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Keywords:

  • Decomposition;
  • drought;
  • flood;
  • heat;
  • immobilization;
  • N mineralization;
  • nitrification;
  • soil N cycling

Abstract

  1. Top of page
  2. Abstract
  3. Introduction
  4. Paradigm Shifts in Terrestrial N Cycling
  5. Temperature and Heat Impacts on N Balance of Forests
  6. Drought Impacts on N Balance of Forests
  7. Flooding Impacts on N Balance of Forests
  8. Impacts on N Uptake
  9. Interactive Effects on N Balance of Forests
  10. General Conclusions
  11. Acknowledgement
  12. References

Forest ecosystems with low soil nitrogen (N) availability are characterized by direct competition for this growth-limiting resource between several players, i.e. various components of vegetation, such as old-growth trees, natural regeneration and understorey species, mycorrhizal fungi, free-living fungi and bacteria. With the increase in frequency and intensity of extreme climate events predicted in current climate change scenarios, also competition for N between plants and/or soil microorganisms will be affected. In this review, we summarize the present understanding of ecosystem N cycling in N-limited forests and its interaction with extreme climate events, such as heat, drought and flooding. More specifically, the impacts of environmental stresses on microbial release and consumption of bioavailable N, N uptake and competition between plants, as well as plant and microbial uptake are presented. Furthermore, the consequences of drying–wetting cycles on N cycling are discussed. Additionally, we highlight the current methodological difficulties that limit present understanding of N cycling in forest ecosystems and the need for interdisciplinary studies.


Introduction

  1. Top of page
  2. Abstract
  3. Introduction
  4. Paradigm Shifts in Terrestrial N Cycling
  5. Temperature and Heat Impacts on N Balance of Forests
  6. Drought Impacts on N Balance of Forests
  7. Flooding Impacts on N Balance of Forests
  8. Impacts on N Uptake
  9. Interactive Effects on N Balance of Forests
  10. General Conclusions
  11. Acknowledgement
  12. References

Many forest ecosystems have developed on marginal soils characterized by growth-limiting availability of the macronutrient N (Rennenberg et al. 1998; Lovett et al. 2004; Chapman et al. 2006). N-limited forests are characterized by a largely closed ecosystem internal N cycle with (i) low external N input from, and N losses to, the atmosphere and hydrosphere and (ii) high ecosystem internal cycling of N mostly derived from depolymerization and mineralization of leaf and root litter and decaying microbial biomass (Schimel & Bennett 2004). In these ecosystems, N cycling is dominated by reduced inorganic and organic N compounds such as ammonium, amino acids and peptides (Lipson & Näsholm 2001). This is different in, for example, riparian forest ecosystems that can rely on significant N input in the form of nitrate from flood- and groundwater sources and, therefore, can be dominated by nitrate nutrition (Gambrell et al. 1991; Holtgrieve et al. 2006).

Several players in forest ecosystems compete for the limiting amount of N liberated by depolymerization and mineralization processes. These players include old-growth trees, natural regeneration, other woody understorey plant species, herbaceous understorey plant species, mycorrhizal fungi, as well as free-living fungi and bacteria, both in close vicinity to and distant from the rhizosphere. Direct competition between these players may be prevented by a separation of N acquisition in space and time, i.e. by the occupation of different soil compartments and/or preferential N uptake in different seasons, but also by different N source preferences, e.g. inorganic versus organic N compounds (Schimel & Bennett 2004; Dannenmann et al. 2009). When direct competition takes place in the acquisition of N, it will be determined at the biochemical level by the kinetic properties and the capacity of uptake mechanisms (Fotelli et al. 2002, 2005), but also by regulatory mechanisms (Gessler et al. 2005) of the competitors, and at the ecosystem level by biotic interactions between these competitors. These biotic interactions may include, for example, the exchange of regulatory compounds and the supply of heterotrophic organisms with organic carbon from autotrophic organisms (Dannenmann et al. 2009).

Climate change interacts with this range of complex processes in a way that is currently not understood, both at the individual process and at the ecosystem level. In this context, steadily increasing atmospheric surface temperatures are supposed to be only slowly translated into increasing soil temperatures because of the buffering capacity of the soil. Therefore, extreme events such as heatwaves, drought periods and flooding are considered more important factors in disturbing ecosystem processes and functions (Rennenberg et al. 2006). Such extreme events have already increased in Central Europe in both frequency and intensity, and have already affected forest ecosystems. Current climate change scenarios indicate that the frequency and intensity of these extreme climate events will further increase during the next century (IPCC 2007) and, therefore, will increasingly affect N acquisition by and competition for N between vegetation components and soil microorganisms in forests. In this review, the present understanding of ecosystem N cycling in N-limited forests and its interaction with climate change stresses, in particular with extreme climate events, is summarized. Although N may also be taken up from the atmosphere, in N-limited forests this input is negligible. Therefore, this review focuses on belowground N uptake. An overview of the effects of climate change-related drivers on selected processes (i.e. heat, drought, flooding) observed in different studies is given in Table 1.

Table 1.   Effects of climate change-related drivers on selected processes observed in different studies.
driverprocesseseffectsreferences
Increasing temperature/heatmicrobial release/consumption of Nincreased N mineralization [RIGHTWARDS ARROW] increased inorganic N availability net nitrification decreasedRustad et al. 2001
Shaw & Harte 2001a,b
N uptakeincrease in transporter activity in plants increase in nitrate uptake in relative proportion to ammoniumMalagoli et al. 2004
Marschner et al. 1991
Gessler et al. 1998
BassiriRad 2000
droughtmicrobial release/consumption of Ndecreased soil microbial activityBorken & Matzner 2009
Schimel et al. 2007
increased DON availability due to dieback of microbial biomass [RIGHTWARDS ARROW] increased N available for plant uptakeBorken & Matzner 2009
Dannenmann et al. 2009
inhibition of gross nitrificatione.g. Smith et al. 2003
Gessler et al. 2005
 
N uptakedecreased nitrate uptake [RIGHTWARDS ARROW] due to decreased nitrate uptake capacity accumulation of amino acids in roots [RIGHTWARDS ARROW] inhibiting N uptake reduction of fine root biomass [RIGHTWARDS ARROW] decreased nutrient absorbing surfaceFotelli et al. 2002
Gessler et al. 2004
Imsande & Touraine 1994
Mainiero & Kazda 2006
Cudlin et al. 2007
Konopka et al. 2007
competition plant–plant/microbial N uptakecompensation of reduced N availability by access from deeper soil layersMcKinley et al. 2009 Dannenmann et al. 2009
reduction of microbial competition in favour of plants 
floodingmicrobial release/consumption of Ninhibition of nitrification [RIGHTWARDS ARROW] increased ammonium and decreased nitrate availabilityOhte et al. 1997
Alaoui-Sosséet al. 2005
enhanced organic/inorganic N transformationse.g. Birch 1960
Reddy & Patrick 1975
Koschorrek & Darwich 2003
N uptakestrong repression of nutrient uptakee.g. Larson et al. 1992
Kreuzwieser et al. 2002
Islam & MacDonald 2009

Paradigm Shifts in Terrestrial N Cycling

  1. Top of page
  2. Abstract
  3. Introduction
  4. Paradigm Shifts in Terrestrial N Cycling
  5. Temperature and Heat Impacts on N Balance of Forests
  6. Drought Impacts on N Balance of Forests
  7. Flooding Impacts on N Balance of Forests
  8. Impacts on N Uptake
  9. Interactive Effects on N Balance of Forests
  10. General Conclusions
  11. Acknowledgement
  12. References

The common view of N cycling in forests has undergone a considerable change, especially in the last decade (for detailed reviews see Schimel & Bennett 2004; Chapman et al. 2006; Jackson et al. 2008; Näsholm et al. 2009). Until the 1990s, the perception and understanding of N cycling in forest ecosystems was dominated by the paradigms that (i) N mineralization (ammonification) is the limiting step in N cycling; (ii) plants take up only inorganic N; and (iii) plants poorly compete for N against microbes and, therefore, use only the N that is ‘left over’ by microbes (Schimel & Bennett 2004). This perception led to the definition of net N mineralization as the sum of N that exceeds the microbial demand and thus is available for uptake by plant roots and for N losses from the ecosystem. Consequently, net rate assays like the buried bag technique (Eno 1960) have been widely used to measure plant available N.

The use of stable 15N isotopes either in 15N tracing experiments (labelling of the source pool of the process to be investigated) or 15N pool dilution experiments (labelling of the sink pool) allowed determination of gross or actual rates of N fluxes in soil, and thus a more holistic view on belowground N cycling. Since the late 1980s, researchers have increasingly applied such 15N techniques and have become aware that net rates may be a poor approximation of the real status of N cycling in soils (Davidson et al. 1991; Hart et al. 1994). Abundant published studies utilizing the 15N pool dilution technique for determination of gross rates of ammonification, nitrification and inorganic N immobilization, highlighted the limitations of net rate assays, and provided an advanced insight in actual N turnover. The 15N pool dilution studies mostly revealed significantly higher rates of gross ammonification and nitrification, while little or no ammonium and nitrate accumulated (Davidson et al. 1992; Hart et al. 1994; Neill et al. 1999; Verchot et al. 2001; Ross et al. 2004). This illustrated the complex dynamics of microbial production and consumption of inorganic N that cannot be evaluated by net rate assays. Furthermore, net rate studies revealed N mineralization rates that were well below estimates of plant N uptake (e.g.Chapin et al. 1988), and other studies even showed negative N mineralization rates (net immobilization) during periods of simultaneous plant N uptake (Nadelhoffer et al. 1984; Hill & Shackleton 1989; Polglase et al. 1992). This was thought to be a consequence of the removal of plants during net rate incubations, thereby annulling the competition for N between soil microorganisms and plants and causing a surplus of N that microbes readily immobilized. In addition, isotope studies, where 15N-labelled compounds were supplied as N sources and 15N recovery in the plant material was determined, showed an accumulation of 15N in plants even under N-limited conditions; the observations imply that plants can in fact successfully compete for N with microbes (Kaye & Hart 1997; Hodge et al. 2000). Thus, the paradigm shift from plants being poor competitors to plants effectively competing for N with microorganisms was developed (Fig. 1).

image

Figure 1.  N cycling in forest ecosystems. Dashed lines: plant processes; solid lines: microbial processes; red dashed and solid lines: competitive processes between plants and microorganisms. Blue lines: hydrological transport pathways. SOM: soil organic matter. Adapted from Schimel & Bennett (2004).

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Since the 1990s, many studies have shown that plants are able to take up organic N, e.g. amino acids, and that organic N uptake can be the dominant N source for forest trees (Dannenmann et al. 2009; review by Näsholm et al. 2009 and references herein). Further studies suggested that mycorrhizal fungi could play a role in the acquisition of organic N by mycorrhizal plants (Jackson et al. 2008; Martin et al. 2008; Dannenmann et al. 2009). This led to the new paradigm of plant N uptake comprising not only the acquisition of inorganic N species, but also – at least in N-limited forest ecosystems – the uptake of organic N forms. The observation that plants take up organic N in a wide range of forest ecosystems (review by Näsholm et al. 2009 and references herein), including boreal forests (Näsholm et al. 1998; Nordin et al. 2001), temperate forests (Schmidt & Stewart 1999; Finzi & Berthrong 2005; Dannenmann et al. 2009), Australian wet forests (Warren & Adams 2007; Pfautsch et al. 2009) tundra (Chapin et al. 1993; Kielland 1994; Schimel & Chapin 1996; Lipson & Monson 1998; Raab et al. 1999; Nordin et al. 2004), arid ecosystems (Schmidt & Stewart 1999; Warren 2006) and grasslands (Bardgett et al. 2003; Weigelt et al. 2003, 2005; Harrison et al. 2007), clearly shows that plants and microbes not only compete for inorganic but also for organic N sources.

The increasing awareness of the role of organic N in plant nutrition and the surprising competitive strength of plants also led to a paradigm shift concerning N mineralization. While the old paradigm was centred around a simple N mineralization term (ammonification of organic N) as the rate limiting step, a more holistic definition of N mineralization takes into account two steps, i.e. (i) the depolymerization of organic macromolecules to bioavailable monomeric dissolved organic N (DON), and (ii) ammonification of bioavailable DON to ammonium. The new paradigm of N cycling in forest ecosystems leaves the core concept unchallenged – that microorganisms are responsible for breaking down complex organic material into plant-available forms. However, it widens the definition of N mineralization into the steps of depolymerization of polymers into monomers and subsequent ammonification to ammonium (Fig. 1).

Biological dinitrogen (N2) fixation and atmospheric N input increases the stocks of reactive, potentially bioavailable N in forest ecosystems, while emission of N gases [i.e. ammonia (NH3), nitric oxide (NO), nitrous oxide (N2O) and N2] and leaching of DON and nitrate remove N from forest ecosystems. Direct N supply to plants via symbiotic N2 fixation, with its huge potential for N input in the magnitude of 100 kg N ha−1·year−1, is widespread in tropical forests but is paradoxically less abundant in N-limited forests of the temperate and boreal zone (Cleveland et al. 1999; Crews 1999; Houlton et al. 2008; Van der Heijden et al. 2008). In contrast, N2 fixation by free-living microorganisms is abundant in temperate and boreal forests, but renders comparably low amounts of N input (<3 kg N ha−1·year−1) into forest ecosystems (Cole 1995; Cleveland et al. 1999; Van der Heijden et al. 2008). N fixed by free-living microorganisms is incorporated into the microbial biomass. Therefore, dieback of microorganisms and subsequent depolymerization is a prerequisite for such N to enter the pools of competitive N partitioning between plants and microorganisms. The last step of denitrification, i.e. the reduction of N2O to N2, converts reactive N back into its inert form and thus closes the N cycle. Denitrification has been systematically underestimated in forest ecosystems in the past, which is primarily associated with the methodological failure of the acetylene inhibition method for denitrification measurements and a lack of direct N2 measurements (Bollmann & Conrad 1997;Davidson & Seitzinger 2006; Groffman et al. 2006; Dannenmann et al. 2008). Thus, denitrification may claim a much more important role in overall nitrate availability in forest soils and the competitive balance of nitrate partitioning between plants and microorganisms than previously thought. In general, these paradigm shifts in N cycling in terrestrial ecosystems enhance the number of processes (see Fig. 1) that have to be considered in order to evaluate the consequences of climate change stresses on cycling of and competition for N in forest ecosystems.

Temperature and Heat Impacts on N Balance of Forests

  1. Top of page
  2. Abstract
  3. Introduction
  4. Paradigm Shifts in Terrestrial N Cycling
  5. Temperature and Heat Impacts on N Balance of Forests
  6. Drought Impacts on N Balance of Forests
  7. Flooding Impacts on N Balance of Forests
  8. Impacts on N Uptake
  9. Interactive Effects on N Balance of Forests
  10. General Conclusions
  11. Acknowledgement
  12. References

Impacts on microbial release and consumption of bioavailable N

The expected increase in temperature within this century will very likely affect nutrient availability in terrestrial ecosystems, since the biological processes of decomposition, N mineralization and nitrification are in general increased with temperature (Stark & Firestone 1996; Hart & Perry 1999; Shaw & Harte 2001a; Emmett et al. 2004; Domisch et al. 2006). A meta-analysis of soil warming studies was conducted by Rustad et al. (2001). Despite the results being variable across the investigated sites, the authors concluded that soil warming in general results in increased net N mineralization and thus increased inorganic N availability in the soil. The temperature-induced increases in net N mineralization may be caused by a proportionally larger temperature response of gross N mineralization than immobilization (Hoyle et al. 2006). This could be related to the rapid assimilation of labile carbon (C) at elevated incubation temperatures, which might limit the C supply to heterotrophic inorganic N immobilizing microorganisms (Hoyle et al. 2006; Cookson et al. 2007). Hence, at increasing temperatures, the balance of gross ammonification and heterotrophic ammonium immobilization may be altered at the expense of immobilization, causing accumulation of inorganic N in the soil.

In a study by Hart & Perry (1999), soil cores were exposed to higher mean annual air (+2.4 °C) and soil (+3.9 °C) temperatures at a depth of 10 cm by transfer along an altitudinal gradient from 1310 to 490 m a.s.l. in an old-growth forest in Oregon, USA. Net N mineralization and nitrification were more than doubled in these cores compared to the in situ transferred controls. This effect could be explained by C availability as a crucial factor in the balance of ammonium consumption by heterotrophic immobilization or autotrophic nitrification. Reduced C availability favours net nitrification ‘at both ends’ because it increases the substrate for autotrophic gross nitrification via reduction of heterotrophic ammonium immobilization and, furthermore, also limits nitrate immobilization by heterotrophic microorganisms (Booth et al. 2005; Dannenmann et al. 2006). The temperature response of nitrification has an optimum in the range 20–35 °C (Barnard et al. 2005). Thus, global warming might be expected to promote nitrification across a wide range of forest ecosystems. However, further studies investigating the warming response of net nitrification revealed somewhat contradictory results. Verburg et al. (1999) found insignificantly increased rates of net nitrification (+50%) and Shaw & Harte (2001b) reported decreased rates (−28%). It remains an open question whether these different results reflect variable responses to warming, e.g. due to different microbial communities involved, or if other environmental controls of net nitrification are inferred. In general, our understanding of the dynamic response of microbial production and consumption of both ammonium and nitrate in forest soils to global warming and heat is hampered, since the available studies were confined to investigation of net rates of N transformations in the soil, which comprise both production and consumption of inorganic N, and therefore do not allow judgement of temperature and heat effects on actual rates of N turnover in forest soils. Thus, there is an urgent need for studies investigating the response of warming effects on gross rates of ammonification and nitrification, as well as microbial immobilization across a wide range of forest ecosystems.

Denitrification, i.e. the stepwise reduction of nitrate to nitrite, nitric oxide, nitrous oxide to molecular dinitrogen, leads to a gaseous loss of N from forest ecosystems and reduces nitrate availability for microbial immobilization, dissimilatory nitrate reduction to ammonium, and plant N uptake. Due to extreme methodological difficulties in the quantification of denitrification, especially concerning the dominant end product N2 (Bollmann & Conrad 1998; Butterbach-Bahl et al. 2002; Davidson & Seitzinger 2006; Groffman et al. 2006; Dannenmann et al. 2008), little is known about actual rates of denitrification in forest ecosystems. The current knowledge on denitrification in forest ecosystems is largely based on the acetylene inhibition method, which has been shown to result in severe, irreproducible underestimation of N2 losses (Bollmann & Conrad 1998;Groffman et al. 2006). The few available direct and reliable measurements of denitrification in forest soils supported the latter hypothesis by revealing unexpectedly high N2 emissions (Butterbach-Bahl et al. 2002; Dannenmann et al. 2008). These N2 losses currently represent the largest uncertainty of the N cycle in forest ecosystems (Galloway et al. 2004; Groffman et al. 2006). Under field conditions, temperature effects on both the individual reduction steps of denitrification and total denitrification rates may often be masked by the domination of environmental controls such as pH, nitrate concentration and soil water content (Dannenmann et al. 2008; Scheer et al. 2009). The temperature optimum of denitrification is thought to be similar to that of nitrification (Barnard et al. 2005). Barnard et al. (2005) performed a meta-analysis of studies investigating warming effects on denitrifying enzyme activities (DEA). Of the six investigated studies, only one reported a significant effect (+44%), while the other five studies could not demonstrate a significant positive response (+<20%). However, since it is unclear how DEA is related to actual rates of denitrification, these studies rarely clarify our understanding of the response of denitrification to warming.

As indicated above, there is increasing evidence that uptake of amino compounds contributes to tree nutrition, not only in boreal but also in other forest ecosystems (e.g. Schimel & Chapin 1996; Lipson & Monson 1998; Näsholm et al. 1998; Schmidt & Stewart 1999; Finzi & Berthrong 2005; Warren 2006; Harrison et al. 2007; Dannenmann et al. 2009). The availability of these amino compounds in forest soils is related to the activity of proteolytic enzymes, as shown for cold temperate forests of the USA (Berthrong & Finzi 2006) and in Alaskan taiga ecosystems (Kielland et al. 2007). Due to the extremely short mean residence time of amino acids in forest soils (Finzi & Berthrong 2005; Berthrong & Finzi 2006), the depolymerization process can directly regulate the availability of monomeric DON compounds in the soil. However, knowledge on production and consumption of amino compounds in soils and potential environmental controls is very scarce (reviewed by Näsholm et al. 2009). Contrary to expectations based on biochemical modelling, a study from taiga ecosystems even implies a negative relationship between temperature and amino acid turnover (Kielland et al. 2007; Näsholm et al. 2009). In view of the lack of knowledge, there is an urgent need for a set of studies investigating potential climate change effects on production and consumption of bioavailable monomeric DON compounds in a wide range of forest ecosystems. These studies should not be restricted to investigation of proteolytic enzyme activities or corresponding gene abundances, but also include measurements on actual flux rates of production and consumption of monomeric organic N compounds.

Impacts on N uptake

Mycorrhizal roots of trees are able to take up inorganic and organic N from the soil solution. Especially in boreal forests, the absorption of organic N by trees seems to play a significant role, albeit quantitative estimations of the role of organic N in plant N nutrition are still highly uncertain (Näsholm & Persson 2001; Warren 2006; Näsholm et al. 2009). Among the inorganic N compounds, ammonium is preferred by most tree species compared to nitrate (Plassard et al. 1991;Kronzucker et al. 1996; Von Wirén et al. 2000; Glass et al. 2002). Low-affinity transport systems (LATS) are responsible for uptake of external nitrate and ammonium concentrations above approximately 1 mM (Siddiqi et al. 1990; Cerezo et al. 2001). At lower concentrations – typical for the soil solutions of natural forest ecosystems – high-affinity transport systems (HATS) are prevalently used (Behl et al. 1988; Camanes et al. 2009). Both, LATS and HATS possess distinct constitutive and inducible components (Glass et al. 2001; Okamoto et al. 2003). LATS have a higher capacity for N uptake than HATS. Thus HATS seem to be important for N acquisition at lower external concentration, whereas LATS are responsible for mass amounts of transport when external concentrations are high (Touraine & Glass 1997). For nitrate uptake, a consistent decrease of HATS activity was observed when root temperature decreased in a range from approximately 25–5 °C (Malagoli et al. 2004). The strong response of HATS to root and soil temperatures may be explained by the putative enzymatic nature of the HATS, as proposed by Glass et al. (1992). LATS are less sensitive to changes in root temperature, supporting the hypothesis that LATS may constitute an ionic channel (Glass et al. 1990). The preference for ammonium compared to nitrate uptake by roots of forest trees in the field can generally be explained by different transport capacities of nitrate and ammonium transporters, but also by different temperature optima for nitrate compared to ammonium net uptake (Gessler et al. 1998; BassiriRad 2000). At lower soil temperatures (up to 10 °C) nitrate uptake is minute, whereas ammonium uptake already reaches >50% of its temperature-dependent maximum in spruce and beech (Gessler et al. 1998). With increasing soil temperature (>10 °C) the preference for ammonium is in general maintained but the relative proportion of nitrate taken up increases (Marschner et al. 1991; Gessler et al. 1998; BassiriRad 2000). However, not only is the uptake preference, i.e. the N ion form is preferentially absorbed, strongly affected by soil temperature, but – depending on latitude, longitude and altitude – also total pedospheric N uptake (as opposed to ‘atmospheric N uptake’) might be restricted by (low) temperatures during particular periods of the growing season (e.g. spring), or temperature might even be the main factor controlling N uptake. It is clearly demonstrated that soil temperatures directly affect nitrate and ammonium uptake of tree species under field conditions (Chapin et al. 1986; Gessler et al. 1998). For spruce and beech growing in temperate forests, net uptake of ammonium showed a seasonal course with maximum uptake rates in mid-summer and strongly restricted uptake rates in spring and autumn, when soil temperatures were below approximately 12 °C (Gessler et al.1998).

Studies investigating different tree species (Millard & Proe 1992; Millard 1996; Stephens et al. 2001; Carswell et al. 2003) showed that the N demand in spring could almost completely be satisfied by remobilized N from the trees’ storage tissues that are filled with N particularly in autumn when the N demand for growth and development is low (Millard 1996; Quartieri et al. 2002). N uptake at the beginning of the growing season may, therefore, be of minor significance (Gessler et al. 1998; Carswell et al. 2003). Thus, with increased spring temperatures the seasonal pattern of N acquisition might change from lower requirement for N in storage tissues as compared to early spring growth. In this way pedospheric N uptake of trees is coupled tightly to whole tree N balance. A cycling pool of amino acids senses the N status and also acts as a signal to adapt N absorption by the roots to N demand (Gessler et al. 2004). To date, it is not known whether potentially lower depletion of N storage pools during spring leads to overall increased growth potential or whether other factors restrict growth, thus inhibiting N uptake during the rest of the growing season. In addition, pedospheric N uptake also depends on transpiration rate – or more specifically on xylem flow-mediated accumulation of amino compounds – regulating N uptake (Gessler et al. 2002). As a consequence, it remains to be proved whether higher soil temperatures in spring increase early season N uptake by roots at all, since transpiration is either low (evergreens) or not present (deciduous trees) at that time.

In the nitrate and ammonium concentration range observed in the soil solution of natural forest ecosystems, N uptake is mediated by active transport processes, consequently temperature optima of net uptake have been observed (Gessler et al. 1998). Under a more-or-less closed canopy, soil temperature in temperate forests is generally below this optimum. For example, Gessler et al. (1998, 2005) observed a soil temperature range between 10 and 16 °C in the upper soil layers of beech and spruce forests during the growing season. An increase in air temperature, and consequently soil temperature, in the range predicted by current climate projections (IPCC 2007) is thus not likely to exceed the temperature optimum of N uptake. However, in open canopies, e.g. after storm events, clear cutting or intensive selective felling, soil temperatures in the uppermost soil layer might approach the optimum for net N uptake (cf. Holst et al. 2004). Further increases in air temperature might thus also reduce N uptake under such conditions.

Impacts on competition between plant species and between plant and microbial uptake

The general increase in net N mineralization with increasing temperature may indicate that the intensity of competition between plants and microorganisms might decrease at higher temperatures due to higher availability of inorganic N in soils. However, competition for inorganic N might be enhanced by a temperature-promoted increase in N acquisition by plants, but, for several reasons, these statements constitute testable hypotheses for further studies. First, very few studies have investigated plant–microbe competition in forest ecosystems, and we are not aware of any study investigating temperature effects on plant–microbe competition in forests in a process-oriented way, i.e. the simultaneous determination of actual process rates of N partitioning in the plant–soil system. Second, the available studies are mainly simple tracer studies rather describing medium- to long-term 15N recovery rates in plant and soil N pools or investigating actual competitive processes of N partitioning between plants and microorganisms. Third, the concept of net rates of N turnover as an index of plant-available N is not compatible with a modern, integrative view of terrestrial N cycling involving organic N uptake by plants as well as successful competition for both organic and inorganic N by plants. Additionally, in the available studies on soil N turnover in forest ecosystems, plants were removed before the experimental incubation began, and thus plant–microbe interactions were eliminated in these kinds of experiment. Hence, future challenges include the development of methodological assays to investigate plant–microbe competition in a process-oriented way without eliminating competition before the experiment begins.

In this context, the gas phase of the soil has to be considered as a path of exchange of signalling compounds between microorganisms and roots. For example, recent studies indicated that nitric oxide (NO) produced and released by soil microorganisms during nitrification and denitrification (Gasche & Papen 2002; Kitzler et al. 2006) can regulate the acquisition of organic N by roots (Simon et al., in preparation); with increasing NO abundance in the gas phase, organic N uptake by beech roots was enhanced. Thus, increasing nitrification activities can be counter-balanced by enhanced uptake capacity of roots for organic N, thereby possibly avoiding a competitive disadvantage in the acquisition of organic N.

Drought Impacts on N Balance of Forests

  1. Top of page
  2. Abstract
  3. Introduction
  4. Paradigm Shifts in Terrestrial N Cycling
  5. Temperature and Heat Impacts on N Balance of Forests
  6. Drought Impacts on N Balance of Forests
  7. Flooding Impacts on N Balance of Forests
  8. Impacts on N Uptake
  9. Interactive Effects on N Balance of Forests
  10. General Conclusions
  11. Acknowledgement
  12. References

Impacts on microbial release and consumption of bioavailable N

Low water potential is a common limitation of microbial activity in soils and may lead to total inhibition of microbial activity in extremely dry soils (Schimel et al. 2007; Borken & Matzner 2009). The effect of drought on N mineralization is more pronounced at high compared to low soil temperatures. The physiological response of soil microorganisms to drought, i.e. a decrease of osmotic potential in the cell, followed by dehydration and, finally, death of microorganisms, is well understood and has recently been reviewed by Borken & Matzner (2009). In general, the decrease of microbial activity under drought conditions seems to be related to the length of the drying period. However, the actual effects on process rates of N turnover, especially on gross rates of ammonification, nitrification, denitrification and microbial immobilization, and the variable sensitivity of these processes to drought stress in various forest ecosystems are still uncertain. A common paradox may be that despite microbial activity strongly decreasing under drought stress, the availability of DON in the soil may strongly increase due to dieback of microbial biomass, thereby promoting N availability for plant uptake (Borken & Matzner 2009; Dannenmann et al. 2009).

Studies manipulating drought effects on soil N pools in situ do not reveal a clear result. For example, Johnson et al. (2002) found no effect of drought (i.e. 33% reduced precipitation) on inline image and inline image in soil solution after several years or even on total N content in the soil after 12 years (Johnson et al. 2008). In the latter study, the authors concluded that the response of total soil N, as a very static parameter to climate change, is a ‘trans-scientific question’ (Rastetter 1996) that can only be assessed by simulation modelling.

Gross nitrification is thought to be inhibited at very low soil moisture levels, to increase with soil moisture to an optimum, and then to decline as the soil becomes saturated (Smith et al. 2003). Optimum soil moisture was in the range of 60–70% of the waterholding capacity in a mountainous beech forest in southern Germany (Dannenmann et al. 2006; Fig. 2). Climate change is expected to reduce the soil water availability far beyond the optimum soil moisture content for nitrification and, thus, should reduce the significance of this process in forest ecosystems under drought conditions.

image

Figure 2.  Dependency of gross nitrification rates on soil microbial biomass N content and soil water availability (expressed as % of the WHC) in a mountainous beech forest of southern Germany. Points are measurements and the mesh is a second order polynomial regression model. R2 of the regression model is 0.44. SW, NE, NW: exposure of slope. N-exposed slopes are characterized by cool-moist microclimate. S-exposed slopes by warm-dry microclimate. C: untreated control stands, T: thinned stands. Taken from: Dannenmann et al. (2006), with permission from Springer Science and Business Media.

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Denitrification an anaerobic process can clearly be expected to decrease with decreasing soil water content; however, aerobic pathways also exist (see Wrage et al. 2001). Furthermore, denitrification may also persist in rather dry soils when oxygen is depleted in soil microsites following high rates of soil respiration. The limited number of studies on actual denitrification rates in forest soils confirmed a positive relationship between soil moisture and total denitrification (Butterbach-Bahl et al. 2002; Gessler et al. 2005; Dannenmann et al. 2008). Overall, our basic mechanistic understanding of denitrification implies that N loss from forest ecosystems will slow under drought conditions. However, further studies are urgently required in order to verify this hypothesis for a broader range of forest ecosystems, in view of the lack of reliable measurements of denitrification rates including N2 emissions from forest soils.

Impacts on N uptake

Beside soil physical and chemical properties, such as soluble inorganic and organic N concentrations in the soil solution, diffusion and mass flow (BassiriRad et al. 1999; Gessler et al. 2005), which are directly dependent on the amount of soil water available, inorganic and organic N availability for trees is also determined by nitrate, ammonium and amino acid uptake kinetics of their (mycorrhizal) roots (Wallenda & Read 1999; Bauer & Berntson 2001; Gessler et al. 2005). In contrast to temperature effects on nitrate and ammonium uptake kinetics and on net uptake rates (e.g. Marschner et al. 1991; Gessler et al. 1998; BassiriRad 2000), not much is known about the effect of increased drought on the physiology of inorganic N transport. Moreover, to our knowledge, there is no publication addressing the effect of water availability on amino acid uptake by tree roots.

Most experimental approaches applied hitherto do not allow distinguishing the effect of reduced water availability on the different processes involved in N nutrition of trees. Assessments of either C or N stoichiometry of tree tissues (Sardans et al. 2008) or the size of the soluble N pool cycling within a tree (e.g. Nahm et al. 2007) allow characterization of the N nutrition status but do not provide hints as to whether changes in the N status are due to changes in soil N availability or root N uptake capacity or both. The application of 15N to the soil either as 15N enriched litter (e.g. Zeller et al. 2000) or in the form of 15N-labelled nitrate and/or ammonium (e.g. Fotelli et al. 2004), and the tracing of its fate have comparable constraints. For example, Fotelli et al. (2002) observed decreased incorporation of 15N-labelled nitrate (>50%) in beech seedlings when the soil water potential decreased from field capacity to c. −0.8 MPa. However, these authors could not exclude that this effect was due to the reduced mobility of nitrate in the soil under drought conditions, and thus not directly related to plant physiology. Mainly, the incubation of intact roots in nutrient solutions as described by Gessler et al. (1998) and Lucash et al. (2007) allows direct characterization of N uptake capacity.

Applying 15N labelling, Gessler et al. (2005) assessed nitrate and ammonium uptake kinetics of non-excised intact mycorrhizal roots of adult beech trees on two aspects (SW and NE exposed) of a site, which strongly differed in soil water availability (cf. Gessler et al. 2001). Whilst KM values for net uptake of both inorganic ions were not different among aspects, there were distinct differences in the maximum rate of nitrate uptake (Jmax), and thus the uptake capacity. At the moister NE aspect, Jmax values were between 50 and 110 nmol·g−1 root FW h−1, which is in the range observed for mycorrhizal beech roots under controlled conditions with optimum water supply (Kreuzwieser et al. 2000). At the drier SW aspect, Jmax was reduced by more than 50% (Gessler et al. 2002) during the whole growing season. As differences in nitrate availability and rooting patterns were ruled out as responsible for the observed inter-aspect difference, a long-term effect on nitrate transport of continued water depletion was assumed (Gessler et al. 2004). As to whether differences in mycorrhizal colonization (cf. Shi et al. 2002), nitrate transporter abundance/activity and/or the expression of different nitrate transporters are responsible for the drought-driven reduction of nitrate uptake remains to be elucidated. Gobert & Plassard (2002) concluded that mainly the fungi determine net nitrate uptake capacity of mycorrhizal roots. Shi et al. (2002) observed not only intense changes in the species composition of fungal partners in beech mycorrhizae as a consequence of drought, but also an accumulation of stress-related metabolites in the mycelia. Thus, the reduced net nitrate uptake capacity at the warm–dry SW aspect may be a likely result of impaired fungal activities.

As already mentioned, transpiration, and thus soil water extraction by tree roots, also affects ion mobility in the soil and might be involved in regulating inorganic N transport into roots. On a diel basis, Gessler et al. (2002) observed net uptake of mycorrhizal roots of beech to be closely related to transpiration rate. When transpiration rates were high, ammonium net uptake had a maximum and vice versa. The authors assumed this effect was not directly related to water transport, but to the transpiration-mediated transport or accumulation of amino compounds in the fine root tissues. When transpiration rates were low, amino acids accumulated in the roots, thereby inhibiting N uptake, as described by Imsande & Touraine (1994). If such transpiration-driven mechanisms are effective on longer time scales, a reduction of N uptake by reduced transpiration under drought conditions has to be expected.

Another factor, which potentially influences the total N uptake of forest stands, is not directly related to the physiology of N uptake. Reduced soil water availability and extended drought periods are known to reduce fine root biomass (Mainiero & Kazda 2006; Cudlin et al. 2007; Konopka et al. 2007), and thus decrease the nutrient-absorbing surface. Since fine root turnover rates of trees can be >1.5 years (Ostonen et al. 2005), drought-induced fine root dieback might affect N acquisition of trees over time periods that exceed the drought event itself.

Impacts on competition between plant species and between plant and microbial uptake

There are various mechanisms of how drought may influence the competition for N between plants and free-living soil microorganisms. These mechanisms either include an alteration of spatial compartmentalization, i.e. a ‘spatial avoidance’ of competition, the avoidance of direct competition by the uptake of different N forms, or altered direct competition for N between plants and microbes. Drought-induced reduction of nutrient availability in soil may be compensated in adult trees via N uptake from deeper soil layers or even the water table (McKinley et al. 2009), provided roots are capable of accessing these N sources. In general, the access of roots to deep soil N reserves or the water table may be a strategy for old-growth trees to avoid competition with understorey species and free-living microorganisms.

Dannenmann et al. (2009) reported a severe breakdown of the biomass of free-living microorganisms after a strong drought period and intense drying/rewetting events in the soil of a mountainous beech forest in southern Germany while the abundance of the most important mycorrhizal fungus, Cenococcum geophilum, was not affected. Dead microbial cells strongly increased the DON concentrations in soil, and trees took up amino acids at high rates. The authors hypothesized that reduced rhizodeposition of labile C compounds contributed to the decline of free-living microorganisms, while C supply to mycorrhizae was sustained. Thus, trees may transiently strengthen their competitive power at the expense of free-living microorganisms under drought stress (Dannenmann et al. 2009). Reduced C availability in the soil alters the balance of major ammonium-consuming microbial processes in forest soils in favour of autotrophic nitrification and at the expense of ammonium immobilization, and additionally inhibits microbial nitrate consumption (Booth et al. 2005; Dannenmann et al. 2006). Therefore, reduced C allocation to soil microorganisms due to reduced root exudation under drought conditions can be expected to increase soil inline image concentrations, as long as water shortage does not completely limit gross ammonification and nitrification. Under such moderate drought conditions, nitrate availability in the soil increases and trees, given that moisture or physiological constraints (see Gessler et al. 2005) do not limit uptake, may have greater access to nitrate. Such an effect may be further promoted by reduced nitrate consumption via denitrification under drought conditions (Butterbach-Bahl et al. 2002; Dannenmann et al. 2008). However, in view of the lack of process-oriented investigations of drought effects on competitive patterns of N partitioning in forests, these statements only constitute testable hypotheses for future studies.

Flooding Impacts on N Balance of Forests

  1. Top of page
  2. Abstract
  3. Introduction
  4. Paradigm Shifts in Terrestrial N Cycling
  5. Temperature and Heat Impacts on N Balance of Forests
  6. Drought Impacts on N Balance of Forests
  7. Flooding Impacts on N Balance of Forests
  8. Impacts on N Uptake
  9. Interactive Effects on N Balance of Forests
  10. General Conclusions
  11. Acknowledgement
  12. References

Besides its intensively discussed effects on temperature, global climate change will also affect precipitation patterns (IPCC 2007). However, in contrast to temperature changes, predictions of future trends in precipitation patterns using regional climate models (RCMs) are highly uncertain. For Europe, it is assumed that future changes in precipitation will differ strongly in their spatial and temporal distribution (Gessler et al. 2006). Modelling approaches indicate a general tendency for higher precipitation in northern Europe but decreasing precipitation in southern Europe (Hegerl et al. 1994; Parry 2000; Räisänen et al. 2004; Christenson et al. 2006). In large parts of Europe, the seasonal distribution of precipitation will shift with decreasing precipitation in summer (up to −30%) and higher amounts of rainfall in late winter and spring (up to 20%) (Bardossy & Caspary 1990; Schönwiese et al. 1993; Kunstmann et al. 2004). Moreover, there is a clear tendency that extremes will increase through rising daily precipitation events (Christensen & Christensen 2003; Giorgi et al. 2004; Frei et al. 2006; Gessler et al. 2006; Kjellström et al. 2007) with the consequence that flooding events may become more frequent and severe for most of Europe (Kundzewicz et al. 2006). Simultaneously, there will be a higher risk of temporary water-logging at sites with clay-rich soils.

Impacts on microbial production and consumption of bioavailable N

Flooding and waterlogging are important environmental factors leading to a reduction in the oxygen concentration in soils. Due to the 105-fold lower diffusion velocity of oxygen in water than in air, the supply of soils with oxygen becomes negligible and, because of a steady consumption of oxygen by soil microorganisms and plant roots, oxygen concentrations quickly decline causing reduced redox potentials in the affected soils (Vartapetian & Jackson 1997). Flooded and waterlogged soils are therefore usually characterized by hypoxia in the upper layer in contact with oxygenated floodwater and are mainly anaerobic in deeper soil layers (Ponnamperuma 1994). Prolonged water saturation of soils can result in dramatic changes in the chemical properties of soils, but also the dynamics of bacterial populations and microbial activities are strongly affected (Lundgren & Söderström 1983). Besides determining soil oxygen availability and redox potential, soil water concentrations also influence nutrient availability to plants and microbes (Gambrell et al. 1991). Soil water concentrations and oxygen availability are therefore important factors controlling N cycling in riparian forest ecosystems (Holtgrieve et al. 2006).

Most of the N abundant in soils is associated with organic matter in the soil humic material. This N pool becomes available to plants by microbial activity in the processes of depolymerization of organic macromolecules, and further mineralization to inorganic N (Fig. 1). Mineralization generally occurs in both aerobic and anaerobic soils, but at different rates (Gambrell et al. 1991). High-soil water concentrations and low-oxygen availability affect the turnover rate of transformation of N and organic matter (Ohte et al. 1997; Franzluebbers et al. 2001; Gilliam et al. 2001; Schuur 2001; Austin 2002; Corre et al. 2002; Holtgrieve et al. 2006), the release of mineral nutrients (Silver et al. 2000; Chadwick et al. 2003), and enhance hydrologic losses of N (Brooks & Williams 1999; Brooks et al. 1999). Nevertheless, under conditions of prolonged anoxia or seasonal waterlogging, the concentrations of ammonium in forest soils might increase (Koschorrek & Darwich 2003; Satti et al. 2003), as the N requirement for anaerobic microbial metabolism is lower than in a normoxic environment, thereby leading to an accumulation of this inorganic N form (Gambrell et al. 1991). Conversion of ammonium to nitrate by nitrifying bacteria is inhibited under these conditions due to the lack of oxygen (Ohte et al. 1997). Also, laboratory experiments have indicated that flooding causes an increase in ammonium concentrations and a sharp decrease in nitrate availability at the same time (Alaoui-Sosséet al. 2005).

Many soil bacteria are aerobic organisms. Under conditions of hypoxia and/or anoxia, in some microbial groups nitrate may replace oxygen as the terminal electron acceptor leading to denitrification and/or nitrate ammonification (Laanbroek 1990). Oxygen availability is, therefore, the most important factor controlling these processes in soils (Robertson 1989; Bollmann & Conrad 1998). Loss of nitrate during oxygen deficiency is achieved by (i) DNRA (dissimilatory nitrate reduction to ammonia), (ii) nitrate dissimilating bacteria reducing nitrate to nitrite, and (iii) denitrifiers reducing nitrate to the gaseous N forms nitrous oxide and/or dinitrogen (Nijburg & Laanbroek 1997; Bodelier et al. 1998; Cai 2002). Consumption of nitrate by microorganisms through denitrification is an important atmospheric loss of N from soils (Henrich & Haselwandter 1997). As a consequence, denitrification leads to enhanced competition for nitrate between roots and soil microorganisms in forest soils (Alaoui-Sosséet al. 2005) and, as a further consequence, it influences nutrient uptake by plants (Drew 1991).

The occurrence of alternate aerobic and anaerobic conditions has been demonstrated to strongly enhance organic and inorganic N transformations in soils (Birch 1960; Patrick & Wyatt 1964; MacRae et al. 1967; Reddy & Patrick 1975; Koschorrek & Darwich 2003). Organic N is converted to ammonium, which is oxidized to nitrate under aerobic conditions; under subsequent anaerobic conditions, the resulting nitrate is denitrified (Reddy & Patrick 1975) and can be lost to the groundwater by leaching or to the atmosphere by nitrous oxide and/or dinitrogen emission (Holtgrieve et al. 2006). However, obviously the soil type plays an important role for N mineralization during flooding. Riparian soils produce the highest mineralization flush during periods of submergence (McIntyre et al. 2009). Another factor that may be associated with flooding is erosion and deposition, which tend to diminish the organic matter pool in soils. As a consequence, mineralization is reduced with lower release of N from this pool.

Impacts on N Uptake

  1. Top of page
  2. Abstract
  3. Introduction
  4. Paradigm Shifts in Terrestrial N Cycling
  5. Temperature and Heat Impacts on N Balance of Forests
  6. Drought Impacts on N Balance of Forests
  7. Flooding Impacts on N Balance of Forests
  8. Impacts on N Uptake
  9. Interactive Effects on N Balance of Forests
  10. General Conclusions
  11. Acknowledgement
  12. References

Anoxia strongly affects terrestrial plants, which as aerobic organisms depend on a steady supply of oxygen to maintain essential processes such as mitochondrial respiration (Vartapetian & Jackson 1997). The tree species’ tolerance to oxygen deficiency, the duration of the anoxic period, the season and a variety of plant internal factors, all determine the degree of damage that may occur (Drew 1991; Kreuzwieser et al. 2004). The development of such damage is a consequence of plant internal disturbances, including impaired carbon and N metabolism.

N is highly abundant in plants and therefore has to be taken up in considerable amounts by the roots (Marschner 1995). It is estimated that the portion of N uptake makes up more than 80% of total ion uptake by the roots of plants (Marschner 1995). Disturbing N uptake may, therefore, cause a loss of vitality of the trees affected. As expected from impaired energy metabolism, and supporting the assumption of minimized oxygen consumption under conditions of hypoxia/anoxia, N uptake by roots of trees is dramatically reduced (Larson et al. 1992; Vaast et al. 1998; Kreuzwieser et al. 2002; Parolin et al. 2002; Islam & Macdonald 2009). Depending on the flooding tolerance of trees, nitrate and ammonium uptake is more strongly reduced in sensitive than in tolerant species (Kreuzwieser et al. 2002). Kreuzwieser et al. (2009) demonstrated that reduced nitrate uptake in poplar is most probably caused by reduced availability of nitrate transporters, since transcript abundance of nitrate transporters disappeared during flooding. Moreover, it has been shown that the reduction of ammonium uptake was more pronounced in flood-sensitive than in more tolerant species, which may be of particular significance since nitrate is consumed by denitrification under hypoxic conditions in soils, and N compounds are usually present mainly as ammonium in submerged soils (Armstrong 1982; Ponnamperuma 1984) as long as nitrate import with the floodwater is not excessive.

Interactive Effects on N Balance of Forests

  1. Top of page
  2. Abstract
  3. Introduction
  4. Paradigm Shifts in Terrestrial N Cycling
  5. Temperature and Heat Impacts on N Balance of Forests
  6. Drought Impacts on N Balance of Forests
  7. Flooding Impacts on N Balance of Forests
  8. Impacts on N Uptake
  9. Interactive Effects on N Balance of Forests
  10. General Conclusions
  11. Acknowledgement
  12. References

Interactive effects of temperature/heat and drought on microbial release and consumption of bioavailable N

Various approaches for investigating combined effects of several parameters altered by climate change, such as temperature increase and drought, have been applied, i.e. (i) observations along natural climatic gradients in ecosystems (McTiernan et al. 2003; Kueppers et al. 2004; Gessler et al. 2005; Dannenmann et al. 2006, 2007), (ii) experimental manipulation of temperature and soil moisture in situ (Peterjohn et al. 1994; Chapin et al. 1995; Harte & Shaw 1995; Marion et al. 1997; Norby et al. 1997; Beier et al. 2008), (iii) exposure of ecosystem components, e.g. soil monoliths or small intact plant–soil microcosms to altered climate conditions in the field (e.g. along altitudinal gradients) (Jonasson et al. 1993; Ineson et al. 1998; Hart & Perry 1999; Hart 2006) or in the laboratory (Billings et al. 1983), (iv) ecosystem modelling (Rastetter et al. 1991), and (v) full-factorial parameterization experiments under controlled conditions in the laboratory involving manipulation of temperature and water availability.

Using closed-top greenhouses, Allison & Treseder (2008) increased soil temperature by 0.5 °C and reduced soil moisture by 22% in an Alaskan boreal forest soil dominated by mature black spruce. As a consequence, inorganic N availability in the soil was slightly increased despite the overall bacterial and fungal abundances declining by over 50%. Hart (2006) transferred soil cores 730 m down-slope from a spruce–fir to a Ponderosa pine forest in the San Francisco peaks in Arizona, USA. This resulted in an increase in mean annual soil temperature of 2.5 °C but also in significantly reduced soil moisture. In this study, the combined temperature and moisture effects increased both net N mineralization and net nitrification by more than 80%, indicating that the effect of increased soil temperature dominated over the reduction in water availability. A contrasting result was obtained by Emmett et al. (2004) and Beier et al. (2008). They conducted a field-scale manipulation experiment in shrublands across a climatic gradient from Denmark to Spain. In the latter study, net N mineralization was clearly more limited by low-soil moisture than promoted by increased temperature. In contrast, the variation in litter decomposition (determined as mass loss in mesh bags) was best explained by temperature and, furthermore, experimental warming resulted in increased decomposition compared to controls. Dannenmann et al. (2007) investigated decomposition and N transformations in the litter layer of a mountainous beech forest ecosystem in southern Germany along an exposure-induced natural climatic gradient. This study confirmed the dominating role of temperature over soil moisture in litter decomposition, since litter mass loss was higher under a warm-dry microclimate compared to a cool-moist microclimate. In contrast, no differences in inorganic N concentrations, and gross and net N transformations were found in the litter layer across the microclimatic gradient, suggesting that moisture and temperature effects on N turnover and N availability outbalanced each other under the different climatic conditions. The variable response of net rates of N turnover to the interaction of temperature increase and reduced water availability in published studies implies that – depending on initial temperature and soil moisture level – either temperature or soil moisture may exert a dominating influence on climate change-induced alterations of inorganic N availability in forest ecosystems. A set of manipulation experiments along typical forest ecosystems based on a standardized experimental design would further contribute to our understanding of the dynamics of inorganic N availability with changing climate.

Gross rates of N turnover have rarely been investigated in climate change experiments. Dannenmann et al. (2006) studied gross and net N transformations during the vegetation period of three consecutive years in mountainous beech forests of southern Germany in different aspects, and thus different microclimate (warm-dry versus cool-moist). Neither mean gross transformations, nor net N transformations and inorganic N concentrations in soil, were significantly different across the natural gradient in climate. This observation may be a consequence of an effect of compensation of lower soil moisture on gross N mineralization by higher temperature (and vice versa; compare Fig. 3), but it remains unclear whether these differences are directly related to climatic factors or may be a consequence of microclimate-induced differences in further soil parameters, e.g. total C and N content and the C:N ratio. This highlights the limitations of investigation of climate change effects along natural climatic gradients (i.e. the ‘space versus time’ concept).

image

Figure 3.  Parameterization of the response of gross ammonification to soil moisture and temperature in the soil of a mixed deciduous forest of northern Germany. Gasche et al. (unpublished results).

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Full-factorial parameterization studies of the dependency of gross rates of N turnover in forest soils on soil moisture and soil temperature under controlled conditions in the laboratory have not yet been published. A recently completed laboratory parameterization of the moisture and temperature response of gross ammonification rates in the uppermost mineral soil of a mixed deciduous forest of the northern German lowlands revealed a decrease of gross ammonification with both decreasing soil temperature and decreasing soil moisture in a temperature range of 2–18 °C and a humidity range between 20% and 75% WHC (Gasche et al. unpublished results, Fig. 3). Considering global change scenarios, this may indicate an increase in gross ammonification due to higher temperatures counterbalanced by a decrease in soil water content.

Interactive effects of temperature/heat and drought on N uptake

As stated above, there is not much information on drought impacts on the physiology of nitrate and/or ammonium uptake and even less information is available on the interaction between drought and temperature. An increase in soil temperature from up to 20 to 25 °C is assumed to increase N uptake capacity of mycorrhizal roots, whereas reduced water availability is likely to have the opposite effect. There is only one study relating the temperature-dependence of nitrate uptake of beech to soil water availability (Gessler et al. 2005). At a cool-moist beech site, 96% of the specific variation of maximum concentration-dependent nitrate uptake could be explained by changes in soil temperature (see above). However, such a temperature dependency of maximum nitrate net uptake was not observed on a warm-dry site with longer periods of limitation in soil water availability to plants (Gessler et al. 2005). In addition, the absolute rates of inorganic N uptake were lower at a given temperature compared to the cool-moist site. There is some evidence that temperature sensitivity of nitrate uptake is highly modulated by the N status of roots or even the whole plant (Smart & Bloom 1991; BassiriRad 2000). In herbaceous species (BassiriRad et al. 1993; Lainéet al. 1993), soil temperature influenced nitrate net uptake only when plants were N starved. The results of Gessler et al. (2005) seem to indicate that also long-term water availability modifies the temperature dependence of nitrate uptake. The current climate projections not only predict higher temperatures but also an increase in the frequency and duration of summer droughts in Central Europe and other temperate zones (IPCC 2007). As a consequence, it may be assumed that a positive effect of increasing air, and consequently soil, temperatures on pedospheric N uptake will be counteracted by this drought response. However, more experimental evidence for interactive effects between temperature and water supply is required to achieve a more general picture.

Interactive effects of temperature/heat and drought on competition between plant species and between plant and microbial uptake

Throughout their life cycle, trees interact with neighbouring herbaceous and woody species, which might or might not differ in life form, physiology and resource requirements. Especially for tree seedlings, which co-exists with other species in the understorey, heat and drought might affect their competitive abilities and, as a consequence, their potential to acquire pedospheric N. For beech, it has been shown that mainly water availability controls the competitive interactions between seedlings and a strong and fast-growing competitor (Rubus fruticosus) (Fotelli et al. 2002). Even under optimum water supply, N uptake of beech was slightly decreased when seedlings had to share resources with a competitor. However, beech seedlings were still able to maintain growth and water balance. In contrast, natural regeneration of beech showed strongly reduced pedospheric N uptake and growth by interference from competitors during a simulated summer drought – as predicted by actual climate models. Increased air temperatures (by approximately 2 °C) intensified the negative competition effects on beech (when light was not limiting the competitor) and further reduced N-uptake by the roots (Fotelli et al. 2005). The observation of reduced ability of beech seedlings to compete for soil N with reduced soil water availability matches observations made for other tree species (Fredericksen et al. 1991; Perry et al. 1994; Backes & Leuschner 2000). However, drought-adapted tree species or provenances might perform better with increased temperatures and drought (e.g. Ducrey et al. 2008), and thus may show stronger competitive abilities.

Interactive effects of heat and drought on competitive patterns of N partitioning between plants and free-living microorganisms in forests have rarely been studied. Gessler et al. (2005) compared gross and net N transformations, as well as inorganic N uptake capacities, of beech under warm-dry versus cool-moist microclimates. In this study, the cool-moist microclimate was a model for current conditions, whereas the warm-dry microclimate represented the predicted climate change conditions in Central Europe. Gross ammonification was even stimulated under the warm-dry microclimate, but net ammonification was consistently close to zero, indicating that strong microbial competition for ammonium was limiting actual access to ammonium by trees both under warm-dry and cool-moist microclimates. Gross nitrification was temporally increased while net nitrification was consistently higher at the warm-dry site. In contrast, denitrification rates were higher under a cool-moist microclimate. Thus, there was an overall positive effect of warm-dry microclimate on microbial re-supply of nitrate. However, beech obviously could not benefit from increased nitrate availability in the soil, since net nitrate uptake capacities were found to be decreased under the warm-dry microclimate. Consequently, the authors concluded that the observed temperature sensitivity of beech is at least partly decoupled from microbial supply of inorganic N and may be a consequence of impaired N acquisition due to physiological constraints. Further integrative studies on interactive effects of increased temperature and drought on the balance of N partitioning between trees and microorganisms are required for a range of forest ecosystems. Such information will be an indispensable prerequisite for sustainable forest management under the predicted climate changes in the 21st century.

Consequences of drying–wetting cycles on microbial release and consumption of bioavailable N

The mechanistic response of soil microorganisms to drying–rewetting cycles and the consequences for net N mineralization has been excellently summarized very recently by Borken & Matzner (2009). Thus, only a short overview on the effect of drying–wetting cycles on net N mineralization is given here. Rewetting of dry soil leads to a net N mineralization flush, which may be promoted by accumulated plant and microbial necromass and lysis of microbial cells (Borken & Matzner 2009). However, the cumulative net N mineralization of drying–rewetting treatments is – compared to permanently moist control treatments – commonly lower, since the short-lived wetting pulses cannot compensate for the low mineralization during drought periods (Borken & Matzner 2009). Thus, the expected climate change-induced increased frequencies and intensities of drying–rewetting cycles can be expected to overall reduce mineral N availability in the soil. Since even single effects of temperature/heat and drought on the availability of amino compounds in the soil have not yet been investigated, there is also no information on interactive effects of climate change conditions.

Consequences of drying–wetting cycles on N uptake

Drying–rewetting events may result in an increase in spatial compartmentalization of plant–microbe competition. While during drying of soil, plants may access N from deeper soil layers, as proposed for arid woodland by Evans & Ehleringer (1994), or from the water table, and microorganisms already suffer from drought stress, microorganisms may profit from small precipitation events wetting the uppermost organic soil layer that is hardly accessibly to tree roots. Cui & Caldwell (1997) observed a strong pulse-like increase in plant available nitrate after rewetting of dry soil over a few days. The uptake capacity of mycorrhizal roots increased within 1 day after the rewetting, whereas new root growth played only a minor role. There is, however, considerable variation among species in their ability to alter uptake kinetics as a response to wetting (BassiriRad & Caldwell 1992a,b; Ivans et al. 2003), and so it remains unclear whether this physiological plasticity is of general importance for N uptake in terrestrial ecosystems. In addition, there is little information on whether a potential rapid expansion of the microbial population and its physiological activity might counterbalance the increased uptake capacity by roots (Hodge 2004).

Lipson & Monson (1998) hypothesized that drying and rewetting events allow plants to have improved access to amino acids by disrupting microbial cells, which decreases the size of competing microbial populations, but increases soil amino acid concentrations. Experiments using an alpine sedge, however, did not show such improvement. Dannenmann et al. (2009) reported that a breakdown of the biomass of free-living microbes following drought and severe drying–rewetting cycles in the soil of a beech forest ecosystem in southern Germany strongly increased the DON availability in soil. Since trees sustained their most important mycorrhizal fungal symbiont, they profited from this situation and took up amino acids and nitrate at high rates.

General Conclusions

  1. Top of page
  2. Abstract
  3. Introduction
  4. Paradigm Shifts in Terrestrial N Cycling
  5. Temperature and Heat Impacts on N Balance of Forests
  6. Drought Impacts on N Balance of Forests
  7. Flooding Impacts on N Balance of Forests
  8. Impacts on N Uptake
  9. Interactive Effects on N Balance of Forests
  10. General Conclusions
  11. Acknowledgement
  12. References

Our understanding of aboveground plant responses to environmental change is becoming clearer while the effects of global change on belowground N transformation by microorganisms are little understood. The lack of knowledge on belowground N cycling in forest ecosystems and its responses to climate change is mainly caused by methodological difficulties. Most of the available studies on microbial N fluxes or plant N uptake have measured N turnover after completely disturbing the plant–soil system. Competitive mechanisms for N between plants and soil microorganisms are likely to be eliminated in these kinds of experiments. This general methodological problem in the investigation of terrestrial N cycling also strongly limits our understanding of the effect of global change factors, i.e. drought, increased temperature and flooding or waterlogging on the competitive balance of N partitioning between plants and microorganisms in forests. Future progress will largely depend on the success of evaluating organic and inorganic N fluxes in intact plant–soil systems at different spatial scales from soil microsites to intact plant–soil microcosms and the ecosystem scale. These interdisciplinary and multiscale studies should focus on simultaneous measurements of all major N fluxes including plant uptake/release of organic and inorganic N compounds as well as microbial N conversion.

In addition to methodological difficulties, a lack of interdisciplinary approaches between plant scientists and soil ecologists/soil microbiologists limit current knowledge on the effects of global change on belowground N cycling in forest ecosystems. Linking plant physiological and soil microbial N cycling, as well as soil hydrological N transport, to more reliable estimates of ecosystem N fluxes will be a major research challenge for the coming years. In particular, this will include further development of methodological approaches and experimental assays that allow for directly assessing N turnover processes in the larger context of intact plant–soil systems where competitive mechanisms and further interactions between microorganisms and plants persist. In this context, analysis of the gaseous environment of the soil as a compartment for the exchange of signals, such as NO, between microbes and plant roots may provide new insight into the regulation of belowground N cycling in forest ecosystems.

Acknowledgement

  1. Top of page
  2. Abstract
  3. Introduction
  4. Paradigm Shifts in Terrestrial N Cycling
  5. Temperature and Heat Impacts on N Balance of Forests
  6. Drought Impacts on N Balance of Forests
  7. Flooding Impacts on N Balance of Forests
  8. Impacts on N Uptake
  9. Interactive Effects on N Balance of Forests
  10. General Conclusions
  11. Acknowledgement
  12. References

Financial support for the studies of the authors from the Deutsche Forschungsgemeinschaft (DFG) and the Bundesminister für Bildung und Forschung (BMBF) is gratefully acknowledged. We thank Magdalena Jaglitz and Ludwig Lipp for excellent help during manuscript preparation.

References

  1. Top of page
  2. Abstract
  3. Introduction
  4. Paradigm Shifts in Terrestrial N Cycling
  5. Temperature and Heat Impacts on N Balance of Forests
  6. Drought Impacts on N Balance of Forests
  7. Flooding Impacts on N Balance of Forests
  8. Impacts on N Uptake
  9. Interactive Effects on N Balance of Forests
  10. General Conclusions
  11. Acknowledgement
  12. References
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Corrigendum after online publication The authors declare no conflicts of interest.