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Background for spatial differentiation in LCA impact assessment - The EDIP2003 methodology 5 Aquatic eutrophication5.1 Introduction Authors: José Potting [29], Arthur H.W. Beusen [30], Henriette Øllgaard [31], 5.1 IntroductionAlthough still of good quality in some regions, European inland waters (fresh waters like rivers and lakes) as well as coastal seas are threatened by a multitude of human activities. The growing number of both users and uses [34] has increased the exploitative pressure on the European waters. This gives more and more conflicts between the various uses and users, and also between man and organisms living in or near those waters. A decline in quantity and quality of the water available for human uses, as well as of the ecological quality of aquatic ecosystems may be the result. A release of pollutants to water may have considerable local impact (in case of a point source), but its main impact seems its contribution to pollution accumulation by other sources upstream and downstream in rivers and finally in coastal waters. Water quality, and eutrophication or nutrient enrichment as part of that, has therewith become a regional topic. Kristensen and Hansen (1994) provide an interesting and comprehensive overview of the state of the environment in European rivers and lakes. EEA (1998) gives a short overview for both marine and inland waters. The focus of this chapter is on life cycle assessment of long-range nutrient enrichment of marine and inland waters. The presently typical way to assess eutrophication in life cycle assessment does not account for spatial variation in emissions, exposure and effect (Section 5.2). This chapter will explore the differences in eutrophying impact from nutrients released at different geographical locations and by different source categories. Section 5.4 describes the CARMEN-model, a model for integrated assessment of eutrophication, which is used to establish spatial resolved exposure factors. The CARMEN-model does not cover an assessment of ecological effect, but Section 5.3 gives a brief review on the issue. Section 5.5 describes how the CARMEN model is used to establish spatial resolved eutrophication factors. The results are presented in Section 5.6 and discussed in Section 5.7. Section 5.8 draws the main conclusions. 5.2 Typical life cycle impact assessmentThe eutrophying impact usually characterised in life cycle impact assessment relates implicitly to eutrophication of aquatic ecosystems. This follows from the modelling of impact that in life cycle assessment typically takes its bases in the composition of aquatic biomass (Wenzel et al. 1997, Heijungs et al. 1992, Lindfors et al. 1995): An increased input of nutrients into an aquatic ecosystem may lead to an increased production of biomass, primarily phytoplankton (see also Section 5.3). The typical composition of phytoplankton, as suggested by Redfield et al. (1993) [35] and commonly used in life cycle impact assessment, consists of carbon, nitrogen and phosphorus in a ratio of 106:16:1. This ratio suggest that 106 mole carbon, 16 mole nitrogen and 1 mole phosphorus is needed for the production of 1 mole of phytoplankton. In life cycle impact assessment of eutrophication, the Redfield ratio is used the other way round to express the contribution of emissions of nitrogen and/or phosphorus to biomass production in terms of the equivalent emission of a reference substance. The reference substance differs between methods, but these methods are further nearly similar. Table 5.1 lists equivalency factors as proposed by Wenzel et al. (1997). Table 5.1. Equivalency factors for nutrient enrichment from Wenzel et al. (1997).
Wenzel et al. (1997) suggest aggregation of the different nitrogen substances separate from those of phosphorus to allow site-considerations in the subsequent weighing of the accumulated substances (see Table 5.1 and Formula 5.1). This separate characterisation of phosphorus and nitrogen does anticipate inland freshwaters typically to be phosphorus limited while marine waters are typically limited by nitrogen. However, Wenzel et al. do also provide factors to add together nitrogen and phosphorus (see Table 5.1 and Formula 5.1): Wenzel et al. (1997) do not distinguish between differences in eutrophying impact from emissions to air, water and soil, and also not distinguish between eutrophication of terrestrial and aquatic ecosystems. Terrestrial ecosystems get eutrophied mainly by airborne emissions [36], [37], however, whereas waterborne nutrients from waste water and originating from agriculture dominate eutrophication of aquatic ecosystems (though stagnant water can be polluted by air depositions as well). The agricultural nutrient data in life cycle inventory usually refer to the amount after plant uptake and binding. This amount can leave the topsoil and may transport to surface water depending on environmental conditions (like soil type or net precipitation). Wenzel et al. (1997) does also not consider variation in environmental conditions. They suppose all types of emissions to fully contribute to eutrophication in general. Similar to Finnveden et al. (1992), we consider eutrophication of terrestrial and aquatic ecosystems as two separate impact categories. This chapter deals with eutrophication of aquatic ecosystems (Chapter 4 addresses terrestrial eutrophication). This chapter focuses on the geographical differences in transport of nutrients through the environment to inland waters and marine waters. The biological aspects of eutrophication will be described first, however. 5.3 EffectThe term "eutrophic" is a biological one and literally means "rich of nutrients", whereas the term "eutrophication" refers to "the process of becoming rich of nutrients" (Dikke van Dalen 1994, Kristensen and Hansen 1994) [38]. This section first elaborates on the nutrients relevant in eutrophication (Section 5.3.1), then describes the potential ecological effects from eutrophication (Section 5.3.2), and finally addresses which nutrients may limit biomass growth (Section 5.3.3). 5.3.1 Relevant nutrientsThe production of biomass in aquatic ecosystems is governed at any time and place by numerous factors simultaneously. However, the overall productivity under otherwise comparable conditions is largely determined by those factors that limit production over a substantial length of time during the main growth period. This long-term limitation is most often – though not always – due to one nutrient being poorly available in the given ecosystem itself and/or not present in sufficient quantities in natural supplies from outside. The concept of "the limiting nutrient" is essential when discussing nutrient enrichment or eutrophication. It states that one nutrient is limiting the growth of primary producers (plants) and thereby indirectly affects the ecosystem and that there is an excess of all other nutrients. An additional amount of the limiting nutrient will lead to increased growth, however, additional amounts of other nutrients will not since they are already in excess. (Finnveden and Potting 1999) Those nutrients most likely to act as limiting factors can be identified by ranking the amount and proportions of all critical elements in aquatic ecosystems against their concentration and proportion in the – unpolluted – water. Hydrogen, oxygen and most salts (like calcium, magnesium, potassium, sodium and chloride) are usually available in abundance and can therefore be eliminated. Carbon dioxide, (bi)carbonates and sulphate are generally available in excess compared to nitrogen and phosphorus in both freshwaters and marine waters. This makes nitrogen and phosphorus the primary candidates for "chronic" nutrient limitation [39]. Both nitrogen and phosphorus exist in many different chemical forms in the aquatic environment. Impact and bio-availability of each form may be different, but they have the ability to interchange. The transformations between the several forms are site-specific and continuous (known as the nitrogen and phosphorus cycle). All forms of nitrogen and phosphorus forms are therefore relevant, and total nitrogen and phosphorus are a good expression for long-term eutrophication. Trace elements may in rare cases be limiting. Primary producers as phytoplankton only need very small quantities of these elements (Hauschild 1998), however, and they are therefore not further addressed here. 5.4 Ecological effectsThe supply of nutrients to water is under natural circumstances in balance with the subsequent growth of biomass such that a stable though variable society of plant and animals is ascertained. Such society is
to some extent able to cope with variations in nutrient supplies. A too large nutrient increase pushes this stable society out of balance, however, and may through a chain of ecological effects provoke a shift
of the biological structure. The effects are generally more apparent in standing or slowly moving waters than in (small) fast flowing rivers (Kristensen and Hansen 1994). The chain of ecological effects is
illustrated in Figure 5.1:
Figure 5.1. Schematic representation of the ecological chain effects as a result of the increase with nutrients or eutrophication of lakes (reproduced from Kristensen and Hansen 1994). See below text for an explanation of the arrows. Standing and shallow clear waters with submerged plants become dominated by phytoplankton after a high nutrient input (→1). Increasing growth of phytoplankton makes the water turbid (→2). The turbidity prevents the light to reach the water-bottom, which makes the submerged plants disappear (→3). The community fish also changes because the predatory fish are unable to see and catch the smaller fish (→4). The fish community becomes dominated by zooplankton-eating species that are more tolerant to the turbidity (→5). Zooplankton eats phytoplankton and the decrease in zooplankton (→6) therefore results in a further increase of phytoplankton (→7). The excess phytoplankton dies and sinks to the bottom (→8). The decay of phytoplankton uses oxygen (→9). The decrease of oxygen leads to further fish kills and disappearance of bottom fauna (→10), and may also evoke a release of phosphorus from the lake bottom (→11). The released phosphorus may be used for a new increase of phytoplankton (→12) after which a new cycle of phytoplankton dying and sinking starts. Topography and the physical/chemical nature of the water influence the impact of nutrients on aquatic biomass growth. It is therefore difficult to define generic nutrient levels above which this growth may become problematic. Kristensen and Hansen (1994) give water quality indications for inland waters and EEA (1998) for marine waters (see Table 5.2). Table 5.2. Water quality categories abstracted from EEA (1998) and Kristensen and Hansen (1994). These levels have to be taken with caution since threshold values are influenced by topography and physical/chemical nature of the water. The nutrient levels refer to the extent of anthropogenic origin rather than to ecological effect.
Based on EEA (1998) Based on Kristensen and Hansen (1994) Drinking Water Directive (80/778/EEC) Kristensen and Hansen (1994) collected and analysed monitoring data about – amongst others – nitrogen and phosphorus concentrations from 800 stations in more than 550 European rivers and 1500 lakes. The annual mean concentrations in the majority of the rivers range from levels of moderate to heavy pollution. Only rivers in northern Europe seem hardly disturbed. Particularly rivers in western Europe show high nitrogen and phosphorus levels. The situation looks somewhat better for European lakes, though the same trends can be observed here. The water quality indications in Table 5.2 relate nutrient levels to the extent of anthropogenic origin rather than to the occurrence of ecological effects caused by biomass growth. Kristensen and Hansen (1994) also describe the biological state or rivers in a number of European countries. Their results are given in Table 5.3 and confirm the trends shown by the monitored data as described in the previous paragraph. EEA (1998) describe the water quality in European seas. They gives annual mean nitrogen and phosphorus concentrations at a number of sample points in west and north European seas and qualitative information about southern seas. Obviously, nutrient concentrations are high and water quality is often poor/bad close to estuaries of rivers (southern part of the North sea, Irish Channel, Waddenzee). The Mediterranean Sea is one of world's most oligotrophic (nutrient-poor) seas. However, eutrophication problems occur in semi-enclosed bays (mainly due to untreated sewage) and uncontrolled expansions of fish farming (Eastern Mediterranean). The northern and west coast of the Adriatic Sea, that receives the nutrient loads of the River Po, is the most endangered in the Mediterranean area. The long residence time of the water in the Black Sea makes this sea highly sensitive to eutrophication. The water quality in this sea and the Sea of Azov is bad, probably due to high inputs from the Danube, Dnieper and Dniester. Table 5.3. Percentage of river reaches in various European countries classified as being of good, fair, poor or bad quality. Of good quality are river reaches with nutrient-poor water, low levels of organic matter, saturated with dissolved oxygen, rich invertebrate fauna, and suitable spawning ground for salmonid fish. River reaches with moderate organic pollution and nutrient content, good oxygen conditions, rich flora and fauna, large fish population are classified as fair. Poor quality river reaches have heavy organic pollution, usually low oxygen concentrations, locally anaerobic sediment, occasional mass occurrence of organism insensitive to oxygen depletion, small or absent fish population, periodic fish kill. Of bad quality are those rivers with excessive organic pollution, prolonged periods of very low oxygen concentration or total deoxygenation, anaerobic sediment, severe toxic input, devoid of fish. (Kristensen and Hansen 1994)
5.4.1 Concept of the limiting nutrientThe concept of "the limiting nutrient" is essential when discussing nutrient enrichment or eutrophication. Generally seawaters are regarded to be "chronic" limited by nitrogen, whereas lakes and larger, slowly flowing rivers are limited by phosphorus. Estuaries can be limited both by nitrogen and phosphorus, whereas small streams and mountainous lakes are usually not limited by nitrogen and phosphorus, but by factors as light and temperature (Miljøstyrelsen 1991, Blau and Seneviratne 1995). The concept of "the limiting nutrient" is a simplification because the limiting nutrient may change over seasons, and also over the years due to earlier loading. (Kristensen and Hansen 1994, Finnveden and Potting 2000) All nutrients can therefore potentially be used for biomass growth, though this maximum will for nitrogen usually not be the case in inland waters and for phosphorus not in marine waters. 5.5 The CARMEN modelCARMEN is an acronym for Cause effect Relation Model to support Environmental Negotiations. It is an integrated assessment model to analyse and evaluate strategies to reduce nutrient loading of inland waters [40] and coastal seas in Europe. The model does actually not contain an assessment of ecological effects, but calculates the change in nutrient loads in ground water, inland waters (river catchments) and coastal seas from changes in nutrient emissions and supplies (i.e. the causes). CARMEN, version 1.0 is used in this study. The causes modelled by CARMEN are atmospheric deposition of nitrogen on soil and coastal seas, phosphorus and nitrogen supply to agricultural soils, and phosphorus and nitrogen in municipal wastewater (see Figure 5.2 and Section 5.4.1). CARMEN models the transport of nutrient by rivers to sea relatively straightforward, but the transport of nutrients from agricultural supply and atmospheric deposition through groundwater drainage and surface runoff, to surface water is modelled spatially resolved over 124320 grid-elements of 10*10 minutes (roughly 100-250km2, depending on the longitude and latitude location of the grid-element). (see Section 5.4.2) Meinardi et al. (1994a,b), Klepper et al. (1995) and Beusen (not published) describe the CARMEN model in detail, but a short overview is given here about estimates of emissions (Section 5.4.1) and transport via groundwater drainage and surface runoff (5.4.2) of nutrients as modelled in CARMEN. 5.5.1 Emission estimatesMain sources for nitrogen in the aquatic environment are – via groundwater drainage and surface runoff – the use of fertiliser and manure in agriculture (diffuse sources), and municipal wastewater and – too a lesser extent – industrial effluent (mainly point sources). Deposition of airborne nitrogen is of some relevance for marine waters, but has usually minor importance for inland waters (except for the indirect contribution via groundwater drainage and surface runoff). Natural nitrogen can be important in regions with low population density. (EEA 1998, Kristensen and Hansen 1994) Figure 5.3 gives an overview of the relative importance of the different sources. Most of the phosphorus loading of inland waters is attributable to discharge from point sources, especially municipal wastewater and – too a lesser extent – industrial effluent. The application of phosphorus in agriculture may also contribute to inland waters through topsoil erosion. Phosphorus will usually not reach the ground water due to strong adsorption to the soil, though it exceptionally occurs after excessive loading on very poor soils (EEA 1998, Klepper et al. 1995, Kristensen and Hansen 1994)
Figure 5.2. Main sources for nitrogen (continuous arrow) and phosphorus (dashed arrow) to soil, groundwater, surface waters and coastal seas addressed in the CARMEN model (Beusen not published). CARMEN models three main sources for nitrogen and phosphorus to surface water (see also Figure 5.2): agriculture, municipal wastewater and atmospheric deposition (only for nitrogen). The several nitrogen and phosphorus supplies have been allocated to each grid-element on the basis of the distribution of land uses in the given grid-element (arable land, grassland, permanent crops, forest, urban area, inland waters, others). 5.5.2 Fate and exposure assessmentThe CARMEN model distinguishes between sources directly and indirectly via inland waters releasing to coastal seas. Oxygen depletion has been a severe problem for several years, but implementation of biological treatment of domestic and industrial wastewater has resulted in that many rivers and lakes are now fairly well oxygenated. However, low oxygen content is a usual phenomenon in the first stretches downstream of point sources for wastewater treatment. Under low oxygen conditions, nitrate is decomposed into nitrogen gas and oxygen used for the biomass decomposition. The nitrogen gas is released to the atmosphere. Removal of nitrogen is modelled for all waters the same in CARMEN by assuming a generic removal of 30% of the nitrogen input (i.e., 70% transports to sea). Figure 5.3. (Figure 9.11 and 9.14 from EEA 1998, p201-202) Phosphorus on its turn may be used for biomass production or temporarily adsorbed in the phosphorus pool in the bottom sediments of lakes and rivers. All phosphorus will remain in the water, however, and in the end transport to marine waters. CARMEN therefore considers no removal of phosphorus (i.e. all of it transports to sea). The strength of CARMEN is a detailed spatial resolved modelling of the transport of nutrients from agricultural supply and atmospheric deposition – through groundwater drainage and surface runoff or topsoil erosion – to surface water. The water flow is the main transport mechanism that brings nutrients – through deep groundwater drainage (nitrogen), runoff (nitrogen) or topsoil erosion (phosphorus) – to surface water. The net precipitation determines the amount of water draining to either groundwater or by surface runoff. The ratio between deep groundwater recharge and surface runoff resulting from net precipitation is determined in CARMEN by: Aquifer type (permeability of subsurface) The flow of deep groundwater to surface water is in addition determined by: Nitrogen transport through deep groundwater and surface runoff to surface water further depends on the amount leaching to the soil water at shallow depth after plant uptake and binding in the topsoil matrix. The share of nitrogen leaching is fixed for arable land by soil type and for agriculturally used grasslands, it is increasing with nitrogen dose for all soils (see also Section 5.7.2.). Phosphorus concentrations in groundwater are usually negligible because of the strong adsorption to the soil. Phosphorus transport to surface water is dominated by erosion of sediment from the topsoil by surface water runoff. (Meinardi et al. 1994/1995) Each of the above factors are modelled by CARMEN spatially resolved over 124320 grid-elements of 10*10 minutes (roughly 100-250km2, depending on the longitude and latitude location of the grid-elements). 5.6 Calculation procedureSite-dependent eutrophication factors have been established with help of the CARMEN model for 32 European countries. The factors relate the amount of nutrient released in a given country to its share to eutrophication of European inland waters [41] and coastal seas (i.e., respectively the share of nutrient indirect to seas through inland waters and the share of nutrient indirectly and directly to seas). The factors quantify the maximum biomass growth these nutrients may contribute to in the receiving water. The factors do not anticipate the deterioration of the water quality as a result of this biomass growth. The eutrophication factors per country were calculated by changing the total amount of either nitrogen or phosphorus from a given source category in one country (other emissions for all countries and other source categories remaining the same). Agricultural supplies of manure or fertiliser to soil, and wastewater releases have been considered as sources for both inland waters and seas. The calculations for coastal seas in addition also addressed atmospheric deposition as a nitrogen source. For each source category, the change in eutrophying loads is calculated spatially resolved over 101 river catchments and 32 coastal seas (see Annex 5.1 and 5.2). Next, the loading increases by this change of one country are accumulated over all river catchments and seas to obtain the factors expressing the share eutrophying respectively inland waters and seas (in kg per kg released). CARMEN models the contributions from nitrogen depositions to eutrophication of inland waters and seas from the actual deposition pattern in Europe. It does not contain the relationships between country emissions and depositions. Huijbregts and Seppälä (2000) recently published data about the nitrogen deposited on European seas as a ratio of the emission in their country of release. These data have been combined with calculations with the CARMEN model to arrive at the loading of separate seas. The data from Huibregts and Seppälä do not consider deposition outside the EMEP model domain, and have therefore been corrected with data from Barrett and Berge (1996) about the percentages exported outside the EMEP model domain. The CARMEN model does not include an assessment whether nutrient loading actually results into biomass growth and what effect this has on the ecological quality of the water. The calculated factors represent the potentially maximum contribution to biomass growth. They are exposure factors rather than expressing ecological effects. 5.7 ResultsThe results of the calculations are abstracted in Table 5.4. Columns 2 to 7 in Table 5.4 give the share of nutrient from a given source category that contributes to eutrophication of inland waters. The last eight columns do the same for marine waters. Table 5.4. The share of nitrogen and phosphorus eutrophying inland waters (Column 2 to 7) and marine waters (Column 8 to 15) for different source categories in different countries (Fert.=fertiliser, Man.=manure, WW=wastewater, NH3=airborne NH3 , NOx=airborne NOx). Hence that the factors for fertiliser and manure relate to nutrient after plant uptake (see Section 5.7.2 for how to arrive at the proper inventory data). There are a number of countries where eutrophication of marine waters by agriculture and wastewater is much smaller than their eutrophication of inland waters (underlined). These numbers relate to countries releasing on seas outside the model domain (they may be replaces with factors for inland waters as best estimate and by a factor 0.7 for nitrogen and a factor 1.0 for phosphor in wastewater). The out-crossed numbers are concluded to be wrong. For Denmark, the factor for marine waters may be lowered to 0.7 for nitrogen and to 1.0 in phosphor in wastewater. The other numbers being striketrough should be replaced by the site-generic factors. Both underlined and out-crossed numbers are excluded from the mean etc.
Hence the factors relating to application of fertiliser and manure refer to the amount of nutrient after plant uptake. This amount is available to leave the topsoil. It is a peculiarity of life cycle assessment to consider the topsoil as part of the economic system or technosphere and the data from inventory analysis should reflect the outputs from the technosphere to the ecosphere. One should thus be careful to apply the factors in Table 5.4 to the proper inventory data (see also Section 5.7.2). The factors in Table 5.4 show eutrophication from fertiliser and manure basically to be similar. There is therefore no reason to employ different factors for these source categories in life cycle assessment. There is neither a difference in the mean factors for agricultural eutrophication of inland waters and marine waters. The eutrophication factors from nitrogen [42] show notable differences for a number of individual countries (i.e. the Netherlands, the Scandinavian countries, the United Kingdom). The main cause is that along the relative long coastline, small water systems are discharging directly to marine waters. Table 5.4 shows the Swiss and Austrian factors for inland waters to deviate from the one for marine waters, while both countries do not have a direct connection to sea. The reason for this deviation is not clear. There are also a number of countries where agricultural eutrophication of seas seems to be much smaller than their eutrophication of inland waters (i.e. Russia, Spain, Portugal, Iceland and Ireland). However, some rivers in these countries release on marine waters that are outside the model domain of CARMEN. The sea factors may here be corrected by replacing them for the inland water factors (which are expected to be a fair estimate). Similar as for nitrogen from agriculture, differences exist between the factors for eutrophication of coastal seas and inland waters by wastewater. These factors express basically the share of nutrient (wastewater) discharged indirectly via inland waters or indirectly plus directly to marine waters. The factors for marine waters and inland waters can be at maximum 0.7 for nitrogen and 1.0 for phosphorus. Hence the Icelandic factors for inland waters are larger then these maximum values and thus probably false. The same applies for Denmark's factors for seas. Small exceedances (like 0.01) of the maximum values (0.7 or 1.00) are expected to be rounding errors. Within source categories, the differences between countries are relatively small. There is a factor 3 for agricultural nitrogen and a factor 7 for phosphorus between lowest and highest factor. The nitrogen factors for the Netherlands and Denmark (both having intensive agriculture) are remarkable low. This is due to the fact that the groundwater recharge of both countries is relative large compared to surface runoff (ratio larger than one). This means that the main part of the nutrients will flow into the groundwater, and due to a time delay of more than 10 years and a mixing with unpolluted groundwater, only a small part of the nutrients will flow out, to the inland waters. The phosphorus factors for Austria and Switzerland are remarkable high. The slope seems here to have a considerable impact on surface runoff and erosion. 5.8 DiscussionSome remarks about the eutrophication factors have already been made in the previous section and will not be repeated here. However, a few additional aspects are discussed in more detail in the following. 5.8.1 Application of the eutrophication factorsThe application of the eutrophication factors in Table 5.5 is very simple. An emission in the product's life cycle is multiplied with the eutrophication factor for that region and substance to derive the estimated eutrophication of that emission. The only additional data required, the country where an emission takes place [43], are in general already provided by current LCI.
The receiving waters can be inland waters on the one hand, and marine waters on the other hand. Both should be regarded in life cycle assessment as subcategories and thus reported separately (i.e. not to aggregate them). It is recommended to refrain from aggregating phosphorus and nitrogen, since nitrogen may at some point leave the aquatic system by nitrification, whereas phosphorus does basically not leave and remain to be available for potential eutrophication. This leads thus all together to two subcategories (inland waters and marine waters) with each two impact indicators (aggregated phosphorus and aggregated nitrogen). The Guidance Document of Hauschild and Potting (2003) based on this Technical Report gives a detailed description of the application procedure for the aquatic eutrophication factors. 5.8.2 Proper inventory dataIt is common practice in life cycle assessment to consider the topsoil of agricultural soils as part of the technosphere. The data in life cycle inventory for nutrient supply in agriculture therefore usually refer to the amount of nutrient available for leaving the topsoil after plant uptake and binding. (Weidema and Meusen 2000). The factors from Table 5.4 also refer to the amount of nutrient available for leaving the topsoil after plant uptake and binding. The eutrophication factors from Table 5.5 can therefore directly be multiplied with that type of inventory data. There is also a number of practitioners who do not consider the topsoil as part of the technosphere or lack data for implementing that, and stop their inventory at the level of nutrient supply. Inventory data reflecting nutrient supplies should first be corrected for plant uptake and binding before applying the eutrophication factors from Table 5.4. The factors in Table 5.5 are based on the relationships employed by CARMEN and can be used for this correction: Table 5.5. The factors expressing the relation between nutrient application and nutrients available for drainage by groundwater and surface runoff or erosion. The factors in this table can be multiplied with the factors in Table 5.4 to obtain the relation between nutrient application and eutrophication of inland waters and coastal seas.
According to Table 5.5, less then 25% of the nitrogen application and 10% of the phosphorus application leaches after plant uptake to soil water at shallow depth and is then available for transport through deep groundwater and surface runoff or erosion. This means that on average less then 12% of the supplied nitrogen and 5% of the supplied phosphorus will end up to be available for biomass growth in inland waters and finally coastal seas. This is in accordance with values given in the literature. (Weidema and Meeusen 2000) A similar situation is at stake for the phosphorus and nitrogen emissions to water. The waterborne nitrogen and phosphorus emissions in life cycle inventory usually refer to the content in effluent after wastewater treatment (unless emissions are released actually directly to surface water). The factors from Table 5.4 refer to releases directly to surface water (after potential purification thus). The eutrophication factors from Table 5.5 can therefore directly be multiplied with this type of inventory data. In case the inventory data do refer to wastewater before purification, the factors in Table 5.6 can be used in connection with Table 5.4 to correct the inventory data. Table 5.6. The expected concentrations (in mg/l) and removal efficiencies (in g/g) for nitrogen and phosphorus in waste after different types of waste water treatment (Hansen, 2000).
5.8.3 Biological oxygen demandDecomposition of organic material used oxygen that is measured as chemical or biological oxygen demand (COD or BOD). Fish kill and disppearance of bottom fauna may occur after prolonged situations of oxygen depletion. This may be the result of prolonged and high nutrient loading. Some methods presently typical for life cycle assessment of aquatic eutrophication therefore suggest considering organic materials, as part of this impact category (Lindfors et al. 1995, Heijungs et al. 1992). This suggestion is not followed here, since organic material often contains nitrogen or phosphorus and is then reflected already in the inventory table as nitrogen and/or phosphorus emission. To also add organic materials to this impact category, would then lead to double counting. Furthermore, organic material is not essentiel for growth of the biomass. Emissions of organic material, which causes BOD/COD, may increase or strengthen the effects and the environmental impacts of nutrient enrichment, especially with regards to oxygen depletion. However, these environmental impact of BOD/COD substances are comparable to the secondary effects of nutrient enrichment. 5.8.4 Related research activitiesDuring and after the work for this chapter, similar research activities were performed by other groups. Huijbregts and Seppälä (2000) used spatially resolved source-receptor matrices to calculate the share of atmospheric emission from source countries that deposits on marine waters. The rest deposits on land, but can transport through deep groundwater and surface runoff to surface water and marine water. The depositions of Huijbregts and Seppälä have been gratefully used here to calculate site-dependent factors that cover the direct and indirect contribution from the atmospheric emissions in all European countries to eutrophication of surface and marine waters. In addition to Huijbregts and Seppälä (2000), Huijbregts and Seppälä (2001) established characterisation factors for aquatic eutrophication, covering fate and potential effect for the Netherlands, Europe and the world. The fate part of these factors seem – at a first sight – less sophisticated than what is proposed in this chapter, whereas the effect factor is similar to what is proposed here. A closer comparison is recommended, however. Bare et al. (2003) calculated similar factors using a similar type of modelling as in this chapter for the federal states in North America. Due to the advanced state of the reporting of the Danish project, it was unfortunately not possible to make a closer comparison here. 5.9 ConclusionsEutrophication factors have been established for 32 countries in Europe to facilitate site-dependent assessment of eutrophication in life cycle assessment (see Table 5.5). The factors relate the country of emission to the potential maximum contribution to biomass growth in the receiving waters. The potential maximum will usually not be the actual contribution. Generally however, seawaters are regarded to be "chronic" limited by nitrogen, whereas lakes and larger, slowly flowing rivers are limited by phosphorus. Interpreting the results from impact assessment can use this and it is therefore recommended to refrain from aggregating nitrogen and phosphorus. Receiving waters can be inland waters and coastal seas (directly or indirect by river discharge to seas). Inland waters and coastal waters are recommended as separate subcategories (aggregating would lead to double counting). The established eutrophication factors are exposure factors rather than expressing the ecological effect of eutrophication and biomass growth. The present state-of-the art in integrated assessment modelling of aquatic eutrophication does not yet allow such effect assessment. The spatial resolved factors in Annex 5.3 to 5.9 may be used for a tentative interpretation in this direction. 5.10 ReferencesBare J.C., Norris G.A., Pennington D.W., McKone T. TRACI: The US EPA's tool for the reduction and assessment of chemical and other environmental impacts. Journal of Industrial Ecology., Vol. 6 (2003), Issue 3/4, pp49-78. Barrett, K. and E. Berge (eds.). Transboundary air pollution in Europe. Part 1: Estimated dispersion of acidifying agents and of near surface ozone. EMEP MSC-W status report (research report no. 321996). Oslo (Norway), Norwegian Meteorological Institute, 1996. Berdowski, J.J.M. and W.J. Jonker. Emissions in the Netherlands. Industrial sectors, regions and individual substances (1992 and estimates for 1993; Publication series emission registration no. 21; in Dutch). The Hague (the Netherlands), Ministry of Housing, Spatial Planning and Environment, 1994. Beusen, A. User manual of CARMEN1 (Draft). Bilthoven (the Netherlands), National Institute of Public Health and Environmental Protection (RIVM), not published. Blau, S. and S. Seneviratne. Acidifiction and eutrophication in life cycle assessments (LCAs). Zürich (Switzerland), Swiss Federal Institute of Technology Zurich, 1995. EEA Europe's environment: The second assessment. Copenhagen (Denmark), European Environmental Agency,1998. Finnveden, G., Y. Andersson-Sköld, M.-O. Samuelsson, L. Zetterberg and L.-G Lindfors. Classification (impact analysis) in connection with life cycle assessment. In: Product life cycle assessment. Principles and methodology (Nord 1992:2). Copenhagen (Denmark), Nordic Council of Ministers, 1992, pp172-231. Finnveden, G. and J. Potting. Eutrophication as an impact category. Int. Journal of Life Cycle Assessment, Vol, 4 (1999), Issue 6, pp311-314. Hansen, O.C: Ole Christian Hansen, unpublished work, 2000. Heijungs, R., J. Guinée, g. Huppes, R.M. Lankreijer, H.A. Udo de Haes, A. Wegener Sleeswijk, A.M.M. Ansems, P.G. Eggels, R. Van Duin en H.P. de Goede. Environmental life cycle assessment of products. Guide and background (ISBN 90-5191-064-9). Leiden (the Netherlands), Centre of Environmental Science of Leiden Univerisity, 1992 Huijbregts, M. and J. Seppälä. Towards region-specific, European fate factors for airborne nitrogen compounds causing aquatic eutrophication. Int. Journal of Life Cycle Assessment, Vol. 5, Issue 2, pp65-67, 2000. Huijbregts, M. and J. Seppälä. Life cycle impact assessment of pollutants causing aquatic eutrophication. Int. Journal of Life Cycle Assessment, Vol. 6, pp339-344, 2001. Jaarsveld, J.A. van, and D. Onderdelinden. TREND: An analtical long-term deposition model for multi-scale purposes (report no. 228603009). Bilthoven (the Netherlands), National Institute of Public Health and Environmental Protection (RIVM), 1990. Klepper, O., A.H.W. Beusen and C.R. Meinardi. Modelling the flow of nitrogen and phosphor in Europe: from loads to coastal seas (RIVM report 461501004). Bilthoven (the Netherlands), National Institute of Public Health and Environmental Protection (RIVM), 1995. Kristensen, P. and H. O. Hansen. European rivers and lakes. Assessment of their environmental state (EEA environmental monographs 1). Copenhagen (Denmark), European Environmental Agency, 1994. Lindfors, L-G, K. Christiansen, L. Hoffman, Y. Virtanen, V. Juntilla, O-J. Hanssen, A. Rønning, T. Ekval and G. Finnveden. Nordic Guidelines on life cycle assessment (Nord 1995; 20). Copenhagen (Denmark), Nordic Council of Ministers, 1995. Meinardi, C.R., A.H.W. Beusen, M.S.J. Bollen and O. Klepper. Vulnerability to diffuse pollution of European soils and groundwater (RIVM report 461501002). Bilthoven (the Netherlands), National Institute of Public Health and Environmental Protection (RIVM), 1994a. Meinardi, C.R., A.H.W. Beusen, O. Klepper and W.J. Willems. Nitrate contamination of European soils and groundwater (RIVM report 461501003). Bilthoven (the Netherlands), National Institute of Public Health and Environmental Protection (RIVM), 1994b. Miljøstyrelsen. Environmental impacts of nutrient emissions in Denmark. Redegørelse fra Miljøstyrelsesn, no. 1). Danish Environmental Protection Agency, Copenhagen, 1991. Paaby H, Møller F, Skop E, Jensen JJ, Hasler B Bruun H, Asman WAH. Omkostninger ved reduktion af næringsstofbelastningen af havområderne – Metode, model, analyse. Faglig rapport fra DMU, nr. 165. Danmarks Miljøundersøgelser, 1996. Redfield, A.C., B.H. Ketchum and F.A. Richards. The influence of organisms on the composition of sea water. In: Proceedings of the 2nd International Water Pollution Conference in Tokyo. Oxford (United Kingdom), Permagon Press, 1993, pp215-243. Samuelsson, M.-O. Life cycle assessment and eutrophication. A concept for calculation of the potential effects of nitrogen and phosphorus. Stockholm (Sweden), Swedish Environmental Research Institute (IVL), 1993. Weidema, B.P. and M.J.G. Meeusen (eds.) Agricultural data for life cycle assessment. Volume 1 and 2. The Hague (the Netherlands), Agricultural Economics Research Institute (LEI), 2000. Wenzel, H., M.Z. Hauschild and L. Alting: Environmental assessment of products. Vol. 1 - Methodology, tools, techniques and case studies, 544 pp. Chapman & Hall, United Kingdom, 1997, Kluwer Academic Publishers, Hingham, MA. USA. ISBN 0 412 80800 5. Wither, P.J.A. Cycling and sources of phosphorus in agricultural systems and to the wider environment: a UK perspective. In: Weidema, B.P. and M.J.G. Meusen (eds.) Agricultural data for life cycle assessment. Volume 1 and 2. The Hague (the Netherlands), Agricultural Economics Research Institute (LEI), 2000, pp13-24. Annex 5.1: The river catchments modelled in CARMEN
Annex 5.2: The coastal seas modelled in CARMEN
Footnotes [29] Institute of Product Development (IPU) in Denmark until 2000, presently at the Center for Energy and Environmental Studies IVEM, University of Groningen [30] The Dutch National Institute of Public Health and Environmental Protection (RIVM) [31] The Danish Technological Institute (DTI) [32] The Dutch National Institute of Public Health and Environmental Protection (RIVM) [33] Institute of Product Development (IPU) in Denmark [34] Main uses being for public water supply, irrigation, industrial processes and cooling, hydroelectric power generation (inland freshwaters), and transport and waste disposal (both coastal and inland freshwaters). [35] Samuelsson lists also other ratios for carbon, nitrogen and phosphor reported in literature. This list shows considerable differences in nitrogen and phosphor ratios, the lowest being 5:1 and the highest being 19:1. However, those ratios are based on samples also containing organic matter originating from other organisms than phytoplankton. Therefore, the Redfield ratio is followed here since it is suggested to be valid for phytoplankton. An alternative might be, as recently suggested by Seppälä (2000), to employ a range of 13 to 19 for nitrogen versus 1 for phosphor around the Redfield ratio. This is not further elaborated here. [36] Terrestrial ecosystems are usually not limited by phosphor, and airborne phosphor has no role worth mentioning in eutrophication of surface waters. For a comparison: Emissions of phosphor to air are approximately 130 kton per year, while phosphor emissions to water exceed 5,000 kton per year in the Netherlands (Berdowski and Jonker 1994). The atmospheric emission is thus less than 3 percent of the waterborne emission, while only a minor part of that atmospheric phosphor will through topsoil erosion reach surface water. [37] Similar to Wenzel et al. (1997), Finnveden et al. (1992) also propose to separately aggregate nitrogen and phosphor. They suggest in addition to report airborne and waterborne nitrogen both separately (maximum scenario for terrestrial eutrophication) as well as summed together (representing a maximum scenario for aquatic eutrophication). All together, Finnveden et al. (1992) distinguish five subcategories: Aggregation of airborne nitrogen emissions (terrestrial eutrophication), aggregation of waterborne phosphor emissions (and organic material), aggregation of waterborne nitrogen emissions (and organic material), aggregation of nitrogen emissions (and organic material) to water and airborne nitrogen emissions, aggregation of all phosphor and nitrogen emissions (and organic material). [38] Eutrophication and nutrient enrichment may thus be taken as synonyms. Other terms sometimes used are "nutrifcation" and "oxygen depletion" and refer to the same group of impacts as covered by eutrophication. The term eutrophication is chosen here and refers in this chapter to aquatic eutrophication only (see Chapter 4 for terrestrial eutrophication). [39] Whereas nature is the predominant source for other nutrients, supply of nitrogen and phosphor largely originates from anthropocentric sources. Therefore, these nutrients are the only ones being amenable to control. This is an additional reason for those nutrients to be important for "chronic" nutrient limitation. [40] CARMEN does not explicitly address eutrophication of lakes, but lakes will usually be part of a river catchment and are thus implicitly covered by CARMEN. [41] The eutrophication factors calculated for rivers are therefore assumed to cover all inland waters (both rivers and lakes thus). This is defensible because the factors represent maximum potentially biomass growth. The maximum potential biomass growth will be larger than the actual expected occurrence of biomass growth both in rivers and lakes. [42] Similar differences probably exist for phosphor, but this is hardly visible because of the small shares. [43] If the emission is known for actually being released to either inland waters or to a coastal sea, this information should of course be used in stead of the factors from Table 5.5. Hence the emission of nitrogen should be multiplied with a factor 0.7 to account for nitrogen leaving the system after denitification.
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