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SBSTTA Review DRAFT GBO4 – Technical Document – Chapter 8 DO NOT CITE

1 Target 8: 2 3 By 2020, pollution, including from excess , has been brought to levels that are not 4 detrimental to ecosystem function and biodiversity. 5 6 7 Preface 8 9 Biodiversity is affected by numerous including heavy metals emitted by , 10 industry and , used in agricultural practice, oil spills and excess of 11 nutrients especially (N) and (P). Given the high importance of N and P 12 impacts on biodiversity and ecosystem services at local to global scales (Fowler et al., 2013, 13 MA, 2005), the main focus of this chapter is on the pollution from N and P. Micro pollutants, 14 e.g. oil spills, plastic debris and use, are treated briefly in the section on status and 15 trends and 2020 projections, but are not treated in subsequent sections due to the lack of 16 long term scenarios. It should also be stressed that we discuss the pollutants separately, 17 while in reality areas might suffer from multiple pollutants. 18 19 Nitrogen - There are many natural processes that generate nitrogen (N) inputs into 20 ecosystems, but industry, transport and have greatly increased N inputs (Fowler 21 et al., 2013, Sutton et al., 2013). It is the excess above natural inputs (the N surplus) that is 22 referred to as N pollution. In non-agricultural terrestrial ecosystems, N pollution primarily 23 arises from wet and dry deposition of N that has been emitted into the air by industry, 24 transport and agriculture (direct effects of N addition to agricultural and 25 aquaculture systems are treated in Target 7). In aquatic ecosystems, N pollution comes 26 primarily from runoff and leaching of from agricultural and to a lesser extent from 27 and wet deposition. 28 29 Nitrogen pollution impacts ecosystems through a variety of mechanisms: the two most 30 important being and acidification (Bobbink et al., 2010, Fowler et al., 2013). 31 Eutrophication is the increase in productivity and accumulation of the load in 32 ecosystems, especially by nitrogen and/or phosphorus, leading to an undesirable disturbance 33 of the balance of organisms in the ecosystem, affecting both terrestrial and aquatic 34 biodiversity (Sutton et al., 2011). Acidification of ecosystems is caused by oxidation of NOx 35 and NH3 in air, and water (Pardo et al., 2011). In nitrogen oxides, , nitrate 36 and ammonium ions can enhance the rate of acidification, where nutrient bases (calcium, 37 magnesium, and potassium) are replaced by acidic elements (hydrogen and aluminum; 38 (Sutton et al., 2011). This may lead to nutrient disorders and to toxic effects in plants (Sutton 39 & Bleeker, 2013). Because N pollution is transported in the air as gases or , 40 atmospheric N pollution can cause eutrophication and acidification in terrestrial and aquatic 41 ecosystems at substantial distances from sources of emissions. 42 43 Nitrogen accumulation is the main driver of species composition changes since many 44 species-rich terrestrial ecosystems are adapted to conditions of low N availability. In these 45 ecosystems excess of nitrogen change competitive relations, and can make conditions 46 unfavorable for some species (Bobbink et al., 2010). Other effects (e.g. direct toxicity of 47 nitrogen gases and aerosols, long-term negative effects of increased ammonium and

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SBSTTA Review DRAFT GBO4 – Technical Document – Chapter 8 DO NOT CITE

1 ammonia availability, soil-mediated effects of acidification, secondary stress and 2 disturbances) are more ecosystem- and site-specific and often play a supporting role 3 (Bobbink et al., 2010). N pollution may also increase the dominance of invasive alien plants in 4 terrestrial ecosystems and decrease the diversity of plant communities worldwide (IAASTD, 5 2009). In addition, N emissions make an important contribution to greenhouse gas balance 6 (Butterbach-Bahl & Dannenmann, 2011). NOx emissions are a major contributor to the 7 formation of tropospheric , while N2O emissions cause depletion of stratospheric 8 ozone. Tropospheric ozone has negative effects on plant productivity and human health. 9 10 Phosphorus - Phosphorus (P) has no significant atmospheric component, except for transport 11 of P attached to soil particles. Next to this, runoff from agriculture fields, sewage water from 12 households containing human excreta and detergents, and wastewater from industries are 13 the main causes of pollution. This primarily impacts aquatic ecosystems through 14 eutrophication, since P is transported more often in water and much less so in air. 15 Traditionally N pollution is thought to primarily impact non-tropical terrestrial ecosystems 16 and coastal zones, while P pollution is traditionally thought to have the largest impacts on 17 freshwater ecosystems. However, recent evidence suggests aquatic ecosystems are often 18 affected by both N and P pollution (Conley et al., 2009b, Elser et al., 2007). For example, 19 eutrophication in lakes, rivers and coastal areas caused by N and P pollution can lead to the 20 generation of oxygen depleted areas (<2 ml O2/L, hypoxia and for coastal systems also known 21 as “dead zones”), the formation of harmful algal blooms (Anderson et al., 2002), increased 22 occurrence of pathogenic microorganisms, and uncontrolled population growth of certain 23 weedy plants or gelatinous zooplankton. 24 25 To regulate N and P pollution national or trans-boundary regulations which are based on the 26 concept of “critical loads” have been developed. Critical loads are thresholds above which N 27 and P pollution has negative effects on biodiversity, ecosystem function or ecosystem 28 services. For each ecosystem type, different levels of N and P are detrimental, and critical 29 loads may therefore differ between countries. For example, critical loads for atmospheric N 30 deposition in the European Union are currently set at 5-40 kg N/ha/yr depending on 31 ecosystem type (Bobbink et al., 2010), while in the United States the ranges from 32 1 kg/ha/yr to nearly 40 kg/ha/yr (Pardo et al., 2011). These thresholds are based on 33 experiments, observations and models. The use of critical loads has been questioned 34 because it is difficult to precisely determine critical loads, since they are highly system 35 specific and some systems appear to be affected even at very low loads (Payne et al., 2012). 36 37 Pesticides - pesticides cover a wide range of compounds including insecticides, fungicides, 38 , rodenticides, molluscicides and nematicides. Ideally a pesticide is lethal to the 39 targeted pests, but not to non-target species. Unfortunately, pesticides can be toxic to a host 40 of other organisms, including birds, fish, beneficial insects, and non-target plants and can be 41 lethal or may cause, among others, neurological and behavior disorders (Mitra et al., 2011). 42 Pesticides can contaminate soil, water, turf, and other vegetation and can therefore be 43 transported over great distances via air and water. Pesticides are persistent and can 44 accumulate in the food chain which can lead to risks to non-target organisms (Aktar et al., 45 2009, Mitra et al., 2011). Pesticides can also cause soil degradation when they are toxic to 46 microorganisms (Aktar et al., 2009). 47

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SBSTTA Review DRAFT GBO4 – Technical Document – Chapter 8 DO NOT CITE

1 Micro pollutants - micro pollutants include plastics, pharmaceuticals, personal care products 2 (PCPs), steroids, hormones, surfactants, perfluorinated compounds (PFCs), various persistent 3 organic pollutants (POPs), nanomaterials and swimming pool disinfection by-products (Farre 4 et al., 2008). These are derived from diverse industrial activities, , 5 waste disposal and many other sources (Travis & Hester, 1991). Often they enter the ocean 6 through waste-water treatment plants (Farre et al., 2008). Some toxic pollutants are natural 7 elements that become widely dispersed by human activities, as in the case of radioisotopes 8 and petrochemicals (Pirrone et al., 2010). Others are xenobiotic compounds synthesized to 9 serve as pesticides, solvents and lubricants. Those micro pollutants can be acutely or 10 chronically toxic to individual organisms or accumulate in food chains and become 11 concentrated in top predators (Persson et al., 2013). Next to this, most of these compounds 12 are long-lived once released into the environment and they are able to circulate widely in the 13 atmosphere and oceans, allowing long-distance transport to sites that are distant from any 14 known source, resulting in high concentrations of many pollutants in polar animal species 15 (e.g. Braune et al., 2005). 16 17 Chemicals – chemicals (e.g. lead, cadmium, cooper, PAHs and PCBs) are emitted via industry, 18 agriculture, households and other human activities. This results in local and diffuse 19 environmental concentrations of chemicals that vary in space and time. Consequently, this 20 leads to a variety of contaminated areas; some with diffuse concentrations of chemicals 21 while hot spots have substantial local . This leads to a highly variable 22 exposure of, amongst others, soil biota. Thereby affecting soil biodiversity, soil functioning 23 and soil processes such as the breakdown and recycling of organic materials, nutrient cycling, 24 , as well as the breakdown of organic contaminants (ecosystem services). 25 26 Through exposure of soil biota (acting in local soil food webs) to different local composition 27 and concentrations of the contaminant mixture biodiversity is affected. Ecotoxicity effects are 28 expected to affect ecosystem function and soil biodiversity at different impact levels. For 29 large diffusely contaminated areas, the impact on soil biota originates from relatively low 30 concentrations of rather complex mixtures of contaminants over wide areas. Hotspots with 31 contamination have high enough concentrations of the mixture of contaminants to impact 32 large parts of the local soil biota in the soil food web. These contaminated hot spots often 33 haven been the result of unintended spills and other local discharges. 34 35 36 1. Are we on track to achieve the 2020 target? 37 38 1.a. Status and trends 39 40 1.a.i. N emissions 41 42 Nitrogen losses to the environment per capita are more than twice the world average in 43 North America (Figure 8.1). The largest fraction of this N pollution comes from food systems, 44 including biofuels. N emissions into the atmosphere from fossil energy use in industry and 45 transport make substantial contributions to N pollution in North America and Europe, but 46 only small contributions in Africa. 47

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SBSTTA Review DRAFT GBO4 – Technical Document – Chapter 8 DO NOT CITE

1 2 Figure 8.1. Average loss of to the environment per inhabitant per continent in 3 2008. The highest nitrogen losses to the environment per capita are highest in North America. Source 4 (INI, 2014). 5 6 Atmospheric N deposition varies greatly across regions (Figure 8.2). Currently, the most 7 heavily polluted areas are the eastern half of North America, Europe and Eastern and 8 Southern Asia (Lamarque et al., 2013). N deposition increased rapidly in Europe and North 9 America over much of the 20th century, but has stabilized or decreased since the 1980s, in 10 part due to regulation of sources of emissions (Sutton et al., 2013). The decrease in N 11 deposition has been particularly strong in some European regions. Nitrogen deposition has 12 increased rapidly in Eastern and Southern Asia over the last several decades due to increased 13 emissions from industry, transport and agriculture (Figure 8.2). 14 15 A. Wet and dry deposition (e.g., ammonia absorption by plants) of reduced forms of N

16 17 B. Wet (e.g., nitric acid in rain) and dry deposition of oxidized forms of N

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20 N deposition in kg/ha/yr 21 Figure 8.2. Multi-model estimates of atmospheric N deposition in 1850, 1980 and 2000 (Lamarque et 22 al., 2013). Eutrophication of terrestrial systems from N pollution depends on total N deposition which 23 is the sum of deposition in A and B. Source: Lamarque et al. (2013). 24 25 The widespread emissions of N have increased N deposition beyond critical loads in large 26 areas (Figure 8.3; Bobbink et al., 2010, PBL, 2012, Posch et al., 2011). Nitrogen deposition

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SBSTTA Review DRAFT GBO4 – Technical Document – Chapter 8 DO NOT CITE

1 currently exceeds estimates of critical loads for large parts of Europe, eastern North America, 2 Eastern and Southern Asia, as well as parts of central Africa and South America. 3

4 5 Figure 8.3. Locations where nitrogen deposition exceeds critical loads. Especially in Europa, eastern 6 USA, India and China nitrogen deposition exceeds the critical load. Figure based on data from (OECD, 7 2012, PBL, 2012) and critical loads on Bobbink et al. (2010), Bouwman et al. (2002). 8 9 1.a.ii P emissions 10 11 The current P losses to the environment from urban wastewater discharge differ 12 considerably between countries. In Northern and Western Europe, 80 to 100% of all laundry 13 detergents are P free, while in Eastern European only 15% to 20%, and in many developing 14 countries even smaller proportions of laundry detergents are P free (van Drecht et al., 2009). 15 Combined with the increasing use of P-free detergents in Europe, North America and 16 Oceania, much of the nutrients in wastewater are removed in treatment plants. As a 17 consequence, P release to has stabilized in recent decades (van Drecht et al., 18 2009). However, with the use of P-based detergents and lacking wastewater treatment, 19 combined with the rapid population growth and sewage connections in South Asia and 20 Africa, the P pollution has been growing rapidly in these regions (211% resp. 300% increase; 21 van Drecht et al., 2009). In Central and South America, the population growth, sewage 22 connection, use of P-based detergents, and lagging wastewater treatments, has resulted in a 23 220% increase of P pollution (van Drecht et al., 2009). 24 25 Due to growing fertilizer use in agriculture, the global P flow to the has increased 26 by at least 75% the last half-century (UNEP, 2011) and has tripled compared to preindustrial 27 levels (Macdonald et al., 2011). This has lead on 71% of the croplands to a P surplus. 28 However 29% of the global crop area has limited availability of P fertilizers, which often leads 29 to soil degradation caused by P depletion (Figure 8.4). The largest deficits were found in 30 South America, northern USA and Eastern Europe (Figure 8.4; Macdonald et al., 2011). The 31 largest P surpluses can be found in East Asia, Western and Southern Europa, coastal USA and 32 Southern Brazil (Macdonald et al., 2011). In these regions, residual P has accumulated in the 33 soil, leading to increased availability of soil P to plants. As a consequence, P fertilizer 34 application rates have gone down, and resulted in many countries in an application rate close 35 to the plant uptake.

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2 3 Figure 8.4. Global map of agronomic P imbalances for the year 2000 expressed per unit of cropland 4 area in each 0.5° grid cell. The P surpluses and deficits are classified according to quartiles globally 5 (0–25th, 25–50th, 50–75th, and 75–100th percentiles; Macdonald et al., 2011). 6 7 1.a.iii. N and P pollution in freshwater ecosystems 8 9 For more than 60 years, enrichment of fresh waters (e.g. rivers, lakes, and the 10 catchment areas draining to them) with N and P from fertilizers and sewage discharges, has 11 been a major causes of impairment (Carpenter et al., 1998, Halpern et al., 12 2008). Nutrient loading now affects virtually all rivers on earth (Vorosmarty et al., 2010) and 13 can result in eutrophication and causing algal and bacterial blooms and loss of light 14 and oxygen from the water (BIP, 2014). Next to this, the declining water quality can cause 15 loss of species and shifts from pollution-sensitive towards pollution-tolerant organisms (BIP, 16 2014). Increase of sediment loads can also interfere with fish respiration, smother bottom- 17 living organism and cover spawning areas (BIP, 2014). 18 19 To determine the effect of water quality on freshwater biodiversity, the Water Quality Index 20 for Biodiversity (WQIB) is used. The WQIB is based on a comprehensive global-scale dataset 21 on water temperature, dissolved oxygen, pH, electrical conductivity (salinity), nitrogen and 22 phosphorus (UNEP-GEMS, 2008). Based on this index, water quality is fair or poor in more 23 than half of all freshwater ecosystems in the Americas, Europe and Africa (Figure 8.5). 24 General declines in the percentage of stations classified as good or excellent were detectable 25 in the Americas and Europe dating back to the 1970s and 1980s, but have stabilized in recent 26 decades (Figure 8.5). Water quality in Oceania appears to have increased in the last decade 27 or two, as the proportion of stations classified as excellent or good has increased. The water 28 quality in Asia and Africa is declining, but regional differences exist. For example the N and P 29 pollution in parts of Indian rivers has significantly declined (Evans et al., 2012). N and P 30 pollution are the dominant factors that underlie low water quality, with approximately 60% 31 of all stations having fair or low water quality based on N and P pollution criteria (UNEP- 32 GEMS, 2008).

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SBSTTA Review DRAFT GBO4 – Technical Document – Chapter 8 DO NOT CITE

1 2 Figure 8.5. Regional Water Quality Index for Biodiversity (WQIB) scores for monitored freshwater 3 systems. Excellent to good scores are indicated in green, fair to marginal and poor scores are 4 indicated in yellow and red respectively. Solid black line indicates number of stations reporting in any 5 given year (UNEP-GEMS, 2008). 6 7 The most severe manifestation of N and P pollution is the appearance of “dead zones” where 8 oxygen levels in the water drop to such low levels that many aquatic organisms are killed. 9 Prior to the 1960s, eutrophication-induced hypoxia was mostly associated with rivers, 10 estuaries, and bays. By the end of the 1980s the so called “dead zones” had developed and 11 expanded in continental seas, such as the Baltic Sea, Kattegat, Black Sea, , and 12 East China Sea (Diaz & Rosenberg, 2008). In inland waters, such dead zones have been 13 recorded in North America’s Lake Erie (Conroy et al., 2011) and Africa’s Lake Victoria 14 (Verschuren et al., 2002). Since the 1960s, the number of hypoxic systems has about doubled 15 every decade and by 2010 hypoxia had become a major worldwide environmental problem 16 with about 650 systems with reports of hypoxia (Figure 8.6). About 10% of these hypoxic 17 systems have improved from management of nutrient and organic matter discharges. 18 Additionally, over 400 coastal sites globally were identified as areas of concern that currently 19 exhibit signs of eutrophication and are at risk of developing hypoxia (Figure 8.6; Conley et al., 20 2011, Diaz et al., 2010). 21

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SBSTTA Review DRAFT GBO4 – Technical Document – Chapter 8 DO NOT CITE

1 2 Figure 8.6. The global pattern of development of coastal eutrophication and hypoxia. Each red dot 3 represents a documented case related to human activities. The number of hypoxic sites is cumulative 4 through time (from Diaz et al., (2013)). An interactive version of this figure can be found at: 5 http://www.wri.org/our-work/project/eutrophication-and-hypoxia/interactive-map-eutrophication- 6 hypoxia 7 8 1.a.iv. Other pollutants 9 10 Micro pollutants - There is a wide range of emerging environmental contaminants or micro 11 pollutants that are not covered by current regulations, but that may pose risks to aquatic 12 species. However, the threat posed by many of these compounds to marine species is poorly 13 studied and often unknown. These include plastic, pharmaceuticals, personal care products 14 (PCPs), steroids, hormones, surfactants, perfluorinated compounds (PFCs), various persistent 15 organic pollutants (POPs), nanomaterials and swimming pool disinfection by-products (Farre 16 et al., 2008). For example plastic debris on the ocean surface has been estimated up to 17 3,520,000 items km2 (Barnes et al., 2009) and up to 100,000 items m2 on some shores 18 (Yamashita & Tanimura, 2007). Plastic debris may act as a vector for 19 (Gregory, 2009) and may entangle or be ingested by marine species. More than 260 species 20 have been reported as being affected by plastic debris.

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1 Oil spills - In the open ocean, small quantities of oil are readily dispersed and degraded 2 (GESAMP, 2007). However, catastrophic releases of oil in coastal waters can be a threat to 3 marine ecosystems and species. Regulation that has followed catastrophic oil spills has been 4 partially successful in reducing the occurrence of such events. Oil spills from tankers have 5 decreased dramatically from an average of 314,000t in the 1970s to 21,000t in the early 6 2000s (Jernelöv, 2010). This has partially arisen as a result of redesign of oil tankers as well 7 as improvements in the management of shipping traffic and navigational technology 8 (Jernelöv, 2010). Operational discharges of oil are still substantial (in the order of 100,000t) 9 but again the trend is downwards (Jernelöv, 2010). However oil spills from pipelines, mainly 10 on land, have increased largely as a result of ageing infrastructure, poor maintenance and 11 lack of prompt remedial action especially in former Soviet Union and West Africa (Jernelöv, 12 2010). One trend for oil exploration and production has been the increasing depth of oil 13 wells (Caineng et al., 2010) and also movement into more extreme environments such as the 14 Arctic. The Deepwater Horizon disaster of 2010 has shown that impacts from catastrophic oil 15 spills can extend into deep-sea ecosystems and negatively impact species such as cold-water 16 corals (White et al., 2012) and other components of benthic communities (Montagna et al., 17 2013). It may take ecosystems to recover from such disaster as little as 3-4 years (for rocky 18 shores; Kingston, 2002). Although in others, recovery can take as long as 10-15 years, as in 19 the case of the Torrey Canyon on the shores of southwestern England (Hawkins et al., 20 2000), and oil may be detectable on shores up to 25 years or more (Kingston, 2002). 21 22 Pesticides1 – Leaching of pesticides is one of the main mechanisms leading to groundwater 23 and surface (Van Dijk et al., 2013). Globally approximately 9,000 species of 24 insects and mites, 50,000 species of plant pathogens, and 8,000 species of weeds damage 25 crops. Therefore about one-third of the agricultural products are produced by using 26 pesticides (Zhang et al., 2011). It is believed that pesticide use is necessary to ensure food 27 safety; however pesticides are highly toxic to humans and the environment (Zhang et al., 28 2011). For example, one of the most significant current debates concerning biodiversity is 29 about the potentially harmful impact of neonicotinoids on pollinating insects like honeybees 30 and bumblebees. There is evidence that the use of this type of systemic pesticide is 31 contributing to large-scale die-off of bee populations, although many other factors appear to 32 play important roles (van der Sluis et al., 2013). Since 1990 pesticide use has been increasing 33 up to two million tonnes per year (De et al., 2014). 45% is consumed by Europe and 25% by 34 the USA (De et al., 2014). However neonicotinoid use as a systemic pesticide in crops was 35 recently suspended in the European Union due its suspected negative impacts on pollinators. 36 However due to the high persistency of this pesticides, the impact on biodiversity will not be 37 reduced immediately. Most frequently used pesticides are in the form of herbicides (Figure 38 8.7), however regional differences exist. For example in India 80% of the used pesticides is in 39 the form of insecticides (De et al., 2014).

1 Taskforce pesticide use will publish a report end of June. If possible we would like to add information from this document in July.

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SBSTTA Review DRAFT GBO4 – Technical Document – Chapter 8 DO NOT CITE

1 2 Figure 8.7. Worldwide consumption of pesticides (De et al., 2014). 3 4 Chemicals2 – Industry, agriculture, households and other human activities have emitted 5 various chemicals and it is expected that global output of chemicals will increase with 85% 6 by 2020 compared to 1995 (HELI, 2004). These chemicals have resulted in many 7 contaminated sites across the globe; chemicals have contaminated soil, water and air 8 especially in developing countries (Blacksmith institute, 2012). Many developing countries 9 suffer from chemical pollution. In China, for example, the chemical pollution is worsening; it 10 has been estimated that about one-tenth of China’s cultivatable land has been 11 contaminated, mostly in economically developed areas (Qi, 2007). Below we discuss lead, 12 chromium and pollution as an example of chemical pollution. 13 14 Mercury pollution can be found worldwide and is released into the environment through 15 mine or by evaporation from gold-mercury amalgams to recover the metallic gold. 16 Biodiversity is heavily affected by this since it is bio accumulative and persist in the 17 food chain (Blacksmith institute, 2012). 18 19 Chromium pollution is caused by tanneries and mining in various African, South American 20 and North Asian countries. The majority of the chromium-contaminated sites can be found 21 in Pakistan and India (Blacksmith institute, 2012). 22 23 Lead can result in air and water pollution and the global production of lead has increased by 24 9% in 2011, due to increases in China, India and Mexico (Blacksmith institute, 2012). Most 25 contaminated sites can be found in Africa, South America, South and Southeast Asia 26 (Blacksmith institute, 2012). To halt the lead pollution, increasing quantities are recycled, 27 however recycling often occurs at poorly controlled facilities making lead processing itself a 28 source of pollution. 29 30 Chemical pollution persists also in developed countries. In Canada, for example, there are 31 more than 10,000 contaminated sites (McKie, 2014) and in Australia more than 150,000 32 contaminated sites (Carbonell, 2013, Ruehl, 2013). At the beginning of 2014, the superfund 33 program of the USA identified 1,319 contaminated sites in the USA (EPA, 2014), however 34 there is some discussion whether this program includes all contaminated sites (Scorecard, 35 2011). In Europe 4.2 potentially contaminated sites are reported per 1,000 inhabitants and 36 about 5.7 contaminated sites per 10,000 inhabitants (van Liederkerke et al., 2014). The 37 distribution of the different contaminants is similar for both liquid and solid matrices. The

2 some more expert input or suggestions for references required

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SBSTTA Review DRAFT GBO4 – Technical Document – Chapter 8 DO NOT CITE

1 most frequent contaminants are mineral oils and heavy metals. Contamination by mineral oil 2 is especially dominant in Belgium (solid matrix: 50%) and Lithuania (solid matrix: 60%), while 3 for Austria (solid matrix: 60%) and the FYROM (solid matrix: 89%) heavy metals 4 predominate. Generally, phenols and cyanides make a negligible overall contribution to total 5 contamination. 6 7 To delineate the spatially and temporally variability of the chemical threats to soil ecosystems 8 the mixture toxic pressure of a site can be determined. It is expressed as multi-substance 9 Potentially Affected Fraction of Species (msPAF; Posthuma et al., 2002). The msPAF can be 10 interpreted as potential loss of species, signifying hazard differences amongst areas. As 11 example, a spatial analysis of emission, fate, exposure and mixture assessment data has been 12 compiled in a map that shows spatial differences in toxic pressures of two metals in Europe 13 (Figure 8.8, Cd and Pb). The results demonstrate that there is variation in the net toxic 14 pressure in top soils across Europe (De Zwart et al., 2010, Van der Voet et al., 2013). 15 Assuming a 1:1 association between toxic pressure and species loss, a loss of biodiversity 16 from Cd‐ and Pb‐deposition is expected to occur only in certain areas in Southern and 17 Eastern Europe (De Zwart et al., 2010, Van der Voet et al., 2013). Throughout Europe, the 18 predicted loss of species due the expected metal (Cd and Pb) loads at steady state soil 19 concentrations with current depositions is lower than 1% (De Zwart et al., 2010, Van der Voet 20 et al., 2013). However note that large uncertainties are involved in 21 estimation, up to a factor 1000 (Unsworth et al., 2006). Also other metals like copper and 22 zinc should be taken into account. 23

24 25 Figure 8.8. The 99th percentile of loss of species at steady state for cadmium (left) and lead (right) (De 26 Zwart et al., 2010). 27 28 29 1.b. Projecting forward to 2020 30 31 A large number of scenarios of future N and P pollution are available. A substantial fraction 32 of these scenarios cover the 2030, 2050 and 2100 time horizons. Scenarios for 2030 have 33 been used in this section when scenarios for 2020 are not available. 2020 pollution is based 34 on interpolating from “current” (often for 2000) N and P emissions to 2030 estimates. 35 1.b.i. Nitrogen and phosphorus emissions

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1 It is foreseen that the N and P losses to the environment will increase in the coming decades 2 in a variety of business-as-usual scenarios (Bouwman et al., 2005, Dentener et al., 2006, PBL, 3 2012, Seitzinger et al., 2010). Several regions have declining N losses to the environment, but 4 these are counterbalanced by large increases in Asia, Central and South America and Sub- 5 Saharan Africa (Figure 8.9; Dentener et al., 2006, PBL, 2012). Similar trends are expected for 6 P pollution; the stabilizing of P emissions in some regions is counterbalanced by the increase 7 of P pollution in others, e.g. Sub-Saharan Africa (Figure 8.9). 8

9 10 Figure 8.9. Global nitrogen and phosphorus surpluses in agriculture per region following the RIO+20 11 “baseline” scenario up to 2050 (PBL, 2012). It is expected that the N and P losses from agricultural 12 fields to the environment will increase in the coming decades. Europe, North America, China and 13 India may have declining N and P surpluses, but these may be counterbalanced by large increases in 14 fertilizer use in parts of Asia, Central and South America and Sub-Saharan Africa. 15 16 1.b.ii. N pollution in terrestrial ecosystems 17 18 The total N fertilizer input per hectare in developing and emerging countries is expected to 19 exceed those of industrialized countries by 2030, due to high cropping intensities that prevail 20 in many tropical countries (Sutton & Bleeker, 2013). In all scenarios developed for the IPCC 21 this leads to a substantial increase of global N deposition, with especially high increases in 22 Southern and Eastern Asia (up to 40-100%; Figure 8.10a, compare with Figure 8.2a; Dentener 23 et al., 2006, Lamarque et al., 2013, Paulot et al., 2013). Regulations on transport and 24 industrial N emissions are assumed to continue to reduce emissions of oxidised forms of N in 25 developed countries, but rising transport and industrial emissions are assumed to increase 26 deposition substantially in Asia (up to 50-100% increase; Figure 8.10b, compare with Figure 27 8.2B; Dentener et al., 2006). 28

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1 A. Scenarios of wet and dry deposition of reduced forms of N in 2030

2 3 B. Scenarios of wet and dry deposition of oxidized forms of N in 2030

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6 N deposition in kg/ha/yr 7 Figure 8.10. Scenarios of atmospheric N deposition in 2030 for three IPCC scenarios for strong 8 (RCP2.6), moderate (RCP 4.5) and weak (RCP 8.5) climate mitigation scenarios. Deposition of reduced 9 forms of N is primarily due to emissions from agricultural systems (panel A). Deposition of oxidized 10 forms of N principally arises from industrial and transport emissions (panel B). Source: (Lamarque et 11 al., 2013). 12 13 The extent to which future N deposition may exceed critical loads in terrestrial ecosystems 14 has been explored for the 2020 to 2030 time horizon in several studies. Bobbink et al. (2010) 15 estimated N deposition for the WWF Global 200 (G200) ecoregions of high conservation 16 value. There are very large differences between scenarios even at this time scale. Assuming 17 an average critical load of 10 kg N/ha/yr (Bobbink et al., 2010), the percent of G200 18 ecoregions above this limit is similar to current levels in the most optimistic scenario (10 - 19 12%; Figure 8.11, inset MFD scenario), but more than double the current levels in the most 20 pessimistic scenario (25%; Figure 8.11, inset A2 scenario). Ecoregions in Southern and 21 Eastern Asia are projected to be the most heavily impacted. 22 23 Also Paulot et al. (2013), using the RCP IPCC scenarios, project N deposition to exceeds 24 critical loads for some ecosystems in protected areas. Even though pollution from oxidized 25 forms of N is projected to stabilize (China) or decline (US), reduced forms increase. For 26 example in the Rocky Mountain area N deposition will exceed the critical loads, since this 27 area is very sensitive to N deposition and has low critical loads. 28

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1 2 Figure 8.11. Map of multi-model estimates atmospheric N deposition in 2030 for the WWF Global 3 200 (G200) ecoregions. N depositions to areas outside the G200 ecoregions are not shown. The map 4 is based the IPCC A2 emissions scenario which assumes rapid population growth with little attention 5 paid to environmentally friendly actions. Inset: Percent of G200 ecoregions having N deposition 6 exceeding 10 kg N/ha/yr which is a rough average of critical loads for terrestrial ecosystems. CLE = 7 scenario assuming current legislation is enforced, MFD = maximum feasible reduction in deposition 8 given foreseeable technologies and A2 = IPCC A2 scenario (corresponds to map). The numbers inside 9 the bars show the number of G200 ecoregions in the area affected by each scenario. Note that 10 projected deposition rates are low in Madagascar, Australia and New Zealand. Source: (Bobbink et al., 11 2010, Dentener et al., 2006). 12 13 Due to current legislations, only 20% of the sites in Europe are protected from N 14 eutrophication. Projections suggest that this will increase to about one third by 2020 based 15 on N deposition reductions due to excisting legislation. Under a maximum reductions 16 scenario about half of the sites would be protected. Even the “maximum technically feasible 17 emission reductions scenario” would not fully protect ecosystems across Europe from N 18 acidification or N eutrophication (Holmberg et al., 2013). 19 20 The potential impact of N deposition on protected areas with high conservation value is 21 estimated by using a “current legislation scenario” for N deposition (Bleeker et al., 2011, 22 Dentener et al., 2006, Dentener et al., 2005). Protected areas in tundra, Mediterranean and 23 mangrove systems are least exposed, but more than half of temperate forests and a third of 24 temperate grasslands and tropical and subtropical broadleaf forests in protected areas are 25 projected to be exposed to N deposition levels that could have negative impacts on 26 biodiversity (>10 kg N/ha/yr; Figure 8.12). However, relatively few protected areas are 27 projected to have N deposition levels exceeding 20 kg N/ha/yr; the largest fractions are in 28 tropical, subtropical and temperate forests. 29 30 Optimistic scenarios suggest that it is possible to stabilize or achieve reductions in N pollution 31 of terrestrial ecosystems over the 2020-2030 time frame in most regions. However, these 32 levels of pollution are high enough in many regions to remain “detrimental to ecosystem 33 functioning and biodiversity”. To meet or exceed these optimistic scenarios, efforts need to 34 be made in many countries to control agricultural N emissions. In addition, to control 35 transport and industrial emissions advanced control techniques could reduce N emissions by 36 50%, however further efforts would need to be made (see section 2).

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1 2

3 4 Figure 8.12. The distribution of deposition classes in protected areas. Red indicates areas where the N 5 depostion in 2030 exceeds 10 kg N/ha/y and where the deposition in 2030 is higher than in 2000. 6 Orange indicates areas where the N depostion in 2030 exceeds 10 kg N/ha/y and where the 7 deposition in 2030 lower than in 2000. Green indicates areas where the N depostion in 2030 exceeds 8 5 kg N/ha/y and where the deposition in 2030 lower than in 2000 (Bleeker et al., 2011). 9 10 1.b.iii. N and P pollution in aquatic ecosystems 11 12 Recent model estimates suggest that N and P loading in rivers by 2030 will differ from 13 current levels substantially depending on socio-economic development pathways (Figure 14 8.13; Qu & Kroeze, 2012, Seitzinger et al., 2010, Strokal & Kroeze, 2013, Suwarno et al., 15 2013). In general, scenarios suggest that agriculture will remain the key driver of N loading to 16 rivers through 2030, while the future of riverine P will depend more heavily on sewage 17 treatment, detergent P content and retention of particulate P in reservoirs (Seitzinger et al., 18 2010). Estimated increases over the last thirty years and projections for the next thirty years 19 for both N and P loading are most striking in Southern Asia (Figure 8.13). In scenarios with 20 high increases of agricultural production and consumption of meat, current trajectories of 21 increasing export of N and P in rivers are predicted to continue or worsen in most regions 22 especially in developing and emerging regions (Figure 8.13, GO 2030; Seitzinger et al., 2010). 23 In scenarios with moderate increases in agricultural production and meat consumption, N 24 export in rivers generally declines and P export generally increases, but only modestly (Figure 25 8.13, AM 2030). Models project that dissolved inorganic forms of N and P will increase, 26 dissolved organic forms of N and P will remain relatively stable and particulate N and P will 27 decline (Figure 8.13).

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1 2 Figure 8.13. River export at continent/regional scale for (A) Dissolved inorganic nitrogen (DIN) and (b) 3 dissolved inorganic phosphorus (DIP) in 1970 (light blue), 2000 (dark blue), and 2030 for the Global 4 Orchestration (GO; the least favorable) and Adapting Mosaic (AM; the most favorable) scenarios of 5 four the Millennium Assessment socio-economic scenarios (MA 2005). Source: (Seitzinger et al., 6 2010). 7 8 Methods similar to Seitzinger et al. (2010) have been applied at national and regional scales 9 including watersheds feeding the Baltic Sea (Strokal & Kroeze, 2013), Chinese coastal zones 10 (Qu & Kroeze, 2012) and Indonesian coastal zones (Suwarno et al., 2013). These studies 11 broadly confirm the global analysis of Seitzinger et al. (2010), but they have also explored a 12 number of variants of scenarios that explore plausible reductions in factors mediating N and 13 P loading to rivers. Large, but plausible reductions in organic and inorganic fertilizer use 14 accompanied by improved removal of N and P from sewage are projected to substantially 15 reduce N and P loading in rivers and coastal zones below “business-as-usual” scenarios and 16 often result in substantial reductions compared current loadings by 2020 (see example for 17 China; Figure 8.14). Despite these improvements, projected N and P pollution remain above 18 limits that would avoid eutrophication (Strokal & Kroeze, 2013). 19

20 21 Figure 8.14. Trends and scenarios of A) dissolved inorganic N (DIN) and B) dissolved inorganic P (DIP) 22 loading in rivers feeding Chinese coastal areas. BS = baseline scenario, RSF = reduction in fertilizer 23 scenario, RSS = improved scenario, RSE = reduction of atmospheric N emissions, 24 RSC = combination of RSF, RSS an RSE. Source: (Qu & Kroeze 2012). 25 26 1.b.iii. Other pollutants 27 28 Total pesticide used (in tonnes per year) is increasing globally. Next to this the potency of 29 many newer pesticides and diversity of pesticides is increasing, as well as the use per unit 30 area. For example in the USA, an increase in pesticide application 5%-14% is estimated

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1 (Koleva & Schneider, 2009). However, the use of some pesticides is expected to decline or 2 stabilize; botanical carbamate, neonicotinoid, inorganic isoxazolidinon and triazine (Koleva & 3 Schneider, 2009; Figure 8.15). The increase of pesticide application is expected to differ 4 between crops; the lowest average increase is expected in fruit and vegetables: 5% in 2030, 5 while the biggest average increase is expected in cereals: 10% (Koleva & Schneider, 2009). 6 For the EU, the median predicated level of application for the 2090 scenario is expected to 7 be 2,4-fold that of 1990 (Kattewinkel et al., 2011). Therefore a minimum increase in the level 8 of application of 22% is predicted (Kattewinkel et al., 2011). 9 10 Despite development of newer pesticide, the most common used pesticides still remain the 11 choice of small farmers because they are cost-effective, easily available, and display a wide 12 spectrum of bioactivity (De et al., 2014). Next to this the toxic effect of pesticides will not 13 disappear directly from the environment after reducing the use. For example, if we stop all 14 DDT emissions now, only in the tropical regions the environmental levels are affected in a 15 significant way (Schenker et al., 2008). In the temperate regions, the environment is still 16 affected for up to 20 years and in the Arctic for about 50 years (Schenker et al., 2008). 17

18 19 Figure 8.15. Statistical extrapolation of insecticide use to 2020. Long dashes represent extrapolation 20 period. Short dashes represent 95% confidence bounds. Horizontal dashed grey line represents 21 model-estimated 2010 value for indicator. Extrapolation assumes underlying processes remain 22 constant. Source: (FAO, 2014) 23 24 Oil spills – The current decreasing trends of oil spills in the ocean are expected to continue 25 (Figure 8.16). This includes spills from accidents involving tankers and accidental loss of oil 26 during operations (GESAMP, 2007). On land there is an expected increase in loss of oil from 27 pipelines and installations (Caineng et al., 2010, Jernelöv, 2010). The trend in large-scale 28 losses of oil to the environment from oil exploration and production at the frontiers of what 29 is currently technically feasible is less easy to predict (Caineng et al., 2010, Jernelöv, 2010)

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1 2 Figure 8.16. Projected oil spills up to 2020 extrapolated from data (reference to be added). 3 4 1.c. Country actions and commitments3 5 6 Most countries have established national targets, or similar objectives, related to Aichi 7 Biodiversity Target 8 in their NBSAPS. These targets are generally in line with the Aichi 8 Biodiversity Target. Further, several countries, such as Ireland, have identified actions to 9 assist with the implementation of their target. The majority of targets which have been 10 developed refer to reducing pollution generally. Few targets refer to reducing excess 11 nutrients specifically. Overall, if implemented, the national targets and other mechanisms 12 contained in the NBSAP would take a significant contribution towards the attainment of the 13 Aichi Biodiversity Target. 14 15 16 2. What needs to be done to reach the Aichi target? 17 18 In this section we only describe actions and costs to reduce N and P pollution, however there 19 are many actions possible to reduce the impacts of from other sources. 20 21 2.a. Actions 22

3 This assessment is based on an examination of the national biodiversity strategies and action plans from the following countries: Australia, Belarus, Belgium, Colombia, Democratic People's Republic of Korea, Dominican Republic, El Salvador, England, The European Union, Finland, France, Ireland, Japan, Malta, Myanmar, Serbia, Spain, Suriname, Switzerland, Timor Leste, Tuvalu and Venezuela. In addition it considers the set of national targets developed by Brazil. This assessment will be further updated and refined to account for additional NBSAPS and as such these initial findings should be considered as preliminary and were relevant a level of confidence has been associated with the main statements. This assessment focuses on the national targets, objectives, priority actions and similar elements included in the NBSAPs in relation to the international commitments made through the Aichi Biodiversity Targets.

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1 Pollution is one of the main direct causes of biodiversity loss. Pollution can take numerous 2 forms as a variety of chemical compounds can cause environmental damage depending on 3 their properties and concentrations. Nutrient pollution has generally stabilized in parts of 4 Europe and North America but is projected to increase in other regions, and remains a 5 significant threat to aquatic and terrestrial biodiversity. Other forms of pollution, such as 6 from other chemicals, pesticides and plastics, are increasing. Despite the progress that has 7 been made, GBO-4 found that, currently, progress is not sufficient to achieve the target. 8 Against this background, possible key actions to accelerate progress towards this target 9 include: 10 11 (a) Developing and enforcing national water quality guidelines and/or 12 concentration thresholds for different pollutants; 13 14 (b) Improving fertilizer use efficiency (Target 7); 15 16 (c) Eliminating phosphates from detergents; 17 18 (d) Enhancing treatment and recycling of sewage and water; 19 20 (e) Conserving and restoring wetlands and other ecosystems which play an 21 essential role in nutrient cycling (Targets 5,11, 14 and 15); and 22 23 (f) Promoting the reuse and recycling of plastics, and the use of biodegradable 24 alternatives, to reduce . 25 26 The programmes of work on agricultural biodiversity, the biodiversity of inland waters and 27 marine and coastal biodiversity provide some guidance on the types of actions that could be 28 taken to reach this target. 29 30 As described above, the main source of N and P pollution is fertilizer use in food production 31 systems, however, fossil fuels and sewages are also a source of N and P pollution. Since 32 fertilizer is essential for increasing food production (Seitzinger et al., 2010) and all scenarios 33 illustrate an increasing of per capita food consumption and GDP, N and P pollution are 34 expected to increase. Thus the challenge is to identify actions that decouple N and P 35 pollution from human development (PBL, 2012). Given that N and P pollution also have many 36 negative impacts on human health, there are substantial opportunities for net benefits by 37 controlling N and P pollution for biodiversity and human well-being (Grinsven et al., 2013) 38 39 There are a number of well-studied actions that can reduce N and P pollution of terrestrial 40 and aquatic ecosystems (Galloway et al., 2008, Seitzinger et al., 2010). Some of the actions 41 are primarily technical, while others require behaviour change. Some actions should result in 42 net gains for nearly all actors (better matching of N and P fertilization with crop demands), 43 while others will require considerable investments for the sectors concerned (e.g., P removal 44 in sewage plants). These actions can intervene in different parts of the N cycle. Some actions 45 focus on the increase of nutrient efficiency use to reduce nutrient loss from agricultural 46 areas. Others aim to change consumer behavior or recycling to reduce nutrient loss. The 47 major actions are outlined below in Table 8.1 and 8.2.

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1 2 3 4 5 Table 8.1. Actions to reduce nitrogen pollution (Sutton et al., 2013) Sector Action Transport Improve fuel efficiency and reduce the amount of

emissions per combustion to reduce vehicle NOx emissions per km. Reduce fossil fuel use by minimizing vehicle use and promoting and increase the use of alternative energy sources.

Industry Reduce NOx emissions, especially from power plants using a variety of proven methods (Frost et al., 2006). Agriculture Increasing nutrient efficiency by adapting N fertilizer application to needs of crops and climate. This will not only reduce nutrient loss but also increase crop production which will help meeting target 7. Decrease production by reducing meat and milk production and consumption where appropriate. Reduce nutrient loss from agricultural sites by erosion control measures, maintain or restore wetlands and vegetation along rivers and streams, increase the use of legume/non-legume mixtures, cover crops, tillage management, buffer strips and minimize use of high N demanding crops such as maize. Minimize emissions from animal housing, manure storage and handling. Couple crop and livestock systems by better integration of animal manure in crop system. Sewage and recycling Expand and improve sewage treatment plants Recycle and reduce waste by improving the overall food supply chain and change consumption patterns. Recycle organic waste containing N by sewage treatment plants. 6 7 Table 8.2. Actions to reduce phosphorus pollution (Sutton et al., 2013) Sector Action Industry Increase efficiency of phosphorus mining and processing. Agriculture Reduce nutrient loss from agricultural sites by erosion control measures, maintain or restore wetlands and vegetation along rivers and streams, increase the use of legume/non-legume mixtures, cover crops, tillage management and buffer strips.

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Minimize emissions from animal housing, manure storage and handling.

Couple crop and livestock systems by better integration of animal manure in crop system. Sewage and recycling Eliminate phosphates from laundry and dishwasher detergents. Recycle organic waste containing P by treating sewage and wastewater before it reaches ground or surface waters and by improving P removal at sewage treatment plants. 1 2 Some of the solutions to N and P pollution are primarily socio-technical. Many countries have 3 regulations concerning sewage treatment, and this is projected to become more systematic 4 in scenarios with strong socio-economic development. Next to this many developed and 5 emerging countries have air quality regulations that aim to reduce NOx emissions from 6 vehicles and industry. These regulations are generally getting more stringent over time, but 7 need to be uniformly enforced. Efforts must be made to get older vehicles out of circulation 8 and to minimize vehicle traffic (Seitzinger et al., 2010). 9 10 Ecosystem services could be used to reduce N pollution. For example wetlands, including 11 constructed wetlands, play a major role in sequestering nutrients by retaining sediment 12 particles (which often have adsorbed phosphorus) and facilitating denitrification of reactive 13 to non-reactive N. However it should be noted that denitrification could also result in the 14 formation of N2O, which can increase climate change. Next to this, wetlands continue to be 15 lost worldwide through drainage and channelization (Carpenter et al., 2011). Protecting 16 (targets 5 and 11) or restoring (targets 14 and 15) wetlands could therefore make an 17 important contribution to achieving target 8. 18 19 However, the most important projected changes in N and P pollution over the coming decade 20 will arise from agriculture. Sustainable agricultural practices are the most important key to 21 reducing N and P pollution. By achieving target 7, through increasing nutrient use efficiency, 22 while simultaneously increasing agricultural production would make a major contribution to 23 achieving target 8. The key for balanced fertilization is efficient recycling of N from animal 24 . Bouwman et al. (2013) estimate a potential reduction of the gross nitrogen balance 25 in 2050 of 25% when N in manure is recycled. Balanced fertilization (defined as a balance 26 between N fertilization input and plant demand) is a promising measure. This measure can 27 reduce N leaching by 39% in 2020 relative to 2000 in Europe (Table 8.3; Oenema et al., 2009). 28 Smaller effects can be achieved with low-protein animal feeding, reduced meat consumption 29 diet, reducing waste and post-harvest losses and ammonia emissions abatement measures 30 (Oenema et al., 2009). 31 32 However, some policies can also cause increasing nutrient pollution, for example biofuel 33 policies. Increased demand for biofuels could increase global fertilizer use to satisfy 34 additional production needs. For example with a 50% increase in the biofuel mandate, N2O 35 emissions from fertilizer use could rise by 1.4% (Mosnier et al., 2012). 36

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1 Nevertheless, due to the wide range of N and P sources and the many ways to improve their 2 management there is no single solution. A package of measures is needed to address 3 nitrogen and phosphorus systems as a whole (See box 8.1; Sutton & Bleeker, 2013). For 4 example in the USA the package of legislations to control nitrogen pollution consists of 5 efforts to decrease sewage pollution, , , nitrogen inputs in coastal systems and 6 runoff from crop and animal production (EPA, 2011). The legislations in the EU combine 7 balanced fertilization, low-protein animal feeding and ammonia emissions abatement 8 measures. This could potentially reduce N leaching by 41% by 2020 in Europe (Table 8.3; 9 Oenema et al., 2009). Nevertheless the N leaching levels remain generally above critical 10 loads levels, illustrating the big efforts needed to be able to reduce nutrient pollution below 11 the critical loads. Next to this there can be unintended consequences when focusing on one 12 pollutant only, since not all practices are effective for both N and P pollution reduction. 13 Therefore an integrated strategy is needed to reduce both the effect of N and P pollution on 14 biodiversity (EPA, 2011). 15 16 Table 8.3. Percentage of N emissions reduction in Europa by 2020 when different policy options are 17 enforced (Oenema et al., 2009). N-

NH3 loss leaching 2000 Reference year (million kg) 2873 2782 2020 Baseline scenario -10% -10% 2020 Balanced fertilizer in designated nitrate vulnerable zones (NVZ) -14% -31% 2020 Balanced fertilizer in whole EU-27 -18% -39% 2020 Balanced fertilizer in NVZ and low-protein animal feeding -20% -34% Balanced fertilizer in whole EU-27 and low-protein animal

2020 feeding and NH3 emissions abatement -31% -41%

2020 NH3 emissions abatement -21% -9% 18 19 Begin Box 8.1 - European nitrogen legislation 20 The EU legislation to reduce nitrogen loading consists of actions to reduce atmospheric 21 deposition and leaching. The three most important pieces of EU legislation for reducing the 22 nitrogen loading to ecosystems are: 23 1. The Nitrates Directive (1991/676/EEC) which caps the total application of N from 24 animal manures to 170 kg N/ha and restricts application of manure and inorganic 25 fertilizer in situations with high risk of N loss. 26 2. The National Emissions Ceilings Directive (NECD 2001/81/EC) caps emissions of NH3 27 and NOx at national levels to reduce acidification and eutrophication. This directive 28 also defines best management practices to reduce ammonia losses. 29 3. The Urban waste water treatment directive (1991/271/EEC) sets targets for N- 30 removal efficiencies. 31 32 Due to these and other regulations, ammonia emissions in the EU27 declined 30% between 33 1980-2011 (EMEP, 2013). On average, the gross nitrogen balance (i.e., an indicator of losses 34 to the environment) decreased in the EU27 between 1980-2005 by 36% (Bouwman et al., 35 2013). Before 2000, when European directives were not yet implemented, the Gothenborg 36 protocol (regulations for reducing acid rain and ) resulted in major reductions in 37 N emissions. Emissions reduction effects of the NECD and Nitrates Directive after the year

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1 2000 were small (Velthof et al., 2014). However, individual EU member states with strict 2 national nitrate and ammonia policies that arose due to serious N pollution problems from 3 intensive agriculture, e.g. Denmark, Belgium and The Netherlands, achieved higher reduction 4 of ecosystem loadings, although levels generally remain well above those that cause 5 ecological damage (Grinsven et al., 2012). Nevertheless, total N loads to rivers in EU 15 have 6 remained relatively high and stable since 1990, although for some rivers such as the Rhine 7 there has been substantial improvement (Bouraoui & Grizzette, 2011). 8 End box 8.1 9 10 2.b. Costs and Cost-benefit analysis 11 12 Reducing N and P pollution results in several costs, however not reducing N and P pollution 13 also leads to high costs (Table 8.4 and 8.5). For example, the damage costs caused by harmful 14 algal blooms due to N and P pollution might range between US$37 million and US$72 million 15 for the fishery industry in the month following a (Compton et al., 2011). When 16 including the damage for recreation, waterfront real estate and recovery of species, 17 eutrophication costs US$2.2 billion per year in the USA. 18 19 The costs of nitrogen pollution in Europe are estimated at €75−€485 billion per year 20 (Grinsven et al., 2013). The environmental costs only range between €13 - €65 billion per 21 year (Erisman et al., 2013). According to the US acid rain program, acidification and damage 22 to material due to N oxides cost US$133 million annually (Compton et al., 2011). In the UK it 23 is estimated that the cost of ecosystem services loss due to nitrogen runoff is €0.3 per kg N 24 (Brink & Grinsven, 2011). This has been estimated that the costs to restore the occurrence of 25 disappeared target species are €2.5 per kilo NOx, €2.3 per kilo NH3 (Brink & Grinsven, 2011). 26 27 As illustrated in Table 8.1 and Table 8.2, P and N pollution can be reduced by several actions, 28 however implementing these actions also incurs substantial expenses. For example, actions 29 to reduce nitrogen runoff from agricultural sources have estimated costs of US$26.80 per kg 30 of nitrogen runoff avoided (Talberth & Gray, 2012). The costs include the implementation of 31 cover crops, riparian forest buffers, grass buffers, water control structures and animal waste 32 management create. 33 34 Next to this, to reduce N and P pollution in water and to reduce the number of dead zones, 35 wetlands should be maintained or restored, upstream wastewater treatments plants should 36 be increased in regions that currently lack treatment or sewage and wastewater can be 37 recycled. Increasing wastewater treatments plants has different costs for N and P pollution. 38 For N pollution it has been estimated that this could cost between US$6.13 - US$9.33 per 39 person per year, resulting in US$0.27-US$0.41 billion per year (Talberth & Gray, 2012). 40 Wastewater treatment plants for P pollution could cost between US$8000-15000 per tonne 41 of phosphorus (koppelaar & weikard, 2013). This is 30-40 times the more than the cost of 42 phosphate rock production (Koppelaar & Weikard, 2013). However, recent technological 43 developments have reduced the cost of wastewater treatments to US$5300-US$8200 tonne 44 phosphorus (koppelaar & weikard, 2013). Also after P has been removed from the water, 45 could cost up to US$ 7280 (Koppelaar & Weikard, 2013). Besides technical 46 solutions, wetlands can also remove P from wastewater. The capital expenditures necessary 47 to implement a constructed are competitive with those for alternative technologies

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1 (Kadlec, 2006). The costs of the land, earthmoving and water conveyance and control are to 2 be compared to these for tanks, pumps, chemicals and sophisticated controls (Kadlec, 2006). 3 Wetlands also need maintenance (inspections, vigor and health of vegetation, damage from 4 muskrats and beavers), however these activities are a factor ten less than required for 5 mechanical treatment (Kadlec, 2006). 6 However, addressing N and P pollution also leads to several benefits. Decreased N and P 7 pollution reduces the costs of treating water, increases human health, increases recreational 8 opportunities, improves fish habitat and health, increases property values, avoids costs 9 associated with dredging and finding water supply substitutes, and increases aesthetic and 10 existence values for biodiversity (Talberth & Gray, 2012, HLP 2014). For example reduced 11 sludge handling and eutrophication could save up to US$ 6.61 per lb. P removed (Seymour, 12 2009). In the UK the decrease of reactive nitrogen in the atmosphere by 25% has resulted in 13 a net benefit (Equivalent Annual Value) of £65 m (£5 m to £123 m, 95% CI; Jones et al., 14 2013). In the USA it was estimated that the commercial timber industry would benefit about 15 US$800 million annually by reduced nitrogen oxides, because nitrogen oxides contribute to 16 ozone formation, which can reduce forest and crop production. This also results in an 17 estimated US$700 million benefit of grain crops (Compton et al., 2011) and US$650 18 prevented agricultural productivity loss per year (Koppelaar & Weikard, 2013). 19 20 Improving sanitation improves human health, increases productivity and prevents deaths. 21 Therefore every dollar spent on improving sanitation results in an average economic benefit 22 of US$8.5 (Hutton & Haller, 2004). Even though increasing sanitation could cost up to 23 US$0.41 billion per year, it would result in a benefit of US$1.8-US$2.87 billion per year 24 (Talberth & Gray, 2012). Overall environmental benefits resulting from treatment of 25 wastewater can result into €50034733 per year or €0.245 per m3 wastewater treated 26 (Hernández-Sancho et al., 2010). 27 In addition to these benefits, the eliminated agricultural subsidies leading to excess nutrients 28 (target 3) could be used to fund the technical and behavioural changes needed to meet this 29 target (US$350 billion per year; Talberth & Gray, 2012). 30 31 Table 8.4. Costs of nitrogen pollution and the costs and benefits of reducing nitrogen pollution 32 (Talberth & Gray, 2012) Cost of nitrogen pollution Cost European nitrogen pollution cost €75 - 485 billion per year European environmental costs €13-65 billion per year Algal blooms for fisheries industry US$37 - US$72 million per Recreation US$2.2 billion annually Acid rain program US$133 million annually Ecosystem services loss €0.3 per kg N

Restore species €2.5 per kilo NOx

Restore species €2.3 per kilo NH3 Costs of action Cost Reducing dead zones by sanitation development US$0.27-$0.41 billion per year Reducing dead zones by reducing N runoff from agriculture US$26.80 per kg nitrogen runoff reduced Benefits from reduced nitrogen pollution Benefit Reduced ozone formation: forestry US$800 million annually

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Reduced ozone formation: grain crops US$700 million annually Reduced reactive nitrogen in atmosphere £65 m Equivalent Annual Value Improved health by improving sanitation US$1.8-US$2.87 billion annually Reduced agricultural subsides US$350 billion annually 1 2 Table 8.5. Cost of phosphorus recycling (Hernández-Sancho et al., 2010, koppelaar & weikard, 2013, 3 Molinos-Senante et al., 2011, Seymour, 2009) Cost of action Cost Non-food recycling US$ 7030 per tonne P Wastewater recycling US$ 6750 per tonne P €2066565 per year No-till farming US$ 1030 per tonne P Capital costs wetlands US$ 50 per tonne P Incineration of nutrient retrieved from wetland US$ 7280 per tonne P Benefits from reduced P pollution Benefits Prevented cost (sludge handling/ eutrophication) US$6.61 per lb. P removed Overall benefits from wastewater treatment in €50034733 per year Spain €0.245 per m3 water treated Overall benefits from wastewater treatment in €906718 per year Valencia €1.022 per m3 wastewater treated Prevented agricultural productivity loss US$ 650 million per year 4 5 6 3. What are the implications for biodiversity in 2020? 7 8 3.i. Terrestrial biodiversity 9 10 The N and P pollution is not expected to decrease by 2020, as well as their impact on 11 biodiversity. Especially (sub)tropical moist broadleaf forest temperate forest and ecosystems 12 in northern Europe and North America are affected (Azevedo et al., 2013b). For example N 13 deposition can cause soil acidification and foliar damage, especially in mosses and lichen, 14 resulting in degradation of plant diversity (Figure 8.17b; Bobbink et al., 2010, Sutton et al., 15 2013, Sutton et al., 2011). High concentrations of atmospheric ammonia result the 16 replacement of local and regional lichen biodiversity with a few species able to profit from 17 these conditions (Sutton et al., 2013). Due to acidification, oligotrophic species can be out- 18 competed by more nitrogen-loving or acid-tolerant plants. Chronically elevated nitrogen 19 deposition can also enhance susceptibility to stress such as frost damage, herbivory or 20 disease (Dise et al., 2011). 21 22 Also the decline of species richness in grassland is expected to continue due to continued N 23 deposition increase. Over the past 70 years, species richness in Great Britain, Germany and 24 the Netherlands have declined significantly due to N deposition N pollution (Figure 8.17a; 25 Stevens et al., 2010). Especially in Southern and Eastern Asia, where high increases of N 26 deposition are expected, species richness in grasslands are expeced to decline. 27 28 The consequences of N deposition in arid zones are rather unclear, although some 29 indications suggest increased invasions of exotic species (target 9). This can be expected in 30 several Mediterranean ecosystems (UNEP, 2004). Most arid zones are low N-deposition

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1 regions, however this does not mean that there are at low risk. The effect of N deposition on 2 species richness seems to be curvilinear (Stevens et al., 2010). This implies that small 3 increases in N deposition will have a larger impact on species richness when background 4 deposition levels are low than when initial deposition levels are higher (above 20kg N ha-1yr-1 5 ;Stevens et al., 2010). Because at high N deposition rates many of the nitrogen-sensitive 6 species have already declined, leaving mainly the less nitrogen-sensitive species (Stevens et 7 al., 2010). Therefore areas which are undammed by N deposition have a higher potential for 8 species loss (Stevens et al., 2010). 9

10 11 Figure 8.17. A) Correlation between observed levels of N deposition in Europe and species richness in 12 acidic grasslands. Source: (Stevens et al., 2010). B) Species richness ratio (values less than one 13 indicate species loss) in experiments with N enrichment. Source: (Bobbink et al., 2010). 14 15 Next to nutrient pollution, pesticide application is also expected to influence species 16 richness. Especially birds, as non-target species, are highly impacted by pesticides. Insect- 17 eating birds are impacted by a decrease in food due to a decline in insect population in 18 agricultural areas (Mitra et al., 2011). Particularly herbicides impact food, nesting and shelter 19 availability for specialist farmland bird species (Chiron et al., 2014, Geiger et al., 2010). 20 However reduced competition with specialist bird species may result in an increase of 21 generalist species and might therefore lead to increased species richness (Chiron et al., 22 2014). 23 24 The Red List Index can be used to track trends in the net impacts of pollution and attempts to 25 manage or control these (Blacksmith institute, 2013). The Red List Index (RLI) shows trends in 26 the survival probability of sets of species. It is based on data from the IUCN Red List – the 27 number of species in each Red List category of extinction risk, and the number moving 28 categories between assessments owing to genuine improvement or deterioration in status 29 (Butchart et al., 2004, 2005, 2007). 30 The RLI can be disaggregated to show trends driven by particular types of threats, showing 31 the net balance between the number of species deteriorating in status owing to negative 32 impacts, and the number improving in status owing to successful attempts to control or 33 reduce pollution impacts. The RLI for birds showing trends driven by pollution has showed 34 shallow declines since 1988, indicating that more species are moving closer to extinction 35 owing to negative impact of pollution (Figure 8.18). Trends for other taxonomic groups are 36 likely to be similar. However, other factors such as unsustainable agriculture, logging and 37 hunting/trapping are much more significant drivers of declines than pollution in those 38 terrestrial groups for which we have trend data at a global scale: birds, mammals and 39 amphibians (see Target 12).

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1 2 Figure 8.18. Statistical extrapolation of Red List Index for birds showing trends driven by the impacts 3 of pollution or its control to 2020. Long dashes represent extrapolation period. Short dashes 4 represent 95% confidence bounds. Horizontal dashed grey line represents model-estimated 2010 5 value for indicator. Extrapolation assumes underlying processes remain constant. 6 7 3.ii. Aquatic biodiversity 8 9 Vörösmarty et al., (2010) estimated that pollution was one of the greatest threats of 10 biodiversity in rivers, on par with water resource development. However the effect of P on 11 aquatic species richness differs between types of water bodies. Rivers can withstand higher P 12 levels before species richness is affected than lakes (Azevedo et al., 2013a). Phytoplankton 13 species are less sensitive to P change than macrophytes (Azevedo et al., 2013a). Species 14 richness in cold-region lakes and rivers has the highest sensitivity to P levels, while temperate 15 streams and xeric lakes have the lowest sensitivity to P increase (Azevedo et al., 2013a). This 16 analysis illustrates the very large negative impacts of pollution on freshwater biodiversity 17 globally and that reductions in pollution levels are urgently needed. 18 19 The expected increase of pollution in aquatic ecosystems can impact on biodiversity through 20 at least three distinct pathways (see Carpenter et al., 1998, 2011, Vörösmarty et al., 2010). 21 First, addition of nutrients and toxic chemicals can lead to inhospitable conditions for many 22 species, for example through eutrophication, the development of “dead zones”, or direct 23 toxicity. Eutrophication may result in mass mortality of organisms, altered biogeochemical 24 cycling, and loss of biodiversity (Diaz & Rosenberg, 2008, Weijters et al., 2009). In the worst 25 cases, where such conditions last for several years and nutrients continue to accumulate in 26 the system, the hypoxic zone will expand and anoxia may be established accompanied by 27 release of H2S by microbial communities (Conley et al., 2009a, Diaz & Rosenberg, 2008). As 28 well as direct mass mortality of organisms, hypoxia or anoxia reduce the available habitat for

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1 marine organisms (habitat compression), potentially affecting the life cycle, local movement 2 and even large-scale migration of affected species (Craig & Crowder, 2005). 3 4 Second, a minority of native species is likely to thrive under polluted conditions. For instance, 5 populations of certain aquatic plants often explode in lakes and rivers as nutrient availability 6 is increased. Though these species may be a natural part of the system, the change in 7 conditions can enable them to dominate the community. An example are harmful algal 8 blooms (HABs), which occur as a result of eutrophication (Andam et al., 2008, Beveridge et 9 al., 2013). Their impacts vary depending on the species of or ciliate, the nutritional or 10 physiological status of the organism, its stage of life history and its concentration in the 11 water. Effects include acute toxicity causing mass mortality of marine life, including 12 invertebrates, fish, and birds but also other organisms such as cetaceans, pinnipeds and 13 sirenians (Bouwman et al., 2005, Landsberg, 2002, Silvagni et al., 2005). 14 15 Finally, nutrient loading and other stressful chemical shifts can enhance the likelihood that 16 aquatic ecosystems will be invaded by exotic species. Such invasions can directly or indirectly 17 harm native species, leading to further losses of biodiversity (more detailed information can 18 be found in the chapter about target 9). 19 20 21 4. What do scenarios suggest for 2050 and what are the implications for biodiversity? 22 23 For many pollutants no projections for 2050 are available. However, a large number of 24 scenarios exist for N and P use and emissions (Bouwman et al., 2005, PBL, 2012, van Drecht 25 et al., 2009). Similar to the projections up to 2020 the 2050, business-as-usual scenarios 26 project that N and P surpluses at the global scale will not decline in the coming decades 27 (Bouwman et al., 2005, PBL, 2012, Seitzinger et al., 2010). 28 29 The expected global increase in N deposition by 2020 is expected to continue, leading to an 30 expected increase of, on average, 35% by 2050 (Balmford et al., 2005, Bouwman et al., 2009, 31 Hardcastle & Hagelberg, 2012, PBL, 2012, van Vuuren et al., 2010). The expected variation of 32 nutrient pollution in world regions by 2020 is expected to be maintained up to 2050. 33 Decreasing trends are expected in North America, Western Europe and Japan, while in India 34 stabilization is expected, due to a decrease in fertilizer inputs and animal manure production 35 (Bouwman et al., 2005). In most developing countries an increase of N and P emissions is 36 expected due to continued population and economic growth, , development of 37 sewer systems, lacking wastewater treatment, and increasing food production (Bouwman et 38 al., 2005). 39 40 If the nitrogen use efficiency is kept at a constant level, the use of nitrogen fertilizers is 41 projected to grow from 82 million tonnes in 2000 to around 120-140 million tonnes in 2050 42 (see “Constant Nitrogen Efficiency” scenario from IAASTD, 2009). However, the projected 43 increase of fertilizer use differs between studies: Galloway et al. (2004) projected that N 44 fertilizer use will range from 140 Tg N/yr to 200 Tg N/yr, while Bouwman et al., (2009) 45 calculated that N fertilizer use will range from 157 Tg/yr to 231 Tg/yr by 2050. This difference 46 in projected fertilizer use is related to enhanced productivity, expansion of agricultural land 47 and growing demand for agricultural products, especially in Sub-Saharan Africa, Southeast

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1 Asia and Latin-America (Bouwman et al., 2005, IAASTD, 2009, PBL, 2012, van Vuuren et al., 2 2010). Increased fertilizer use does not always have a negative influence on the environment, 3 as agriculture without fertilizer use on nitrogen deficient soils can lead to soil degradation. 4 Increased fertilizer use at moderate levels may therefore prevent in 5 nitrogen deficient soils. 6 7 Similar to the 2020 projections the transport and industrial emission are expected to rise up 8 to 52 Tg N yr-1 (Galloway et al., 2004). However these nitrogen emissions will be lower in 9 scenarios with emphasis on climate change mitigation than in baseline scenarios, due to a 10 shift from fossil fuels to renewable fuels (PBL, 2012). 11 12 The P use increase is expected to slow down after 2020 due to an expected stabilizing global 13 population, higher efficiency rates in fertilizer use and improved agricultural management 14 (Van den Berg et al., 2011, van Vuuren et al., 2010). Nevertheless, the P surplus is expected 15 to range from 18 Tg a_1 to 35 Tg a-1 by 2050 (Bouwman et al., 2009). Similar to 2020 the 16 increase of global P use is primarily attributed to fertilizer use (40% increase between 2010- 17 2050), which projected to increase in developing regions and decrease in Europe and 18 Northeast Asia (Van den Berg et al., 2011, van Vuuren et al., 2010). Also P emissions from 19 detergents and in livestock increase in all scenarios in developing countries, due to an 20 increase in laundry and dishwasher detergents in Asia, South and Central America (Van den 21 Berg et al., 2011, van Vuuren et al., 2010). 22 23 Up to 2050 wastewater treatment will keep influencing N and P emissions. With the 24 expected population growth in developing countries, sewage N and P discharge to surface 25 water is projected to increase substantially between 2000 and 2050, especially in southern 26 Asia, where increases of N and P emissions up to a factor 4–5 are foreseen. This results from 27 a combined effect of increasing numbers of people, urbanization, and enhanced sewerage 28 connectivity. Also North America (with high levels of N and P removal), rapid population and 29 economic growth may also lead to increasing sewage N and P discharge to surface water (van 30 Drecht et al., 2009). Similar to the 2020 scenarios the optimistic scenarios for the 31 development of wastewater treatment systems expect N and P effluents to remain above 32 limits that would avoid eutrophication. 33 34 As outlined in section 2, several actions can reduce N and P pollution. Several scenarios have 35 explored the effect of the key actions on the N and P pollution in 2050. These scenarios can 36 be divided into three types of scenarios: i) increasing resource efficiency of crops and 37 livestock ii) reducing post-harvest losses and promoting dietary changes and iii) recycling of 38 fertilizer and human excreta. 39 40 In the first type of scenarios, integrating animal manure better into agricultural systems can 41 reduce P fertilizer use by 10% in the USA and by 12% in Europe (Van den Berg et al., 2011). 42 Technological developments might result in even higher nutrient efficiency, which can 43 decrease the N and P surplus, as shown by the “Rio+20 Global technology” scenario (Figure 44 8.19 and 8.20; PBL, 2012). However, due to increased food demand the IAASTD (2009) 45 indicates that the fertilizer use is still expected to increase up to 110-120 million tonnes in 46 2050. 47

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1 In the second type of scenarios, a decrease in meat consumption lowers the demand for 2 fertilizer, since reduced meat demand in human diets also leads to decreased requirements 3 for animal feed and associated fertilizer use (van Vuuren et al., 2010). In the PBL “Rio+20 4 Consumption Change” scenario it is projected that the reduction of meat consumption 5 reduces fertilizer use in industrialized countries. Global nitrogen use is reduced by 10% and 6 phosphorus use by 11% (Figure 8.19 and 8.20); however in many regions the reduction is 7 bigger (e.g. for phosphorus 16% in North America and 22% in western and central Europe; 8 Bouwman et al., 2009, PBL, 2012).

9 10 Figure 8.19. The regional N surplus in 2050 based on the RIO+20 “baseline” and three alternative 11 socio-economic scenarios (PBL, 2012). Due to higher nutrient efficiency (scenario “global 12 technology”) and by reducing human excreta and fertilizer surplus (scenario “decentralized solution”) 13 nitrogen surpluses can be reduced. Reducing meat consumption (scenario “consumption change”) 14 leads to a reduction of global nitrogen surpluses of 10%.

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1 2 Figure 8.20. The regional P surplus in 2050 based on the RIO+20 “baseline” and three alternative 3 policy scenarios (PBL, 2012). Due to higher nutrient efficiency (scenario “global technology”) 4 phosphorus surpluses can be reduced. Reducing human excreta and fertilizer surplus (scenario 5 “decentralized solution”) could lead to global reduction of P surpluses by 15%. Reducing meat 6 consumption (scenario “consumption change”) allows global P surpluses to be reduced by 11%. 7 8 In the third type of scenarios, recycling of human N and P from households results in a lower 9 N and P increase compared to the "business-as-usual" scenarios (Bouwman et al., 2005, van 10 Vuuren et al., 2010). About 30% (3.1 million tonnes) of the P requirement for crop 11 production can be delivered by animal manure and human excreta (van Vuuren et al., 2010). 12 This could lead to a reduction of global P fertilizer use of 15% (Van den Berg et al., 2011). In 13 developing countries with nutrient deficit soils, better integration of animal manure is 14 considered difficult because inorganic fertilizer use is minimal and animal manure already 15 plays an important role in sustaining crop production. 16 17 All three types of scenarios reduce the N pollution, and thereby the effect on biodiversity 18 (Figure 8.21). The PBL “Rio+20 Consumption Change” scenario indicates that the impact of 19 nitrogen deposition on biodiversity (as measured by Mean Species Abundance) can be 20 reduced by 50% due to the recycling of N (PBL, 2012). Nevertheless the impact of N 21 deposition on biodiversity at the global level is limited (about 0.4 – 0.7% reduced MSA loss). 22

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1 2 Figure 8.21. MSA loss due to nitrogen pollution in 2050 in different scenarios. 3 4 5 5. Uncertainties 6 7 The trend in nutrient deposition values are well studied in Europe and the USA. However, 8 more routine measurement in South America and Africa are needed to reduce uncertainties 9 (Lamarque et al., 2013). The model estimations for the nutrient deposition also include 10 uncertainties. Dentener et al., (2006) illustrated that the wet deposition of nitrate is in 11 relative good accordance with the measurements, with 60–70% of the model calculated 12 deposition agreeing within ±50% with measurements. However in South America the models 13 have a low spatial correlation and in India the nitrate deposition is with a factor 2 14 underestimated (Dentener et al., 2006). The NHx deposition estimates show a small bias in 15 all continents. However in East Asia the measurements are 50% higher than the model 16 estimates, while in India the model overestimates the measurements with a factor 2 17 (Dentener et al., 2006). This is also reflected in the determination of the effect of nitrogen 18 deposition on biodiversity. The effect of nitrogen deposition has well been studied in 19 temperate ecosystems. However more research is needed to establish the sensitivity of 20 tropical and sub-tropical ecosystems (Bobbink et al., 2010). 21 22 Also a major concern in determining the global water quality is ensuring good geographic 23 representation of monitoring stations and temporal coverage of the same water quality 24 parameters (UNEP-GEMS, 2008). For the WQIB data of approximately 100 countries have 25 been included, however the reporting of data is inconsistent. Not all countries provide yearly 26 or do not provide all quality parameters (UNEP-GEMS, 2008). In addition, some countries 27 only supply data from few monitoring stations, or only from impacted sites with very little 28 data from non-impacted or baseline sites (UNEP-GEMS, 2008). 29 30 31 32 33 34

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1 6. Dashboard – Progress towards Target

Element Current Comments Confidence Status Pollution has been brought to levels No clear Highly variable between pollutants. Low that are not detrimental to evaluation ecosystem function and biodiversity Pollution from excess nutrients has Nutrient use levelling off especially in High been brought to levels that are not Europe in N America, but at levels that detrimental to ecosystem function are still detrimental to biodiversity. and biodiversity Still rising in other regions. Very high

regional variation. 2 3 Compiled by Jennifer van Kolck and Rob Alkemade, with contributions from Paul Leadley, 4 Peter McIntyre, Marc Metian, Robert Diaz, Alex Rogers, Hans van Grinsven, Stephanie 5 Januchowski-Hartley, Michiel Rutgers, Leo Posthuma, Kees Versluijs and Mark Huijbregts. 6 Extrapolations: Derek Tittensoor 7 NBSAPs and National Reports: Kieran Mooney / CBD Secretariat 8 Dashboard: Tim Hirsch 9 10 Version history: 12 Dec 2013 JvK, 18 Dec 2013 CBK, 16 Jan 2014 JvK, 19 Feb 2014 JvK, 25 Feb 11 2014 PL, 29 Apr. 2014 JvK 12 13 14 7. References 15 16 Aktar W, Sengupta D, Chowdhury A (2009) impact of pesticides use in agriculture: their benefits and hazards. 17 interdisc. toxicol., 2, 1-12. 18 Andam KS, Ferraro PJ, Pfaff A, Sanchez-Azofeifa GA, Robalino JA (2008) Measuring the effectiveness of protected 19 area networks in reducing deforestation. Proceedings of the National Academy of Sciences, 105, 16089- 20 16094. 21 Anderson DM, Glibert PM, Burkholder JM (2002) Harmful algal blooms and eutrophication: Nutrient sources, 22 composition, and consequences. Estuaries, 25, 704–726. 23 Azevedo LB, Van Zelm R, Elshout PMF et al. (2013a) Species richness-phosphorus relationships for lakes and 24 streams worldwide. Global Ecology and Biogeography, 22, 1304-1314. 25 Azevedo LB, Van Zelm R, Hendriks AJ, Bobbink R, Huijbregts MaJ (2013b) Global assessment of the effects of 26 terrestrial acidification on plant species richness. environmental pollution, 174, 10-15. 27 Balmford A, Green R, Scharlemann JPW (2005) Sparing land for nature: exploring the potential impact of 28 changes in agricultural yield on the area needed for crop production. Glob. Change Biol., 11, 1594-1605. 29 Barnes DKA, Galgani F, Thompson RC, Barlaz M (2009) Accumulation and fragmentation of plastic debris in 30 global environments. Philosophical transactions of the royal society, 364, 1985-1998. 31 Beveridge M, Thilsted S, Phillips M, Metian M, Troell M, Hall S (2013) Meeting the food and nutrition needs of 32 the poor: the role of fish and the opportunities and challenges emering from the rise of aquaculture. 33 Journal of fish biology, 83, 1067-1084. 34 Bip (2014) Water quality in freshwater systems. 35 http://www.bipindicators.net/language/enus/indicators/ecosystemintegrityservices/waterquality/2010, 36 Biodiversity indicators partnership. 37 Blacksmith institute (2012) The world's worst pollution problems: assessing helath risks at 38 sites. (eds Mills-Knapp S, Traore K, Ericson B, Keith J, Hanrahan D, Caravanos J), New York, Blacksmith 39 institue Green Cross Switzerland. 40 Blacksmith institute (2013) State of the world’s birds, Cambridge, UK, BirdLife International. 41 Bleeker A, Hicks WK, Dentener F, Galloway J, Erisman J-W (2011) N deposition as a threat tot he world's 42 protected areas under the convention on biological diversity. environmental pollution, 159, 2280-2288.

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