Institute for Environment and Sustainability, Life Cycle Thinking and Assessment Source: European Commission, 2010. Life Cycle Thinking and Assessment - Our thinking - life cycle thinking. Available: http://lct.jrc.ec.europa.eu/index_jrc [Accessed: 20/03/2013] Figure B.2: Joint Research Centre, Institute for Environment and Sustainability Source: European Commission, n.d. Joint Research Centre, Institute for Environment and Sustainability, Life Cycle Thinking and Assessment. Available: http://ies.jrc.ec.europa.eu/our-activities/support-for-eu-policies/life-cyclethinking-and-assessment.html [Accessed: 20/03/2013] 69 Figure B.3: United States Environmental Protection Agency Source: United States Environmental Protection Agency, n.d. Risk Management Sustainable Technology, Life Cycle Perspective. Available: http://www.epa.gov/nrmrl/std/lifecycle.html [Accessed: 20/03/2013] Figure B.4: Mobile phone life cycle Source: UNEP/SETAC, 2009. Life Cycle Management: How business uses it to decrease footprint, create opportunities and make value chains more sustainable. United Nations Environmental Programme/Society of Environmental Toxicology and Chemistry - Life Cycle Initiative. P.4 70 Raw material acquisition Resources, e.g. raw materials energy land resources Processes Transports Manufacture Use Emissions to Air Water ground Waste Management Figure B.4: The life cycle model Source: Baumann, H., Tillman, A.M., 2004. The Hitch Hiker’s Guide to LCA. Studentlitteratur, Lund. P. 20 Figure B.4: The life cycle perspective Source: Rex, E., 2008. Marketing for Life Cycle Thinking. PhD Thesis. Environmental Systems Analysis. Department of Energy and Environment, Chalmers University of Technology Göteborg, Sweden. Page 3 71