Impacts of European livestock production: nitrogen, sulphur, phosphorus and greenhouse gas emissions, land-use, water eutrophication and biodiversity

Livestock production systems currently occupy around 28% of the land surface of the European Union (equivalent to 65% of the agricultural land). In conjunction with other human activities, livestock production systems affect water, air and soil quality, global climate and biodiversity, altering the biogeochemical cycles of nitrogen, phosphorus and carbon. Here, we quantify the contribution of European livestock production to these major impacts. For each environmental effect, the contribution of livestock is expressed as shares of the emitted compounds and land used, as compared to the whole agricultural sector. The results show that the livestock sector contributes significantly to agricultural environmental impacts. This contribution is 78% for terrestrial biodiversity loss, 80% for soil acidification and air pollution (ammonia and nitrogen oxides emissions), 81% for global warming, and 73% for water pollution (both N and P). The agriculture sector itself is one of the major contributors to these environmental impacts, ranging between 12% for global warming and 59% for N water quality impact. Significant progress in mitigating these environmental impacts in Europe will only be possible through a combination of technological measures reducing livestock emissions, improved food choices and reduced food waste of European citizens.


Introduction
Nowadays agricultural land occupies about 180 million hectares or 42% of the land area of the European Union 10 from which a great portion is used as grassland and for cultivating feed (FAO 2006). Historically, livestock helped to transform inedible materials (grass and waste) into high quality food. However today livestock production systems affect air quality, global climate, soil quality, biodiversity and water quality (Sutton et al 2011b(Sutton et al , 2011c, by altering the biogeochemical cycles of nitrogen, phosphorus and carbon. In particular, reactive nitrogen (N r ) plays a key role in several environmental impacts (N r represents all forms of nitrogen other than N 2 , including ammonia (NH 3 ), nitrogen oxides (NO x ), nitrous oxide (N 2 O), and N losses to water bodies). Nitrogen cascades or recycles through crop and livestock production systems, in form of feed for livestock and of manure to grow crops, as illustrated in figure 1, leading to several un-intended flows that give rise to environmental concerns (Sutton et al 2011a.
The emissions from the livestock sector contribute to five major environmental impacts: (i) The emissions of ammonia (NH 3 ) and nitrogen oxides (NO x =NO+NO 2 ) contribute to the formation of secondary particulate matter (PM) and tropospheric ozone, both with serious impacts on air quality. Across Europe, ammonium in particles may account for 5-15% of total PM 2.5 (Putaud et al 2010). Loss of statistical life expectancy due to exposure to PM 2.5 is estimated at 6-12 months for large parts in Europe (Amann et al 2011).
(ii) Emissions from the livestock sector affect radiative forcing in many ways; long-lived greenhouse gases (GHGs) occur as methane (CH 4 ), nitrous oxide (N 2 O) and carbon dioxide (CO 2 ) from the use of fossil fuels or through land use and land use change. Emissions of NO x contribute to tropospheric ozone formation-an important greenhouse gas in its own right-which reduces carbon sequestration through damage to vegetation (Ainsworth et al 2012, Simpson et al 2014. Conversely, aerosols produced by NO x -driven photochemistry also affect climate, generally having a cooling effect (Simpson et  Estimates of the share of GHG emissions from land use changes to total livestock emissions range between 9%-35% (FAO 2006, Lesschen et al 2011, Weiss and Leip 2012.
(iii) Terrestrial biodiversity is affected by livestock production through land use (including historic land use changes), ammonia emissions and consequent deposition and climate change. Intensively managed grassland and arable land used to grow livestock feeds have a low biodiversity, while extensive grazing avoids shrub encroachment or reforestation and helps maintain landscapes of high biodiversity. Habitat changes and land fragmentation can lead to truncation of migratory routes or the replacement of native species with invasive ones (Reid et al 2010). N r deposition reduces species richness through eutrophication, acidification, direct foliar impacts, and exacerbation of other stresses (Dise et al 2011, Bobbink et al 2010. (iv) Finally, livestock production has a role in the deterioration of the quality of freshwater and coastal water, increasing losses of N and phosphorus (P) to the water system. There is evidence that concentrations of 25 mg L −1 nitrate (NO 3 − ) in drinking water are related to an increase of incidences of colon cancer by about 3% (van Grinsven et al 2010). The stoichiometric excess of nitrogen and phosphorus in coastal water with respect to silica (Si), can enhance water eutrophication (Billen and Garnier 2007, Grizzetti et al 2011, Voß et al 2011. In this study we quantified the contribution of livestock production systems to the above mentioned five impacts combining new data based on a cradle-togate Life Cycle Assessment (LCA) calculation with results of other studies on emissions at the European scale. The results are discussed with respect to total emissions in the EU27. Although several compounds are included in the analysis, this paper will give particular attention to reactive nitrogen flows in agricultural production systems.

Overall approach
We estimated emissions of CO 2 , CH 4 , N 2 O, NH 3 , NO x , SO 2 , and N and P losses to the hydrosphere related to livestock production in Europe and interpreted the results in the light of the five major environmental threats. This was done by (i) extending the LCA approach in the agro-economic Common Agricultural Policy Regionalised Impact (CAPRI) modelling system (Britz and Witzke 2012) 11 , developed for the assessment of GHG emissions and N footprints, to provide cradle-to-farm gate LCA data for reactive nitrogen, in combination with (ii) model results for other compounds (P and SO 2 ) to provide a comprehensive picture of the environmental impact of EU livestock production. Table 1 gives an overview of emission sources quantified indicating how they are linked to livestock production systems. As the table indicates, emissions caused by livestock rearing were calculated for EU27 only. Emissions from imported livestock products were not included in the analysis, while further emissions associated with cultivation, processing and transport of (non-forage) feed might occur within or outside the EU.
The CAPRI N-LCA approach The emissions of GHGs (N 2 O and CH 4 ) and reactive nitrogen (N r ) to air (NH 3 , NO x ) and water (N leaching and runoff) and other N flows (such as trade of food and feed) were estimated with CAPRI, using the cradle-to-gate LCA approach implemented for the estimation of GHG emissions (Weiss and Leip 2012) and further extended to include the estimation of N footprints per product group (Leip et al 2014b). Here we present new data that have been calculated when further extending the model with flows of all N r . All new data presented here include emissions occurring within Europe and outside the EU territory for imported feed and/or land use change (LUC; LUC emissions were calculated using the land use transition probabilities of scenario II in Weiss and Leip (2012), table A1). Also included are credits for carbon sequestration in managed grassland as well as emissions due to foregone carbon calculation in croplands according to Weiss and Leip (2012). To analyze international N flows to and from EU27, the detailed trade data matrix of the trade module in the FAOSTAT database (FAOSTAT 2014) was used taking into account the N content to the 504 commodities involved (vegetal, animal and fiber products) (Lassaletta et al 2014). Internal flows among EU 27 countries were calculated and subtracted. The net N import to EU27 for each world country was estimated by the difference between total imports and exports. A general description of the CAPRI model and its relevant modules can be found in Britz et al (2010), Jansson and Heckelei (2011), Leip et al (2011bLeip et al ( , 2011d, Britz and Wizke (2012) and Perez-Dominguez et al (2012).
In the CAPRI LCA module (Weiss and Leip 2012, Leip et al 2014b, total agricultural emissions E agri were estimated as the sum of flows caused by agricultural production activities, plus emissions caused in earlier phases of the products life cycle, including energy use or land use change. Supplementary Information S1 gives detailed results of the N-LCA for the main six vegetable and six livestock product groups to which the data were aggregated. Differently from Weiss and Leip (2012) and in accordance to Leip et al (2014b) the allocation of flows from primary crop products to secondary products (e.g. soya to soybean oil and soybean cakes) is done by mass. The allocation of emissions from feed production to specific livestock products makes use of the animal budget module in CAPRI where energy and protein requirements are matched with domestic and imported feed supply, and data on farm expenditures for feed Witzke 2012, Leip et al 2011d). In a first step, emissions from crop activities are converted into emission intensities and allocated to animal activities and in a second step to animal products (Weiss and Leip 2012).
For our purpose, E agri was divided into emissions related to livestock production E lvst and those related to production of crops for other purposes (food, fuel, fibre) E crop . E lvst includes emissions from livestock production systems E anim (e.g. CH 4 emissions from enteric fermentation, emissions from energy use for milking etc) plus the emissions from feed production, transport and processing E feed .
For each crop product, allocation to food or feed is done on the basis of the market balance which is available in CAPRI at the national level (see supplementary information S2). The share of total quantities of flows allocated to food crops was calculated on the basis of total domestic production (gross production) minus the quantity used for feed, while the share of the flow used to feed is allocated to livestock products based on feed intake quantities. Emission sources considered are given in table 2.1 by Westhoek et al (2015).While this study is restricted to quantifying emissions from EU27 agriculture, E feed includes both emissions from domestic feed production E feedeu and emissions from imported feed products E feedrow :

Global warming
The contribution to global warming was assessed as the sum of direct GHG emissions (CO 2 , N 2 O and CH 4 ), indirect N 2 O emissions and C sequestration. Direct N 2 O and CH 4 emissions were from CAPRI-LCA, using a global warming potential of 298 kg CO 2 eq (kg N 2 O) −1 and 25 kg CO 2 eq (kg CH 4 ) −1 (IPCC 2007). Indirect N 2 O emissions were estimated as 1% of the emitted N (IPCC 2006). Indirect C sequestration in forests was calculated using an average carbon uptake (sequestration) factor of 35 kg C per kg N deposited for European boreal and temperate forests (de Vries et al , 2009) and a fraction of 0.25 kg N deposited on forest per kg N emitted, based on EU27 total NH 3 -N emissions from agriculture and NH 3 -N deposition on forests (de Vries et al 2011b). Other 'cooling effects'(Butterbach-Bahl et al 2011) were not included in the quantification.

Air quality
The contribution to air quality impacts was assessed on the basis of the sum of NH 3 and NO x emissions, as calculated with CAPRI N-LCA.

Soil acidification
Contribution to soil acidification was assessed on the basis of the sum of SO 2 , NH 3 and NO x emissions. NH 3 and NO x emissions were calculated with CAPRI N-LCA. Total SO 2 emissions were obtained from the EDGAR data base (European Commission 2011). SO 2 emissions caused by agricultural activities were approximated by the ratio of CO 2 emissions from agriculture related to energy use (Weiss and Leip 2012) and total energy CO 2 emissions in EU27 (EEA 2011). This gives a share of about 6%, which is associated with livestock products (4%) and vegetable products (2%), in accordance with the global estimate of FAO (FAO 2006).
The contribution of agricultural sources to emissions of acidifying substances was estimated on the basis of acidity equivalents (Schöpp and Posch 2003). This method converts emissions of S, NO x and NH y to acidity equivalents on the basis of the molecular weight m (64, 46, and 17 for SO 2 , NO x , and NH 4 + , respectively) and the charge per

Terrestrial biodiversity
The contribution of agriculture, livestock and feed to loss of relative mean species abundance (MSA) was estimated using shares of land use and emissions of NH 3 , NO x , and net GHG exchange, accounting for C sequestration, as calculated with the CAPRI-LCA. The data are linked to the estimates of the effect of the main drivers for biodiversity loss as calculated with the GLOBIO-model (Alkemade et al 2009, Kram and Stehfest 2012, van Vuuren et al 2015. This model gives an absolute loss of 65% MSA caused by land conversion into arable, grazing and forestry (35%, 15% and 14%, respectively), and to pressures such as N deposition (2%), climate change (3%) and land fragmentation (30%) (for details see supplementary information S3).

Water quality
The contribution of agriculture, livestock and feed to N r losses to the hydrosphere was derived from the results of CAPRI N-LCA. Contribution of livestock and feed to dissolved inorganic phosphorus (DIP) losses has been calculated by applying the share of P in fertilizers (mineral fertilizer and manure) per crop from CAPRI LCA to Global NEWS results on total and agricultural flows of DIP (Mayorga et al 2010). We were unable to quantify the role of agriculture in the load of particulate phosphorus (PP).
A quantification of the impact of N and P losses was done by combining an analysis of potential risk of eutrophication, based on the Indicator for Coastal

Results
The role of trade in N emissions in EU27 and other world regions N flows from EU27 to other world regions and vice versa for the year 2004 are illustrated in figure 2. Much of the proteins grown in Europe are used to feed livestock. From a total of 16.4 Tg N produced on agricultural land in the year 2004 only 2.4 Tg N yr −1 (about 15%) were supplied for direct human consumption or further processing. Most of it was used as animal feed (8.8 Tg N yr −1 or 54%) or returned to the soil as crop residue (5.1 Tg N yr −1 or 31%). Furthermore, we estimate that livestock received 4.2 Tg N yr −1 from imports or industry (Leip et al 2014b, see also details in supplementary information S2).
According to FAO trade statistics, EU27 was in 2004 a net importer of agricultural products with soybean products for animal feed produced in Argentina, Brazil and USA representing 84% of the total net imports of EU27 (see figure 2; for a comparison of EU27 and global agricultural structure and emissions see supplementary information S5) that entails significant trade of embodied cropland surface (MacDonald et al 2015). According to calculations based on FAOSTAT (2014) data, in 2004 about 70% of the European livestock production was used for intranational consumption and 18%-27% (respectively for chicken and cattle meat, expressed in N) was traded between EU27 countries with significant associated embodied GHG emissions (Caro et al 2014). The EU was thus close to self-sufficiency for meat and dairy products, but the share of pig meat production was much higher than in the rest of the world, while the share of ruminant meat was significantly lower (22% versus 29% globally).
The environmental impact of agriculture and livestock production in EU27 Table 2 shows the results for total agricultural emissions from the EU27 agricultural sector and emissions related to livestock production, feed production and imported feed. Values are provided for NH 3 , NO x , SO 2 , the combined effect of the three pollutants converted to acidity equivalents, GHG emissions (CH 4 , N 2 O, and CO 2 ), C-sequestration, water pollution by emissions of N and P (as dissolved inorganic P, DIP), the land use and the contribution to the loss of the MSA. Values reported in table 2 refer to the year 2004. Detailed results of the LCA calculation (N r and GHG emissions) are given in supplementary information S1.
Total agricultural emissions as compared with the total EU27 emissions from all sources  are given in table 3. Both total and agricultural emissions refer to emissions from EU27 territory. Therefore, agricultural emissions in table 3 do not include emissions associated with imported feed (see table 2).

Air quality
Agricultural sources of NH 3 from manure management, and manure and mineral fertilizers on soils totalled 2.8 Tg N yr −1 in 2004. The contribution of livestock production to total agricultural emissions was particularly high for NH 3 (82%) due to the importance of manure management. The share of NH 3 emissions linked to livestock feed was 41% of agricultural emissions, of which about 8% occurred outside Europe.
Total agricultural NO x emissions at 0.46 Tg N yr −1 were dominated by emissions from fossil fuel used for farm operations and during processing or transport of animal feed. The share of energy related emissions was higher for crop products (88%) than for livestock products (77%) with an overall contribution to emissions of 85% (0.39 Tg N yr −1 ). As there were only small NO x emissions from livestock production systems, most of the emissions were related to feed production, processing and transport (0.23 Tg N yr −1 ) and we estimated that about 51% of those occurred outside the EU territory or were linked to feed transport.
For the sum of NH 3 and NO x emissions, the share of agricultural emissions from livestock was 80% due to the dominance of NH 3 emissions. 42% of emissions were related to feed production, and 10% were associated with feed imports.

Global warming
The direct emissions of GHGs from the agriculture sector itself in 2003-2005 was 483 Tg CO 2eq yr −1 , contributing about 10% of total anthropogenic GHG emissions in the European Union (EEA 2011). However, we estimated emissions of more than twice that amount when including associated emissions that agriculture causes in other sectors, such as energy, industry, or land use and land use change (Weiss and Leip 2012). Overall, 81% of total European agricultural emissions (including associated emissions and emissions from outside of the EU27) were caused by livestock production. As much as 39% of agricultural emissions were estimated to occur outside the EU territory or from associated emissions. This includes especially feed imports, feed transport and emissions from land use change. Carbon sequestration induced by N deposition on forests was found to reduce agricultural emissions by about 100 Tg CO 2eq yr −1 (i.e., 10%). As agricultural N emissions are closely linked to manure management (see supplementary information S1), the N benefit for carbon sequestration was mainly located within the EU.

Soil quality
Emissions of acidity equivalents were dominated by NH 3 which accounted for about 85% of the acidity equivalent emissions for livestock (including associated emissions).

Terrestrial biodiversity
Expressed in terms of MSA, we estimated that overall agriculture, through arable and grazing and emissions of N and GHG, caused a loss of 34% MSA, i.e., more than half of the overall loss of biodiversity (Alkemade et al 2009). Of this agriculture related loss, 76% was estimated to be caused by livestock, with most of this through feed production.

Quality of inland and coastal water
Nitrogen Diffuse N losses from agricultural systems were estimated at 6.0 Tg N yr −1 . This represents percolation of nitrate and organic nitrogen below the rooting zone in agricultural soils, including both cropland and pasture land, and run-off from soils or barn yards. 73% of these emissions were associated with livestock, which was dominated by feed production. The share of leaching and runoff occurring outside of the EU territory was estimated at 10%.

Phosphorus
Diffuse losses of DIP from agriculture were estimated at 0.025 Tg P yr −1 , while weathering in agricultural systems contributes an estimated additional 0.003 Tg P yr −1 (Mayorga et al 2010). By far the largest share of net P input (P input minus P crop removal) was retained in the soil, which is considered a benefit as long as this leads to increased soil fertility, i.e., with low erosion. We do not have an estimate of agricultural dissolved organic phosphorus (DOP) or PP, as it is very difficult to distinguish sources for DOP and particularly for PP export. However, most likely the contribution of agriculture is much higher for PP than for DIP while PP dominates P export. The data presented in table 2 relate to DIP only and are therefore to be regarded as a conservative estimate for the total contribution of agriculture to P flows to coastal areas in Europe. Phosphorus losses from livestock were entirely attributed to feed production, with livestock DIP representing 73% of total agricultural losses, even though some additional losses from animal housing or manure storage systems might occur.
We estimate that in Europe the livestock sector accounts for 23%-47% of the nitrogen river load to coastal waters, and 17%-26% of the phosphorus river loads, where the lower limit is calculated considering the contribution of manure alone and the upper limit taking into account manure applications plus mineral fertilizer (see supplementary information S4).

Discussion
To our knowledge, this is the first study to estimate the contribution of livestock production systems, feed and Table 2. Share of the livestock sector, feed production and feed imports on the emissions of pollutants due to agriculture in EU27 with relevance for air quality, global warming, soil quality, biodiversity and water quality for the year 2004.

Air and soil quality
Air and soil quality Air quality Global warming feed imports to total agricultural emissions and their related environmental impact at a comparable level of detail. Plausibility of results have been discussed in depth with regard to GHG emissions (Weiss and Leip 2012) and N-footprints (Leip et al 2014b). Estimates of the share of N r emissions however are different from those given in Leip et al (2014b), as the authors calculated the shares on the basis of domestic consumption (human consumption or processing) while in this study we calculated the share on the basis of total production. Below, we first discuss uncertainty aspects of our emission estimates and estimates of emission shares, followed by options to reduce the environmental impact.  countries, while national inventories in Europe are constructed with a large variety of methods and data quality (Leip 2010). National estimates would be of higher quality than CAPRI estimates from countries with good data (Leip et al 2014b), but some countries still need to improve their methodology (EEA 2014). Furthermore, CAPRI uses ammonia abatement measures from the GAINS model (Klimont and Winiwarter 2011) which may not have been considered in national inventories .

Uncertainty of emission estimates
In comparison with 4.4 Tg N yr −1 agricultural NH 3 emissions in the EDGAR data base, our estimate of 2.6 Tg N yr −1 for EU27, excluding emissions from imported feed, is lower. The reason for this might be the lower excretion estimates, although the NH 3 emissions are in line with estimates by the MITERRA model.
While 85% of NO x emissions were related to energy use, only 0.07 Tg N yr −1 were from non-energy sources. A quality check of the total agricultural emission estimate for NO x is difficult as no comparable study exists including both NO x budget flows and NO x emissions related to energy consumption in agricultural systems. Energy consumption in agriculture is calculated in CAPRI with a dedicated energy module which is also used for GHG emission estimates (Kempen and Kraenzlein 2008, Weiss and Leip 2012); this is also the basis of the estimated contribution of SO 2 emissions. The share of agricultural NO x and NH 3 to total emissions is within −6% to +16% of earlier estimates .
Our estimate for the share of agricultural GHG emissions to total GHG emissions based on table 3 is about 13%. It ranges between the value of the official EU GHG inventory (10%, EEA 2014) and other estimates on the shares of agriculture or even livestock production on total GHG emissions , FAO 2006, Weiss and Leip 2012. The official GHG inventory considers only emissions reported in the agriculture sector, whereas LCA studies also include emissions from Land Use Change (LUC) and from imported feeds, which amounted to 39% of total agricultural emissions (see table 2).
Our estimate for N 2 O emissions from agricultural soils is considerably lower than official estimates; a comparison of N 2 O emissions between various models (de Vries et al 2011b) showed overall satisfying agreement. No methodology is able to capture the huge variability of N 2 O emissions caused by changing soil and climate conditions. In view of the general lack of experimental observations, even process-based models are not able to achieve a closer match than independent calculations using inverse methods .
LUC is certainly one of the most difficult sources to quantify, as it requires data (or good assumptions) on how much LUC is occurring as a consequence of EU agricultural and livestock production, as well as what kind of LUC is triggered. Indeed, the debate on the best method to estimate LUC emissions from agricultural products is still ongoing (European Commission 2013). The method developed in the CAPRI model (Leip et al 2010, Weiss andLeip 2012) was based on the assumptions that the agricultural market is very fluid and no differentiation between direct and indirect LUC is possible. The approach considers only LUC linked to an expansion of harvested area, very similar to the methods proposed by recent guidelines (Food SCP RT 2013, leap 2014. We used unique LUC factors for imports from a country outside the EU as weighted average for all importing countries accounting for globally connected and substitutable trade flows. We are aware of the debate on the permanency of carbon sequestration in grassland (Smith 2014), however the approach by Weiss and Leip (2012) is based on the observation that enhanced carbon sequestration rates in grassland are observed also after the 20 years equilibrium time usually used by IPCC methodologies (IPCC 2006), which is also consistent with recent simulations with the CENTURY model (Lugato et al 2014).
Carbon sequestration in forests has been estimated earlier at the scale of the EU27 for the year 2000 by multiplying an estimated N deposition caused by agricultural NH 3 emissions of 0.61 Tg N yr −1 with a C response of 50 kg C per kg N deposited leading to a C sequestration near 30 Tg C yr −1 or 112 Tg CO 2 yr −1 (de Vries et al 2011a), being very close to our estimate of 104 Tg CO 2 yr −1 . In our study the estimated N deposition was larger (0.82 Tg N yr −1 ) while the C:N response was estimated at 35 kg C per kg N deposited.
Estimates of agricultural N-leaching range from 2.0 to 5.7 Tg N yr −1 (de Vries et al 2011b), the higher value being also found in the EU GHG inventory (EEA 2014). Possible reasons for these differences-in addition to those already discussed-are available calibration data for nitrate concentrations which might neglect flows of organic nitrogen to water, and the split of total N between the highly uncertain N 2 emissions and N leaching/runoff (de Vries et al 2011b). Our estimate of nitrate leaching at 5.4 Tg N yr −1 is consistent with the estimate of the European Nitrogen Budget  which is used in table 3 for total N input to water. Although livestock dominated overall agriculture P flows (73%), agriculture is responsible for only 10% of the total riverine P export (table 3). This is because point sources from human wastewater dominate (accounting for 0.21 Tg P yr −1 out of a total riverine DIP export of 0.25 Tg P yr −1 ). The contribution of agriculture to the total P load (including DIP, dissolved organic P, DOP and particulate P, PP) may however be larger, specifically for PP, which is about 40% of the total P export to waters in Europe (50% is DIP and 10% is DOP). The share of agriculture in PP export is determined by (i) surface runoff of P in particles of fertilizer and manure, (ii) agricultural practices (e.g. tillage) that affect the erosion rate and (iii) elevated P contents of soil material eroded from agricultural fields due to application of P fertilizers and animal manure. The effects of agriculture on PP export are likely to occur much faster than the strongly delayed effect of DIP export through the soil systems but these mixed contributions make it very hard to assess what the agricultural contribution to the PP load is.
Finally, we have used the MSA indicator as a measure for terrestrial biodiversity. MSA represents an index of the naturalness of an ecosystem. Compared with more traditional measures (e.g. monitoring species changes), this measure has two main advantages: it is possible to attribute biodiversity loss to certain sectors (PBL 2014), and the effect of alternative scenarios on biodiversity can be quantified (PBL 2012(PBL , 2014. The patterns of change indicated by the MSA are largely similar to those indicated by other measures as the living planet index as developed by WWF or red list indices (SCBD 2014). A limitation of the MSA indicator is that it does not yield comprehensive information on the actual distribution and abundance of species, such as the status of endangered or threatened endemic species.

Uncertainty of the estimated shares
While some uncertainty is associated with the individual emission estimates, other parameters might dominate the uncertainty of the estimated shares. For example, while the estimate of N 2 O emission factors might be associated with an uncertainty of up to 50% (at the European scale), the uncertainty around the estimated share of crops that is used as feed determines the uncertainty of livestocks' contribution on total agricultural N 2 O emissions. A bias in the total feed translates directly into a bias of the estimated share of emissions from livestock production, as it not only determines the share of crops used as feed, but also the amount of CH 4 emissions from enteric fermentation in ruminants and manure excretion (which is calculated on the basis of animal retention data) and consequent emissions from manure management. This value is obtained from statistical sources (market balances) and is further constrained by energy and nutrient requirement calculations for major livestock types.
No data are available whether farmers prefer domestically produced crops or imported crops for feed; therefore this value is highly uncertain. Because of the lack of information, we considered equal preference. However, this uncertainty concerns only primary crops that are used for both food and feed, which make only 12% of the total feed, while the rest comes from non-marketable crops (82%), such as grass, fodder maize and beet, or secondary crops (6%). Nonmarketable crops are all domestically produced; secondary feed stuff is dominated by imported soya bean.
Options for reducing the environmental impact of livestock production There are two main routes to reduce the environmental impacts of livestock production: (i) technical measures (reduce emissions intensity/ land use intensity and (ii) lower livestock production in the EU with demand side measures, i.e., a reduction of food losses and wastes and/or dietary shifts.
Our study presents a 'status quo' analysis (attributional LCA) and does not examine emissions without (or with less) livestock production (case ii). What would happen if livestock production is reduced in Europe has been discussed in depth in Westhoek et al (2014Westhoek et al ( , 2015. Based on the observation that the intake of protein as well as saturated fats by European inhabitants is far above the maximum recommended level (WHO 2007, the authors showed that reducing the consumption of meat, dairy and eggs in the EU27 by 50% would lead to a decrease of N r emissions by 40% and a reduction of GHG emissions by 25%-40% with expected substantial health benefits (Westhoek et al 2014. Those results hold for two contrasting scenarios on the use of the 'freed' land that would not anymore be required for feed production, i.e., a 'greening scenario' with enhanced production of bio-energy and a 'high price' scenario with increasing export of cereals. They can be regarded as conservative scenarios, as beneficial environmental effects outside Europe had not been quantified, such as the subsequent prevention of land conversion outside Europe (Stehfest et al 2013), or the reduction of GHG emissions from bio-energy production (or other options such as afforestation).
From the production side many technical, structural or policy mitigation options are being discussed, addressing feed production (e.g., precision agriculture and agronomic nitrogen use efficiencies), livestock production (e.g. grazing and feeding management and feed supplements, improved herd structures), or housing and manure management

Conclusions
This analysis shows that, while agricultural activities are a major source of pollutants and land use change, livestock production systems dominate the environmental consequences. For the five threats considered here, livestock production contributed between 73% (water quality) to about 80% (biodiversity, air quality, soil acidification and global warming) of the overall agricultural impact.
The results point to the fact that in Europe serious efforts in mitigating the major environmental problems for Europe from agriculture need to address the livestock sector. While technical measures can clearly contribute significantly to emission reductions, they cannot alone be sufficient (Bellarby et al 2013, Bajželj et al 2014, Eshel et al 2014, Leip et al 2014a, Witzke et al 2014, Pierrehumbert and Eshel 2015, Vanham et al 2015. The issues of what European citizens eat and their food waste also need to be addressed. For example, recent scenarios showed that all these actions would be necessary to achieve a stabilization in global N 2 O emissions (UNEP 2013).
Moreover, while a shift of production from Europe to other world regions might make Europe 'cleaner', this would possibly come at the cost of higher emission intensities in other regions of the world where production systems might be less optimised (FAO 2010, Cederberg et al 2011; this could increase the environmental footprint of products consumed in Europe, unless additional actions were taken to address this.
Our study shows that there are intimate links between key environmental threats, emissions of N r to the environment, the production of animal products and our diet choices.