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Resources, Conservation and Recycling 78 (2013) 54– 66

Contents lists available at SciVerse ScienceDirect

Resources, Conservation and Recycling

journa l h om epa ge: www.elsev ier .com/ locate / resconrec

eview

ife cycle assessments of biodegradable, commercial biopolymers— critical review

adeleine R. Yates ∗, Claire Y. Barlownstitute for Manufacturing, Department of Engineering, University of Cambridge, 17 Charles Babbage Road, Cambridge CB3 0FS, United Kingdom

r t i c l e i n f o

rticle history:eceived 21 November 2012eceived in revised form 25 June 2013ccepted 29 June 2013

eywords:iopolymerife cycle assessmentolylactic acidolyhydroxyalkanoate

a b s t r a c t

Biopolymers are generally considered an eco-friendly alternative to petrochemical polymers due to therenewable feedstock used to produce them and their biodegradability. However, the farming practicesused to grow these feedstocks often carry significant environmental burdens, and the production energycan be higher than for petrochemical polymers. Life cycle assessments (LCAs) are available in the litera-ture, which make comparisons between biopolymers and various petrochemical polymers, however theresults can be very disparate. This review has therefore been undertaken, focusing on three biodegradablebiopolymers, poly(lactic acid) (PLA), poly(hydroxyalkanoates) (PHAs), and starch-based polymers, in anattempt to determine the environmental impact of each in comparison to petrochemical polymers. Rea-sons are explored for the discrepancies between these published LCAs. The majority of studies focused

tarch based polymers only on the consumption of non-renewable energy and global warming potential and often found thesebiopolymers to be superior to petrochemically derived polymers. In contrast, studies which consideredother environmental impact categories as well as those which were regional or product specific oftenfound that this conclusion could not be drawn. Despite some unfavorable results for these biopolymers,

the immature nature of these technologies needs to be taken into account as future optimization andimprovements in process efficiencies are expected.

© 2013 Elsevier B.V. All rights reserved.

ontents

. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 55

. Background . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 552.1. Life cycle assessment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 552.2. Biopolymers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 56

. Review of LCAs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 563.1. Polylactic acid . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 563.2. Polyhydroxyalkanoates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 583.3. Starch-based polymers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 60

. Waste management options . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 60

. Recycling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 62

. Discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 626.1. The Benefits of Biodegradability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 626.2. Analysis of discrepancies in results . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 626.3. From LCA data to a material choice . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 65

6.4. Future improvements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

. Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

References. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

Abbreviations: AP, acidification potential; COD, chemical oxygen demand; DAR, depleHG, green house gases; GCV, gross calorific value; GWP, global warming potential; HTP,

se; PCL, polycaprolactone; PE, polyethylene; PET, polyethylene terephthalate; PHA, pololypropylene; PS, polystyrene; TPS, thermoplastic starch; PVA, polyvinylalcohol.∗ Corresponding author at: Alan Reece Building, 17 Charles Babbage Road, Cambridge C

E-mail address: mry23@cam.ac.uk (M.R. Yates).

921-3449/$ – see front matter © 2013 Elsevier B.V. All rights reserved.ttp://dx.doi.org/10.1016/j.resconrec.2013.06.010

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 65. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 65. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 65

tion of abiotic resources; EP, eutrophication potential; FFE, fossil fuel equivalents;human toxicity potential; LCA, life cycle assessment; NREU, non renewable energyyhydroxyalkanoates; PLA, polylactic acid; POC, photochemical ozone creation; PP,

B3 0FS, United Kingdom. Tel.: +44 1223 766 402.

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M.R. Yates, C.Y. Barlow / Resources, C

. Introduction

Petrochemical plastics have many technical advantages thatave seen them replace other materials in many applications (Davisnd Song, 2006). Their light-weight, resistance to corrosion, andow temperature processing often result in energy savings (Colwillt al., 2010), however these materials also have environmentalisadvantages. Unlike the materials they have replaced, such aslass and metal, the recycling of plastics has been less successfulue to difficulties in identification and sorting and the presence ofarious other materials and additives such as fillers and plasticiz-rs (Davis and Song, 2006). As a result of this and their longevitynd widespread use, particularly in disposable products, manag-ng plastic waste can be a major problem. Being organic materials,hey have the potential to be used for energy production but uncon-rolled or poorly managed incineration can result in hazardousmissions (Hopewell et al., 2009) and this suffers considerable pub-ic opposition. Additionally, when plastic waste enters the naturalnvironment as litter through poor waste management or incorrectisposal, this poses threats to wildlife (Rao, 2010). The large vol-mes and bright colors of plastic waste make them highly visible inhe waste stream and as litter (Davis and Song, 2006). Such mattersoster a negative emotional response and give plastics considerablead press.

Biopolymers have emerged as potential alternatives, some ofhich are available commercially while others remain under

esearch. The term biopolymer refers to naturally occurring long-hain molecules but also materials which have been derived fromhese or bio-based monomers (Song et al., 2009). Most of theseill also be biodegradable (Davis and Song, 2006), however this isot necessarily the case. A biopolymer made from annually renew-ble resources which will biodegrade can appear to solve the majorroblems associated with plastics. The feedstock is no longer of fos-il origin, so improving resource security (Shapouri et al., 1995) andt the end of its life, the polymer is capable of biodegrading to leaveo waste product.

This paper discusses the environmental impacts based on lifeycle assessment (LCA) studies of three biopolymers: poly(lacticcid) (PLA), the poly(hydroxyalkanoates) (PHAs), and starch basedolymers. These are examples of biopolymers which can be pro-uced using renewable bio-based feedstocks and are capable ofiodegradation given appropriate conditions. Other biopolymerso exist, some of which have greater commercial importance athis point. These include bio-based polymers which have tradition-lly been produced from petrochemical feedstocks such as PE andET (Shen et al., 2009). These bio-based polymers are chemicallydentical to their petrochemical equivalents meaning they haveimilar properties and currently have a greater potential for sub-titution (Shen et al., 2009). However, as with their petrochemicalquivalents, they are not biodegradable. The LCAs reviewed in thisaper have mainly focused on non-renewable energy use (NREU)nd global warming potential (GWP). A limited number of studiesave also evaluated other environmental impacts. The results showhat, although biopolymers can bring some reduction in environ-

ental burdens, they are far from a perfect solution at this point inime.

. Background

.1. Life cycle assessment

LCA is a framework which can be used to assess the environ-ental impacts of a product throughout its life starting from the

xtraction of raw materials from the earth and ending at the wasteroducts being returned to the earth. An LCA involves collecting

vation and Recycling 78 (2013) 54– 66 55

information on the inputs and outputs, such as emissions, waste,and resources, of a process (life cycle inventory) and translatingthese to environmental consequences (using impact assessmentmethodologies) such as contribution to climate change, smogcreation, eutrophication, acidification, and human and ecosys-tem toxicity (Rebitzer et al., 2004). The final results can bepresented in these impact categories or grouped together intodamage categories (e.g. ‘human health’, ‘ecosystem quality’, and‘resources’ used in the Ecoindicator-99 methodology) followingnormalization. Normalization can be performed by dividing impactcategory scores by the average person’s annual contribution, allow-ing for the combination of categories, whose scores now haveno units. Since some environmental impacts can be consideredmore important than others, a weight can also be assigned tothe normalized impact score. However the decision on relativeimportance is subjective and therefore has no scientific basis(BSI, 2006a) and there is no consensus on proposed methods(Thrane and Schmidt, 2007). LCA has been standardized to a cer-tain extent by the International Organization for Standardization(ISO) although these guidelines leave much to be interpreted andselected by the LCA practitioner. As a result, the comparison ofresults from different studies may not be possible and in any case,need to be made with caution for a number of reasons includ-ing methodological differences (such as functional unit, allocationmethod, and impact assessment methodology) and inventory dataused.

For petrochemical polymers, the life cycle starts with the extrac-tion of the raw materials required, including the fossil feedstock.The fossil feedstock is combined with (non-renewable) energyrequirements to calculate the depletion of fossil fuels, usuallypresented as NREU. This is possible by representing the feed-stock as energy rather than a material input by multiplying theamount consumed by its heat of combustion (BSI, 2006b). LCI dataon petrochemical plastics based on European averages is avail-able from PlasticsEurope which was used in many of the studiesreviewed for comparison with biopolymers. LCAs on biopolymersderived from agricultural products include the cultivation of thecrop used. This includes the fuel required for farming activi-ties such as plowing, agrochemical application, and harvesting. Itshould also include the manufacture and transport of the mate-rials required such as fertilizer, herbicides, and pesticides. Landuse and water consumption may also be important factors, asare nitrogen based emissions from fertilizer use. Other processesto include for typical biopolymers include milling and produc-tion.

The studies reviewed here have used LCA to make comparisonsbetween biopolymers and the petrochemical plastics they could bereplacing. Some are full life cycle assessments from cradle-to-gravewhile others have only considered the impacts up to the factorygate of polymer pellets or a particular product (known as eco-profiles). Studies which stop at pellet or product production havethe advantage of being more easily compared to other studies, how-ever the omission of down-stream processes could be significantand change the preferred material depending on its application anddisposal route. For product eco-profiles and full LCAs, the manufac-ture of the specific item is an additional step. This was often omittedfrom comparative studies since the environmental impacts wereassumed to be equivalent for all polymers if the same processeswere used. Similarly, the use phase and, less frequently, transporta-tion have been omitted in cradle-to-grave studies. However, if themass requirements for a product differed depending on the mate-rial used, this was taken into account. The impact of end-of-life

processes is also included in cradle-to-grave studies. The optionsconsidered varied between studies but have included landfill, incin-eration, industrial and home composting, anaerobic digestion (AD)and recycling.

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6 M.R. Yates, C.Y. Barlow / Resources, C

.2. Biopolymers

This review will concentrate on three biodegradable biopoly-ers which have received considerable attention in the academic

iterature: PLA, PHAs, and starch based polymers.PLA is synthesized by either condensation polymerization or

zetropic dehydrative condensation of lactic acid or by ring-pening polymerization of lactide, an intermediate produced fromhe PLA pre-polymer (Averous, 2008). PLA’s monomer, lactic acid,an be obtained by both chemical synthesis and fermentation ofarbohydrates (Averous, 2008). While PLA can be produced fromarious feedstocks, corn (maize) has been desirable since it is cheap,bundant, and contains large quantities of sugar, all of which aressential for PLA to be competitive on a large scale with petrochemi-al polymers (Landis, 2010). It is, however, an energy and chemicalntensive crop (Landis et al., 2007) which can be associated withonsiderable environmental burdens. Sugarcane has also been useds a feedstock which has lower environmental impacts, largelyue to the current practice of burning the bagasse (residues fromilling) to generate electricity which can displace fossil-derived

lectricity (Patel et al., 2006). For PLA derived from either of theseeedstocks, the LCA includes the cultivation and harvesting of cornr sugarcane, milling to obtain dextrose, and conversion into lacticcid, lactide, and finally PLA.

The mechanical performance of PLA has been described asigher than that of PS and comparable to that of PET, although

t has poorer barrier properties (Scaffaro et al., 2011). Applica-ions include packaging, textiles, disposable serviceware and paperoatings (Rudnik, 2008).

Degradation of PLA is by hydrolysis, after which the oligomersan be metabolized by microorganisms (Drumright et al., 2000). Forndustrial composting, hydrolysis at temperatures above 58 ◦C for

weeks is required, while in nature, such as marine environments,egradation is poor (Rudnik, 2008). For this reason, PLA may notffer any advantage over petrochemical plastics with respect to theroblem of plastic litter.

PHAs are polymers synthesized inside microorganisms (Khannand Srivastava, 2005), augmented by a carbon source which is mostommonly derived from an agricultural product, typically corn.lternatively, PHAs may be obtained directly from genetically mod-

fied (GM) plants (Kurdikar et al., 2000). In both cases, the PHA isecovered using solvent extraction (Mooney, 2009; Kurdikar et al.,000). LCAs of PHAs include an agricultural and milling stage sim-

lar to that for PLA and production which includes fermentationnd separation (Akiyama et al., 2003). The most common PHA, theomopolymer PHB, has properties which resemble PP (Mooney,009). Due to their physical properties including being stiff andrittle, applications are limited but include packaging, flushable

tems, and coatings (Rudnik, 2008).Thermoplastic starch (TPS) is obtained by the destructurization

f native starch in the presence of a plasticizer (Patel et al., 2003).PS can be used either on its own or in combination with otherolymers for improved mechanical properties. For example, Nova-ont, a manufacture of starch based polymers, produces MaterBi,

material composed of different proportions of starch and petro-hemical but biodegradable polymers such as PCL and PVA (Gironind Piemonte, 2011). Applications include foams (for loose-fill),gricultural films, moldable products, and shopping bags (Rudnik,008).

Although biopolymers do not rely on fossil resources as a feed-tock, the agricultural, milling, and production stages consumearge amounts of energy, currently mainly derived from fossil fuels.

here are additional environmental burdens associated with thegricultural stage such as nitrogen emissions from fertilizers whichesult in eutrophication. Non-renewable energy consumption cane improved by incorporating what is known as ‘the integrated

vation and Recycling 78 (2013) 54– 66

system’. This involves combining up to 60% of corn stovers (residueslike stalks, leaves, and cobs) with the corn grain for polymer produc-tion, after which the remaining residues are incinerated to producesteam and electricity to run the process (Kim and Dale, 2005). How-ever, this system may increase soil erosion and reduce the soil’scarbon sequestering ability (Kurdikar et al., 2000).

Other issues which may be raised concerning agriculture-derived biopolymers are finite land resources, the resultingcompetition with food crops (Scott, 2000), and their vulnerability tocrop failure from flooding or drought. Crank et al. (2005) state that,if biopolymer consumption has a high growth rate, there may besome conflict of interest with bioenergy crops around 2050. Addi-tionally, Colwill et al. (2012) have predicted that all crop and grazingland, and cleared forest land would not be sufficient to meet thedemand for food, liquid fuels, and plastics (assuming all fuels andplastics are derived from agricultural products due to exhaustionof fossil resources) in 2050 in a high consumption, low productivityscenario. However, more realistic scenarios analyzed indicate thatdemand could be met with only some clearing of existing forestland (Colwill et al., 2012). If grass and forest land do require clear-ing in order to make space for biopolymer feedstock cultivation,the indirect greenhouse gas emissions associated with this need tobe taken into account, which may cancel out any carbon savings(Piemonte and Gironi, 2011).

3. Review of LCAs

3.1. Polylactic acid

Four of the LCA studies on PLA reviewed use original raw data.NatureWorks, a US company with a large-scale PLA productionfacility published LCA results based on their technology using cornas a feedstock (Vink et al., 2003) followed by 2 updates (Vinket al., 2007, 2010). Bohlmann (2004) and Uihlein et al. (2008) alsouse corn as a feedstock while Groot and Borén (2010) use sugar-cane grown in Thailand. All studies reported results for NREU andGWP and found PLA to be superior to petrochemical plastics. Thiscan be seen in Fig. 1 where PLA’s NREU is lower than that for allpetrochemical polymers, regardless of the inventory data used forthese. The GWP is also lower in many cases, although the rangeof scores reported for some petrochemical polymers (in particularPP) demonstrate that conclusions will depend on the polymer usedfor comparison and the inventory data used for this. The data inthe tables and figures are given on a weight basis which the stud-ies by NatureWorks (Vink et al., 2003, 2007, 2010) and Groot andBorén (2010) have used to make comparisons. This is not alwaysappropriate since different material properties can mean differentmass requirements for certain applications. However, Tabone et al.(2010) has shown that results are relatively similar when polymersare compared by volume rather than weight. Despite increasedimpacts associated with PLA when a greater mass is required (e.g.for yoghurt containers studied by Bohlmann (2004) and drinkingcups studied by Uihlein et al. (2008)), these studies have shownthat a preference for PLA remains.

Another study which was not an LCA found that the produc-tion of PLA from food waste did not require substantially moreenergy than that from a homogeneous feedstock (Sakai et al., 2004),indicating that PLA produced via this route could have lower envi-ronmental impacts than petrochemical plastics also.

Other LCA studies on corn derived PLA have been product-specific and use LCI data from NatureWorks to some extent (see

notes in Table 1 for details). While some of these studies haveassumed a greater mass of PLA for the functional unit used, the datain Table 1 only show the results on an equal mass basis to allow forcomparison with other studies. Piemonte (2011) and Gironi and

M.R. Yates, C.Y. Barlow / Resources, Conservation and Recycling 78 (2013) 54– 66 57

0

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ig. 1. GWP and NREU for PLA studies and petrochemical polymers. Sequestration

over values reported in various comparative LCA studies and those provided by Pl

iemonte (2010) found that PLA products had lower NREU andWP than the petrochemical polymer products they were com-ared with. Madival et al. (2009) also reported lower NREU for PLAlamshell containers compared with PS and PET, however the GWPas higher than that for PS containers. It should be noted that this

tudy incorporated the renewable energy credits (RECs) discussedn Vink et al. (2007) which results in lower values for GWP andREU. The use of RECs is discussed in Section 6.2.

In contrast to these results, Hermann et al. (2010) found that PLAsed for both inner and outer food packaging had a greater envi-onmental impact than PP, the reference material, based on theategories NREU, GWP, AP, EP, POF, and water use. However, thistudy included the energy requirements for plastic film productionaken from existing manufacturers who use less energy-efficient,mall scale lines for novel materials such as PLA in comparison to

onventional plastics (Hermann et al., 2010). Suwanmanee et al.2013) also found that PLA thermoform boxes had greater environ-

ental impacts in comparison to PS using the categories GWP, AP,nd POC. In this case, the high GWP is due to the incorporation of

able 1REU and GWP results for LCA studies on PLA and a selected number of petrochemical po

Reference No. Product

Vink et al. (2003)a 1 PLA pellets

Vink et al. (2007)b 2 PLA pellets

Vink et al. (2010)c 3 PLA pellets

Bohlmann (2004)d 4 PLA yoghurt cSuwanmanee et al. (2013)e 5 PLA thermofoMadival et al. (2009)f 6 PLA clamshelGironi and Piemonte (2010)g 7 PLA water boSakai et al. (2004)h 8 PLA pellets (fGroot and Borén (2010)i 9 PLA pellets (sPlasticsEurope (2011)j 10 PET

PlasticsEurope (2012)k 11 PS

PlasticsEurope (2008a) 12 HDPE

PlasticsEurope (2008b) 13 PP

ll results are given per kg polymer pellet/resin. The feedstock for PLA is corn grain unlesa Based on engineering estimates.b Excludes wind energy credits.c Values for current technology only. Reductions due to improved fermentation technod Does not use standardized methodology. Based on PEP (process economics program)e CO2 sequestration from corn plants not included. Uses some data from Vink et al. (20f Not including container manufacture. Uses some data from Vink et al. (2007) with wig Not including energy for bottle production. Uses some data from Vink et al. (2007) reh Production and transport energy only.i Total primary energy requirements (including renewable energy used in production)j Using data from 2005.k General purpose PS. Using data from 2002.

to CO2 uptake by plants used as the feedstock. Ranges for petrochemical polymersurope.

indirect emissions from land use change which contribute 81–91%of this impact category (Suwanmanee et al., 2013).

For categories other than GWP and NREU, Groot and Borén(2010) also found PLA to have a higher environmental impact thanpetrochemical polymers with higher values for AP and EP than allpetrochemical polymers studied (see Table 2 and Fig. 2) and highervalues for POC and HTP than all polymers except PS. Fig. 2 showsthat AP and EP are greater for PLA (as reported by Groot and Borén(2010)) than that reported for all petrochemical polymers in allstudies reviewed. Madival et al. (2009) compared PLA to PS andPET and report higher impacts in AP, respiratory organics and inor-ganics, as well as higher values for ozone layer depletion and EPcompared with PS (Madival et al., 2009). However, the AP and EPfor PLA differ substantially to those reported by Groot and Borén(2010) and Gironi and Piemonte (2010) and the conclusion that EP

is greater than that for PS would depend on the inventory data usedfor PS (see Fig. 2).

With respect to ecotoxicity, PLA had a lower score than PET andPS (Madival et al., 2009), a result also found by Gironi and Piemonte

lymers for comparison.

NREU (MJ) GWP (kg CO2 eq)

54.1 1.850.9 2.0242.2 1.24

ontainers 56.7 0.74rm boxes 51.1 4.0l containers 32.4 1.96ttles 55.4 1.12ood waste) 44.4ugarcane) 30.55 0.5

82.3 3.4987.4 3.4076.7 1.973.4 2.0

s otherwise stated.

logy. data.10).nd energy credits.placing wind-with fossil fuel-derived energy.

= 54 MJ.

58 M.R. Yates, C.Y. Barlow / Resources, Conservation and Recycling 78 (2013) 54– 66

Table 2AP and EP results for LCA studies on PLA and a selected number of petrochemical polymers for comparison.

Reference No. Product AP (g SO2 eq) EP (g PO43− eq)

Madival et al. (2009) 1 PLA clamshell containers 38.3 0.18Groot and Borén (2010) 2 PLA from sugarcane 21 5.0Gironi and Piemonte (2010) 3 PLA water bottles 11.88 7.70

PET 9.59 3.04PlasticsEurope (2011)a 4 PET 15.59 1.03Madival et al. (2009)b 1 PET 11.1 2.09

PS 18.8 0.00789PlasticsEurope (2012)c 6 PS 11.48 0.72PlasticsEurope (2008a) 7 HDPE 6.39 0.43PlasticsEurope (2008b) 8 PP 6.13 0.74

All results are given per kg polymer pellet/resin. The feedstock for PLA is corn grain unless otherwise stated.a Using data from 2005.b The results for AP and EP reported by Madival et al. (2009) for PLA and petrochemical polymers are substantially different to other studies indicating a probable error in

c

(i9Ptet

wpstspwmpioocbtsth

alculation.c General purpose PS. Using data from 2002.

2010). Despite this, when other factors related to ecosystem qual-ty were taken into account by grouping using the Ecoindicator9 methodology, PLA had a greater impact than PET (Gironi andiemonte, 2010). Uihlein et al. (2008) also used the Ecoindica-or 99 methodology which showed higher impacts for PLA in thecosystem quality and human health categories in comparisono PS.

The results show that up to pellet formation, comparisons byeight indicate a lower GWP and NREU for PLA compared toetrochemical polymers and this remained the case when factorsuch as additional mass requirements for specific products wereaken into account. However, Hermann et al. (2010) has demon-trated that PLA products may not offer environmental savings atresent due to the facilities used to process them. Additionally, evenhen PLA was found to have lower NREU and GWP, the environ-ental impact in other categories was often higher than that for

etrochemical plastics. Further details on the results of these stud-es can be found in Table 3 and Figs. 1 and 2. These, along withther tables and figures in this review, only contain informationn studies which provide the results in a format that allows foromparison with other studies (i.e. impacts provided or able toe calculated on a mass basis and excluding studies using rela-

ive or normalized scores only). Due to a lack of studies reportingcores for other impact categories using the same methodology andherefore units, only scores for NREU, GWP, AP, and EP are givenere.

0

5

10

15

20

25

30

35

40

AP

(g S

O2eq

/kg

Pol

ymer

)

1 2 3 PE T PS HDPE PPStudy Numbe r

Fig. 2. AP and EP for PLA studies a

3.2. Polyhydroxyalkanoates

LCA studies on PHAs have been based on simulations andestimates from laboratory and small facility scale-ups since nolarge-scale commercial facilities exist. One exception exists wheredata has been collected from a facility in Iowa (Kim and Dale,2008). Older studies have been brief and provided little detail ontheir methodology and assumptions. The three papers written orco-authored by Gerngross (Gerngross, 1999; Gerngross and Slater,2000; Gerngross, 2001) along with those by Lynd and Wang (2003),Heyde (1998), and Khoo et al. (2010) found that PHA productionrequired more fossil fuel than petrochemical plastics. The age ofthese studies should be noted as they are unlikely to representthe technology currently available, including reductions in energyrequirements (Khoo et al. (2010) has used data from Gerngross(1999)). Tabone et al. (2010) also shows that PHA from corn grainhas higher GWP and NREU than PP and PE by volume, however theirdata source is unclear. Kurdikar et al. (2000) found that PHA fromGM corn had a higher GWP than petrochemical polymers (although,as seen in Fig. 3, large variation exists depending on the sourceof production energy assumed). Table 4 and Fig. 3 shows that theNREU reported in these studies is up to almost double that found

by more recent LCAs.

More recent studies (e.g. Akiyama et al., 2003; Kim and Dale,2005, 2008; Gurieff and Lant, 2007; Yu and Chen, 2008; Hardinget al., 2007) found that PHAs can have lower NREU and/or GWP than

0

1

2

3

4

5

6

7

8

9

EP

(g P

O43

- eq/k

g P

olym

er)

1 2 3 PET PS HDPE PPStud y Number

nd petrochemical polymers.

M.R. Yates, C.Y. Barlow / Resources, Conservation and Recycling 78 (2013) 54– 66 59

Table 3NREU and GWP results for LCA studies on PHA and a selected number of petrochemical polymers for comparison. All results are per kg polymer.

Reference No. Product NREU (MJ) GWP (kg CO2 eq)

Gerngross (1999) 1 PHA 81Heyde (1998)a 2 PHA 66

PHA from sugar beet 80Khoo et al. (2010)b – PHA carrier bags 81 34Akiyama et al. (2003)c 3 PHA 59–68 0.5–1.4

PHA from soybean oil 42–62 −0.24 to 0.82Kurdikar et al. (2000)d 4 PHA from GM corn −4 to 5.7Pietrini et al. (2007)e 5 PHA composite (sugarcane) −22.7 −3.1

PHA composite 38.6 0Kim and Dale (2005)f 6 PHA (Gerngross technology) 107 31.5l 4.1–0.28l

PHA(Metabolix technology) −24.9l 3.2–0.77l

PHA (Akiyama technology) 69 12.3l 1.6–1.93l

Kim and Dale (2008)g 7 PHA 2.5 −2.3Harding et al. (2007) 8 PHA from sugarcane 44.7 1.9Yu and Chen (2008)h 9 PHA from black syrup 44 0.49Kendall (2012)i 10 PHA from organic waste 49–76 3.1–5.1PlasticsEurope (2011)j 11 PET 82.3 3.49PlasticsEurope (2012)k 12 PS 87.4 3.40PlasticsEurope (2008a) 13 HDPE 76.7 1.9PlasticsEurope (2008b) 14 PP 73.4 2.0

All results are given per kg polymer pellet/resin. The feedstock for PLA is corn grain unless otherwise stated.a No raw data; insufficient information about basis for calculations. NREU range depends on source of electricity.b GWP estimated from figure but unusually high (high value for PP of 19 kg also reported indicating a potential error).c Ranges provided based on different production conditions and parameters (e.g. temperature, yield).d PHA in stover, corn grain not used. Range reflects different sources of energy.e Data taken from BREW report (Patel et al., 2006) which uses a variety of past literature and confidential sources of information.f Uses own data for agricultural stage based on no-till farming practices.g Most of energy for milling and fermentation obtained by burning stover in a cogeneration power plant.h Black syrup is a by-product of ethanol production from corn stover. Cradle-to-grave (landfill).i Uses fermentation data from Harding et al. (2007) (lower impacts) and Akiyama et al. (2003) (higher impacts) and pre-fermentation data from other studies. Includes

emission factors due to land use change.

pPhuadiaPdpb

Fp

j Using data from 2005.k General purpose PS. Using data from 2002.l Using the integrated system.

etrochemical plastics. Additionally, Pietrini et al. (2007) found thatHB-based composite materials had lower NREU and GWP thanigh impact PS when used for a cathode ray tube monitor housingsing PHA data from the BREW project (Patel et al., 2006). Thesere shown as studies 3 and 6–10 in Fig. 3 which also shows that,epending on the petrochemical polymer used for comparison, the

nventory data used for this, and assumptions made regarding yieldnd production of PHA, material preference may change. Similarly,atel et al. (2006) have studied 12 different PHA materials from 3

ifferent feedstocks and the large range in results show that thereference for these over petrochemical polymers when comparedy weight varies (results not shown here).

-5

-4

-3

-2

-1

0

1

2

3

4

5

6

GW

P (k

gCO

2eq

/kg

Pol

ymer

)

3 4 5 6 6* 7* 5 8 9 10

3 P

ET PS

HD

PE PP

Corn

*Integrat

Other feedstock s

Study Number

ig. 3. GWP and NREU for PHA studies and petrochemical polymers. Ranges for petrochemrovided by PlasticsEurope.

Mixed culture PHA from industrial waste water has also beenshown to have a lower GWP than HDPE, assuming a high concen-tration or COD (Gurieff and Lant, 2007) although the units make itdifficult to compare with other studies. The results from Kendall(2012) for PHA from waste show a lower NREU than petrochemicalplastics while their results for corn-derived PHA were much higherthan those reported by other authors (shown by the range in study10, Fig. 3), the reasons for which are unclear.

One final LCA study on PHA looked at its production using glu-

cose from corn, cheese whey (a waste product of cheese making),and GM corn (Zhong et al., 2009). The units and categories makeit difficult to compare with other studies and the results are not

ed Sys tem

-20

0

20

40

60

80

100

120

NR

EU

(MJ/

kg P

olym

er)

1 2 3 5 6

6* 7* 2 5 8 9 10 3 P

ET PS

HD

PE PP

Stud y Number

CornOther feedstocks

ical polymers cover values reported in various comparative LCA studies and those

60 M.R. Yates, C.Y. Barlow / Resources, Conservation and Recycling 78 (2013) 54– 66

Table 4AP and EP results for LCA studies on PHA and a selected number of petrochemical polymers for comparison. All results are per kg polymer.

Reference No. Product AP (g SO2 eq) EP (g PO43− eq)

Harding et al. (2007) 1 PHA from sugarcane 24.9 5.19Kendall (2012)a 2 PHA from organic waste 16–28 0.54–5.0PlasticsEurope (2011)b 3 PET 15.59 1.03PlasticsEurope (2012)c 4 PS 11.48 0.72PlasticsEurope (2008a) 5 HDPE 6.39 0.43PlasticsEurope (2008b) 6 PP 6.13 0.74Harding et al. (2007) HDPE 22.5 0.81Harding et al. (2007) PP 48.8 5.84

All results are given per kg polymer pellet/resin. The feedstock for PLA is corn grain unless otherwise stated.a Uses fermentation data from Harding et al. (2007) (lower impacts) and Akiyama et al. (2003) (higher impacts) and pre-fermentation data from other studies. Includes

e

cph9e

TmdDtAKtst

wtEtrFcumP

3

amafi(aetffttwwtsn

mission factors due to land use change.b Using data from 2005.c General purpose PS. Using data from 2002.

ompared to petrochemical plastics. The authors found that PHAroduction using glucose (from corn) or whey as the feedstockas comparable environmental impacts in all three Ecoindicator-9 categories while PHA from transgenic corn has much highernvironmental burdens (Zhong et al., 2009).

Four studies considered categories other than NREU and GWP.hese have not always been calculated or presented in ways thatake comparison between studies possible, therefore very limited

ata for EP and AP only are shown in table and Fig. 4. Kim andale (2005) found that PHB had higher AP and EP than PS while

he results from Kendall (2012), shown in Fig. 4, also show higherP and EP in comparison to a number of petrochemical polymers.hoo et al. (2010) also found that PHA bags had higher AP and POC

han PP. When compared by volume rather than weight, PHA washown to have higher impacts than PP and PE in all categories otherhan POC (Tabone et al., 2010).

In contrast, Harding et al. (2007) reported lower AP and EP (asell as ozone depletion, HTP, ecotoxicity, and POC) in comparison

o PP but higher AP and EP compared to PE. However, the AP andP for PP and PE reported by Harding et al. (2007) are much greaterhan the values reported in the most recent reports by PlasticsEu-ope (PlasticsEurope, 2008a,b) so these findings are not reflected inig. 4 which shows PHA to have a higher AP and EP than all petro-hemical polymers used for comparison. The reasons for this arenclear but could be due to updated eco-profiles for these poly-ers since Harding et al. (2007) claim to have used data from

lasticsEurope and the same impact assessment methodology.

.3. Starch-based polymers

Fewer LCA studies have been found on starch-based materi-ls, however several review papers have been published whichake reference to many reports in the German language which

re not available to the authors of this review. The impact resultsrom these studies vary, but in contrast to the PLA and PHA stud-es, this was not sufficient to alter the material preference. Bastioli2001) states that MaterBi compost bags perform better than papernd PE bags, implying the use of NREU and/or GWP impact cat-gories. With respect to NREU and GWP, Piemonte (2011) foundhat MaterBi (34% starch derived from corn) food packages per-ormed better than PE and PET while Gironi and Piemonte (2011)ound that MaterBi (36% starch) shopping bags also performed bet-er than PE even though an additional 16 g of material was requiredo obtain equivalent mechanical properties. Both of these studiesere cradle-to-gate. These studies do not present their results in a

ay that allows for comparison between studies on a weight basis,

herefore they are not presented here. Tabulated results from othertudies not available can be found in Shen and Patel (2008) and areot replicated.

Patel et al. (2003) and Shen and Patel (2008) review studies onpure TPS, and blends with PCL and PVA, most of which used PE forcomparison. The authors conclude that starch based polymers havebetter environmental profiles than PE in all categories studied otherthan EP, although they have higher impacts in other categorieswhich are often outside the scope of LCAs (Patel et al., 2003). Thiswas also the case for the MaterBi shopping bags studied by Gironiand Piemonte (2011) which had higher impacts in the ecosys-tem quality and human health damage categories. With respectto NREU and GWP however, all studies showed that starch basedpolymers had a lower impact, often substantial, than petrochemi-cal polymers (Shen and Patel, 2008). This included cradle-to-gravestudies which take into account greater mass requirements and dis-posal, however it was noted that these starch-based polymers couldnot compete with recycled petrochemical polymers. Additionally,Piemonte and Gironi (2011) show that the 10% reduction in GWPachieved through replacement of PE with Mater-Bi shopping bagswould take over a century to be realized if land-use change relatedcarbon emissions are taken into account (although it was stressedthat the methodology for calculating these emissions is still underdevelopment). Similar results would also be expected if this factoris taken into account for other biopolymers using agriculture-basedfeedstocks.

4. Waste management options

For biodegradable polymers, disposal options include landfill,incineration, AD, home and industrial composting, and recycling.Some of the studies discussed above include the waste manage-ment stage in their LCA either quantitatively or qualitatively whileone study looked purely at the disposal stage. Where landfill is con-sidered, the assumed level of degradation for materials differedbetween studies from 0% (Bohlmann, 2004) up to 85%. Materialdegradation in landfill was assumed to be anaerobic, producingmethane, a more potent greenhouse gas than CO2 (Khoo and Tan,2010) and assumptions on capture and use and the efficiencies ofthese processes differ. Further details on the assumptions made inthe studies considering the waste management stage are outlinedin Table 5.

Hermann et al. (2011) performed an LCA on the waste dis-posal stage for a variety of biodegradable materials including PLA,PHAs, and MaterBi, looking at GWP and NREU only. While impactscores varied slightly for the different materials, the conclusionson treatment preference remained the same with industrial com-posting found to be the worst option despite carbon credits forthe compost produced. Home composting, on the other hand was

comparable to incineration (with energy recovery), while AD wasthe preferred option for all materials studied. Similar findings werepresented by Piemonte (2011) and Gironi and Piemonte (2010).Gironi and Piemonte (2010) found that incineration of PLA was

M.R. Yates, C.Y. Barlow / Resources, Conservation and Recycling 78 (2013) 54– 66 61

0

5

10

15

20

25

30

AP

(g S

O2eq

/kg

Pol

ymer

)

1 2 PE T PS HD PE PP0

1

2

3

4

5

6

EP

(g P

O43

- eq/k

g P

olym

er)

1 2 PE T PS HD PE PP

dies a

pLcoas2p

TS

N

Study Numbe r

Fig. 4. AP and EP for PLA stu

referred over landfill and composting based on GWP and NREU.andfill had lower energy requirements but a greater GWP thanomposting. However this is based on the assumption that 85%f the PLA will degrade in landfill and that the methane gener-ted will be captured to generate electricity while studies have

hown that PLA does not degrade in landfill (e.g. Kolstad et al.,012). Piemonte (2011) found that AD was the preferred dis-osal option for PLA and MaterBi, followed by incineration, then

able 5ummary of assumptions used, when mentioned, for waste disposal stage.

Reference Landfill Incineration AD

Hermann et al.(2010)

With gas recovery With energy recovery NM

Hermann et al. (2011) NC Carbon credits for power andheat generation. 11% GCVpower exported, net export ofheat 22% of GCV (author statesthis is significantly lower thanthe efficiencies possible innewer, optimized systems)

35% carassigneIndividN2O em36% elegeneratefficienexportecredits

fossil delectricdisplac

Khoo and Tan(2010)

50% decompositionof PHA. No gasrecovery.

Energy utilized but efficiencynot mentioned.

NC

Piemonte (2011) NC Energy utilized but efficiencynot mentioned.

85% dePLA andassumerecoverefficienconvertelectric

Gironi andPiemonte (2010)

85% degradation ofPLA, 25% recoveryof gas. 36%efficiency forelectricitygeneration fromgas burning.

26 MJ of electric and 53.3 MJ ofthermal energy from 1000 PLAbottles.

NC

C, not considered; NM, assumptions and efficiencies not mentioned.

Study Number

nd petrochemical polymers.

industrial composting when considering energy requirements,GWP, and the Ecoindicator-99 damage categories. However, thedegree of degradation and therefore methane gas produced in AD inboth of these studies was based purely on assumptions and analo-gies due to lack of data. Additionally, Hermann et al., 2011 found

that the carbon footprint of incineration would be reduced by 50%if state-of-the-art waste-to-energy technology were used whichmay indicate incineration as the best future option (Hermann et al.,

Home composting Industrialcomposting

Recycling

NC NM NC

bon stored in soil. Carbon creditsd for soil conditioner replacement.ual perspective: no nitrogen credits orissions applied.

NC

ctricityioncy, 28%d – carbonassigned forerivedityement

Carbon credits forpeat replacement.temp. ≤35 ◦C

Carbon credits forpeat and strawreplacement, ratio1:3. temp. 50–60 ◦C

NC 1/3 compostreplaces peat

NC

gradation of MaterBid. 95% gased, 36%cy ining toity

NC 50% of compostdisplaces 20% ofsynthetic fertilizerused foragricultural stage.60% degradation.95% of thisdegrades to CO2,5% into CH4

90% used to makea)lower gradeproduct which isincinerated afteruse or b) sameproduct

NC 60% degradation.95% of thisdegrades to CO2,5% into CH4

100% closed-looprecycling withsame efficiency asPET

6 onser

2wc

pwwcLitTaeeeic

5

atocerbcsmMcaficHi

nitm2PPer(pKrmcatTtc

6

6

a

2 M.R. Yates, C.Y. Barlow / Resources, C

011). This was based on combined heat and power (CHP) plantshich may not be appropriate technologies in areas without suffi-

ient demand for heat.In contrast, Khoo and Tan (2010) found that industrial com-

osting was the environmentally preferred option for PHA bagshen compared with landfill and incineration. This conclusionas drawn using emissions data from other studies and the GWP

ategory as well as the normalized value for GWP, AP, and POC.andfill resulted in the greatest environmental burdens although its important to note that, unlike other studies, no landfill gas is cap-ured at the existing site in Singapore used in this study (Khoo andan, 2010). Due to the increased potential for methane emissionsnd lack of energy credits from the utilization of landfill gas to gen-rate electricity in this case, the environmental impacts would bexpected to be greater. The preference for composting over incin-ration in this case could be due to poor energy recovery efficiencyn the particular incinerator under study, although this cannot beonfirmed as no value for efficiency was provided.

. Recycling

In a limited number of studies, recycling was also considereds an end-of-life option, and this was found to be environmen-ally superior. Piemonte (2011) found that closed-loop followed bypen-loop recycling were the best options for PLA and MaterBi foodontainers based on the Ecoindicator-99 damage categories. How-ver, AD (and incineration for PLA) had lower values for GWP thanecycling. Gironi and Piemonte (2010), who found that PLA waterottles had a lower overall impact than PET up to the factory-gate,oncluded that this would only remain the case in a cradle-to-gravetudy if 100% closed-loop recycling were possible. Using the sameethodology, Gironi and Piemonte (2011) found that compostedaterBi had a higher environmental impact than PE bags that were

losed-loop recycled with an efficiency of 90%, even though it had lower impact up to the factory gate. However, the practical dif-culties of recycling waste such as PE bags and also the fact thatlosed-loop recycling is currently only feasible for PET bottles andDPE milk containers (Hopewell et al., 2009) have not been taken

nto account.The studies which consider recycling of biopolymers also do

ot consider the practicalities of this option and little informations available in the literature on their recyclability. There is evidencehat recycling is possible with Erema GmbH claiming their equip-

ent is suitable for PLA recycling (Erema Plastic Recycling Systems,010). REPLA perform closed-loop recycling for post-industrialLA and are working on making this possible for post-consumerLA also (Reclay Group, 2013). However, the results from Scaffarot al. (2011) indicate that PLA may only be suitable for closed-loopecycling a limited number of times, or not at all. La Mantia et al.2002) concluded that MaterBi could be recycled into the sameroduct, although they do not specify how many times whileendall (2012) states that there is currently no technology forecycling PHAs. Finally, although recycling biopolymers may beore favorable energetically than composting, the sorting and

leaning processes may make this impractical (Kale et al., 2007)nd if biodegradation of these biopolymers has been triggered,hey will be unsuitable for recycling (Davis and Song, 2006).hese sources indicate that recycling is possible and research intohis is ongoing, however they are faced with similar issues asonventional polymer recycling.

. Discussion

.1. The Benefits of Biodegradability

Biodegradability appears to be viewed as a positive materialttribute with regards to environmental impact. For example,

vation and Recycling 78 (2013) 54– 66

Tabone et al. (2010) list this as a green design metric. One reason forsuch views could be the fact that these materials may be suitable forcomposting and AD at the end of their life, thereby reducing wastesent for incineration or landfill and the environmental impactsassociated with these processes. However, composting and AD alsohave negative impacts on the environment. In fact, the literaturereviewed here has demonstrated that incineration (with sufficientenergy recovery efficiency), an end-of-life option which does notrequire biodegradation of the materials, could have lower envi-ronmental impacts than those which do. This makes the currentenvironmental benefits of biodegradation questionable. It shouldbe stressed that there are large uncertainties related to the wastemanagement stage of LCA studies (Hermann et al., 2010). One rea-son is the lack of data on the extent of biodegradation of differentbiopolymers in the different environments which is important indetermining their suitability for that disposal route as well as theemissions generated and energy recovered (for methane capturedfrom landfill and AD). Further research into the extent of biodegra-dation of the different biopolymers is required as well as LCAstudies which consider a greater range of impact categories in orderto make a more comprehensive judgment. Other factors which maybe difficult to incorporate into an LCA study may also be impor-tant. For example, despite the unfavorable results of composting,Hermann et al. (2011) note that the end product of this, compost,may be a vital source of soil carbon in the future while Kale et al.(2007) highlights benefits of compost including moisture reten-tion, reduced erosion, and disease and pest control which can notbe achieved using alternative products. However, one could arguethat there are currently other feedstocks for composting which aremore appropriate.

6.2. Analysis of discrepancies in results

Studies on the same biopolymer have reported results that oftenvary substantially, sometimes resulting in conflicting conclusionson material preference. The discrepancies arise from a number ofsources, discussed in detail in this section. Some of these are relatedto the assumptions made about system boundaries and allocationmethods. There may also be a range of processes under analysis,resulting in genuinely different impacts. A major source of vari-ability arises from geographical differences in electricity generationmethods.

Determining the cause of the discrepancies has not been pos-sible in all cases due to lack of information in many studies.Although methodology is discussed, particular details regardingdata sources and choice of allocation methods are not always spec-ified. Few studies have provided impact scores for individual lifecycle stages which also increases the difficulties in identifyingthe source and reasons behind different results. When such detailhas been reported, it is often only for NREU and/or GWP. Somestudies have only presented their results as relative scores in refer-ence to the polymer used for comparison (e.g. Uihlein et al., 2008;Hermann et al., 2010) which does not allow for comparison withother studies. In these cases, the functional unit and the polymerused for comparison becomes more important since the propertiesand environmental impacts of petrochemical polymers differ fromone another. For example, Uihlein et al. (2008) have used PS as thepetrochemical polymer for comparison, which has higher environ-mental impacts than polyolefins such as PE (see Tables 1 and 3).In addition, the environmental impacts quoted for petrochemicalpolymers are not always consistent between studies and this couldhave an effect on final results. The range of impact scores for some

petrochemical polymers reported from various studies as well asthe data available from PlasticsEurope are shown in Figs. 1 and 3.

Discrepancies between studies can be attributed to differencesin numerous life cycle stages, as well as the system boundaries

onser

utebreaDgimadpdta

sNBemtTofaGsaidc

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evDreoWeTP(Teicgtsfc2

M.R. Yates, C.Y. Barlow / Resources, C

sed. The variability in agricultural and disposal practices andechnologies modeled in different studies will result in differ-nt environmental impacts reported. Additionally, the boundaryetween the technical and natural systems is not as well-defined,esulting in different boundary settings between studies. Hermannt al. (2010) also notes that there are large uncertainties associ-ted with the estimation of emissions from end-of-life processes.espite this, the agricultural and end-of-life stages are not thereatest sources of discrepancies in the most commonly stud-ed impact categories since their contribution to overall results is

uch less than the production stage. Discrepancies in the millingnd production stage can result from different allocation methods,ifferent production technologies, sources of energy and the incor-oration of credits from any energy produced, and the quality of theata used. All of these factors will be discussed further in relationo each material. The discussion is limited to differences in NREUnd GWP.

Impact results for PLA production, including the agriculturaltage for feedstock production, is provided in the articles foratureWorks (Vink et al., 2003, 2007, 2010) and PURAC (Groot andorén, 2010) on a mass basis. Different feedstocks grown in differ-nt geographical locations are used and the production methodsay differ although there is limited information on this. Both of

hese can explain differences in reported environmental impact.he level of detail provided only allows for a comparison of the GWPf different life cycle stages which shows that, despite these dif-erences, the impacts for combined agricultural and milling stagesre comparable, with sugar production having a slightly higherWP. However, it is standard practice to burn the bagasse in Thaiugar mills to generate steam and electricity which is taken intoccount as carbon credits (Groot and Borén, 2010). When this isncorporated, the GWP for sugar feedstock is lower than for cornextrose, despite the lower carbon sequestration assumed for thisrop (based on values reported in these studies).

The GWP for the production stage reported by Vink et al. (2010)as higher than that used by Groot and Borén (2010). Differences

n impacts in the production stage could be due to real variationss a result of technological differences including energy and rawaterial requirements, and waste produced. However, it could also

e a result of differences in data quality, source of electricity, andllocation methods applied. With respect to data quality, one studyay base their energy use on recordings from continuous measure-ents while another may estimate this from a breakdown of annual

onsumption.Other studies on PLA have used some data from NatureWorks,

ither from their published articles or the inventory in the ecoin-ent database v2.0 and v2.2 (Ecoinvent Centre Code of Practiceata, 2007; Ecoinvent Centre Ecoinvent data, 2010), both of which

eportedly use PLA6 as described in Vink et al. (2007) (Hermannt al., 2010; Gironi and Piemonte, 2010). Despite this, differencesccur depending on the date of the study, and the amount of Nature-orks data used. NatureWorks have published three LCA studies,

ach presenting reductions in environmental impact per kg PLA.herefore, older studies may overestimate the NREU and GWP ofLA. Some studies have included the RECs discussed by Vink et al.2007), while others replace this with fossil fuel derived electricity.he effect of RECs on LCA results has been discussed by Hermannt al. (2010) who argue that it should not be used when compar-ng technologies since any producer can purchase these. In thisase, electricity from wind farms put into the regular electricityrid is ‘purchased’, however there is no change in the actual elec-ricity delivered to the production site. Since the generation of the

ame quantity of electricity has lower primary energy requirementsrom wind than from fossil fuels, a reduction of 17 MJ/kg PLA islaimed with additional reductions in GWP and AP (Hermann et al.,010). This does not reflect any changes in technology and should

vation and Recycling 78 (2013) 54– 66 63

therefore not be used when comparing two materials. The energyrequirements given in MJ also differ substantially for fossil-derivedand other sources of electricity. The report by Boustead (2005) illus-trates how dramatically different the NREU of a material can bedepending on electricity generation parameters; this has also beendemonstrated by Khoo et al. (2010) and Suwanmanee et al. (2013).Therefore, major discrepancies can result between studies basedon their choice of electricity source. Electricity sources used in thePLA studies varied and included average European electricity fromSimaPro2007 (Hermann et al., 2010), average Thai electricity (Grootand Borén, 2010), and US grid electricity mix from Ecoinvent v2.0(Gironi and Piemonte, 2010).

Some studies may have used the farming and milling inven-tory data from the NatureWorks studies while others appear to useseparate sources. In these cases the allocation of co-products in themilling stage differ from system expansion (Uihlein et al., 2008),economic (Groot and Borén, 2010), and mass based (Gironi andPiemonte, 2010). Sensitivity analyses have indicated that this mayresult in small discrepancies but does not change the final resultssignificantly.

Differences in impacts from the agricultural stage are alsolikely. Lack of details in many studies makes it difficult to iden-tify methodological differences which affect the results. One sourceof GWP differences is the inclusion or exclusion of CO2 seques-tration from photosynthesis. This is included in the studies byVink et al. (2010) and Groot and Borén (2010) but excluded inthe study by Suwanmanee et al. (2013) which is significant whencomparing the amount sequestered to the total emitted through-out the lifecycle of the material. The incorporation of LUC canalso result in large changes in the overall GWP as shown bySuwanmanee et al. (2013). However, studies using other method-ologies have shown this to have little influence on the final results(e.g. Kendall, 2012), therefore the significance of its exclusion frommost studies is unknown. Discrepancies in results which representreal differences in environmental impact occur as a result of theassumed location of agricultural production since energy require-ments and practices can vary significantly. For example, the energyuse for corn cultivation in Nebraska is double that in Minnesotawhere high yields are achievable without irrigation (Shapouri et al.,1995).

LCAs on PHAs are subject to the same variations in agriculturaldata used as PLA since the same feedstocks can be used. Differ-ent allocation methods in the milling stage also exist. However,many of these studies have either considered different feedstocksor used agricultural data directly from earlier studies and it appearsthat differences in the agricultural stage do not contribute dramat-ically to the differences in reported results. Akiyama et al. (2003)does show that PHA from corn oil has lower impacts as a result oflower energy requirements in the agricultural stage and a greateryield of PHA per kg feedstock, while the NREU from ‘raw materi-als’ (including the agricultural and milling stage) for sugar-derivedPHA is similar to that for corn as originally reported by Gerngross(1999) and used in many other studies. However, the effect of differ-ences in yield and reduced NREU requirements due to the burningof bagasse for process steam and electricity cannot be extractedfrom the data provided on sugar derived PHA (Harding et al., 2007).Studies using other feedstocks do not provide the detailed datarequired to determine where discrepancies arise. Kim and Dale(2005) perform LCAs for three production technologies and reporthigher NREU than that presented in the original studies (Gerngross,1999; Akiyama et al., 2002) but it is unclear whether this is purelya result of using data for the agricultural stage which is more com-

prehensive than that used in other studies.

From the data available, the greatest discrepancies result fromdifferent production technologies with differing energy require-ments. More recent studies report lower NREU for production

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hich could represent improvements in technology. However itould also be attributed to the assumption that the residues fromermentation are burnt to produce steam and electricity; this isot taken into account by Gerngross (1999). Kim and Dale (2008)eported much lower NREU than any other study, as seen in Table 2,ased on data from an existing facility however this is explained byhe use of the integrated system to power the process. There is alsovidence of the use of different energy accounting methodologieshich can result in substantial differences in results. Using the mass

f fossil fuels required for PHA production and conversion param-ters provided by Gerngross (1999), a value of 80 MJ is obtained,hich has been used in other studies such as Khoo et al. (2010).owever, Harding et al. (2007) have calculated energy require-ents of 50.4 MJ/kg PHA based on the mass of fossil fuels quoted

y Gerngross (1999). Therefore, the reduced energy requirements44.7 MJ) of sugarcane-derived PHA reported by Harding et al.2007) may be largely due to methodological differences ratherhan technological improvements made in-between the publica-ion of the two studies.

Some LCAs for both PLA and PHA have been product specific.ost of these have taken into account the additional mass require-ents to make an equivalent product using biopolymers. While

his increased the environmental burdens of biopolymer productsn comparison to the petrochemical polymers used for comparison,

aterial preference was only affected in one case in the study byietrini et al. (2007) where PHA-based composites were used forar panels. In this case, the increased mass of the product was ofreater significance as it results in greater fuel requirements duringhe ‘use’ stage of the car. Other studies have focused on disposableroducts such as food packaging where the use-phase is assumed toave no or equivalent environmental burden, regardless of materialr mass used.

Some product-specific studies have also considered differencesn product manufacture while others assume this to be identical.his only had an influence on the results in the study by Hermannt al. (2010) as a result of using real data collected from film pro-uction lines which are less energy efficient for the small-scaleperations used for PLA at the time of data collection. While thisighlights that PLA products may currently have higher energyequirements, this is not related to the material itself, rather theacilities used for down-steam processing.

Product-specific studies were also more commonly location-pecific. In such cases, different transport requirements forifferent materials were taken into account. The different productass requirements would also be incorporated resulting in greater

mpacts from transportation for products using a greater mass ofolymer due to increased fuel requirements. Where given, resultshow that the contribution from transport is small compared toroduction, therefore this also does not contribute significantlyo discrepancies. One exception is the study by Madival et al.2009) who report significant contributions for transport of poly-

er resin and finished containers such that this influences theesults depending on the assumed location of resin producer rel-tive to the container manufacturer. This study considered theransportation by truck for long distances over 4000 km across theSA, a much greater distance than those assumed in other stud-

es.Studies which include the end-of-life stage have demonstrated

hat this is not a major contributor to the environmental impactstudied. Therefore, while discrepancies exist between studieselated to this stage, the overall contribution is small, with recyclingeing an exception which can change the material preference.

iscrepancies in this stage are due to different assumptions on

he extent of degradation of biopolymers, allocation of emissionsrom composting, and differences in technology such as incineratornergy recovery efficiency and landfill gas capture.

vation and Recycling 78 (2013) 54– 66

The data on starch-based polymers have mainly been obtainedfrom review articles as the original articles are not available inEnglish. As a result, it is difficult to discuss where differences inthe results presented in the reviews arose from. The two originalstudies reviewed here were cradle-to-grave and assessed differ-ent products with the results presented for the entire life-cycleonly as relative scores making comparisons difficult. However, theauthor of one has also co-authored the second paper and there-fore it is highly likely that identical methodologies have been usedand any differences in results are due to product-specific require-ments. The same database, ecoinvent v2.2 (Ecoinvent Centre, 2010)is used for the inventory data which reportedly uses environmen-tal performance data on MaterBi from Novamont using corn as afeedstock (Piemonte, 2011). Many of the studies discussed in thereview paper by Patel et al. (2003) have also used MaterBi mate-rials as the starch polymer, however there are different grades ofthis material containing various proportions of different biodegrad-able but petrochemical polymers (PVA and PCL). The starch-basedmaterials studied have ranged from 100% down to 40% starch con-tent and feedstocks have included corn, wheat, and potato (Patelet al., 2003). The effect of increasing the proportion derived frompetrochemical sources was shown to be an increase in energyrequirements and GWP (Patel et al., 2003). The use of differentfeedstocks grown in different countries is also expected to be asource of discrepancies in the results of these studies. Addition-ally, the studies presented in the reviews were published 10 ormore years earlier than those by Piemonte (2011) and Gironi andPiemonte (2011) in which time technology and data quality mayhave improved. While the feedstock, geographic location, and ageof the study are likely to contribute to discrepancies in results, theproportion of starch in the material may be the most significantfactor.

In summary, there are many sources of discrepancies, not allof which can be identified due to lack of detail. They include truedifferences in environmental impact as well as differences causedby methodological choices. These exist in all life cycle stages, how-ever, since the production stage is the dominating contributor toNREU and GWP, it is likely to be the greatest contributor to varia-tions in results. Due to the complexity of LCAs and the differencesresulting from location and product specific aspects, the scope forimproving their comparability may be limited. However, inclusionof intermediate data in all studies would be helpful. For exam-ple, impact scores for the agricultural stage could be provided,with all assumptions and geographical modifications clearly stated.While discrepancies caused by geographical differences in practice,technology, and climate, and by product specific requirements can-not be overcome, differences in methodology could potentially beeliminated or reduced to minimize confusion in understanding theoutcome of different studies.

Perhaps the most important aspect relates to electricity sources.Since this, along with steam generation, have been reportedas dominating contributors to most impact categories (Hardinget al., 2007), results are likely to be sensitive to any changesin energy supply. Energy consumed, in Joules (J), in the form ofelectricity is heavily affected by the fuel mix used, and efficien-cies in production and delivery which differ between regions andcountries (Boustead, 2005). This results in discrepancies betweenstudies, but will also affect the results within studies when mate-rials being compared are assumed to be produced in differentlocations, therefore using different electricity sources. For truecomparability of LCAs on materials from a technological view-point, energy requirements should perhaps also be represented in

alternative units. At the very least, a further breakdown of energyrequirements should be provided such that electricity use can becompared more easily and the effects of different sources are moreclear.

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.3. From LCA data to a material choice

Comparative studies which only looked at NREU and GWP coulduickly draw conclusions on material preference. This becomesore difficult when additional impact categories are considered

ince lower environmental impacts in some categories often comet the expense of increased burdens in other categories. As men-ioned in Section 2.1, combining the results into a final score tondicate material preference does not have any scientific basis andherefore many studies have discussed the results without draw-ng solid conclusions on material choice. The studies by Gironind Piemonte (2010), Piemonte (2011), and Gironi and Piemonte2011) have taken the results one step further using normalizationnd grouping into the Ecoindicator-99 damage categories and pre-enting these in mixing diagrams. The mixing diagrams allow forisualization of the preferred material for all possible combinationsf weighting placed on each category. Each point inside the triangleepresents a combination of weights assigned to the three damageategories adding up to 100% (Gironi and Piemonte, 2011) and thereferred option at each point is shown. If one material dominateshe area inside the diagram, one may be more confident in suggest-ng it has lower overall environmental impacts in comparison. If theiagram is more evenly split, a decision would be more difficult.

.4. Future improvements

Since the production stage was identified as the major contribu-or to environmental burdens of PLA and PHA, the use of renewablenergy has been suggested as a future improvement (Kurdikar et al.,000; Vink et al., 2003; Groot and Borén, 2010; Hermann et al.,010; Gurieff and Lant, 2007; Khoo et al., 2010), impacting AP asell as NREU and GWP. One feasible example of renewable energyse is the integrated system described earlier.

Another suggested improvement has been the use of alternativeeedstocks, which might be either waste- or by-products or lessnergy and chemical intensive crops (Kurdikar et al., 2000; Vinkt al., 2003; Groot and Borén, 2010; Hermann et al., 2010). Largecale production based on waste may be logistically difficult. Inddition, studies such as those by Kendall (2012) and Zhong et al.2009) indicate little improvement in the environmental perfor-

ance of waste-product derived biopolymers.Lastly, improved farming practices have been suggested includ-

ng no-tilling approaches (Kim and Dale, 2005), efficient fertilizerpplication by precision knifing, and the planting of buffer zonesround crops (Landis, 2010).

Gironi and Piemonte (2010) point out that biopolymerechnologies are new and are being compared to mature, opti-

ized technologies. Biopolymers made using environmentalest-practice methods may have lower impacts than petrochemicallastics, but this is for the future and demands further development.dditionally, further improvements in petrochemical polymers

s also possible, something that has generally been dismissedased on the assumption that the technologies have already beenptimized. The latest eco-profile of PET shows significant reduc-ions in all impact categories including a 16% reduction in NREUnd 50% reduction in AP, explained largely by process efficiencymprovements in the production of purified terephthalic acid (PTA)PlasticsEurope, 2011). Similarly, the updated eco-profile for PS hasignificant reductions in environmental impacts thought to be dueo a number of factors including the source of benzene, strictermission controls, and changes in external energy supply (grid elec-

ricity) (PlasticsEurope, 2012). Therefore, while improvements arexpected for biopolymers, simultaneous improvements in petro-hemical polymers may reduce or remove the comparative benefitsf biopolymers.

vation and Recycling 78 (2013) 54– 66 65

7. Conclusion

This review of existing LCAs on PLA, PHA, and starch based poly-mers has demonstrated that reductions in NREU and GWP can beachieved, however higher impacts in other categories were com-monly reported making it difficult to determine which materialscould be considered least detrimental to the environment. Thesituation is further complicated by discrepancies in the reportedresults. Some of these can be attributed to methodological dif-ferences while others reflect different environmental impacts dueto geographical location and differences in the processes gener-ating the same end product. The current picture is confusing anddefinitive conclusions are difficult to draw although the studiesreviewed suggest that these biopolymers may not necessarily bemore environmentally friendly than the petrochemical polymersthey could replace at this time. However, trends in studies showthat the environmental profile of these biopolymers is improvingand may continue to do so in the future.

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