1-s2.0-S1470160X1300383X-main (1)

8/13/2019 1-s2.0-S1470160X1300383X-main (1)

1/19
http://localhost/var/www/apps/conversion/tmp/scratch_7/dx.doi.org/10.1016/j.ecolind.2013.10.012mailto:[email protected]:[email protected]://www.elsevier.com/locate/ecolindhttp://www.sciencedirect.com/science/journal/1470160Xhttp://localhost/var/www/apps/conversion/tmp/scratch_7/dx.doi.org/10.1016/j.ecolind.2013.10.012http://localhost/var/www/apps/conversion/tmp/scratch_7/dx.doi.org/10.1016/j.ecolind.2013.10.012

8/13/2019 1-s2.0-S1470160X1300383X-main (1)

2/19
http://localhost/var/www/apps/conversion/tmp/scratch_7/dx.doi.org/10.1016/j.ecolind.2013.10.012

8/13/2019 1-s2.0-S1470160X1300383X-main (1)

3/19

Please cite this article in press as: Curtis, C.J., et al., The future of uplandwater ecosystems of the UK in the 21st century: A synthesis.Ecol. Indicat. (2013), http://dx.doi.org/10.1016/j.ecolind.2013.10.012

ARTICLE IN PRESSG Model

ECOIND-1708; No.of Pages19

C.J. Curtis et al./ Ecological Indicators xxx(2013) xxxxxx 3

Fig. 1. (a)Extent ofMMHhabitat withinthe UK predominantly uplands butalso includingvery fragmented lowland heaths(usedwith permission from theUnited Nations

Environment ProgrammeWorld ConservationMonitoring Centre and NERC. This material is based upon Crown Copyright and is reproduced with the permission of Land& Property Services under delegated authority from the Controller of Her Majestys Stationery Office, Crown copyright and database rights, EMOU206.2). (b) Exceedanceofacidity critical loads for freshwaters for 198688 using a critical ANC value of 20equiv.L1, showing wide extent of critical load exceedance during the 1980s (Source:Jane Hall, CEH Bangor). Thedataset comprises a grid-based national survey of themost sensitive headwater lakes or streams in each 10km square (20km in non-sensitivelowland regions), combinedwith later regional surveys from some of themost impacted uplandareas (Curtis et al., in press-a). Note the close correspondencebetween thedistribution of exceeded sites andMMHhabitat.

cycles), soil erosion and pollutant behaviour (specifically tracemetals and POPs);

land-use and land management change related to changesin forestry practices, animal husbandry and arable expansion,associated with socio-economic and environmental pressures,potentially causing changes in acidity, nutrient status, thermalregime and soil erosion; and

invasive species, both aquatic and riparian. These have beenrecognised as potential threats to upland water ecosystems inother parts of the world, e.g. North America (Rahel and Olden,2008)andpresentathreatintheUKbutasyetthereisnoevidencefor their presence in UK uplandwaters.

The spatial extent of potential impact from one of the biggesthistorical threats, acid deposition, is best expressed in termsof the exceedance of critical loads for acidity (Fig. 1b). Criticalload models for surface waters employ empirically based chem-ical relationships between hydrochemistry and deposition alongwith mass balance principles to determine the maximum loadof acid deposition that will not cause a critical chemical thresh-old to be crossed (Henriksen and Posch, 2001). For the UK the

critical chemical threshold is a mean annual acid neutralising

capacity of 20equiv.L1 for the great majority of sites (bothlakes and streams), linked to a 90% probability of undamagedbrown trout populations (Lien et al., 1996; Curtis et al., 2000, inpress-a). Exceedance of critical loads indicates that at long-termsteady state between deposition and surface waters, ANC willdecline below the critical value (Fig. 1). Critical loads data for UKsurface waters feed into the international integrated assessment

modelling efforts carried out under the auspices of the UN ECEConvention on Long Range Transboundary Air Pollution and theUK is one of only five countries which currently submit freshwa-ter critical loads (Curtis et al., in press-a). Almost all of the mostsensitive water bodies modelled in the UK are located in MMHhabitats.

Here we consider the evidence from this Special Issueand elsewhere for continuing and changing threats to thechemical and biological quality of upland freshwater ecosys-tems and their structure and function. We identify knowledgegaps and data requirements to assist policymakers, land-scape and environmental managers in protecting upland waterecosystems and maintaining their healthy functioning for thebenefits of both biodiversity and the human populations down-

stream.
http://localhost/var/www/apps/conversion/tmp/scratch_7/dx.doi.org/10.1016/j.ecolind.2013.10.012http://localhost/var/www/apps/conversion/tmp/scratch_7/dx.doi.org/10.1016/j.ecolind.2013.10.012

8/13/2019 1-s2.0-S1470160X1300383X-main (1)

4/19




4 C.J. Curtis et al./ Ecological Indicators xxx (2013) xxxxxx

Fig.2. Schematicdiagram showing environmentalpressuresonuplandwaters (cli-mate change, air pollution and land-use) and their key interactions. All pressuresultimately impact on macronutrient cycling, structure and function of terrestrialandaquatic ecosystems, andaffect theformandexport fluxes ofmacronutrients aswell as pollutants from catchments.

2. Future threats to uplandwaters

Most of the pressures or drivers identified above interact witheach otherandhenceare notindependentin their effects onuplandwater ecosystems and their catchments (Fig. 2). For example,changes in rainfall, temperature, storminess andcloud cover affectphotochemical reactions as well as the relative balance betweendry, occult andwet deposition of pollutants. Climate and land-usefactors (e.g. overgrazing, commercial forestry) will affect catch-ment hydrology andvulnerability to soil erosion. Airpollution andlanduse or land covermay interact, forexample by increasing pol-lutant deposition due to scavenging of air pollutants by forests, orbycausing a reduction in cover of certain plantssuch asbryophytesandlichens dueto acid deposition or excessnutrient inputs (Curtis

et al., 2005).

2.1. Air pollution and upland waters

Since uplandwatersoccur upstreamof themost intensive agri-cultural and industrial activity, contamination of their catchmentsis usually dominated by atmospherically transported pollutantsincluding sulphur and nitrogen compounds and toxic substances.Recently, industrially derived hydrochloric acid has also beenshown to have made a significant contribution to the atmosphericpollution load over much wider areas of the uplands than previ-ously thought (Evans et al., this volume). Nitrogen deposition mayact as an agent of acidification and/or of nutrient enrichment ofnaturally nutrient poor ecosystems in the uplands; it is therefore

considered in both potential roles below.

2.1.1. Acid deposition

Levelsof acid deposition touplandwaters have declinedgreatlyin the UK and many parts of Europe and North America, but theproblem has not disappeared.Major reductions in acid depositionin the UK over the last 2040 years have been described in therecently published RoTAP Report (2012) as well as the paper byCurtis and Simpson (this volume). Sulphur emissions decreasedby 94% from a 1970 peak up to 2010, with the UK meeting itstarget under the EU National Emission Ceilings Directive (NECD)(RoTAP, 2012). AcrosstheUKasawhole,resultingdecreasesintotalsulphur deposition from the start of monitoring in 1986 to 2006amountedto 80%, comprising a93%reduction indrydepositionand

a 57% decrease in wet deposition. Dry deposition shifted from the

dominant component of total S deposition in the 1980s to a minorcomponent by 2008 (RoTAP, 2012). Reductions in non-marine sul-phate in bulk deposition appear to be the main cause of increasingrainfall pH since the late 1990s (Curtis and Simpson, this volume)and while not always linear, show a broadly consistent decline.Since wet deposition has decreased less than dry deposition, andbecause total deposition reductions tend to be largest closer toemission sources in central and southeast England, the smallestreductions in total S deposition are found in the uplands wherewetdeposition is dominant (RoTAP, 2012; CurtisandSimpson, thisvolume).

Emissionsof oxidisednitrogen (NOx) decreasedby 58%between1970 and 2010, meeting the NECD target for 2010 (RoTAP, 2012).Much smaller reductions in emissions of ammonia (NH3) wereachieved from a later baseline, just 21% between 1990 and 2010.Reductions in total deposition of nitrogen have been much smallerthan for sulphur, with only a 13% decline between 1988 and 2008,despiteemissionreductionsof50%forNOx and18%forNH3 overthesameperiod.Thenon-linearitiesbetweenemissionsanddepositionof nitrogen compounds are poorly understood, but are thought tobe related to changes in chemical processing of NOx in the atmo-sphere (RoTAP, 2012). Temporal patterns in the bulk depositionof nitrate and ammonium across the UK uplands include periodsof both significant increases and decreases over the monitoringperiod,withfewsitesshowingsignificant,monotonictrends(Curtisand Simpson, this volume).

Responses in the chemistry and ecology of upland waters tothe decreases in acid deposition are described in a number ofthe accompanying papers in this special issue and summarisedin Battarbee et al. (this volume-a). Indicators of chemical recov-ery include major reductions in non-marine sulphate, reductionsin labile aluminium and increases in ANC,while changes in nitratearemore heterogeneousanddissolved organic carbon(DOC)levelsare increasing (Fig. 3: Monteith et al., this volume). There has beenconsiderable debate about the reasons for increasing DOC, but thedecline in acid deposition is known to be a major factor (Monteithet al., 2007; Evans et al., 2012).

As a result of the chemical improvement, aquatic plant andani-mal communities are now beginning to recover. Benthic diatompopulations show statistically significant trends away from acidtolerant taxa (Flower et al., 2010). Newaquatic plant species haveappeared in seven of the lake sites and five of the stream sites inthe Acid Waters Monitoring Network. However, species compo-sition at ten sites has not changed significantly during the 20 yearmonitoringperiod, including somesiteswithsignificant improvingtrends in surface water chemistry (Shilland and Monteith, 2010).Benthic invertebrate changes atmostsites remain fairlymodestbutcommunity composition has changed significantly at about half ofthe sites in the Network (Murphy et al., this volume). For somesites the invertebrate community has yet to respond to chemicalimprovement. Salmonid fish populations are also showing some

signs of improvement,with labile aluminiumbeing the singlebestpredictor of the presence of fry (Malcolm et al., this volume).Dynamic modelling of the water chemistry of sites in the Acid

Waters Monitoring Network has provided valuable insights intothe baseline or reference chemical conditions at these sites as wellas future predictions of timescales of recovery (Helliwell et al.,this volume-a, this volume-b). The degree of recovery in ANC ateach AWMN site is shown relative to the MAGIC-modelled refer-enceANC value (using1860 as thepre-industrial referenceyear) inFig. 4. In order to control for effects of short-term variation, threeyearmeansareused forthe initial (generally 19891991) andmostrecent (20082010) three years of monitoring (cf. Kernan et al.,2010). While all sites show an increase in ANC towards the ref-erence value over the period of monitoring, recent average ANC

values atmost sites still fall short of the referencevalue (Fig. 4).

8/13/2019 1-s2.0-S1470160X1300383X-main (1)

5/19





Fig. 3. AnnualmeanDOCconcentrationsin nine AWMN lakes(19892011). Reductionsin concentrationssince2006 largely reflect a return fromparticularly elevated levelsin 2006 following a drought. The longer term trend remains significantly upward.

Diatomdata strongly support these results. Comparison of lakediatom species composition in recent sediment trap samples withpre-acidification diatom assemblages from sediment cores showsthat while there has been some recovery in acid-sensitive diatomspecies (as noted above from epilithon samples), current assem-blages are still very different from the diatom assemblages thatoccurred in theacidifiedlakesprior to thefirst signsof acidificationin the 19th century (Battarbee et al., this volume-b).

The patterns of chemical and biological recovery observed inuplandwatersof theUK areconsistent with broader scale patternsaroundEurope andNorthAmerica reported for long termmonitor-ing sites within the ICP Waters Programme (Stoddard et al., 1999;Skjelkvle et al., 2005; Skjelkvle and DeWit, 2011). The evidence

for onlymodest recovery fromacidification in the UK is consistentwith national scale modelling of critical loads for five north-westEuropean countries that shows critical load exceedances will stillbe widespread following the full implementation of the Gothen-burg Protocol by 2020, includingrecent revisions to theProtocol in2012 (Fig. 5, Curtis et al., in press-a).

Hence it is clear thatwhile some improvement inwater qualityand, to a lesser degree, biology has occurred in response to largereductions in acid deposition over the last 20 years, upland lakesand streams in the most impacted and acid-sensitive regions ofthe UK and other northern European countries will remain acidi-fied into the foreseeable future. It is likely that over the long term(decades) only a global switch away from fossil fuels as a major

Fig. 4. Initial (c. 19891991,measured), recent (20082010,measured) and reference(1860, MAGICmodelled)values of acid neutralising capacityequiv.L1 insites oftheUK Acid WatersMonitoring Network with continuous recent data (after Kernan et al., 2010). Note that initial data do not reflect the peak of acidification in the UKwhichoccurred some years prior to the onset onmonitoring. See Battarbee et al. (this volume-a) for further details of AWMN sites. Sites are ordered by decreasing referenceANC

values.

8/13/2019 1-s2.0-S1470160X1300383X-main (1)

6/19

8/13/2019 1-s2.0-S1470160X1300383X-main (1)

7/19





Fig. 6. (a) Small Water in the English Lake District, with thedrinkingwater supply reservoirHawesWater in the background (Photo: C.J. Curtis). (b) Sedimentcore analysisshows a strongdecline in thestableisotope15N at SmallWater in recent decades suggestive of anthropogenic N inputs (data from G.L. Simpson; Curtis andSimpson, 2011).

of N deposition and climate on planktonic and benthic diatoms orchrysophytes found that both drivers appeared to contribute tochanges in the sediment record, butwere inconclusive in attribut-ing their relative importance (Curtis et al., 2009; Pla et al., 2009;Simpson andAnderson, 2009).

For example, the upland tarn Small Water in the English LakeDistrict is not acid-sensitive but nitrate concentrations have beenelevated substantially by N deposition (Fig. 6, Curtis and Simpson,

2011). Analysis of sediment cores from Small Water showed astrongdeclineinthestableisotope 15NsuggestiveofanthropogenicN inputs (see above), while the winter phytoplankton was dom-inated by the mesotrophic diatom Asterionella formosa which isunusual foran oligotrophicupland lake (Curtis andSimpson,2011).Bioassays of phytoplankton production at this site indicated co-limitation by N and P throughout the growing season (Curtis andSimpson, 2011).

Excess N availability may compromise the integrity of uplandwater ecosystems in several ways. First, shifts in the relativeabundance of planktonic diatom species towards those favou-ring mesotrophic waters have been widely recorded in alpinelakes, with some evidence from UK upland waters (see aboveFig. 6). Similarly, Battarbee et al. (this volume-b) describe unex-

pected changes in the (largely benthic) diatom flora of someAWMN lakes recovering from acidification that maybe indicativeof nutrient enrichment by enhanced nitrate inputs. Nutrientaddition experiments have also demonstrated responses in thecomposition of epilithic periphyton in addition to shifts in pro-ductivity (Maberly et al., 2002). There are suggestions of a shiftfrom N-fixing epilithic cyanobacteria in low N deposition areasof Sweden to non N-fixing species in high deposition areas,although the role of organic carbon and light climate cannotbe ruled out (Liess et al., 2009). Some studies in upland lakes,for example the AWMN site Loch Coire Fionnaraich, have foundevidence of increased relative abundance of small chrysophytesin P rather than N limited lakes where they have a competitiveadvantage in P uptake (Pla et al., 2009). It was further speculated

that this phenomenon could be linked to the increases in small

planktonic cladocera observed by Kattel et al. (2006) at Loch CoireFionnaraich.

Effects of P limitation (induced by excess N availability) onzooplankton andmacroinvertebrate grazers have been suggested.Phosphorus-limited algae are poor quality food for zooplankton,and their development may therefore induce effects through-out aquatic food webs (Elser et al., 2009b). Nitrogen:phosphorusrecycling ratios in macroinvertebrates were much greater in high,

relative to low, N deposition lakes of Sweden (Liess et al., 2009).Ultimately, increasedN loadingmayreducethebiodiversity of lakephytoplankton because producer diversity tends to be low whenone nutrient is present in excess relative to others (Elser et al.,2009b).

Evidence for ecological changes in remote lakes has been usedto support the application of nutrient N critical loads, primarily inNorthAmerica(e.g.Baronetal.,2011;Sarosetal.,2011;Nanusetal.,2012). These studies suggest very low critical loads for nutrient Nto alpine lakes in the US, ranging from 1.4 to 6.0kgN ha1 year1.Changes in diatom communities may already have occurred inresponse toN inputs in lakeswheremeasured late summernitrateconcentrations are

8/13/2019 1-s2.0-S1470160X1300383X-main (1)

8/19





concentrations and CO2availability due to acidification or liming,reduced light availability to isoetids dueto smothering byfilamen-tous green algae, or enhanced phytoplankton production and/orincreases in floating leaved plants due to eutrophication.

Empirical nutrient N critical loads for softwater macrophyteshave been suggested to be in the range of 310kgNha1 year1

(Bobbink and Hettelingh, 2011) for EUNIS Category C1.1 (perma-nent oligotrophic) and C1.4 (dystrophic) lakes. The lower end oftherange is recommended foralpine lakes (which arenotdefined)while the upper end is recommended for Atlantic softwater lakes.As with empirical critical loads based on phytoplankton changein alpine lakes, the nutrient N critical loads for oligotrophic lakes(based on softwater macrophytes) are currently exceeded acrossmost of theUK(Curtis and Simpson, 2011).

Overall, there is a growing body of evidence that N depositiontoUKuplandwaters is likely tohave induced orenhancedP limita-tion of primary production in both the plankton and the epilithon.Effectsmay include changes in algal assemblages, rates of nutrientcycling and aquatic food webs. It is also likely that pressures onsoftwater macrophyte communities may have caused changes insome sites but the level of threat to isoetid communities is largelyunknown.

2.1.3. Trace metals and POPs

Fossil fuel combustion leads to the emission and long rangetransport not only of acid gasesbut also of toxic metals such asHgandPb, andpersistent organic pollutants (POPs). Toxic tracemetalsand POPs are also produced by a range of other industrial pro-cesses. Such toxic air pollutants have contaminateduplandwatersthroughout most of the world, including remote Arctic (high lati-tude) and alpine (high altitude) regions where, despite often greatdistances from sources, volatile contaminants such as POPs mayaccumulatethrough condensation dueto lowtemperatures.Heavymetals and POPs have been found at significant concentrations inSvalbardinthehighArctic(Rognerudet al., 2002), Tibet (Yangetal.,2010) and high mountain lakes around Europe, especially at theAWMN site Lochnagar in Scotland (Yang et al., 2002a,b;Vives et al.,

2004a; Rosseland et al., 2007), which is the most highly studiedmountain lake in the UK for trace metals and POPs. The few otherstudies conducted in the UKhave also shown evidence of contam-ination in remote lakes (Rose et al., 2012) reflecting the intenseindustrialisation of the UK over the last 200 years, but the spa-tial extent of the relatively high levels of contamination as seen atLochnagar remains largely unknown.

A wide range of trace metals have been identified at elevatedconcentrations in remote lakes. For example, studies of fish gillsand kidneys fromnine remote European lakes found high gill con-centrations of Al, Pb andCd and thehighest kidney concentrationsof Pb and Cd at Lochnagar (Rosseland et al., 2007). Mercury issubject to both bioaccumulation and biomagnification, and fishmuscleconcentrationshavebeenfoundtoexceedtheWorldHealth

Organisation (WHO) dietary recommendation of 0.3gg

1 in pis-civorous Arctic charr in the high Arctic site Arresjen, Svalbard(Rognerud et al., 2002), despite low sediment and water concen-trations, which is indicative of biomagnification. Lower muscleconcentrations of Hg (still approaching WHOlimits for older fish)were found inbrown trout at Lochnagar, butthe sampled fishwerenon-piscivorous (as shown by stomach content analysis) and theHg accumulation at the sitewas the highest of all non-piscivorousfish studied (Rosselandet al., 2007).

In thesamestudies ofEuropeanmountain lakes,a rangeof POPswere also analysed, including polycyclic aromatic hydrocarbons(PAHs), polybrominated diphenyl ethers (PBDEs) and organochlo-rines such as polychlorinated biphenyls (PCBs). Many of thesecompounds are endocrine disruptors. All lakes sampled revealed

concentrations of PCBs and DDTs in fish liver and muscle which

Fig. 7. Percentage reductionin decadal full basin inventoryof Pb andHg foruplandlakes (redrawn from Rose et al. , 2012). Note that negative values for eroded sitesindicate an increase in inventories.

were comparable with fish from low altitude systems (Rosselandet al., 2007). Similarly, an earlier study by Rose et al. (2001) foundlevelsof thepesticidetoxaphene in sediment from Lochnagartobehigher than at any other site receiving only atmospheric sourcesand comparable to levels found in the U.S. Great Lakes where thereare important riverine inputs. The highest levels of PBDEs in theEuropeanmountain lakes studywere found infishmuscle andliverfrom Lochnagar, but concentrations in all thehigh mountain lakeswere relatively lowcomparedwithfish from lower altitudes(Viveset al., 2004b).

The burden of trace metals and POPs in the above studies offish from remote lakes led to visible toxic effects in histologicalexaminations of liver tissues, including the trout from Lochnagar(Rosselandet al., 2007). Hence there is clear evidence, albeit from asmallnumberof studysites, forcontaminationofremotelakeswithtracemetalsandPOPs atconcentrations ofecological relevanceandpotentially of concern for human consumption of fish caught inthese remote lakes.

Many of the most toxic pollutants emitted by human activityhave been banned or restricted in recent decades with resultantlarge reductions in emissions, but some are still in current useand there remains a legacy of trace metals and POPs in soils andsediments including in upland areas (e.g. Rose et al., 2001, 2012;Yang et al., 2002a,b; Grimalt et al., 2004a,b). A future threat to

upland waters is therefore posed from these legacy pollutantsif environmental change leads to their remobilisation. In particu-lar, trace metals and POPs immobilised in upland catchment soilsmaybe remobilised by soil erosion, induced by land-use or climatechange impacts (Rose et al., 2004, 2012). Theremayalso bepoorlyunderstood changes to the toxicity of some PAHs under chang-ing ultraviolet exposure regimes due to increasing DOC trends(Schindler, 2009).

Roseetal.(2012)argue thatmanyofthe changesassociatedwithpredicted climate change, such as increased winter rainfall, pro-longed droughts andmore high-intensity rainfall events are likelyto lead to increasedsoil erosion andtransport of terrestrially boundpollutants intosurfacewaters. Fig. 7 compares thereduction in fullbasin inventories for lake sediment Pb and Hg in catchments with

thin soils, erodedsoils andnon-eroded soils. Catchments with thin

8/13/2019 1-s2.0-S1470160X1300383X-main (1)

9/19

8/13/2019 1-s2.0-S1470160X1300383X-main (1)

10/19





Fig.8. Severegully erosion onBleaklow, theheadwatersof theRiver Etherow inthePeak District of north-central England. In places the depth of eroded peat is morethan 2m down to bedrock.

Photos: C.J. Curtis.

increaseddegradationover recentdecadeshasresulted in less than50% remaining in a favourable condition (Bain et al., 2011). Causesof damagehave included land drainage, acid deposition, overgraz-ing and poorly managed burning. Physical damage to peatlandshas been proposed as a cause of higher concentrations of DOC insurface waters (Bain et al., 2011) although to date it has been dif-ficult to distinguish between effects of catchment-specific factorsandtheregional effectof decliningacid deposition in these heavilymanipulated systems (see Section 3.1).

There are moves towards large scale changes to land-use andhabitat management for restoring peatlands which could includeblocking drainage ditches, changing grazing regimes andmodify-ing burning practices (Reed et al., 2009; Bain et al., 2011). In someareas direct interventions are being implemented, including gully

blocking, removal of woodlands and revegetation of eroded peatsurfaces.While no longer widely practised on peatland, the subsidised

cutting of drainage ditches in the late 20th century resulted inmajorchangestocatchmenthydrologyinlargepartsoftheuplands,affecting local water tables, flood hydrographs, drying out anderosion of areas downslope of ditches (Ramchunder et al., 2009).Associatedwaterquality impacts include increasedsuspendedsed-iment loads (Holden et al., 2007a) and DOC fluxes (Mitchell, 1990;Wallage et al., 2006). More recently, measures have been taken toreverse thedamage andrestore peatlandsthrough blockinggullies,grips and drains. Drain-blocking in peatlands has been found insome studies to reducesuspendedsediment (Holdenet al., 2007b),DOC and colour in surface waters (Armstrong et al., 2010; Wilson

et al., 2011), while others found little impact on stream DOC (e.g.

Worrall et al., 2007). Effects of drain blocking on aquatic biotaare poorly understood (Ramchunder et al., 2009). A more recentissue has been the practice of some new wind farm constructionprojects to clear fell forest and use the mulch to block moorlanddrains, which maypose a threat to surface water quality in termsof nutrients andacidity (I.Malcolm, pers. comm.).

Rotational burning of upland sites is a long established practiceforcreatinga mosaicof differentagedvegetation standsto improvehabitats for grouse as well as sheep and deer (Ramchunder et al.,2009). Inthe21stcenturytherehasbeenanincreaseintheextentofnew burns, reflecting in part new incentives such as the MoorlandRegenerationProgrammeof2001 (Ramchunder et al., 2009;Yallopet al., 2010). Some studies indicate that the extent of new moor-land burn (

8/13/2019 1-s2.0-S1470160X1300383X-main (1)

11/19





Table 1

Regional range of projected changes from long term mean baseline (19611990) in UK air temperature and precipitation under medium emissions scenarios for the 2080s(UKCP09) Source: http://ukclimateprojections.defra.gov.uk/21730. Accessed 25th October 2013.

Probability Meanwinter t emperature ( C) Mean summer temperature (C) Meanwinter p recipitation ( %) Mean summerp recipitation ( %)

10% 0.91.6 1.52.1 +1 to +6 56to 2950% 2.23.0 3.03.9 +11 to +23 28to 1290% 3.64.7 4.96.5 +24 to +54 0 to +8

Sheep are able to crop vegetation at close to ground level,whereas cattle remove it less selectively using a ripping actionthat often leaves a longer, more heterogeneous sward (Armstrong,1998) and any management changes resulting in a change in thedominantgrazinganimal, or especially in animal density,mayhaveconsequences for soil erosion (Albon et al., 2007). The effects of aswitch from sheep to cattle grazing on upland water quality havebeentestedexperimentallyforanumberofyearsattheLochLaidongrazing experiment (Shilland et al., 2011). While this has resultedin substantial shifts in thecomposition of terrestrial vegetation, nosignificant differences in water quality have been detected overthe first two decades of the experiment between cattle-grazedand control catchments (Shilland et al., 2011). Changes in diatomassemblages during the course of this experiment were similar at

both cattle-grazed and control catchments, both indicating shiftstowards less acid tolerant taxa suggesting that thereduction in aciddeposition that has taken place hasbeen themost important influ-ence onwaterquality. This findingsupports theconclusions from anumber of palaeoecologicalstudies (e.g.Battarbeeet al., 1985) thatdemonstratedevidence for catchmentvegetationchange related toland-use change but no evidence for impacts onwater quality.

2.3. Climate change impacts on upland waters

Climate change affects thehydrological, physical, chemical andbiological characteristics of all freshwaters and is thereby a keyinfluence on biogeochemical cycling (nutrients, major ions, DOC,organic pollutants, metals), food webs andbiodiversity (Curtisand

Simpson, 2007). While pressures on upland ecosystems from sul-phur and N deposition are declining, climate change is expectedto have an increasing impact on upland biogeochemical cycling,the behaviour of pollutants stored in catchment soils, land useand land management, and hence water quality over the comingdecades. These interactions are illustrated conceptually in Fig. 2.The potential impacts of climate change on upland waters havebeen investigated in a number of international research projectsincluding EMERGE (Catalan et al., 2009; Kernan et al., 2009) andEuro-limpacs (e.g.Whiteheadet al., 2009). Predictions of thedirecteffects of climate change on aquatic ecosystems are very complex,andoncecombinedwithotherhuman impactsposeanevengreaterchallenge.

The latestUKclimate projections,UKCP09(Murphy et al.2009),

provide ranges of projected winter temperature increases by the2080s for theUK,which are summarised for regionally aggregateddata in Table 1. The percentiles refer to the probability of differ-ent increases: the50th percentile is thecentral estimateof change,while the 10thpercentile is very likely tobeexceededand the 90thpercentile is very likely not to be exceeded (Murphy et al., 2009).It should be noted that UK climate has been relatively stable overthe period the AWMNhas been operating to date, and while datafrom the network are already being used to examine the effects ofinterannual climate variability on biogeochemistry and biota, fur-ther years ofmonitoring will be necessarybeforeeffects of climatechange can begin to be assessed. Furthermore, there are very fewlong-term data series for water temperature in the UK.

UKCP09 projects UK mean winter air temperature increases in

all parts of the UK with 50th percentile projections of 2.23.0

C

increases and a 1090th percentile range of 0.94.7

C with themedium emissions scenario. The lowest regional increases areprojected for northern and eastern Scotland and the highestfor southern and eastern England. For mean summer air tem-peratures the 50th percentile projections are increases in therange 3.03.9C (1090th percentile 1.56.5 C) with the smallestincreases in northern Scotland and the greatest in southern andeastern England.Hence thegreatest air temperature increasesmaybeinthelowlandeasternandsouthernregionsoftheUKwhilemostMMHcatchments are in regions with lower projected increases.

Winter precipitation may increase in all regions, with a 50thpercentilerange of 1123%increaseanda1090thpercentilerangeof 154%increases. Thesmallest increasesareprojected foreasternScotland and Northern Ireland while the largest are projected for

south-westEngland.Overall,decreasesinsummerprecipitationaremore likely, with a 50th percentile range of 1228% reductions.The 10th percentile of model runs project decreases in the range2956% while the 90th percentile of runs project small increasesin the range 08%. The largest changes are modelled for southernEngland and the smallest for northern Scotland.

Here we consider two aspects of these climate change impactson upland waters; direct changes in terms of temperature andhydrology, and biogeochemical cycles in upland catchments,specifically the carbon cycle.

2.3.1. Direct impacts of climate change in the uplands: hydrology,

temperature and aquatic ecology

Extensiveanalysis of weather station data over several decades

indicates that upland areas are experiencing greater climaticchange than lowland areas in the UK (Burt and Holden, 2010;Holden and Rose, 2011), with stronger increases in winter mini-mum temperatures and increasingwinter precipitation.

Increasing temperature may impact upland lake ecosystems inseveral ways, through decreases inwinter icecover, changes to thethermal structureof lakesandtheirmixing behaviour, andphysicalchanges brought about by temperature increase, which can havesignificant impacts on surface water chemistry and biology. Theremay also be a feedback in terms of increasing DOC concentrations(see Section 3.1) which could affect absorbance of solar radiationand hence lead to earlier and more frequent onset of ice melting.

Mountain lakes are very sensitive indicators of environmentalchange and have responded to climate change both chemically,

through pH changes in response toweathering inputs from catch-ments (Psenner and Schmidt, 1992; Sommaruga-Wgrath et al.,1997; Koinig et al., 1998; Mast et al., 2011) and biologically, asindicated by palaeolimnological records of changes in lake phyto-plankton (e.g.Battarbeeet al., 2002;Catalan et al., 2002;Smolet al.,2005; Curtis et al., 2009). Similar changeshave also been recordedinArcticlakes(e.g.Prowseet al., 2006;Hobbsetal., 2010;Holmgrenet al., 2010).

The most alpine site in the UK AWMN, Lochnagar, lies withina maritime altitude zone deemed to be particularly sensitive toclimatic influences on ice cover duration (Thompson et al., 2007;Kettle and Thompson, 2007). While currently freezing for about3 months of the year, it was likely to have experienced around 6weeks extra ice cover during the Little Ice Age, while just modest

warming from the present could mean very limited opportunities
http://localhost/var/www/apps/conversion/tmp/scratch_7/dx.doi.org/10.1016/j.ecolind.2013.10.012http://ukclimateprojections.defra.gov.uk/21730http://ukclimateprojections.defra.gov.uk/21730http://localhost/var/www/apps/conversion/tmp/scratch_7/dx.doi.org/10.1016/j.ecolind.2013.10.012

8/13/2019 1-s2.0-S1470160X1300383X-main (1)

12/19





Fig. 9. Modelled effects of climate warming on lake ice-coverduration in Scotland.

Source: Kettle and Thompson (2007); with kind permission from Springer ScienceandBusiness Media.

for ice formation in future (Fig. 9, Thompson et al., 2007). Previ-ousclimate changeprojections (UKCIP02)suggestedan increase insummermaximumtemperaturesby24C by2080whilemonthlyminimainwinterwouldrarelybebelow0C(KettleandThompson,2007). Even lowemission scenariospredicteddrastic reductions insnowfall at Lochnagar due to temperature increases, although theoverall change in precipitation (as rainfall) is unknown. The grow-ingseason forprimaryproducerswithinthe lochwasalsopredictedto increase,perhapsbyasmuch as50%under high emissionscenar-

ios (Kettle and Thompson, 2007). More recent climate projectionsunder UKCP09 suggest even greater warming than UKCIP02 forbothwinter and summer (Murphy et al., 2009).

The UK uplands experience high windspeeds and upland lakestend thereby to be well mixed. Temporary stratification can occurin summer during calm periods (e.g. Curtis and Simpson, 2011;Shilland et al., 2012), so most lakes can be classified as beingpolymictic. Only one site in the AWMN, Loch Chon, stratifies con-sistently each year (Shilland et al., 2012). The extent to which themixingcharacteristics ofUK lakeswill changewilldependonfuturechanges in summerair temperature andonwind strength, and thealtitude and exposure of individual sites. Overall wemight expecta shift tomore stablesystemswith stratification takingplacemorefrequently and for longer periods of time as summer wind speeds

decline slightlywhile temperatures increase at sites likeLochnagar(Kettle and Thompson, 2007).In running waters there are concerns for the future for cold

stenothermic taxa, most notably salmonid fish such as salmon(Salmo salar) and brown trout (Salmo trutta), in terms of poten-tial changes in geographic range in response to changing watertemperatures andflows(e.g.Jonssen and Jonssen, 2009;Elliott andElliott, 2010; Moore et al., 2012). Climate warming resulted in anupward shift of thermal habitat for brown trout in Swiss alpinerivers and streams, indicating a loss of habitat due to physicalbarriers to upstreammigration and linked to an increase in tem-perature dependent kidney disease (Hari et al., 2006). Predictionsof lake and stream warming in northern North America indicatetemperatureswill increase beyond lethal levels for cold stenother-

mic fish taxa like lake trout andwhitefish (Schindler, 2009; Cunjak

et al., 2013). In Scotland, Hrachowitz et al. (2010) demonstratedthat small upland streams with no forest cover weremost vulner-able to the adverse effects of projected climate change on watertemperature regime suitability for salmonids. Clews et al. (2010)foundsignificant relationships between juvenile salmonid popula-tion densities and antecedent summer climate in the River Wyein Wales over 20 years, with hot, dry summers associated withreductions in salmondensity.

Aquatic macroinvertebrates are also sensitive to the effects ofclimate change. Local extinctions of sensitive macroinvertebratespecies under 3 C warming at Llyn Brianne in Wales were pre-dictedbyDuranceandOrmerod(2007),whiletemperaturewasalsofound to be themost importantdeterminant ofmacroinvertebrateassemblages in northern streams acrossa gradient fromDenmarkto Greenland and northern Sweden under the North Watch study(Friburg et al., 2013). A key unknown is the adaptive capacity ofthe impactedmacroinvertebrate species, butwith increasing tem-peratures, metabolic rates will increase with likely impacts on thefunctioning of whole ecosystems (Tetzlaff et al., 2013a).

As well as changes in water temperature, changes in hydro-logical regimes are also anticipated under climate change, withpoorlyunderstood impacts on thestructureandfunctionofuplandwater ecosystems. Hydrological modelling at one of the AWMNsites, Dargall Lane, indicates that most UKCP09 projections wouldresult in major changes in flow seasonality with increased win-ter discharges and reduced summer flows (Thompson, 2012).Decreases in snowfall have been projected for Scottish uplandstreams including the AWMN site Allt aMharchaidh, with asso-ciated reductions in snowmelt inputs but increased winter flows(Capell et al., 2013). Such hydrological changes are likely toresult in important hydrochemical and biological changes inaquatic ecosystems; indeed, Tetzlaff et al. (2013b) describenorthern rivers studied under the North Watch project, whichincluded the AWMN sites Allt aMharcaidh and Allt na Coirenan Con, as being on the cusp of major changes in ecohydrol-ogy.

In the North Watch study, temperature seems to be a more

important determinant of change for macroinvertebrate commu-nities (Friburg et al., 2013) while geographic range suitability forsalmonmay be impacted by changes in both hydrology and tem-perature (Tetzlaff et al., 2005;Clewset al., 2010;Mooreet al., 2012;Cunjak et al., 2013). Lower flows and higher temperatures duringsummer, as predicted forupland catchments, are likely to increasestresson salmonids,while largefloodsmaywashoutsalmonideggsdepositedin spawninggravels (Clews etal.,2010;Thompson,2012;Cunjak et al., 2013).

Finally, catchment vegetation is also expected to respond tochanges in climate, with resultant influences onwater quality andecohydrologywhich are poorly understood (Tetzlaff et al., 2013b).The composition and distribution of catchment vegetation maybe affected by changing temperature and precipitation regimes,

withphysiologicalresponsesthatmayinclude,forexample,alteredrooting patterns and phenology (Tetzlaff et al., 2013b). There arelikely to be complex feedbacks between catchment hydrology andvegetation change (for example through stomatal control ofwaterlosses and changed evapotranspiration) which are difficult to pre-dict, as well as impacts on biogeochemical cycling of carbon andN.

2.3.2. Climate change impacts on catchment biogeochemical

cycles

The macronutrient cycles of upland catchments are intimatelylinked to both climate and vegetation type. Since vegetation coverand typewill also respond directly to changes in climate there arepotential feedbacks in terms of nutrient cycles, all of which will

impact on upland aquatic ecosystems.

8/13/2019 1-s2.0-S1470160X1300383X-main (1)

13/19





Considerable scientific attention is currently being devoted tothe quantification of the major carbon fluxes in peatlands (manyof which are located in the uplands in the UK) to determine theirfuture role as potential sinks or sources of carbon under climatechange. The key question is whether peatlands will continue tosequester carbon and build peat (a net sink) or will degrade andrelease old carbon (net source).

Climate change impacts on upland peatlands may depend onwhether they are intact and healthy or degraded, since damagedpeatlandsare considered tobe less resilientto climate change(Bainetal.,2011). Climatechangemay leadtoincreasederosion,floodingand risk ofwildfires (Bain et al., 2011)which could increase carbonloss both to the atmosphere and to surface waters. Increased peaterosion, transport of particulate organic carbon and sedimenta-tion impacts on downstream fisheries, e.g. for salmon, could result(House et al., 2010; Bain et al., 2011).

Carbon losses from peatlands may be increased both directlyby the influence of higher temperatureson organic matter decom-position, and by droughts, whereby increased oxygen penetrationresults in greater phenol oxidase activity and reduced concen-trations of the phenolic compounds which retard decomposition(FennerandFreeman, 2011). Davidson and Janssens (2006)arguedthat thedecomposition rate ofmore refractory organicmattersuchas peat is particularly sensitive to increasing temperature as aresult of the relatively high activation energies of such substrates,a hypothesis supported by incubation experiments ofCraine et al.(2010) on soils acrossNorth America.

A modelling study of c. 1000 Norwegian headwater lakes sug-gested that increased terrestrial vegetation cover under climatechange could lead to major increases in organic carbon in lakes(Larsen et al., 2011). Such an effect could be additive to the de-acidificationmediatedincreasesinDOCdescribedabove (e.g.Evanset al., 2012).

Analysis of bioclimatic envelope (niche) models suggests thatmore than half of UK peatlands will experience change by 2050(Clark et al., 2010) while over the longer term, the area of activelygrowing blanket bogs could shrink by 84% from the present day

(Gallego-Sala et al., 2010). However, otherdynamic process-basedmodel applications gave mixed results with respect to the netstorage or release of carbon, showing the critical importance ofmonitoring the response of peatlands to climate change for reduc-inguncertainties inmodelling and futurepredictions (House et al.,2010).

The carbon cycle is not the only aspect of upland catchmentbiogeochemistry to be impacted by climate change. Whiteheadet al. (2009) reviewed climate variables that can affect the acidi-fication process, including higher temperatures, summer drought,wetter winters, reduced snow cover and changes in hydrologicalpathways. Climate change is likely to lead to changes in the inci-dence of storm events that are known to deposit large sea-saltloads onto upland catchments, leading to temporarydisplacement

of acid cations stored in catchment soils and episodic acidification(Skjelkvle et al., 2007; Evans et al., 2008b; Monteith et al., thisvolume). Drought and rewetting cycles can be important mecha-nisms for the oxidation of reduced sulphur compounds in organicsoilsandreleaseofacidity (Clarketal., 2005,2009; Schindler, 2009)and in uplandwaters in parts of the USA, climate induced moder-ation of sulphate release from catchments soils has overtaken Sdeposition as themain determinant of streamwater sulphate con-centrations (Mitchell and Likens, 2011). Such processes are likelyto be important in areas of the UK uplands such as the North YorkMoors where slow release of stored S occurs (Evans et al., thisvolume).

Climatic controls on the leaching of nitrate from upland soilshave been demonstrated through synchronous patterns in stan-

dardised nitrate concentrations and the North Atlantic Oscillation

(NAO) Index (Monteith et al., 2000, this volume). This observa-tion is consistent with independent evidence from stable isotopesthat a major proportion of the nitrate in upland waters hasbeen produced microbially (Curtis et al., 2012), since climateis an important determinant of mineralisation and nitrificationprocesses.

Henceit isclear that climateinduced changes tobiogeochemicalcycles in the uplands will influence surface water fluxes of carbon(both particulate anddissolved), N andpotentially sulphur aswell.The water quality implications of all these changes have been dis-cussed previously in the context of land-use change and recoveryfromacidification above.

3. Synthesis and recommendations

The diverse drivers of change affecting upland water ecosys-tems result in complex, interacting responses in terms of uplandcatchmentbiogeochemistry,aquatic chemistryandaquaticecology(see Fig. 2) and future directions of change are therefore difficultto predict. Outcomes will vary from site to site depending on thecontinuing,althoughdeclining,influenceofaciddeposition andtheemergence of newpressures, especially from climate change.

3.1. The legacy of acid deposition

Although acid deposition has declined substantially its legacywill continue to influence surface waters in the UK for manydecades to come, as soils slowly release S and N compounds andtoxic tracemetalsthrough leachinganderosionandbeginto regainbuffering capacity through the process of geological weathering.While partial chemical recovery of acidified watersheds is alreadyclear, pH and labile aluminium concentrations in more impactedwaters areunlikely to reach conditions required tomeet good eco-logical status for several decades. Where soils become saturatedbyN as predicted by FAB modelling, recovery may be even furtherdelayed.

Chemical and biological recovery processes will both be influ-enced by changes in temperature andhydrology brought about byclimate change and changes in uplandmanagement. Any increasein winter precipitation, or increase in sea-salt deposition broughtabout by increased storminess, is likely to depress pH and raiselabile aluminium concentrations in recovering waters by favou-ring more lateral flow paths and stimulating displacement ofacid cations from soils by marine base cations. Meanwhile, anychanges in the extent and species composition of upland forestswill continue to influence atmospheric acid interception rates,while changes in terrestrial N demand brought about by fellingand replanting will result in pulsed releases of acidifying nitrate.Of continuing concern is the extent to which persistence of occa-sional highly acid episodes brought about by these interactions

may serve as a ceiling for biological recovery (Lepori et al., 2003;Kowaliketal.,2007).DatafromAWMNstreamsitessuggestthattheseverityof acidepisodesis decliningat leastas rapidlyasmeanacid-ity (Monteith et al., this volume), but themagnitude of occasionalspikes in labile aluminium and hydrogen ion concentrations maystill act as a barrier for the return of some acid-sensitive species.

A further consequence of acid deposition on uplandwaters hasbeen changes in DOC concentration, although the interpretationof DOC trends has not been without controversy. Data from theUK Acid Waters Monitoring Network provided some of the firstevidence for recent large regional-scale increases in uplandwaterDOC fluxes (Freeman et al., 2001; Evans et al., 2006). Subsequentdebate over drivers of increasing DOC, whether reductions inacid deposition, climate change (see Section 2.3.2), N deposition

or land-use change (see Section 2.2.2), has prompted various

8/13/2019 1-s2.0-S1470160X1300383X-main (1)

14/19





national-scale investigations (e.g. Evans et al., 2006; DeWit et al.,2007; Erlandsson et al., 2010; SanClements et al., 2012), a largeinternational study of DOC trends involving sites across the ICPWaters network (Monteith et al., 2007) and field experiments(Evans et al., 2012). Together these provide highly persuasiveevidence that soil acidity, and hence acid deposition, exerts amajor temporal control on the mobility of DOC and that reducedacid deposition is theprimary driverof increasing DOCtrendsseenin streams and lakes draining the uplands of the UK and otheracidification-impacted regions. This conclusion has importantimplications for our understanding of the reference conditions ofacidified lakes in terms of pre-industrial DOC concentrations andpH (Evans et al., 2008a; Erlandsson et al., 2011), as it suggeststhat organic acidity may have partially countered changes inmineral acidity from atmospheric pollutants. Consequently, thepre-industrial pH of surface waters in some areas may not havebeen as high, and recovery in pHmay therefore not be asmarked,asdynamicmodels such asMAGIC oncepredicted (Battarbeeet al.,2005).

The link established between rising DOC and falling acid depo-sition is highlypertinent to the interestsofwatersupplyindustriesserving most of the north and west of the UK. Rising levels of DOCare increasing water treatment costs substantially, as removal ofDOCis necessarypriorto chlorination.Deposition-driven increasesin thesolubilityof soil organicmatterare likelytomakeDOCexportmore vulnerable to the range of climatic factors and soil degrada-tion described in earlier sections. Levels of DOCreaching thewatersupply system in these regions are therefore likely to at least besustained, or even further increased, in future as soils continue torecover from acidification, and attempts to reduce fluxes throughlocal catchment management will need to take into account thisdominant regional effect.

Another factor linked to air pollution with the potential toinfluence DOC levels in upland waters is the eutrophying effectof N deposition on terrestrial ecosystems. Pregitzer et al. (2004)demonstrated that nitrate additions of 30kgNha1 year1 tohardwood forests, comparable to high N deposition regions of

the UK uplands, increased soilwater leaching fluxes of nitrate,DOC and also DON. Findlay (2005) suggested a similar mech-anism may be responsible for increased DOC fluxes into theHudson River, possibly due to changes in soil enzyme activity.Even if terrestrial primary productivity across the UK uplands iseffectively N-saturated across the region today, the legacy of his-torical N deposition is likely to be affecting nutrient cycling. Itis therefore possible that DOC concentrations and fluxes may beapproaching, or already have exceeded, pre-industrial levels as aresult of the combined effects of N-enhanced NPP and an aciddeposition-induced return towards pre-industrial organic mattersolubility.

Regardless of the driver, in upland waters where levels ofplanktonic algae and suspended sediment are often relatively low,

DOC concentration is normally the dominant determinant of thedegree of penetration of photosynthetically active radiation (PAR)through the water column. Rising DOC concentrations are there-fore reducing PAR penetration of upland lakes, where in somecases benthic primary production may be limited by light ratherthan nutrients (Karlsson et al., 2009). Dissolved organic matteralso has an important function as a sunscreen, limiting the expo-sure of biota to harmful ultraviolet radiation (Schindler, 2009).Furthermore, changes in water colour may affect lake freezingdynamics.

The ecosystem services provided by upland waters and theircatchments in termsof carbonstorage,aquatic biodiversityand theprovision of high quality drinking water are evidently affected bymultiple interacting factors. Climate change and land-use clearly

play important roles in themobilisation of carbon from catchment

soils and as acid deposition continues to decline there is likely tobe a shift in the relative importance of the key drivers of change inDOC.

3.2. Emerging pressures on upland waters

It is clear that the main pressures on uplandwaters are shiftingin relative importance and presenting new challenges to scien-tists, policymakers, landscape and environmental managers. Asatmospheric deposition of pollutants continues to decrease, cli-mate change is set to continue, with major changes for uplandwaters anticipated in temperature, seasonal flowpatterns and therole of snowfall, snowmelt and lake ice-cover. Concomitantly, landuse and management interventions are also expected to change.While forestry practices have moved away from large scale use ofconifers in acid-sensitive areas, there is nevertheless a clear inten-tion in some regions, as in Scotland, to significantly increase theplanted area and this may lead to conflicting interests with otherstakeholders in the uplands such as groups working to restoresalmonid fisheries in previously acidified areas.On theother hand,mitigation efforts to reduce the impacts of increasing temper-atures on upland stream ecosystems may include riparian treeplanting.

Even as acid deposition is decreasing across Europe and NorthAmerica, there is increasing evidence for nutrient-N inducedchanges in phytoplankton assemblagesofArcticand alpinelakes atdepositionlevelslowerthanthosecurrentlyexperiencedanywherein the UK. Likewise, changes in aquatic macrophyte communitiesin upland waters have been attributed to N deposition in variousEuropean countries, but not yet in the UK. In both cases, convinc-ing evidence for such changes in the UK uplands is still lacking. Inpart thismay be due to the availability of historical data, the mainimpacts of N deposition possibly having occurred several decadesprior to theestablishmentof robustmonitoring systems,butpalae-olimnological studies of upland lakes in high N deposition areasof the UK have yet to provide conclusive evidence of significantenrichment during this period (Curtis and Simpson, 2011). Poten-

tial changes due to changing nutrient regimes may also have beenmasked, or inhibited, by the confounding effects of acidification(Fig.1b).DepositionofbothNandS,aswellastracemetalsandtoxicorganic compounds, is however still declining and future issuesare likely to revolve around the legacy of cumulative depositionloadsovermanydecades in thecontextofa changingclimate in theuplands. The possibility that climate change could lead to increas-ingmobilisationofsuchpollutantsstoredinuplandcatchmentsoilsis a very real one.

Changes in perceivedpressuresmaysometimes result in seem-ingly conflicting benefits and disbenefits. Hence the reductionin acid deposition that is promoting the return of acid-sensitiveaquatic species would also appear to be resulting in increasedwater treatment costs due to increasing trends in DOC, and poten-

tially, depending on the fate of the fluvial carbon, a net lossof soil carbon to the atmosphere. On the other hand, there arealso potential co-benefits which policymakers should aim tomax-imise. For example, reduced exploitation and physical restorationof peatlands can improve water quality and reduce UK carbonemissions.

There are complex interactions between changes in atmo-spheric deposition, land use, climate change and catchmenthydrology with poorly understood feedbacks, as discussed inSection 3.1 above in the context of DOC trends. Neverthe-less, a wide range of studies indicate ongoing, but changing,impacts on sensitive biota in aquatic ecosystems, from phyto-plankton and aquatic macrophytes to macroinvertebrates andsalmonid fish, leading to changes in food webs and in ecosystem

functioning.

8/13/2019 1-s2.0-S1470160X1300383X-main (1)

15/19





Giventhewiderangeof often interactingandevolvingpressuresthat uplandwater ecosystems are facing, an integrated ecosystemapproachwillbenecessarytobalancetheneedsofconservationandthe wider stakeholder community. Such an integrated approach,along with the establishment of resilient ecological networks,is indeed recommended in the UK Government White Paper onthe Environment (H.M. Government, 2011). A further target in theWhite Paper is to identify where land can be managed to delivermultiple benefits, including improved water quality, flood alleviation

and biodiversity. Integrated monitoring of upland aquatic ecosys-tems is a prerequisite for understanding and implementing theseapproaches.

3.3. Shifting baselines and novel recovery endpoints

While the pressure from deposition of pollutants may bedecreasing, climate change over the foreseeable future will havea strong influence on upland freshwater ecosystems limiting theextent to which restoration targets based on the reference condi-tion concept can be attained. In the absence of such a benchmarkit is difficult to define an ecological endpoint that managerscan use for guidance. The recent Government White Paper (H.M.Government, 2011, section 2.69) states that the aim for the major-ity of UKwater bodies is to achieve good ecological status as soonas possible and for as many as possible to attain good ecologicalstatus by 2027. This target is set under a section titled Restoringnature in our rivers and water bodies. However, restoring natureis a troublesome concept.

Battarbeeet al.(thisvolume-b)andMalcolm et al. (thisvolume)raise in particular the issue of difficulties associated with defin-ing chemical andespecially biological endpoints for uplandwatersrecovering from acidification. Critical loads of acidity are set rel-ative to a target ANC value which makes no direct reference tobaselinevalues,soachievementofcriticalloads,asenvisagedunderthe UN ECE Convention on Long-Range Transboundary Air Pollu-tion, does not imply any return to baseline or reference conditions(Kernan et al., 2010). Dynamic models like MAGIC can be used

to hindcast a notional pre-industrial, unimpacted (with respect toacid deposition) chemical baseline (Helliwell et al., this volume-b)and attempts have been made to link these to reference biologi-cal assemblages using modelling approaches (e.g.Juggins, 2001).Palaeolimnological records provide a more direct (if only partial)indication of pre-acidification communities. Indications fromsedi-ment trapdataatAWMNsites suggest that trajectories of recoveryindiatomcommunities arenotnecessarilyreturningtowardsthesebaseline conditions but potentially towards entirely novel assem-blages (Battarbee et al., this volume-b). Dynamic modelling ofrecovery from acidification indicates that baseline conditions forsoil base saturation (and hence resilience to acid episodes) areunlikely to be achieved by 2100 (Helliwell et al., this volume-b).Furthermore, increasingDOCtrends dueto recovery fromacidifica-

tion inuplandsoilssuggeststhatmanyuplandwatersarereturningtowards a more highly coloured natural baseline, implying a shiftin the relative roles of benthic and planktonic primary production(Monteithetal.,thisvolume). Thereforethekeyquestion forsettingrecovery targets is whether these chemical or biological baselinesare achievable or even appropriate in the context of a changingclimate (Fig. 10).

3.4. The requirements for integrated monitoring of upland waters

Many authors have highlighted thedearthof high quality, long-termmonitoring sites which can provide the integrated chemicaland biological datasets required to investigate the ongoing andfuture changes in upland water (or indeed any other) ecosystems.

For example, Tetzlaff et al. (2005) point out that for studies linking

Fig. 10. Schematic diagram of potential recovery outcomes in the context ofdeclining acidification pressures and other dynamic drivers of change indicatinghowpalaeolimnological (dashed line) andmonitoring data (solid line) canbe com-bined to track change from pre-acidification reference conditions to the presentday. Potential future change is shown by dotted lines indicating theuncertainty ofmovingback to thereference (a) or towards a novel endpoint (c).

Modified fromBattarbee et al. (2012).

hydrologicaleventstobiological responses, in this case forjuvenilefeeding and adult spawning of salmon, high resolution flow dataare required and daily meandatamay be insufficient. Ramchunderet al. (2009) identify a major lack of research in the investigationof upland stream ecological changes in response to drainage,drain-blocking and burning. The urgent need to continue withexisting long-term monitoring programmes in the uplands, suchas the Countryside Survey, Environmental Change Network andAcid Waters Monitoring Network, has been highlighted by Dise(2009), Bain et al. (2011) and Tetzlaff et al. (2013a). Upland watersmonitoring addresses a series of very specific research needs,highlighted above, in small headwater streams and lakes whicharelargely neglectedunderthe EUWaterFrameworkDirective and

bynational regulators dueto a historical focus on point sourcepol-lution. Nevertheless, these small uplandwaterbodies fulfil a rangeof ecosystemserviceswhicharevery disproportionate to their size.

From therangeof scientificfindingsdrawnfrom theAWMNandother long-term upland monitoring studies reported in this spe-cial issue (and see Battarbee et al., this volume-a, this volume-b)there canbe littleargument that their high quality long-terminte-grated datasets have made, and continue to make, a substantialcontribution to UK environmental scientific capability. The abil-ity to compare andcontrast ecological andhydrochemical datasetsacross years and between sites delivers information and raisesnew environmental change hypotheses that can then be testedexperimentally. The accumulated data have clearly allowed theAWMN to go beyond its original tight remit to provide precise

and accurate data on upland environments more widely. Giventhe long-term time series and high quality of data provided bythe AWMN and other important datasets presented in this spe-cial issue, these networks will be well placed to contribute to ourunderstanding of emerging issues relating to the use of and dam-age to upland ecosystems. The AWMN in particular is used here asan example of the typeof integratedmonitoring that is required toaddress the research questions posed above and integrated man-agement approaches proposed in the Government White Paper(H.M. Government, 2011).

Clearly, long-term monitoring networks need to be continu-ally modified to adapt to changing circumstances. The AWMN,for example, was designed to address questions relating to thechemical and ecological impacts of acid deposition on acid-

sensitive upland aquatic environments, incorporating both lakes

8/13/2019 1-s2.0-S1470160X1300383X-main (1)

16/19
http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0155http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0155http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0155http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0155http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0155http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0155http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0155http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0155http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0155http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0155http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0155http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0155http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0155http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0155http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0155http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0155http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0155http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0155http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0155http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0155http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0150http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0145http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0145http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0145http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0145http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0145http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0145http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0145http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0145http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0145http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0145http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0145http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0145http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0145http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0145http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0145http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0145http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0145http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0145http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0140http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0140http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0140http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0140http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0140http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0140http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0140http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0140http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0140http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0140http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0140http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0135http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0130http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0125http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0120http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0115http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0110http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0110http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0110http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0110http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0110http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0110http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0110http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0110http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0110http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0110http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0110http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0110http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0110http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0110http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0110http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0105http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0100http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0095http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0095http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0095http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0095http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0095http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0095http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0095http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0095http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0095http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0095http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0095http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0095http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0095http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0095http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0095http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0090http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0090http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0090http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0090http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0090http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0090http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0090http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0090http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0090http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0090http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0090http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0090http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0090http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0090http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0090http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0090http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0090http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0090http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0090http://www.rivm.nl/ccehttp://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0080http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0080http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0080http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0080http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0080http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0080http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0080http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0080http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0080http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0080http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0080http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0075http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0070http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0060http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0060http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0060http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0060http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0060http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0060http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0060http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0060http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0060http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0060http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0060http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0060http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0045http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0045http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0045http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0045http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0045http://refhub.elsevier.com/S1470-160X(13)00383-X/sbref0045http://refhub.elsevier.com/S1470-160X(13)

1-s2.0-S1470160X1300383X-main (1)

Documents

Transcript of 1-s2.0-S1470160X1300383X-main (1)