Data for Europe for example show a decline of 5.4% of stratospheric O3 concentration per decade since the 1980s when measured in winter and spring, with an improving trend over the period 1995–2000. However, in later years low concentration records were broken on an almost annual basis. To date, the largest ‘ozone hole’ in human history was observed in 2006 with 29.5 million km2 over Antarctica, but even in 2015 its largest spread still reached 28.2 million km2. The largest Arctic ‘ozone hole’ ever was observed in 2011. Impacts of stratospheric ozone depletion are essentially linked to reduced absorption of solar radiation in the stratosphere leading to increased UV radiation intensities at the planet surface, of which three broad (wavelength) classes are distinguished: UV-C, UV-B and UV-A. The impact of UV radiation on living organisms depends on its wavelength, the shorter the more dangerous


The impact of UV radiation on living organisms depends on its wavelength, the shorter the more dangerous. UV-C is the most dangerous wavelength range, but almost completely filtered by the ozone layer. UV-B (wavelengths 280–315 nm) is of the most concern due to ozone layer depletion, while UV-A is not absorbed by ozone. Depending on duration and intensity of exposure to UV-B, impacts on human health are suspected to include skin cancer, cataracts, sun burn, increased skin cell ageing, immune system diseases, headaches, burning eyes and irritation to the respiratory passages. Ecosystem effects are linked to epidermal damage to animals (observed e.g. in whales), and radiation damage to the photosynthetic organs of plants causing reduced photosynthesis, leading to lower yields and crop quality in agricultural produce and loss of phytoplankton, the primary producers of aquatic food chains, particularly in the polar oceans

Hauschild et al., LCA: Theory and Practice


Additionally, UV-B accelerates the generation of photochemical smog, thereby stimulating the production of tropospheric ozone, which is a harmful pollutant (see Sect. 10.10). Fig. 10.10 Evolution of the hole in the ozone layer over Antarctica in September from 1980 to 2015 (Source NASA Ozone Watch 2016, public domain, http://ozonewatch.gsfc.nasa.gov/ monthly/climatology_09_SH.html) R.K. Rosenbaum et al.


10.7.2 Environmental Mechanism Stratospheric ozone concentrations result from a balance between O3 formation and destruction under the influence of solar (UV) radiation, temperature and the presence of other chemicals. The annual cycle of ozone destruction over the poles develops under the presence of several influencing factors with its intensity directly depending on their combined intensity: (1) meteorological factors (i.e. strong stratospheric winds and low temperature) and (2) the presence of ozone depleting chemicals. Meteorological factors involve the formation of the “polar vortex”, a circumpolar stratospheric wind phenomenon, in the polar night during the polar winter, when almost no sunlight reaches the pole. This vortex isolates the air in polar latitudes from the rest of Earth’s atmosphere, preventing ozone and other molecules from entering. As the darkness continues, the air inside the polar vortex gets very cold, with temperatures dropping below −80 °C


As the darkness continues, the air inside the polar vortex gets very cold, with temperatures dropping below −80 °C. At such temperatures a special type of clouds, called Polar Stratospheric Clouds (PSC), begins to form. Unlike tropospheric clouds, these are not primarily constituted of water droplets, but of tri-hydrated nitric acid particles, which can form larger ice particles containing dissolved nitric acid in their core as temperature continues to drop. The presence of PSC is crucial for the accelerated ozone depletion over the polar regions because they provide a solid phase in the otherwise extremely clean stratospheric air on which the ozone-degrading processes occur much more efficiently. Chemical factors involve the presence of chlorine and bromine compounds in the atmosphere as important contributors to the destruction of ozone. The majority of the chlorine compounds and half of the bromine compounds that reach the stratosphere stem from human activities


The majority of the chlorine compounds and half of the bromine compounds that reach the stratosphere stem from human activities. Due to their extreme stability, chlorofluorocarbons (CFCs) are not degraded in the troposphere but slowly (over years) transported into the stratosphere. Here, they are broken down into reactive chlorine radicals under the influence of the very energy-rich UV radiation at the upper layers of the ozone layer. One chlorine atom can destroy very high numbers of ozone molecules, before it is eventually inactivated through reaction with nitrogen oxides or methane present in the stratosphere. The degradation and inactivation scheme is illustrated in a simplified form for a CFC molecule in Fig. 10.11


The degradation and inactivation scheme is illustrated in a simplified form for a CFC molecule in Fig. 10.11. When they are isolated in the polar vortex and in the presence of PSC, these stable chlorine and bromine forms come into contact with heterogeneous phases (gas/liquid or gas/solid) on the surface of the particles forming the PSC, which breaks them down and release the activated free chlorine and bromine, known as “active” ozone depleting substances (ODS). These reactions are very fast and, as explained, strongly enhanced by the presence of PSC, a phenomenon which was neglected before the discovery of the ‘ozone hole’. While this describes the fate mechanism leading to stratospheric ozone reduction, Fig. 10.12 shows the impact pathway leading to ozone depletion in the Life Cycle Impact Assessment


Cl• O ClO• Strong UV light O3 O2+O NO2 ClONO2 CFC Fig. 10.11 Degradation of ozone catalysed by chlorine in the stratosphere (simplified) Halocarbon emissions (CFCs, HCFCs, Halons) Depletion of stratospheric ozone by catalysed cleavage Increased UVB radiation intensity at Earth’s surface Damage to ecosystems (terrestrial vegetation, phytoplankton, production of wood, crop and fish) Damage to human health (skin cancer, cataract, immune system depression) N2O emissions Decreased abundance of stratospheric ozone Fig. 10.12 Impact pathway for stratospheric ozone depletion R.K. Rosenbaum et al.


stratosphere from man-made emissions of long-lived halocarbons and nitrous oxide as used by most LCIA methods. The midpoint indicator used without exception in all LCIA methods to calculate characterisation factors is the Ozone Depletion Potential (ODP). In a similar manner as the Global Warming Potential (GWP), it evaluates the potential of a chemical to destroy the ozone layer based on a model from the World Meteorological Organization (WMO 2014)


In a similar manner as the Global Warming Potential (GWP), it evaluates the potential of a chemical to destroy the ozone layer based on a model from the World Meteorological Organization (WMO 2014). The ODP essentially expresses the global reduction in stratospheric O3 concentration CO3 due to an ozone depleting substance i relative to the global reduction of stratospheric O3 concentration CO3 due to 1 kg of CFC-11 (CFCl3), and is hence expressed in CFC-11 equivalents: ODPi 1⁄4 DCO3ðiÞ DCO3ðCFC  11Þ ð10:5Þ 10.7.3 Emissions and Main Sources The halogen compounds in the stratosphere are mostly originating from very stable industrial halocarbon gases used as solvents or refrigerants (the chlorinated CFCs or freons), or fire extinguishers (the brominated halons)


Groups of anthropogenic ODS are: bromochloromethanes (BCM), CFCs, carbon tetrachloride, hydrobromofluorocarbons (HBFCs), hydrochlorofluorocarbons (HCFCs), tetrachloromethane, 1,1,1-trichloromethane, methyl bromide, methyl chloride and halons. The main uses of ODS during the last century were: fire extinguishing systems (halon), plastic foams, propellant gas in spray cans, fumigate and pesticides (methyl bromide), metered-dose inhalers (MDIs), refrigeration and air-conditioning and solvent degreasing. Natural ozone depleting substances are CH4, N2O, H2O and halogenated substances with sufficient stability and/or release rates to allow them to reach the stratosphere


Natural ozone depleting substances are CH4, N2O, H2O and halogenated substances with sufficient stability and/or release rates to allow them to reach the stratosphere. All ozone depleting substances have two common characteristics, being: • Chemically very stable in the lower atmosphere • Capable of releasing chloride or bromide under UV radiation (photodissociation) The phasing-out of production and use of the concerned substances has been successfully enforced under the Montreal protocol, which was signed in 1987 and led to phasing-out of consumption and production of ODS by 1996 in developed countries and by 2010 in developing countries


If continuously respected, this effort should lead to the cessation of the annual appearance of the ‘ozone hole’ around 2070, the delay being due to the facts that (1) we are still emitting decreasing amounts of relevant substances (mostly during the end-of-life treatment of old refrigeration and air-conditioning systems) and (2) they are very persistent and may Life Cycle Impact Assessment


take decades to reach the poles and hence continue their adverse effects for a prolonged time. When significant emissions or dominating impacts of ODS are observed in LCIs or LCA results nowadays, it is likely because the data originate from references before the phase-out and hence it is most likely an artefact due to obsolete data, unless the end-of-life treatment of old refrigeration and air-conditioning systems are an important component of the LCA. 10.7.4 Existing Characterisation Models Without any exception, all existing LCIA methods use the ODP as midpoint indicator (although not all of them have the most recent version). For endpoint characterisation, different midpoint-to-endpoint models are applied that relate ozone depletion to increased UV radiation and ultimately to skin cancer and cataract in humans


For endpoint characterisation, different midpoint-to-endpoint models are applied that relate ozone depletion to increased UV radiation and ultimately to skin cancer and cataract in humans. All endpoint LCIA methods characterise impacts on human health, but only the Japanese method LIME additionally considers impacts on Net Primary Productivity (NPP) for coniferous forests, agriculture (soybean, rice, green pea, mustard) and phytoplankton at high latitudes. For further details see Chap. 40 and Hauschild and Huijbregts (2015). 10.8 Acidification 10.8.1 Problem During the 1980s and 90s, the effects of acidification of the environment became clearly visible in the form of a pronounced lack of health especially among conifers in many forests in Europe and the USA, resulting locally in forest decline, leading to accelerated clearing of whole forests


Clear acidic lakes without fish go right back to the beginning of the twentieth century, occurring locally for example in Norway and Sweden as a result of human activities, but the extent of the problem increased dramatically in more recent times, and during the 1990s there was serious acidification in more than 10,000 Scandinavian lakes. Metals, surface coatings and mineral building materials exposed to wind and weather are crumbling and disintegrating at a rate which is unparalleled in history, with consequent major socio-economic costs and loss of irreplaceable historic monuments in many parts of the industrialised world. The acidification problems were one of the main environmental concerns in Europe and North America in the 1980s and 90s but through targeted regulation of the main sources in the energy, industry and transportation sectors followed by liming to restore the pH of the natural soils and waters, it is no longer a major concern in these regions


In China, however, acidification impacts are dramatic in R.K. Rosenbaum et al.


some areas due to the extensive use of coal-fired power generation using sulphur-rich coal. 10.8.2 Environmental Mechanism Acidification of soil or aquatic ecosystems can be defined as an impact which leads to a fall in the system’s acid neutralising capacity (ANC), i.e. a reduction in the quantity of substances in the system which are able to neutralise hydrogen ions added to the system. ANC can be reduced by: 1. Addition of hydrogen ions, which displace other cations which can then be leached out of the system 2. Uptake of cations in plants or other biomass which is collected and removed from the system Particularly the former is relevant for acidification impacts in LCA. Acidification occurs naturally over time, but it is greatly increased by man-made input of hydrogen ions to soil and vegetation. The main source is air-borne emissions of gases that release hydrogen when they are degraded in the atmosphere or after deposition to soil, vegetation or water


The main source is air-borne emissions of gases that release hydrogen when they are degraded in the atmosphere or after deposition to soil, vegetation or water. Deposition is increased during precipitation events where the gases are dissolved in water and come down with rain, which can be rather acidic with pH values down to 3–4 in cases of strong air pollution (“acid rain”). The most important acidifying man-made compounds are: Sulphur oxides, SO2 and SO3 (or jointly SOx), the acidic anhydrides of sulphurous acid H2SO3 and sulphuric acid H2SO4, respectively, meaning that upon absorption of water from the atmosphere they form these very strong acids which both release two hydrogen ions when deposited: SO2 þ H2O ! H2SO3 ! 2H þ þ SO3 2 SO3 þ H2O ! H2SO4 ! 2H þ þ SO4 2 Nitrogen oxides, NO and NO2 (or jointly NOx) that are also acidic anhydrides as they can be converted to nitric and nitrous acids by oxidation in the troposphere


NO is oxidised to NO2 primarily by reaction with ozone (see Sect. 10.10): NO þ O3 ! NO2 þ O2 NO2 can be oxidised to nitric acid, HNO3 or HONO2: Life Cycle Impact Assessment


NO2 þ OH þ M ! HONO2 þ M where OH is hydroxyl radical present in the atmosphere and M is an inactive body which can remove surplus energy. Ammonia, which is in itself a base (absorbing hydrogen ions via the reaction NH3 þ H þ ! NH4 þ ), but upon complete mineralisation through nitrite, NO2 þ , to nitrate, NO3 releases one net proton: NH3 þ 2O2 ! H þ þ NO3  þ H2O Strong acids like hydrochloric acid, HCl or sulphuric acid, H2SO4, which release their content of hydrogen ions as soon as they are dissolved in water and thus also are strongly acidifying. Because of their high water solubility, the atmospheric residence time of these acidifying substances is limited to a few days, and therefore acidification is a regional effect with its extent limited to the region around the point of emission. When acidifying compounds deposit on plant leaves or needles, they can damage these vital plant organs and through this damage the plants


When acidifying compounds deposit on plant leaves or needles, they can damage these vital plant organs and through this damage the plants. When the acidifying compounds reach the soil, protons are released in the soil where they may lower the pH of the soil water and cause release of metal ions bound in the soil. Some of these metals are toxic to the plants in the soil, others are essential for plant growth, but after their release, they wash out, and the availability of these metals to plants may then become limiting for plant growth. The result is stress on the plants through root and leaf damage and after prolonged exposure the plants may die as a direct consequence of this or through diseases or parasites that benefit from the weakened constitution of the plant. Lakes are also exposed to the acidification, in particular through the acidified soil water leaching to the lake


Lakes are also exposed to the acidification, in particular through the acidified soil water leaching to the lake. When the pH of a lake drops, the availability of carbon in the water in its dominating form around neutral pH, which is HCO3, is converted to dissolved CO2. The solubility of toxic metals is increased, in particular aluminium which may precipitate on the gills of fish at pH 5. The phytoplankton and macrophyte flora gradually change and also the fauna is affected. Humic acids that give the lakewater a brown colour are precipitated, and the acidified lakes appear clear and blue. The sensitivity to acidification is strongly influenced by the geology and nature of the soil


The sensitivity to acidification is strongly influenced by the geology and nature of the soil. Calcareous soils with a high content of calcium carbonate are well buffered meaning that they will resist the change in pH by neutralising the input of hydrogen ions with the basic carbonate ions: H þ þ CaCO3 ! Ca2 þ þ HCO3  H þ þ HCO3  ! H2O þ CO2 As long as there is calcium carbonate in the soil, it will thus not be acidified. R.K. Rosenbaum et al.


Soils that are rich in clay are also resistant to acidification through their ability to adsorb the protons on clay mineral surfaces under release of metal ions, while sandy soils are more sensitive to acidification. The sensitivity of an ecosystem towards acidification can be described by its critical load—“A quantitative estimate of an exposure to one or more pollutants below which significant harmful effects on specified sensitive elements of the environment do not occur according to present knowledge” (Nilsson and Grennfelt 1988). Critical loads are high in calcareous regions like the Mediterranean and low in e.g. granite rock regions like most of Scandinavia. Incorporating the environmental mechanism described above, the impact pathway of acidification is illustrated in Fig. 10.13. Oceanic acidification is the process of dissolution of CO2 into seawater leading to a slight lowering of the pH in the open oceans as a consequence of increasing concentrations of CO2 in the atmosphere


Dissolution of CO2 in water generates carbonic acid, a rather weak acid (think soda water), which releases protons according to CO2 þ H2O ! H2CO3 ! HCO3  þ H þ The slightly lowered pH is deleterious to coral reefs, which should be included in endpoint characterisation. CO2 is the only important contributor to oceanic acidification and inclusion of this impact category on midpoint level therefore offers little additional information to the LCIA that already considers climate change, we will hence not discuss it further here. 10.8.3 Emissions and Main Sources Sulphur dioxides and nitrogen oxides are the man-made emissions that contribute the most to acidification. Historically metal smelters of the mining industry have been strong sources of local acidification with large localised emissions of sulphur oxides. Today, the main sources of both SOx and NOx are combustion processes in thermal power plants, combustion engines, waste incinerators and decentralised furnaces


Today, the main sources of both SOx and NOx are combustion processes in thermal power plants, combustion engines, waste incinerators and decentralised furnaces. For sulphur oxides, the level of emissions depends on the sulphur content of the fuels. Since nitrogen is abundant in the atmosphere and hence in all combustion processes using air, emissions of nitrogen oxides are mainly determined by conditions of the combustion process and possible treatment of the flue gases through catalysers and filters. As response to the serious problems with acidification in Europe and North America in previous times, regulation now ensures that sulphur content is removed from the fuels, that important combustion activities like thermal power plants and waste incinerators have an efficient neutralisation of the flue gases before they are released, and that combustion engines have catalysers lowering the NOx content of the exhaust gases. Life Cycle Impact Assessment


Ammonia is also an important contributor to acidification in some regions and the main sources are all related to agriculture using NH3 as a fertiliser, and to animal husbandry, in particular pig and chicken farms, with ammonia emissions from stables and dispersion of manure. Mineral acids like HCl and H2SO4 rarely appear as elementary flows in life cycle inventories but they may be emitted from some industrial processes and also from waste incinerators with inefficient flue gas treatment


Airborne emissions of SOx, NOx, NH3 and other acidifying substances Atmospheric fate and transport, wet and dry deposition Deposition on vegetation Deposition on land Damage to leaves and needles Buffering and change in soil pH Leaching of H+ Leaching of metal ions Toxicity to roots Acid stress and toxicity in lakes Effects on forests and vegetation Reduced metal availability Changes in species composition Damage to freshwater ecosystems Damage to terrestrial ecosystems Damage to crops Damage to forestry Fig. 10.13 Impact pathway for acidification R.K. Rosenbaum et al.


10.8.4 Existing Characterisation Models The acidification potential depends both on the potency of the emitted gas and on the sensitivity of the receiving environment in terms of buffering capacity of the soils and sensitivity of the ecosystems to acidification as expressed by their critical load. While the difference between the contributing gases is modest—within a factor 5–10 across substances, the difference between sensitivities in different locations can be several orders of magnitudes depending on the geology and soil characteristics. Early characterisation models were site-generic and only incorporated the difference in ability to release protons, but newer models incorporate more and more of the cause–effect chain in Fig. 10.13 and model e.g. the area of ecosystem in the deposition area that becomes exposed above its critical load


10.13 and model e.g. the area of ecosystem in the deposition area that becomes exposed above its critical load. This requires a site-dependent LCIA approach where the characterisation factor is determined not just per emitted substance but also per emission location. Characterisation factors may be expressed as absolute values or as an equivalent emission of a reference substance which in that case is usually SO2. For further details see Chap. 40 and Hauschild and Huijbregts (2015). 10.9 Eutrophication 10.9.1 Problem Nutrients occur naturally in the environment, where they are a fundamental precondition for the existence of life. The species composition and productivity of different ecosystems reflect the availability of nutrients, and natural differences in the availability of nitrogen and phosphorus are thus one of the reasons for the existing multiplicity of species and of different types of ecosystems


Ecosystems are dynamic, and if they are affected by a changed availability of nutrients, they simply adapt to a new balance with their surroundings. Originally, eutrophication of aquatic environments, such as rivers or lakes, describes its eutrophic character (from the Greek word “eu”—good or true, and “trophein”—feed), meaning nutrient-rich. From the 1970s the term was used to describe the slow suffocation of large lakes. It now has a meaning close to dystrophic, i.e. poor conditions and low in oxygen, supporting little life. An aquatic ecosystem in strong imbalance is named hypertrophic, when close to a natural equilibrium it is called mesotrophic, and when healthy it is called oligotrophic. The perhaps most prominent effect of eutrophication in lakes, rivers and the coastal sea are lower water quality including low visibility or for stronger situations massive amounts of algae in the surface layers of those waters


Eutrophication essentially describes the enrichment of the aquatic environment with nutrient salts leading to an increased biomass production of planktonic algae, gelatinous zooplankton and higher aquatic plants, which results in the degradation of (organoleptic) water quality (e.g. appearance, colour, smell, taste) and an altered Life Cycle Impact Assessment


species composition of the ecosystem. It may also lead to the development of toxic phytoplankton, dynophysis, cyanobacteria or blue-green algae. When the algae die, they sink to the bottom where they are degraded under oxygen consumption. As a consequence, the concentration of dissolved oxygen decreases (hypoxia), which results in biodiversity loss (flora and fauna). Ultimately, if the process is not stopped, this will turn a lake into a swamp, that will gradually become grassland and forest. This process occurs naturally but over a much longer time horizon. For terrestrial systems, the most significant environmental problem in relation to nitrogen compound loading is changes in the function and species composition of nitrogen-poor (and nitrogen limited) ecosystems in heathlands, dune vegetation, commons and raised bogs as a result of the atmospheric deposition of nitrogen compounds. Forestry and agriculture may also be affected by reduced yields via damage to forests and crops


Forestry and agriculture may also be affected by reduced yields via damage to forests and crops. This section however focuses on aquatic eutrophication. 10.9.2 Environmental Mechanism The food chain in aquatic ecosystems can be distinguished into three trophic levels: primary producers (algae and plants producing biomass via photosynthesis), primary consumers (species consuming algae and plants, the vegetarians) and secondary consumers (species consuming primary consumers, the carnivores). In addition to sunlight, growth of primary producers (algae and higher plants) requires all of the elements which enter into their anabolism (i.e. their synthesis of the molecules which constitute the organisms’ cells). A molecular formula for the average composition of an aquatic organism is C106H263O110N16P (Stumm and Morgan 1981). Apart from the elements represented in this formula, minor quantities of a large number of other elements are required, e.g


Apart from the elements represented in this formula, minor quantities of a large number of other elements are required, e.g. potassium, magnesium, calcium, iron, manganese, copper, silicon and boron (Salisbury and Ross 1978). In principle, the availability of any of these elements can determine the potential extent of the growth of the primary producers in a given system. The elements entering in greatest quantities into the primary producers (as in all other living organisms) are carbon, C, hydrogen, H and oxygen, O. The availability of water can limit growth in terrestrial plants, but the availability of one of the three basic elements is rarely a limiting factor in the growth of primary producers. The other elements which enter into the construction of the primary producers are nutrients, as the availability of these elements in sufficient quantities is necessary to ensure growth


The other elements which enter into the construction of the primary producers are nutrients, as the availability of these elements in sufficient quantities is necessary to ensure growth. The nutrients are classified as macronutrients (>1000 lg/g dry matter in plants) and micronutrients (<100 lg/g dry matter in plants) (Salisbury and Ross 1978). In rare cases, growth is limited by the availability of one of the micronutrients, but very small quantities of these elements are required by the primary producers, and these elements are therefore limiting only on very poor soils. Of the macronutrients, sulphur is added to all ecosystems in fair quantities in most of the industrialised world by the atmospheric deposition of sulphur R.K. Rosenbaum et al.


compounds from flue gases resulting from energy conversion based on fossil resources. Calcium, potassium and magnesium occur in lime and clay, respectively, which exist in large quantities in soils. In practice, one of the two last macronutrients, nitrogen and phosphorus, is therefore almost always the limiting element for the growth of primary producers, and it is therefore reasonable to regard only the elements nitrogen and phosphorus as contributors to nutrient enrichment. In many lakes, phosphorus deficiency, or a combination of nitrogen and phosphorus deficiencies, is typically limiting growth, and their addition promotes algal growth. In coastal waters and seas, nitrogen is often the limiting nutrient. Substances which contain nitrogen or phosphorus in a biologically available form are therefore classified as potential contributors to nutrient enrichment


Substances which contain nitrogen or phosphorus in a biologically available form are therefore classified as potential contributors to nutrient enrichment. As is evident from the formula for the average composition of aquatic organisms, the ratio of nitrogen to phosphorus is of the order of 16. If the concentration of bioavailable nitrogen is significantly more than 16 times the concentration of bioavailable phosphorus in an ecosystem, it is thus reasonable to assume that phosphorus is the limiting nutrient, and vice versa. Since most of the atmosphere consists of free molecular nitrogen, N2, further addition of N2 will not have any effect, and it is also not directly bioavailable. N2 is therefore not classified as contributing to nutrient enrichment. For aquatic eutrophication, the starting point of the cause–effect chain is the emission of a compound containing either nitrogen (N) or phosphorus (P)


For aquatic eutrophication, the starting point of the cause–effect chain is the emission of a compound containing either nitrogen (N) or phosphorus (P). Increased availability of nutrients will primarily increase the growth of algae and plants, especially in summer with abundant sunlight. This algae growth is visible as rivers, lakes or coastal waters turn turbid in summer. Eventually, the algae will sink to the bottom where they are decomposed by degraders like bacteria under consumption of oxygen in the bottom layer. With the sunlight being increasingly blocked from reaching deeper water layers, the build-up of a temperature gradient causes stratification in deep lakes and some coastal waters in the summer months. In the marine environment, stratification is determined by density differences between salt water flowing in from the sea and brackish water flowing out from river deltas and fjords. Such stratification prevents effective mixing of the water column


Such stratification prevents effective mixing of the water column. If fresh oxygen-rich water from the surface does not find its way to the bottom layers, the oxygen concentration near the bottom will gradually be reduced until the bottom-dwelling organisms move away or die. As the oxygen concentration approaches zero, poisonous substances such as hydrogen sulphide, H2S, are formed in the sediments, where they accumulate in gas pockets which, when released again, kill those organisms exposed to them. The main cause–effect chain as shown in Fig. 10.14 can be summarised as: • Emission of N or P containing substances • Growth and blooming of algae and higher plants increases • Sunlight no longer reaches lower water layers, which creates a temperature gradient with increasing depth • This supports a stable stratification of water layers reducing the transport of fresh oxygen-rich surface water to deeper layers Life Cycle Impact Assessment


• Oxygen is steadily depleted in bottom layers, which leads to suffocation of bottom-dwelling species and fish • This is additionally accelerated by the oxygen consuming decomposition of the dead species and sedimented dead algae • The aquatic medium becomes hypoxic and finally anoxic, favouring the formation of reducing compounds and noxious gases (mercaptans, methane) Deposition on land Deposition on water Atmospheric fate and transport Transport/removal of N and P in water compartments Airborne emissions of N-compounds Waterborne emissions of N and P compounds Waterborne emissions of dissolved organic compounds (DOC, COD, BOD) Terrestrial eutrophication Aquatic eutrophication Transport/removal in water compartments Increased nutrient concentration in aquatic compartments (P in freshwater, N in marine) Increased leaching to groundwater Increased/critical exposure of aquatic ecosystems Increased/critical exposure of terrestrial ecosystems Increased alga growth Damage to terrestrial


N in marine) Increased leaching to groundwater Increased/critical exposure of aquatic ecosystems Increased/critical exposure of terrestrial ecosystems Increased alga growth Damage to terrestrial ecosystems Damage to freshwater ecosystems Changes in species composition Damage to crops Damage to forestry Chronic effects on vegetation Oxygen depletion in bottom strata of lakes and coastal waters Damage to fishing Damage to marine ecosystems Fig


10.14 Impact pathways for terrestrial and aquatic (freshwater and marine) eutrophication [adapted from EC-JRC (2011)] R.K. Rosenbaum et al.


In a tripartite division of environmental impact categories into global, regional and local, eutrophication is considered a local to regional impact. As a consequence of the above explanations, impact potentials are highly dependent on local conditions, e.g. whether the recipient of the emission will support the requisite conversion of the emission (e.g. mineralisation of organic nitrogenous compounds), or whether the recipient is limited in nitrogen or phosphorus, while both elements are always considered potential contributors to eutrophication. The calculation of characterisation factors for a nutrient enriching substance consists of an assessment of the number of moles of nitrogen or phosphorus which can be released into the environment from one mole of the substance emitted. This can be expressed in the form of two nutrient enrichment equivalents, as kg N-equivalents and kg P-equivalents


This can be expressed in the form of two nutrient enrichment equivalents, as kg N-equivalents and kg P-equivalents. The possible consequences of eutrophication are often irrespective of whether nitrogen or phosphorus is the causing agent. In some situations it can therefore be desirable to reduce the complexity of the results of the environmental assessment by expressing eutrophication as one equivalent, so that the contributions for nitrogen and phosphorus are aggregated. In this case the impact potential may also be expressed as an equivalent emission of a reference substance (e.g. NO3 as one of the most important nutrient enriching substances). Aggregation of N and P potentials requires an assumption concerning the magnitude of the ratio N/P between these two elements in living organisms. As explained above a molar ratio of 16 can be used for nitrogen:phosphorus in living material


As explained above a molar ratio of 16 can be used for nitrogen:phosphorus in living material. One mole of phosphorus (in an area where the availability of phosphorus limits growth) therefore contributes as much to eutrophication as 16 mol of nitrogen (in an area where the availability of nitrogen limits growth). The aggregate nutrient enrichment potential for nitrogenous substances is then calculated as the emission’s N potential multiplied by the gram/mol molecular weight of the reference substance (e.g. NO3 of 62 g/mol). The P potential for phosphorous-containing substances is multiplied by 16 times the gram/mol molecular weight of the reference substance. The primary receiving compartment for agricultural emissions is mainly freshwater where some of the nitrogen may be removed on the way to the marine systems by denitrification in rivers and lakes converting the nitrogen into molecular N2 which is released to the atmosphere


Loading of freshwater with nitrogen is thus greater than the quantity conveyed to the marine areas via rivers and streams. Phosphorous compounds do not undergo this kind of conversion but phosphate forms insoluble salts with many metals and this may lead to some removal through accumulation of phosphorus in lake sediments. Phosphorus accumulated in the sediments of rivers and streams during drier periods may later be washed out into the marine environment when the water flow increases, e.g. after a thunderstorm. 10.9.3 Emissions and Main Sources Due to the use of inorganic fertilisers and manure, agriculture is a significant source of phosphorus and nitrogen emissions in the form of phosphates and nitrates, Life Cycle Impact Assessment


respectively, affecting groundwater via percolation and surface water via runoff and leaching processes, and of ammonia emitted to air and deposited on land nearby. Oxides of nitrogen may be emitted from incineration processes. Point sources in the form of wastewater treatment plants for households (e.g. from polyphosphates in detergents) and industry as well as fish farming are important sources of phosphorus and nitrates. Apart from man-made emissions, natural sources include leaching and runoff of nitrogen and phosphates. The natural addition of nutrients to terrestrial areas is believed to consist mainly of atmospheric deposition of oxides of nitrogen and ammonia while some natural plant species also possess the ability to fixate atmospheric nitrogen


Emissions of organic materials can lead to oxygen consumption by bacteria degrading this organic matter and thus contributing to oxygen depletion similarly to what is observed as a result of the nutrient enrichment of lakes and coastal waters. However, this is a primary effect and is strictly speaking not part of the nutrient enrichment mechanism. Therefore, emissions of BOD (biological oxygen demand— substances which consume oxygen on degradation) or COD (chemical oxygen demand) may additionally be characterised by some LCIA methods considering oxygen depletion (hypoxia) in water as a common midpoint for both mechanisms. Most LCIA methods are currently based on the N/P ratio and typically do not classify BOD or COD as contributing to nutrient enrichment and thus eutrophication. In large parts of the industrialised world organic matter emissions are only of local significance in watercourses and for occasional emissions of untreated effluent


In large parts of the industrialised world organic matter emissions are only of local significance in watercourses and for occasional emissions of untreated effluent. 10.9.4 Existing Characterisation Models The essential evolutions during the last decade were related to improved fate modelling, distinguishing P-limited (freshwater) and N-limited (marine) ecosystems, introduction of a midpoint effect factor in the more recent methods, and characterisation models becoming global and spatially more detailed. Midpoint LCIA methods usually propose units in P- and N-equivalents such as kg P-eq or kg PO43-eq and kg N-eq or kg NO3-eq. For endpoint characterisation most models use Potentially Disappeared Fraction of species (PDF) in [m2 years], except LIME which uses Net Primary Productivity (NPP) loss. For further details see Chap. 40 and Hauschild and Huijbregts (2015)


For further details see Chap. 40 and Hauschild and Huijbregts (2015). 10.10 Photochemical Ozone Formation This impact category appears under a number of different names in the various LCIA methods: (tropospheric) ozone formation, photochemical ozone formation or creation, photo oxidant formation, photosmog or summer smog. There are minor R.K. Rosenbaum et al.


differences, but in essence they all address the impacts from ozone and other reactive oxygen compounds formed as secondary contaminants in the troposphere by the oxidation of the primary contaminants volatile organic compounds (VOC), or carbon monoxide in the presence of nitrogen oxides (NOx) under the influence of light. VOCs are here defined as organic compounds with a boiling point below 250 °C (WHO 1989). NOx is a joint name for the nitrogen monoxide NO and nitrogen dioxide NO2. 10.10.1 Problem The negative impacts from the photochemically generated pollutants are due to their reactive nature which enables them to oxidise organic molecules in exposed surfaces. Impacts on humans arise when the ozone and other reactive oxygen compounds, which are formed in the process, are inhaled and come into contact with the surface of the respiratory tract, where they damage tissue and cause respiratory diseases


Impacts on vegetation arise when the reactive compounds attack the surfaces of plants or enter plant leaves and cause oxidative damage on their photosynthetic organs. Impacts on man-made materials are caused by oxidation and damage many types of organic materials which are exposed to ambient air. It is thus not the VOCs per se which cause the environmental problems associated with photochemical ozone formation, but the products of their transformation in the troposphere which is the lower stratum of the atmosphere, from the surface of the earth to the tropopause 8–17 km above us. Direct toxic effects on humans from VOCs are treated separately in the impact category human toxicity (see Sect. 10.12). Apart from a general increase in the tropospheric ozone concentration, photochemical ozone formation may cause smog episodes on a more local scale in and around cities with a combination of large emissions and the right meteorological conditions


During smog episodes, the concentrations of ozone and other photooxidants reach extreme levels causing immediate damage to human health. 10.10.2 Environmental Mechanism The photochemical formation of ozone and other reactive oxygen compounds in the troposphere from emissions of VOCs and NOx follows rather complex reaction schemes that depend on the nature of the specific organic compound emitted. A simplified presentation of the fundamental elements of the schemes is given in Fig. 10.15 and can be summarised as: 1. VOCs (written as RH) or CO react with hydroxyl radical OH• in the troposphere and form peroxy radicals, ROO• 2. The peroxy radicals oxidise NO to NO2 Life Cycle Impact Assessment


3. NO2 is split by sunlight with formation of NO and release of free oxygen atoms 4. Free oxygen atoms react with molecular oxygen O2 to form ozone Both VOCs and nitrogen oxides are thus needed for the photochemical ozone formation and both contribute to the formation of ozone and other oxidants. VOC and NOx sources are very heterogeneously distributed across Europe. VOC emissions involve hundreds of different organic compounds, depending on the nature of the source and activity causing the emission. This means that at the regional level, photochemical formation of ozone is highly non-linear and dynamic with the influence of meteorological conditions and on top of this the interaction between the different VOCs from both anthropogenic and natural sources like forests, and a large number of different reaction products. A further complication arises because NO may react with the formed ozone, abstracting an oxygen atom to give oxygen and NO2


A further complication arises because NO may react with the formed ozone, abstracting an oxygen atom to give oxygen and NO2. This means that depending on the conditions, NO may locally have a negative ozone formation potential and hence a negative characterisation factor for this impact category. Rather than a permanent removal of ozone this reaction of NO leads to a geographic displacement of the ozone formation since the NO2 thus formed can later cause ozone formation again following the scheme in Fig. 10.15, just in a different location. The ozone formation requires the reaction between a hydroxyl radical and a bond between carbon and hydrogen or another carbon atom in a VOC molecule. The relative strength of a volatile organic compound in terms of ozone formation potential per unit weight thus depends on how many such bonds it contains. The strength grows with the number of double or triple bonds and declines with the content of elements other than carbon and hydrogen


The strength grows with the number of double or triple bonds and declines with the content of elements other than carbon and hydrogen. The following general ranking can be given from high to low ozone formation potential: RH, CO RO light OH• ROO• NO2 NO OH• O2 O3 Fig. 10.15 Simplified presentation of the photochemical formation of ozone R.K. Rosenbaum et al.


1. Alkenes (decreasing with chain length) and aromatics (increasing with the degree of alkyl substitution, decreasing with the length of the chain in the substituted alkyl group) 2. Aldehydes (the strongest is formaldehyde; benzaldehyde has no or even a negative ozone formation potential) 3. Ketones 4. Alkanes (almost constant from a chain length of three carbon atoms and upwards), alcohols and esters (the more oxygen in the molecule, the weaker) 5. Halocarbons (decreasing with the degree of halogen substitution and the weight of the halogen element) Animals and humans are mainly exposed to the photochemical oxidants through inhalation of the surrounding air, and the effects therefore appear in their respiratory organs. Ozone is detectable by its odour at a concentration of ca


Ozone is detectable by its odour at a concentration of ca. 20 ppb in pure air, but only at somewhat higher concentrations we start to see acute symptoms like increased resistance of the respiratory passages and irritation of the eyes, followed at even higher concentrations by more serious effects like oedema of the lungs, which can lead to long-term incapacity. Smog episodes with extreme concentrations of photochemical oxidants in urban areas are known to cause increased mortality. Chronic respiratory illness may result from long-term exposure to the photochemical oxidants. Plants rely on continuous exchange of air between their photosynthetic organs (leaves or needles) and the atmosphere to absorb the carbon dioxide which is needed for photosynthesis. Ozone and other photooxidants enter together with the air and through their oxidative properties damage the photosynthetic organelles, leading to discolouration of the leaves followed by withering of the plant


The sensitivity of the plant varies with the season and also between plant species, but considerable growth reductions are observed in areas with high ozone concentrations during the growth season. Agriculture yield losses of 10–15% have been estimated for common crop plants. Figure 10.16 summarises the impact pathway for photochemical ozone formation linking emissions of VOCs, CO and NOx to the resulting damage to the areas of protection. 10.10.3 Emissions and Main Sources In some cases the emissions of individual substances are known, but in the case of oil products the emissions will often be composed of many different substances and will be specified under collective designations like VOCs or nmVOCs (non-methane VOCs, i.e. VOCs apart from methane which is typically reported separately due to its nature as a strong greenhouse gas) and sometimes also HCs (hydrocarbons), or nmHCs (non-methane hydrocarbons, i.e. hydrocarbons excluding methane). Life Cycle Impact Assessment


The most important man-made emissions of VOCs derive from road traffic and the use of organic solvents, which during 2000–2010 in Europe amounted to around 40% of the total man-made nmVOC emissions. A further 7% derives from industrial processes and 10% are fugitive emissions (Laurent and Hauschild 2014). VOCs are also emitted in large quantities from vegetation, in particular forests, but unless a man-made manipulation of the natural system affects its emissions of VOCs, these will not be reported in an LCI and hence not dealt with in the impact assessment. Carbon monoxide is emitted from combustion processes with insufficient oxygen supply. These include road traffic and various forms of incomplete combustion of fossil fuels or biomass in stationary systems. Nitrogen oxides are also emitted from combustion processes in transport, energy- and waste incineration systems


Nitrogen oxides are also emitted from combustion processes in transport, energy- and waste incineration systems. Photochemical oxidation of VOCs Increase in tropospheric ozone concentration Sun light Airborne emissions of VOCs and CO Airborne emissions of NOx Increased/critical exposure of vegetation Increased/critical exposure of humans Chronic effects on vegetation Acute effects on vegetation Chronic effects on humans Damage to ecosystems Damage to crops Damage to forestry Damage to human health Acute effects on humans Fig. 10.16 Impact pathway for photochemical ozone formation [adapted from EC-JRC (2011)] R.K. Rosenbaum et al.


10.10.4 Existing Characterisation Models The complexity of the underlying reaction schemes and the high number of individual contributing substances for which photochemical ozone formation characterisation factors must be calculated calls for simplification in the characterisation modelling. Existing characterisation models apply one of two approaches: The first alternative is to simplify the non-linear and dynamic behaviour of the photochemical oxidation schemes by modelling one or a few typical situations in terms of meteorology, atmospheric chemistry and concomitant emissions of other air pollutants. For each individual VOC, characterisation factors may then be presented for each situation or in the form of a weighted average across the situations. The second alternative is to ignore the variation between individual VOCs and concentrate on getting the spatial and temporal specificities well represented in the characterisation model


The second alternative is to ignore the variation between individual VOCs and concentrate on getting the spatial and temporal specificities well represented in the characterisation model. This approach leads to spatially (and possibly temporally) differentiated characterisation factors for VOCs (as a group, ignoring variation in strength between individual substances), CO and NOx. Often methane is treated separately from the rest of the VOCs (which are then termed non-methane VOCs or nmVOCs) due to its very low characterisation factor which really distinguishes it from the majority of the other VOCs. The first approach is adopted in characterisation models based on the POCP (Photochemical Ozone Creation Potential) or MIR (Maximum Incremental Reactivity) concepts. The second approach is adopted in regionally differentiated models which attempt to capture the non-linear nature of the ozone formation with its spatially and temporally determined differences. For further details see Chap


For further details see Chap. 40 and Hauschild and Huijbregts (2015). 10.11 Ecotoxicity The contents of this section have been modified from Rosenbaum, R.K.: Ecotoxicity, appearing as Chapter 8 of Hauschild M. Z. and Huijbregts M. A. J. (eds.) LCA Compendium—The Complete World of Life Cycle Assessment—Life Cycle Impact Assessment, pp 139–162. Springer, Dordrecht (2015). 10.11.1 Problem About 500 years ago Paracelsus stated that ‘All substances are poisons; there is none which is not a poison. The right dose differentiates a poison and a remedy’. Today’s toxicology science still agrees and adheres to this principle and in consequence any substance emitted may lead to toxic impacts depending on a number of driving factors: (1) emitted quantity (determined in the LCI), (2) mobility, Life Cycle Impact Assessment


(3) persistence, (4) exposure patterns and bioavailability and (5) toxicity, with the latter four considered by the characterisation factor. This shows that toxicity is not the only parameter that determines the potential ecotoxic impact of a chemical in the environment as it first has to reach and enter a potential target organism. For example, a substance may be very toxic, but never reach any organism due to its short lifetime in the environment (e.g. rapid degradation) or because it is not sufficiently mobile to be transported to a target organism and ends up bound to soil or buried in sediment, in which case it contributes little to ecotoxic impacts. On the other hand, another substance may not be very toxic, but if it is emitted in large quantities and over prolonged periods of time or has a strong environmental persistence, it may still cause an ecotoxic impact. Chemical emissions into the environment will affect terrestrial, freshwater, marine and aerial (i.e


Chemical emissions into the environment will affect terrestrial, freshwater, marine and aerial (i.e. flying and gliding animals) ecosystems depending on the environmental conditions of the place and time of emission and the characteristics of the substance emitted. They can affect natural organisms in many different ways, causing increased mortality, reduced mobility, reduced growth or reproduction rate, mutations, behavioural changes, changes in biomass or photosynthesis, activity etc. 10.11.2 Environmental Mechanism As shown in Fig. 10.17, the environmental mechanism of ecotoxic impacts of chemicals in LCA can be divided into four consecutive steps. 1. Fate modelling estimates the increase in concentration in a given environmental medium due to an emission quantified in the life cycle inventory 2. The exposure model quantifies the chemical’s bioavailability in the different media by determining the bioavailable fraction out of the total concentration 3


The exposure model quantifies the chemical’s bioavailability in the different media by determining the bioavailable fraction out of the total concentration 3. The effect model relates the amount available to an effect on the ecosystem. This is typically considered a midpoint indicator in LCA, as no distinction between the severity of observed effects is made (e.g. a temporary/reversible decrease in mobility and death are given the same importance) 4. Finally, the severity (or damage) model translates the effects on the ecosystem into an ecosystem population (i.e


Finally, the severity (or damage) model translates the effects on the ecosystem into an ecosystem population (i.e. biodiversity) change integrated over time and space All four parts of this environmental mechanism are accounted for in the definition of the substance-specific and emission compartment-specific ecotoxicity characterisation factor CFeco: CFeco 1⁄4 FF  XFeco  EFeco  SFeco ð10:6Þ where FF is the fate factor, XFeco the ecosystem exposure factor, EFeco the ecotoxicity effect factor (midpoint effects), and SFeco the ecosystem severity factor (endpoint effects). Each of these four elements of the environmental mechanism of R.K. Rosenbaum et al.


ecotoxicity, and thus its characterisation factor, is described in the following sections. Some LCIA methods also directly combine EFeco and SFeco into a single damage factor, directly calculating an endpoint characterisation factor. For midpoint characterisation, SFeco is simply omitted and CFeco is then the midpoint ecotoxicity characterisation factor. A method for toxic impact assessment of chemicals in the framework of LCA must be able to cover the very large number of potentially toxic substances in the inventory in terms of available characterisation factors. It must also be based on integration of the impact over time and space as LCI data are typically not spatially and/or temporally differentiated, and the characterisation factor must relate to a mass flow and not require any information about concentrations of the substance as this information is not available in the LCI


To be compatible with the effect model, the fate model must translate chemical emissions calculated in the life cycle inventory into an increase in concentration in the relevant medium. In the Individual species 1,2,...n Multiple species & ecosystem Algae Crustacean Fish Individual species Trophic level, e.g. Ecosystem level Damage to marine ecosystems Damage to freshwater ecosystems Damage to terrestrial ecosystems Environmental fate (transformation and distribution between environmental compartments) Ecotoxic effects Emissions to air, water and soil Fig. 10.17 General scheme of the Impact pathway for ecotoxicity [adapted from EC-JRC (2011)] Life Cycle Impact Assessment


characterisation modelling this leads to the use of fate models assuming steady-state conditions. The fate model predicts the chemical behaviour/distribution in the environment accounting for multimedia (i.e. between environmental media and compartments) and spatial (i.e. between different zones but within the same compartment or medium) transport between environmental compartments (e.g. air, water, soil). This is accomplished via modelling of (thermodynamic) exchange processes such as partitioning, diffusion, sorption, advection, convection—represented as arrows in Fig. 10.18—as well as biotic and abiotic degradation (e.g. biodegradation, hydrolysis or photolysis), or burial in sediments. Degradation is an important loss process for most organic substances, but may also lead to toxic breakdown compounds


biodegradation, hydrolysis or photolysis), or burial in sediments. Degradation is an important loss process for most organic substances, but may also lead to toxic breakdown compounds. The rate by which the degradation occurs can be derived from the half-life of the substance in the medium and it depends both on the properties of the substance and on environmental conditions such as temperature, insolation or presence of reaction partners (e.g. OH radicals for atmospheric degradation). The basic principle underlying a fate model is a mass balance for each compartment leading to a system of differential equations which are solved simultaneously, which can be done for steady-state or dynamic conditions. A life cycle inventory typically reports emissions as masses emitted into an environmental compartment for a given functional unit


A life cycle inventory typically reports emissions as masses emitted into an environmental compartment for a given functional unit. The mathematical relationship between the steady-state solution for a continuous emission and the time-integrated solution for a mass of chemical released into the environment has been demonstrated (Heijungs 1995; Mackay and Seth 1999). Figure 10.18 shows the overall nested structure of the USEtox model which is a widely used global scientific consensus model for characterisation modelling of human and ecotoxic impacts in LCA. Further details on fate modelling principles in the USEtox model can be found in Henderson et al. (2011) and Rosenbaum et al. (2008). Exposure is the contact between a target organism and a pollutant over an exposure boundary for a specific duration and frequency. The exposure model accounts for the fact that not necessarily the total (‘bulk’) chemical concentration present in the environment is available for exposure of organisms


The exposure model accounts for the fact that not necessarily the total (‘bulk’) chemical concentration present in the environment is available for exposure of organisms. Several factors and processes such as sorption, dissolution, dissociation and speciation may influence (i.e. reduce) the amount of chemical available for ecosystem exposure. Such phenomena can be defined as bioavailability (“freely available to cross an organism’s cellular membrane from the medium the organism inhabits at a given time”), and bioaccessibility (“what is actually bioavailable now plus what is potentially bioavailable”). The effect model characterises the fraction of species within an ecosystem that will be affected by a certain chemical exposure. Effects are described quantitatively by lab-test derived concentration-response curves relating the concentration of a chemical to the fraction of a test group that is affected (e.g


Effects are described quantitatively by lab-test derived concentration-response curves relating the concentration of a chemical to the fraction of a test group that is affected (e.g. when using the EC50— the Effect Concentration affecting 50% of a group of individuals of the same test species compared to a control situation). Affected can mean various things, such as increased mortality, reduced mobility, reduced growth or reproduction rate, R.K. Rosenbaum et al.


mutations, behavioural changes, or changes in biomass or photosynthesis. These are the effects that may be observed during standardised laboratory-based ecotoxicity tests, and the results are specific for each combination of substance and species. Toxic effects are further distinguished into acute, sub-chronic and chronic toxicity (including further sub-groups like sub-acute, etc.). Acute toxicity describes an adverse effect after a short period of exposure, relative to the lifetime of the animal (e.g. <7 days for vertebrates, invertebrates or plants and <4 days for algae). Chronic toxicity is based on exposure over a prolonged period of time covering at least one life cycle or one sensitive period (e.g.  32 days for vertebrates,  21 days for invertebrates,  7 days for plants and  4 days for algae). When relating to freshwater ecosystems, the question arises what exactly we mean by that


32 days for vertebrates,  21 days for invertebrates,  7 days for plants and  4 days for algae). When relating to freshwater ecosystems, the question arises what exactly we mean by that. In LCIA, a freshwater ecosystem is typically seen as consisting of at least three trophic levels: 1. Primary producers, converting sunlight into biomass via photosynthesis (i.e. phytoplankton, algae) 2. Primary consumers, living off primary producers (i.e. zooplankton, invertebrates, planktivorous fish) 3. Secondary consumers at the upper end of the aquatic food chain (i.e. piscivorous fish) air air agricultural soil natural soil freshwater ocean Global scale agricultural soil urban air natural soil freshwater Coastal marine water Continental scale Fig. 10.18 The USEtox fate model [taken from Rosenbaum et al. (2008)] Life Cycle Impact Assessment


It should be noted that only impacts on cold-blooded species in freshwater ecosystems are currently considered. There is no minimum requirement established, which trophic levels should be covered by a characterisation factor for terrestrial or marine ecosystems and available methods usually extrapolate from freshwater data or use the relatively few data available directly for these ecosystems. There is often a large variation of sensitivity to a given substance between different species in the freshwater ecosystem. This is described by a species-sensitivity distribution (SSD) curve, which hence represents the sensitivity of the entire ecosystem to a substance—see Fig. 10.19. The SSD is constructed using the respective geometric mean of all available and representative toxicity values for each species. This curve represents the range of sensitivities to exposure to a given substance among the different species in an ecosystem from the most sensitive to the most robust species


This curve represents the range of sensitivities to exposure to a given substance among the different species in an ecosystem from the most sensitive to the most robust species. The ecotoxicity effect factor is then calculated using the HC50—Hazardous Concentration at which 50% of the species (in an aquatic ecosystem) are exposed to a concentration above their EC50, according to the SSD curve (see Fig. 10.19). The dimension of the effect factor is PAF—Potentially Affected Fraction of species, while the unit is typically m3/kg. The ecotoxicological effect factor of a chemical is calculated as: EFeco 1⁄4 0:5 HC50 ð10:7Þ 1.0 0.2 0.5 0.8 Environmental concentration (mg/l) Cumulative EC50 distribution Potentially affected fraction of species (PAF) HC50 0.1 0.01 Fig. 10.19 Species-sensitivity distribution (SSD) curve representing the sensitivity of the ecosystem to a chemical substance R.K. Rosenbaum et al.


The HC50 value can be determined from the SSD curve but is often, more conveniently, calculated as the geometric mean of the EC50 values per species s, respectively: log HC50 1⁄4 1 ns  X s log EC50s ð10:8Þ where ns is the number of species. A damage model, incorporating the severity of the effect, goes even further along the cause–effect chain and quantifies how many species are disappearing (instead of ‘just’ affected) from a given ecosystem. Disappearance may be caused by mortality, reduced proliferation or migration, for example. 10.11.3 Emissions and Main Sources Chemicals are a main pillar of our industrialised economy, they are used in virtually any product around the globe and therefore numerous, used in large quantities and emitted from nearly all processes that an LCI may contain. Ecotoxity is very different from any other (non-toxicity) impact category when it comes to the number of potentially relevant elementary flows


Ecotoxity is very different from any other (non-toxicity) impact category when it comes to the number of potentially relevant elementary flows. Whereas no other (non-toxicity) impact category—with the exception of photochemical ozone formation—exceeds 100 contributing elementary flows (and related characterisation factors), the toxicity categories are facing the challenge of having to characterise several tens of thousands of chemicals with huge differences in their abilities to cause toxic impacts


The CAS registry currently (end of 2016) contains more than 124 million unique organic and inorganic structures (www.cas.org/about-cas/cas-fact-sheets) of which roughly 200,000 may play an industrial role as reflected by the ever increasing number of more than 123,000 substances registered in the European Classification and Labelling Inventory Database which contains REACH (Registration, Evaluation, Authorisation and Restriction of Chemical substances) registrations and CLP (Classification, Labelling and Packaging of substances and mixtures) notifications so far received by the European Chemicals Agency (ECHA: http://echa. europa.eu/information-on-chemicals/cl-inventory-database). Current LCIA models cover around 3000 substances for aquatic ecotoxicity. 10.11.4 Existing Characterisation Models Characterisation methods like EDIP account for fate and exposure relying on key properties of the chemical applied to empirical models


10.11.4 Existing Characterisation Models Characterisation methods like EDIP account for fate and exposure relying on key properties of the chemical applied to empirical models. Mechanistic models and methodologies have been published accounting for fate, exposure and effects Life Cycle Impact Assessment


providing cardinal impact measures. Among these methods are IMPACT 2002 (used in IMPACT 2002+) and USES-LCA (used in CML and ReCiPe). All these methods adopt environmental multimedia, multipathway models employing mechanistic cause–effect chains to account for the environmental fate, exposure and effects processes. However, they do not necessarily agree on how these processes are to be modelled, leading to variations in results of LCA studies related to the choice of LCIA method. Based on an extensive comparison of these models followed by a consensus-building process, the scientific consensus model USEtox (UNEP/SETAC toxicity consensus model) was developed with the intention to solve this situation by representing a scientifically agreed consensus approach to the characterisation of human toxicity and freshwater ecotoxicity (Hauschild et al. 2008; Rosenbaum et al. 2008; Henderson et al. 2011)


2008; Rosenbaum et al. 2008; Henderson et al. 2011). It has been recommended and used by central international organisations like the United Nations Environment Program UNEP, Society of Environmental Toxicology and Chemistry SETAC, the European Commission and US-EPA to characterise human and ecotoxicity in LCIA. Among the existing characterisation models on midpoint level, three main groups can be distinguished: (1) mechanistic, multimedia fate, exposure and effect models, (2) key property-based partial fate models and (3) non-fate models (EC-JRC 2011). According to ISO 14044 (2006b) “Characterisation models reflect the environmental mechanism by describing the relationship between the LCI results, category indicators and, in some cases, category endpoints


[...] The environmental mechanism is the total of environmental processes related to the characterisation of the impacts.” Therefore, ecotoxicity characterisation models falling into categories (2) and (3) do not completely fulfil this criterion. Caution is advised regarding their use and most importantly the interpretation of their results, which should not be employed without prior in-depth study of their respective documentation. Having said that, depending on the goal and scope of the LCA, they may still be an adequate choice in some applications, and indeed may agree quite well with the more sophisticated multimedia-based models. Ecotoxicity endpoint modelling is still in an early state and much research needs to be performed before maturity is reached


Ecotoxicity endpoint modelling is still in an early state and much research needs to be performed before maturity is reached. The authors of the ILCD LCIA handbook concluded that “For all the three evaluated endpoint methods (EPS2000, ReCiPe, IMPACT 2002+), there is little or no compliance with the scientific and stakeholder acceptance criteria, as the overall concept of the endpoint effect factors is hardly validated and the endpoint part of the methods is not endorsed by an authoritative body. [...] No method is recommended for the endpoint assessment of ecotoxicity, as no method is mature enough.” (EC-JRC 2011). When interpreting the results of existing methods, it is important to keep in mind that many aspects are not or only very insufficiently covered. This includes elements like terrestrial and marine ecotoxicity as well as toxicity of pesticides in pollinators. For further details see Chap. 40 and Hauschild and Huijbregts (2015). R.K. Rosenbaum et al.


10.12 Human Toxicity As explained in Sect. 10.11, both toxicity impact categories have a number of things in common, like main emissions and sources, modelling principles, model structure and even some of the models used in the characterisation are identical between the human toxicity and ecotoxicity impact categories. Notably the fate model used is the same in LCIA methods using mechanistic characterisation modelling, which is the majority of existing methods. Therefore, only those parts that are specific for human toxicity and different from ecotoxicity will be discussed here. It is recommended to first read Sect. 10.11 in order to understand the main underlying principles not repeated hereafter. 10.12.1 Problem Human toxicity in LCA is based on essentially the same driving factors as ecotoxicity: (1) emitted quantity (determined in the LCI), (2) mobility, (3) persistence, (4) exposure patterns and (5) human toxicity, with the latter four considered by the characterisation factor


The respective mechanisms and parameters are certainly different and specific for human toxicity, notably for the exposure modelling, where many factors capturing human behaviour, such as dietary habits, influence human exposure pattern. Chemical exposure of humans can result from emissions into the environment which will affect the whole population, but also from the many chemical ingredients in products released during their production, use, or end-of-life treatment and thus affecting workers or consumers. Chemical emissions are responsible for, or contribute to, many health impacts such as a wide range of non-cancer diseases as well as increased cancer risks for those chemicals that are carcinogenic. 10.12.2 Environmental Mechanism Modelling the toxicological effects on human health of a chemical emitted into the environment, whether released on purpose (e.g


10.12.2 Environmental Mechanism Modelling the toxicological effects on human health of a chemical emitted into the environment, whether released on purpose (e.g. pesticides applied in agriculture), as a by-product from industrial processes, or by accident, implies a cause–effect chain, linking emissions and impacts through four consecutive steps as depicted in Fig. 10.20. The cause–effect chain links the emission to the resulting mass in the environmental compartments (fate model) and on to the intake of the substance by the overall population via food and inhalation exposure pathways (human exposure model), and to the resulting number of cases of various human health risks by comparison of exposure with the known dose-response relationship for the Life Cycle Impact Assessment


chemical (toxic effect model) and finally their damage to the health of the overall population. In the characterisation modelling, the links of this cause–effect chain are expressed, similarly to Eq. 10.6, as factors corresponding to the successive steps of fate, exposure, effects and severity: CFhh 1⁄4 FF  XFhh  EFhh  SFhh ð10:9Þ where CFhh is the human health characterisation factor, FF the fate factor, XFhh the human exposure factor, EFhh the human toxicity effect factor (midpoint effects) and SFhh the human health severity factor (endpoint effects). Some LCIA methods also directly combine EFhh and SFhh into a single damage factor, directly calculating an Indoor air home or workplace Fish Agricultural produce Meat Cancer diseases Non cancer diseases Damage to human health Environmental fate (transformation and distribution between environmental compartments) Outdoor emissions to air, water and soil Emissions to indoor air Toxic effects Human exposure Drinking water Milk Air Fig


10.20 General scheme of the impact pathway for human toxicity [adapted from EC-JRC (2011)] R.K. Rosenbaum et al.


endpoint characterisation factor. For midpoint characterisation, SFhh is simply omitted and CFhh is then the midpoint human toxicity (i.e. not human health) characterisation factor. The midpoint human toxicity characterisation factor [number of cases/kgemitted] expresses the toxic impact on the global human population per mass unit emitted into the environment and can be interpreted as the increase in population risk of disease cases due to an emission into a specific environmental compartment. The endpoint human health characterisation factor [DALY/kgemitted] quantifies the impact on human health in the global population in Disability-Adjusted Life Years (DALY) per mass unit emitted into the environment. DALY is a statistical measure of population life years lost or affected by disease (or other influences) and is used among other by the World Health Organisation. The fate model is, without exception, the same as for ecotoxicity


The fate model is, without exception, the same as for ecotoxicity. Logically, the environment in which a chemical is transported, distributed and transformed is the same, no matter who will be affected. Therefore, for the sake of consistency, all LCIA methods that cover human toxicity are using the same fate model as for ecotoxicity, but of course different exposure and effect models, as this will be specific for the targeted organism (human or ecosystem species). The fate model is therefore the same as described in Sect. 10.11. The exposure model relates the amount of chemical in a given environmental compartment to the chemical intake by humans (exposure rates). It can be differentiated into direct intake (e.g. by breathing air and drinking water), indirect intake through bioconcentration processes in animal tissues (e.g. meat, milk and fish) and intake by dermal contact


by breathing air and drinking water), indirect intake through bioconcentration processes in animal tissues (e.g. meat, milk and fish) and intake by dermal contact. An exposure pathway is defined as the course a chemical takes from the environment to the exposed population, for example through air, meat, milk, fish, water or vegetables. Exposure pathways can be further aggregated into exposure routes, such as inhalation of air, ingestion of food including drinking water and other matter such as soil particles and dermal exposure. The human exposure model is designed for assessing human exposure to toxic chemical emissions applying realistic exposure assumptions and being adapted to take spatial variability into account. In LCIA human exposure is always assessed at the population level


In LCIA human exposure is always assessed at the population level. The intake fraction iF is calculated as the product of fate and exposure factor (iF = FF * XFhh [kgintake/kgemitted]) and it can be interpreted as the fraction of an emission that is taken in by the overall population through all exposure routes, i.e. as a result of food contamination, inhalation and dermal exposure. A high value, such as iF = 0.001 for dioxins, reflects that humans will take in 1 part out of 1000 of the mass of a chemical released. Dioxins are very efficient in exposing humans as reflected by the high intake fraction. For other chemical emissions, intake fraction values typically lie in the range of 10−10 to 10−5


Dioxins are very efficient in exposing humans as reflected by the high intake fraction. For other chemical emissions, intake fraction values typically lie in the range of 10−10 to 10−5. The effect model relates the quantity of a chemical taken in by the population via a given exposure route (inhalation and ingestion, respectively, dermal uptake is not currently modelled in LCIA) to the toxic effects of the chemical once it has entered the human organism and can be interpreted as the increase in the number of cases of a given human health effect (e.g. cancer or non-cancer diseases) in the exposed population per unit mass taken in. The two general effect classes, cancer and Life Cycle Impact Assessment


non-cancer, each cover a multitude of different diseases, so this is a simplification reflecting the fact that it is very difficult to predict the many underlying human toxicity endpoints from the animal dose-response curves from laboratory experiments with test animals which are normally the basis of the effect factor. The severity factor represents adversely affected life years per disease case (DALY/case), distinguishing between differences in the severity of disabilities caused by diseases in terms of affected life years, e.g. discriminating between a lethal cancer and a reversible skin irritation. It is quantified by the statistically determined, population-based years of life lost (YLL) and years of life disabled (YLD) due to a disease. 10.12.3 Emissions and Main Sources The relevant emissions and main sources are identical to those of the ecotoxicity impact category and discussed in Sect. 10.11. 10.12.4 Existing Characterisation Models Again here, Sect


10.11. 10.12.4 Existing Characterisation Models Again here, Sect. 10.11 contains a discussion on existing characterisation models, which largely applies also to the human toxicity impact category. In USEtox, the units of the two human toxicity midpoint indicators for non-cancer and cancer are Comparative Toxic Unit for humans CTUh [disease cases]. They can be added up to a single human health indicator, but then the interpretation needs to consider that this intrinsically assumes equal weighting between cancer and non-cancer effects (which includes equal weighting between e.g. a reversible skin rash and non-reversible death). Human health endpoint indicators in USEtox are given in the Comparative Damage Unit for human health CDUh [DALY]. In accordance with the purpose of endpoint modelling, this indicator better represents the distinction of the severity of different effects


In accordance with the purpose of endpoint modelling, this indicator better represents the distinction of the severity of different effects. When interpreting human toxicity indicators from existing methods, it is important to be aware that these only provide indicators for global population exposure to outdoor and indoor emissions, while human toxicity for occupational exposure of workers or direct exposure related to product use for consumers are not yet covered by USEtox and the other characterisation models, despite their very high relevance. Products of special interest in this context are cosmetics, plant protection products, textiles, pharmaceuticals and many others, that may in particular contain substances having toxic properties and have the potential to cause mutagenic, neurotoxic or endocrine disrupting effects. This is the subject of ongoing research and will be included in LCIA methods once the models are mature and operational. For further details see Chap


This is the subject of ongoing research and will be included in LCIA methods once the models are mature and operational. For further details see Chap. 40 and Hauschild and Huijbregts (2015). R.K. Rosenbaum et al.


10.13 Particulate Matter Formation In existing LCIA methods, health impacts from exposure to particulate matter (PM) as impact category is referred to by different terms (e.g. ‘particulate matter/respiratory inorganics’ in ILCD, ‘respiratory effects’ in IMPACT 2002+, ‘human health criteria pollutants’ in TRACI, or ‘particulate matter formation’ in ReCiPe). Although causing mainly toxicity-related health effects, exposure to PM is considered a separate impact category in most LCIA methods. This is mainly due to a number of important differences between the characterisation of PM formation and that of human toxicity. These differences include the complex atmospheric chemistry involved in the formation of secondary PM from different precursor substances which requires a different fate model


These differences include the complex atmospheric chemistry involved in the formation of secondary PM from different precursor substances which requires a different fate model. Furthermore, different emission heights are important to consider, global monitoring data for PM air concentrations are used, and the effect assessment is based on exposure-response functions mostly derived from epidemiological evidence, which is not possible for most toxic chemicals due to missing emission locations and exposure- or dose-response information. 10.13.1 Problem A large number of studies including the global burden of disease (GBD) study series consider particulate matter (PM) to be a leading environmental stressor contributing to global human disease burden (i.e. all diseases around the world) via occupational and household indoor exposure as well as urban and rural outdoor (ambient) exposures


all diseases around the world) via occupational and household indoor exposure as well as urban and rural outdoor (ambient) exposures. In 2013, outdoor PM pollution accounted for 2.9 million deaths and 70 million DALY, and household PM pollution from solid fuels accounted for 2.9 million deaths and 81 million DALY (Forouzanfar et al. 2015). With that, outdoor and household PM pollution combined contributed in 2013 with 71% to premature deaths attributable to all environmental risk factors and with 19% to premature death attributable to all risk factors (i.e. including behavioural etc.). This means that exposure to PM accounts on average for 1 out of 5 premature deaths worldwide


including behavioural etc.). This means that exposure to PM accounts on average for 1 out of 5 premature deaths worldwide. Thereby, exposure to PM is associated in epidemiological and toxicological studies with various adverse health effects and reduction in life expectancy including chronic and acute respiratory and cardiovascular diseases, chronic and acute mortality, lung cancer, diabetes and adverse birth outcomes (Fantke et al. 2015). PM can be distinguished according to formation type (primary and secondary) and according to aerodynamic diameter (respirable, coarse, fine and ultrafine). Primary PM refers to particles that are directly emitted, e.g. from road transport, power plants or farming activities. Secondary PM refers to organic and inorganic particles formed through reactions of precursor substances including nitrogen oxides (NOx), sulphur oxides (SOx), ammonia (NH3), semivolatile and volatile organic compounds (VOC)


Secondary particles include sulphate, nitrate and Life Cycle Impact Assessment


organic carbonaceous materials and can make up to 50% of ambient PM concentrations. Respirable particles (PM10) have an aerodynamic diameter less than 10 μm, coarse particles (PM10–2.5) between 2.5 and 10 μm, fine particles (PM2.5) less than 2.5 μm, and ultrafine particles (UFP) less than 100 nm (WHO 2006). PM2.5 is often referred to as the indicator that best describes the component of PM responsible for adverse human health effects (Lim et al. 2012; Brauer et al. 2016). 10.13.2 Environmental Mechanism Characterising health impacts from exposure to PM associated with emissions of primary PM or secondary PM precursor substances builds on the general LCIA framework for characterising emissions of air pollutants (see Fig. 10.2). The impact pathway for health impacts from PM emissions is illustrated in Fig. 10.21 and starts from primary PM emissions or secondary PM precursor substances emitted into air


10.2). The impact pathway for health impacts from PM emissions is illustrated in Fig. 10.21 and starts from primary PM emissions or secondary PM precursor substances emitted into air. As for the toxicity impact categories, combining all factors from emission to health impacts or damages yields the characterisation factor for particulate matter formation (CF) with units [disease cases/kgemitted] at midpoint level (i.e. excluding SF) and [DALY/kgemitted] at endpoint level: Mass emitted to air [kg PM or precursor emitted] Human health impacts [Disability-adjusted life years, DALY] Time-integrated mass in air [kg PM in air day] Mass inhaled [kg PM inhaled] Disease incidences [cases] Impact Pathway Fate factor [day] Exposure factor [1/day] Effect factor [cases/kg PM inhaled] Severity factor [DALY/case] Intake fraction [kg PM inhaled/ kg PM or precursor emitted] [DALY/kg PM inhaled] Characterisation factor [DALY/kg PM or precursor emitted] Intermediate and final LCIA output metrics Fig


10.21 Schematic impact pathway and related output metrics for characterising health impacts from particulate matter (PM) exposure in life cycle impact assessment [adapted from Fantke et al. (2015)] R.K. Rosenbaum et al.


CF 1⁄4 FF  XF  EF  SF ð10:10Þ Emissions are expressed as mass of PM or precursor substance released into air. From there, the impact pathway follows different distribution processes within and between air compartments and/or regions (indoor, outdoor, urban, rural, etc.) yielding a time-integrated mass of PM in the different air compartments and/or regions. Relating the time-integrated PM mass in air to the mass of PM or precursor substance emitted yields the fate factor (FF) with unit kg in air integrated over one day per kg emitted. A certain fraction of PM mass in air is subsequently inhaled by an exposed human population. This fraction is expressed by the exposure factor (XF) describing the rate at which PM is inhaled with unit kg PM inhaled per kg PM in air integrated over one day. Multiplying FF and XF yields the cumulative PM mass inhaled by an exposed population per kg PM or precursor emitted expressed as human intake fraction (iF)


Multiplying FF and XF yields the cumulative PM mass inhaled by an exposed population per kg PM or precursor emitted expressed as human intake fraction (iF). Inhaling PM mass may then lead to a cumulative population risk referred to as expected disease incidences in the exposed human population and typically assessed based on PM air concentration. Relating PM concentration in air to cumulative population risk yields the exposure-response or effect factor (EF) with unit disease cases (e.g. death for mortality effects) per kg PM inhaled. Finally, disease incidences are translated into human health damages by accounting for the disease severity expressed as disability-adjusted life years (DALY) that include mortality and morbidity effects. Linking health damages to disease incidences yields the severity (or damage) factor (SF) with unit DALY per disease case


Linking health damages to disease incidences yields the severity (or damage) factor (SF) with unit DALY per disease case. For characterising health impacts from emissions of PM or precursor substances, several aspects influence emission, fate, intake and health effects. Regardless of the modelling setup (spatial vs. archetypal; including or disregarding indoor sources and/or secondary PM formation, etc.), main influential aspects are spatiotemporally variable population density and activity patterns, background PM concentration in air, background disease rate and background severity, emission location (e.g. indoor vs. outdoor or urban vs. rural) and emission height, as well as potential nonlinearity in the disease-specific exposure-response relationship. The effect of using a non-linear exposure-response curve in the calculation of CFs following the marginal and average approach is illustrated in Fig


The effect of using a non-linear exposure-response curve in the calculation of CFs following the marginal and average approach is illustrated in Fig. 10.22 for two distinct background concentration scenarios, where the difference between marginal and average approach is increasing with increasing background concentration for an exposure-response curve of supralinear shape. 10.13.3 Emissions and Main Sources Substances considered in the different LCIA methods to contribute to health impacts from PM are typically one or more PM fractions (PM10, PM10–2.5, PM2.5) Life Cycle Impact Assessment


and PM precursor substances (mostly NOx, SO2 and NH3) and in some cases also carbon monoxide (e.g. IMPACT 2002+) or non-methane volatile organic compounds (e.g. ReCiPe). Relevant emission sources of PM (and/or precursors) are for example road traffic, stationary emissions from coal/gas-fired power plants or indoor emissions from solid fuels combustion. Several emission sources are ground-level sources (e.g. road traffic and household combustion), while others are considered to occur at higher stack levels (typically stationary emission sources, e.g. power plants). 10.13.4 Existing Characterisation Models In LCIA, archetypal impact assessment scenarios (e.g. urban, rural) are often used instead of spatialized or site-specific scenarios, especially when emission locations are unknown or fate, exposure and/or effect data do not allow for spatial differentiation. Such archetypal approach and related intake fractions were proposed by Humbert et al


Such archetypal approach and related intake fractions were proposed by Humbert et al. (2011) with population density (urban, rural and remote) and emission height (ground-level, low-stack and high-stack emissions) as main determinants of PM and precursor impacts. The UNEP/SETAC Life Cycle Initiative established a task force to build a framework for consistently quantifying health effects from PM exposure and for recommending PM characterisation factors for application in LCIA with fine particulate matter (PM2.5) as representative indicator. First recommendations from this task force focus on the integration of 100 120 140 160 180 200 220 240 annual deaths per 105 persons fine parƟculate maƩer air concentraƟon [μg/m3] exposure-response curve marginal slope average slope working point 1 working point 2 Fig


10.22 Illustration of using a non-linear exposure-response curve for health effects from fine particulate matter exposure with dashed and dotted lines as approaches for calculating marginal and average (between working point and theoretical minimum-risk concentration) characterisation factors, respectively, at different background concentrations in air as working points. Exposure-response curve based on data from Apte et al. (2015) R.K. Rosenbaum et al.


indoor and outdoor environments, the archetypal approach capturing best the dominating differences between urban and rural areas and a number of other improvements (Fantke et al. 2015). Most LCIA characterisation methods addressing particulate matter formation follow the framework described in this section. There are some methods, however, that characterise impacts from particles as part of the ‘human toxicity’ impact category (e.g. CML 2002 and EDIP 2003), while most methods (including all methods developed after 2010) characterise human toxicity impacts from chemicals and impacts from particles as separate impact categories, mainly due to the differences in available data that allow using more refined models and less generic assumptions for the impact assessment of particle emissions. The most recent characterisation models—all damage-oriented—include work by van Zelm et al


The most recent characterisation models—all damage-oriented—include work by van Zelm et al. (2008) providing characterisation factors for primary and secondary PM10 for Europe based on a source receptor model, work by Gronlund et al. (2015) giving archetypal characterisation factors for primary PM2.5 and secondary PM2.5 precursors based on US data and work by van Zelm et al. (2016) proposing averaged primary and secondary PM2.5 characterisation factors for 56 world regions based on a global atmospheric transport model. However, none of the currently available approaches includes indoor sources, is able to distinguish emission situations at the city level or considers the non-linear nature of available exposure-response curves, which is why further research is needed for this impact category. For further details see Chap. 40 and Hauschild and Huijbregts (2015). 10.14 Land Use 10.14.1 Problem Land use refers to anthropogenic activities in a given soil area


For further details see Chap. 40 and Hauschild and Huijbregts (2015). 10.14 Land Use 10.14.1 Problem Land use refers to anthropogenic activities in a given soil area. Examples of land use are agricultural and forestry production, urban settlement and mineral extraction. The land use type in a specific area can be identified by the physical coverage of its surface, for example tomato crop grows in open-field orchards or under greenhouses, artificial surfaces with infrastructure are the expression of human settlements and open-pits are a sign of ore extraction. There is thus a direct link between land use and land cover, which is used to analyse land use dynamics and landscape change patterns. Soil is a finite resource, which contributes to the environmental consequences of its use. Soil loss actually occurs quantitatively with the average soil formation rate being extremely low compared to the soil depletion rate


Soil loss actually occurs quantitatively with the average soil formation rate being extremely low compared to the soil depletion rate. It also affects qualitative soil attributes, because degrading takes place via unsustainable management practices for the highest quality soils, which are those able to fulfil a greater diversity of purposes. As soil or land surface available at a given time is limited, land-use competition between resource users for occupying the same space often Life Cycle Impact Assessment


arises. This drives continuous changes in land uses. Croplands, pastures, urban areas and other land-use-intensive, human activities have expanded worldwide in the last decades at the expense of natural areas to satisfy our growing society’s needs for food, fibre, living space and transport infrastructure. Such changes transform the planet’s land surface and lead to large and often irreversible impacts on ecosystems and human quality of life (EEA 2010). For example, forest clearing contributes to climate change with the release of carbon from the soil to the atmosphere. The loss, fragmentation and modification of habitats lead to biodiversity decline. Land use change alters the hydrological cycle by river diversion and by modifying the portion of precipitation into runoff, infiltration and evapotranspiration flows (Foley et al. 2005)


Land use change alters the hydrological cycle by river diversion and by modifying the portion of precipitation into runoff, infiltration and evapotranspiration flows (Foley et al. 2005). After soil surface conversion, inappropriate management practices on human-dominated lands can also trigger a manifold of environmental effects on soil physical properties. In agricultural lands, mechanised farming can induce soil compaction, which affects aquifer recharge and the natural capacity of the soil to remove pollutants. Erosion is also a spread environmental concern of intensive agricultural practices. In urban and industrial areas, soil has been replaced by concrete surfaces and all its functions annulled. The Millennium Ecosystem Assessment (2005) provides a comprehensive description of how human land-use activities affect biodiversity and the delivery of ecological functions


The Millennium Ecosystem Assessment (2005) provides a comprehensive description of how human land-use activities affect biodiversity and the delivery of ecological functions. Some ecological effects of land use are: • Biodiversity decrease at the ecosystem, species and genetic levels • Impacts on local and regional climate regulation due to changes in land cover and albedo, e.g


tropical deforestation and desertification may locally reduce precipitation • Regional decline in food production per capita due to soil erosion and desertification, especially in dry lands • Rise in flood and drought risks through loss of wetlands, forests and mangroves • Change in the water cycle by river diversion and by greater appropriation of freshwater from rivers, lakes and aquifers to be used for irrigation of areas converted to agriculture To sum up, land-use activities (including land conversion and land use itself) cause noticeable damages on biodiversity and on the performance of soil to provide ecological functions as illustrated in Fig. 10.23. These ecological functions upon which human well-being depends are also referred to as ecosystem services (Millenium Ecosystem Assessment 2005), and together with biodiversity loss are the focus of the LCIA land-use impact category


10.14.2 Environmental Mechanism The LCIA land-use impact category covers a range of consequences of human land use, being a receptacle (or ‘bulk’) category for many impact indicators. It does not R.K. Rosenbaum et al.


assess nutrients, pesticides and any other types of emission to the ecosphere which are characterised by the corresponding emission-based impact category (e.g. eutrophication for emission of nutrients, ecotoxicity for emission of pesticides). Their inclusion in the land-use category would lead to double counting of the same impact. The general land-use environmental mechanism follows the model of Fig. 10.24. It shows the cause–effect chain from the elementary flow (i.e. land transformation or land occupation) to the endpoint damages on human health and ecosystems as well as available soil resources. Land transformation refers to the conversion from one state to another (also known as land use change, LUC) and land occupation to the use of a certain area for a particular purpose (also known as land use, LU)