Time Horizons A central issue in the discussion of the handling of temporary carbon storage in LCA is the use of time horizons. When assessing climate change impacts in LCA, a time horizon of 100 years is often used for the characterisation factor, reflecting the time horizon adopted in the Kyoto Protocol. In approaches for assessing the value of temporary carbon storage, the use of such a time horizon is by some interpreted as implying that impacts occurring after this time should not be included in the assessment (e.g. Moura-Costa 2002; Clift and Brandão 2008). To distinguish between these interpretations, this latter interpretation, of disregarding all impacts occurring after the time horizon, is here referred to as ‘accounting period’. It is important to understand the major difference between the time horizon used in the Kyoto Protocol and such an accounting period


It is important to understand the major difference between the time horizon used in the Kyoto Protocol and such an accounting period. The 100-year time horizon adopted in the Kyoto Protocol includes impacts of the first 100 years of every greenhouse gas emission, regardless of when the emission occurs, cutting off only the so-called ‘tails’ of the emissions that are left 100 years after the emission. For CO2 emissions, for example, approximately 20% stays in the atmosphere for many thousand years (Archer et al. 1997). The approach is illustrated in Fig. 30.2: In contrast to this, the accounting period, suggested by some for temporary carbon storage crediting, starts counting from the time the carbon is stored and cuts off every impact occurring after the accounting period, regardless of when along the accounting period the carbon is released again

Hauschild et al., LCA: Theory and Practice


This means that some of the impacts from an emission will be ‘hidden’ or completely disregarded due to the accounting period, which is a very different result than using the time horizon as adopted in the Kyoto Protocol (Jørgensen and Hauschild 2013; Brandão et al. 2013). For example, if carbon is stored for 30 years before being released, using the 100-year time horizon will still give a 100 year impact inclusion whereas the 100-year accounting period will only give an inclusion of 70 years impact. The difference in the two ways of interpreting time horizons is illustrated in Fig. 30.3. 0.5 Part of CO2 emission leŌ in the atmosphere Years since emission Cut off of atmospheric CO2 'tail' aŌer 100 years from emission, using the 100-year Ɵme horizon Fig. 30.2 Implication of the 100-year time horizon


30.2 Implication of the 100-year time horizon. The black line indicates the fraction of a CO2 emission left in the atmosphere as a function of time after emission, whereas the grey area represents the ‘tail’ of the emission, which is cut off due to the use of the 100-year time horizon (modified from: Jørgensen and Hauschild 2013) S.V. Hjuler and S.B. Hansen


The use of a 100-year accounting period can thus not be justified with reference to the normal use of a 100-year time horizon in LCIA of climate change impacts. As the choice is essential for the outcome, it should be ensured that it properly reflects the real climate change mitigation value of temporarily storing carbon (e.g. Jørgensen and Hauschild 2013; Brandão et al. 2013). Another time-related issue is the lifetime of the product, which determines the length of temporary carbon storage and thus influences the potential value of the storage relative to the product carbon footprint. Since product lifetimes are generally very short compared to the duration of impacts from GHG emissions, temporary carbon storage does not change much in terms of long-term climate impacts (Jørgensen and Hauschild 2013). Note that terminology for above-discussed types of time horizons differ in the literature, and may often not be clearly distinguished


Note that terminology for above-discussed types of time horizons differ in the literature, and may often not be clearly distinguished. Tipping Points The issue of long-term climate change implications is, however, not the only relevant issue when considering the potential role of carbon storage in mitigating climate change. Due to increased global warming, there is also an issue of urgency, with the risk of passing critical climate change levels, so-called climatic ‘tipping points’. A tipping point is a point where a structural change occurs in a system, which starts a ‘chain reaction’, meaning that it is no longer external forcing, but rather internal mechanisms in the system, which drives the process of change (IPCC 2007). Crossing such a tipping point is expected to lead to dramatic climate system changes which may be virtually irreversible (IPCC 2007), meaning that it is not realistic to return to the situation as it was before crossing the tipping point


Mitigation of the rise in atmospheric GHG concentration is urgently required if the passing of expected climatic tipping points should be avoided. Thus, even short-term storage of carbon may have value in terms of climate change mitigation, if it can add to staying below such climatic tipping points, until more permanent solutions are reached (Jørgensen et al. 2015). It is therefore suggested to distinguish between long-term and short-term impacts when addressing the climate change 10 15 20 25 30 35 40 45 50 55 60 65 70 75 80 85 90 95 100 105 Years of impact included Years of product storage before release Years of impact inclusion aŌer GHG release following temporary storage 100-year TH 100-year AP Fig. 30.3 Difference in using a 100-year time horizon (TH) and a 100-year accounting period (AP)


30.3 Difference in using a 100-year time horizon (TH) and a 100-year accounting period (AP). Inclusion of impacts after a GHG emission, following a certain time of temporary carbon storage in bio-products depending on whether a 100-year TH or AP is used LCA of Biofuels and Biomaterials


mitigation value of temporary carbon storage (Jørgensen and Hauschild 2013; Jørgensen et al. 2015). 30.3.2 Biodegradability There is a widespread misconception that all biomaterials are biodegradable, making them more sustainable than, and distinguishing them from, petrochemical materials. There are several mistakes in this belief. In the first instance, biomaterials are not necessarily biodegradable, while some petrochemical materials can be biodegradable. Biodegradability is a material property, which depends on the molecular structure of the material, not the feedstock (PlasticsEurope 2013). In fact, materials can be identical once produced, regardless of whether they were made from a biomass or petrochemical feedstock. Second, biodegradability does not inherently equal sustainability


In fact, materials can be identical once produced, regardless of whether they were made from a biomass or petrochemical feedstock. Second, biodegradability does not inherently equal sustainability. Even if biodegradability does increase environmental sustainability of a certain material in the disposal stage, there may be other impacts in the life cycle of that biodegradable product that cancel out or overshadow the benefit. However, in some cases, being biodegradable may also in itself add no benefit, or may even have a damaging effect, relative to the environmental sustainability profile. In some countries, plastic is incinerated to produce heat and power, thereby replacing alternative fuels that are often fossil. The concept of using biomass first for products and subsequently for energy recovery through incineration is referred to as ‘carbon cascading’ (Weiss et al. 2012)


The concept of using biomass first for products and subsequently for energy recovery through incineration is referred to as ‘carbon cascading’ (Weiss et al. 2012). If instead composting biodegradable materials, the benefits from the composting need to be counterbalanced against what is lost from the heat and power recovery from waste incineration. A study of the biopolymer polylactic acid (PLA) has shown that product-specific GHG emissions may vary by 20% depending on whether or not energy recovery from incineration is included (Hermann et al. 2011). In case biodegradable materials end up being incinerated, which is likely under the current waste management situation in a country like Denmark, it does not matter whether they are biodegradable. If biodegradation takes place under anaerobic conditions, the carbon in the material will be converted into a mixture of methane and CO2 by digestion, rather than just converted to CO2 as under aerobic conditions in e.g. composting


composting. If anaerobic digestion takes place in a biogas plant or a landfill with capture of landfill gases, the methane can be used to substitute fossil energy. However, if taking place in landfills where the methane is not captured, or in poorly managed home composting systems leading to anaerobic conditions (Song et al. 2009), as applicable in many places in the world, the methane production may lead to higher overall climate change impacts from biodegradable materials than for non-biodegradable ones, due to the relative high characterisation factor of methane compared to CO2 S.V. Hjuler and S.B. Hansen


(e.g. Patel et al. 2005). When conducting studies on biodegradable polymers, composting is often considered as the waste handling method, whereas the option of digestion is rarely included (Patel et al. 2005). Some studies suggest that composting biodegradable materials can be more beneficial than incineration, if the compost is used for agricultural carbon soil replenishment (e.g. Weiss et al. 2012; Hermann et al. 2011). For the issue of potential storage of carbon in the composted material, however, it is most likely that between 80% and all of the initially sequestered carbon will be released during composting of biodegradable polymers, due to their ability to rapidly decompose (Patel et al. 2005) in which case there will not be much left for soil replenishment and storage in the soil


2005) in which case there will not be much left for soil replenishment and storage in the soil. Furthermore, composting biomaterials only have a nutritional value for the soil if the biomaterial in question includes nutrients such as nitrogen, phosphorous or potassium, which is normally not the case. If, on the other hand the biomaterial does include nitrogen, then composting includes the risk of potential emissions of N2O, which due to its large global warming potential can reverse any benefits from the compost (IPCC 2006). Finally, for using biodegradable products in compost and soil replenishment, all substances in the bio-products have to comply with requirements for this use, such as heavy metal thresholds (e.g. Song et al. 2009). Biodegradability is therefore a property like many others, which may or may not have a positive impact on the environmental profile of a product when considering all life cycle impacts under the relevant circumstances


30.4 Land Use Change, LUC Three major drivers for land use change (LUC) exist globally: (1) Timber harvesting, (2) Infrastructure development, (3) Agricultural/horticultural expansion (Kim et al. 2009). Often, the LUC is due to a combination of the above. Biofuels and biomaterials are by default directly linked to the agri- and horticultural sectors and direct and indirect land use change will in most cases contribute significantly to the life cycle impacts of bio-based products. In some cases, LUC impacts can be higher than the impacts of the rest of the life cycle impacts combined. Potential LUCs related to bio-product crops are: 1. Forest or other virgin land to bio-product crop 2. Agricultural crop to bio-product crop 3. Fallow land to bio-product crop 4. Urban land to bio-product crop (happens only on a very small scale and is not dealt with here) LCA of Biofuels and Biomaterials


30.4.1 Direct and Indirect Land Use Change Direct land use change (DLUC) is the man-made conversion of a piece of land from a previous land use (like another crop type or forest) to a bio-product crop. Indirect land use change (ILUC) occurs when land carrying a crop ‘A’ is converted to a bio-product crop and market demand for product(s) that are derived from the earlier crop ‘A’ (directly or through other indirect crop displacements) drives conversion of marginal land elsewhere to meet that demand. This process is shown in Fig. 30.4. DLUC is often followed by ILUC if the DLUC takes place on previously cultivated land. Marginal land describes the land ultimately most likely to be converted. Unless there is an unlikely decline in demand for agricultural land in the region, the marginal land will be virgin land like forest or abandoned land like secondary forest or fallow


Unless there is an unlikely decline in demand for agricultural land in the region, the marginal land will be virgin land like forest or abandoned land like secondary forest or fallow. Example 30.3 If rapeseed oil in Europe is used for biodiesel production, less vegetable oil will be available on the market. Palm oil, which is produced mainly in Malaysia and Indonesia, is currently the cheapest and fastest growing vegetable oil crop in the world and as such it is the marginal oil crop (i.e. the oil most likely to be produced if vegetable oil demand increases). Thus, the supply deficit in vegetable oil when rapeseed oil is used as fuel will most likely be covered by additional production of palm oil in Malaysia and Indonesia. The marginal land in Malaysia and Indonesia is to a large extent forest, so the production of biodiesel from rapeseed in Europe will likely result in deforestation in South East Asia


The marginal land in Malaysia and Indonesia is to a large extent forest, so the production of biodiesel from rapeseed in Europe will likely result in deforestation in South East Asia. Note that in this example no direct land use change occurs when producing fuel from rapeseed oil as only the use of the crop changes. As ILUC is highly dependent on market forces, and it creates scenarios that are beyond the influence of farmers and bio-product producers, it becomes highly uncertain to predict or model, and a formal consensus has not been reached for a methodology to include ILUC in LCA. 30.4.2 Impacts of Land Use Change Historical and current LCA literature on bio-products focuses mainly on carbon emissions from LUC, but other impacts like biodiversity loss, water use impacts and biogeophysical impacts like albedo changes are starting to attract more interest


Biodiversity The loss of biodiversity is a major environmental concern, and habitat loss and (local) species extinction due to a changed land use is expected to be the main driver of biodiversity changes in terrestrial ecosystems (Sala et al. 2000). However, S.V. Hjuler and S.B. Hansen


approaches on how to handle the consideration of biodiversity impacts in LCA are still on a rather preliminary level (e.g. Chaudhary et al. 2015; de Baan et al. 2013; Koellner et al. 2013; Michelsen 2008). The current lack of consensus on an assessment method (which over time may be solved by the UNEP-SETAC Life Cycle Initiative’s effort to establish a global consensus on LCIA models and methods) and the lack of geographical specific data are main obstacles for obtaining quantitative and reliable results of biodiversity impacts in LCA. However, due to the importance of this aspect, obtaining qualitative or rough quantitative results of best available approaches is considered better than disregarding biodiversity impacts. Albedo Albedo is a measure of how much of the incoming solar radiation is reflected by a surface. Albedo values vary with seasons and differ a lot depending on the type of land cover


Albedo Albedo is a measure of how much of the incoming solar radiation is reflected by a surface. Albedo values vary with seasons and differ a lot depending on the type of land cover. Surfaces covered in snow and ice thus have much higher albedo than darker surfaces, for instance areas covered with forest (e.g. Cherubini et al 2012a). The effect of increased albedo is cooling, whereas decreased albedo of a surface leads to warming. As LUC can lead to substantial changes in the albedo of an area, which can play an important role in terms of climate impacts, albedo changes is increasingly included in climate impact assessments of biofuel systems (e.g. Bright et al. 2012a, b; Cherubini et al. 2012a)


Bright et al. 2012a, b; Cherubini et al. 2012a). Reliability of results depends on availability Biofuel and feedstock replacing other crop AlternaƟve/syntheƟc producƟon Replacing other crop Induced impacts Other market changes Crop price increases (Similar/marginal) Crop planted elsewhere Less demand for replaced crop MiƟgated or no indirect impact Yes No (or insufficient) Intensified agricultural producƟon of the crop* No (or insufficient) Yes Replacing marginal virgin land (forest) Fig. 30.4 Potential ILUC effects can be very complex making accurate predictions very difficult. The figure provides a simplified overview of the potential consequences of LUC. ILUC can be avoided or mitigated by less demand for—or intensified agricultural production of—the replaced crop. Asterisk Intensified agricultural production can induce impacts through e.g. increased use of chemicals and soil degradation LCA of Biofuels and Biomaterials


of relevant geographical data for similar land uses. Other biogeophysical effects exist as well, but on a global scale the albedo effect is the dominating direct climate forcing of these, especially in locations with seasonal snow cover (Claussen et al. 2001; Randerson et al. 2006; Bala et al. 2007). 30.4.3 Carbon Pools Through the conversion of land a number of carbon pools are affected (Germer and Sauerborn 2008): • Above ground biomass: All living plants above ground • Below ground biomass: All living biomass in the soil; e.g. roots • Litter and deadwood: Fallen trees, leaves and branches • Soil carbon: Organic carbon residues left in the soil after degradation of biomass The various carbon pools vary greatly depending on climate, vegetation and soil type. The net carbon balance for LUC is the difference in the four carbon pools between the previous land use and the bio-product crop land use


The net carbon balance for LUC is the difference in the four carbon pools between the previous land use and the bio-product crop land use. LUC can thus result in net emission as well as a net sequestration of carbon depending on the carbon stored in the previous land use and the bio-product crop land use. Note that carbon sequestration credits are the difference in carbon stock at the beginning and the end of the assessment time horizon. If the land is cleared in the last crop cycle of the assessment time horizon then no permanent carbon sequestration should be credited. Thus, unless there is documented reason to believe that the land will be left with a certain carbon pool at the end of the assessment time horizon then no permanent carbon sequestration should be credited. In the case of plantations, which are replanted, e.g. every 30 years, it can be argued that there is temporary carbon storage for these 30 years


In the case of plantations, which are replanted, e.g. every 30 years, it can be argued that there is temporary carbon storage for these 30 years. However, as the carbon is sequestered and stored approximately linearly over the 30 year period, the average retention of the carbon is only half of that, i.e. 15 years. If the goal of the study requires inclusion of temporary carbon storage, then the temporary carbon storage model from ILCD (European Commission 2010) could be applied (see Sect. 30.3.1). Applying this, the temporary carbon storage in the plantations should be credited 15% of the full carbon storage potential (see application in e.g. Hansen et al. 2014). It must be clearly stated that the storage is biogenic. If organic residues of the bio-product crop (e.g


Hansen et al. 2014). It must be clearly stated that the storage is biogenic. If organic residues of the bio-product crop (e.g. trunks) are treated and stored in a product thereby preventing the release of the fixed carbon, then this temporary carbon storage should not be inventoried under LUC, but rather under the production stage of the bio-product, in which the residue is generated. Although biomass is often considered CO2 neutral as the carbon has been captured recently, this is not the case if the biomass is from virgin sources, such as virgin forest, where there has been carbon equilibrium for thousands or millions of years, or if such land has been cleared for enabling the production of the biomass. So even though the individual trees/plants may not be old, the carbon stored in the land must be considered a permanent carbon pool, which is lost if the forest is converted to S.V. Hjuler and S.B. Hansen


agriculture. All virgin forest conversion emissions should thus be inventoried as fossil carbon dioxide (European Commission 2010). Before a forest is converted to agricultural land it is most often logged for valuable timber. The extraction of trees suitable for timber has a market value and can in attributional LCA be considered as belonging to a separate product system. The carbon stored in the timber trees should thus not be counted in the LUC emissions in the bio-product system. In a consequential approach it is necessary to consider whether the timber is substituting timber harvesting somewhere else, in which case the carbon stored in a similar quantity of marginal timber can be subtracted from the total carbon loss from the LUC including the timber


Special local conditions, which are not covered specifically in the ILCD Handbook (European Commission 2010), can occur as per following example: Example 30.4 In the case of palm oil derived biodiesel from Malaysia, the Malaysian government has selected areas still under forest cover for timber extraction and future development, e.g. oil palm plantations. These areas have alreadybeen or will beharvested for timber byselective logging.Clear cutting is only used if a plantation is established immediately after the logging. The extracted timber would thus be cut whether an oil palm plantation is established or not and thus felling of that timber in the virgin forest should in this case not be allocated to the palm oil biodiesel, but to the timber production. The emissions from a logged-over forest would thus be more suitable to use in the assessment of the palm oil production


The emissions from a logged-over forest would thus be more suitable to use in the assessment of the palm oil production. However, if the logged-over forest was left idle it would likely recover and the actual loss of carbon when converting to oil palm is thus that of a recovered logged forest. ILCD recommends that the emissions from land use change should be allocated to the first 20 years of agricultural use or in the case of plantations the first planting cycle, which is often 20–30 years (more details on timing issues can be found in Sect. 30.4.4). Applying this amortization period, the emissions from conversion of forest (which is harvested for timber before conversion) should be those of a logged-over forest after 20–30 years of recovery. The emissions from the logged-over forest should still be counted as fossil emissions as per above whereas the carbon sequestered during recovery is biogenic. Note that with a few changes in details or assumptions (e.g


Note that with a few changes in details or assumptions (e.g. that the logged forest dies rather than recovers), this example could turn out differently. ILCD (European Commission 2010) labels the emissions from conversion of secondary forests, crops and plantations to bio-products crops as biogenic carbon dioxide. The IPCC (2006) provides generic carbon data for various land uses under various climatic conditions. It is recommended to use IPCC data when no site or regional specific data is available (European Commission 2010). However, it must be stressed that—as also highlighted in IPCC (2006)—carbon stocks even in forests of similar type and geography vary significantly and using generic/mean values would result in large uncertainties. LCA of Biofuels and Biomaterials


30.4.4 Timing issues and Payback Time Immediate Emissions Immediate LUC emissions are emissions taking place during the conversion of the land. Above and below ground biomass as well as litter and deadwood are included. In accordance with the ILCD (European Commission 2010) guidelines, LUC emissions for conversion to sub-annual, annual and bi-annual crops should be attributed to the first 20 years of the new land use, unless it is known that the new land use period will be shorter than 20 years. Thus, for any given year the production at that piece of land should be allocated 1/20 of the total immediate LUC emissions. For a plantation with trees or palms, it is recommended that the emissions are allocated to the first cycle of trees/palms, i.e. 20–30 years. Long-Term Emissions The degradation or sequestration of soil carbon takes place over a number of years until a new equilibrium is established


20–30 years. Long-Term Emissions The degradation or sequestration of soil carbon takes place over a number of years until a new equilibrium is established. As for the immediate emissions, it is recommended to allocate the emissions/sequestration over 20 years unless the actual period, when 90% of the emissions/sequestration have occurred is known, in which case that number is used as n in Eq. (30.2) (European Commission 2010). As opposed to the immediate emissions, the long-term emissions cannot be allocated linearly over the allocation period. Instead, larger fractions of the total emissions are allocated to the earlier years: x 1⁄4 100  2 ð Þ= n þ 1 ð Þ  n  i ð Þ=n; ð30:2Þ where • x is the % of the total carbon inventory allocated to year i • n is the amortization period for the total carbon allocation, e.g. 20 years • i is the number of years after the LUC (i  n) Example 30.5 Peat soils are soils with very high organic contents


20 years • i is the number of years after the LUC (i  n) Example 30.5 Peat soils are soils with very high organic contents. They are created through biomass accumulation under water logged, anoxic conditions, which slows the organic degradation and makes the accumulation rate higher than the degradation rate. When peat land is converted to agricultural land the peat is drained, thus exposing the peat to oxygen and increasing the rate of decomposition. Without proper land management, the oxidation of the peat can continue for several decades and emit carbon dioxide in quantities several times larger than the immediate emissions from land clearing depending on the site specific soil conditions. Special considerations must thus be given to the management practices as well as site specific soil conditions when conducting an assessment of a bio-product crop planted on peat soil. Whereas the largely homogenous peat in temperate climate regions has been studied S.V. Hjuler and S.B. Hansen


intensively, the large localised variations in tropical peat composition as well as effects of diverse management practices of cultivated peat land have not yet been fully understood in terms of potential and actual carbon emissions. The large ranges of emissions from cultivated tropical peat are evident from IPCC (2014) and these ranges may increase further as more data is produced from the numerous peat composition and management practices scenarios. Note that peat land conversion to palm oil plantation is not uncommon in, e.g. Indonesia. As palm oil is the marginal vegetable oil on the world market, this can have impacts on all other biodiesel feedstocks through ILUC, as explained in Sect. 30.4.1. For both immediate emissions and long-term emissions no allocation to individual years is needed if the assessment uses an average of the emissions from the bio-product crop of a period equal to or longer than the amortization period


In that case, the total LUC and ILUC emissions are simply added to the rest of the life cycle emissions (European Commission 2010). Payback Time Payback time is often used to describe the time it takes a bio-product to pay back the carbon emissions related to LUC; the so-called carbon debt. Almost all bio-products save carbon emissions compared to fossil products when (indirect) LUCs are not taken into consideration, but the picture can change when including these. By dividing the total LUC carbon emissions for a hectare of land converted to bio-product crop production with the carbon savings from replacing the fossil products with bio-products produced from that hectare for a year (not including LUC emissions) it is calculated how long it will take before the bio-product system actually starts saving carbon emissions. Example 30.6 A land use change has resulted in a total emission of 50 tonne CO2-eq/ha


Example 30.6 A land use change has resulted in a total emission of 50 tonne CO2-eq/ha. 1 tonne biodiesel is produced per ha/year and the carbon emissions from the biodiesel production are 1.7 tonne CO2-eq/tonne biodiesel. The emissions for extraction, refining and combustion of fossil diesel are 2.7 tonne CO2-eq/tonne biodiesel equivalent (biodiesel has a lower heating value than fossil diesel, so less than 1 tonne diesel is needed per tonne biodiesel equivalent). The net saving for the biodiesel production is thus 1 tonne CO2eq/tonne biodiesel. Paying back the land use change emissions will thus take 50 years. Only then will the biofuel system actually start saving CO2 emissions. (Note that values used in the example are generally realistic values but can vary from case to case.) LCA of Biofuels and Biomaterials


30.5 Concluding Remarks and Perspectives • Competition for Biomass Due to limitations in land availability, only the most optimistic outlooks expect availability of enough sustainable biomass to cover future replacement of all replaceable fossil fuels and products with bio-products, while still covering demands for food and feed of a growing population (see e.g. Jørgensen 2014). Therefore, the biomass use needs to be based on prioritisation. • Integrated Biorefineries As mentioned in Sect. 30.2, the biorefinery concept is already to some extend existing today and breakthroughs within this area are expected in the coming years (Weiss et al. 2012). Biorefineries are expected to provide the optimal utilisation of biomass for production of both biofuels and biomaterials, resulting in a maximised utilisation of the biomass feedstock. However, there are today still significant challenges in reducing production cost


However, there are today still significant challenges in reducing production cost. • Maturity of Competing Technologies While the petrochemical industry has been optimised through a long time period and is based on mature technology, production of many bio-products is still relying on rather new technologies. Some types of biofuels can to a certain extent make use of conventional processes, but this is often not the case for biomaterials (Weiss et al. 2007). Furthermore, many bio-products are not yet available in commercial scale. Thus, there is a rather large potential for optimising efficiency, while substantial improvements are also expected for the integrated production of biofuels and biomaterials in biorefineries (Patel et al. 2005). Major reduction potentials in environmental impacts from biomass feedstock production have also been identified to be obtainable if changing agricultural management practices


2005). Major reduction potentials in environmental impacts from biomass feedstock production have also been identified to be obtainable if changing agricultural management practices. For example, changing to no-tillage farming, including cultivation of winter crops or changing to extensive farming, as well as utilising agricultural residues for biofuel or biomaterial production, except the fraction that is essential for preserving soil organic carbon (Weiss et al. 2012). However, it should be kept in mind that some changes in farming practices can also lead to lower productivity and thus need for additional land use. • Future Bio-product Feedstocks As mentioned, most current general conclusions on environmental sustainability of bio-products relative to fossil products relate to 1G, whereas it is expected that the environmental performance of 2G will have substantial potential to improve


The potentials lie both in reduced GHG emissions and non-renewable energy use when substituting fossil products, compared to 1G (e.g. Dornburg et al. 2008) and in reduced impacts in terms of water use and water quality, unless the feedstock S.V. Hjuler and S.B. Hansen


crops are from irrigated plantations (Gnansounou 2010). Compared to feedstocks for 1G bio-products, feedstocks for 2G bio-products generally have lower need for agricultural input such as fertiliser, pesticides and irrigation (e.g. Dohleman et al. 2010) have higher soil organic carbon sequestration potentials (e.g. Anderson-Teixeira et al. 2009) and can reduce erosion (e.g. Somerville et al. 2010). Furthermore, feedstock for 2G bio-products can decrease the land use compared to that for 1G bio-products, and especially the use of agricultural land. A major part of the reason for improvement expectations for 2G is that it can utilise biomass feedstock types which can grow on land not suited for agricultural production, as well as agricultural and forest residues. Such feedstocks are not competing with global food production, and in the case of residues, this feedstock is in some cases even seen as a waste product


Such feedstocks are not competing with global food production, and in the case of residues, this feedstock is in some cases even seen as a waste product. However, currently much of such residues are left in the fields/forests, adding to soil carbon replenishment. The aspect of soil carbon replenishment should thus be considered when assessing the sustainably available amount of crop residues, e.g. by considering a minimum amount of residues to either be left in the field or be earmarked for biochar production through pyrolysis, which also produces bio-oil. Biochar is mostly non-degradable carbon, which will potentially stay in the soil for millennia to avoid soil degradation and act as carbon sequestration (Lehmann 2007). Bio-products based on algae biomass (3G) also present some interesting perspectives


Bio-products based on algae biomass (3G) also present some interesting perspectives. Promising features of algae biomass as feedstock for bio-products are that they do not compete for agricultural land, there is a large production potential, and CO2 capture and biomass yields are high. Furthermore, a range of products are available from microalgae, including hydrogen, biogas, biodiesel, bioethanol and other starch based bio-products (Posten and Schaub 2009). However, there are also challenges, e.g. in terms of power requirements and production cost, and improvements are needed in order to make algae as feedstock for bio-based production a sustainable and commercially viable reality (e.g. Sander and Murthy 2010). Also, algae feedstock may still show higher impacts in most environmental impact categories, including fossil energy use and GHG emissions, compared to conventional crops (Clarens et al. 2010). As it is a rather new technology, it is hard to predict the actual potential


2010). As it is a rather new technology, it is hard to predict the actual potential. • Reference Fossil Product Feedstock Conventional petrochemical reference products are typically oil-based. However, as conventional crude oil resources are moving towards depletion, other petrochemical options will take over in the future. Shale gas and oil sands have received increased attention as oil prices have made these options feasible to exploit. These feedstocks, however, seem to have much larger impacts on the environment than conventional oil. Production of synthetic crude oil from oil sands is generally reported to have substantially higher GHG emission impacts per barrel compared to production of conventional crude oil, however large variations in results exist (e.g. National Energy Technology Laboratory 2008; Charpentier et al. 2009), and the life cycle water consumption for shale gas and oil sands is LCA of Biofuels and Biomaterials


approximately double that of conventional oil (King and Webber 2008), while also affecting quality and availability of land (Jordaan 2012). Thus, if these are the fossil resources that will become the future conventional feedstock against which to compare the bio-products, the latter will become more competitive in terms of environmental performance. Acknowledgements The authors gratefully acknowledge the helpful inputs from Arnaud Hélias (Montpellier SupAgro/INRA) and Anthony Benoist (CIRAD). Anderson-Teixeira, K.J., Davis, S.C., Masters, M.D., Delucia, E.H.: Changes in soil organic carbon under biofuel crops. GCB Bioenergy 1, 75–96 (2009) Archer, D., Kheshgi, H., Maier-Reimer, E.: Multiple timescales for neutralization of fossil fuel CO2. Geophys. Res. Lett. 24(4), 405–408 (1997) Bala, G., Caldeira, K., Wickett, M., Phillips, T.J., Lobell, D.B., Delire, C., Mirin, A.: Combined climate and carbon cycle-cycle effects of large-scale deforestation


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Author Biographies Susanne Vedel Hjuler has been working with LCA since the early 2010s. Particularly experienced within LCA of biomaterials, climate change related LCIA method development, as well as experienced LCA practitioner. Participated in the UNEP/SETAC task force on global warming LCIA. Main LCA interests are climate change, biomass use, as well as use of LCA in practice. Sune Balle Hansen LCI expert working intermittently on applied LCA of biofuels and edible oils since 2004 with focus on palm oil and palm oil derived biodiesel. Special LCA research attention has been given to the inclusion of residue use and land use change aspects. S.V. Hjuler and S.B. Hansen


Chapter 31 LCA of Chemicals and Chemical Products Peter Fantke and Alexi Ernstoff Abstract This chapter focuses on the application of Life Cycle Assessment (LCA) to evaluate the environmental performance of chemicals as well as of products and processes where chemicals play a key role. The life cycle stages of chemical products, such as pharmaceuticals drugs or plant protection products, are discussed and differentiated into extraction of abiotic and biotic raw materials, chemical synthesis and processing, material processing, product manufacturing, professional or consumer product use, and finally end-of-life. LCA is discussed in relation to other chemicals management frameworks and concepts including risk assessment, green and sustainable chemistry, and chemical alternatives assessment. A large number of LCA studies focus on contrasting different feedstocks or chemical synthesis processes, thereby often conducting a cradle to (factory) gate assessment


A large number of LCA studies focus on contrasting different feedstocks or chemical synthesis processes, thereby often conducting a cradle to (factory) gate assessment. While typically a large share of potential environmental impacts occurs during the early product life cycle stages, potential impacts related to chemicals that are found as ingredients or residues directly in products can be dominated by the product use stage. Finally, methodological challenges in LCA studies in relation to chemicals are discussed including the choice of functional unit, defining the system boundaries, quantifying emissions for many thousands of marketed chemicals, characterising emissions in terms of toxicity and other impacts, and finally interpreting LCA results. The chapter is relevant for LCA students and practitioners who wish to gain basic understanding of LCA studies of products or processes with chemicals as a key aspect. P. Fantke (&)  A


The chapter is relevant for LCA students and practitioners who wish to gain basic understanding of LCA studies of products or processes with chemicals as a key aspect. P. Fantke (&)  A. Ernstoff Division for Quantitative Sustainability Assessment, Department of Management Engineering, Technical University of Denmark, 2800 Kongens Lyngby, Denmark e-mail: pefan@dtu.dk A. Ernstoff EPFL Innovation Park, Quantis International, Bât. D, 1015 Lausanne, Switzerland © Springer International Publishing AG 2018 M.Z. Hauschild et al. (eds.), Life Cycle Assessment, DOI 10.1007/978-3-319-56475-3_31


31.1 LCA and Chemicals: Introduction and Context 31.1.1 Chemicals and Their Relevance in Society Chemicals are everywhere. Almost every second a new entry is added to the list of more than 100 million unique chemical substances registered in the Chemical Abstracts Service (CAS; www.cas.org), the world’s authority on chemical information. Since industrialisation, the welfare of modern society largely builds on extensively mining minerals and fossil fuels including coal, petroleum and natural gas to produce large quantities of synthetic chemicals (‘synthetic’ simply means man-made and should not be confused with ‘artificial’, which implies that a chemical does not occur naturally). Consequently, the enormity and diversity of the chemical industry is astounding and poses various challenges for the management of environmental and human health impacts related to chemicals production and use. In this chapter, we outline important aspects to know about chemicals in the context of LCA


In this chapter, we outline important aspects to know about chemicals in the context of LCA. Fundamentally, chemicals are substances composed of one or more atoms, and make up every material thing on earth—including our bodies. The atomic composition of chemicals classifies them essentially as ‘organic’ (chemicals with molecules built on a skeleton of interlinked carbon atoms and primarily consisting of carbon, oxygen, and hydrogen) and ‘inorganic’ (chemicals with molecules generally lacking carbon-to-carbon bonds, but instead based on the rest of the elements, including metals). In this sense, ‘organic’ has nothing to do with ‘organic food’ or ‘organic farming’ or ‘organic lifestyle’ as these terms generally refer to promoting sustainability. The atomic composition, molecular structure and ionisation (positive/negative charge) all influence chemical reactivity and behaviour in the environment as well as in living organisms


The atomic composition, molecular structure and ionisation (positive/negative charge) all influence chemical reactivity and behaviour in the environment as well as in living organisms. Because of this, chemical behaviours can be predicted and tested, and chemicals can be designed by industries to fulfil biological (e.g. medical) and physical (e.g. solvent) functions. Chemicals may also be classified according to functional groups (e.g. alcohols, amines, acids and bases), structural groups (e.g. polycyclic aromatic hydrocarbons), physical structure (e.g. nanotubes), feedstock sources (e.g. petrochemicals derived from fossil fuels, biochemicals derived from starch- and sugar-based feedstocks), physicochemical properties (e.g. volatile, lipophilic), use function (e.g. surfactants, warfare agents), means of creation (e.g. reaction intermediates, metabolites), main economic sector (e.g. cosmetics, agrochemicals), toxicity endpoints (e.g


surfactants, warfare agents), means of creation (e.g. reaction intermediates, metabolites), main economic sector (e.g. cosmetics, agrochemicals), toxicity endpoints (e.g. carcinogens, neurotoxins, endocrine disruptors), and other aspects. Established nomenclatures or patents can be used to name chemicals. Most chemicals have an assigned CAS Registry Number except some metabolites of natural processes or grouped chemicals, such as polychlorinated dibenzofurans. CAS numbers are the most discriminant method for chemical reference. Of the chemicals registered by CAS, more than ten thousand are currently in commercial use, some with annual production volumes of millions of tonnes, while most chemicals are produced at levels of less than thousand tonnes per year. Worldwide, the production of chemicals has risen to several hundred million tonnes per year and P. Fantke and A. Ernstoff


sales were valued in 2013 at 3156 billion Euro with an average annual growth of 10.3% between 2003 and 2012 (CEFIC 2014). China dominates world chemical sales with a share of 33.2% followed by the European Union (16.7%), USA (14.8%), and Japan (4.8%) in 2013. Databases, such as the European Chemicals Agency (ECHA) Registered Substances database (http://echa.europa.eu/ information-on-chemicals), the Household Product Database (http:// householdproducts.nlm.nih.gov), the Hazardous Substances Data Bank through ToxNet (http://toxnet.nlm.nih.gov), and the Chemical and Product Categories Database (http://actor.epa.gov/cpcat) attempt to keep track of chemicals, their uses, properties and/or toxicity, but large data gaps still remain. Several major environmental and health concerns associated with chemicals have led to various shifts in the global chemicals market


Several major environmental and health concerns associated with chemicals have led to various shifts in the global chemicals market. As an example, potentially toxic and highly persistent polychlorinated biphenyls (PCBs) have been replaced by chlorinated paraffins in various applications. While PCBs have been primarily produced in USA and Europe with a total historical production volume of 1.3 million tonnes between 1930 and 1995, chlorinated paraffins are currently almost exclusively produced in China and reach production volumes of more than one million tonnes per year (Fantke et al. 2015). The chemical and pharmaceutical industries are a major driver of the welfare of modern society and scientific progress. These industries rely on the extraction, purification and synthesis of both naturally occurring and artificial chemicals and are among the largest and most influential economic sectors at the global scale. Main production segments are petrochemicals (e.g


Main production segments are petrochemicals (e.g. benzene, styrene), consumer chemicals (e.g. detergents, fragrances and flavours), speciality chemicals (chemicals used for providing a special performance or effect, e.g. paints, dyes, adhesives), basic inorganics (fertilisers, industrial gases like nitrogen and oxygen), and polymers (e.g. plastics, synthetic rubbers and fibres). One of the largest segments is the production of organic chemicals with, e.g. formaldehyde, aromatics, acids, alcohols and esters providing the building blocks for drugs, agrochemicals, cosmetics and many other applications. Along with societal advantages, the rise of chemical industries has also caused various undesirable consequences. Health impacts associated with air pollution are increasing worldwide and there is currently insufficient information to fully assess the impacts of chemicals on humans and the environment


Health impacts associated with air pollution are increasing worldwide and there is currently insufficient information to fully assess the impacts of chemicals on humans and the environment. Rachel Carson’s book Silent Spring published in 1962 documented the detrimental impacts of chemicals on wildlife and humans, especially related to using synthetic organic pesticides, and marked a major change in public awareness that eventually inspired regulation of industry and for example the creation of the United States Environmental Protection Agency. Since that time, a remarkable amount of research correlates and demonstrates impacts on human and ecosystem health as well as the environment (e.g. the ozone layer) caused by intentional and unintentional chemical releases both indoors and outdoors. Some reported impacts are directly related to the chemical industry, whereas other impacts are related to the use or disposal of chemicals by other industries


Some reported impacts are directly related to the chemical industry, whereas other impacts are related to the use or disposal of chemicals by other industries. In the following sections, we overview strategies for chemical management, focusing particularly on life-cycle assessments of chemicals production processes and chemical products. LCA of Chemicals and Chemical Products


31.1.2 Chemicals Management in Relation to LCA Depletion of the ozone layer by chlorofluorocarbons used as refrigerants and solvents, soil and water pollution with heavy metals from ore mining and processing, pesticide emissions and residues in food, the formation of dioxins by incomplete combustion processes, and leaching of fertilisers into groundwater are just examples of the many problems associated with chemical releases to the environment. Hence, managing human and environmental risks posed by chemicals that are potentially toxic or may lead to other impacts is a major concern of regulators, industries, consumers and other stakeholders. As a consequence, the chemicals industry is one of the most regulated industries with main focus on regulating chemicals in consumer products and minimising chemical emissions to the indoor (workplace, public buildings and household) and outdoor environments along product life cycles


In the context of chemicals management, risk is defined as the probability of a chemical to cause an adverse effect (hazard) occurring as a result of a given contact between the chemical and humans or ecosystems (exposure). Risks associated with chemical emissions from a given product or process can arise at a later time after a chemical emission or exposure has occurred and depend on chemical background concentrations due to all release sources. In LCA, information on emission location and time as well as information on background concentrations, e.g. from sources outside the considered product system, is usually not available. Hence, modelled impacts in LCA are not interpreted in terms of actual risk, i.e. real environmental effects, but in terms of ‘potential impacts’ (Chap. 10) used as environmental performance indicators for comparing and optimising products or systems with respect to a defined functional unit (Hauschild 2005)


10) used as environmental performance indicators for comparing and optimising products or systems with respect to a defined functional unit (Hauschild 2005). However, models applied in LCA can also be applied in other fields of research and can be advanced and adapted to consider background concentrations as well as spatiotemporal resolution (e.g. daily or seasonal changes), and in such cases estimated potential impacts can be interpreted as estimates of actual risk. Chemicals management occurs from local to global scale, from specific product– chemical combinations to entire industries and from raw material acquisition to waste handling, depending on the intended purpose


The Montreal Protocol on Substances that Deplete the Ozone Layer (http://ozone.unep.org) and the Stockholm Convention on Persistent Organic Pollutants (POPs; www.pops.int) are examples of global chemicals management treaties, whereas the Registration, Evaluation, Authorisation, and Restriction of Chemicals (REACH) is a recent example of an international legislative framework for managing industrial chemicals in the European Union. At all levels and scopes, effective chemicals management relies on assessment tools and guiding principles to ensure consistency and the achievability of defined goals. There are many examples of chemicals assessment tools and guidance, such as risk assessment, green and sustainable chemistry, chemical alternatives assessment, life cycle assessment, and a market for entrepreneurs to create industry-specific interfaces and applications. In the following sections, risk assessment (Sect. 31.1.3), green and sustainable chemistry (Sect. 31.1.4), and P


In the following sections, risk assessment (Sect. 31.1.3), green and sustainable chemistry (Sect. 31.1.4), and P. Fantke and A. Ernstoff


chemical alternatives assessment (Sect. 31.1.5) are discussed as commonly used chemical management tools that have both complementary and overlapping aspects with LCA as illustrated in Fig. 31.1. 31.1.3 Risk Assessment and Safety Chemical risk assessment or chemical safety assessment is implemented in various regulatory frameworks and is one of the most widely used chemicals management tools. Risk assessment (‘How risky is a situation?’) as an integral part of risk management (‘What shall we do about it, if a situation is risky?’) essentially emerged at the start of the nineteenth century from studying hazards and risks associated with different occupations. Risk assessment mainly consists of hazard identification, dose-response assessment, exposure assessment and risk characterisation. Depending on the context, ‘risk’ and ‘safety’ have different meanings with regulatory policy commonly seeking to minimise risk while optimising safety


Depending on the context, ‘risk’ and ‘safety’ have different meanings with regulatory policy commonly seeking to minimise risk while optimising safety. In this context, risk is generally defined as the probability of harm, whereas safety is described as the absence of harm (Embry et al. 2014). Chemical ‘safety’ is defined by legislators or regulators and can vary from country to country, and evolves over time as science progresses. In this sense, ‘safe’ Chemical Risk Assessment Chemical AlternaƟves Assessment Green and Sustainable Chemistry Environmental Life Cycle Assessment of chemicals, products and services Sustainability assessment tools Regulatory chemicals management tools Chemical-related criteria (hazard, exposure, toxicity, ...) Other environmental criteria (energy and resources efficiency, ...) Other criteria (technical feasibility, socioeconomics,...) Product and material design Fig


31.1 Conceptual relationships of main chemical management tools LCA of Chemicals and Chemical Products


is not synonymous with ‘natural’ as it is often perceived. In fact, using the word ‘natural’ is misleading in the context of chemical safety (and LCA) and there are many naturally occurring chemicals that have very harmful properties like arsenic, nicotine or radon. As a consequence, we need to acknowledge that it is not always the ‘natural’ chemicals or solutions that are most ‘environmentally friendly’—a common misconception in different science-policy fields and among consumers. Defining safety thresholds, e.g. chemical concentrations in different environmental media (e.g. ambient air, soil, water) or in food, is a common strategy in chemical risk assessment, and generally refers to levels below which a situation is considered ‘safe’ by a risk manager, meaning that any risk below threshold is regarded as ‘acceptable’


As an example, chemical exposure resulting in one additional cancer case or less over lifetime in a population of one million people is regarded as an acceptable risk, i.e. safe, in the U.S. (van Leeuwen and Vermeire 2007). Using units like ‘part per million’ (ppm) as in one cancer case in a million or ‘part per billion’ is common for assessing (very small) amounts of chemicals in the environment. To get an impression of how much one ppm actually is, we can use 1 teaspoon of salt (5.5 g) in 5.5 tonnes of potato chips corresponding to one part of salt per one million parts of potato chips


To get an impression of how much one ppm actually is, we can use 1 teaspoon of salt (5.5 g) in 5.5 tonnes of potato chips corresponding to one part of salt per one million parts of potato chips. Thresholds are also applied when managing environmental systems and for developing chemical pollution control strategies, such as allowable nutrient releases from wastewater treatment plants or setting greenhouse gas emission targets, or in the context of ‘planetary boundaries’ in an attempt to assess if the pressure from chemical pollution (analogous to the amount of receiving environment required to dilute pollution below a threshold level) exceeds a planetary boundary (analogous to the amount of receiving environment available) for a ‘safe operating space’ for human activities (MacLeod et al. 2014). Chemical pollution levels have recently been expressed as ‘chemical footprints’ that can be compared with respective planetary or other boundaries for chemical pollution (Posthuma et al


2014). Chemical pollution levels have recently been expressed as ‘chemical footprints’ that can be compared with respective planetary or other boundaries for chemical pollution (Posthuma et al. 2014) to assess how companies or nations perform with respect to different chemicals management issues. Risk assessment approaches take a receptor perspective (Fig. 31.2, right-side box), where thresholds are set in order to protect specific receptors, i.e. exposed humans or ecosystem species. In a receptor perspective, all relevant sources of a chemical or target chemicals are typically considered. In contrast, impact assessment tools in LCA are generally not receptor-oriented or threshold-based. This is because LCA takes a ‘producer’ (or ‘emitter’) perspective (Fig


In contrast, impact assessment tools in LCA are generally not receptor-oriented or threshold-based. This is because LCA takes a ‘producer’ (or ‘emitter’) perspective (Fig. 31.2, left-side box) by comparing potential impacts relative to each other across compared products and life cycle stages, aiming at minimising impacts considering various receptors (entire human populations, freshwater ecosystems, marine ecosystems, etc.). Differences and commonalities of risk assessment and LCA have been contrasted elsewhere (e.g. Bare 2006; Pennington et al. 2006), and there are several attempts combining, blending or integrating both concepts (Harder et al. 2015). An increasing number of chemicals is approved for use in commerce, e.g. in food contact materials, but many chemicals lack adequate information to characterise risks (Neltner et al. 2013). In response, high-throughput screening (‘first tier’ P. Fantke and A. Ernstoff


assessments) of chemical risks has emerged as a strategy for prioritising and ranking chemicals for more in-depth study (‘higher-tier’ assessments). First-tier screening usually relies on ranking chemicals with respect to hazard (e.g. chemical toxicity) combined with estimates of exposure. ‘High-throughput’ refers to processing dozens to thousands of chemicals via resource efficient methodologies, such as robotic in vitro bioassays (instead of animal in vivo experiments) and low-tier computational models relying on databases (instead of data-intensive complex and time-consuming modelling). LCA impact assessment models have been used in high-throughput risk screening offering dual purpose and a promising area of interdisciplinary overlap to manage chemical risks (e.g. Shin et al. 2015). 31.1.4 Green and Sustainable Chemistry ‘Green chemistry’ is a concept that was coined by the U.S


Shin et al. 2015). 31.1.4 Green and Sustainable Chemistry ‘Green chemistry’ is a concept that was coined by the U.S. Environmental Protection Agency in the early 1990s in response to the Pollution Prevention Act and increasing attention to chemical pollution. This concept builds upon a set of 12 Principles of Green Chemistry defined by Anastas and Warner (1998) aiming at reducing or eliminating hazardous substances in the design, manufacture and application of chemical products. Thereby, ‘green’ refers to more environmentally benign (less hazardous) chemicals. The concept of ‘sustainable chemistry’ is broader than the scope of green chemistry and strives towards ‘eco-efficiency’. In addition to chemical hazards, sustainable chemistry centrally focuses on optimising the use of finite resources, while reducing environmental impacts of chemical production (OECD 2012)


Sustainable chemistry—sometimes also referred to as R t Ɵ P d / iƩ Ɵ Receptor perspecƟve [Chemical-specific hazards and risks] Single chemical exposes individual Producer/emiƩer perspecƟve [Product- or service-specific impacts] Single or mulƟple chemicals expose Single chemical exposes individual receptor via mulƟple sources Exposure and related risks esƟmated with respect to Single or mulƟple chemicals expose mulƟple receptors over life cycle of single product or service Exposure and related impacts esƟmated with respect to environmental chemical concentraƟons CompariƟve ranking of risks based Exposure and related impacts esƟmated per unit of chemical emiƩed or used ComparaƟve ranking of impacts CompariƟve ranking of risks based on exposure and toxicity potenƟal Risk assessment; high throughput risk screening ComparaƟve ranking of impacts based on funcƟonal service or use Life cycle assessment; chemical alternaƟves assessment high-throughput risk screening Single chemical


high throughput risk screening ComparaƟve ranking of impacts based on funcƟonal service or use Life cycle assessment; chemical alternaƟves assessment high-throughput risk screening Single chemical alternaƟves assessment Single product MulƟple sources Exposed i di id l MulƟple Exposed Single chemical product over life cycle sources, e.g


products individual receptor MulƟple chemicals p mulƟple receptors Fig. 31.2 Examples and underlying characteristics of dichotomous perspectives followed in different chemicals management approaches LCA of Chemicals and Chemical Products


sustainable chemistry and engineering—is rooted in the concept of Sustainable Development established in Rio de Janeiro in 1992 at the United Nations Conference on Environment and Development and is guided by 9 Principles of Green Engineering postulated at the Sandestin conference (Abraham and Nguyen 2003). Green and sustainable chemistry are concepts focusing on the technological approaches aiming at the reduction of resource consumption and pollution prevention in chemical production processes rather than focussing on the assessment of chemicals in the environment. Hence, green and sustainable chemistry—often relying on comparing qualitative or semi-quantitative indicator results—are primarily applicable in the design phase of products to guide innovation and to support sustainable production goals. Green chemistry in relation to LCA has been discussed in more detail elsewhere (e.g. Anastas and Lankey 2000)


Green chemistry in relation to LCA has been discussed in more detail elsewhere (e.g. Anastas and Lankey 2000). In summary, compared to green and sustainable chemistry, LCA aims at fully quantifying potential impacts associated with a chemical product or production system over its entire life cycle. Using LCA in early stages of chemical product and process design of various sectors including emerging technologies (e.g. bio- and nanotechnologies) has provided insight into the relationship between chemical and process parameter selection and related impacts on humans and the environment (Kralisch et al. 2015). LCA results have moreover demonstrated that quantitative methods are needed to assess the environmental performance of ‘green’ chemicals (Tufvesson et al. 2013)


2015). LCA results have moreover demonstrated that quantitative methods are needed to assess the environmental performance of ‘green’ chemicals (Tufvesson et al. 2013). This is especially relevant as green chemistry usually focuses on optimisation of (production) processes, including some specific end-of-life problems related to chemicals, which may still risk sub-optimisation when a full life cycle perspective is lacking. Using LCA in early product development stages, for example before a product has been created and marketed, comes with methodological and practical challenges, such as low data availability, uncertainty related to future product applications, and unclear scale of production for a changing market. Therefore, LCA has mostly been applied to chemical products and processes that are already well established and operational at the market scale, which often leads to LCA supported decision making being reactive instead of proactive


31.1.5 Chemical Alternatives Assessment Chemical alternatives assessment (CAA) aims to identify, compare and select safer alternatives to substitute (replace) harmful chemicals in materials, processes and products on the basis of their hazards, performance and economic viability (Hester and Harrison 2013). CAA emerged from the U.S. Environmental Protection Agency’s Design for Environment (DfE) program in the late 1990s to promote less hazardous chemicals in various products and applications, and to avoid unintended consequences of harmful alternatives resulting in incremental improvements or even ‘regrettable substitution’ situations (Fantke et al. 2015). Ideally, CAA tools P. Fantke and A. Ernstoff


would evaluate hazard, exposure, life cycle and social impacts, economic feasibility and technical performance of alternative solutions, and consider chemicals, materials, products or technologies, and behavioural changes as potentially viable solution options. In reality, however, most CAA tools focus only on comparisons of hazard scores and exclusively consider chemicals as potential solutions. Several existing CAA tools have been compiled into the OECD Substitution and Alternatives Assessment Toolbox (www.oecdsaatoolbox.org). The concept of ‘acceptable risk’ (as applied in risk assessment) is usually avoided in CAA in order to support selecting relatively less hazardous chemicals and materials in products (Whittaker 2015). Despite the current focus on assessing chemical hazard, including exposure, life cycle, and social considerations are lately also gaining more attention (Jacobs et al


Despite the current focus on assessing chemical hazard, including exposure, life cycle, and social considerations are lately also gaining more attention (Jacobs et al. 2016), focusing the CAA discussion around using more quantitative and chemical function-based methods and tools (Tickner et al. 2015). However, the need for rapid screening of numerous viable alternative solutions prevents CAA from simply adopting the use of LCA tools due to high complexity and data demand. CAA is mainly used to identify and evaluate solutions to hazardous chemicals in products that have been targeted for market phase-out, and to inform early product development to minimise reliance on hazardous chemicals. With that, CAA takes the ‘producer’ perspective similarly to LCA (Fig. 31.2, left-side box), focusing on the impact of chemicals and their alternatives on various receptors


With that, CAA takes the ‘producer’ perspective similarly to LCA (Fig. 31.2, left-side box), focusing on the impact of chemicals and their alternatives on various receptors. The main difference between CAA and LCA is that while CAA focuses on seeking for viable alternatives to harmful chemicals, LCA considers the life cycle of whole products or processes not focusing specifically on the content of one or more chemicals that might be considered ‘hazardous’, but instead evaluating the overall product or process environmental performance. 31.2 LCA Applied to Chemicals Chemicals play a central role in the LCA framework for different reasons. Hundreds of chemical emission (inventory) flows typically occur along the life cycle of products or systems (Fig. 31.3) and are quantified as part of the Life Cycle Inventory (LCI; see Chap. 9) phase. Chemicals are also often precursors of product materials, and input for manufacturing and disposal processes


31.3) and are quantified as part of the Life Cycle Inventory (LCI; see Chap. 9) phase. Chemicals are also often precursors of product materials, and input for manufacturing and disposal processes. Chemical emissions associated with energy conversion during manufacturing, transport of goods and end-of-life treatment processes often dominate overall emission profiles for many product categories that can be characterised in the Life Cycle Impact Assessment (LCIA; see Chap. 10) phase. Chemicals contribute to nearly all LCIA impact categories affecting human health and ecosystem quality as two main areas of protection in LCA (Hauschild et al. 2013). In LCIA, chemicals contribute to global warming, stratospheric ozone depletion, formation of photochemical ozone in the troposphere, air pollution (via respiratory particles and precursors), aquatic and terrestrial acidification and eutrophication, and last but not least human toxicity and aquatic and terrestrial ecotoxicity


Only a handful LCA of Chemicals and Chemical Products


of chemicals are associated with the majority of abovementioned impact categories, such as carbon dioxide, methane and other greenhouse gases contributing to global warming impacts or ammonia, nitrogen oxides, phosphate and some other nitrogen and phosphorus containing chemicals contributing to aquatic eutrophication. In contrast, thousands of chemicals can be characterised as potentially toxic to humans and/or ecosystems (Rosenbaum et al. 2008). The generic life cycle stages shown in Fig. 31.3 are applicable to a chemical product (e.g. pharmaceutical or dye) or material (e.g. polymer), from raw materials extraction to product disposal, often referred to as ‘cradle to grave’ (Fig. 31.3, stages A–F). A ‘cradle to grave’ LCA study can provide valuable insight regarding which stages dominate the impacts throughout a product life cycle. Some of these life cycle stages, however, may not be relevant or may be assumed to be equal in two compared systems depending on the goal (Chap


Some of these life cycle stages, however, may not be relevant or may be assumed to be equal in two compared systems depending on the goal (Chap. 7) and scope (Chap. 8), and the product system under study. For example, the ‘material processing’ stage may Chemical synthesis/ processing Product manufacturing Raw materials extracƟon Product applicaƟon/use Product disposal System boundaries Input: energy, resources (e.g. water), materials, chemical substances A B D F Material processing C E Output: chemical emissions, secondary funcƟons, co-products Relevant quesƟons for different life cycle stages... A: Fossil or bio-based resources; virgin or recycled materials; alternaƟve use of resource or material (e.g. food) or waste (e.g. shell or husk); locaƟon of extracƟon (e.g. rain forest, ocean, conflict zone); secondary funcƟons (e.g. algae waste treatment)? B: Reactor configuraƟon (e.g. batch or conƟnuous); number of reacƟon steps or unit operaƟons; solvents; catalysts (e.g


algae waste treatment)? B: Reactor configuraƟon (e.g. batch or conƟnuous); number of reacƟon steps or unit operaƟons; solvents; catalysts (e.g. enzymes, rare metals)? C: Machine lubricants; chemical contents and concentraƟons? D: Machine lubricants; solvents; adhesives? E: Product longevity; secondary funcƟons; exposure during use; nature of use (e.g. amount used; addiƟonal requirements such as hot water); societal funcƟon (e.g. to cure a disease)? F: Special handling (e.g. pharmaceuƟcals); disposal locaƟon; process (e.g. incineraƟon or landfil); reuse opportunity; circularity of waste streams? A-D and F: Waste handling (e.g. municipal or industrial; type); co-products (e.g. materials; energy from incineraƟon); exisƟng or new infrastructure; efficiency; worker exploitaƟon or occupaƟonal exposure? Fig. 31.3 Generic life cycle stages and system boundaries for chemical products or materials and LCA-related questions


31.3 Generic life cycle stages and system boundaries for chemical products or materials and LCA-related questions. In some cases, chemical processing may be followed by material production (e.g. polymers) before manufacturing a product (e.g. plastic bottles), while in other cases, chemicals (e.g. solvents) may be directly added to products or product manufacturing processes. Underlined topics are mostly lacking methods or not included in environmental LCA studies P. Fantke and A. Ernstoff


not be relevant in cases where a chemical is directly added into a product as an ingredient, such as fragrances in cosmetics or detergents in cleaning products. As another example, the ‘product application/use’ or ‘product disposal’ stages may not be relevant for comparing the environmental performance of chemical synthesis or production processes as long as the compared processes do not influence the chemical amount used in a product or for product disposal. An LCA study from raw material extraction to chemical product manufacturing, i.e. without considering product use and disposal stages, is referred to as ‘cradle to gate’ (Fig. 31.3, stages A–D), where ‘gate’ refers to the manufacturing or production facility (which could be the ‘gate’ of a chemical or product ‘factory’, depending on the focus of the study)


31.3, stages A–D), where ‘gate’ refers to the manufacturing or production facility (which could be the ‘gate’ of a chemical or product ‘factory’, depending on the focus of the study). In Table 31.1, different assessment scopes for LCA studies focusing on chemicals in materials, products and processes are contrasted and associated with relevant chemicals management questions. LCA can help identify a variety of impacts associated with chemical production, use, and disposal, that are either intrinsic to a chemical (e.g. toxicity potential) or related to supporting industrial chemical processes (e.g. water consumption, greenhouse gas emissions). The main uses of LCA for managing chemicals and chemical processes are to compare impacts between products or services, or to identify ‘hot spots’ within a life cycle that contributes greatly to the impacts of a product or service. With respect to chemicals, LCA can be applied to various combinations of the generic life cycle stages in Fig


With respect to chemicals, LCA can be applied to various combinations of the generic life cycle stages in Fig. 31.3 depending on the LCA study goal and chosen system boundaries. In some cases, individual life cycle stages and associated inputs or outputs may be found to be irrelevant for the system considered or question asked. The chemical industry developed a guidance document to support the assessment of the environmental performance of chemical products based on attributional LCA, i.e. referring to process-based modelling and excluding market-mediated effects (WBCSD 2014). In the following sections, an overview is given of how LCA has been applied to consider these various life cycle stages and the general lessons learnt from these studies. Thereby, LCA can be used to compare impacts at the level of chemicals in materials, products and formulations or at the level of chemical synthesis and production processes


Thereby, LCA can be used to compare impacts at the level of chemicals in materials, products and formulations or at the level of chemical synthesis and production processes. 31.2.1 Chemicals in Materials, Products, and Formulations A subset of materials, products, formulations (combination or mixture of chemicals) and processes are intrinsically reliant on the functionality of key chemical ingredients. In this section, main trends in the application of LCA- or LCA-based methodologies are summarised. This may include also partial LCA studies, e.g. methods only considering a subset of life cycle stages (i.e. cradle to gate or gate to gate), with focus on chemicals in materials, products and formulations. LCA of Chemicals and Chemical Products


LCA studies have focused on pharmaceuticals (e.g. De Soete et al. 2014), cleaning products (e.g. Van Lieshout et al. 2015) and pesticide formulations (e.g. Geisler et al. 2005) as examples of products where chemicals provide the main product functions. Other LCA studies on chemicals with in-product functions include studies focusing on flame retardants in electronics (Jonkers et al. 2016), nano-materials used in bandages and cosmetics (Botta et al. 2011), and polymers used in food packaging (Hottle et al. 2013). Chemicals required for industrial processes have also been assessed in LCA studies, including industrial solvents (Zhang et al


2011), and polymers used in food packaging (Hottle et al. 2013). Chemicals required for industrial processes have also been assessed in LCA studies, including industrial solvents (Zhang et al. 2008) and chemicals used for the production of treated water, oil and Table 31.1 Relevant life cycle assessment scopes and life cycle stages for selected chemicals management questions and example studies Chemicals management questions Assessment scopes and considered life cycle stages Example studies What is the environmental performance of different products with respect to chemical emissions? • Cradle to grave • Stages A–F (Fig. 31.3) • Focus on chemicals consumption and emissions • Cleaning products (Van Lieshout et al. 2015) • Textiles (Roos et al. 2015) What are the environmental profiles of the production of different chemicals? • Cradle to (factory or consumer) gate • Stages A–D or a subset of these stages (Fig. 31.3) • Focus on chemical manufacturing • Pharmaceuticals (Wernet et al


31.3) • Focus on chemical manufacturing • Pharmaceuticals (Wernet et al. 2010) Which stage of a chemical product life cycle contributes most to environmental impacts? • Cradle to grave • Hotspot analysis including stages A–F (Fig. 31.3) • Focus on chemicals as products • Plant protection products (Geisler et al. 2005) Which chemical synthesis and/or manufacturing processes contribute most to environmental impacts? • Cradle to (factory) gate • Hot-spot analysis including stages A–B or A–C (Fig. 31.3) • Focus on chemical manufacturing • Pharmaceuticals (De Soete et al. 2014) • Nano-materials (Pati et al. 2014) Which life cycle stage of a chemical in a product contributes most to human exposure? • Cradle to grave • Partial LCA (only human e.g. exposure estimates) including stages A–F (Fig. 31.3) • Focus on chemicals in products • Cosmetics (Ernstoff et al


exposure estimates) including stages A–F (Fig. 31.3) • Focus on chemicals in products • Cosmetics (Ernstoff et al. 2016a) Which feedstock provides the most environmentally friendly substrate for biochemical synthesis? • Cradle to (factory) gate • Stages A or A–B (Fig. 31.3) • Focus on chemicals and raw materials consumption • Acrolein (Cespi et al. 2015) • PET (Akanuma et al. 2014) P. Fantke and A. Ernstoff


gas, printing paper and dyed textiles (e.g. Alvarez-Gaitan et al. 2013; Parisi et al. 2015). When analysing LCA studies on chemical-based functions, a few generalisations emerge. For example, it is important to consider life cycle thinking early on in the design phase of products and processes whenever possible and it has been shown that simplified tools may help in this process (e.g. De Soete et al. 2014). Furthermore, it has been demonstrated that hybridised LCA tools or metrics can be useful to improve communication and management for specific stakeholders (e.g. Alvarez-Gaitan et al. 2013). Several LCA studies indicate that being sceptical of services deemed ‘green’ or ‘sustainable’ is crucial, especially when an LCA has not yet been performed. Case studies on, e.g. ‘green’ solvents (Zhang et al. 2008) or ‘sustainable’ bio-based chemicals and materials (e.g. Hottle et al


Case studies on, e.g. ‘green’ solvents (Zhang et al. 2008) or ‘sustainable’ bio-based chemicals and materials (e.g. Hottle et al. 2013) demonstrate that materials and products guided by principles of ‘sustainability’, ‘eco-friendliness’ or ‘green chemistry’ can have significant, but often disregarded or unassessed, environmental impacts. An example is given in Fig. 31.4, where environmental life cycle impacts of petro- and bio-based polymers are contrasted based on data from Hottle et al. (2013). 0% 20% 40% 60% 80% 100% PLA TPS HDPE LDPE PET PP PS PLA TPS HDPE LDPE PET PP PS PLA TPS HDPE LDPE PET PP PS PLA TPS HDPE LDPE PET PP PS Global warming Ozone depleƟon EutrophicaƟon AcidificaƟon Normalized impacts 60% 80% 100% zed impacts 0% 20% 40% 60% PLA TPS HDPE LDPE PET PP PS PLA TPS HDPE LDPE PET PP PS PLA TPS HDPE LDPE PET PP PS PLA TPS HDPE LDPE PET PP PS Ecotoxicity Human toxicity cancer Human toxicity noncancer Respiratory inorganics Normalized imp Fig


31.4 Impact scores for LCAs of two bio-based polymers (dark bars; PLA Polylactic acid, TPS Thermoplastic starch) compared to petroleum-based polymers (light bars; HDPE High-density polyethylene, LDPE Low-density polyethylene, PET Polyethylene terephthalate, PP Polypropylene, PS Polystyrene) per kg of produced granule, normalised for each category to the polymer with highest impacts (based on data from Hottle et al. 2013) LCA of Chemicals and Chemical Products


According to this study, bio-based polymers lead to higher impacts than petro-based polymers for several impact categories, which contradicts assumptions that bio-based automatically implies ‘green’ or ‘sustainable’ (see also Chap. 30). Higher impacts for bio-based polymers are mainly associated with feedstock-related agricultural emissions of fertilisers (eutrophication) and pesticides (human toxicity and ecotoxicity), as well as deforestation (impacts related to changes in land use). However, the relative importance (i.e. contribution to overall environmental impacts) of the different impact categories also needs to be considered when evaluating the overall environmental performance of different polymers or other chemical products and processes. Often products are referred to as ‘green’ or ‘sustainable’ based on a single environmental issue (e.g. reducing greenhouse gas emissions), or based on following the principles of green chemistry in chemical design only


reducing greenhouse gas emissions), or based on following the principles of green chemistry in chemical design only. However, chemical products that are claimed to be ‘green’ or ‘sustainable’ may in fact lead to greater impacts on the environment or humans than the conventional alternatives. For example, ‘eco-friendly’ food packaging made of plant fibres may increase exposure and environmental emissions of highly hazardous fluorinated chemicals (Yuan et al. 2016), and ‘green’ solvents can have higher impacts across many impact categories when compared to conventional solvents (Zhang et al. 2008). Furthermore, the production of bio-based raw materials (such as corn, sugar cane, or soy for feedstock) may or may not be associated with lesser greenhouse gas emissions and consumptions of fossil resources, but may have equal or greater impacts in other impact categories (e.g


land use, toxicity related to using pesticides, eutrophication related to using fertilisers) than fossil-based materials (see Chap. 30 for further details). These phenomena are commonly referred to as burden shifting (e.g. between environmental issues or compartments). Identifying these is a fundamental application principle unique to LCA. LCA is a tool that can be useful for comparing products and processes for identifying such burden shifting and how to minimise impacts across a variety of impact categories. However, it is important always to ensure as a practitioner that all relevant chemical emissions are inventoried and all impact pathways are characterised. These general cautions are also relevant for LCA studies focusing on chemical synthesis and production processes as discussed in the following section. 31.2.2 Chemical Synthesis and Production Processes LCA is a useful tool for improving existing processes and designing new processes for the synthesis and production (Fig