Scholarly article on topic 'Informing policy to protect coastal coral reefs: Insight from a global review of reducing agricultural pollution to coastal ecosystems'

Informing policy to protect coastal coral reefs: Insight from a global review of reducing agricultural pollution to coastal ecosystems Academic research paper on "Earth and related environmental sciences"

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Marine Pollution Bulletin
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Abstract of research paper on Earth and related environmental sciences, author of scientific article — Frederieke J. Kroon, Britta Schaffelke, Rebecca Bartley

Abstract The continuing degradation of coral reefs has serious consequences for the provision of ecosystem goods and services to local and regional communities. While climate change is considered the most serious risk to coral reefs, agricultural pollution threatens approximately 25% of the total global reef area with further increases in sediment and nutrient fluxes projected over the next 50years. Here, we aim to inform coral reef management using insights learned from management examples that were successful in reducing agricultural pollution to coastal ecosystems. We identify multiple examples reporting reduced fluxes of sediment and nutrients at end-of-river, and associated declines in nutrient concentrations and algal biomass in receiving coastal waters. Based on the insights obtained, we recommend that future protection of coral reef ecosystems demands policy focused on desired ecosystem outcomes, targeted regulatory approaches, up-scaling of watershed management, and long-term maintenance of scientifically robust monitoring programs linked with adaptive management.

Academic research paper on topic "Informing policy to protect coastal coral reefs: Insight from a global review of reducing agricultural pollution to coastal ecosystems"

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Informing policy to protect coastal coral reefs: Insight from a global review of reducing agricultural pollution to coastal ecosystems

Frederieke J. Kroon a,b'*, Britta Schaffelke b, Rebecca Bartleyc

a CSIRO Ecosystem Sciences, P.O. Box 780, Atherton, Qld 4883, Australia b Australian Institute of Marine Science, PMB 3, Townsville MC, Townsville, Qld 4810, Australia c CSIRO Land and Water, 41 Boggo Road, Dutton Park, Qld 4102, Australia




Article history:

Available online 26 June 2014






Water quality

The continuing degradation of coral reefs has serious consequences for the provision of ecosystem goods and services to local and regional communities. While climate change is considered the most serious risk to coral reefs, agricultural pollution threatens approximately 25% of the total global reef area with further increases in sediment and nutrient fluxes projected over the next 50 years. Here, we aim to inform coral reef management using insights learned from management examples that were successful in reducing agricultural pollution to coastal ecosystems. We identify multiple examples reporting reduced fluxes of sediment and nutrients at end-of-river, and associated declines in nutrient concentrations and algal biomass in receiving coastal waters. Based on the insights obtained, we recommend that future protection of coral reef ecosystems demands policy focused on desired ecosystem outcomes, targeted regulatory approaches, up-scaling of watershed management, and long-term maintenance of scientifically robust monitoring programs linked with adaptive management.

Crown Copyright © 2014 Published by Elsevier Ltd. This is an open access article under the CC BY-NC-ND

license (

1. Introduction

The continuing degradation of coral reefs around the world (Bruno and Selig, 2007; De'ath et al., 2012; Gardner et al., 2003) has serious consequences for the provision of ecosystem goods and services to local and regional communities. While climate change is considered the most serious risk to coral reefs around the world, agricultural pollution threatens approximately 25% of the total global reef area (Burke, 2011) (Fig. 1). To ensure the future of coral reefs, the 2012 Consensus Statement on Climate Change and Coral Reefs has called for the immediate management of local anthropogenic pressures including reducing land-based pollution (12th International Coral Reef Symposium, 9-13 July 2012). Attempts are being made to reduce land-based pollution to coral reefs (Brodie et al., 2012; Richmond et al., 2007), however, these efforts are impeded by a current paucity of studies demonstrating whether improvements to coral reef health are realized following watershed management. For the next 50 years, riverine fluxes of sediment, nitrogen (N) and phosphorus (P) to tropical coastal areas

* Corresponding author. Present address: Australian Institute of Marine Science, PMB 3, Townsville MC, Townsville, Qld 4810, Australia. Tel.: +61 7 4753 4159; fax: +61 7 4772 5852.

E-mail addresses: (F.J. Kroon), (B. Schaffelke), (R. Bartley).

are projected to increase (Mackenzie et al., 2002). It is therefore timely to inform coral reef policy using insights gained from global cases that were successful in reducing agricultural pollution to coastal ecosystems.

Here, we synthesize successful examples of reduced agricultural pollution that could be used as a model to improve coral reef water quality, with the assumption that improved water quality will result in a concomitant improvement in ecological health of coral reefs. Previous reviews of the problem of coastal eutrophication (Boesch, 2002; Cloern, 2001) do not include recent reports on reduced fluxes of sediment and nutrients at end-of-river (Chu et al., 2009; Duarte et al., 2009; GEF-UNDP, 2006; Pastuszak et al., 2012; Stalnacke et al., 2003; Windolf et al., 2012), and associated declines in nutrient concentrations and algal biomass in receiving coastal waters (Carstensen et al., 2006; Duarte et al., 2009; Jurgensone et al., 2011; Oguz and Velikova, 2010). Our review focuses on restoration of diffuse fluxes of freshwater, suspended sediment, and nutrients, while acknowledging the presence of other pollutants (e.g. pesticides, herbicides and heavy metals) and their potential impact on coral reef resilience (Van Dam et al., 2011). We first summarize the global evidence for changes in freshwater flow regimes and terrestrial pollutant fluxes to coastal and coral reef environments. Next, we outline how coral reefs are affected by resultant changes in water quality. We then examine the effectiveness of land-based efforts aimed at restoring 0025-326X/Crown Copyright © 2014 Published by Elsevier Ltd.

This is an open access article under the CC BY-NC-ND license (

Fig. 1. Global classification of the threat from watershed-based pollution to coral agriculture delivered by rivers to coastal waters (data from Burke, 2011).

more natural fluxes to coastal and coral reef environments and reversing ecosystem degradation. We conclude with the insights gained into effective management of agricultural pollution from multiple global examples where reductions of land-based pollution to coastal ecosystems have been achieved. Because patterns in coastal water quality data following land use change display similar trends globally (Boesch, 2002; Cloern, 2001; Mackenzie et al., 2002; Syvitski et al., 2005), we envisage that the insights from effective management examples in non-tropical systems can be successfully transferred to coral reefs.

2. Alteration of terrestrial freshwater, sediment and nutrient fluxes to coastal waters

Globally, humans have altered terrestrial fluxes of freshwater (Vórósmarty and Sahagian, 2000), sediment (Syvitski et al., 2005), and nutrients (Mackenzie et al., 2002) to coastal marine waters, including to coral reef environments (Hendy et al., 2002; Hungspreugs et al., 2002; McCulloch et al., 2003; Prouty et al., 2009; Yamazaki et al., 2011). Natural river flow regimes, including magnitude, frequency, duration, timing, and rate of change, have been modified through surface water diversion, dam construction, aquifer mining, and wetland drainage and deforestation (Vórósmarty and Sahagian, 2000). This includes modification of flow regimes in tropical coastal catchments upstream from coral reefs in both the Atlantic (Porter et al., 1999) and Indo-Pacific (Pena-Arancibia et al., 2012). Impoundments and diversion of surface water enhance evaporation and reduce run-off, altering the magnitude and timing of freshwater flows (Vórósmarty and Sahagian, 2000). In contrast, the loss of water storage capacity associated with wetland drainage and deforestation results in lower evaporation, increased runoff, and more variable hydro-graphs (Vórósmarty and Sahagian, 2000). The resulting changes in long-term net runoff have modified coastal salinity, nutrient stoichiometry and biogeochemistry (Cloern, 2001), including on coral reefs (Porter et al., 1999).

Fluxes of terrestrial sediment to coastal marine waters have been modified by humans around the world (Syvitski et al., 2005). Increases in these fluxes are due to soil erosion, associated with changes in surface runoff, deforestation, coastal development, urbanization, agricultural practices, and mining. In tropical coastal regions, annual fluxes of suspended sediment have increased by approximately 1.3 times, with 16% of the current flux retained in impoundments. This is exemplified in the Great Barrier Reef region, where a large proportion of terrestrial sediment is trapped by multiple reservoirs (e.g. 10-90% depending on flow in the Burdekin Falls Dam, (Lewis et al., 2009)). Notwithstanding, annual fluxes of

reefs, using an index based on estimated erosion and nutrient fertilizer runoff from

terrestrial sediment to the lagoon are still estimated to have increased more than fivefold since European settlement (Kroon et al., 2012) (Fig. 2). On the other hand, reductions in sediment fluxes to coastal areas are primarily due to retention within impoundments (Syvitski et al., 2005). Reservoirs now retain 26% of the global sediment flux, resulting in an overall 10% decrease compared to the prehuman sediment load (Syvitski et al., 2005). Overall, these changes in terrestrial sediment fluxes to coastal ecosystems directly affect habitat formation of benthic environments through enhanced sedimentation or coastal erosion.

Global fluxes of nitrogen (N) and phosphorus (P) to coastal areas have increased due to human activities (Cloern, 2001; Galloway et al., 2008), with a doubling of riverine, reactive N and P fluxes in the preceding 150 years (Galloway et al., 2004; Mackenzie et al., 2002), and a rise in atmospheric deposition of N from land to coastal areas (Galloway et al., 2004). Increases in these fluxes to the coastal zone are due to agricultural crop and livestock production, fertilizer application, discharge of urban and industrial sewage, and fossil fuel burning (Galloway et al., 2008), as well as removal of the ecosystems' filtering and buffering capacity (e.g. riparian zones and floodplain wetlands, (Verhoeven et al., 2006). Further substantial increases in riverine fluxes of N and P to coastal areas are projected (Galloway et al., 2004), particularly in tropical regions (Mackenzie et al., 2002). Nutrient loadings to the Great Barrier Reef lagoon, for example, have increased 6-fold for N and 9-fold for P since European settlement in the 19th century (Kroon et al., 2012) (Fig. 2). Excess nutrient inputs to coastal areas increase net primary production and lead to eutrophication (Cloern, 2001), which in extreme cases causes widespread hypoxia (Diaz and Rosenberg, 2008), and contribute to loss of ecosystem diversity, structure and functioning (Lotze et al., 2006).

3. Impacts of land based pollution on coral reefs

Modification of terrestrial pollutant fluxes, and consequent declines in reef water quality have resulted in detrimental impacts on physical, ecological and physiological processes of reef-building corals (Coles andJokiel, 1992; Fabricius, 2011). Compared to other terrestrial pollutants, the effects of changes in freshwater fluxes on coral reefs have received relatively little attention. Proxy records from coral cores indicate both enhanced (Hendy et al., 2002) and reduced (Prouty et al., 2009) freshwater fluxes into tropical waters since the late 19th century. Cases of coral mortality, bleaching and disease, associated with reduced salinity due to extreme rainfall, land runoff, and groundwater discharge, have been documented on coral reefs around the world (Coles and Jokiel, 1992). Conversely, reduced freshwater fluxes may result in increased

Fig. 2. River flood plumes carrying watershed-based pollutants reaching the Great Barrier Reef, Australia, captured by MODIS Aqua on 10 February 2007.

salinity in coastal embayments, detrimentally affecting downstream coral communities (Porter et al., 1999). While corals can generally tolerate short changes in salinity, longer exposure to salinities outside their normal range leads to reduced growth and production, with mortality occurring at both low and high salinities (Coles and Jokiel, 1992). The exact responses (including acclimation) depend on the coral species, the magnitude of salinity change compared to background levels, and the exposure time (Berkelmans et al., 2012). However, it is currently unknown whether adverse effects of salinity on coral reefs have become more frequent or extensive with alteration of freshwater flow regimes to tropical coastal waters.

Cores of reef sediment and corals have indicated both increases (McCulloch et al., 2003) and decreases (Hungspreugs et al., 2002) in terrestrial sediment fluxes to coral reefs since the 1900s. Increases in sediment fluxes can result in smothering of coral reef organisms due to the settling of suspended sediment (sedimentation), as well as in reduced light availability for photosynthesis due to turbidity caused by suspended sediment in the water column (Fabricius, 2011). Sedimentation can lead to profound changes in coral populations affecting all life history stages. High sedimentation rates may reduce larval recruitment by making the settlement substratum unsuitable (Dikou and van Woesik, 2006). After settlement, sediment composition and short-term sedimentation affect the survival of coral recruits, and inhibits growth of adult corals through reduced photosynthesis and production (Fabricius, 2011). Extensive or excessive sediment exposure can also result in coral disease (Sutherland et al., 2004) and mortality

(Victor et al., 2006), and concomitant phase shifts to macro-algal dominance have been observed (De'ath and Fabricius, 2010; Dikou and van Woesik, 2006). Recovery is possible from short-term or low levels of sedimentation (Fabricius, 2011) as the polyps of many coral species exhibit sediment rejection behavior comprising of ciliary currents, tissue expansion, and mucus production (Stafford-Smith and Ormond, 1992). The exact responses to sedimentation depend on the coral species, duration and amount of sedimentation, and sediment types (Fabricius, 2011).

Enriched signatures of N isotopes in coral cores and tissues indicate increased fluxes of terrestrial N to coral reefs from agricultural and sewage run-off since at least the 1970s (Jupiter et al., 2008; Marion et al., 2005; Yamazaki et al., 2011). Likewise, cores of reef sediment and corals have indicated an increase in terrestrial phosphorus fluxes to coral reefs in the 20th century, associated with soil erosion, sewage, aquaculture and mining operations and harbor development (Chen and Yu, 2011; Dodge et al., 1984; Harris et al., 2001; Mallela et al., 2013). Corals are mostly adapted to low-nutrient environments and increases in primary production and eutrophication due to enhanced nutrient loads can detrimentally affect corals (Fabricius, 2011). Direct effects of increased nutrients are generally at the physiological level but so far there is little evidence that this leads to coral mortality. However, indirect effects of nutrient pollution are profound. For example, phototrophic hard corals can be out-competed by other benthic primary producers in high nutrient environments, leading to the establishment of macro-algae. High nutrient availability generally leads to increases in phytoplankton populations which in extreme

cases reduce benthic light availability and cause seasonal hypoxia (Diaz and Rosenberg, 2008). Resultant organic enrichment can cause a shift to heterotrophic and/or filter feeding communities, and plays a role in driving population outbreaks of the coral-eating crown-of-thorns starfish (Fabricius, 2011), one of the main causes of coral cover declines on the Great Barrier Reef (De'ath et al., 2012). Overall, eutrophication can result in increased coral disease (Sutherland et al., 2004; Vega Thurber et al., 2013) and mortality, and contribute to loss of coral diversity, structure and function, including phase shifts to macroalgae (Fabricius, 2011).

4. Restoration of terrestrial freshwater, sediment and nutrient fluxes to coastal waters

The reduction of siltation and eutrophication of coastal marine ecosystems by better managing agricultural sources at local and regional scales is a challenge for coastal communities around the world (Boesch, 2002; Cloern, 2001), including those bordering coral reefs (Brodie et al., 2012). Globally, substantial effort is going into re-establishing environmental flows (Postel and Richter, 2003). In headwater catchments, more natural flow regimes are being reinstated through, for example, including high flows in dam releases (Rood et al., 2005) and removing small dams and weirs (Stanley and Doyle, 2003). Ecological outcomes in downstream reaches have been documented within a year, and include formation of new river channels, restored riparian vegetation, and improved fish passage and spawning habitat (Rood et al., 2005; Stanley and Doyle, 2003). Restoration of more natural flow regimes to coastal marine waters is being attempted through, for example, removal of large dams (Service, 2011), buying back irrigation water (Pincock, 2010) or agricultural land (Stokstad, 2008), and restoration of coastal floodplains (Buijse et al., 2002). Such larger-scale interventions have only commenced in recent years, and consequently, we were unable to find any documented examples of restored freshwater flow regimes into coastal waters (Table 1a). Nevertheless, while it is expected that freshwater flows should return to more natural regimes almost immediately, recovery of

associated physical and biological processes may take years to decades (Hart et al., 2002).

Despite significant investment in sediment erosion and transport control measures (Bernhardt et al., 2005), we found only one documented example of reductions in net fluxes of sediment reaching coastal marine waters following land-based restoration efforts (Tables 1b and 2). In China, targeted management of pollutant sources and pathways to conserve soil and water has contributed to reducing sediment fluxes to coastal waters (Chu et al., 2009). Soil and water conservation programs in China were first legislated in the 1950s following concern about local agricultural and industrial productivity and flooding downstream (Shi and Shao, 2000). Implementation at large spatial scales (e.g. 0.92 M km2 of land terracing, tree and grass planting, and construction of sediment trapping dams), mostly in the Yellow and Yangtze basins, has reduced sediment fluxes to coastal waters by an estimated 11.5 Gt during 1959-2007 (Chu et al., 2009).

Terrestrial fluxes of N and P to coastal waters have been reduced following management of point sources, such as waste water treatment plants, phosphate mines and P-detergents (Boesch, 2002; Cloern, 2001) (Tables 1c and 2). For example, regulation has reduced the contributions from waste water treatment plants and industrial discharges to total annual average N and P loads to the Danish coast from ~50% to <10%, and from 59% to ~20%, respectively, over 14 years (Carstensen et al., 2006). The nutrient regulation in Denmark followed lobster mortality in coastal waters in the 1980s which was attributed to algal blooms and hypoxia induced by agricultural nutrient run-off (Windolf et al., 2012). Similar declines in nutrient loads from point sources have resulted in reductions in coastal nutrient and chlorophyll a (chl a) concentrations (Greening and Janicki, 2006), enhanced benthic irradiance (Greening and Janicki, 2006), seagrass recovery (Tomasko et al., 2005), and concomitant decline in macroalgae (Cardoso et al., 2010; Vaudrey et al., 2010), including on coastal coral reefs (Laws and Allen, 1996; Smith et al., 1981). Further recovery, including to a coral-reef dominated state, may be partly constrained by nutrient sources other than point sources (Hunter

Table 1

Published evidence on the effectiveness of land-based management to restore more natural fluxes of (a) freshwater, (b) suspended sediment, and (c) nutrients, to coastal marine environments, resulting in improved coastal water quality and ecosystem condition.

Land-based management regime

Evidence for change


Riverine fluxes

Coastal water quality

Coastal ecosystem condition

Large dam removal

Coastal floodplain restoration Buy back irrigation water Buy back agricultural land

Source control, e.g. terraces, revegetation

Transport control, e.g. sediment trapping dams

Point sources STPs, P mines, P-free detergent

Diffuse sources Application control measures, e.g. reduction in N and P fertiliser use, decrease in livestock numbers and manure use, changes in land use

Transport control measures, e.g. wetland restoration

None reported effects likely to be immediate None reported effects likely to take yrs/decades

Yellow and Yangtze rivers

Yellow and Yangtze rivers

Various examples from USA, Europe

Danube, Daugava, Elbe, Leilupe, oder and Vistula rivers, Denmark, The Netherlands


None reported effects likely to take yrs/decades

None reported effects likely to take decades/centuries

Decline in nutrients, turbidity, Chl a

Decline in nutrients and phytoplankton biomass

Partial recovery of aquatic

communities, e.g. seagrass and coral

Concomitant changes in flora and fauna, but no complete recovery

None reported effects likely to take decades/centuries

Hart et al. (2002), Stokstad (2008), Pincock (2010), Buijse et al. (2002), Service (2011)

Shi and Shao (2000), Chu et al. (2009)

Boesch, (2002), Smith et al. (1981), Hunter and Evans (1995), Laws and Allen (1996), Tomasko et al. (2005), Greening and Janicki (2006)

Mee (2001), Stàlnacke et al. (2003), Carstensen et al. (2006), GEF-UNDP (2006), Duarte et al. (2009), Oguz and Velikova (2010), Hansen and Petersen (2011), Jurgensone et al. (2011), Pastuszak et al. (2012), Windolf et al. (2012) Windolf et al. (2012)

and Evans, 1995), as well as obscured by increases in human population, changes in diffuse sources and variation in freshwater discharge (Williams et al., 2010).

Reducing diffuse source loads becomes increasingly important where point source discharges comprise only a small percentage of the total N and P loads, such as in the Great Barrier Reef (GBRMPA, 2009). Major recent reviews provide recommendations to reduce excessive or inappropriate input of N and P from diffuse sources such as agriculture, fossil-fuel and animal husbandry (Canfield et al., 2010; Elser and Bennett, 2011; Galloway et al., 2008; Vitousek et al., 2009). Deliberate management of agricultural diffuse pollution has contributed to reducing nutrient fluxes to coastal waters in Denmark (Windolf et al., 2012) and The Netherlands (Duarte et al., 2009) within decades (Tables 1c and 2). Moreover, decreasing nutrient fluxes have been measured in several Eastern European rivers, namely the Danube, Daugava, Elbe, Leilupe, Oder and Vistula rivers, in the years following economic decline and associated drop in agricultural subsidies in the early 1990s (Duarte et al., 2009; GEF-UNDP, 2006; Mee, 2001; Pastuszak et al., 2012; Stalnacke et al., 2003). While not the result of a dedicated management strategy, these Eastern European examples demonstrate the magnitude of change required in agricultural management to reduce nutrient fluxes at end of river within timeframes of ten to twenty years. Subsequent declines in nutrient concentrations and phytoplankton biomass have been reported in the Western Dutch Wadden Sea and South East North Sea (Duarte et al., 2009), the Danish straits (Carstensen et al., 2006; Duarte et al., 2009), the Gulf of Riga (Jurgensone et al., 2011), and the Black Sea (Oguz and Velikova, 2010), respectively. Whilst the Danish straits and the Black Sea also show some concomitant changes in flora and fauna (Hansen and Petersen, 2011; Oguz and Velikova, 2010), complete recovery to pre-impact conditions has not been reported.

Finally, restoration of coastal ecosystems' filtering and buffering capacity is expected to enhance sediment and nutrient retention and assimilation during catchment transport processes. Improving an ecosystem's buffering capacity, for example through restoration or creation of wetlands (Verhoeven et al., 2006) and riparian zones (Tomer and Locke, 2011), can result in full recovery of N storage and cycling processes within 25-30 years (Moreno-Mateos et al., 2012). If critical nutrient loads are surpassed, however, undesirable phase-shifts can occur in these wetland and riparian ecosystems

(Verhoeven et al., 2006), potentially reducing the systems' capacity for nutrient cycling (Cardinale, 2011). Establishing more natural drainage and vegetation patterns is expected to further increase hydraulic, sediment, and nutrient residence times and enhance the opportunity for landscape mitigation of terrestrial fluxes (Burt and Pinay, 2005; Whalen et al., 2002). Enhancing an ecosystem's filtering capacity, for example through restoration of native seagrass (McGlathery et al., 2012) or oyster beds (Schulte et al., 2009), will contribute to deposition of suspended sediment, nutrient cycling and water filtration (Cloern, 2001; McGlathery et al., 2012) and may significantly reduce total sediment and nutrient loads to receiving waters (Cerco and Noel, 2010). However, despite significant investments in improving ecological filtering and buffering capacity (Bernhardt et al., 2005; Moreno-Mateos et al., 2012; Whalen et al., 2002), concomitant reductions in total pollutant loads to coastal marine waters have not been documented and may take decades to centuries.

5. Reversal of coral reef degradation

Commensurate with the lack of evidence of restored flow regimes and sediment fluxes to tropical coastal marine waters, the resultant ecological outcomes for coastal coral reefs remain unknown. Corals have the capacity to recover from short-term exposure to both low and high salinity (Coles and Jokiel, 1992), as exemplified by partial (Goreau, 1964) and complete (Egana and Disalvo, 1982) recovery of corals within days to months following mass expulsion of zooxanthellae after heavy rainfall. Similarly, following short-term or low levels of sedimentation, structural (i.e. polyp re-colonization) (Wesseling et al., 1999) and functional (i.e. photosynthetic activity) (Philipp and Fabricius, 2003) recovery within days to weeks has been demonstrated for some, but not all, coral species. Coral growth recovered within weeks following short-term enrichment of N, and of N and P combined, but not of P (Ferrier-Pages et al., 2000). It is unlikely for such swift recovery to occur following restoration of more natural freshwater, sediment and nutrient fluxes, given that coral ecosystem processes would have been chronically impacted for years to decades, if not centuries.

The well-known case of Kane'ohe Bay, Hawaii, is the only example demonstrating partial reversal of coral reef degradation

Table 2

Global examples of reductions in terrestrial fluxes of sediment and nutrients to coastal marine waters, following quantified changes in watershed management with timeframes of detection.

Pollutant Location Scale of change Flux reduction Detection timeframe (yrs) Source

Reduction in river pollutant fluxes

Sediment Yellow river Soil and water conservation ~17,000 km2 23% (0.3 Gt yr-l) 28 Chu et al. (2009)

Yangtze river Soil and water conservation >84,000 km2 12% (60 Mt yr-l) 17

Nutrients Danube river Reduction in fertiliser use (N 50%, P 70%), and livestock numbers (>50%) 60% (N), 50% (P) 8 Mee (2001), GEF-UNDP (2006)

Elbe river Reduction in agricultural N surplus (>50%), and industrial/STP emissions (N 50%, P 32%) 30% (N, P) 15 Hussian et al. (2004)

Lielupe river Reduction in fertilizer use (N 90%, P 95%), and livestock numbers (70%) Significant declines in nutrient levels 11 Stalnacke et al. (2003)

Oder river Reduction in fertiliser use (NPK 43.1%)*, livestock 25% (N), 65% (P) 20 Jankowiak et al. (2003), Pastuszak

Vistula river numbers (20.6%)*, and STP emmissions (N 27-60%, P 6173%) 20% (N), 15% (P) et al. (2012)

Denmark Reduction in fertiliser use (N 50%), and industrial/STP 51% (N), 64% (P) 20 Kronvang et al. (2008), Hansen and

emissions (N 74%, P 88%) Petersen (2011), Windolf et al. (2012))

* For whole of Poland for period 1988-1997.

following a reduction in terrestrial nutrient fluxes. Following sewage diversion in 1978, turbidity, nutrients and chlorophyll a concentrations, as well as macroalgae biomass, declined within months (Laws and Allen, 1996; Smith et al., 1981). In the next few decades, coral cover more than doubled and subsequently stabilized, however, further recovery may at least be partly constrained by nutrient sources other than sewage outfalls, by modified freshwater and sediment fluxes resulting from historical and recent changes in the Bay and its catchments (Hunter and Evans, 1995), and by additional impacts of introduced macroalgae (Conklin and Smith, 2005).

To reverse coral reef degradation, it is critical to define the different ecosystem states of a coral reef system, and understand the ecological processes that drive the change from one state to another. This relates to the concept of resilience, i.e. the capacity of an ecosystem to absorb perturbations before it shifts to an alternative state with different species composition, structure, processes and functions (Folke et al., 2004). For coral reefs, multiple alternative states can exist and have been documented for coral reefs, generally dominated by organisms other than reef-building coral (Gardner et al., 2003; Hughes et al., 2010; Mumby et al., 2007). Chronic environmental pressures such as changes in terrestrial fluxes of freshwater, sediment, and nutrients (De'ath and Fabricius, 2010; Dubinsky and Stambler, 1996; Fabricius, 2011) reduce resilience by decreasing the threshold at which the coral-dominated state shifts into a different state. A return to the more desirable coral-reef dominated state by reducing chronic drivers of change such as land-based pollution may be difficult to achieve due to the inherent stability of the degraded state, known as hysteresis (Mumby and Steneck, 2011).

6. Reducing agricultural pollution

We identified multiple examples in the global literature where reductions of land-based pollution to coastal ecosystems have been achieved (Table 2). Most examples comprise reduced nutrient fluxes from point sources, such as waste water treatment plants, through legislative mandates and regulatory enforcement (Boesch, 2002; Cloern, 2001). More recent examples also include studies demonstrating reduced sediment and nutrient fluxes from agricultural land use (Chu et al., 2009; Duarte et al., 2009; GEF-UNDP, 2006; Pastuszak et al., 2012; Stálnacke et al., 2003; Windolf et al., 2012). These examples provide us with the following insights into effective management of agricultural pollution.

First, the desired outcomes of agricultural management for coral reef ecosystems need to be clearly defined, and underpinned by knowledge of the processes that determine the trajectories of ecosystem recovery. The substantial large-scale and long-term decline in coral reef condition over recent decades (Bruno and Selig, 2007; De'ath et al., 2012; Gardner et al., 2003) has, in part, been linked to agricultural pollution. Attempts to reverse this decline, however, are generally constrained to improving agricultural and land-based pollution per se (Brodie et al., 2012; Richmond et al., 2007) without due consideration of the effort required to achieve desired outcomes for coral reefs. Consequently, many management efforts are not targeting the critical sources and ecological processes that underpin the pollution problem being remedied (Palmer, 2009). Similar to temperate systems, a return to a particular past state may be unlikely, and other perturbations such as climate change, overfishing, and invasion by non-native species may prevent a simple reversal of coastal ecosystem degradation following improvements to upstream water quality (Duarte et al., 2009; Jurgensone et al., 2011; Oguz and Velikova, 2010). Hence, when linking the implementation of agricultural management targets to ecosystem condition in reef waters, a range

of possible outcomes with associated trajectories should be considered (Palmer, 2009; Perry and Smithers, 2011).

Second, management approaches that have resulted in reduced agricultural pollution to coastal ecosystems have all been non-voluntary (Boesch, 2002; Chu et al., 2009; Cloern, 2001; GEF-UNDP, 2006; Pastuszak et al., 2012; Stálnacke et al., 2003; Windolf et al., 2012), indicating that voluntary approaches alone may not be sufficient to achieve improvements. These reductions were achieved through legislation and regulation supported by long-term political commitment (e.g. China, Denmark) (Shi and Shao, 2000; Windolf et al., 2012) or declining economic subsidies, fertilizer use and livestock numbers following the collapse of the Soviet Union (eastern Europe) (GEF-UNDP, 2006; Jankowiak et al., 2003; Pastuszak et al., 2012; Stálnacke et al., 2003). In Denmark, for example, five national action plans were implemented and enforced to improve waste water treatment, and regulate N fertilizer and manure use over two decades (Kronvang et al., 2008; Windolf et al., 2012). Each action plan in turn has been underpinned by a variety of descriptive, and sometimes quantitative, policy measures, such as strict timeframes for manure storage, a 10% cut in the optimal N quota to crops, and a decrease in arable land (Windolf et al., 2012). While specific details may differ in tropical countries, the examples from China and Europe indicate that targeted regulatory policy approaches can greatly enhance the protection of downstream coral reef ecosystems from land-based pollution.

Third, management efforts to control agricultural pollution need to be at relevant spatio-temporal scales to achieve desired ecological outcomes on downstream coral reefs. The magnitude of effort required to obtain significant pollution reductions is exemplified in non-tropical systems, including (i) (unintended) large cuts in pollutant sources (e.g. ~95% cut in fertilizer use and ~70% drop in livestock numbers in Latvian rivers (Stálnacke et al., 2003)), (ii) application at large spatial scales (e.g. 84,000 km2 of land terracing, tree and grass planting, and construction of sediment trapping dams in China (Chu et al., 2009)), and (iii) adaptive implementation over decadal time frames (e.g. >25 years in Denmark (Windolf et al., 2012)) (Table 2). Across all European rivers, substantial decreases in the nutrient input from agriculture contributed to nutrient load reductions at end-of-river. The Chinese and Danish cases further demonstrate that targeted and simultaneous implementation of a combination of measures will augment reductions of pollutant fluxes at watershed outlets. Enhanced targeting and upscaling of management efforts in agricultural systems will improve the condition of coral reef ecosystems, whilst also preventing further detrimental impacts from predicted increases in sediment and nutrient fluxes in the next 50 years.

Finally, sustained monitoring at appropriate spatio-temporal scales is required to ascertain whether agricultural management results in desired improvements of downstream coral reef ecosystems. Importantly, these monitoring programs should be driven by the development of critical questions and objectives, a conceptual understanding of linkages between desired outcomes and land-based pollution (Bartley et al., 2014), robust statistical design, and adaptive review cycles (Lindenmayer and Likens, 2009). In complex systems such as coral reefs, this would maximize the probability of detecting trends following management intervention, which could take years to decades even in comprehensively monitored systems (Darnell et al., 2012; Meals et al., 2010). Importantly, consideration of desired outcomes for coral reefs in monitoring programs will focus efforts towards detecting change in relevant metrics. For example, specific biological indicators have been identified that link changes in marine water quality to changes in the condition of coral reef ecosystems (Cooper et al., 2009). Similar metrics in upstream watersheds will enable the

assessment of progress early in the management phase and alert managers to potential unintended consequences, e.g. multiple sediment trapping dams (Chu et al., 2009) may act as flow and fish migration barriers. This emphasizes the need to maintain long-term monitoring programs to provide feedback on ecosystem condition, linked with adaptive management programs (Lindenmayer and Likens, 2009; Meals et al., 2010).

7. Conclusion

The protection of coral reefs from human pressures on regional and local scales, such as increased fluxes of freshwater, sediments and nutrients, is particularly pertinent in the context of global environmental changes, such as rising sea temperatures, ocean acidification, increase in severity of tropical storms and sea level rise (Anthony et al., 2011; Carpenter et al., 2008; Pandolfi et al., 2011). Recent research has confirmed the ongoing degradation of coral reef ecosystems around the world (Bruno and Selig, 2007; De'ath et al., 2012; Gardner et al., 2003), but global examples of watershed management demonstrating the halting or reversing of coral reef decline are not readily available. Our global review demonstrates that transformative change in agricultural management for coastal ecosystem outcomes is achievable. For coral reef ecosystems, future protection demands policy focused on desired ecosystem outcomes, targeted regulatory approaches, upscaling of watershed management, and long-term maintenance of scientifically robust monitoring programs linked with adaptive management. Implementing these recommendations will increase the resilience of desired, coral-dominated states within a timeframe (years to decades) where more extreme perturbations associated with climate change are expected.


We thank the CSIRO, AIMS and anonymous reviewers for constructive comments on earlier versions of the paper, K. Guidetti and J. MacKeen for literature searches, Matthew Slivkoff (In-situ Marine Optics) for the processed satellite image, and Irena Zagorskis and 'Reefs at Risk' for the global threat map. We acknowledge the financial support from the CSIRO and the Australian Institute of Marine Science.


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