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Nitrogen transformations in microenvironments of river beds and riparian zones
Ecological Engineering 24 (2005) 447–455
Nitrogen transformations in microenvironments of river beds and riparian zones Niels Peter Revsbech ∗ , Jacob Peter Jacobsen, Lars Peter Nielsen Department of Microbial Ecology, University of Aarhus, Bd. 540, DK-8000 Aarhus C, Denmark Accepted 10 January 2005
Abstract The soil of flooded riparian zones, the rhizosphere of riparian plants, biofilms at solid surfaces in the river, and the surface layer of sediments all constitute important environments for the oxidative or reductive transformations of inorganic nitrogen compounds. The exact microzonation and coupling of the processes have recently been studied intensively with 15 N enrichment methods and microsensors for NH4 + , NO2 − , NO3 − , and N2 O. Microsensor analyses of gradients in sediments and biofilms have shown that nitrate production takes place in an aerobic surface zone that has a maximum thickness of a few millimeters in most shallow-water sediments and may be as thin as 100 m in biofilms from very eutrophic environments. In the anoxic zone, denitrification is also concentrated in a zone of maximum a few millimeters, and typically half of the nitrate produced by nitrification is denitrified while the other half escapes to the water. The supply of nitrate from above is primarily controlled by the oxic layer acting as a diffusion barrier, and therefore denitrification is generally a linear function of the nitrate concentration in the water. The overlying water is thus a much more important source of nitrate for denitrification if the concentration is high. The rate and location of denitrification are also affected by bioturbating animals, benthic microphytes, plants, and bacteria performing dissimilatory nitrate reduction to ammonium (DNRA). © 2005 Elsevier B.V. All rights reserved. Keywords: Denitrification; Nitrification; 15-N isotope; Isotope pairing; Microsensors; Nitrate; Sediment; Wetland
1. Introduction Several methods have been used to study nitrogen transformations in biofilms and sediments. One of the oldest approaches is the acetylene inhibition technique (Yoshinari and Knowles, 1976; Sørensen, 1978) by which acetylene is injected into the substrate (e.g. sed∗
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iment) to inhibit the action of nitrous oxide reductase so that N2 O is the end product of denitrification instead of N2 gas. The spatial distribution of the denitrification process can then be analyzed by sectioning of the sample and subsequent analysis of the N2 O content in the various fractions, although this only results in a very coarse spatial resolution due to the rapid diffusion at a mm–cm scale of N2 O during the assay. The acetylene inhibition technique has the advantage that it is relatively simple, requiring little equipment
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in addition to a sensitive N2 O analyzer, as for example a gas chromatograph equipped with an ECD detector. Acetylene, however, is not a perfect inhibitor. The acetylene inhibition may not be total, allowing some N2 formation (Dalsgaard and Bak, 1992), and more seriously, acetylene also inhibits nitrification, and thereby reduces or even eliminates the substrate for denitrification in some systems (Seitzinger et al., 1993). Despite the high background concentrations of N2 , it is also possible in tightly controlled setups to measure total denitrification directly as N2 evolution in sediment without any addition of inhibitor (Cornwell et al., 1999). In the following, we describe the study of nitrogen transformations in the microenvironments associated with the oxic–anoxic interface of sediments, biofilms, and flooded soil. Some of the studies were done in marine or artificial systems, but we believe that the results are valid for freshwater habitats also. The techniques chosen for these studies are profile recordings by use of microscale N-species sensors and 15 N isotope analysis by the so-called isotope pairing technique.
2. Materials and methods 2.1. Isotope pairing technique The 15 N isotope has been used to analyze the nitrogen cycle in aquatic environments for many years (e.g. Knowles and Blackburn, 1993), but a break-through in terms of quantification of nitrification and denitrification came with the invention of the isotope pairing technique (Nielsen, 1992) where the distribution of 15 N14 N and 15 N15 N formed by denitrification is analyzed after addition of 15 NO3 − . By the isotope pairing technique it is possible to quantify nitrification, denitrification based on nitrate from the overlying water, and coupled nitrification/denitrification based on nitrate (or nitrite) formed within the substrate. The isotope pairing technique results in reliable data for all stratified environments where labeled nitrate can be added to an oxic overlying water, and where the oxygen in the overlying water or oxygen formed by photosynthesis at the surface of the substrate is the only source of oxygen for nitrification. An example of a system where the isotope pairing method does not work is the oxic rhizosphere around for example rice roots (Liesack
Fig. 1. Hypothesized profiles of 14 NO3 − and 15 NO3 − in a sediment during incubation with the isotope pairing technique for measuring denitrification. The added 15 NO3 − diffuses to the anoxic zone and is denitrified together with the native 14 NO3 − coming from nitrification in the oxic zone and from the water column. The dinitrogen species 14 N14 N, 14 N15 N and 15 N15 N are formed according to the mixing ratio of the isotopes. Accumulation of the two 15 N-labeled species are measured and from that rates of denitrification are calculated. The contribution from nitrification is specified from the difference in isotope mixing ratios in the overlying water and in the denitrification zone.
et al., 2000), as there is no diffusional mixing of 15 NO − in the overlying water with the 14 NO − 3 3 formed in the rhizosphere. The principle of the isotope pairing technique is illustrated in Fig. 1, where profiles in a sediment of naturally occurring 14 N isotope and added 15 N are drawn. The 15 N isotope is added as 15 NO3 − , and both this 15 NO3 − and the 14 NO3 − produced in the oxic surface sediment diffuse down into the anaerobic zone where it is reduced to a mixture of 14 N14 N, 14 N15 N, and 15 N15 N. In this example, the concentration gradients and hence the diffusional fluxes for 14 NO3 − and 15 NO − are identical at the oxic–anoxic interface, and 3 simple recombination then gives us the result that this particular situation will result in a formation of two 14 N15 N for every 14 N14 N and 15 N15 N. The amount of formed 14 N14 N is very difficult to measure directly due to the high atmospheric background, but based on our knowledge about the random recombination we can calculate this amount when we know the amounts of the other isotopic N2 species. Knowing the amount and relative distributions of 15 N15 N and 14 N15 N formed during an incubation we can thus calculate the rate of denitrification, and if the specific labeling of the NO3 − pool in the overlying water is known we can specify the contribution from coupled nitrification–denitrification within the sediment. By
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analysis of the isotopic dilution of the NO3 − pool during the incubation it is also possible to calculate the total rate of nitrification. Rates of DNRA may be obtained from analysis of the 15 NH4 + formed during the incubation. 2.2. N2 O microsensors The accumulation of N2 O after inhibition of nitrous oxide reductase with acetylene may be a convenient measure of denitrification. It is, however, only recommendable to use N2 O accumulation as a measure when NO3 − from the water overlying the sediment or biofilm is the dominating source of substrate for denitrification, as acetylene inhibits NO3 − formation by nitrification. A good spatial resolution of the N2 O accumulation and thus the denitrifying activity can be obtained by use of N2 O microsensors. Two types are described, one of which measures O2 and N2 O simultaneously in the same spot (Revsbech et al., 1988), and one with an ascorbate oxygen trap that only quantifies N2 O (Andersen et al., 2001). The last type is identical in design to the NO3 − biosensor described later (Fig. 3), except that the mass of immobilized bacteria is substituted with an alkaline ascorbate solution. The two types of N2 O sensors both have their advantages and disadvantages. The advantage by the combined O2 and N2 O sensor is that the location of the denitrifying zone in relation to the oxic–anoxic boundary can be determined at high accuracy. The sensor is, however, difficult to make and is not commercially available. A major problem by this design is also that most sensors exhibit interference from acetylene and that they only work for a few days after construction. The N2 O sensor with ascorbate oxygen trap is much simpler to make, and it is commercially available (www.unisense.com). It has the advantages that the interference from acetylene is negligible and that the lifetime is several months. The relatively slow response of up to 1 min for 90% is, however, a disadvantage.
in Fig. 2. The advantage of the LIX microsensors is that at least the NH4 + and the NO3 − sensors can be made with extremely small tips (<1 m). LIX microsensors are also relatively easy to make, and several LIX compounds are commercially available (e.g. Fluka). The disadvantages are that LIX microsensors often have lifetimes of only a few days, and that they suffer from interference from other common ions. Due to interferences none of the mentioned LIX sensors can be applied in brackish or marine waters. The signal stability is also often a problem, but a coating of the sensor tips with a protein layer may improve the stability (de Beer, 2000). The LIX-type microsensors are presently our only option for getting high-resolution information about the micro-distribution of NH4 + . A review of LIX sensors was presented by de Beer (2000).
2.3. LIX-type microsensors
2.4. Biosensors for NO2 − and NOx −
Microsensors based on various types of Liquid Ion eXchangers (LIX) may be used to analyze for NH4 + , NO2 − , and NO3 − . The general design of a LIX type microsensor with or without coaxial shielding is shown
The availability of a very stable and sensitive N2 O sensor has made it possible to make microscale (tips down to 15–20 m) and macroscale biosensors for NO2 − and NOx − (i.e., the sum of NO2 − and NO3 − ).
Fig. 2. Liquid Ion eXchanger (LIX) type microsensor. Based on different LIX mixtures such sensors may sense the nitrogen species NH4 + , NO2 − , and NO3 − and many other ionic species (from Jensen et al., 1993).
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3. Results and discussion 3.1. Denitrification in sediments as studied by use of N2 O microsensors
Fig. 3. Biosensor for nitrate + nitrite (NOx − ). The diffusion flux of N2 O into the built-in electrochemical N2 O sensor is proportional to the slope (illustrated with straight line) of the N2 O profile, which is again proportional to the NOx − concentration in the medium. Oxygen is preferred as an electron acceptor by the bacteria and is consumed right behind the tip membrane. The tip diameter may vary from 15 to 5000 m depending on the purpose.
The biosensor principle is illustrated in Fig. 3. The biosensors are based on the principle that NO2 − or NO3 − pass through the ion-permeable membrane at the tip and into a mass of immobilized denitrifying bacteria. Here the NO2 − or NOx − are reduced to N2 O by the bacteria that are supplied with electron donors by diffusion from the bulk medium reservoir inside the sensor. The N2 O formed is subsequently quantified by the built-in electrochemical sensor. Whether the sensor functions as a NO2 − sensor or as a NOx − sensor depends on the type of bacterium used inside the sensor. When a nitrous reductase deficient strain of Agrobacterium radiobacter is used the sensor will sense NOx − (Larsen et al., 1997), and when Stenotrophomonas nitritireducens is used the sensor will sense NO2 − (Nielsen et al., 2002). Sensors made this way may detect sub-micromolar concentrations, and the sensitivity may be further improved by applying a positive charge to the sensor tip, whereby the negatively charged NO2 − and NO3 − ions are transported across the membrane by electrophoresis. This electrophoretic signal control (ESC) principle (Kjær et al., 1999) only works ideally, however, when the membrane is relatively impermeable. By highly ion-permeable membranes such as commercial dialysis membranes, the calibration curve becomes non-linear by use of ESC.
The accumulation of N2 O in a river sediment as a function of time after addition of acetylene is shown in Fig. 4A. The NO3 − in the overlying water was in this case as high as 1250 M, and nitrification within the sediment could only play a minor role as a source of NO3 − for denitrification. The distribution of denitrifying activity as obtained from a diffusion-reaction simulation model is shown in Fig. 4B. The points in Fig. 4A are the measured data and the curves in Fig. 4A are the N2 O profiles predicted by the profile of denitrifying activity shown in Fig. 4B. In this particular case, a non-steady state diffusion-reaction model had to be applied, but in most cases reaction rates can be calculated from microsensor data by a simpler steady-state model (Berg et al., 1998). The data presented in Fig. 4 suggest that the vast majority of the denitrifying activity occurs in the anoxic sediment. A low activity seems, however, to be present in the layers with <20 M O2 concentration, and such low denitrification activities at low oxygen have also been found in other investigations (Dalsgaard and
Fig. 4. The N2 O profiles in a river sediment at 9, 28, and 55 min after addition of 10 kPa of acetylene are shown in (A). The curves drawn at the same figure are the concentration profiles that would be result from the denitrification profile shown in (4). The denitrification profile was fitted to obtain the best possible simulation of the data. Also shown in (B) are a measured oxygen profile and a modeled NO3 − profile suggesting 1170 M NO3 − in the water (from Christensen et al., 1989).
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Revsbech, 1992; Blackburn et al., 1994). It is noteworthy that the zone of denitrification was only 4 mm even with this exceptionally high nitrate concentration in the water. With a more moderate nitrate concentration (115 M), the denitrification zone was only 0.5 mm in this sediment (Christensen et al., 1989). Data such as those shown in Fig. 4 have been recorded in many sediments and biofilms, but they often do not represent a satisfactory picture of the denitrifying activity in the sediment. The reason is that most sediments are not adequately represented by a onedimensional representation. One of the major reasons for that is the presence of burrowing infauna. The fauna creates a three-dimensional network of irrigated burrows that may extend to considerable depth in the sediment. An example of N2 O accumulation in a sediment with abundant infauna (amphipods Corophium sp. and polychaetes Nereis sp.) was presented by Binnerup et al. (1992). Initially, accumulation of N2 O appeared to be similar to the one shown in Fig. 4 with denitrification occurring in a subsurface horizon, but after some time a secondary N2 O producing zone associated with animal burrows appeared in deeper strata. It was not possible to estimate the denitrifying activity in the deeper layers by diffusion-reaction modeling, but comparison of diffusional fluxes of N2 O and bulk accumulation in the system indicated that the majority of denitrifying activity may be associated with animal burrows. Later unpublished experiments with microscale NOx − biosensors have shown that also the majority of nitrifying activity in sediments may be associated with animal burrows. In freshwater sediments, it is typically insect larvae like Chironomids that are important in bioturbation (e.g. Svensson, 1998). The importance of taking sediment or biofilm heterogeneity into account by microsensor analysis cannot be over-emphasized. Microsensors are ideal for a study of mechanisms, but in many cases the heterogeneity may be so large that a quantification of overall activity is difficult to obtain by such local determinations, and other methods (e.g. 15 N-enrichment, N2 fluxes or nitrogen mass balances) should be used to obtain overall rate estimates. 3.2. Nitrification in biofilm studied by LIX sensors An example of profiles measured with LIX sensors in a biofilm aggregate is shown in Fig. 5. The diffu-
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Fig. 5. Profiles of O2 , NH4 + , NO2 − , and NO3 − in a nitrifying aggregate from a bioreactor. The O2 profile was measured with a Clarktype oxygen microsensor while the other profiles were measured with LIX-type electrodes (from Schramm et al., 1998).
sive boundary layer around the aggregate analyzed in Fig. 5 was about 200 m thick, and the major part of the chemical gradients actually were in the diffusive boundary layer. The nitrifying zone was situated near the surface of the aggregate where the concentrations of NH4 + and O2 were highest, and the zone was only slightly more than 100 m thick. There was no accumulation of the intermediate NO2 − in the nitrifying zone, and the net result of the nitrification can thus be seen as a rise in NO3 − . No anoxic denitrification zone was present in this biofilm and generally it seems that biofilms on macrophytes and stones in moderately eutrophied rivers and lakes do not support significant rates of denitrification (Eriksson and Weisner, 1999; Sørensen et al., 1988). 3.3. Nitrification and denitrification in sediment studied by the NOx biosensor By use of the microscale biosensors it is possible to get a very detailed picture of the NOx − or NO2 − transformations in a sediment or biofilm. An example of such analysis is given in Fig. 6 which shows the oxygen and NOx − profiles in a shallow-water pond sediment. The pond is an old water-mill pond at a creek dominated by agricultural run-off, and the NO3 − concentration in the water is high during most of the year.
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Fig. 6. Profiles of O2 (), NO3 − (), rates of NOx − consumption (dark bars), and rates of NOx − production (light bars) rates in a diatomcovered pond sediment affected by agricultural run-off. Profiles were measured in the dark (A) and during illumination at 110 mol photons m−2 s−1 (B).
At the shallow site analyzed the sediment was covered by a film of benthic diatoms. The profiles measured in the dark (Fig. 6A) show that no nitrification occurred during the night, and that NO3 − was consumed, presumably by denitrification, in the anoxic layers below 1-mm depth. During illumination (Fig. 6B) the situation was very different, as oxygenic photosynthesis increased the oxygen penetration to about 4.2 mm. There was a consumption of NOx − in the diatom layer in the upper 1 mm, followed by a production of NOx − (nitrification) in the oxic layers below the diatom layer. The production of NOx − was highest right above the oxic–anoxic interface where ammonium was supplied from below (ammonium profiles were not measured). Anoxic conditions and hence denitrification were now present a depths greater than 4.2 mm. A thorough discussion of a similar set of data as those shown in Fig. 6 was presented by Lorenzen et al. (1998). It has been a general finding by microsensor studies that the major part of the nitrate gradient is above the denitrification zone, indicating that it is the transport of nitrate to the zone rather than the microbiological potential within the zone that limits the
denitrifying activity (Christensen et al., 1989). This explains why denitrification rates in stream and lake sediments are proportional to water column nitrate concentrations far beyond the Km values of the bacteria (e.g. van Kessel, 1977; Andersen, 1977; Christensen et al., 1990; Cooper and Cooke, 1984). 3.4. Nitrogen transformations around roots of riparian plants The rhizosphere of plants growing in water-logged soils such as wet meadows or river banks have often been postulated to be very active in nitrification and denitrification (e.g. Reddy et al., 1989). The oxygen leakage from roots of several plants including reed (Phragmites) and rice (Oryza) have been investigated by use of oxygen microsensors, and often very thin oxic zones are found around the roots (Fig. 7). Sensors for NOx − may in principle be used in a similar manner to study nitrification and NOx − consumption associated with the rhizosphere, but molecular studies on the rice rhizosphere from many different rice paddies and soils showed only low numbers of nitrifiers (M. Nicolaisen
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Fig. 8. Variations in nitrate concentration in a flooded meadow. The two arrows indicate storm events when nitrate-rich stream water flushed into the meadow. Following these events nitrate decreased exponentially with depletion rates of 0.8 day−1 , respectively. Fig. 7. Oxygen and pH profiles around a 0.5-mm thick secondary young root of a rice plant growing in rice paddy soil at 30 ◦ C. Older roots exhibited much less oxygen leakage. The pH profile did not show any major effect on soil pH of the root (from Revsbech et al., 1999).
et al., unpublished results). Parallel application of 15 Nbased techniques also gave evidence of very low rates of nitrification and denitrification in the rice rhizosphere (N. Risgaard-Petersen, unpublished results). Calculations showed that peak NO3 − concentrations would be <1 M, so that profiles could not be determined. It is possible that earlier data on relatively very high transformation rates associated with rice roots were due to exceedingly high application of nitrogen fertilizer. Under realistic growth conditions the plants are N-limited, and the nitrifying bacteria have to compete with the plant about any available ammonium. Wastewater treatment based on coupled nitrification–denitrification in the rhizosphere of reed was a popular concept a few years ago, but also in this context it has turned out that the oxygen leakage from the roots is too low to have any practical implications (Brix, 1997).
tions in a Danish flooded meadow were investigated by mass balances and by the isotope pairing technique. Fig. 8 shows the nitrate concentrations in the water around two storm events where the meadow was flooded by nitrate-rich stream water. After the events the concentration dropped exponentially, thus confirming that sediment nitrate reduction is proportional to nitrate concentration in the water column as indicated by the microscale results. The data also showed that the exchange of water between the stream and this particular flooded meadow was very poor in between the storm events. A more open connection to the stream would obviously enhance the reduction of nitrate in this meadow. Seasonal variations of denitrification were
3.5. Fate of NO3 − in a flooded meadow as determined by the isotope pairing technique During periods with high water discharge, rivers and creeks often flood the surrounding meadows, and the large aquatic environments thereby created are potential sites for denitrification. The nitrogen transforma-
Fig. 9. Dissimilatory nitrate reduction to ammonia (DNRA) as a fraction of total dissimilatory nitrate reduction vs. sediment oxygen uptake in a flooded meadow. The data points represent single box incubations during nine campaigns from April to July.
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Fig. 10. Estimated vs. measured rates of total dissimilatory nitrate reduction in the sediment of a flooded meadow. The rate estimates were based on nitrate concentration, temperature, and an assumed constant nitrate reduction rate fuelled by nitrication. Same dataset as in Fig. 9.
more intensively studied in box samples using the isotope pairing technique. It appeared that rates of DNRA could be high and sometimes even be more important than denitrification in the dissimilatory reduction of nitrate. The percentage of DNRA was found to correlate with oxygen uptake which could be taken as a measure of total metabolism in the sediment or soil (Fig. 9). This correlation is consistent with the notation that all known denitrifying bacteria are facultative aerobes while DNRA bacteria only have other anaerobic processes as alternatives for nitrate reduction (Tiedje et al., 1984). In periods of high sediment metabolism, the anaerobic mineralization processes dominate and therefore the potential for DNRA rises relative to the denitrification potential. A presence of DNRA may blur the general correlation between denitrification and water column nitrate concentration described above. In the present study, it was actually found that total dissimilatory nitrate reduction (denitrification plus DNRA) and not denitrification alone correlated well with the nitrate concentration after some temperature correction and inclusion of a constant low nitrate reduction rate based on nitrification (Fig. 10). The magnitude and regulation of the DNRA process are still poorly studied and hard to predict, and this lack of knowledge strongly limits the development of general models of nitrogen removal by denitrification in riparian systems.
Acknowledgements The guidance and support from Carl C. Hoffmann in the study of flooded meadows is greatly acknowledged. The work with N-biosensors was supported by the Icon project under the EU 5th framework programme, contract EKV1-CT-2000-00054.
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Nitrogen transformations in microenvironments of river beds and riparian zones Niels Peter Revsbech ∗ , Jacob Peter Jacobsen, Lars Peter Nielsen Department of Microbial Ecology, University of Aarhus, Bd. 540, DK-8000 Aarhus C, Denmark Accepted 10 January 2005
Abstract The soil of flooded riparian zones, the rhizosphere of riparian plants, biofilms at solid surfaces in the river, and the surface layer of sediments all constitute important environments for the oxidative or reductive transformations of inorganic nitrogen compounds. The exact microzonation and coupling of the processes have recently been studied intensively with 15 N enrichment methods and microsensors for NH4 + , NO2 − , NO3 − , and N2 O. Microsensor analyses of gradients in sediments and biofilms have shown that nitrate production takes place in an aerobic surface zone that has a maximum thickness of a few millimeters in most shallow-water sediments and may be as thin as 100 m in biofilms from very eutrophic environments. In the anoxic zone, denitrification is also concentrated in a zone of maximum a few millimeters, and typically half of the nitrate produced by nitrification is denitrified while the other half escapes to the water. The supply of nitrate from above is primarily controlled by the oxic layer acting as a diffusion barrier, and therefore denitrification is generally a linear function of the nitrate concentration in the water. The overlying water is thus a much more important source of nitrate for denitrification if the concentration is high. The rate and location of denitrification are also affected by bioturbating animals, benthic microphytes, plants, and bacteria performing dissimilatory nitrate reduction to ammonium (DNRA). © 2005 Elsevier B.V. All rights reserved. Keywords: Denitrification; Nitrification; 15-N isotope; Isotope pairing; Microsensors; Nitrate; Sediment; Wetland
1. Introduction Several methods have been used to study nitrogen transformations in biofilms and sediments. One of the oldest approaches is the acetylene inhibition technique (Yoshinari and Knowles, 1976; Sørensen, 1978) by which acetylene is injected into the substrate (e.g. sed∗
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iment) to inhibit the action of nitrous oxide reductase so that N2 O is the end product of denitrification instead of N2 gas. The spatial distribution of the denitrification process can then be analyzed by sectioning of the sample and subsequent analysis of the N2 O content in the various fractions, although this only results in a very coarse spatial resolution due to the rapid diffusion at a mm–cm scale of N2 O during the assay. The acetylene inhibition technique has the advantage that it is relatively simple, requiring little equipment
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in addition to a sensitive N2 O analyzer, as for example a gas chromatograph equipped with an ECD detector. Acetylene, however, is not a perfect inhibitor. The acetylene inhibition may not be total, allowing some N2 formation (Dalsgaard and Bak, 1992), and more seriously, acetylene also inhibits nitrification, and thereby reduces or even eliminates the substrate for denitrification in some systems (Seitzinger et al., 1993). Despite the high background concentrations of N2 , it is also possible in tightly controlled setups to measure total denitrification directly as N2 evolution in sediment without any addition of inhibitor (Cornwell et al., 1999). In the following, we describe the study of nitrogen transformations in the microenvironments associated with the oxic–anoxic interface of sediments, biofilms, and flooded soil. Some of the studies were done in marine or artificial systems, but we believe that the results are valid for freshwater habitats also. The techniques chosen for these studies are profile recordings by use of microscale N-species sensors and 15 N isotope analysis by the so-called isotope pairing technique.
2. Materials and methods 2.1. Isotope pairing technique The 15 N isotope has been used to analyze the nitrogen cycle in aquatic environments for many years (e.g. Knowles and Blackburn, 1993), but a break-through in terms of quantification of nitrification and denitrification came with the invention of the isotope pairing technique (Nielsen, 1992) where the distribution of 15 N14 N and 15 N15 N formed by denitrification is analyzed after addition of 15 NO3 − . By the isotope pairing technique it is possible to quantify nitrification, denitrification based on nitrate from the overlying water, and coupled nitrification/denitrification based on nitrate (or nitrite) formed within the substrate. The isotope pairing technique results in reliable data for all stratified environments where labeled nitrate can be added to an oxic overlying water, and where the oxygen in the overlying water or oxygen formed by photosynthesis at the surface of the substrate is the only source of oxygen for nitrification. An example of a system where the isotope pairing method does not work is the oxic rhizosphere around for example rice roots (Liesack
Fig. 1. Hypothesized profiles of 14 NO3 − and 15 NO3 − in a sediment during incubation with the isotope pairing technique for measuring denitrification. The added 15 NO3 − diffuses to the anoxic zone and is denitrified together with the native 14 NO3 − coming from nitrification in the oxic zone and from the water column. The dinitrogen species 14 N14 N, 14 N15 N and 15 N15 N are formed according to the mixing ratio of the isotopes. Accumulation of the two 15 N-labeled species are measured and from that rates of denitrification are calculated. The contribution from nitrification is specified from the difference in isotope mixing ratios in the overlying water and in the denitrification zone.
et al., 2000), as there is no diffusional mixing of 15 NO − in the overlying water with the 14 NO − 3 3 formed in the rhizosphere. The principle of the isotope pairing technique is illustrated in Fig. 1, where profiles in a sediment of naturally occurring 14 N isotope and added 15 N are drawn. The 15 N isotope is added as 15 NO3 − , and both this 15 NO3 − and the 14 NO3 − produced in the oxic surface sediment diffuse down into the anaerobic zone where it is reduced to a mixture of 14 N14 N, 14 N15 N, and 15 N15 N. In this example, the concentration gradients and hence the diffusional fluxes for 14 NO3 − and 15 NO − are identical at the oxic–anoxic interface, and 3 simple recombination then gives us the result that this particular situation will result in a formation of two 14 N15 N for every 14 N14 N and 15 N15 N. The amount of formed 14 N14 N is very difficult to measure directly due to the high atmospheric background, but based on our knowledge about the random recombination we can calculate this amount when we know the amounts of the other isotopic N2 species. Knowing the amount and relative distributions of 15 N15 N and 14 N15 N formed during an incubation we can thus calculate the rate of denitrification, and if the specific labeling of the NO3 − pool in the overlying water is known we can specify the contribution from coupled nitrification–denitrification within the sediment. By
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analysis of the isotopic dilution of the NO3 − pool during the incubation it is also possible to calculate the total rate of nitrification. Rates of DNRA may be obtained from analysis of the 15 NH4 + formed during the incubation. 2.2. N2 O microsensors The accumulation of N2 O after inhibition of nitrous oxide reductase with acetylene may be a convenient measure of denitrification. It is, however, only recommendable to use N2 O accumulation as a measure when NO3 − from the water overlying the sediment or biofilm is the dominating source of substrate for denitrification, as acetylene inhibits NO3 − formation by nitrification. A good spatial resolution of the N2 O accumulation and thus the denitrifying activity can be obtained by use of N2 O microsensors. Two types are described, one of which measures O2 and N2 O simultaneously in the same spot (Revsbech et al., 1988), and one with an ascorbate oxygen trap that only quantifies N2 O (Andersen et al., 2001). The last type is identical in design to the NO3 − biosensor described later (Fig. 3), except that the mass of immobilized bacteria is substituted with an alkaline ascorbate solution. The two types of N2 O sensors both have their advantages and disadvantages. The advantage by the combined O2 and N2 O sensor is that the location of the denitrifying zone in relation to the oxic–anoxic boundary can be determined at high accuracy. The sensor is, however, difficult to make and is not commercially available. A major problem by this design is also that most sensors exhibit interference from acetylene and that they only work for a few days after construction. The N2 O sensor with ascorbate oxygen trap is much simpler to make, and it is commercially available (www.unisense.com). It has the advantages that the interference from acetylene is negligible and that the lifetime is several months. The relatively slow response of up to 1 min for 90% is, however, a disadvantage.
in Fig. 2. The advantage of the LIX microsensors is that at least the NH4 + and the NO3 − sensors can be made with extremely small tips (<1 m). LIX microsensors are also relatively easy to make, and several LIX compounds are commercially available (e.g. Fluka). The disadvantages are that LIX microsensors often have lifetimes of only a few days, and that they suffer from interference from other common ions. Due to interferences none of the mentioned LIX sensors can be applied in brackish or marine waters. The signal stability is also often a problem, but a coating of the sensor tips with a protein layer may improve the stability (de Beer, 2000). The LIX-type microsensors are presently our only option for getting high-resolution information about the micro-distribution of NH4 + . A review of LIX sensors was presented by de Beer (2000).
2.3. LIX-type microsensors
2.4. Biosensors for NO2 − and NOx −
Microsensors based on various types of Liquid Ion eXchangers (LIX) may be used to analyze for NH4 + , NO2 − , and NO3 − . The general design of a LIX type microsensor with or without coaxial shielding is shown
The availability of a very stable and sensitive N2 O sensor has made it possible to make microscale (tips down to 15–20 m) and macroscale biosensors for NO2 − and NOx − (i.e., the sum of NO2 − and NO3 − ).
Fig. 2. Liquid Ion eXchanger (LIX) type microsensor. Based on different LIX mixtures such sensors may sense the nitrogen species NH4 + , NO2 − , and NO3 − and many other ionic species (from Jensen et al., 1993).
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3. Results and discussion 3.1. Denitrification in sediments as studied by use of N2 O microsensors
Fig. 3. Biosensor for nitrate + nitrite (NOx − ). The diffusion flux of N2 O into the built-in electrochemical N2 O sensor is proportional to the slope (illustrated with straight line) of the N2 O profile, which is again proportional to the NOx − concentration in the medium. Oxygen is preferred as an electron acceptor by the bacteria and is consumed right behind the tip membrane. The tip diameter may vary from 15 to 5000 m depending on the purpose.
The biosensor principle is illustrated in Fig. 3. The biosensors are based on the principle that NO2 − or NO3 − pass through the ion-permeable membrane at the tip and into a mass of immobilized denitrifying bacteria. Here the NO2 − or NOx − are reduced to N2 O by the bacteria that are supplied with electron donors by diffusion from the bulk medium reservoir inside the sensor. The N2 O formed is subsequently quantified by the built-in electrochemical sensor. Whether the sensor functions as a NO2 − sensor or as a NOx − sensor depends on the type of bacterium used inside the sensor. When a nitrous reductase deficient strain of Agrobacterium radiobacter is used the sensor will sense NOx − (Larsen et al., 1997), and when Stenotrophomonas nitritireducens is used the sensor will sense NO2 − (Nielsen et al., 2002). Sensors made this way may detect sub-micromolar concentrations, and the sensitivity may be further improved by applying a positive charge to the sensor tip, whereby the negatively charged NO2 − and NO3 − ions are transported across the membrane by electrophoresis. This electrophoretic signal control (ESC) principle (Kjær et al., 1999) only works ideally, however, when the membrane is relatively impermeable. By highly ion-permeable membranes such as commercial dialysis membranes, the calibration curve becomes non-linear by use of ESC.
The accumulation of N2 O in a river sediment as a function of time after addition of acetylene is shown in Fig. 4A. The NO3 − in the overlying water was in this case as high as 1250 M, and nitrification within the sediment could only play a minor role as a source of NO3 − for denitrification. The distribution of denitrifying activity as obtained from a diffusion-reaction simulation model is shown in Fig. 4B. The points in Fig. 4A are the measured data and the curves in Fig. 4A are the N2 O profiles predicted by the profile of denitrifying activity shown in Fig. 4B. In this particular case, a non-steady state diffusion-reaction model had to be applied, but in most cases reaction rates can be calculated from microsensor data by a simpler steady-state model (Berg et al., 1998). The data presented in Fig. 4 suggest that the vast majority of the denitrifying activity occurs in the anoxic sediment. A low activity seems, however, to be present in the layers with <20 M O2 concentration, and such low denitrification activities at low oxygen have also been found in other investigations (Dalsgaard and
Fig. 4. The N2 O profiles in a river sediment at 9, 28, and 55 min after addition of 10 kPa of acetylene are shown in (A). The curves drawn at the same figure are the concentration profiles that would be result from the denitrification profile shown in (4). The denitrification profile was fitted to obtain the best possible simulation of the data. Also shown in (B) are a measured oxygen profile and a modeled NO3 − profile suggesting 1170 M NO3 − in the water (from Christensen et al., 1989).
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Revsbech, 1992; Blackburn et al., 1994). It is noteworthy that the zone of denitrification was only 4 mm even with this exceptionally high nitrate concentration in the water. With a more moderate nitrate concentration (115 M), the denitrification zone was only 0.5 mm in this sediment (Christensen et al., 1989). Data such as those shown in Fig. 4 have been recorded in many sediments and biofilms, but they often do not represent a satisfactory picture of the denitrifying activity in the sediment. The reason is that most sediments are not adequately represented by a onedimensional representation. One of the major reasons for that is the presence of burrowing infauna. The fauna creates a three-dimensional network of irrigated burrows that may extend to considerable depth in the sediment. An example of N2 O accumulation in a sediment with abundant infauna (amphipods Corophium sp. and polychaetes Nereis sp.) was presented by Binnerup et al. (1992). Initially, accumulation of N2 O appeared to be similar to the one shown in Fig. 4 with denitrification occurring in a subsurface horizon, but after some time a secondary N2 O producing zone associated with animal burrows appeared in deeper strata. It was not possible to estimate the denitrifying activity in the deeper layers by diffusion-reaction modeling, but comparison of diffusional fluxes of N2 O and bulk accumulation in the system indicated that the majority of denitrifying activity may be associated with animal burrows. Later unpublished experiments with microscale NOx − biosensors have shown that also the majority of nitrifying activity in sediments may be associated with animal burrows. In freshwater sediments, it is typically insect larvae like Chironomids that are important in bioturbation (e.g. Svensson, 1998). The importance of taking sediment or biofilm heterogeneity into account by microsensor analysis cannot be over-emphasized. Microsensors are ideal for a study of mechanisms, but in many cases the heterogeneity may be so large that a quantification of overall activity is difficult to obtain by such local determinations, and other methods (e.g. 15 N-enrichment, N2 fluxes or nitrogen mass balances) should be used to obtain overall rate estimates. 3.2. Nitrification in biofilm studied by LIX sensors An example of profiles measured with LIX sensors in a biofilm aggregate is shown in Fig. 5. The diffu-
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Fig. 5. Profiles of O2 , NH4 + , NO2 − , and NO3 − in a nitrifying aggregate from a bioreactor. The O2 profile was measured with a Clarktype oxygen microsensor while the other profiles were measured with LIX-type electrodes (from Schramm et al., 1998).
sive boundary layer around the aggregate analyzed in Fig. 5 was about 200 m thick, and the major part of the chemical gradients actually were in the diffusive boundary layer. The nitrifying zone was situated near the surface of the aggregate where the concentrations of NH4 + and O2 were highest, and the zone was only slightly more than 100 m thick. There was no accumulation of the intermediate NO2 − in the nitrifying zone, and the net result of the nitrification can thus be seen as a rise in NO3 − . No anoxic denitrification zone was present in this biofilm and generally it seems that biofilms on macrophytes and stones in moderately eutrophied rivers and lakes do not support significant rates of denitrification (Eriksson and Weisner, 1999; Sørensen et al., 1988). 3.3. Nitrification and denitrification in sediment studied by the NOx biosensor By use of the microscale biosensors it is possible to get a very detailed picture of the NOx − or NO2 − transformations in a sediment or biofilm. An example of such analysis is given in Fig. 6 which shows the oxygen and NOx − profiles in a shallow-water pond sediment. The pond is an old water-mill pond at a creek dominated by agricultural run-off, and the NO3 − concentration in the water is high during most of the year.
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Fig. 6. Profiles of O2 (), NO3 − (), rates of NOx − consumption (dark bars), and rates of NOx − production (light bars) rates in a diatomcovered pond sediment affected by agricultural run-off. Profiles were measured in the dark (A) and during illumination at 110 mol photons m−2 s−1 (B).
At the shallow site analyzed the sediment was covered by a film of benthic diatoms. The profiles measured in the dark (Fig. 6A) show that no nitrification occurred during the night, and that NO3 − was consumed, presumably by denitrification, in the anoxic layers below 1-mm depth. During illumination (Fig. 6B) the situation was very different, as oxygenic photosynthesis increased the oxygen penetration to about 4.2 mm. There was a consumption of NOx − in the diatom layer in the upper 1 mm, followed by a production of NOx − (nitrification) in the oxic layers below the diatom layer. The production of NOx − was highest right above the oxic–anoxic interface where ammonium was supplied from below (ammonium profiles were not measured). Anoxic conditions and hence denitrification were now present a depths greater than 4.2 mm. A thorough discussion of a similar set of data as those shown in Fig. 6 was presented by Lorenzen et al. (1998). It has been a general finding by microsensor studies that the major part of the nitrate gradient is above the denitrification zone, indicating that it is the transport of nitrate to the zone rather than the microbiological potential within the zone that limits the
denitrifying activity (Christensen et al., 1989). This explains why denitrification rates in stream and lake sediments are proportional to water column nitrate concentrations far beyond the Km values of the bacteria (e.g. van Kessel, 1977; Andersen, 1977; Christensen et al., 1990; Cooper and Cooke, 1984). 3.4. Nitrogen transformations around roots of riparian plants The rhizosphere of plants growing in water-logged soils such as wet meadows or river banks have often been postulated to be very active in nitrification and denitrification (e.g. Reddy et al., 1989). The oxygen leakage from roots of several plants including reed (Phragmites) and rice (Oryza) have been investigated by use of oxygen microsensors, and often very thin oxic zones are found around the roots (Fig. 7). Sensors for NOx − may in principle be used in a similar manner to study nitrification and NOx − consumption associated with the rhizosphere, but molecular studies on the rice rhizosphere from many different rice paddies and soils showed only low numbers of nitrifiers (M. Nicolaisen
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Fig. 8. Variations in nitrate concentration in a flooded meadow. The two arrows indicate storm events when nitrate-rich stream water flushed into the meadow. Following these events nitrate decreased exponentially with depletion rates of 0.8 day−1 , respectively. Fig. 7. Oxygen and pH profiles around a 0.5-mm thick secondary young root of a rice plant growing in rice paddy soil at 30 ◦ C. Older roots exhibited much less oxygen leakage. The pH profile did not show any major effect on soil pH of the root (from Revsbech et al., 1999).
et al., unpublished results). Parallel application of 15 Nbased techniques also gave evidence of very low rates of nitrification and denitrification in the rice rhizosphere (N. Risgaard-Petersen, unpublished results). Calculations showed that peak NO3 − concentrations would be <1 M, so that profiles could not be determined. It is possible that earlier data on relatively very high transformation rates associated with rice roots were due to exceedingly high application of nitrogen fertilizer. Under realistic growth conditions the plants are N-limited, and the nitrifying bacteria have to compete with the plant about any available ammonium. Wastewater treatment based on coupled nitrification–denitrification in the rhizosphere of reed was a popular concept a few years ago, but also in this context it has turned out that the oxygen leakage from the roots is too low to have any practical implications (Brix, 1997).
tions in a Danish flooded meadow were investigated by mass balances and by the isotope pairing technique. Fig. 8 shows the nitrate concentrations in the water around two storm events where the meadow was flooded by nitrate-rich stream water. After the events the concentration dropped exponentially, thus confirming that sediment nitrate reduction is proportional to nitrate concentration in the water column as indicated by the microscale results. The data also showed that the exchange of water between the stream and this particular flooded meadow was very poor in between the storm events. A more open connection to the stream would obviously enhance the reduction of nitrate in this meadow. Seasonal variations of denitrification were
3.5. Fate of NO3 − in a flooded meadow as determined by the isotope pairing technique During periods with high water discharge, rivers and creeks often flood the surrounding meadows, and the large aquatic environments thereby created are potential sites for denitrification. The nitrogen transforma-
Fig. 9. Dissimilatory nitrate reduction to ammonia (DNRA) as a fraction of total dissimilatory nitrate reduction vs. sediment oxygen uptake in a flooded meadow. The data points represent single box incubations during nine campaigns from April to July.
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Fig. 10. Estimated vs. measured rates of total dissimilatory nitrate reduction in the sediment of a flooded meadow. The rate estimates were based on nitrate concentration, temperature, and an assumed constant nitrate reduction rate fuelled by nitrication. Same dataset as in Fig. 9.
more intensively studied in box samples using the isotope pairing technique. It appeared that rates of DNRA could be high and sometimes even be more important than denitrification in the dissimilatory reduction of nitrate. The percentage of DNRA was found to correlate with oxygen uptake which could be taken as a measure of total metabolism in the sediment or soil (Fig. 9). This correlation is consistent with the notation that all known denitrifying bacteria are facultative aerobes while DNRA bacteria only have other anaerobic processes as alternatives for nitrate reduction (Tiedje et al., 1984). In periods of high sediment metabolism, the anaerobic mineralization processes dominate and therefore the potential for DNRA rises relative to the denitrification potential. A presence of DNRA may blur the general correlation between denitrification and water column nitrate concentration described above. In the present study, it was actually found that total dissimilatory nitrate reduction (denitrification plus DNRA) and not denitrification alone correlated well with the nitrate concentration after some temperature correction and inclusion of a constant low nitrate reduction rate based on nitrification (Fig. 10). The magnitude and regulation of the DNRA process are still poorly studied and hard to predict, and this lack of knowledge strongly limits the development of general models of nitrogen removal by denitrification in riparian systems.
Acknowledgements The guidance and support from Carl C. Hoffmann in the study of flooded meadows is greatly acknowledged. The work with N-biosensors was supported by the Icon project under the EU 5th framework programme, contract EKV1-CT-2000-00054.
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