river bed carbon and nitrogen cycling: state of play and some new directions

Science of the Total Environment 434 (2012) 143–158

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Review

River bed carbon and nitrogen cycling: State of play and some new directions

Mark Trimmer a,⁎, Jonathan Grey a, Catherine M. Heppell b, Alan G. Hildrew a, Katrina Lansdown b,Henrik Stahl c, Gabriel Yvon-Durocher a

a School of Biological and Chemical Sciences, Queen Mary University of London, E1 4NS, UKb Department of Geography, Queen Mary University of London, E1 4NS, UKc Scottish Marine Institute Oban, Argyll, PA37 1QA, UK

⁎ Corresponding author.E-mail address: [email protected] (M. Trimme

0048-9697/$ – see front matter © 2011 Published by Eldoi:10.1016/j.scitotenv.2011.10.074

a b s t r a c t
a r t i c l e i n f o
Article history:Received 15 April 2011Received in revised form 30 September 2011Accepted 21 October 2011Available online 8 June 2012

Keywords:CarbonNitrogenBiogeochemistryDenitrificationMethaneChemosynthesis

The significance of freshwaters as key players in the global budget of both carbon dioxide and methane hasrecently been highlighted. In particular, rivers clearly do not act simply as inert conduits merely piping car-bon from catchment to coast, but, on the whole, their metabolic activity transforms a considerable fraction ofthe carbon that they convey. In addition, nitrogen is cycled, sometimes in tight unison with carbon, with ap-preciable amounts being ‘denitrified’ between catchment and coast. However, shortfalls in our knowledgeabout the significance of exchange and interaction between rivers and their catchments, particularly the sig-nificance of interactions mediated through hyporheic sediments, are still apparent. From humble beginningsof quantifying the consumption of oxygen by small samples of gravel, to an integrated measurement of reachscale transformations of carbon and nitrogen, our understanding of the cycling of these two macro elementsin rivers has improved markedly in the past few decades. However, recent discoveries of novel metabolicpathways in both the nitrogen and carbon cycle across a spectrum of aquatic ecosystems, highlights theneed for new directions and a truly multidisciplinary approach to quantifying the flux of carbon and nitrogenthrough rivers.

© 2011 Published by Elsevier B.V.

Contents

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1441.1. Renewed interest in macronutrient cycling in rivers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1441.2. Scope of the review . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 144

2. Linking geomorphology, hydrology and biogeochemistry in benthic sediments and the hyporheic zone . . . . . . . . . . . . . . . . . . . 1452.1. The role of benthic sediments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1452.2. Hydrological exchange in permeable catchments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1452.3. The landscape to river continuum . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 145

3. Carbon and nitrogen metabolism in rivers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1463.1. Quantifying carbon metabolism, new directions and some functional metrics . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1463.2. Relationships between temperature and the carbon cycle . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1473.3. Pathways of N removal: the old and the new . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 148

3.3.1. Riverine and hyporheic ‘denitrification’: a brief overview . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1493.3.2. Quantifying the reduction and removal of nitrate in riverbed sediments . . . . . . . . . . . . . . . . . . . . . . . . . . 1503.3.3. Reach scales measurements of denitrification and estimates of the global rate in rivers . . . . . . . . . . . . . . . . . . . 150

4. River bed methanogenesis, methane oxidation, chemosynthesis and the third way . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1514.1. River bed methanogenesis and methane oxidation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1514.2. Methane based chemosynthesis and the third way . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 151

5. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 154References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 154

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144 M. Trimmer et al. / Science of the Total Environment 434 (2012) 143–158

1. Introduction

1.1. Renewed interest in macronutrient cycling in rivers

The cycling of carbon in freshwaters (lakes, streams and rivers)has attracted renewed interest in the last few years (Battin et al.,2008, 2009; Cole et al., 2007; Tranvik et al., 2009). Where once theglobal carbon cycle was thought to be the sole domain of the fluxand storage of carbon between the atmosphere, terrestrial and ocean-ic pools, recent revisions to the global carbon budget have highlightedthe significant contribution that freshwater ecosystems make to theglobal carbon cycle, despite their comparatively small size (Battin etal., 2008, 2009; Cole et al., 2007; Tranvik et al., 2009). It is now esti-mated that close to 3 Pg C y−1 are transported, mineralised and bur-ied in inland freshwaters, a budget which is comparable to therecognised terrestrial carbon sink for anthropogenic CO2 emissions,also of ~3 Pg C y−1 (Battin et al., 2009; Cole et al., 2007). In addition,others have drawn attention to the considerable role that inlandfreshwaters make to the global methane (CH4) budget, with emis-sions of CH4 from freshwaters being at least comparable to the terres-trial CH4 sink; yet the role of rivers in this freshwater emission of CH4

remains poorly defined (Bastviken et al., 2011; Tranvik et al., 2009).Hence our understanding of the carbon cycle on Earth will only becomplete if we include the flux of carbon through inland freshwaters(Battin et al., 2009; Cole et al., 2007).

That is not to say that the biogeochemical cycling of carbon in riv-ers and freshwater ecosystems, as a whole, has not been recognisedand studied extensively (Brunke and Gonser, 1997; Cole and Caraco,2001; Tank et al., 2010; Webster and Meyer, 1997), but more thatthe true flux of carbon through freshwaters has not been fully appre-ciated in a global context (Battin et al., 2008; Cole et al., 2007; Tranviket al., 2009). In particular, the lateral exchange of carbon between themain river channel, floodplains, riparian areas and fringing wetlandsand how this exchange modulates and transforms the transport ofcarbon from catchment to coast needs to be considered (Battin etal., 2009; Cole et al., 2007).

For example, in their revision of the ‘active pipe’ hypothesis, orig-inally put forward by Cole et al. (2007), Tranvik et al. (2009) arguethat close to 50% of the ~3 Pg C y−1 conveyed to inland waters(lakes, streams and rivers) from the land is lost as CO2, ~20% is buriedin sediments and some 30% is exported to the sea. The import ofallochthonous carbon is, therefore, key to the production of rivers,streams and many lakes (Cole and Caraco, 2001; Grey et al., 2001;Meyer, 1989; Richey et al., 2002; Tank et al., 2010). The fraction spe-cifically respired to CO2 in rivers and streams combined is thought tobe some 0.35 Pg C y−1 or some 12% of the total freshwater carbonbudget (Battin et al., 2008). In contrast, this cycling is ignored bythe IPCC, whose global models merely pipe terrestrially derived car-bon straight from a catchment to the oceans (IPCC. Solomon,Working Group, 2007). This assumption, which is clearly in need ofrevision, may in part be due to the fact that organic carbon leachedfrom the land was once regarded as recalcitrant and old (and someof it certainly is) and, as a consequence, was not capable of being sub-sequently ‘transformed’ through heterotrophic metabolism on reach-ing inland waters (Battin et al., 2008; Caraco et al., 2010; Cole andCaraco, 2001). Hence, research aimed at quantifying both the fluxand dynamics of the lateral exchange of organic carbon betweenfloodplains and rivers is timely and novel (Battin et al., 2008).

An increase in both the frequency and severity of storms, as onemanifestation of climate change, could increase the flux of carbonfrom the catchment. Such a process may already explain some ofthe pronounced losses of organic carbon from soils across the UK(Bellamy et al., 2005), which is consistent with the increased concen-tration and use of dissolved organic and inorganic carbon in lakes andrivers (Evans et al., 2005; Jones et al., 2001; Worrall et al., 2004). Fur-ther, the increased siltation and clogging of gravel beds (colmation),

loss of salmonid spawning grounds (Schalchli, 1992; Soulsby et al.,2001) and substantial methanogenesis in chalk streams, that are os-tensibly of good ecologically status (Sanders et al., 2007), may alsobe linked, in part, to elevated erosion. Conversely, increasingly flashyflows could ultimately increase the export of carbon from the catch-ment to the coast as retention in streams, rivers and estuaries isdecreased (Cole et al., 2007). How the full effects of changes in tem-perature, season, precipitation and runoff will propagate through riv-ers and their surrounding catchments will only be revealed by futureresearch but some interesting predictions relating to the direct effectsof temperature and indirect consequences of changes in riparianvegetation and light are being discussed (Mulholland et al., 2009b;Yvon-Durocher et al., 2010a).

There is clearly a need for a better understanding of the biogeo-chemistry of carbon in rivers, yet carbon is not cycled in isolation(Arango and Tank, 2008; Böhlke et al., 2009; Hill et al., 2000a;Manzoni et al., 2010). The familiar, yet arguably simplistic, Redfieldratio tells us that primary production requires a balanced flux of Cand the macro nutrients N and P at a ratio of 106C to 16N to 1P. Al-though there are comparable bodies of literature covering the bulkflux (or instantaneous load) of either N or P through aquatic ecosys-tems, we probably know more about specific aspects of the cyclingof N relative to P. This latter disparity in knowledge may reflect thesimple fact that we have more tools at our disposal (stable isotopes,inhibitors, pure cultures of microbes that mediate specific parts ofthe N cycle, more easily defined and quantifiable inorganic species)to study the dynamics of the N cycle compared to P. This focus on Nhas, in turn, tended to concentrate on the removal of anthropogenicnitrate via ‘denitrification’ at the fine (e.g. discrete samples of gravelor piezometers) and catchment-wide scales, rather than the completeN cycle (Burgin and Hamilton, 2007).

In a synthesis of the global magnitude of denitrification across avariety of landscapes it is estimated that rivers ‘denitrify’ annuallysome 30% of the total N that they receive, and that together ground-waters, lakes and rivers are responsible for some 20% of total globaldenitrification (Seitzinger et al., 2006; Wollheim et al., 2008).Hence, just as others would argue that rivers are not simply inert con-duits merely ‘piping’ carbon through the landscape, the same is clear-ly true for N (Battin et al., 2009; Cole et al., 2007; Davidson andSeitzinger, 2006; Groffman et al., 2006; Mulholland et al., 2008). In-deed, in a comprehensive analysis of available data for nitrate andboth dissolved and particulate organic carbon (DOC, POC) concentra-tions, Taylor and Townsend (2010) quite clearly demonstrated thatnitrate (a dominant form of anthropogenic nitrogen pollution) accu-mulates as the availability of organic carbon declines right along thehydraulic continuum from catchment to coast. Hence these two keymacronutrients are closely cycled, yet they are often studied inisolation.

1.2. Scope of the review

In this review we focus our attention on three broad themes. First,whereas the lateral and bidirectional exchange of water between riv-ers and the wider catchment is recognised, the overall effect of thisexchange on the net export of carbon and nitrogen from catchmentto coast is poorly constrained (Battin et al., 2009; Davidson andSeitzinger, 2006). Hence, we consider rivers in their broader land-scape setting via exchange processes and the potential implicationsthat this exchange has for carbon and nitrogen cycling in the riverto landscape continuum. Second, we look at quantifying the metabo-lism of both carbon and nitrogen in rivers, considering both estab-lished principles and some more novel aspects (metabolisms,techniques and the potential effects of warming). For example, recentdiscoveries of novel metabolic pathways in the N cycle mean that theproduction of N2 gas is no longer the sole domain of denitrification inaquatic ecosystems and the identification of hitherto unknown sinks

145M. Trimmer et al. / Science of the Total Environment 434 (2012) 143–158

for carbon in rivers has altered our understanding of the cycling ofboth of the key macro elements in rivers (Battin et al., 2009; Chapinet al., 2006; Ettwig et al., 2010; Thamdrup and Dalsgaard, 2008;Trimmer et al., 2009a). Thirdly, we highlight our novel findings onchemosynthetic production coupled to the oxidation of methane inthe chalk streams of southern England, which, if found to be wide-spread, could represent a paradigm shift in how we view productionin rivers.

2. Linking geomorphology, hydrology and biogeochemistry inbenthic sediments and the hyporheic zone

2.1. The role of benthic sediments

Although rivers can be conceptualised into the water column andthe bed, by far the greatest biogeochemical activity takes place in theriver bed, either at or just below the surface; though this is may notalways the case in large rivers (Brunke and Gonser, 1997; Clapcottand Barmuta, 2010; Fellows et al., 2006; Findlay, 1995; Oliver andMerrick, 2006). This is because the concentration of organic matterand associated microorganisms in sediments is typically several or-ders of magnitude greater than that in the overlying water (Findlay,2010; Tank et al., 2010). In addition, organic material and organismshave lower residence times in suspension than in the bed materialof a reach. Natural variation in surface flow patterns creates a ‘mosaic’of erosion and depositional areas which influences the distribution offine and coarse sediments, in turn, generating variability in river bedpermeability (Kasahara and Wondzell, 2003; Hatch et al., 2010;Malard et al., 2002). Regions of low flow facilitate sedimentation ofentrained particulate matter (organic or inorganic). In addition,stands of aquatic macrophytes can alter flow patterns, retard watervelocity and stimulate the sedimentation of considerable amountsof particulate material or entrap material moving along the riverbed (Cotton et al., 2006; Sand-Jensen, 1998). From a hydrologicalperspective, permeable benthic and hyporheic sediments can be con-sidered as reactive sieves or filters through which flowing watercarries particulate and dissolved material. So where the river bed ispermeable, entrained suspended particles and dissolved materialcan be transported into the bed on strong downwelling flows to bedeposited and processed at depth (Findlay, 1995; Hatch et al.,2010). Hence, the sediments are a focal point for material carried bythe river, concentrating both allochthonous and autochthonous or-ganic matter (Boulton et al., 1998; Cole et al., 2007; Meyer, 1989;Tank et al., 2010; Webster et al., 1999). Organic matter (particulateor dissolved, POC and DOC, respectively) in the sediment provideselectron donors to support an array of respiratory and fermentativereactions, with the former consuming oxygen and, in turn, alternativeelectron acceptors such as nitrate and sulphate once the oxygen is di-minished or fully depleted (Baker et al., 1999, 2000; Morrice et al.,2000). Whether the organic material enters the benthos from above,through lateral exchange, or from the groundwater, the rates of me-tabolism supported tend to be greatest at the point of delivery, decay-ing away there after (Brunke and Gonser, 1997; Hill et al., 1998a;Morrice et al., 2000). Although in this simple sense both DOC andPOC provide electron donors to sustain benthic metabolism, as out-lined, it is important to appreciate that the way in which both areconveyed along the aquatic continuum from catchment to coastand, in turn, how they are either assimilated, respired or buried arequite fundamentally different (Battin et al., 2008).

2.2. Hydrological exchange in permeable catchments

In permeable catchments, the subsurface vertical and lateral ex-changes between river water, underlying groundwater and the ripar-ian zone and/or floodplain import and export dissolved andparticulate organic matter, oxygen and nutrients. This drives cycles

of the key biotic macronutrients (C, H, N, P, S) and raises the produc-tivity of the system via a ‘dynamic ecotone’ at or beneath the sedi-ment surface (Baker et al., 2000; Gilbert et al., 1990). Hydrologicalflow paths are hierarchically nested in the river landscape from kilo-metres in length, linking terrestrial landscapes and river beds,through hundreds of metres (e.g. flows through channel bars and me-anders) to flow paths at the metre scale within the river bed arisingfrom subsurface hydrologic flow through and under geomorphicfeatures such as pool/riffle sequences or beds of submerged macro-phytes (Boulton et al., 2010). The interactions between geomorphol-ogy and the water level variations in the river and surroundinglandscape control the length, direction and hence network of subsur-face flows through benthic and hyporheic sediments, and influencethe extent and rates of biogeochemical processing (Poole, 2010). Forexample, the contribution of the hyporheic zone to the biogeochemi-cal budget of a river as a whole is governed by the balance betweensurface and subsurface flow and the intensity of the subsurface pro-cesses (Baker and Vervier, 2004; Boulton et al., 1998; Findlay,1995). Modelling and experimental studies have indicated thatshort metre-scale path lengths and short residence times appear todominate hyporheic flows in gravel-bed river systems, with an in-verse power–law relationship between frequency of occurrence andflow path length/residence time (Gooseff et al., 2003; Haggerty etal., 2002; Poole et al., 2008). Whether such frequency distributionshold for river systems located within other permeable substrate set-tings (e.g. sandstone and chalk geologies of the UK) remains to beseen.

Irrespective of landscape setting, the biogeochemical conse-quences of different distributions of flow path length in benthic andhyporheic sediments will depend on the respective rates of chemi-cal/biological reactions and advective flow. For fast rates of reactionand relatively slow flow, the distribution of products and reactantsalong a hydrological flow path may be largely independent of pathlength. However, for products/reactants involved in slower reactions,path length may be an important control on the evolution of porewater chemistry (Poole et al., 2008). Thus we can envisage thatthere will be a threshold, dependent on reaction relative to transportrate, at which path length becomes an important control on the bio-geochemical evolution of a hyporheic flow path. By definition, thisthreshold will be different for each chemical or biologically mediatedreaction. Similarly we can consider threshold in relation to residencetime. For example, in the gravel bar of a riffle-pool sequence,Zarnetske et al. (2011) observed a water residence threshold ofc. 7 h at which the hyporheic zone turned from a unit of net nitrateproduction to net nitrate removal. Such research highlights the im-portance of combining thermodynamic and hydrological approachesto improve our understanding of spatial and temporal biogeochemis-try in the hyporheic zone.

2.3. The landscape to river continuum

As indicated above, lateral exchange is not just limited to hypor-heic flows within the channel and riparian environment. Benthic bio-geochemistry will also be influenced by the lateral export of materialfrom the wider surrounding landscape (hillslopes, wetlands andfloodplains) through both overland and subsurface flow processes,examples of which are provided later in this manuscript. For the pur-poses of understanding biogeochemical processing in rivers, Fisher etal. (2004) conclude that rivers and their valleys should be consideredas a hydrologically integrated component of the landscape and notsubdivided into terrestrial and aquatic compartments. SimilarlyPinay et al. (2002) define rivers as open ecosystems linked longitudi-nally, laterally and vertically by hydrological and geomorphologicalprocesses. Both perceptual models of the landscape highlight the im-portant role that integrated catchment-scale management has to playin tackling the water quality and river habitat issues caused by


macronutrient enrichment. Pinay et al. (2002) emphasise the impor-tance of three factors regulating the transfer and cycling of macronu-trients into rivers at the landscape scale: mode of delivery, length andduration of contact time between flow and sediments (akin to flowpath length and residence time in the hyporheic literature) and theimportance of floods and droughts. Both Pinay et al. (2002) andFisher et al. (2004) highlight the importance of flooding for resettingsediment structure and texture both within the channel and flood-plain; consequently such disturbances alter lateral flow paths andresidence times of nutrients within the landscape, with potential con-sequences, as previously discussed for nutrient fluxes along thecatchment to coast continuum.

In summary, the extent to which the dynamics of river biogeo-chemistry are controlled by fluvial geomorphology and hydrologicalconnectivity is now an active research focus in river ecology. The im-portance of the interactions between geomorphology, ecology andhydrology, and the consequences for biogeochemical processing arenow recognised across a wide range of spatial and temporal scales.An important contemporary research challenge is to determine theextent to which lateral exchanges of material in different landscapesettings influence longitudinal river bed carbon and nitrogen fluxesand cycling (Boulton et al., 2010; Datry et al., 2007; Gilbert et al.,1990; Krause et al., 2010).

3. Carbon and nitrogen metabolism in rivers

3.1. Quantifying carbon metabolism, new directions and some functionalmetrics

The headwaters of rivers are usually supersaturated with carbondioxide and sometimes with methane, and this subsequently declineswith distance downstream as the river water re-equilibrates with theatmosphere (Cole and Caraco, 2001; Hope et al., 2001; Jones andMulholland, 1998b). Such supersaturation is a product of respirationin the surrounding catchment, conveyed to the river by lateral ex-change, and respiration in the river itself driven by autochthonousand allochthonous carbon (Cole et al., 2007; Richey et al., 2002;Tank et al., 2010; Webster and Meyer, 1997). Some have used the su-persaturation of rivers with carbon gases as an integral measure ofcatchment respiration but it is hard to distinguish between the terres-trial and aquatic sources of carbon gases in rivers (Cole et al., 2007;Jones and Mulholland, 1998a). Hence, the measurement of such su-persaturation and or ‘outgassing’ of carbon gases from rivers doesnot tell us much about the true magnitude or dynamics of carbon cy-cling in either compartment of the river and its catchment (Cole et al.,2007).

Beyond the fact that rivers are supersaturated with carbon gases, amore complete understanding of the biogeochemical cycling of car-bon requires quantification of ‘metabolism’ within the river itself(Tank et al., 2010). Here ‘metabolism’ is taken as the overall balancebetween carbon produced through Gross Primary Production (GPP)and that lost or respired through (total) Ecosystem Respiration (ER)by the autotrophs and other biota. These measures of whole systemmetabolism provide an integral measure of both carbon (energy)flux and broader ecosystem processes in a river (Izagirre et al.,2008). The ratio of respiration to production indicates whether ornot a river is broadly ‘autotrophic’, where GPP>ER, or heterotrophic,and ER>GPP. If GPP>ER then this would also indicate net ecosystemproduction (NEP), and ER>GPP that river metabolism is supportedby inputs of carbon from the catchment and is almost always thecase (Meyer, 1989; Tank et al., 2010).

Although these metabolic processes are directly concerned with ei-ther the fixing of inorganic carbon (photosynthesis) into sugars or theremineralisation of organic carbon (respiration) back to CO2, few au-thors have quantified either component directly as a flux of CO2 eitherinto or out of sediments, or the river as a whole. Most have inferred

such carbon metabolism indirectly from changes in oxygen concentra-tion (consumption being used as a measure of respiration and the ox-idation of organic matter to CO2) or through the production ofoxygen, primary production and the fixation of CO2 (del Giorgio andWilliams, 2005; Johnson et al., 2010). This is in part due to the difficultyof measuring accurately a significant change in the concentration ofCO2, either in the river water itself or across the water/air interface, rel-ative to that for oxygen but this is changingwith the application of newtechnologies (del Giorgio and Williams, 2005; Johnson et al., 2010).These common concepts of GPP, ER and NEP, integral to the conceptof whole river carbon metabolism, may not capture all of the carbonflux as some carbon may ‘leak’ from the river (Chapin et al., 2006).Chapin et al. (2006) argue that leaching and lateral transfer of dis-solved organic and inorganic carbon, release of volatile organics and ef-flux of methane and carbon monoxide (amongst others), can allcontribute to the overall net ecosystem carbon budget (NECB) whichextends beyond the concept of NEP (Middelburg et al., 1996; Richand King, 1999; Ruuskanen et al., 2011). Such processes really do em-phasise Battin et al.'s (2009) notion of a truly ‘boundless’ carbon cycle,yet the role of such carbon loss processes in freshwater aquatic ecosys-tems has barely been considered (Aufdenkampe et al., 2011).

The consumption of oxygen provides us with a measure of aerobicrespiration coupled to the oxidation of organic carbon compounds. Inaddition, through the re-oxidation of a variety of reduced inorganicchemical species (e.g. NH4

+, NO2−, H2S, H+, Fe2+, Mn2+), it can also

‘integrate’ anaerobic respiratory pathways and give us an overallmeasure of ‘ecosystem respiration’ (Glud, 2008). The consumptionof oxygen can, in turn, be used to estimate the production of CO2 byassuming that 1 mol of O2 consumed equates to 1 mol of C oxidisedi.e., a respiratory quotient (RQ) of 1:1, but see Glud (2008) for arange from 0.8 to 1.2 and a fuller discussion. Conversely, oxygen pro-duction is taken as a measure of photosynthesis, with its yield beingproportional to the amount of CO2 fixed.

The simplest form of oxygen change measurement is that of theBiological (or Biochemical) Oxygen Demand or BOD assay, whichgives a measure of the potential of a sample of water to consume ox-ygen over a fixed time at a standard temperature, and is a staple partof many routine monitoring exercises. It does, however, only providean indication of the quantity and quality of organic material in thewater itself, telling us nothing about the bed, the system as a wholeor any real functional dynamics. In the vast majority of rivers, theconsumption of oxygen will be dominated by the sediments, and sev-eral have measured this consumption simply by incubating isolatedand homogenized samples of sediment in the laboratory (Hill et al.,1998b, 2000b, 2002). Although such a simple assay can be deployedquickly and cheaply across hundreds of field sites, it only measurespotential benthic respiration and may not reflect true benthic respira-tion i.e. that realised where advective flow, redox and hydraulic gra-dients are intact.

Alternatively, others have used some variant of a benthic chamberto measure both production and consumption of oxygen under condi-tions more representative of those in situ, enabling estimates of pri-mary productivity and respiration to be derived. Such techniquescontinue to be modified (Bott et al., 1997; Bunn et al., 1999; Uzarskiet al., 2001). There are caveats associated with benthic chambers, asthere are with all methods employed to study any aspect of biogeo-chemistry (Glud et al., 2009; Young and Collier, 2009). Isolating thebed from river flow will restrict the advective flow of water throughthe interstices and probably underestimate true rates of benthic oxy-gen consumption (Cook et al., 2007). There are other concerns overlight and nutrient concentration, though the latter seem minor insome fine sand sediments (Trimmer et al., 2009b).

A technique that can integrate processes over comparatively largescales, such as an entire reach, will have advantages over those thatoperate at the scale of a single patch of gravel or periphyton, particu-larly if it can also keep intact advective flow through the bed and


deeper hyporheic. To this end, a non-invasive method of measuringdiel changes in oxygen in the water column alone (either single ordual stations) to integrate whole river, or reach scale, metabolismby the sediments and water column combined was developed (Bottet al., 1978; Marzolf et al., 1994; Odum, 1956). Though this has ad-vantages over benthic chambers the open diel change technique re-quires careful measurement of the rate of re-aeration across thewater to air interface and any ingress of groundwater along a gainingreach needs also to be taken into account (Demars et al., 2011; Halland Tank, 2005). However, quantifying re-aeration and groundwaterinflow involves complex tracer studies that provides stream reach av-erages for these parameters, whereas the dissolved oxygen is typical-ly only measured at one or two points along the reach. This leads to adiscrepancy between the empirical data and the underlying assump-tions of this method (Demars et al., 2011). A technique that circum-vents these problems and limitations is the Eddy Correlation (EC)technique, which has been used for atmospheric flux measurementsfor over 60 years, but was only pioneered in the aquatic environmentby Berg et al. in, 2003. EC has recently been used for determining ben-thic oxygen exchange in coastal sandy and muddy sediments and, thelargely unexplored, rocky substrata (Berg et al., 2003, 2007; Glud etal., 2010) as well as in deep ocean and lake sediments (Berg et al.,2009; McGinnis et al., 2008). The typical ‘footprint’ of an EC flux is a~40 m long and ~1 m wide oval shaped area upstream of the EC in-strument, although this ‘footprint’ can vary several fold dependingon site specific parameters (Berg et al., 2007). Since the EC techniqueprovides a direct non-invasive measurement of the ‘true’ in situ fluxof oxygen it is unaffected by the potential artefacts of the oxygendiel change technique and benthic chamber incubations describedabove.

Furthermore, current progress in predicting the importance oflateral exchanges for river metabolism is limited by the needs of si-multaneous measurements of hydrology and metabolism at the ap-propriate scales. Battin et al.(2003) estimated that exchange withthe hyporheic zone can contribute between 40% and 90% of river me-tabolism, but the effect of wider landscape-scale lateral exchange onwhole river metabolism is as yet undetermined.

To date the application of the widely established open diel oxygenchange technique has highlighted various factors that appear to influ-ence whole river metabolism. These include light, nutrient availabilityand the quality and quantity of organic matter (Grimm and Fisher,1986; Mulholland et al., 2001; Webster and Meyer, 1997). This pro-gress in understanding the local factors which influence river metab-olism camemostly from studies of single rivers or regions. However, amore sophisticated assessment of how the interactions betweenproximal and distal controls (such as land use and geographic region)affect river metabolism has recently been revealed by measuringwhole river metabolism in nine rivers in each of eight regions acrossthe continental United States and Puerto Rico (Bernot et al., 2010).This wide scale measurement of diel oxygen change, combined withsophisticated statistical modelling, revealed complex regulation ofboth GPP and ER, which ultimately suggested that the impact of agri-culture and urban land use can mask regional differences in river me-tabolism (Bernot et al., 2010). With such clear impacts fromanthropogenic activities in the landscape, some have proposed theuse of whole river metabolism as a functional metric for assessingecosystem or river ‘health’, which in the long term could offer consid-erable time and cost benefits over more routinely applied structuralmetrics (Wright et al., 1998; Fellows et al., 2006; Young et al.,2008). Indeed, through an extensive analysis of published GPP andER data from 213 reference and 82 impacted rivers, Young et al.(2008) developed a robust framework for assessing and distinguish-ing between rivers of poor, satisfactory or healthy status.

Whilst there are clear advantages to the use of whole-river metab-olism, benthic chambers do still have a use. Benthic chambers can beused to study dynamics associated with a particular ‘patch’ of the

river bed that the open diel technique would miss. For example, it iswell known that macrophytes can alter the flow regime of rivers,with water velocity being markedly reduced within a stand comparedto that in the open channel, aiding the deposition of fine particulatematerial (Cotton et al., 2006; Sand-Jensen, 1998). Such depositioncan stimulate very high rates of benthic respiration, ranking somepatches of these chalk streams as poor according to Young et al.(2008), and resulting in an efflux of CO2 and CH4 comparable tothat in some heavily impacted estuaries and peat bogs (Sanders etal., 2007; Trimmer et al., 2009b). In turn, the measured efflux of CO2

could account for a considerable proportion (up to 57%) of the organiccarbon and associated N being mineralised whilst it is resident be-neath the plants, before the remainder is presumably exporteddown stream (Trimmer et al., 2009b). Such benthic respirationwould ‘show up’ in an open diel curve measurement but only benthicchambers could identify the ‘hot-spots’ and true nature of the impactsand fate of mobilized soil from the catchment in the river bed(Walling, 2005).

3.2. Relationships between temperature and the carbon cycle

Recent work has indicated that increases in temperature forecastby the IPCC (Solomon, Working Group, 2007) for the coming decadesmay have profound implications for the cycling of carbon in aquaticecosystems due to the differential temperature dependencies of car-bon fixation by gross primary production (GPP) and carbon minerali-sation by ecosystem respiration (ER). For example, Yvon-Durocher etal. (2010b) showed that warming of 4 °C reduced the carbon seques-tration capacity of freshwater mesocosms by 13%, shifting them to-wards net heterotrophy (i.e. net sources of CO2 to the atmosphere)because ER responded more strongly to temperature than GPP. Fur-ther, they found that methanogenesis responded even more stronglythan ER or GPP, with 20% more of the GPP being accounted for bymethane emissions with 4 °C of warming (Yvon-Durocher et al.,2011). Moreover, they found that the precise magnitude of theseshifts could be predicted by the average activation energies of photo-synthesis, respiration and methanogenesis (e.g. ~0.32 eV, 0.65 eV and0.85 eV respectively, where 1 eV=96.49 kJ mol−1).

Here, using 296 measurements of whole river metabolism (com-piled from the literature) we obtained similar results for the regionalto global scale temperature dependence of ER in rivers to those foundin our freshwater mesocosms (Fig. 1 and see Yvon-Durocher et al.,2010a). Similarly, studies by Acuna et al. (2008) and Demars et al.(2011) also demonstrated that the temperature dependencies ofwhole river metabolism, and benthic river metabolism respectively,also conformed to the average activation energy for respiration pre-dicted by the metabolic theory of ecology (Brown et al., 2004).

Our work hints at the tantalising prospect that understanding theeffects of temperature on the carbon cycle may be less complex thanone may at first expect. However, although temperature is often a keydriver of ecological processes in freshwaters, it rarely acts in isolation,and if we are to completely understand the multifarious implicationsof climate change – of which temperature change is only one element– comprehensive, long term assessments of the interactions betweentemperature, hydrological processes, catchment delivery of nutrientsand land-use change will be key to accurately forecasting the re-sponse of ecosystem metabolism to climate change in freshwaters(Bernot et al., 2010; Marcarelli et al., 2010; Mulholland andWebster, 2010). In contrast to our strong respiration–temperature re-lationship, previous analyses of ER in rivers (or streams) argued foronly a weak or modest dependency between temperature and ER(Young et al., 2008). This difference can be reconciled by the factthat Young et al. (2008) focus on site level respiration–temperaturerelationships. At the site level the relationship between seasonaltemperature change and ER can be modulated by other factors that

-2 -1 0 1 2

-4

-3

-2

-1

0

1

2

Standardised Temperature : 1 kTc 1 kT

Sta

ndar

dise

d R

ate

: ln(

RT

OT

(T)

RT

OT(T

C))

0 10 20 30

E = 0.5eVRiver metabolism (n = 263)

Temperature ( C)

Fig. 1. The temperature dependence of river metabolism. Data were compiled fromBott et al.(1985); Dawson et al.(2001); Kowalczewski and Lack(1971); Uehlin-ger(2006) by digitising the figures. The short-term temperature dependence of riverrespiration was analysed as follows. Seasonal respiration rates RTOT, and temperature,T, for multiple rivers were analysed by expressing river respiration as a standardisedrate,ln Rs(T)=ln[RTOT(T)/RTOT(TC)]=E(1/kT−1/kTC), where RTOT(TC) is the respirationrate in a river at some fixed temperature, TC, here we have used 288 K=15 °C, T is ab-solute temperature (K), k is the Boltzmann constant (8.62×10−5 eV K−1), and E is anactivation energy (1 eV=96.49 kJ mol−1). Data were standardised by fitting a linearmixed-effects model of the form lnRTOT(T)=E(1/kTC−1/kT)+ lnRTOT(TC) using thenlme package in R version 2.11.1. In this analysis, E and ln RTOT(Tc) were both treatedas random variables that varied by river. This modelling procedure provided an overallestimate for the global scale activation energy, E=0.50 (95% CI=0.36 to 0.62 eV), it isimportant to appreciate that this was not derived from the standardised fluxes whichwere used only to facilitate visualisation of the data. It also provided site-specific esti-mates for E and ln RTOT(Tc), which were used to generate the standardised rates of riverrespiration, In RS(T), plotted in the figure.


seasonally covary with temperature— e.g. timing of allochthonous in-puts, discharge etc.

Rivers and estuaries receive allochthonous organic carbon firstfixed in the terrestrial realm, which fuel metabolism and heterotro-phic biomass production far beyond that which could be sustainedfrom autochthonous production alone (Battin et al., 2008; Meyer,1989). Moreover, recent work highlights that a significant proportionof these carbon subsidies are bioavailable and, therefore, rates of car-bon mineralisation and CO2 efflux to the atmosphere, may be partic-ularly responsive to temperature in rivers and streams (Battin et al.,2008; Caraco et al., 2010; Richey et al., 2002). In contrast, for ecosys-tems where the mineralisation of organic carbon by ER is tightlycoupled to its fixation by GPP – e.g. most terrestrial ecosystems(Allen et al., 2005), or the open ocean (del Giorgio and Williams,2005) – long-term increases in temperature are expected to resultin a steady state between ER and GPP (Allen et al., 2005). In theopen ocean or on the land no positive feedback between warmingand the carbon cycle is expected because, despite the fact that respi-ration responds more strongly than GPP to increases in temperature,catabolic metabolism cannot burn more carbon than is first fixed inthat ecosystem (Yvon-Durocher et al., 2010a).

Such carbon dynamics, however, cannot be said to hold for mostaquatic ecosystems, especially not rivers which receive significantsubsidies of organic carbon from their surrounding catchments(Cole and Caraco, 2001). Thus, with global warming, in rivers whereER and GPP are decoupled by allochthonous carbon subsides, thereis the potential for them to act as even greater net sources of CO2 tothe atmosphere, because ER is not constrained by the weaker temper-ature dependence of GPP (Battin et al., 2008; Yvon-Durocher et al.,

2010a). Finally, even though CH4 is an atmospheric trace gas itsgreenhouse warming potential (GWP) is 20 times that of CO2. Recentrevisions to the inland freshwaters carbon gas budgets (lakes and res-ervoirs) argue that CH4 is on a par with CO2 in terms of GWP but ouranalysis suggests that it could potentially become dominant with onlymoderate warming. There is a paucity of data for CH4 emissions fromrivers, compared to lakes, yet with their allochthonous inputs of or-ganic carbon their potential for elevated emissions of CH4 under fu-ture warming could be significant (Bastviken et al., 2011; Tranvik etal., 2009; Yvon-Durocher et al., 2011).

3.3. Pathways of N removal: the old and the new

A variety of microbial metabolic pathways and interacting abioticreactions transform and regulate the bioavailability of N in aquaticecosystems, including N fixation, ammonification, denitrification, an-aerobic ammonium oxidation (anammox), nitrification, dissimilatoryreduction of NO3

−to NH4+ (DNRA), inorganic N assimilation into plant

or microbial biomass (immobilization), NH3 volatilization, NH4+ ad-

sorption and desorption (Reddy and DeLaune, 2008; Thamdrup andDalsgaard, 2008). Some of these microbial processes are supportedby the reduction of the oxidised species of N, such as nitrate, in anoxicsediments and waterlogged soils, whilst others derive energy fromthe aerobic oxidation of reduced species such as ammonium(Painter, 1970; Shapleigh, 2006). Here we focus our attention on pro-cesses that represent a net sink for N. This is because the availabilityof nitrogen can limit primary production and hence the availabilityof energy at the base of the food web, but also because N removalvia ‘denitrification’ throughout the landscape can help mitigate theimpacts of anthropogenic N further downstream (Seitzinger et al.,2006).

Formerly the production of N2 gas was thought to be mediatedsolely by denitrification, i.e. the biological conversion of bioavailablenitrate (NO3

−+NO2−or NOx

−) into nitrous oxide (N2O) and di-nitrogen gas (N2), the latter being unavailable to the vast majorityof fauna and flora in the biosphere. With the discovery of anammoxas an alternative route of N2 production in marine and estuarine sed-iments whereby ammonium is oxidised by nitrite, and even more re-cently the anaerobic oxidation of methane, also at the expense ofnitrite, in freshwater sediments, this perception has changed(Ettwig et al., 2010; Thamdrup and Dalsgaard, 2002). However, notonly do these ‘new’ metabolic processes alter our perception of theN cycle in all aquatic ecosystems but they undermine the principleson which many of the quantitative techniques employed to measurethe N cycle are built (Risgaard-Petersen et al., 2003). For example,whilst the application of 15N became popular for studying N cyclingin estuarine and coastal sediments during the late 1980s and early1990s, the earlier technique of acetylene blockage (Yoshinari et al.,1977; Sørensen, 1978), employed to measure denitrification as an ac-cumulation of nitrous oxide in marine sediments, remained popularin the freshwater literature for a good deal longer (see the reviewby Mulholland and Webster, 2010 and Arango et al., 2007); thoughthis bias is changing with the large scale application of 15N acrossmultiple rivers and catchments (Bohlke et al., 2004; Mulholland etal., 2004, 2008, 2009a).

It is important to appreciate, in the case of measuring the produc-tion of N2, that the anammox reaction simply does not make nitrousoxide. Further, some authors have argued that from a net N2 produc-tion perspective the metabolic pathway involved in the production N2

is of little consequence (Chang and Devol, 2009). However, denitrifi-cation is predominantly a heterotrophic pathway, requiring a supplyof organic carbon, and, in contrast, anammox is chemoautotrophic,yet what regulates the production of N2 via either pathway remainspoorly defined (Thamdrup and Dalsgaard, 2008). If we are to have afully integrated understanding of C, N and P cycling from catchmentto coast, then we must recognise these new biogeochemical


pathways and develop techniques to measure them accordingly(Battin et al., 2009; Davidson and Seitzinger, 2006). Although theanammox reaction is widespread in marine and estuarine sediments,and the putative bacteria responsible have been widely characterisedin freshwater habitats, its role in the production of N2 gas remainspoorly characterised in both lentic and lotic freshwaters (Dalsgaardet al., 2005; Hamersley et al., 2009; Nicholls and Trimmer, 2009;Penton et al., 2006; Schubert et al., 2006; Zhu et al., 2010).

3.3.1. Riverine and hyporheic ‘denitrification’: a brief overviewThere are already numerous reviews covering all aspects of re-

search carried out to date on ‘denitrification’ in rivers and their adja-cent landscapes (Mulholland and Webster, 2010; Pina-Ochoa andAlvarez-Cobelas, 2006; Seitzinger et al., 2006; Wollheim et al.,2008). Much of this research has employed widely differing tech-niques to quantify denitrification at different temporal and spatialscales. Despite these differences, and the difficulties remainingin quantifying the true magnitude of denitrification in all aquaticecosystems, it is apparent that considerable amounts of fixed N areprocessed biologically and ultimately removed as it journeysfrom catchment to coast (Battin et al., 2009; Cole et al., 2007;Davidson and Seitzinger, 2006; Groffman et al., 2006; Mulholland etal., 2008).

With regard to the regulation of denitrification, there is much ev-idence to suggest that denitrification activity is governed either bythe availability of organic electron donors (particulate or dissolved)and/or electron acceptors e.g. nitrate or nitrite at the habitat or finescale (Arango et al., 2007; Baker and Vervier, 2004; García-Ruiz etal., 1998; Jones, 1995; Mulholland et al., 2008). It is important to ap-preciate that the concentration of nitrate will seldom be low enoughin the sediment pore water to limit the specific rate of denitrificationsensu Michaelis–Menten, more that the total availability of nitratelimits the size of the denitrification zone sensu Liebig's Law (Meyeret al., 2005). At the reach scale (10s of metres) on the other hand,the proportion of nitrate removed by ‘denitrification’ is governedmore by river discharge and total nitrate flux, where high dischargeequates to a low residence time or contact time between sediments(surface or hyporheic) and the load of nitrate per unit area of sedi-ment (Alexander et al., 2009; Böhlke et al., 2009; Seitzinger et al.,2006). Even though denitrification may be operating at its apparent‘Vmax’ in the field, high loads (and/or low residence times) can resultin a comparatively small turnover of the total nitrate load. From ariver continuum perspective, the relative significance of headwatersto lower reaches for N retention can be rationalised as a function ofnitrate loading. Lower reaches gain in significance as higher nitrateloadings saturate the headwaters and a greater proportion of theoverall nitrate load is exported downstream (Mulholland andWebster, 2010).

Despite considerable effort and progress to date, there are stillrecognised gaps in our knowledge about the true magnitude ofdenitrification, or overall production of N2, questions that may onlybe answered by a truly integrated and multidisciplinary approach toquantifying the flux of N from catchment to coast (Böhlke et al.,2009; Davidson and Seitzinger, 2006).

As outlined above, the hyporheic zone is a well recognised biogeo-chemical feature of permeable catchments, attracting increased re-search interest in recent years (Krause et al., 2010). The broadecological significance of water and nutrient exchange through thehyporheic zone has been highlighted, and it is a widely discussedhabitat and refugium for a range of organisms (Brunke and Gonser,1997; Robertson andWood, 2010; Schmid-Araya, 1994). More specif-ically, the hyporheic zone can function as either a source or sink fornitrogen, depending upon the speciation of nitrogen present, theavailability of reactive substrates and prevailing oxygen status(Storey et al., 2004 and references therein). A good deal of researchhas focused on understanding the dynamics of nitrate production

and consumption within hyporheic zones across a variety of river sys-tems (Duff and Triska, 2000; Hill et al., 1998a; Holmes et al., 1996).The clear ecological significance of the limitation of primary produc-tion by N has stimulated research into ‘denitrification’ in the hypor-heic zone. As a facultative anaerobic respiratory pathway,denitrification is broadly dependent upon the absence of oxygen,but appears to occur within anoxic micro-sites in otherwise appar-ently well oxygenated sediments (Baker et al., 2000; Storey et al.,1999; Triska et al., 1993).

More recently, the focus has widened to include the applied as-pects of the role of hyporheic sediments in attenuating anthropogen-ically derived nitrate as the surface and groundwaters mix andexchange (Mulholland et al., 2008, 2009a; Rivett et al., 2008). Indeed,some argue that the attenuation of nitrate in the hyporheic zonecould be vital for maintaining the ecological status of groundwater-fed rivers (Smith et al., 2004). However, with climate change, regionalrainfall distributions, water resource and nutrient yields are likely tochange, which could reduce the effectiveness of the hyporheic zonein nutrient mitigation (Wilby et al., 2006). If climate change deliverswarmer, drier summers then this could affect biogeochemical proces-sing in the hyporheic zone of groundwater-fed rivers, particularlyunder baseflow conditions, yet our understanding of nutrient trans-formations and the inter-relationships with groundwater fluxthrough this zone is still lacking (Conant et al., 2004; Gooseff et al.,2005; Hill et al., 1998a).

Despite the need to understand the biogeochemical significance ofvertical and lateral exchange of waters through the hyporheic zone,and any concomitant removal of fixed N via the production of N2

gas, most research on hyporheic denitrification has focussed on theupper few centimetres of sediments, and as such ignores the rolethat groundwater flows may play in the chemical reactivity of thiszone (Wondzell, 2006). This focus on the surface sediments is inpart justified, as microbial metabolism in general appears to be limit-ed by the availability of organic carbon at depth. Hence, any ‘denitri-fication’ activity will probably be concentrated towards the surface,where the availability of both electron donors and acceptors is likelyto be greatest (Baker et al., 2000; Holmes et al., 1996). However, italso reflects the logistical challenges of recovering sediment fromdepth, especially in armoured river beds (Storey et al., 2004). A fewstudies have demonstrated considerable potential for denitrificationat greater than 25 cm into the riverbed (LeFebvre et al., 2004).

Recently, the irreversible reduction of resazurin to the stronglyfluorescent compound resorufin, arising chiefly from aerobic respira-tion, has been used at the whole reach scale to assess the relativeimportance of sediment–water interaction in stream ecosystems(Haggerty et al., 2009). Theoretically the reaction could also be har-nessed to assess changes in metabolic activity in the hyporheic withdepth, and thus indicate where the potential for denitrification oc-curs. Similarly, the hydrolysis of fluorescein diacetate (FDA) hasbeen used to detect metabolically active zones in surface and hypor-heic sediments of streams (Atkinson et al., 2008; Battin, 1997). Futuredevelopments of these ‘smart’ tracer (sensu Haggerty et al., 2008)techniques may give rise to a suite of useful tools by which total mi-crobial activity and potential for nitrogen transformations in thehyporheic zone can be assessed from reach to patch scale.

Further, a differing response to simple organic substrates acrosssediment strata, suggests that the fate of nitrate and dissolved organiccarbon within the hyporheic zone could depend upon where in thesediment profile the final point of processing occurs (Pfenning andMcMahon, 1997; Sobczak et al., 2003). We still know comparativelylittle about ‘denitrification’ or other pathways of nitrate reduction atdepth in the hyporheic zone, or beneath contrasting habitats suchas macrophyte stands and or riffle/pool sequences. An increase infloods and flow could bury organic carbon at depth in the hyporheiczone, resulting in denitrification occurring at a greater distancealong hydrological flow path.


3.3.2. Quantifying the reduction and removal of nitrate in riverbedsediments

Denitrification of nitrate to N2 gas is not the only possible route forthe reduction of nitrate in river sediments, but the relative lack ofknowledge relating to alternative pathways of dissimilatory nitrate re-duction may be due to the former widespread use of the acetyleneblock technique (Duff and Triska, 2000; Sørensen, 1978; Yoshinari etal., 1977). This perspective is changing, however, as the cost and techni-cal challenges of applying the 15N technique aremet (Bohlke et al., 2004;Mulholland et al., 2004, 2008; Mulholland and Webster, 2010). Whilstacetylene block gives a relative measure of denitrification potential, itcannot capture N2 production via anammox, nor can it measure directlyany dissimilatory reduction of nitrate to ammonium (DNRA) or assimi-latory reduction of nitrate (Duff and Triska, 2000; Kelso et al., 1999;Risgaard-Petersen et al., 2003). Even though the latter pathways of ni-trate reduction have rarely been measured directly in freshwater sedi-ments, and the existing evidence is equivocal, we may risk neglectingthese pathways as being of little importance in subsurface environments(Bohlke et al., 2004; Burgin and Hamilton, 2007; Rivett et al., 2008).

Historically, in addition to acetylene block, the production of N2

gas has been measured directly (mainly by researchers in NorthAmerica) as a flux of N2, or by changes in the ratio of N2 to Ar (Kanaet al., 1994; Seitzinger et al., 1980). Alternatively, it has been quanti-fied indirectly by measuring the production of 15N2 gas after applica-tion of 15NO3

−(mainly by European groups), where the turnover ofthe 15N spike is assumed to be proportional to that of any ambient14NO3

−and the concomitant production of 14N2 gas (Nielsen, 1992).Indeed, Nielsen's, 1992 isotope pairing technique (IPT) is probably themost widely applied and frequently cited 15N isotope tool to date. De-spite the widespread application of either of these two techniques tothe quantification of N2 production in estuarine and marine sediments,comparatively few measurements have been made in rivers. Further,the discovery of anammox inmarine sediments undermined the centralassumptions on which the original IPT was built, and the technique hasrecently been revised (Risgaard-Petersen et al., 2003; Thamdrup andDalsgaard, 2002; Trimmer et al., 2006). In comparison to carbon though,the overall metabolism of which is often inferred indirectly fromchanges in oxygen, the approach for measuring a flux of N2 is more di-rect, be it as a flux of 14N2 or that traced in its heavier form of 15N2.

The great majority of studies to date which have applied somevariant of Nielsen's (1992) original isotope pairing technique or its re-visions (Risgaard-Petersen et al., 2003; Trimmer et al., 2006) haveaimed at measuring the production of N2 gas in estuarine and marinesediments (Steingruber et al., 2001). Such sediments are typically co-hesive mud or muddy-sand, having a clearly defined sediment–waterinterface, across which the exchange of solutes is governed by molec-ular diffusion. Thus, in its simplest form, the IPT would be suitable forquantifying the production of N2 gas in soft bottomed rivers. Howev-er, in permeable sandy (median particle size>200 μm) or coarser(gravels, cobbles, boulders) riverbed sediments the exchange of sol-utes and particles will be driven by advective transport, which cangreatly enhance both the metabolism and filtration of detritus perunit area of sediment (Rusch and Huettel, 2000).

Whether or not the biogeochemical cycling of N in permeableriver bed sediments is to be studied genuinely in situ, or in the labo-ratory, this interstitial flow should be included and or replicated inthe experimental design. So far the application of techniques to cap-ture the influence of advective flow on both oxygen consumptionand N2 production, using the IPT, has received particular attentionin coarse sandy marine sediments, where it was demonstrated thatboth the consumption of oxygen and denitrification increased two-and seven-fold, respectively, with increasing advective flow (Cooket al., 2006, 2007; Gihring et al., 2010).

Such advective flow is obviously central in modulating hyporheicprocesses (Findlay, 1995; Hill et al., 1998a; Palmer, 1993) and there-fore some have advocated the use of the open diel oxygen change

technique for quantifying whole reach or whole river metabolism,as discussed previously. Few have attempted such an open dielchange technique for quantifying reach scale production of N2

(Laursen and Seitzinger, 2002, 2004, 2005), with most of the researchinto river bed production of N2 being based on the incubation of dis-crete samples of sediment in the laboratory or some form of ‘push–pull’ technique to quantify production of N2 in situ (Duff and Triska,1990; Garcia-Ruiz et al., 1998; García-Ruiz et al., 1998; Hill et al.,1998a; Sanders and Trimmer, 2006). Sheibley et al. (2003) replicated‘flow’ in recovered intact sediment cores (20 cm of medium- tocoarse-grained sands) to measure denitrification, whilst others haveused more artificial repacked filtration columns to measure potentialnitrate removal only (Mermillod-Blondin et al., 2005). Where intactcore recovery is more difficult, in gravels and cobbles, for instance,‘colonisation and recovery’ corers have also been used to investigatedenitrification and attempts have also been made to recreate thehyporheic zone in flow through mesocosms (Storey et al., 2004).

3.3.3. Reach scales measurements of denitrification and estimates of theglobal rate in rivers

It is from such fine scale studies that we have gained an improvedunderstanding of the factors which regulate denitrification at thehabitat scale. As with many aspects of environmental/ecological sci-ence, however, the challenge is one of scaling-up to make the datamore realistic and representative of the real world or, in this case, areach or whole river. In a series of papers in the early 2000s Laursenand Seitzinger developed and assessed a combined open diel andmodelling technique for quantifying reach scale production of N2

and N2O gases, alongside parallel measurements of dissolved oxygen(Laursen and Seitzinger, 2002, 2004).

In short, this technique seems most sensitive and suitable forquantifying the production of N2 in shallow reaches under low windvelocities. However, with a limit of detection of between 30 and100 μmol N m−2 h−1 it should be more than capable of resolving de-nitrification in agricultural catchments, where inputs of nitrate arehigh and denitrification likely to be intense (Laursen and Seitzinger,2005; Sanders and Trimmer, 2006). Laursen and Seitzinger (2005)also highlighted the problems associated with ingress of N2 enrichedgroundwater along gaining reaches in permeable catchments (as perthe open diel oxygen change technique) and suggested that supple-mentary knowledge relating to the groundwater characteristics ofany particular study reach would be beneficial (Mariotti et al.,1988). Others developed a combined whole reach 15N and modellingapproach, which appears capable of resolving lower rates of denitrifi-cation (~10 μmol N m−2 h−1) and is better suited to distinguishingbetween genuine in situ production of N2 and that brought in by in-gress from groundwater (Bohlke et al., 2004; Mulholland et al.,2004). Since these trials, the technique has been used in at least 72rivers across several regions of the USA, measuring rates of denitrifi-cation at >100 μmol N m−2 h−1 where nitrate concentration was>70 μmol L−1 (>1 mg N L−1) (Mulholland et al., 2008, 2009a).

Whole reach 15N techniques have also been used to study spatial dy-namics of nitrogen transformations at the patch or sub-reach scale. Inthis approach whole-stream steady state injections of 15NO3

−and a con-servative tracer are performed (Mulholland et al., 2004) and key sol-utes, 15N isotope samples (e.g. 15NO3

−, 15N2O and 15N2) and electricalconductivity are measured in hyporheic piezometers of interestthroughout the steady state time period (Zarnetske et al., 2011). Al-though these field experiments do not directly measure in-situ ratesof N transformation in the hyporheic, they do enable researchers toidentify regions of nitrate production and removal in the hyporheiczone under real flow conditions. Therefore, these techniques may helpus to answer questions about the relative importance of differenthyporheic units or patches for nitrate removal in permeable river beds.

Despite its impressive capabilities, the whole river 15N techniqueof Mulholland et al. requires considerable amounts of 15N, which


would be very costly in rivers heavily enriched with nitrate and withhigh discharge. For example, the LINX II protocol requires enrichingthe δ15N of the stream NO3

−pool to at least 20,000‰. Many riversand streams across the developed world have a long-term meanNO3

−concentration of >5 mg L−1 (~360 μM) which, at modest flowsof only 0.1 m3 s−1, would require a staggering £1.3 M worth ofK15NO3

−! In addition, the whole river 15N technique may actuallyonly capture surface ‘denitrification’ if the spike of 15NO3

−is not fullymixed with 14NO3

−throughout the river bed, despite its potential tointegrate fully hyporheic activity and, hence, whole river denitrifica-tion (Bohlke et al., 2004; Mulholland et al., 2004). In addition, as thetechnique only measures the production of 29N2 it cannot be usedto distinguish between N2 coming from either canonical denitrifica-tion or that potentially from anammox (Risgaard-Petersen et al., 2003).

The whole river 15N technique seems more sensitive and bettersuited to a wider variety of physical and chemical river conditions,compared to the open diel N2 technique. In addition, the rates of de-nitrification measured by the 15N methodology are much closer toprevious areal estimates of denitrification from across a variety ofaquatic ecosystems (Fennel et al., 2009; Trimmer and Engström,2011; Steingruber et al., 2001). In comparison, those reported byLaursen and Seitzinger(2004) are considerably higher, at 310 to15,910 μmol N m−2 h−1, though there are a few examples of similarlyhigh rates (McCutchan et al., 2003; Sjodin et al., 1997).

Recent large scale compilations (n=609, (Fennel et al., 2009)n=62, (Trimmer and Engström, 2011)), primarily of data from estu-arine and coastal sediments, have suggested a fairly consistent ratioof between 0.07 and 0.09 mol of N2–N produced per mol of O2 con-sumed. There are many more published data on whole reach dielchanges in oxygen than for N2 production. In this paper we selecteddata (n=263; see Fig. 1 and references below) that were readilyavailable and which also encompassed a wide range of metabolic ac-tivity (Bott et al., 1985; Dawson et al., 2001; Kowalczewski and Lack,1971; Uehlinger, 2006). Our compilation covers rates of whole reachrespiration from 66 to 14,719 μmol O2 m−2 h−1, with a median valueof 3084 μmol O2 m−2 h−1and largely covers the range of respirationdata used by Young et al (2008) to describe rivers as of being in eithera healthy, satisfactory or poor condition. If we assume an averageratio of 0.08 mol of N2–N produced per 1 mol of oxygen consumed,we would predict rates of N2 production across these rivers of be-tween 5 and 1178 μmol N m−2 h−1, with a median value of247 μmol N m−2 h−1. This median value is practically identical tothat (241 μmol N m−2 h−1) calculated as part of a synthesis of deni-trification across aquatic environments (Pina-Ochoa and Alvarez-Cobelas, 2006), and suggests that whole river metabolism is a goodpredicator of the ‘denitrification’ potential of a river as well as otheraspects of nutrient cycling (Bernot et al., 2010; Meyer et al., 2007).

Further, the mean ratio of 0.08 mol of N2–N produced per 1 mol ofoxygen consumed (or N gas produced per unit respiration) can beused to estimate a global total for denitrification in rivers. For example,Battin et al. (2008) estimated the global rate of respiration in riversand streams to be 0.35 Pg C y−1 or 2.9×1013 mol CO2 y−1 and whichhere we assume is equivalent to the amount of oxygen consumed(RC=1:1 as above). At a ratio of 0.08:1, the global total for river andstream respiration of 2.9×1013 mol CO2 y−1 would be equivalent to2.3×1012 mol N y−1 or 33 Tg N y−1. This 33 Tg N y−1 is practically iden-tical to the value of 35 Tg N y−1, modelled independently using N load-ings and mean water residence times for rivers (Seitzinger et al., 2006).

4. River bed methanogenesis, methane oxidation, chemosynthesisand the third way

4.1. River bed methanogenesis and methane oxidation

It has long been recognised that many rivers are supersaturatedwith the dominant species of carbon gases, namely carbon dioxide

and methane, although less is known about the true provenance ofthese gases or indeed their final fate (Cole et al., 2007; Hope et al.,2001; Jones and Mulholland, 1998b). The well irrigated gravel bedof many rivers has been regarded as too oxic to sustain methanogen-esis but ingress of fine particulate material can block the gravel inter-stices (colmation), restricting the flow of water and delivery ofoxygen and ultimately generating sufficiently anoxic conditions toenable methanogenesis to develop (Brunke and Gonser, 1997; Jonesand Mulholland, 1998b; Schalchli, 1992). Sanders et al. (2007)reported rates of methane emission from an English chalk stream(River Frome), ranging from 0.02 to 45 μmol CH4 m−2 h−1, whichwere comparable to, and in some cases exceeded, those reported forpeat bog hollows (1.6–96 μmol CH4 m−2 h−1) and hummocks(0.12–14 μmol CH4 m−2 h−1) (Nedwell and Watson, 1995) andrepresented one of the first accounts of seasonal methanogenesis ina river bed itself.

An interesting aspect of these observations was that the methanewas predominantly transported to the atmosphere through the stems(via the aerenchyma or gas exchange channels, namely the aerenchy-ma) of the aquatic macrophyte Ranunculus penicillatus (var. calcare-ous: R.W. Butcher) which, in itself, plays an important role inrestricting flow and enhancing the deposition of fine sediments(sands, silt and clay) (Sand-Jensen, 1998). Deposition of fines restrictsthe exchange of oxygen into the bed, most likely aiding the develop-ment of anoxic sediments and allowing methanogenesis in the firstplace (as described above). In addition, the efflux of methane trans-ported via the plant stems was at least two orders of magnitudegreater than that via ebullition (release of bubbles) from the sedi-ment surface, and the potential for methane production over the sed-iment as a whole (~14 cm) was far greater than that emitted to theatmosphere via the plants (~750 μmol CH4 m−2 h−1 compared to amaximum 45 μmol CH4 m−2 h−1). A large proportion (~94%) of themethane potentially produced at depth never appears to escapefrom the sediment, being most likely oxidised to CO2 by the aerobicmethane oxidising bacteria (MOB) in the upper-most oxic layers ofsediment. Methane oxidation appears as important in chalk streamsin mitigating the efflux of methane to the atmosphere as it is inother aquatic ecosystems (King, 1996; Roslev and King, 1996).Hence any increase in river flow, of sufficient magnitude to mobilisethe bed sediments and associated plants could release any methanetrapped at depth, by-passing the potential for oxidation and elevatingthe efflux of methane to the atmosphere from 6% to 100% of that pro-duced at depth.

Riverbeds are often a complex mosaic of substrata, and the highestconcentrations of methane measured were associated with either thesediment deposited in the lee of macrophyte stands or in marginaldepositional areas, whereas the ‘cleaner’ gravels associated withopen water were deplete of methane relative to the overlying watercolumn (Sanders et al., 2007; Trimmer et al., 2009a, 2010). The net ef-fect is that the riverbed consists of mixed patches acting as either netsinks or sources of methane (Fig. 2 and see Malard et al., 2002). How-ever, recent evidence suggests that methane could play a hithertounsuspected role as an energy source at the base of river food webs(Trimmer et al., 2009a, 2010).

4.2. Methane based chemosynthesis and the third way

A cornerstone of lotic ecology is that secondary production is sup-ported either by ‘autochthonous’ photosynthetic primary productionor by ‘allochthonous’ production imported from the catchment, usu-ally as leaf litter, fine particulate organic matter or dissolved organiccarbon but all of it ultimately based on terrestrial primary production(e.g. Allan, 1995; Bunn et al., 2003; Hynes, 1970). Chemosynthesis, inwhich the production of reduced carbon compounds is driven by theoxidation of inorganic substrates rather than by light energy, repre-sents an alternative to photosynthesis. Well characterised forms of

O2 concentration ( mol L-1)

0 50 100 150 200 250 300 350 400

CH

4 co

ncen

trat

ion

(nm

ol L

-1)

101

102

103

104

105

Surface water

Fig. 2. Pore-water concentrations of methane as a function of dissolved oxygen in dif-ferent substrate types in the River Lambourn: triangles — silt and sand (filled, b30 cm;and open, >30 cm depth); filled circles — coarse gravels; open circles — fine gravelsand sand; filled squares — surface water. The depletion in methane in the open gravels(combined) is significant relative to that in the surface water (t-test, t=11.05,df=165, pb0.001) whereas shallow silts in the lee of macrophytes are hot-spots ofmethanogenesis. The graphed methane data are reproduced from Trimmer et al.(2010), Copyright by the American Society of Limniology and Oceanography, Inc. Theimage is courtesy of James Pretty.

Trimmer et al.

-42 -40 -38 -36 -34 -32 -30

Epiphyton Ber

SPM

Agapetus

Baetis

Gammarus

Epilithon

Epiphyton Ran

Berula

Ranunculus

Potamopyrgus

Ephemerella

Simulium

Silo

FPOM

13C (‰ vs. VPDB)

Fig. 3. Stable isotope values (δ13C‰ vs. VPDB; mean±95% CI) for macroinvertebrates(filled circles, n ranges from 36 to 141) and putative food sources (open squares,n ranges from 10 to 34) from the River Lambourn. FPOM: fine particulate organicmatter; SPM: suspended particulate matter; Ber: Berula; Ran: Ranunculus. Reproducedfrom Trimmer et al. (2009a), copyright by the American Society of Limniology andOceanography, Inc.


the reduced inorganic substrates required are ammonium, nitrite,various sulphide and sulphur species, hydrogen, ferrous iron andmethane, each being exploited by a characteristic family of aerobicchemolithotrophic α-proteobacteria. More recently, however, newforms of chemolithotrophy, in which ATP production is coupled tothe anaerobic oxidation of ammonium to N2 gas at the expense of ni-trite, have been discovered (see above), and there is emerging inter-est in chemolithotrophic nitrification mediated by members ofCrenarchaeota within the domain Archaea (Dalsgaard et al., 2005;Labrenz et al., 2010; Lam et al., 2007).

The notion of chemosynthetic production is nothing new, datingback to the pioneering Russian microbiologist Sergio Winogradsky,who even in the late 19th century was able to apportion the yield ofnitrifying bacterial biomass to the amount of ammonium substrateoxidised. That said, there was a great deal of popular media interestin 1977 when images of bizarre 2 m high tubeworms and giantclams came up from the depths of the Pacific to reveal significant pro-duction, indeed whole chemosynthetic ecosystems, coupled to theoxidation of sulphur belching forth from ‘black smokers’.

Methane-derived carbon has a typically 13C-deplete isotope signa-ture, and is thus isotopically distinct from most other basal resourcesin aquatic ecosystems, allowing its contribution to food webs to betraced with confidence. For example, from deep sea vents and meth-ane seeps, δ13C values for gastropod snails and sea stars indicate al-most 100% reliance upon chemosynthetic production coupled tomethane oxidation (MacAvoy et al., 2002). Methane-derived carbonhas since been found to be important to the food web of lakes, espe-cially those which stratify during the summer and support methano-genesis in the sediments. Bioturbating midge larvae (Chirnomidae)can then exploit methane oxidising bacteria that assimilate thismethane resulting in larval biomass comprising up to 60% methane-derived carbon (Grey et al., 2004; Jones et al., 2008; Jones and Grey,2011). However, few would suspect that chemosynthesis would besignificant in the classic chalk rivers of southern England, though re-cent research suggests that this is the case (Pretty et al., 2006;Trimmer et al., 2009a).

The δ13C values of common aquatic invertebrates and their puta-tive food sources in an English chalk stream (the River Lambourn)suggested that free-living invertebrates, such as Gammarus sp. andSimulium sp., reflected that of the dominant photosynthetic primaryproduction, whereas the cased larvae of two caddisflies, Agapetus sp.and Silo sp., were consistently too isotopically light throughout theyear to be feeding on this resource (Fig. 3). Our calculations suggestedthat a mean annual average of about 11% of the carbon in the caddislarvae was derived from methane, with the remainder coming frommore conventional epilithon. The contribution of methane-derivedcarbon could rise to 20–30%, depending on the degree of fractionationby MOB, during autumn, when the difference between the δ13C ofAgapetus sp. and epilithon was greatest (Trimmer et al., 2009a).

Besides a depleted δ13C signature in the caddisfly larvae, indicat-ing the incorporation of methane derived carbon, we also found ac-tive methane oxidation on the caddis-cases, and simultaneousmethane oxidation and photosynthesis by gravels from the wider riv-erbed itself (Trimmer et al., 2009a, 2010). Hence, chemosynthesis,coupled to the oxidation of methane, extends far beyond the immedi-ate vicinity of the larval case, and the true potential and fate of che-mosynthesis in the gravel biofilm more generally is unknown.

Does chemosynthesis, and in particular methanotrophy, representa generally important ‘third way’ of energy production to supportriver food webs and communities? This would be in addition to thetwo conventional sources of photosynthetically produced carbon(authochonous and allochthonous). If so, this would signal a genuineparadigm shift in the way we view lotic ecosystems and their role inthe wider terrestrial ecosystem. If methane is produced in situ, basedon reduced inorganic substrates and organic carbon already in theriver, then this could be categorised as authochthonous, but chemo-synthetic, primary production. If the methane is imported from thecatchment, from adjacent wet soils/sediments or from the groundwa-ter, then this could be viewed as a form of allochthonous chemosyn-thetic energy. In the case that the methane is from the groundwater,and of ancient origin, then we would have a genuine subsidy of con-temporary river food webs (and thus of the many terrestrial animalsthat rely on aquatic production) from across the ages (Darling andGooddy, 2006).

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CH4 DIC

Epilithon(+ CH4 to CO2)

Epilithon (- CH4 to CO2)

Agapetus fuscipes

MOB

Mixing model x = [c - b]/[a - b]

x = 20% CH4-derived

x = 14% CH4-derivedc ba

δ13C (‰)

Fig. 4. Schematic of simple two-source mixing model outputs using δ13C values ofmethane oxidising bacteria (MOB; chemosynthetic production) and epiphyton (photo-synthetic production) as end-members and assuming Agapetus sp. is a mixture of thetwo. However, epiphyton may also include a proportion of methane-derived carbonfrom recycled CO2 and so the model output (x) is calculated under two scenarios:accounting for a methane-derived component; and without.


A reliance on chemosynthetic energy by the cased grazing caddis,Agapetus sp. and Silo sp. might seem less dramatic if they were minorplayers in the benthos and overall production of chalk rivers. Agapetussp. reaches densities of 1650 m−2 in the gravels of the RiverLambourn, however, where they are the dominant benthic taxon(Wright, 1992), and in other rivers their abundance can be evenhigher (e.g. the German Breitenbach: 18,770±5760 m−2; (Becker,1990)). Therefore, Agapetus sp. could make a significant contributionto the flow of energy through the food web, a substantial part ofwhich is derived from the oxidation of methane. Cased caddis likeAgapetus sp. can be viewed as ‘armoured grazers’, apparently relative-ly well protected against many predators. Their most likely enemiesare perhaps planarian flatworms or leeches, that might enter thecase directly, small benthic fish such as the bullhead (Cottus gobio)that can potentially ingest the whole organism, including the case,and where it is present the dipper (Cinclus cinclus), which pickscased caddis from small stony rivers.

What is curious about this potentially important role for chemo-synthetic production is that the energy requirements of any herbi-vore/detritivore in chalk rivers could easily be matched bytraditional net photosynthetic primary production (NPPP). The iconiclowland chalk rivers of southern England are famously productive, afunction of the stability of the flow within perennial reaches, in con-junction with clear water of a high mineral content promoting highproduction of microphytobenthos and macrophytes. Indeed, exten-sive growths of the dominant macrophytes Ranunculus spp. andBerula spp. can cover up to 80% of the main river channel in latesummer (Cotton et al., 2006). In the River Lambourn, for instance,secondary production by invertebrates has been estimated at23 g dry mass m−2 y−1 in the open water gravel beds, and evenmore within the macrophyte stands (65 g dry mass m−2 y−1; (Todand Schmid-Araya, 2009)). If we assume that 50% of this dry mass iscarbon, then these values equate to ~11 and 32 g C m−2 y−1 for theopen gravel and macrophyte stands, respectively which, with an en-ergy transfer efficiency of 10% between trophic levels, would requireNPPP of 110 and 320 g C m−2 y−1 in each habitat, respectively. Mea-surements of oxygen evolution and consumption by biofilm on thegravel from the Lambourn, and other rivers, give typical values forNPPP in the order of ~5 g C m−2 d−1 or 1825 g C m−2 y−1, which isclearly far in excess of that required to sustain the estimates of sec-ondary production (Trimmer et al., 2010; Uehlinger and Brock,2005). The same is true for the macrophyte stands, where the produc-tion of Ranunculus sp. has been measured at 410 g C m−2 y−1, withadditional contributions from the epiphytes (Shamsudin and Sleigh,1995; Trimmer et al., 2009b). However, photosynthetic productionis seasonally limited, whereas chemosynthetic production is availableall year round (Trimmer et al., 2010) and, because groundwater-fedrivers are comparatively thermally stable, such an alternative energypathway could continually supplement secondary production. Thus,the relative importance of methane-derived carbon may increase atparticular times of year; for example to fish like the bullhead that pro-vision their eggs over winter. Further, if present in the ground water,methane and MOB potentially provide an energy source to inverte-brates in the dark, labyrinthine interstices beneath the bed, wheremany of them are found in their young stages.

There is clearly a plentiful supply of ‘conventional’ organic carbonat the base of the food web in chalk streams, and such a reliance onchemosynthetic production amongst the cased caddis flies may be aconsequence of their particular behaviour rather than a shortage ofresources per se. Although published measurements of methane oxi-dation by caddis cases, along with very low (depleted) δ13C values,are so far restricted to Agapetus sp. and Silo sp. (Trimmer et al.,2009a), we have made similar measurements for another cased cad-dis (Drusus sp.). In addition, such consistently low δ13C for larval cad-dis flies (Glossosoma spp., like Agapetus an armoured grazer in thefamily Glossosomatidae) relative to bulk epilithon, by up to 10‰

have also been reported elsewhere (McNeely et al., 2006). Further,(Cox and Wagner, 1989) and (Cavanaugh et al., 2004) both found ev-idence for glossosomatid caddis larvae grazing on material on theirown cases or those of conspecifics. Thus, material grazed from thecase might be seen as ‘emergency rations’, where food is short or lar-vae are crowded, as they frequently are. This material would includeMOB, thus less available to caseless species.

Hitherto, the importance of methane-derived carbon has beenpresented simply as a direct subsidy, via methanotroph biomass, toa few (case-bearing) primary consumers, and estimated using two-source isotope mixing models (see Trimmer et al., 2009a). Suchmodels have inherent assumptions and most discussion as to theirvalidity in estimating contributions to consumer biomass stem fromthe trophic fractionation factor applied (France and Peters, 1997;Grey et al., 2001). However, the simplest premise in this instance re-mains that the consumer represents a proportional mixture of carbonfrom two end members; methanotroph biomass and epilithic bio-mass (see Fig. 4).

In comparison to δ13C data from aquatic plants sampled from awide range of aquatic ecosystems (typically>−30‰; Finlay andKendall, 2008), photosynthetic producers that were analysed fromchalk rivers were substantially more depleted (means of −38‰ to−34‰; Trimmer et al., 2009a, unpublished data), so perhapsmethane-derived carbon supports a much greater part of the foodweb indirectly. The simplest explanation might be that the samplescollected to represent epilithon contain a proportion of methanogenicand methanotrophic consortia, as in the biofilms on smaller calibregravels (Trimmer et al., 2009a). However, this would not accountfor the apparent 13C-depletion of the higher plants, unless they alsocontained methane oxidising bacteria. Potential sites are the rhizo-sphere of the Ranunculus sp., which can actively exchange oxygenand methane (Sanders et al., 2007) and which are known sites ofmethane oxidation (Calhoun and King, 1998; van der Nat andMiddelburg, 1998), or within the aerenchyma, which in a similarmanner has been shown to provide Sphagnum sp. with 10–15% ofit's carbon (Raghoebarsing et al., 2005).

The concentration of dissolved inorganic carbon (DIC) in the RiverLambourn was consistently high, on average 5.1 mmol DIC L−1, andwith a similarly constant δ13C of −15.1‰ (Trimmer et al., 2009a).On this basis, macrophyte and epiphyte biomass should exhibit δ13Cvalues between −35 and −33‰, because plants typically fractionate

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2o consumer

Carbon source

Basal resource

1o consumer

MOB

CH4 DICAtmosCO2

Periphyton

Leaf litter

δ13C (‰)

Fig. 5. Simplified schematic representation of alternative energy pathways in chalkstreams (using typical δ13C values) via cased caddis such as Agapetus sp. or shredderssuch as Gammarus sp. to secondary consumers such as Cottus gobio. Autochthonousperiphyton via the dissolved inorganic carbon (DIC); allochthonous material such asleaf litter stemming from an atmospheric CO2 source; or via methane oxidising bacteria(MOB).


their source of inorganic carbon by a further 18–20‰ during photo-synthesis (e.g. Marshall et al., 2008 and references therein), yet weconsistently measured values substantially lower than this. Althoughthe bulk of ΣDIC in chalk rivers comprises HCO3

−, groundwater dom-inated flow is constantly supersaturated with methane and it hasbeen shown that methane oxidation is a continuous process in thegravels (Trimmer et al., 2009a), so there is potential for the resultantCO2 produced (which will also be isotopically light) to be incorporat-ed into macrophytes, their epiphytes or epilithic algae/cyanobacteria,as a preferred source of DIC. Thus, the previous estimates of the directmethane-derived carbon subsidy (up to 30% in Agapetus; (Trimmer etal., 2009a)), which were modelled using epilithic δ13C values, may bean underestimation when one considers that the epilithic photosyn-thetic biomass may already incorporate methane-derived carbon in-directly via the DIC (see Fig. 5). This could explain why the entirefood webs of groundwater-fed rivers are typically more 13C-depletedthan rivers reliant upon allochthonous leaf litter or photosynthesisbased upon (non-recycled) inorganic carbon.

Hitherto, our research into methane as a novel chemosyntheticcarbon subsidy to chalk streams has focussed on the well charac-terised pathway of aerobic methane oxidation (Murrell andRadajewski, 2000; Trimmer et al., 2009a, 2010). Alternatively, meth-ane oxidation can also proceed anaerobically at the expense of sul-phate, a process that has been fairly well characterised in a varietyof marine systems (Knittel and Boetius, 2009; Treude and Ziebis,2010). There is also evidence for anaerobic methane oxidation insulphate-rich freshwater lakes, but the role of alternative electron ac-ceptors such as Fe-and Mn-oxides cannot be ruled out (Schubert etal., 2011). What could offer the greatest potential for anaerobic meth-ane oxidation in these chalk streams could be an entirely novel me-tabolism, only recently enriched from freshwater sediments(Candidatus Methylomirabilis oxyfera), where methane is oxidisedat the expense of nitrite (Ettwig et al., 2010). As well as being super-saturated with methane the gravel beds of the Lambourn are alsosites of nitrate reduction, acting as sources of both nitrite and nitrousoxide to the overlying water, so the substrates and potential are cer-tainly there (Pretty et al., 2006).

5. Conclusions

Rivers are not simple inert conduits merely piping carbon andnitrogen from catchment to coast. Some 0.35 Pg C y−1 is estimated

to be respired to CO2 across the global network of streams and rivers,even though some of that carbon is conveyed from the land and maybe very old. A challenge for future research will be to quantify the ex-tent to which lateral exchange influences the flux of carbon and nitro-gen through rivers, in particular what fraction of their respectivemetabolisms and/or storage occurs either on the land or in the riverand, in turn, how these dynamics change with catchment geologyand in response to future hydrological/climate scenarios.

In turn, the availability of that carbon correlates negatively withthe accumulation of nitrate along the catchment to coast continuum.This is most likely because, just as some of that carbon is respired viaaerobic respiration, it is also available to support denitrification andhence the removal of nitrate. Indeed, in a new synthesis we predictthe global total amount of denitrification in rivers as a fraction ofthe carbon they respire, deriving a value (33 Tg N y−1) which is prac-tically the same as that modelled previously using an entirely inde-pendent approach (35 Tg N y−1). Thus, the turnover and availabilityof carbon in rivers do indeed appear to be a good indicator of their de-nitrification capacity and hence their potential to remove nitrate. Inaddition, residence time, within a reach or at depth in the hyporheiczone, has a strong influence on habitats acting as either sinks orsources of N; the dynamics of which are beginning to be probedwith both fluorescent ‘smart tracers’ and the in situ application of 15N.

A large amount of data are available for whole river carbon metab-olism, data which can provide managers with functional metric ‘tools’to assess river quality in light of the impact of changing land use, forexample. Such ‘tools’ are firmly grounded in the simple concepts ofGPP and ER but these may need some reconsideration if a system‘leaks’ carbon, through emissions of volatiles or methane, for exam-ple. Drawing on such a wealth of data at the same scale as that forGPP and ER cannot be done for denitrification. Even though the recentwhole reach application of 15N advances our understanding ofdenitrification in streams, its cost is likely to prohibit its wide scaleuse in research, let alone its routine application as a managementtool. Better measures of reach scale metabolism may be availablesoon through the application of eddy correlation for example, andthis, coupled to a robust relationship between respiration and theproduction of N2 gas, may offer one way forward for up-scaling deni-trification. In addition, too much effort may have been focused ondenitrification in the past and the comparatively poorly studied alter-native pathways of nitrate reduction (assimilation, DNRA, anammox)may be dismissed before being assessed fully.

Our meta-analysis indicates a strong response by respiration inrivers to temperature and, as a consequence, future warming maysee rivers generate more CO2 per unit of carbon conveyed. In addition,ingress of particulate material via erosion can lead to significantmethanogenesis, which is predicted to respond even more stronglyto raising temperatures than aerobic respiration. Whether such atemperature response will be mirrored by key metabolisms in the Ncycle (denitrification and N2-fixation) is currently unknown but atleast the autotrophic and heterotrophic branches of the N cycle mayhave similar temperature characteristics to that of GPP and ER, re-spectively. Finally, the discovery of new metabolisms that couplethe anaerobic oxidation of methane and ammonium to nitrite, andthe realisation that a considerable fraction of production appearsdriven by the oxidation of methane in some lowland rivers, ensurethat the biogeochemistry of carbon and nitrogen in rivers willcontinue to stimulate novel research for years to come.

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river bed carbon and nitrogen cycling: state of play and some new directions

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